Porewater Distribution and Benthic Flux Measurements of Mercury and Methylmercury in the Gulf of Trieste (Northern Adriatic Sea)

Estuarine, Coastal and Shelf Science (1999), 48, 415–428 Article No. ecss.1999.0466, available online at http://www.idealibrary.com on

Porewater Distribution and Benthic Flux Measurements of Mercury and Methylmercury in the Gulf of Trieste (Northern Adriatic Sea) S. Covellia, J. Faganelib, M. Horvatc and A. Brambatia a

Department of Geological, Environmental and Marine Sciences, University of Trieste, Via E. Weiss 2, 34127 Trieste, Italy b Marine Biological Station, Fornace 41, 6330 Piran, Slovenia c Jozef Stefan Institute, Jamova 39, 1000 Ljubljana, Slovenia Received 3 August 1998 and accepted in revised form 12 January 1999 The Gulf of Trieste is one of the most mercury-contaminated areas in the Mediterranean Sea. It is characterized by high mercury inputs from the Isonzo River whose tributary, the Idrijca River, drains the mercury mining area of Idrija in Slovenia where extraction activity has taken place for nearly 500 years. This appears, therefore, to be one of the most suitable sites for studying processes that affect Hg cycling in the marine environment and for determining whether sediments might act as secondary sources of mercury species in the water column. Porewater seasonal distributions of total dissolved Hg (HgT) and methylmercury (MeHg) were investigated. Using in situ benthic chambers it was possible to determine benthic fluxes of HgT and MeHg at the water–sediment interface throughout the year. Benthic fluxes were also compared with diffusive fluxes calculated from porewater profiles. The results indicate that, following hypoxic conditions which occurred in late summer in the sea-bottom layer, highest benthic effluxes and porewater concentrations of Hg and MeHg appeared during autumn and winter. This was probably due to the transition from rapid sulphate reduction in late summer to cooler temperatures, higher oxygenation of the bottom water layer, and lower microbial activity which is well suited for Hg transformation, accumulation and flux. A tentative budget based on benthic flux measurements indicates that 75% of HgT is buried in the sediment whereas 25% of HgT, approximately 23% in methylated form, is annually recycled and released at the water–sediment interface.  1999 Academic Press Keywords: porewaters; dissolved Hg; methylmercury; benthic chamber; benthic fluxes; budget; northern Adriatic Sea

Introduction Biogeochemical processes involving mercury compounds in aquatic environments are of great importance especially in those areas where mercury inputs are particularly high. The considerable interest in this element is due to the well-known toxicity of its major organic form, methylmercury (MeHg), the accumulation of MeHg in biota and its biomagnification in aquatic food-chains. The first pathway for human exposure is the consumption of contaminated fish (Fitzgerald & Clarkson, 1991). Investigations in aquatic ecosystems have focused on natural and human-induced mercury sources, the chemical speciation of Hg, and the physicochemical conditions that control mercury cycling and consequent bioavailability. Mining operations in areas rich in cinnabar ore or native mercury (Baldi & Bargagli, 1984; Benoit et al., 1994) or gold and silver mining areas using Hg amalgamation processes (Lacerda et al., 1993; Bonzongo et al., 1996) mobilize Hg for transport away 0272–7714/99/040415+14 $30.00/0

from its source. Although the flow of elemental Hg (Hg0) into the atmosphere is an important pathway of Hg cycling and mobility (Fitzgerald, 1989), riverine inflows are often the primary source of mercury input and dispersion into lakes and estuaries, either dissolved or as suspended fine particles (Cranston, 1976; Figueres et al., 1985). In some instances, it has been discovered that marine sediments contaminated by industrial effluents, e.g. chlor-alkali plants, may be secondary sources of mercury to aquatic ecosystems even though discharge has been strongly reduced or has even ceased (Bothner et al., 1980; Krom et al., 1994). However, there are cases where the release of Hg, entering the sea as sewage sludge, is rendered negligible by natural processes, thus revealing sediments as important sinks for this metal (Baldi & Bargagli, 1984). Several factors are necessary to understand mercury biogeochemistry in the aquatic environment. When mercury is introduced as HgS, the extremely low solubility and strong resistance to oxidative processes  1999 Academic Press

416 S. Covelli et al. 13°30

13°45

Monfalcone

N

Austria

Is

on

zo

45°45

ri

ve

Italy

r

Italy

AA1

Grado

Idrija Ljubljana

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–10 m

–20 m

Is on zo

CZ Adriatic Sea

Gorizia

Rizana

Koper

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Slovenia

Dr

ag

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Trieste Piran Adriatic Sea Grado

ia

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Croatia

F 1. Location of sampling site AA1 in the Gulf of Trieste (northern Adriatic Sea).

limit its recycling and toxicity. Nevertheless, both laboratory and field research have shown that in favourable environmental conditions inorganic mercury may be transformed into highly volatile elemental form (Hg0) and organomercury compounds. Microbial transformations play a key role by methylating ionic Hg (Hg2+ ), degrading MeHg, and reducing Hg2+ to Hg0 (Barkay, 1992). In the anaerobic environment, methylating activity has been attributed to sulphate-reducing bacteria whereas many other microbes methylate ionic Hg in culture (Compeau & Bartha, 1985). According to the experimental data from estuarine sediments, synthesis and stability of MeHg are enhanced under anaerobic, low-salinity conditions. In contrast, aerobic, high-salinity conditions are either less favourable for MeHg synthesis or are more suitable for its destruction (Compeau & Bartha, 1984). In this paper, a study of mercury at the water– sediment interface in the Gulf of Trieste (northern Adriatic Sea) is presented. Owing to this area having the highest concentrations of mercury observed in the whole Mediterranean basin, due to constant input for nearly 500 years, it may be considered as a ‘ natural laboratory ’, suitable for investigating mercury recycling in the coastal marine environment. Using in situ benthic chambers we determined benthic fluxes of Hg species at the water–sediment interface and compared them with diffusive fluxes. A tentative budget for total and organic mercury at the water–sediment interface was constructed to decode the amount of mercury

released and potentially bioavailable to the water column.

