Mercury Diagenesis in the Saguenay Fjord
By
Geneviève Bernier
August, 2005
A thesis submitted to the Office of Graduate and Postdoctoral Studies in partial fulfillment of the requirements for the Degree of Master of Science
Earth and Planetary Sciences McGill University Montréal, Québec, Canada
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While these forms may be included Bien que ces formulaires in the document page cou nt, aient inclus dans la pagination, their removal does not represent il n'y aura aucun contenu manquant. any loss of content from the thesis. ••• Canada ACKNOWLEDGEMENTS
1 am grateful to the following people, whose help and continuous support are the foundation on which this thesis was built. First and foremost, to Dr. A. Mucci, 1 wish to express my most sincere appreciation and thanks for his guidance and endless patience during the tortuous path of this project, his commitment to his students, as well as for his cheerful stewardship of the field work conducted in the Saguenay Fjord and St. Lawrence. This was a considerable leaming experience for me on many levels, and 1 wish to thank him for giving me the opportunity to continue on with this project. Thanks also must be expressed to C. Guignard, for her support throughout, her considerable problem solving abilities, for maintaining the standards of rigorous laboratory practice, and for many excellent recipes.
1 also wish to thank G. Keating for her technical assistance and S. Musc10w for her cheerful presence and assistance during pyrite analyses, as well as to P. Collin, L. Barazzuol, C. Magen, P. Benoit, G. Chaillou, A. Villegas, S.T. Kim, C. Mann, and numerous others for the camaraderie, assistance, and exchange of ideas that made this project such a pleasant and stimulating experience.
This study would not have been possible without fmancial backing from the Fonds pour la Formation des Chercheurs et l'Aide a la Recherche du Québec (FCAR). The Department of Earth and Planetary Sciences at McGill also provided financial support in the form of teaching assistantships and, more importantly, continuous moral and administrative support in the forms of Anne, Carol, and Kristy. Thanks also go to the captain and crew of the Alcide C. Horth, for their tireless work and for making field research such a pleasant experience.
My gratitude also goes to J. Hartzler who, during the last stages of this undertaking, provided support and understanding that went beyond the bounds of a roommate's duty; and to S. Rahman for the sanity control, stimulating discussions, and late-night lab fun and laughter during our laboratory years together.
Finally, my greatest thanks go to my family. To my brothers, for the cheerfulness and considerable support. To my mom, for her common sense, for always having an eye on the big picture, and always keeping my priorities straight. And to my dad, for the enthusiasm, the endless pep talks and moral support, and for being there at the very best oftimes.
ii CONTRIBUTIONS OF AUTHORS
This is a manuscript-based thesis divided in three chapters. The first chapter is a general introduction which provides an overview of the present knowledge on the chemistry of mercury and more specifically on the topics presented in this work. The third section includes the author's (Geneviève Bemier's) general conclusions and recommendations for future work. The second chapter was written in manuscript format and will be submitted to the scientific journal Applied Geochemistry. The author and Constance Guignard were responsible for aIl the experimental and analytical work. The interpretation of the results is solely the responsibility of the author who did benefit from critical comments and suggestions from Prof. Mucci. Prof. Alfonso Mucci, supervisor of this M.Sc. project, contributed heavily towards the edition of the manuscript.
Funding for this project was provided by Natural Sciences and Engineering Research Council of Canada (NSERC) strategic and research grants to Prof. Mucci. Financial support to the author was provided by the Fonds pour la Formation des Chercheurs et l'Aide a la Recherche du Québec (FCAR) in the form of a post-graduate scholarship, and by the Department of Earth and Planetary Sciences at McGiH University in the form of teaching assistantships.
iii TABLE OF CONTENTS
ACKNOWLEDGEMENTS II
CONTRIBUTIONS OF AUTHORS III
TABLE OF CONTENTS IV
LIST OF FIGURES VII
ABSTRACT IX
RÉSUMÉ X
CHAPT ER 1 1 INTRODUCTION 1 Sources of Mercury 2 Speciation of Mercury in the Surficial Cycle 4 History and Source of Mercury Contamination to the Sediments of the Saguenay Fjord 6 Diagenetic Behaviour of Mercury 7 Mercury mobility in marine sediments 7 Factors controlling mercury methylation and demethylation 9 Objectives 11
CUAPTER2 14 MERCURY GEOCHEMISTRY IN THE SEDIMENTS OF THE SAGUENAY FJORD: Diagenetic behaviour following a catastrophic depositional event 14 Abstract 15 Introduction 16 Fjord Morphology and Hydrographie Characteristics 18 Materials and Methods 20 Sampling history 20
iv ~.~ Sampling method 21 Analytical methods 22 Results and Discussion 23 Pre-flood conditions 23 Flood mate rial 28 Post-flood geochemistry 29 Summary and Conclusions 48
!:HAl!IERJ ~l Summary and conclusions 51 Recommendations for future work 55
REFERENCES 58
APPENDIX1 72
-----'. APPENDIX2 74
v TABLE OF FIGURES
Figure 1: Map of the Saguenay Fjord showing the sampling sites. The number in brackets is the approximate thickness of the flood layer in centimeters. From Mucci et al. (2000) ...... l9.
Figure 2: Geochemical characteristics and steady-state conditions of sediments recovered at SAG-30. a) Vertical distribution ofTHgs (ng goI) on five sampling dates over a 6-year period; b) Vertical distribution of THgd (ng LoI) at SAG-30 from a core recovered in 2001 (SAG-30/2001L ...... ~~
Figure 3: Geochemical characteristics of pre-flood sediments at SAG-09 from a core recovered in 1995 (SAG-09/1995): a) Corg (wt%), b) Cinorg (wt%) and c) THgs (ng goI) depth profiles...... ~9.
Figure 4: Geochemical characteristics of pre-flood sediments at SAG-05 from a core recovered in 1991 (SAG-05/1991): a) Corg (wt%), b) Cinorg (wt%), c) THgs (ng goI), d) AVS (~mol goI) and e) pyrite (FeS2; ~mol goI) depth profiles. The short-dashed lines mark the approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide deposit...... ~~
Figure 5: Geochemical characteristics of sediments at SAG-09 from a core recovered three weeks after the 1996 flood (SAG-09/1996post-flood): a) Corg (wt%), b) Cinorg (wt%) and c) THgs (ng goI). The stippled line marks the approximate location of the contact between the flood material and indigenous sediment ...... }l
Figure 6: THgs depth profiles at SAG-09 from cores recovered in a) 1997 (SAG- 09/1997), b) 1998 (SAG-09/1998), c) 1999 (SAG-09/1999(1)), d) 1999 (SAG-09/1999(2), e) 2000 (SAG-09/2000), f) 2001 (SAG-09/2001) and g) 2002 (SAG-09/2002). The stippled lines mark the approximate location of the upper and lower boundaries of the flood deposit ...... }~
Figure 7: Geochemical characteristics of sediments at SAG-09 from a core recovered in 2001 \SAG-09/2001): a) AVS (~mol gol), b) THgs (ng gol) and c) THgd (ng LO ). The stippled lines mark the approximate location of the upper and lower boundaries of the flood deposit...... }}
Figure 8: Geochemical characteristics of sediments at SAG-09 from a core recovered in 2002 (SAG-09/2002): a) AVS (~mol gol), b) THg (ng gol) and c) THg (ng LoI). The stippled lines mark the approximate location of the upper and lower boundaries of the flood deposit...... ~.3
vi Figure 9: Geochemical characteristics of sediments at SAG-05 from a core recovered three weeks after the 1996 flood (SAG-05/1996post-flood): a) C org (wt%), b) Cinorg (wt%) and c) THgs (ng goI) depth profiles. The short dashed line marks the approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide deposit whereas the stippled line corresponds to the contact between the 1996 flood deposit and the indigenous sediment ______.. ______}~
Figure 10: Geochemical characteristics of sediments at SAG-05 from cores recovered in: a) 1997 (SAG-05/1997), b) 1998 (SAG-05/1998), c) 1999 (SAG- 05/1999(1), d) 1999 (SAG-05/1999(2», e) 2001 (SAG-05/2001) and f) 2002 (SAG-05/2002). The short-dashed and stippled lines mark, respectively, the approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide and 1996 flood deposits ______~_9.