Environmental setting The Gulf of Trieste, located in the north-eastern part of the Adriatic Sea, is a 500 km2 shallow-water basin, reaching only 25 m depth in the central sector (Figure 1). The main freshwater and sediment supply originates from the Isonzo River in the north-west whereas secondary terrigenous sources are the Rosandra, Rizana and Dragonia streams in the south-east. The water circulation in the gulf is mostly affected by the anticlockwise northern Adriatic circulation system, by tides, with an approximate range of 0·5 m, and by the prevalently north-north-eastern wind pattern (Olivotti et al., 1986). The salinity of the waters in the mid-gulf, typically marine, ranges between 33 and 38·5 (using the Practical Salinity Scale) in the surface layer and between 36 and 38·5 at the bottom. Intensive seasonal fluctuations of surface salinity are observable in late spring in parallel with the increased Isonzo freshwater inflow which is spreading over the whole basin as a less dense layer (Ogorelec et al., 1991). Water density stratification during late summer often causes hypoxic or even anoxic events in the bottom layer in the central deepest part of the gulf (Faganeli et al., 1991a). The textural variation in sediments of the Gulf of Trieste appears to be influenced by the Isonzo riverine inputs. Suspended sediment is transported

Mercury and methylmercury in the Gulf of Trieste 417

symmetrically from the river mouth as a spreading plume. The distribution pattern of deposits, therefore, shows a progressive decrease in grain size from very sandy pelites near the river mouth to pelites in the mid-gulf (Brambati et al., 1983). The sediment accumulation rate, based on 210Pb activity, in the central part of the Gulf of Trieste is about 1 mm yr
1, but increases to about 2·5 mm yr
1 in the direction of the Isonzo River mouth (Ogorelec et al., 1991). The northern and western marine areas are dominated by carbonate material derived from the Isonzo River inflow whereas the south-eastern most part shows significant amounts of quartz feldsphatic material derived from flysch erosion. The biogenic component is abundant and mostly composed of brittle stars, sponges and tunicates. The sediments are bioturbated further by polychaetes and bivalves. Surficial sediment is populated by microalgal mats mostly composed of diatoms. The Isonzo River is the main source of Hg in the gulf, since its tributary, the Idrijca River, drains the Idrija mercury mining area in western Slovenia (Faganeli et al., 1991b). Extraction activity from the second largest Hg mine in the world was in operation for nearly 500 years but it has been gradually closed during the last 10 years. Over 5106 metric tons of Hg ore were excavated and it has been estimated that 73% of the Hg mined was recovered (Gosar et al., 1997). Although the ultimate source of Hg in the Idrija region is from base deposits, the greater part of Hg that is found in soil, surficial and deep sediments, and along the banks of the two rivers is derived from Hg remobilized by mining activity, including smelting. It is well documented that volatile forms of Hg are undoubtedly an important pathway of metal transport, dispersion and deposition throughout atmospheric precipitations (Bloom & Watras, 1989) and this is thought to occur also in the Idrija region. Concentrations of Hg in air samples exceeded 2500 ng m
3 during active mining periods, but even today, airborne Hg levels near the closed smelter and around mine shafts are very high, exceeding 300 ng m
3 (Gosar et al., 1996). Recent Hg measurements performed on the Idrjia stream sediments by Gosar et al. (1997) show values ranging mostly between 100 and 300 mg kg
1 with maxima exceeding 1000 mg kg
1 near the town of Idrija. Thus, not only the Isonzo–Idrijca river system but also the Gulf of Trieste has received, and still receives, large quantities of mercury. Several investigations on Hg content in the gulf sediments (Kosta et al., 1978; Bussani & Princi, 1979; Donazzolo et al., 1983) indicated maximum values of about 25–30 ìg g
1 at the Isonzo River mouth. It was found that an exponential

decrease of Hg content occurs which is proportional to the distance from the river mouth. The Hg concentration of 0·2 ìg g
1, obtained in the southern part, is very close to the estimated natural geochemical background value of 0·1 ìg g
1 for the Gulf of Trieste (Faganeli et al., 1991b). Materials and methods Sampling Experimental fieldwork was performed six times (approximately bimonthly) from September 1995 to October 1996 at the sampling site AA1 (4540.33 N, 1335.67 E, 21 m of depth) located in the central part of the gulf. Virtually undisturbed short cores (five replicates) were collected by scuba diving, inserting a perspex tube (6·5 cm i.d.; 40 cm length) into the sediment. The cores were immediately transported to the laboratory and the supernatant water collected (about a 5-cm thick layer above the water–sediment interface). The sediment was extruded and sectioned in 1, 1·5 and 2 cm sections down to a depth of 18 cm working in a glovebox filled with N2 to maintain inert conditions. Porewaters were extracted from wet sediment by centrifugation at 4000g for 20 min at in situ bottom-water temperature. The resulting porewaters were successively recovered in an inert atmosphere, transferred in pre-acid-cleaned containers and stored in a deep freezer until analysis. Solid phase samples were freeze-dried, homogenized with a mortar and pestle and sieved through a 420 ìm screen to remove coarse shell debris. Benthic chamber experiment Benthic fluxes were assessed in the same period as core collection, by deploying an in situ transparent benthic chamber on the sea-floor, isolating an area of sediment surface and overlying water. The box-shape chamber (505029 cm), open at the bottom side, was constructed from Plexiglas. Two stopcocks for water sampling and a flexible membrane consisting of a polyethylene bag for replacing the volume removed during sample collection, were fastened on the chamber top side. The chamber was equipped with a stirring mechanism which consisted of a rotating bar (30 cm long; 5 rpm speed) inside the chamber connected through a magnet to an electromotor, coupled with 12 V batteries, located in a separate housing on the top of the chamber. The chamber was gently placed on the sea bottom by a scuba diver. A plastic skirt, fitted to the outside of the chamber, controlled its penetration into sediment to a depth of

418 S. Covelli et al.

about 7·5 cm. Water samples from the benthic chamber were periodically, at t=0 and after 1, 6, 24 or 48 h, collected by scuba with 50 ml syringes, immediately transferred to the acid-pre-cleaned containers and stored frozen. Benthic fluxes of Hg dissolved species were calculated from the difference in concentrations, ÄC=Cf
C0, over the experiment time, Ät=(tf
t0): F=(Cf
C0) (V/A)/(tf
t0) where tf and t0 are the final and starting times, respectively, V is the benthic chamber volume and A is the sea-bottom area covered (Santschi et al., 1990). Analyses The textural composition of the sediment samples was defined by dry-sieving (sandy fraction) and by using a Micromeritics Sedigraph 5000 ET Particle Size Analyser for silty and clay components, following procedures reported by Covelli and Fontolan (1997). Porosity (Ö) was calculated by measuring the weight loss of sediments dried overnight at 110 C to constant weight and calculating Ö= (MW/ñW)/(MS/ñS)+(MW/ñW) where MW is the weight of water lost on drying, MS is the weight of dry sediment, ñW =1·025 is the water density whereas ñS is the sediment density determined on five specimens (0–1, 2–3, 5–6, 9–10, 15–16 cm) with a gas Multipicnometer (mod. Quantachrome Corp.). The total mercury in freeze-dried sediment samples was determined following the procedure of CV AAS (Ure & Shand, 1974). After overnight decomposition with suprapur HNO3 in PTFE vessels at 120 C, SnCl2 and hydroxylamine sulphate were added directly to the vessels and Hg vapours were flushed by air into the measuring cell of a Varian AAS (mod. 1250). The total carbon and sulphur contents were analysed with a CHNS Carlo Erba (mod. EA 1108) elemental analyser at a combustion temperature of 1020 C. The organic carbon and the total nitrogen in the sediments were determined on freeze-dried and homogenized samples, after acidification with 1 M HCl, using a Perkin Elmer C–H–N elemental analyser at a combustion temperature of 920 C (Hedges & Stern, 1984). Water samples, thawed at room temperature and filtered through a 0·45 ìm Millipore HA cellulose nitrate filter, were analysed for total mercury (HgT)

and methylmercury (MeHg). Reactive mercury (HgR) determination was performed only in September and November 1995. For HgR determination, a 5-ml filtered subsample was transferred into a reduction cell and SnCl2 was used as a reductant solution. Reduced Hg was measured by double amalgamation CV AFS with a detection limit of 0·2 ng litre
1 (Gill & Fitzgerald, 1987). A 10-ml subsample was transferred into precleaned teflon vials and 0·25 ml of BrCl oxidizing solution was added. After 30 min, water samples were subjected to HgT measurement using SnCl2 and double amalgamation CV AFS with a detection limit of 0·5 ng litre
1 (Bloom & Crecelius, 1983). The remaining water sample was acidified with 5 ml suprapur HCl and MeHg was extracted into methylene chloride (25 ml). The water phase was discharged and MeHg re-extracted from organic solvent to deionized water. MeHg was detected using aqueous phase ethylation, preconcentration on Tenax, GC separation and CV AFS measurement with 0·05 ng litre
1 detection limit (Horvat et al., 1993; Liang et al., 1994). Quality control of the results for total Hg and MeHg analysis in samples was performed by analysis of certified reference material (CRMs) obtained by the International Atomic Energy Agency (IAEA-356, Polluted Marine Sediment; Horvat et al., 1993) and by regular participation in intercomparison exercises for determination of total Hg and MeHg. The precision expressed as the relative standard deviation of at least three replicates varied from 3 to 5% for determination of total Hg and MeHg in sediments and up to 10% for water samples. The detection limit for measurement of total Hg in water is 0·05 ng litre
1 and for MeHg 0·01 ng litre
1 using at least 20 ml of water sample. Most of the samples were analysed in duplicates. Each batch of samples was accompanied by at least three blank samples and duplicates of appropriate CRMs. Diffusive fluxes ( J) across the water–sediment interface were determined using Fick’s first law by constructing a linear gradient of the porewater concentration of solute (C) in the 0–1 cm layer (z) and supernatant water: J=
ÖDS(äC/äz)z=0 where Ö is the porosity calculated in the 0–1 cm depth interval, DS is the whole-sediment diffusion coefficient and (äC/äz) is the concentration gradient. DS is assumed to be Ö2D0, following suggestions of Ullman and Aller (1982) for sediment with Öd0·7. D0 is the molecular diffusion coefficient for mercury in sea water considered to be equal to 510
6 cm2 s
1 (Bothner et al., 1980; Gobeil & Cossa, 1993; Mason

Mercury and methylmercury in the Gulf of Trieste 419 Porosity (φ) 0.65 0.75

0.55 0 –3 1 2.597 g cm 2 3

0

20

80 100

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2.674

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(%) 40 60

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2.727

2.669

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15

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2.674

(a)

sand silt clay

ln 210Pb (Bq kg–1) 3 3.5 4 4.5 5

13 14

18 (b)

F 2. Vertical profile of sediment density and porosity (a), textural variability (b) and AA1.

et al., 1993) and according to general values for heavy metal ions (Li & Gregory, 1974). Results and discussion Solid phase The sediment density at the sampling point AA1 ranges from 2·597 to 2·727 g cm
3 and the porosity decreases with depth from the surficial value of 0·776 to 0·597 at a depth of 18 cm [Figure 2(a)]. The first 18 cm are texturally rather uniform. Sediments consist of sandy pelites with the sandy fraction ranging between 10·8 and 15·1% whereas, within the pelitic component, the silty fraction (43·1–64·9%) is always predominant over the clay component (21·3–42·7%) [Figure 2(b)]. Mean size (3·2–6·9 ìm) is an index of very fine material. Small grain-size changes, particularly in silt/clay proportion, may be from normal fluctuations of terrigenous inputs from the Isonzo River. The total S content ranged between 0·240·03 and 0·140·02% with a maximum at a depth between 12 and 14 cm (Figure 3). The total C concentrations were in the range between 5·850·28 and 5·310·25% with lower values in deeper layers. The same trend was observed for organic C and total N

5.5

(c) 210

Pb activity (c) at sampling site

contents. The highest content of organic C occurred in the surficial layer (1·150·12%) whereas in the deepest it reached the minimum (0·740·08%). The total N concentrations decreased almost exponentially from the surficial maximum of 0·160·01% to minimum (0·100·01%) at a depth between 15 and 18 cm. The C/N ratio is rather constant (about 7 by weight; Figure 3) with depth indicating an approximately parallel degradation of organic C and N. The total Hg in the surficial sediment showed a progressive slight decreasing trend from the 6·5– 80 cm deep interval (3·640·52 ìg g
1) to the topmost layer of the sediment column (2·82 0·38 ìg g
1). At greater depths the mercury contents were variable with the highest concentration observed in the 12–14 cm deep layer (3·840·49 ìg g
1). Lacking direct measurements of MeHg in the solid phase, we speculate that MeHg in sediments at AA1 is about 1·8% of the total Hg according to data recently obtained from surficial sediments of the Gulf of Trieste where organo-mercury, ranging between 0·2 and 60 ng g
1, is positively correlated with the clay fraction (Horvat et al., 1996). It is worth noting that field surveys indicate the proportion of MeHg in sediments ranging normally between 0·1 and 2·1% of the total mercury (Kudo et al., 1977).

420 S. Covelli et al. Hg (µg g–1)

Depth (cm)

1.0 1.5 2.0 2.5 3.0 3.5 4.0 4.5 5.0

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18

S (%) 0.05

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0.20

C total (%) 0.25

0.30

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18

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N total (%) 0.8

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Depth (cm)

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0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18

0.14

C/N (by weight) 0.16

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0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18

F 3. Vertical distribution of mean values (SD, N=6) for total Hg, total S and C, organic C, N contents and C/N ratio (by weight) in the sediments at the sampling site AA1.