Figure Il: Geochemical characteristics (or signatures) of sediments at SAG-05 from a core recovered in 2001 (SAG-05/2001): a) AVS (Ilmol g'l), b) THgs (ng gol) and c) THgd (ng LoI). The approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide and 1996 flood deposits are marked as in Figure 10 _._ ...... ____ .. ______.... ______.. ____ .... ______.... ______. __ ._. __ .~.7.
Figure 12: Geochemical characteristics of sediments at SAG-05 from a core recovered in 2002 \SAG-05/2002): a) AVS (Ilmol gol), b) THgs (ng gol) and c) THgd (ng LO ) depth profiles. The approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide and 1996 flood deposits are marked as in Figure 10. ______... ______~.7.
Figure 13: Pyrite-associated mercury in the sediments of a core taken at SAG-30 in 2002 (SAG-30/2002): a) THgs (ng goI) and pyrite-associated mercury (py- Hg; ng g.l pyrite); b) pyrite (Ilmol gol): c) % DOP______}9.
Figure 14: Pyrite-associated mercury in the sediments of a core taken at SAG-05 in 2002 (SAG-05/2002): a) THgs (ng gol) and pyrite-associated mercury (py Hg; ng gol pyrite); b) pyrite (Ilmol gol) and AVS (Ilmol gOI*O.Ol); c) % DOP. The approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide and 1996 flood deposits are marked as in Figure 1O. ______. ____ . ______.. ______. _.. _.. ______. _.. ______. ______. ___ . ______4..Q
Figure 15: Solid and pore water methyl mercury (MeHg) and total mercury (THg) profiles in sediment cores recovered at SAG-30 in a) 1999, b) 2000 and c) 200 l, ______. _.. ______. ______.. ______. _____ . ______. _.. ______.4.. ~
Figure 16: Solid and pore water methyl mercury (MeHg) and total mercury profiles (THg) in sediment cores recovered at SAG-09 in a) 1999 and b) 2001. The
vii stippled lines mark the approximate location of the upper and lower boundaries of the flood deposit...... 4~
Figure 17: Solid and pore water methyl mercury (MeHg) and total mercury profiles (THg) in sediment cores recovered at SAG-05 in a) 1999 and b) 2001. The approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide and 1996 flood deposits are marked as in Figure 1O ...... ~~
viii ABSTRACT
The mass f10w event resulting from the 1996 flood in the Saguenay region buried the mercury-contaminated indigenous sediments at the head of the Saguenay Fjord under up to 50 cm of postglacial deltaic sediments. This study investigated whether this deposit served as an efficient geochemical barrier to the remobilization of mercury and its availability to the benthic organisms. The vertical distributions of total mercury and methyl mercury (MeHg) in the sediments and pore waters were measured in box cores recovered at three stations along the main axis of the Saguenay Fjord and in the Baie des Ha!Ha!, in successive years between 1996 and 2002. The total solid mercury profile time-series shows that most of the mercury remobilized from the contaminated, indigenous sediments was trapped below or slightly above the former sediment water interface. Strong correlations with acid-volatile sulphide profiles and extractions of pyrite-associated mercury indicate that most of this mercury was co-precipitated with authigenic iron sulphides. The mercury that was not sequestered by iron sulphides, diffused into the flood layer where it was scavenged by organic matter or methylated. Mercury sequestration at SAG-05 occurred within the older indigenous sediments, in contrast to SAG-09 where it occurred at or above the original sediment-water interface. The sediments are richer in organic matter, more reducing and, thus, establishment of suboxic conditions and precipitation of authigenic iron sulphides occurred more rapidly. The abundance of mercury at the former sediment-water interface and the low dissolved LH2S concentrations, buffered by acid-volatile sulphide precipitation, both favored mercury methylation. A strong correlation between the distribution of acid-volatile sulphides and methyl mercury in the sediment also reveal that the former may serve as a sink for the latter. Throughout the sediment cores, sediment-water partitioning of MeHg as weil as Hg(II) is controlled in great part by the residual organic matter content of the sediment.
ix RÉSUMÉ
À la tête du fjord du Saguenay, un écoulement de masse résultant de l'inondation de 1996 dans la région du Saguenay a enterré les sédiments indigènes contaminés en mercure sous des sédiments deltaïques post-glaciaux atteignant jusqu'à 50 centimètres d'épaisseur. Cette étude a évalué l'efficacité de ce dépôt comme barrière géochimique à la remobilisation du mercure, ainsi que la biodisponibilité de ce dernier aux organismes benthiques. Les distributions verticales du mercure total et du méthyle mercure dans le sédiment et eaux porales ont été mesurés à partir de carottes-boîtes récupérées à trois stations le long de l'axe principal du fjord de Saguenay et dans la Baie des Ha!Ha, entre 1996 et 2002. La série temporelle des profils de mercure total dans le sédiment illustre la séquestration de la majeure partie du mercure remobilisé près de l'ancienne interface eau-sédiment. De fortes corrélations avec les profils de sulfures de fer, ainsi que les extractions du mercure associé à la pyrite, indiquent que la majeure partie du mercure remobilisé a été co-précipité avec ces sulfures de fer authigènes. Le mercure non-séquestré a diffusé dans la couche de dépôt où il a été adsorbé par la matière organique ou méthylé. Les fortes concentrations de mercure à l'ancienne interface eau-sédiment ainsi que les faibles concentrations de rH2S dissous, tamponnées par la précipitation de sulfures de fer, ont favorisé la méthylation du mercure. Une forte corrélation entre la distribution des sulfures de fer et le méthyle mercure dans le sédiment révèle que ceux-ci peuvent également servir de piège au MeHg. Par ailleurs, le fractionnement sédiment-eau du MeHg, aussi bien que celui du Hg(II), est contrôlé en grande partie par la teneur en matière organique résiduelle du sédiment.
x CHAPTER 1
INTRODUCTION
Mercury is a trace element, with a terrestrial abundance of 50 ng/g (Jonasson and Boyle, 1971). It is part of Group 12 in the periodic table, which also contains elements such as zinc and cadmium. It has many of the properties of that group, including a marked chalcophile and oxyphile character (Jonasson and Boyle, 1971). The public and scientific interest Hg has generated in the last 50 years is mainly due to environmental characteristics it shares with the organochlorine and other hydrophobic organic chemicals: its volatility, persistence and potential for bioaccumulation (e.g., Macdonald and Bewers, 1996; Mason et al., 1996; Schroeder and Munthe, 1998; Turner et al., 2001). Because of its volatility, elemental mercury, Hg(O), is transported through the gaseous as weil as the dissolved and particulate phases, with the gaseous phase representing a major pathway for regional and global atmospheric transport (e.g., Macdonald and Bewers, 1996). The residence time of Hg(O) in the atmosphere is on the order of one year and equal to four times the intra-hemispherical mixing time of -3 months (Warneck, 1988; Sierm and Langer, 1993), except in marine environments where its residence time in the marine boundary layer is estimated at about 10 days (Hedgecock and Pirrone, 2004).