It seems reasonable that the lower mercury content in recent sediments follows the reduced mercury riverine input due to the gradual cessation of mining activity in the Idrjia district. In parallel with Hg, the total S concentration increases with depth, due to the increased sulphate-reduction activity (Hines et al., 1997) suggesting that mercury sulphide and other sulphur compounds could be the most important form of metal in the solid phase in this marine area. Total dissolved mercury in porewaters It should be noted that HgT profiles in porewaters was not proportional to the amount of Hg associated with the solid phase. This result suggests that Hg concentration in porewaters is not controlled by an exchange equilibrium between the two phases. During the first experiment performed in September 1995 [Figure 4(a)], the HgT content detected in porewaters ranged from 3·40 to 14·39 ng litre
1, without any appreci-

able variability throughout the 18 cm of the sediment core. The bottom water overlying sediment (supernatant) showed high concentration of total mercury (11·23 ng litre
1) similar to that found in porewaters below the water–sediment interface. The concentration profile of dissolved HgT in porewaters collected in November 1995 exhibits the greatest variability and the most peculiar trend. From a concentration of 13·58 ng litre
1 in the first centimetre, the dissolved HgT increased sharply to a subsurface maximum peak of about 107·47 ng litre
1 at a depth of 3·5–5 cm, before decreasing to values of less than 10 ng litre
1 below a depth of 6·5 cm. The maximum value of dissolved HgT was about 14 times higher than the Hg content in the supernatant water overlying the sediment and 10 times higher than the concentration present in the deepest layers. The increased concentrations of dissolved HgT just below the water–sediment interface could be related to the solubility changes of hydrated Fe and Mn oxides/

Mercury and methylmercury in the Gulf of Trieste 421

Depth (cm)

0

Hg tot (ng litre–1) 20 40 60 80 100 120

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18

(a)

Sep. '95 Nov. '95 Feb. '96 May '96 Jul. '96 Oct. '96

0

MeHg (ng litre–1) 1 2 3 4 5 6 7 8 9 1011 12

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18

(b)

Sep. '95 Nov. '95 Feb. '96 May '96 Jul. '96 Oct. '96

0

Hg reactive (ng litre–1) 20 40 60 80 100 120

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18

(c)

Sep. '95 Nov. '95

F 4. Concentration profiles of total dissolved Hg (a), MeHg (b) and reactive Hg (c) in porewaters at sampling site AA1 in the period September 1995–October 1996.

(a) MnHCl (ppm) solid phase 370 380 390 400 410 420 430 440 450

(c) Carbon oxidation (µM day–1)

(b) Dissolved Fe and Mn (µM) 0

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FeHCl (%) solid phase

F 5. Sedimentary depth profiles of solid (a) and dissolved (b) iron and manganese at site CZ. Sulphate and iron reduction rates (c) are presented as carbon oxidation equivalents (Hines, unpubl. data).

hydroxides as a consequence of the low-oxygen and low-sulphide conditions. It is well known that these compounds, insoluble in oxidizing conditions, are efficient scavengers for many soluble heavy metals such as mercury (Lockwood & Chen, 1973). In suboxic conditions these oxides are used in microbial degradation of organic matter and are solubilized (Froelich et al., 1979). This behaviour was observed in sediments of Piran Bay (Hines et al., 1997) as well as in mid-gulf at the sampling station CZ [Figure 5(b); Hines, unpubl. data] located a short distance from the experimental site AA1, in an area of sea-bottom geochemical homogeneity. The consequences of this

would be an increase of mercury concentration in porewaters in the oxygen-depleted zone. On the other hand, exchange reactions with solid phase, diffusion and complexation with organic matter and products of local sulphide oxidation resulting from bioturbation (Gagnon et al., 1996) could also be operative. The precipitation of insoluble mercury sulphide and the coprecipitation of mercury with iron sulphides may provide mechanisms for the decrease in dissolved mercury below the depth of 6–7 cm (Gobeil & Cossa, 1993). Measurements of reactive Hg performed in the first two experiments [Figure 4(c)] indicated that part of

422 S. Covelli et al.

the mercury in porewaters was easily reducible. This determination is a suitable measure of the Hg substrate available for methylation, elemental Hg formation, and other conversion processes in natural waters (Mason et al., 1993). The HgR fraction in porewaters ranged approximately between 10 and 53% of the HgT in September 1995 and from about 10 to 87% in November 1995, when the greatest part of this fraction was detected along with the maximum peak of HgT at a depth of 3·5–5·0 cm. In winter (February 1996), the HgT values ranged between 10·90 and 40·80 ng litre
1 with the maximum observed at a depth of 6·5–8·0 cm. Considering this concentration vertical profile it is possible that part of the dissolved mercury in this period was still related to the highest values which appeared in autumn. In the water overlying the sediment, mercury concentration (8·32 ng litre
1) was only about one-third lower than in porewaters of the first layer. Dissolved mercury concentrations appeared to be lower in late spring (May 1996), with values ranging from 7·63 to 24·60 ng litre
1, and only slightly higher than those observed at the end of the summer 1995. Maximum concentration appeared at the top of the sediment core followed by a vertical decreasing gradient. The HgT concentration in the supernatant was about six times lower (4·12 ng litre
1) than in porewaters below the water–sediment interface. The lowest values of HgT in porewaters, ranging from 2·0 to 8·7 ng litre
1, and in the supernatant (0·9 ng litre
1) were observed in the summer season (July 1996). The last campaign in October 1996 showed HgT concentrations similar to those reported for September 1995. It seems that environmental conditions in the warmer period, when sulphatereducing activity is high (Hines et al., 1997), negatively influences the mobility of Hg in the dissolved phase. Whereas it was demonstrated that sulphatereducing bacteria are responsible for much of the methylation in aquatic environments (Compeau & Bartha, 1985; Berman et al., 1990; Choi et al., 1994), it was also recognized that hydrogen sulphide, which evolves from dissimilative sulphate reduction, can convert ionic Hg into metacinnabar (HgS) or involve MeHg transformation into dimethylmercurysulphide. This, in turn, can be decomposed into volatile dimethylmercury and metacinnabar (Baldi et al., 1995), thus reducing Hg content in porewaters. It is interesting to note that HgT concentrations in the bottom water layer (about 1 m above the seabottom) of the Gulf of Trieste were in general 4 to 25 times lower than those present in the supernatant waters during our experiments. Recent measurements performed on several sampling stations in this coastal

area showed HgT concentrations ranging between 0·18 and 4·90 ng litre
1, with the highest values found in the surface layer close to the freshwater inputs of the Isonzo River (Horvat et al., 1996; Horvat, 1997). These results are very close to those reported for other contaminated river mouths, such as the Rhone delta (Cossa & Martin, 1991), the Pettaquamscutt Estuary (Mason et al., 1993), the St. Lawrence Estuary (Cossa et al., 1988) and the Gironde estuary (Figueres et al., 1985), although in these cases mining activity was not claimed to be the main Hg source. The total mercury concentrations in porewaters in our study are one order of magnitude higher than the values reported by Gobeil and Cossa (1993) for the Laurentian Trough (0·5–12·8 ng litre
1). Sediments in that estuarine area exhibit surficial mercury contents about 30 times lower (0·180 ìg g
1) in comparison with our experimental site AA1, whereas porewaters were 14 times enriched in mercury relative to the overlying bottom water in which mercury concentration, for unfiltered samples, did not exceed 0·8 ng litre
1. Conversely, our porewater values are comparable to those reported for the highly contaminated Bellingham Bay (U.S.A.) oxidizing sediments which contain mercury between 1·0 and 5·4 ìg g
1 (Bothner et al., 1980). On the contrary, the dissolved mercury concentrations in the Gulf of Trieste are significantly lower than those obtained by Bothner et al., (1980) for Bellingham Bay (0·25– 3·80 ìg litre
1) reducing surficial sediments with high mercury contents (32–71 ìg g
1) and located very close (500 m) to the contamination source of a chloroalkali plant. However, it should be kept in mind that differences and similarities of mercury behaviour in aquatic environments should be carefully evaluated in relation to substrate material and to the prevalent chemical form of available mercury. Mercury species in industrial effluents, mainly from chloro-alkali plants, might have different mobility and adsorption/ desorption characteristics with respect to mercury sulphide supposed, in our case, to be the prevalent metal compound in the sediments of the Gulf of Trieste.