The various species of mercury (Le., Hg(O), Hg(lI) salts, organic mercurials) found in the environment have different levels of toxicity and mobility, with methyl mercury being the most toxic. The health risks caused by bioaccumulation of methyl mercury in fish were first exposed after the 1950's outbreak of mercury poisoning in Minamata Bay, Kyushu Island, Japan. The consumption of mercury contaminated fish and other seafood harvested from the Bay caused numerous symptoms of brain damage and the death of several ~-----', hundred villagers (D'itri, 1992; Jahanbakht et al., 2002). Mercury and methyl
1 ~ mercury poisoning can lead to physiological and neurological damage such as kidney damage, impaired motor skills, mood and personality alterations or mental derangement, reduced reproductive success or even coma and death (Compeau and Bartha, 1985; Gilmour et al., 1992; Marvin-DiPasquale et al., 2000). Over the past 40-50 years, mercury has caused the death of approximately 1,400 people and has affected some 20,000 others worldwide, making it one the few metals to have caused human death due to food contamination (D'itri, 1992; Jahanbakht et al., 2002).
Sources of mercury
Uncontaminated surficial sediments usually contain less than 0.4 IJg total Hg/g (e.g., Fôrstner and Wittmann, 1981, Thompson-Roberts and Pick, 2000). Since the beginning of the industrial age (past 200-300 years), background Hg concentrations have increased, from a factor of 2 to 3 in sediment cores (e.g., Nriagu and Pacyna, 1988; Rada et al., 1989; Siemr and Langer, 1993; Swain et al., 1992; Lockhart et al., 1995; Lucotte et al., 1995; Asmund and Nielsen, 2000), up to over one order of magnitude in mid-latitude glacial ice cores (e.g., Schuster et al., 2002). Furthermore, measurements carried out on ombrotrophic peat bogs reveal that the increase could be as high as 1 to 2 orders of magnitude when samples older than 300 years and pre-1800 atmospheric pollution are considered (e.g., Bindler, 2003).
Local industrial inputs and long-range atmospheric transport of mercury are recognized as the two processes responsible for the contamination of aquatic systems (e.g., Loring, 1975; Loring and Bewers, 1978; Lindqvist et al., 1991; Fitzgerald et al., 1998; Louchouarn and Lucotte, 1998). At regional to global scales, the primary vector of Hg to pristine environments is the atmosphere. Mercury enters the atmosphere naturally from the erosion of rocks, volatilization from water, volcanic vents, flora, etc., as the native metal Hg(O), Hg(II), particulate Hg and HgO as weil as dimethylmercury, monomethylmercury (Rudd,
2 .~ 1995; St. Louis et al., 1995; Weber et al., 1998) and other volatile organomercury compounds (Schuster et al., 2002). Over the past 100 years, an increasing number of anthropogenic sources have contributed to atmospheric mercury loading and these now account for about 60 to 70% of the global atmospheric input (Schuster et al., 2002). Major industrial sources of atmospheric Hg include fossil fuel (Le., oil and coal) burning, mining and ore roasting, chlor-alkali plants, incineration of municipal and medical waste and sewage sludge, and non-ferrous metal manufacturing (e.g., Jonasson and Boyle, 1971; Macdonald and Bewers, 1996). The incineration of solid and domestic waste is a dominant source of anthropogenic atmospheric mercury in North America (40%), Central and South America (34%), Africa (30%) and Western Europe (28%). The burning of coal, principally for electrical power generation, represents the dominant source in the developing countries of Asia (42%), eastern Europe and the former USSR (58% during 1983-1989,40% during 1990-1992) (Pirrone et al., 1996; and references herein). Other sources include crematoria, fluorescent lamp breakage, cement manufacturing, dental laboratory, degassing of latex paint, etc. (Pirrone et al., 1996).
Global anthropogenic emissions of Hg to the atmosphere were estimated at 3600 tlyr in 1983, compared to 2600 tlyr for natural emissions (Nriagu and Pacyna, 1988). A more recent estimate of the global anthropogenic mercury emissions is on the order of 1500-2000 tons per year (Mason et al., 1994; Pirrone et al., 1996, 1998). Of this estimate, approximately 332 tlyr were emitted from North America, and 73 tlyr from Central and South America in 1992 (Pirrone et al., 1996). Asia alone accounts for about 46% of the global anthropogenic mercury emissions (Pirrone et al., 1996). Between 1990 and 1995, mercury emissions in Canada were reduced from 34-39 tJyr to 11-15 tJyr (Nriagu, 1994; Environment Canada, 1997, 1998; Allan, 1998). Base metal operations are the largest source of atmospheric Hg in Canada, making up 40% of the Canadian anthropogenic emissions. The eastern Canadian provinces (Québec, New Brunswick, Nova Scotia, Prince Edward Island and Newfoundland) emitted an
3 .~~. estimated 2.8 t or 20-25% of the total 1995 emissions. U.S. and Canadian atmospheric mercury emissions result in wet mercury deposition in southern Québec of 7.6 !-Ig/m 2 per year (Pilgrim et al., 2000).
Speciation of mercury in the surficial cycle
At the local and regional scale, mercury enters the surficial (Le., exogenic) cycle through atmospheric deposition, erosion of the drainage basin and anthropogenic input (Jonasson and Boyle, 1971). Mercury in the surficial cycle is found either in the particulate (e.g. from weathering of rocks containing Hg-bearing minerais, particulate deposition from the atmosphere) or dissolved state. In the latter, mercury is found as Hg(O), Hg(l) and Hg(lI) (e.g., Jonasson and Boyle, 1971). Hg(l) is mostly found as a binuclear (Hg-Hg)2+ species, which is unstable under natural aqueous conditions and reverts to Hg(O) and Hg(II). As the oxidation state of Hg increases, its solubility generally increases. Elemental Hg is sparingly soluble (56 ng/g at 25°C) and volatile (1.22*10-3 mm at 20°C) in water (Schuster, 1991). It is found at variable concentrations in natural water (e.g., Lindberg et al., 2000; Poissant et al., 2000) and is often removed from the aqueous phase through volatilization (Benes and Havlik, 1979). Hg(lI) is the dominant species in oxygenated natural waters. Its speciation further depends on redox and pH conditions, as weil as on cr and S(II) concentrations.
Hg(OHh and HgCI2 would be the expected dominant dissolved species in most natural waters (Jonasson and Boyle, 1971). Mercury, as a groupe 12 transition metal, has a large, easily polarized electron field and, consequently, forms stable complexes with such ligands (Pearson, 1963). The stability of mercury complexes with a number of these ligands varies according to: RS->HS >CN-»r>OH->B(>Cr>NH3»F>N03->sol- (Smith and Martell, 1976; Dyrssen and Wedborg, 1991; Bono, 1997). Accordingly, mercury is found in nature as a number of soluble compounds such as sulphate, nitrate, various chloride,
4 carbonate, hydroxide, and ammonia complexes, as weil as a number of organometallic complexes (Jonasson and Boyle, 1971). Furthermore, mercury(lI) is a d10 transition metal and typically exhibits octahedral geometry. Because of the inefficient shielding of the nucleus by 4f and 5d shells, it tends to form uncharged complexes with anions of moderate electronegativity, like Cr. Neutral chloro-complexes of other group 12 metals are generally ionic, soluble, less hydrophobie and less stable (Turner, 1996; Turner et al., 2001). Whereas both Hg(l) and Hg(lI) can form inorganic complexes, only Hg(lI) can form covalent bonds with carbon to produce organic species su ch as methyl mercury (Kaplan et al., 2002).
ln reality, because of its strong affinity for thiolic and other functional groups, most of the dissolved inorganic mercury is bound to dissolved organic matter (DOM) in natural freshwaters (Back and Watras, 1995; Watras et al., 1995; Sjëblom et al. 2000). Dissolved methyl mercury, on the other hand, is bound to DOM only in highly humic waters (Sjëblom et al., 2000). Similarly, Hg(ll) rarely exists in sediment pore water as a free ion due to its propensity to form complexes or sorb to solids, especially organic matter (e.g., Schuster, 1991; Bloom and Lasorsa, 1999; Turner et al., 2001; Hintelmann and Harris, 2004). Partition coefficients of Hg(lI) between water and sediments (Kc!, defined as [Hg(II)]s/[Hg(lI)]d) vary greatly according to the substrate (log Kd range from 4.5 to 6; Hintelmann and Harris, 2004) and increase with salinity (Turner et al., 2001). Turner et al. (2001) propose that this is due to the hydrophobie characteristics of Hg(ll), especially in the presence of organic matter. Speciation of mercury is thus further complicated by adsorption onto inorganic particulate matter, co precipitation with metal oxides, hydroxides and sulphides, and by adsorption or chelation/organic binding by humic substances (Jonasson and Boyle, 1971).
5 History and source of mercury contamination to the sediments of the Saguenay Fjord
From 1947 to 1976, a chlor-alkali plant was in operation at Arvida on the shores of the Saguenay River, 24 km upstream from the head of the Fjord. Wastewater discharge from this plant and local paper mills was responsible for most of the mercury contamination in the Saguenay River and the adjoining Fjord (Loring, 1975; Loring and Bewers, 1978; Barbeau et al., 1981; Smith and Loring, 1981). Mercury discharge from the chlor-alkali plant increased throughout most of its period of operation, reaching a peak in the late 1960's and early 1970's. During the lifetime of the plant, approximately 300x1 03 kg of mercury were discharged to the fjord of which an estimated 120x103 kg now reside in the sediments of the Saguenay Fjord as a result of scavenging by settling terrigenous particulate organic matter (e.g., Pocklington, 1973; Loring, 1975).
ln the 1970's, mercury contamination in the aquatic fauna of the Saguenay Fjord was documented (e.g., Bligh 1970, 1972; Tam and Armstrong, 1972). Reports of high levels of mercury in fish and shrimp (0.5-10 j.Jg/g) lead to restrictions on commercial fishing of shrimp, crab and cod in the fjord (Smith and
Loring, 1981). In 1972, the implementation of federa~ regulations that govern wastewater discharge to aquatic systems and the application of remedial actions greatly reduced the amount of Hg released by chlor-alkali plants throughout Canada and eventually lead to the closure of the Arvida plant in 1976. These actions, in turn, lead to lowered fluxes of mercury to the sediments (e.g., Smith and Loring, 1981; Gobeil and Cossa, 1984; Louchouarn and Lucotte, 1998) and a dramatic decrease of mercury levels in the biota (Cossa and Desjardins, 1984; Desjardins, 1989).
Despite the relatively high sedimentation rates (Le., 0.2 to >10 cm/yr; Smith and Walton, 1980; Barbeau et al., 1981) and rapid burial of the highly contaminated sediments (Le., in excess of 12 j.Jg Hg/g DW; Loring, 1975) in the Saguenay Fjord, mercury concentrations in surface sediments have remained
6 /~- relatively high. Surface sediments collected between 1964 and 1978 showed
mercury concentrations ranging from 0.16 to 12 ~g/g (Loring and Bewers, 1978; Barbeau et al., 1981). These decreased progressively thereafter: between 1978
and 1986, surface sediments concentrations ranged from 0.01 to 1.2 ~g/g (Pelletier et al., 1989; Gobeil and Cossa 1984, 1993) whereas cores collected in
the early 1990's revealed total mercury levels on the order of 0.2 to 0.8 ~g/g (e.g., Gagnon et al., 1997). These concentrations remain weil above the pre
industrial levels of 0.15 ~g/g (Loring, 1975; Barbeau et al., 1981; Smith and Loring, 1981).
Diagenetic behaviour of mercury
Mercury mobility in marine sediments
The early diagenetic mobility of many metals such as Hg, As, and Se is defined by the degradation or alteration of their carrier phase(s) during burial, in response to physico-chemical changes (e.g., change in redox potential and pH) or biological activity. Mercury in the water column is mostly scavenged by settling organic matter and iron oxides, its main carrier phases (e.g., Loring, 1975; Coquery et al., 1997; Quemerais et al., 1998; Gobeil et al., 1999; Laurier et al., 2003). Loring (1975) demonstrated that particulate organic matter scavenges up to 70-90% of the mercury in the sediments of the Saguenay Fjord. Accordingly, the fate of mercury during burial is greatly affected by the degradation of organic matter during early diagenesis.
ln addition to organic matter, oxide minerais such as Mn- and Fe-oxides, are also effective scavengers of trace metals such as mercury (e.g., Jenne, 1968; Farrah and Pickering, 1978; Gobeil and Cossa, 1993). Oxidized manganese and iron accumulate at the sediment-water interface (Froelich et al., 1979). Under suboxic and sulphate-reducing conditions, these solid phases are
7 .~ reduced chemically or microbially and Fe(lI) and Mn(lI) are released to the pore waters. In turn, the trace metals adsorbed to these solids via surface complexation of the hydroxyl ligands (Forbes et al., 1974) are liberated to the pore waters. The mercury released to the pore waters by the degradation of the organic matter and the dissolution of the oxides follows various pathways. A large fraction of the dissolved Hg may be associated with colloidal dissolved organic matter (DOM) (Guentzel et al., 1996; Stordal et al., 1996; Gagnon et al., 1997). The predominance of Hg(II)-organic matter complexes, as weil as the formation of complexes with sulphides and polysulphides such as HgSl- (Lu and Chen, 1977; Dyrssen, 1985; Paquette and Helz, 1995), may depress the activity 2 coefficient of Hg + and limit the precipitation of HgS (e.g., Gagnon et al., 1997). Dissolved mercury can be re-adsorbed by residual particulate organic matter, it can diffuse to the oxic zone and be adsorbed onto metal oxides (Gag non et al., 1997) and particulate organic matter, it can be adsorbed/co-precipitated with authigenic sulphides such as pyrite and acid-volatile sulphides precipitated in the sulphate reduction zone (Jean and Bancroft, 1986; Hyland et al., 1990; Huerta Diaz and Morse, 1992) or it can be methylated (Gilmour and Henry, 1991; Gagnon et al. 1996, 1997). Hg can also be released to pore waters upon the oxidation and dissolution of Hg-rich iron sulphides by oxygen introduced to the sediment through bioturbation and bioirrigation.