Methylmercury in porewaters Seasonal variations of organomercury concentrations in porewaters [Figure 4(b)] were generally well correlated with dissolved HgT. The lowest concentrations of MeHg were observed in September 1995 (0·27– 0·61 ng litre
1), and July (0·10–0·80 ng litre
1) and October 1996 (below detection limit–0·83 ng litre
1).

Benthic fluxes The highest benthic effluxes of HgT were observed in early autumn (November 1995) and early spring [February and March 1996, Figures 6(a) and 7], i.e. in periods of higher oxygen concentration at the sea-bottom layer and lower temperature (Figure 8). In these periods the benthic fluxes of HgT ranged between 5 and 6 ìg m
2 day
1. Similar to HgT, the MeHg benthic fluxes were also highest in November 1995, reaching about one-third of the HgT fluxes.

80 (a)

March 1996 Hg tot MeHg

60 50 40 30 20 10 0

10

20 30 Time (h)

40

1.20

8 7

12 11 10 9 8 7 6 5 4 3 2 1 0 50

(b)

July 1996 Hg tot MeHg

6 5 4

1.00 0.80 0.60

3

0.40

2 0.20

1 0

MeHg (ng litre–1)

Hg tot (ng litre–1)

70

Hg tot (ng litre–1)

The MeHg percentage of HgT for these dates was, on average, always less than 11%. The MeHg concentration profile in November 1995 showed a similar increasing trend with depth as that mentioned for dissolved HgT. The increase of MeHg content was clearly localized at a depth between 1 and 5 cm, where the highest concentration was detected (10·43 ng litre
1). The percentage of MeHg relative to HgT ranged between about 4 and 58%. The maximum value did not appear in the depth interval of the highest MeHg content but was deeper (6·5–8·0 cm). The organo-mercury concentrations analysed in porewaters in the winter (February 1996) appeared lower (2·29–4·39 ng litre
1) than those observed in the 1–5 cm depth interval in early autumn. Although, similar to what occurred for dissolved HgT, the characteristic vertical profile was still recognizable. The percentages of MeHg varied from one-tenth to one-third of the total mercury content. The results from the late spring samples (May 1996) revealed a further general decrease of MeHg content (0·85–3·50 ng litre
1). The MeHg percentages in the HgT were instead comparable to that observed in February 1996. The highest percentages of MeHg in porewaters were always observed during low-temperature periods at the sea-bottom. The MeHg concentrations in unfiltered samples in the water column, collected in several stations of the Gulf during September 1995, were up to 12 times higher in the bottom layer compared with the seasurface (Horvat, 1997). The MeHg concentrations in the bottom water layer (0·005–0·062 ng litre
1) were almost five times lower than those found in pore and supernatant waters. This would indicate that the recent sediment could be a source of dissolved and particulate MeHg for the water column of the Gulf of Trieste. However, either degradation processes involving the organomercury form in the upper water layers, or methylating processes active in the low-oxygen region of the water column cannot be excluded (Mason et al., 1993).

MeHg (ng litre–1)

Mercury and methylmercury in the Gulf of Trieste 423

5

10 15 Time (h)

20

0.00 25

F 6. Examples of the evolution of total dissolved Hg and MeHg in the overlying water during benthic flux experiments in March 1996 (a) and July 1996 (b).

However, we have to consider that the experiment performed in February 1996 was unfortunately interrupted after 6 h, so the measured fluxes were probably underestimated. The released Hg species in these periods can affect the whole water column because of the vertical homogenization normally present from November to April. On the other hand, summertime was characterized by low HgT and MeHg fluxes [Figures 6(b) and 7]. The O2 content in the benthic chamber was checked during the experiments and results showed that oxygen never dropped down to hypoxic conditions and only 30% of the initial concentration was lost in 24 h (Bertuzzi et al., 1997). Bothner et al. (1980) measured an average efflux of 1·210
5 ng cm
2 s
1 (about 10 ìg m
2 day
1) from highly contaminated anoxic sediments to oxygenated overlying waters in Bellingham Bay close to the chloro-alkali plant effluents. Fluxes to overlying water, that was allowed to become anoxic, were more than twice as high at the same location. The authors suggested that maintaining oxygenated conditions could be a way to reduce the mercury flux to the water column. However, this would not be the case for the Gulf of Trieste.

424 S. Covelli et al. 7000

6456

HgT, MeHg (ng m–2 day–1)

6000

5297

5000

Hg tot MeHg

4000 3000

2370

2000 1000 0

645 53

8

63

27

431 314

760 90 –2

3

–105

–1000

Sep. 95

Oct. 95

Nov. 95

Feb. 96

Mar. 96

May 96

Jul. 96

Oct. 96

F 7. Annual variations of total dissolved Hg and MeHg benthic fluxes at sampling site AA1 in the period September 1995–October 1996.

25

120 110

23

oxygen temperature

21

100

17 15

80

13

70

Temperature (°C)

O2 saturation (%)

19 90

11 60 9 50 40

7 Aug. Sep. Oct. Nov. Dec. Jan. Feb. Mar. Apr. May Jun. Jul. Aug. Sep. Oct. 1995–96

5

F 8. Annual variations of O2 saturation percentage and temperature in the bottom water layer at sampling site AA1 during the period September 1995–October 1996.

Our results indicate that autumn (November), following the oxygen depletion in late summer in the bottom layer, is the period of the highest benthic fluxes and porewater concentrations of Hg and MeHg. It seems that the transition from rapid sulphate reduction to cooler temperatures (Figure 8) and lower microbial activity is well suited for Hg transformations, accumulation and flux. It might be expected that sulphate reduction, when most active in late summer (Hines et al., 1997), would remove Hg available for methylation and other transformations as sulphide. During winter, when temperature and microbial activity are lower, the increased penetration

of oxygen allows for a predominance of O2 and metal reduction in surficial sediments of the Gulf of Trieste. In the warm periods, when microbial and bioturbation activity are high and bottom-water O2 content is high, the activities of infauna maintain subsurface cycling of redox-sensitive elements which enhances the importance of metal reductions in sedimentary biogeochemical processes. During warm periods, when bottom water becomes depleted in O2, the decreased subsurface cycling of metals and sulphur allows for sulphate reduction to dominate the decomposition of sedimentary organic matter in the Gulf (Hines et al., 1997). Thus, it was