The distribution of dissolved Fe(lI) and reactive (1 N HCI extractable) Fe(llI) phases is affected by factors such as variations in the organic carbon content of the sediment. High organic carbon contents and high rates of sulphate reduction result in their precipitation as iron sulphides such as pyrite and acid volatile sulphides (AVS) (Goldhaber and Kaplan, 1980; Gagnon et al., 1995). In the presence of abundant reduced iron and/or easily reducible Fe(lIl) phases such as encountered in the Saguenay Fjord and the St. Lawrence Estuary, sulphate reduction will likely lead to the preferential precipitation and accumulation of AVS (Mucci and Edenborn, 1992; Gagnon et al., 1995). Huerta Diaz and Morse (1992) have reported significant pyritization of trace metals in
8 -, sulfidic sediments. AVS solids (Le., amorphous iron sulphides, mackinawite and poorly crystallized greigite) have been shown to serve as a sink for many trace metals (Morse and Arakaki, 1993). Given its affinity for sulphide minerais (Jean and Bancroft, 1986; Hyland et al., 1990), a large proportion of the total mercury in the sediments can be associated with authigenic iron sulphides. Whereas pyrite is highly resistant to oxidation, iron monosulphides (AVS) are highly reactive and oxidize within minutes of exposure to oxygen (e.g. Moore et al., 1988; Huerta Diaz and Morse, 1992), releasing the associated trace metals to the surroundings (Moore et al., 1988; Morse, 1994; Gagnon et al., 1997).
Evidence of mercury remobilization during its early diagenesis includes Hg concentrations in sediment pore waters that are several times higher than those measured in the overlying waters (Bothner et al., 1980; Gobeil and Cossa, 1993; Matty and Long, 1995; Gagnon et al., 1996; Gill et al., 1999), significant production of methyl mercury in anoxic sediments (Gag non et al., 1996), as weil as Hg associated with reactive Fe phases such as Fe-oxides and acid-volatile sulphides, as evidenced by strong correlations between their respective vertical distribution in sediments (Gobeil et al., 1999) and during partial chemical extractions.
Factors controlling mercury methylation and demethylation
Methyl mercury is formed from the biotic and abiotic transformation of inorganic mercury, Hg (II), to CH3Hg+ (referred to as MeHg thereafter) (Callister and Winfrey, 1986; Korthals and Winfrey, 1987; Gilmour et al., 1992; Ramai et al., 1993). It is toxic to microorganisms at lower concentrations than Hg(lI) «0.05 IJg/L, compared to 1.0 IJg/L; Jonas et al., 1984; Batten and Scow, 2003). It is also more volatile than Hg(II). In aqueous systems, it can originate from atmospheric deposition and/or from methylation by sulphate reducing bacteria (SRB) in anoxic sediments (Compeau and Bartha, 1985; Gilmour et al., 1992; Marvin-DiPasquale et al., 2000; Zelewsky et al., 2001). Sources of atmospheric
9 ~. Me Hg are not weil characterized. It could originate from industrial activity (Rudd, 1995; Zelewsky et al., 2001) or be produced in the atmosphere from the degradation of dimethylmercury (e.g., Fitzgerald and Mason, 1995).
Although it occurs in both oxic and anoxic conditions, microbial methylation of mercury is favoured by the presence of high organic matter concentrations and reducing conditions. It is, however, inhibited by high pore water sulphide concentrations that may accumulate as a result of active sulphate reduction, as they lead to the precipitation of iron sulphides that adsorb/co precipitate Hg(lI) (Bartlett and Craig, 1981; Corn peau and Bartha, 1985; Gilmour and Henry, 1991). In freshwater sediments, low sulphate concentrations in the pore waters and overlying waters can also limit methylation (Gilmour and Henry, 1991). High methylation rates in the sediment column are therefore expected near the oxic-anoxic boundary where high concentrations of biodegradable carbon and other nutrients (N, P), high sulphate-reduction rates, low [H2S], and a high particle density are encountered. Sorne of the methyl mercury formed under these conditions is absorbed onto organic matter and other solid phases, or is complexed by dissolved and colloidal compounds (Guentzel et al., 1996; Gagnon et al., 1997).
Methyl mercury can be degraded through numerous pathways, either abiotically (for example, through photodegradation, Sellers et al., 1996) or biotically. Biotic reductive degradation is characterized by the production of CH4. Aerobic biotic degradation, in particular, is a detoxification response by bacteria possessing genes of the mer-operon (Tsai and Oison, 1990). Oxidative demethylation, on the other hand, is characterized by the production of C02 and represents a cometabolism of MeHg by heterotrophic bacteria (Oremland et al., 1991). Marvin-DiPasquale et al. (2000) determined that high concentrations of methyl mercury in sediments stimulate the microbial populations that actively degrade MeHg via mer-detoxification, whereas oxidative demethylation is the predominant degradation pathway for sites with low Me Hg concentrations. The
10 .~.. overall or steady state methyl mercury concentrations in sediments reflect the balance between production and degradation rates (Compeau and Bartha, 1985; Marvin-DiPasquale et al., 2000).
Objectives
Reclamation of mercury-contaminated areas is generally a long-term pursuit. Furthermore, remediation techniques for mercury contamination have not been fully investigated, nor have they been extensively applied to the field. Active techniques such as electrokinetics (e.g., Cox et al., 1996; Thoming et al., 2000) and in-situ extraction (e.g., Wasay et al., 1995) appear to be ineffective. For example, electrokinetic processes involve passing a low intensity electric current between a cathode and an anode inserted in the contaminated sediments. This causes the migration of dissolved ions and small charged particles towards the electrodes. This technology has been used for metals such as copper, zinc, lead, arsenic, cadmium, chromium and nickel (e.g., Mulligan et al., 2001). However, in the case of mercury, this process is ineffective because of the low solubility of mercury and its compounds, even when a lixiviant is injected to enhance the oxidation of mercury (e.g., Cox et al., 1996). Biologically-mediated remediation techniques, such as bioleaching, phytoremediation or bioremediation, have only recently been investigated (e.g., Dey and Patke, 2000; Mulligan et al., 2001). Methods such as natural attenuation and stabilization (isolation and containment in order to minimize remobilization, for example by burial under the natural sedimentation regime, by capping, or by addition of an agent that immobilizes the metals), seem to be the only effective ways of reclaiming contaminated areas. This is especially true for contaminated sediments. Investigations of natural attenuation in lakes (e.g., Gbondo-Tugbawa and Driscoll, 1998), estuaries (e.g., Sager, 2002), bays (Hosokawa, 1993) and lagoons (Degetto et al., 1997) have been undertaken. Such remediation projects often include steps for contamination source removal and/or isolation (e.g., by dredging, sand capping, etc.). In Minamata Bay, Japan, a remediation project
11 spanning 23 years (from 1977 to 1990) was elaborated for the reclamation of 1.5 million m3 of sediments in a 2 km 2 area (Hosokawa, 1993). Steps included dredging of the contaminated sediments and their isolation by sand capping. Such undertakings require a detailed knowledge of the distribution, speciation and mobility of mercury in the sediments as weil as continued monitoring after isolation to detect any remobilization.
ln the Saguenay area, the source of contamination (Le. a chlor-alkali plant) was removed and river and fjord sediments were left to slowly decontaminate with time (Le. burial of contaminated sediments under the natural sedimentation regime). Between July 19 and 21 1996, an exceptional meteorological event in the Saguenay region (Yu et al., 1997) resulted in the precipitation of 200 mm of rain over an area of 5000 km2 (Lapointe et al., 1998). This event led to the collapse of dikes at lakes Kénogami and Ha!Ha! and flooding of the area (Lapointe et al., 1998). In the western part of the Saguenay Fjord and the Baie des Ha!Ha!, the flooding caused mass flows that displaced an estimated 15x106 m3 of postglacial marine clays and deposited a layer up to 50 cm thick over the mercury-contaminated sediments of that area. These flood sediments contain much lower levels of mercury and other trace metals (Pelletier et al., 1999; this study) than the indigenous sediments.