Mercury and methylmercury in the Gulf of Trieste 425 T 1. Estimated diffusive fluxes of total dissolved Hg and MeHg at sampling site AA1 in the period September 1995–October 1996

Sampling September 1995 November 1996 February 1996 March 1996 May 1996 July 1996 October 1996

HgT (ng m
2 day
1)

MeHg (ng m
2 day
1)


2·12 11·05 6·41 4·05 41·28 9·89 7·29

0·42
0·60 4·39 5·30
1·77 1·28 0·01

estimated that in the mid-gulf, near CZ station, the percentage of organic carbon oxidation [Figure 5(c)] attributed to sulphate reduction in comparison with iron reduction for the upper 3 cm of sediment was 100% in September 1993 when O2 dropped to about 40% of saturation (Hines, unpubl. data). The subsequent decrease of sulphide production plus a slight oxidation of the sediments (higher oxygen content in the bottom-water layer, Figure 8) might make favourable conditions for Hg cycling in this part of the year. In addition, concerning the MeHg production rate, measured benthic fluxes should be considered as net MeHg production rates due to several competing processes which determine the ratio between methylation and demethylation. The diffusive effluxes of HgT (Table 1) were always observed, except in late summer (September 1995), when a small influx (
2·11 ng m
2 day
1) appeared consequent to a slightly negative concentration gradient across the water–sediment interface. The highest diffusive fluxes of 41·28 and 11·05 ng m
2 day
1 occurred in late spring (May 1996) and autumn (November 1995), respectively. Lower effluxes were calculated during winter (February and March 1996) and summer (July 1996) seasons. Very low MeHg diffusive fluxes were observed between September and November 1995 whereas the highest effluxes, comparable to values observed for HgT in the same season, were reported in February and March 1996 (4·39 and 5·30 ng m
2 day
1). It is worth noting that the primary production of benthic microalgae, phytoplankton and resuspended benthic microalgae in the bottom-water layer along with benthic oxygen consumption in the same period (Bertuzzi et al., 1997) are not linked to the observed mercury fluxes. Comparison of HgT benthic with diffusive fluxes showed surprisingly that the latter were more than 100 time lower in autumn and winter periods when benthic fluxes were the highest. A possible explanation

could be that the first sediment sampling interval (0–1 cm) chosen to apply Fick’s first law is too large to be representative, thus underestimating the real gradient between the sediment and water column. However, it should be noted that processes other than simple concentration gradients must account for Hg release from sediments. Bioturbation activity of benthic infauna increases fluxes of organic and inorganic complexes at the water–sediment interface up to 2–10 times more than those calculated on the basis of molecular diffusion (Rutgers van der Loeff et al., 1984). Previous analyses performed in the sediments of the Gulf of Trieste showed that bioturbating activity is intensive especially in warmer periods of the year (Cermelj et al., 1997). The same phenomenon occurs also in the mid-gulf until occasional depletion of O2 in bottom waters appeared to cause a decrease in irrigation/reworking activity by infauna due to the toxic effect of sulphides produced by the sulphate reduction (Hines et al., 1997). Our diffusive fluxes are higher when environmental conditions are more favourable for the benthic infauna (low sulphate reduction and high oxygen), i.e. May and November, but even this feature is still not sufficient to explain high benthic fluxes which seem to be more a consequence of described biogeochemical processes occurring in surficial sediments involving sulphate reduction and changing oxidation conditions.

Tentative budget of Hg at the water–sediment interface Considering the porewater vertical profiles, it seems that the total mercury variability in sediment at site AA1 is not highly affected by early diagenetic processes involving mercury redistribution between solid and liquid phases. Concentrations of dissolved HgT in porewaters are about three orders of magnitude lower than the mercury in sediment solid phase. However, inorganic and organic mercury compounds may be involved in chemical transformations within the sediment column following redox condition changes and they might lead to partial release into the overlying water column. The fate of Hg at the water–sediment interface can be determined from the Hg budget. A tentative annual benthic Hg budget for the sampling point AA1 can be constructed from the bimonthly measured fluxes, and assuming steady-state conditions. The HgT deposition rate (S) in the surficial sediment layer can be estimated from the average recent accumulation rate (R) of 1·8 mm yr
1 based on 210Pb activity [Figure 2(c) M. Frignani, pers. comm.] using the relation S=ñScR(1
Ö) (Suess & Djafari, 1977) where ñS is the dry density of surficial

426 S. Covelli et al.

MeHg HgT Benthic fluxes 0.75

0.17

Net accumulation rate

fluxes measured using benthic chambers could be overestimated. Current ongoing research on Hg methylation and demethylation processes in sediments of this area will clarify this discrepancy and, in addition, will help understand the real contribution of these processes to the Hg budget in the Gulf.

2.96

WATER

Conclusions SEDIMENT 2.21

0.04

Burial

F 9. Tentative annual budget of total Hg and MeHg (mg m
2 yr
1) based on benthic flux measurements at the water–sediment interface in the central area of the Gulf of Trieste (sampling site AA1).

sediment, c the HgT concentration and Ö the porosity [Figure 2(a)]. The resulting tentative budget, based on benthic flux measurements (Figure 9) shows that 25% of HgT settling down to the sea-bottom is annually recycled at the water–sediment surface and 75% is buried in the sediment. Comparing our data with MeHg accumulation rate of Coquery et al. (1996) in the area of the Northern Adriatic Sea (0·75 mg m
2 yr
1) affected by the river Po plume, it appears that the central sector of the Gulf shows a fourfold higher annual metal deposition (2·96 mg m
2 yr
1). This result is more relevant if it is considered that the sedimentation rate in the outer Po delta zone is about fourfold higher than that at the AA1 sampling station. Approximately 23% of the HgT released as benthic fluxes, corresponding to 6% of deposited HgT, occurs in methylated form. Burial MeHg is estimated to comprise 1·8% of the total Hg present in the solid phase. MeHg effluxes account for more than fourfold of MeHg bound to sediments. In these calculations it was supposed that the whole amount of MeHg associated to sediments, porewaters and related fluxes is produced by bacterial methylation activity in situ. Assuming that part of MeHg may have a fluvial origin and settles down at the sea-bottom associated into inorganic and organic particles and considering that particulate MeHg can constitute, on average, up to 3% of particulate HgT (Horvat, 1997), it can be estimated that at least about 50% of the released MeHg could be produced by in situ methylation. On the other hand, if diffusive fluxes are taken into consideration, the total Hg annually released from sediment would account only for less than 1% of that estimated from benthic flux measurements. Thus, the

(1) The concentrations of total dissolved mercury in porewaters at the sampling site AA1 in the Gulf of Trieste are in general three times lower than in the solid phase in the first 18 cm. The total easily reducible Hg and MeHg represent, on average, 30 and 17%, respectively, of total dissolved Hg in porewaters. (2) Higher total dissolved Hg and MeHg in porewaters observed in autumn and winter are a consequence of lower oxygen and low-sulphide conditions. These would increase metal concentration in the dissolved phase just a few centimetres below the water–sediment interface in the oxygen-depleted zone. (3) Hg cycling in sediments of the Gulf of Trieste is limited to the uppermost layers. Benthic fluxes were the highest in autumn and winter and were is probably related to the transition from rapid sulphate reduction to cooler temperatures and lower microbial activity, suitable conditions for Hg transformation, accumulation and flux. (4) Hg annual budget, based on benthic flux measurements at the water–sediment interface, in the mid-Gulf indicates that 25% of HgT is recycled annually from the water–sediment surface and 75% is buried in the sediment. Approximately 23% of the released HgT is in the methylated form, whereas the burial of MeHg represents 1·8% of total Hg in the solid phase. (5) Sediments in the central part of the Gulf of Trieste may be considered to be an important secondary source of mobile forms of mercury in this marine environment. Implications for the trophic chains and the behaviour of this toxic element in different locations of the Gulf should be further investigated.

Acknowledgements The authors thank the staff of Laboratorio di Biologia Marina of Trieste for skilful help in the sampling operation and for technical assistance.

Mercury and methylmercury in the Gulf of Trieste 427

References Baldi, F. & Bargagli, R. 1984 Mercury pollution in marine sediments near a chlor-alkaly plant: distribution and availability of the metal. Science of the Total Environment 39, 15–26. Baldi, F., Parati, F. & Filippelli, M. 1995 Dimethylmercury and dimethylmercury-sulfide of microbial origin in the biogeochemical cycle of Hg. Water, Air, and Soil Pollution 80, 805– 815. Barkay, T. 1992 Encyclopedia of Microbiology Vol. 3. Academic Press, San Diego, pp. 65–74. Benoit, G., Schwantes, J. M., Jacinto, G. L. & Goud-Collins, M. R. 1994 Preliminary study of the redistribution and transformation of HgS from cinnabar mine tailings deposited in Honda Bay, Palawan, Phillippines. Marine Pollution Bulletin 28, 754–759. Berman, M., Chase, T. & Bartha, R. 1990 Carbon flow in mercury biomethylation by Desulfovibrio desulfuricans. Applied Environmental Microbiology 56, 298–300. Bertuzzi, A., Faganeli, J., Welker, C. & Brambati, A. 1997 Benthic fluxes of dissolved inorganic carbon, nutrients and oxygen in the Gulf of Trieste (Northern Adriatic). Water, Air and Soil Pollution 99, 305–314. Bloom, N. S. & Crecelius, E. A. 1983 Determination of mercury in seawater at sub-nanogram per liter levels. Marine Chemistry 14, 49–59. Bloom, N. S. & Watras, C. J. 1989 Observations of methylmercury in precipitation. The Science of the Total Environment 87/88, 199–207. Bonzongo, J. C., Heim, K. J., Chen, Y., Lyons, W. B., Warwick, J., Miller, G. C. & Lecher, P. J. 1996 Mercury pathways in the Carson River–Lahontan Reservoir system, Nevada, USA. Environmental Toxicology and Chemistry 15, 677–683. Bothner, M. H., Jahnke, R. A., Peterson, M. L. & Carpenter, R. 1980 Rate of mercury loss from contaminated estuarine sediments. Geochimica Cosmochimica Acta 44, 273–285. Brambati, A., Ciabatti, M., Fanzutti, G. P., Marabini, F. & Marocco, R. 1983 A new sedimentological textural map of the northern and central Adriatic Sea. Bollettino di Oceanologia Teorica ed Applicata 1, 267–271. Bussani, M. & Princi, M. 1979 Mercurio nel Golfo di Trieste. Bollettino dei Chimici dell’Unione Italiana dei Laboratori Provinciali 5, 224–237. Cermelj, B., Bertuzzi, A. & Faganeli, J. 1997 Modelling of pore water nutrient distribution and benthic fluxes in shallow coastal waters (Gulf of Trieste, northern Adriatic). Water, Air and Soil Pollution 99, 435–444. Choi, S. C., Chase, T. & Bartha, R. 1994 Metabolic pathways leading to mercury methylation in desulfovibrio desulfuricans LS. Applied Environmental Microbiology 60, 4072–4077. Compeau, G. & Bartha, R. 1984 Methylation and demethylation of mercury under controlled red-ox, pH, and salinity conditions. Applied Environmental Microbiology 48, 1203–1207. Compeau, G. & Bartha, R. 1985 Sulphate-reducing bacteria: principal methylators of mercury in anoxic estuarine sediment. Applied Environmental Microbiology 50, 498–502. Coquery, M., Cossa, D., Gobeil, C., Azemard, S., Sanjuan, J., a1Magand, O. & Horvat M. 1996 The significance of mercury and methylmercury profiles in coastal sediments. In Book of Abstracts, IV International Conference ‘Mercury as a Global Pollutant’, 4–8 August 1996, Hamburg, Germany (Ebinghaus, R., Petersen, G. & von Tu¨mpling, U. eds). GKKS Research Centre, Geesthacht, Germany, p. 501. Cossa, D. & Martin, J.-M. 1991 Mercury in the Rhone delta and adjacent marine areas. Marine Chemistry 36, 291–302. Cossa, D., Gobeil, C. & Courau, P. 1988 Dissolved mercury behaviour in the Saint Lawrence estuary. Estuarine, Coastal and Shelf Science 26, 227–230. Covelli, S. & Fontolan, G. 1997 Application of a normalization procedure in determining regional geochemical baselines, Gulf of Trieste, Italy. Environmental Geology 30, 34–45.