It has been proposed (Pelletier et al., 2003; Tremblay et al., 2003; Mucci et al., 2003) that the flood layer could serve to isolate the mercury-contaminated sediments from the water column and the biota. The objectives of this study were to verify this hypothesis by:
• Monitoring the extent of mercury remobilization in the sediment column with time and, thus, verify the efficiency of the flood layer at capping the Hg-Iaden sediments;
• Determining the speciation and phase associations of mercury in the sediments and pore waters after the flood, in order to identify
12 which species are mobile and which phases (e.g. iron sulphides) are responsible for its sequestration. In turn, this allowed us to identify the loci of Hg methylation in the sediments and, thus, provide a better understanding of the bioavailability of mercury.
To evaluate the fate of mercury in the Saguenay Fjord sediments, total mercury analyses were performed on the solid and dissolved phases in cores collected at three stations: SAG-30, SAG-OS, and SAG-09. Sediments at SAG- 30, in the deepest basin of the fjord, were not affected significantly by the flood event and serve to define the steady-state conditions that prevailed before the 1996 flood. SAG-OS and SAG-09, situated in the North Arm of the Fjord and in the Baie des Ha!Ha! respectively, are both within the area affected by the flood event but are characterized by different sedimentation regimes. Solid phase analyses were carried on cores collected every year between 1996 and 2002 at these stations. The dissolved phase (Le. pore water) was isolated and analyzed on cores collected between 2000 and 2002. Cores were recovered using a 0.06m2 Ocean Instruments Mark Il box corer. Sub-sampling was performed on deck in a nitrogen-purged glove box. Pore waters were extracted using Reeburgh-type squeezers. Mercury analyses were performed using a gold amalgamation pre-concentration procedure followed by cold vapor atomic fluorescence spectrometry (CVAFS). Methyl mercury analyses of the solid and dissolved phases were carried out on cores collected in 2000-2002. Solid MeHg was extracted using a KOH in methanol solution whereas dissolved MeHg was measured directly following ethylation by sodium tetraethylborate, gas chromatographie separation, thermal decomposition and detection by CVAFS.
13 CHAPTER2
MERCURY GEOCHEMISTRY IN THE SEDIMENTS OF THE SAGUENAY FJORD: DIAGENETIC BEHAVIOUR FOLLOWING A CATASTROPHIC DEPOSITIONAL EVENT.
Geneviève Bernier, Alfonso Mucci, and Constance Guignard
Department of Earth and Planetary Sciences, McGiII University, 3450 Université, Montréal, QC, Canada H3A 2A7
To be submitted to Applied Geochemistry
14 Absfracf
The mass flow event resulting from the 1996 flood in the Saguenay region buried the mercury-contaminated indigenous sediments at the head of the Saguenay Fjord under up to 50 cm of postglacial deltaic sediments. This study investigated whether this deposit served as an efficient geochemical barrier to the remobilization of mercury and its availability to benthic organisms. The vertical distributions of total mercury and methyl-mercury in the sediments and pore waters were measured in box cores recovered at three stations along the main axis of the Saguenay Fjord and in the Baie des Ha!Ha!, in successive years between 1996 and 2002. The total solid mercury profile time-series show that most of the mercury remobilized from the contaminated, indigenous sediments was trapped below or slightly above the former sediment-water interface. Strong correlations with the vertical distribution of acid-volatile sulphides and results of extractions of pyrite-associated mercury indicate that most of this mercury was co-precipitated with authigenic iron sulphides. The mercury that was not sequestered by iron sulphides, diffused into the flood layer where it was scavenged by organic matter or methylated. Mercury sequestration at SAG-05 occurred within the older indigenous sediments, whereas it occurred at or even above the former sediment-water interface at SAG-09. The sediments at SAG- 05 are richer in organic matter, more reducing and, thus, establishment of suboxic conditions and precipitation of authigenic iron sulphides occurred more rapidly. The abundance of mercury at the former sediment-water interface and the low dissolved LH2S concentrations, buffered by acid-volatile sulphide precipitation, both favored mercury methylation. A strong correlation between the distribution of acid-volatile sulphides and methyl mercury in the sediment also reveals that the former may serve as a sink for the latter. Throughout the sediment cores, sediment-water partitioning of MeHg as weil as Hg(lI) is controlled in great part by the residual organic matter content of the sediment.
15 /_-.., Introduction
Mercury, like lead and cadmium, is included in the "Red-List" of priority pollutants and in List 1 of the EEC Dangerous Substances Directive (1967). It is directly responsible for the death of more than 1400 individuals worldwide over the past 40-50 years (D'ltri, 1992; Jahanbakht et al., 2002). Its toxicity to humans is enhanced by its volatility and the bioaccumulation of methyl mercury, the most toxic Hg species, through the food chain (Bloom, 1992; Macdonald and Bewers, 1996; Schroeder and Munthe, 1998). Its volatility allows for long-range transport through the atmosphere and contamination of otherwise pristine environments (e.g., Lucotte et al., 1995). The residence time of Hg in the atmosphere is on the order of one year and equal to four times the intra-hemispherical mixing time of -3 months (Warneck, 1988; Sierm and Langer, 1993), except in marine environments where its residence time in the marine boundary layer is estimated at 10 days (Hedgecock and Pirrone, 2004). The main vector of mercury to humans is through the aquatic food chain and the ingestion of fish and seafood. The health risks to humans caused by bioaccumulation of methyl mercury in fish were first exposed after the 1950's outbreak of mercury poisoning in Minamata Bay, Japan (D'itri, 1992; Jahanbakht et al., 2002). Mercury poisoning can le ad to physiological and neurological disorders such as impaired motor skills or mental derangement, weakened immune response, reduced reproductive success and even coma and death (Compeau and Bartha, 1985; Gilmour et al., 1992; Marvin DiPasquale et al., 2000; Zelewsky et al., 2001).
Reclamation of mercury-contaminated areas is generally a long-term process. In lacustrine and estuarine sediments, remediation projects usually include natural attenuation through contamination source removal and/or isolation (e.g., by sand capping). In cases where the sediments are to be isolated from the overlying water column, it is critical to have detailed knowledge of the speciation and distribution of mercury in the sediments as weil as to
16 continuously monitor their evolution after isolation to detect remobilization of the mercury.
Between 1947 and 1976, the Saguenay Fjord was contaminated by the activity of a chlor-alkali plant at Arvida, 24 km upstream from the head of the fjord (Figure 1). This plant discharged an estimated 300x103 kg of mercury during its lifetime (Le., 1947 to 1976), 120x103 kg of which now reside in the sediments of the Saguenay Fjord (Loring, 1975; Smith and Loring, 1981). In 1971, reports of severe mercury contamination of the aquatic fauna (e.g., Tam and Armstrong, 1972) led to restrictions on commercial fishing of shrimp, crab and cod in the Saguenay Fjord (Cossa and Desjardins, 1984; Desjardins, 1989). In 1972, the implementation of federal regulations that govern wastewater discharge to aquatic systems and the application of remedial actions greatly reduced the amount of Hg released by chlor-alkali plants throughout Canada and eventually led to the c10sure of the Arvida plant in 1976. Despite these actions and the relatively high sedimentation rates (Le., 0.2 to >10 cm/yr; Smith and Walton, 1980; Barbeau et al., 1981) in the Saguenay Fjord, mercury concentrations in surface sediments have remained high (Le., 0.2 to 0.8 1-19/g) , higher than the pre industrial levels of 0.15 1-19/g (Loring, 1975; Barbeau et al., 1981; Smith and Loring, 1981; Pelletier et al., 1989; Gobeil and Cossa, 1993; Gagnon et al., 1997).
Between July 19 and 21 1996, an exceptional meteorological event in the Saguenay region (Yu et al., 1997; Lapointe et al., 1998) resulted in widespread flooding in the steep catchments along the Saguenay River, the western part of the Saguenay Fjord and the Baie des Ha!Ha!. The flash flood caused mass flows that displaced an estimated 15 x 106 m3 of postglacial marine clays and debris and deposited a layer up to 50 cm thick over the mercury-contaminated sediments of the fjord in that area (Tremblay et al., 2003). These flood sediments contain much lower levels of mercury than the indigenous sediments (Pelletier et al., 1999) they covered. It has been proposed that, since the flood
17 event, the flood layer has served to isolate the contaminated sediments from the water column and the benthic organisms that live close to the new water sediment interface (Pelletier et al., 2003). In this study we investigated the validity of this hypothesis by monitoring the speciation and the extent of mercury remobilization in the fjord sediments after the flood. To document the fate of mercury in the Saguenay Fjord sediments, analyses were performed on the solid and dissolved phases recovered from cores collected at three stations: SAG-30, SAG-05, and SAG-09, situated along the main axis of the fjord and in the Baie des Ha! Ha!, from 1996 to 2002.
Fjord morph%gy and hydrographie eharaeferisfies
The Saguenay Fjord is a 93 km long, 1-6 km wide U-shaped submerged valley, bounded by sheer, vertical walls that reach up more than 300 m above the water line. Situated approximately 150 north-east of Quebec City on the north shore of the St-Lawrence Estuary, it connects with the estuary at Tadoussac through a 20 m deep sill. Its bottom morphology is characterized by three basins separated by two sills at 60 and 120 m deep, located approximately 20 km and 30 km, respectively, from the mouth of the fjord. The outer basin is up to 250 m deep whereas the inner basin has a depth of 275 m. The latter branches out at the head of the fjord into two shallow arms: the Baie des Ha!Ha! to the southwest and the North Arm to the northwest (Figure 1). The North Arm extends towards Chicoutimi and the Saguenay River, the fjord's main tributary (Smith and Loring, 1981; Schafer et al., 1990; Stacey and Gratton, 2001).
The water column is characterized by a sharp pycnocline separating two distinct water masses. The thick bottom layer is well-mixed and oxygenated, with waters penetrating landward from the St. Lawrence Estuary (Thérriaut and Lacroix, 1977; Siebert et al., 1979). Salinity in this layer is approximately 30.5 %0 (Syvitski and Schafer, 1996; Mucci et al., 2000), with temperatures ranging from
18 /~ 0.4 to 1.7 oC (Fortin and Pelletier, 1995). The surficial layer consists of brackish waters (S-0-10 %0) resulting from the turbulent mixing of the outflow from the Saguenay River and the underlying marine waters. The thickness of this layer decreases while its salinity increases progressively towards the mouth of the fjord. The surface water temperatures range from freezing in the Winter to 16°C in the Summer (Fortin and Pelletier, 1995). Detailed hydrographie characteristics of the fjord can be found in Schafer et al. (1990).
Saguenay fJord
/--,
1(00' l!!!!!!-.l-....e::..:;=.,;~_-J 10'00' 1 1 Figure 1: Map of the Saguenay Fjord showing the sampling sites. The number in brackets is the approximate thickness of the flood layer in centimeters. From Mucci et al. (2000).
The Saguenay Fjord is part of the Saguenay-Lac-Saint-Jean watershed, 2 with a drainage basin of 85 500 km . Erosion of glaciomarine clays supplies much of its suspended load (Locat and Leroueil, 1988). Particulate organic matter originates from both natural and anthropogenic sources but mostly from pulp and paper mill discharges (Pocklington and Leonard, 1979; Louchouarn et al. 1997, 1999). Sedimentation rates decrease seaward and range from 2-7 cm/yr at Saint-Fulgence to <0.2 cm/yr in the deep inner basin as deposition is almost entirely a function of hypopycnal processes (Smith and Walton, 1980; Perret et al., 1995).
19 .~ Sediments that accumulate under normal conditions in the Fjord are generally bioturbated and consist of dark grey silty clays to clayey silts (e.g., Smith and Walton, 1980; St-Onge and Hillaire-Marcel, 2001). Their organic carbon content ranges from 0.5 to 3% in mass. The sediments are thus highly reducing and the oxygen penetration depth rarely exceeds 5 mm (Deflandre et al.,2000). Adjacent to the mouth of the Saguenay River, Spring runoff deposits consist of dark organic-rich sandy layers that contrast with the finer-grained material that settles during the low runoff Fall and Winter periods (Smith and Loring, 1981).
The Saguenay Fjord has been the site of a number of recent, major landslides involving Champlain marine clayey silt deposits from the late Wisconsinan period (8,000-11,000 BP). Aside from the 1996 mass flows, other notable events include the May 4th 1971 landslide at St. Jean Vianney and the 1924 slide at Kenogami. Triggered by a period of heavy rainfalls, the St. Jean Vianney slide displaced an estimated 6.9x106 m3 of sediment in a layer that can be detected up to 30 km downstream of the landslide area.
Materials and Methods
Sampling History Throughout the 1990s and up to 2002, cores were collected at a number of stations along the main axis of the Saguenay Fjord. This study focuses on data obtained from three sites: SAG-5, SAG-09 and SAG-30 (Figure 1). SAG-05 (48°24'N 70 049'W) is situated in the North Arm in approximately 90 m of water. The sedimentation rate (w) at this site has been estimated at -1 cm/yr (Smith and Walton, 1980; Mucci and Edenborn, 1992). SAG-09 (48°21'N 70047'W) is in the middle of the Baie des Hal Hal at a depth of approximately 150 m (w = <0.2 cm/yr; Barbeau et al., 1981) whereas SAG 30 (48°21'N 700 23'W) is in the /', deepest part of the inner basin in -270 m of water (w = <0.2 cm/yr; Smith and Walton, 1980).