Cranston, R. E. 1976 Accumulation and distribution of total mercury in estuarine sediments. Estuarine, Coastal and Shelf Science 4, 695–700. Donazzolo, R., Hieke Merlin, O., Menegazzo Vitturi, L., Orio, A. A., Pavoni, B. & Rabitti, S. 1983 Toxic elements in bottom sediments of Northern Adriatic Sea (Italian area). Thalassia Yugoslavica 19, 111–119. Faganeli, J., Pezdic, J., Ogorelec, B., Herndl, G. J. & Dolenec, T. 1991a The role of sedimentary biogeochemistry in the formation of hypoxia in shallow coastal waters (Gulf of Trieste, Northern Adriatic). In Modern and Ancient Continental Shelf Anoxia. Tyson, R. V. & Pearson, T. H., eds), The Geological Society, London, pp. 107–117. Faganeli, J., Planinc, R., Pezdic, J., Smodis, B., Stegnar, P. & Ogorelec, B. 1991b Marine geology of the Gulf of Trieste (northern Adriatic): geochemical aspects. Marine Geology 99, 93–108. Figueres, G., Martin, J. M., Meybeck, M. & Seyler, P. 1985 A comparative study of mercury contamination in the Tagus estuary (Portugal) and major French estuaries (Gironde, Loire, Rhoˆne). Estuarine, Coastal and Shelf Science 20, 183–203. Fitzgerald, W. F. 1989 Atmospheric and oceanic cycling of mercury. In Chemical Oceanography Vol. 10 (Riley, J. P. & Chester, R., eds). Academic Press, London, pp. 151–186. Fitzgerald, W. F. & Clarkson, T. W. 1991 Mercury and monomethylmercury—present and future concerns. Environmental Health Perspectives 96, 159–166. Froelich, P. N., Klinkhammer, G. P., Bender, N. A., Luedtke, G. R., Heath, G. R., Cullen, C., Dauphin, P., Hammond, D., Hartman, B. & Maynard, V. 1979 Early oxidation of organic matter in pelagic sediments of the eastern equatorial Atlantic: suboxic diagenesis. Geochimica Cosmochimica Acta 43, 1075– 1090. Gagnon, C., Pelletier, E., Mucci, A. & Fitzgerald, W. F. 1996 Diagenetic behaviour of methylmercury in organic-rich coastal sediments. Limnology and Oceanography 41, 428–434. Gill, G. A. & Fitzgerald, W. F. 1987 Picomolar mercury measurement in seawater and other materials using stannous chloride reduction and two-stage gold amalgamation with gas phase detection. Marine Chemistry 20, 227–243. Gobeil, C. & Cossa, D. 1993 Mercury in sediments and sediment pore water in the Laurentian Trough. Canadian Journal of Fishery and Aquatic Science 50, 1794–1800. Gosar, M., Pirc, S., S y ajn, R., Mashyanov, N. M. & Sholupov, S. E. 1996 Mercury in the air over Idrija—report of measurements on 24th September 1994. In Idrija as a Natural and Anthropogenic Laboratory (Miklavcic, V., ed.) Mercury Mine Idrija, Slovenia, pp. 26–29. Gosar, M., Pirc, S. & Bidovec, M. 1997 Mercury in the Idrijca river sediments as a reflection of mining and smelting activities of the Idrija mercury mine. Journal of Geochemical Exploration 58, 113–125. Hedges, J. I. & Stern, J. H. 1984 Carbon and nitrogen determinations in carbonate-containing solids. Limnology and Oceanography 29, 657–663. Hines, M. E., Faganeli, J. & Planinc, R. 1997 Sedimentary anaerobic microbial biogeochemistry in the Gulf of Trieste, Northern Adriatic Sea: Influences of bottom water oxygen depletion. Biogeochemistry 39, 65–86. Horvat, M. 1997 Mercury behaviour in estuarine and coastal environment. In Water Pollution IV, Modelling, Measuring and Prediction (Rajar, R. & Brebbia, C. A., eds). Computational Mechanics Publications, Southampton, pp. 161–173. Horvat, M., Bloom, N. S. & Liang, L. 1993 Comparison of distillation with other current isolation methods for the determination of methyl mercury compounds in low level environmental samples. Part II. Water. Analytica Chimica Acta 282, 153–168. Horvat, M., Mandic, V., Liang, L., Bloom, N. S., Padberg, S., Lee, Y. H., Hintelmann, H. & Benoit, J. 1994 Certification of methyl-

428 S. Covelli et al. mercury compounds concentration in marine sediment reference material, IAEA-356. Applied Organometallic Chemistry 8, 533– 540. Horvat, M., Faganeli, J., Planinc, R., Prosenc, N., Azemard, S., Coquery, M., Sirca, A., Rajar, R., Byrne, A. R. & Covelli, S. 1996 Mercury Pollution in Trieste Bay, IV International Conference ‘Mercury as a Global Pollutant’, 4–8 August 1996, Hamburg, Germany (Ebinghaus, R., Peterson, G. & von Tu¨mpling, U. eds) GKSS Research Centre, Geesthacht, Germany, p. 491. Kosta, L., Ravnik, V., Byrne, A. R., Stirn, J., Dermelj, M. & Stegnar, P. 1978 Some trace elements in the waters, marine organisms and sediments of the adriatic by neutron activation analysis. Journal of Radioanalytical Chemistry 44, 317–332. Krom, M. D., Kaufman, A. & Hornung, H. 1994 Industrial mercury in combination with natural Pb210 as time-dependent tracers of sedimentation and mercury removal from Haifa Bay, Israel. Estuarine, Coastal and Shelf Science 38, 625–642. Kudo, A., Akagi, H., Mortimer, D. C. & Miller, D. R. 1977 Equilibrium concentrations of methylmercury in Ottawa river sediments. Nature 270, 419–420. Lacerda, L. D., Carvalho, C. E. V., Rezende, C. E. & Pfeiffer, W. C. 1993 Mercury in sediments from the Paraiba do Sul river continental shelf, S. E. Brazil. Marine Pollution Bulletin 26, 220–222. Li, Y.-H. & Gregory, S. 1974 Diffusion if ions in sea water and in deep-sea sediments. Geochimica Cosmochimica Acta 38, 703–714. Liang, L., Horvat, M. & Bloom, N. S. 1994 An improved speciation method for mercury by GC/CVAFS after aqueous phase ethylation and room temperature precollection. Talanta 41, 371– 379.

Lockwood, R. A. & Chen, K. J. 1973 Adsorption of Hg(II) by hydrous manganese oxides. Environmental Science and Technology 8, 459–462. Mason, R. P., Fitzgerald, W. F., Hurley, J., Hanson, A. K., Donaghay, P. L. & Sieburth, J. M. 1993 Mercury biogeochemical cycling in a stratified estuary. Limnology and Oceanography 38, 1227–1241. Ogorelec, B., Misˇic, M. & Faganeli, J. 1991 Marine geology of the Gulf of Trieste (northern Adriatic): sedimentological aspects. Marine Geology 99, 79–92. Olivotti, R., Faganeli, J. & Malej, A. 1986 Impact of ‘ organic ’ pollutants on coastal waters, Gulf of Trieste. Water Science and Technology 18, 57–68. Rutgers van der Loeff, M. M., Anderson, L. G., Hall, P. O. J., Iverfeldt, A., Josefson, A. B., Sundby, B. & Westerlund, S. F. G. 1984 The asphyxiation technique: An approach to distinguish between molecular diffusion and biologically mediated transport at sediment–water interface. Limnology and Oceanography 29, 675–686. Santschi, P., Hohener, P., Benoit, G. & Bucholtz-ten Brink, M. 1990 Chemical processes at the sediment–water interface. Marine Chemistry 30, 315–369. Suess, E. & Djafari, D. 1977 Trace metal distribution in Baltic Sea ferromanganese concretions: inferences on accretion rates. Earth Planetary Science Letters 35, 49–54. Ullman, W. J. & Aller, R. C. 1982 Diffusion coefficient in nearshore marine sediments. Limnology and Oceanography 27, 552–556. Ure, A. M. & Shand, C. A. 1974 The determination of mercury in soils and related materials by cold vapour atomic absorption spectrometry. Analytica Chimica Acta 72, 63–77.