20 Stations SAG-05 and SAG-09 are both within the area affected by the flood event but are characterized by different sedimentation regimes and sediment geochemistry. Station SAG-30 received only a thin discontinuous layer of the flood material that could only be detected visually in 1996 and, thus, the sediments at this station were not affected by this event (Mucci et al., 2003). For this reason it is used as a reference station to represent the steady-state conditions in the fjord throughout the sampling period. In addition, cores taken before and after the flood at stations SAG-05 and SAG-09 serve to define the initial conditions and the temporal evolution of the sediment chemistry following the flood.
Sampling method
Forty-five to fifty-cm long sediment cores were recovered using a 0.06m2 Ocean Instruments Mark Il box-corer with minor disturbance to the sediment-water interface. Upon recovery, the cores were transferred to and sub-sampled in a customized glove box (Edenborn et al., 1986). The glove box was purged by a continuous flow of nitrogen that limited sediment oxidation during sub-sampling. Vertical sub-sampling of the core was carried out according to a set grid, typically every 0.5 cm for the first cm, every 1 cm over the next 5 to 6 cm, and every 2 to 5 cm over the remainder of the core. As each layer was exposed (Le., by lowering the front plate of the box), Eh measurements and solid sub-samples were taken. Samples for solid- and pore water-phase analyses were transferred to pre weighted plastic scintillation vials and pore water squeezers respectively. The solids were freeze-dried and homogenized by grinding using an agate mortar and pestle. The water content of the sediments and the salinity of the pore waters were used to calculate the sediment porosity. Additional solid samples were taken using mini-cores (Le. 13mL polyethylene screw cap test tubes with their distal ends cut off and a 10 cc syringe plunger) at a rate of two per sampling interval and frozen for acid-volatile sulphide (AVS) and sediment methyl-mercury
21 /~ analyses. Pore waters were extracted with Reeburgh-type squeezers (Reeburgh, 1967) and filtered in-line through a glass microfiber filter and a 0.45
IJm Type HA Millipore R filter before being collected in acid-cleaned 60 cc plastic syringes. The squeezers were washed in tap water and air-dried before being re-used in order to avoid cross-contamination between samples. Pore waters destined for total mercury and methyl mercury analyses were recovered only between 1999 and 2002. The pore waters were then transferred to a number of glass and plastic containers, including acid-cleaned Teflon bottles for total mercury and methyl mercury analyses. The latter were acidified with a 1% equivalent volume of Seastar Ultrapure concentrated HCI.
Analytical methods Analytical procedures for total dissolved and solid-phase mercury determinations were described in detail by Gagnon et al. (1996). Briefly, for solid-phase total
/~-~ mercury [THg]s analyses, micro-wave acid digestion (HCI:HN03; 1: 10) of approximately 0.1g of freeze-dried sediment in Teflon reactors was followed by the reduction of Hg(II) to Hg(O) in a stannous chloride solution, pre-concentration of Hg(O) on a gold-coated quartz sand column, thermal desorption and detection by cold vapor atomic fluorescence spectrometry (CVAFS). Total dissolved mercury concentration, [THg]d, measurements were performed using a similar gold amalgamation pre-concentration and CVAFS detection procedure (Gill and Fitzgerald, 1987) following cold oxidation of the organic Hg compounds by a BrCI solution (Bloom and Crecelius, 1983) and reduction in a stannous chloride solution. When possible, duplicate analyses of the samples were carried out and a reproducibility of better than 10% was obtained, with an average relative standard deviation of 2.8% for samples recovered in 2002. The accuracy, as determined from replicate analyses of the BEST (NRC Canada) sediment standard, was better than 15% in 2001 and 5.8% in 2002. The limit of detection was of 5.1 ng/g, based on 3 times the standard deviation of procedural blanks.
22 Analytical procedures for dissolved and solid-phase methyl-mercury ([MeHg]d and [MeHg]s, respectively) determinations are based on the methods of Sloom and Fitzgerald (1988) and Sloom (1989) as modified and described by Gagnon et al. (1996). [MeHg]d was extracted from the freshly collected and refrigerated pore waters with methylene chloride. The solve nt was evaporated and MeHg back-extracted in a buffered (pH = 4.9) aqueous solution before ethylation. MeHgs was extracted from the frozen-fresh sediments with a 25% KOH in methanol solution for twenty-four hours. Following the extraction, the MeHg in an aliquot of the extract was ethylated with sodium tetraethylborate. The volatile alkylmercury compounds isolated from the pore waters and solids were collected on a Tenax™ trap , separated by isothermal gas chromatography, thermally decomposed on a quartz column and the evolved elemental mercury determined by CVAFS. Duplicate sample and methyl mercury standard (Le., NRC DORM-1 and laboratory prepared MeHgd external standards) analyses yielded a reproducibility of better than 10%.
AVS concentrations were determined on the freshly frozen sediments (Le., mini-cores) according to the procedure described by Chanton and Martens (1985) and Hsieh and Yang (1989) adapted by Gagnon et al. (1995). Pyrite content was determined on the freeze-dried sediments using the method elaborated by Lord (1982). Other analyses (e.g., pore water Mn and Fe, reactive Mn and Fe) performed on these cores are described in detail and results reported in Deflandre et al. (2002) and Mucci et al. (2003).
Resulfs and Discussion
Pre-flood Conditions Steady-state conditions in the sediments of the Saguenay Fjord are defined here by the geochemical signatures of cores recovered prior to the July 1996 flood r-, (Le., SAG-05/1991, SAG-09/1995) or from sites unaffected by the event (Le., SAG-30). The vertical distributions of a number of species (e.g., Corg , Cinorg, AVS
23 """, and pyrite) in these cores were reported previously (Mucci et al., 2003). They show the accumulation of reactive Fe and Mn (1 N HCI-extractable) solids in the thin oxic layer (Le., < 5 mm) near the sediment-water interface at ail three sites. These reflect the remobilization of Fe(lI) and Mn(lI) following the reductive dissolution of authigenic and detrital Fe- and Mn-oxides under suboxic and anoxic conditions (e.g., Froelich et al., 1979; Mucci and Edenborn, 1992; Mucci et al., 2003). At SAG-30, C (Cinorg and Corg), dissolved and 1N HCI-extractable solid Fe and Mn profiles before and after the flood (see Figure 2 in Mucci et al., 2003) support the premise that this site received only a thin film of the flood material and that the event had little effect on the site's sediment geochemistry.
Similarly, the THgs profiles (Figure 2) in cores taken at the same site before and after the flood further support this hypothesis and illustrate the progressive burial of sediment-bound mercury under steady-state conditions.
Surface sediments in cores recovered in the fjord before the 1996 flood
contain total mercury (THgs) concentrations that range from around 100 ng g-1 in the Baie des Ha!Ha! (Figure 3) to 300 ng g-1 at SAG-30 (Figure 2). Although low relative to the period of peak Hg discharge (up to 12000 ng g-1; values obtained by Loring and Bewers (1978) for cores recovered between 1964 and 1976), these surficial concentrations remain high compared to the pre-industrial values of 25-100 ng g-1 estimated by Barbeau et al. (1981) and Louchouarn and Lucotte (1998). It has been proposed (Gagnon et al., 1997) that these high surface sediment concentrations could reflect the deposition of contaminated sediment or soil eroded upstream during periods of high runoff. At SAG-30 and SAG-09, the
bell-shaped patterns of sediment THgs accumulation (Figures 2 and 3) most likely record the history of mercury discharge to the fjord, assuming that little or no diagenetic remobilization of mercury occurs. This interpretation is supported by 210Pb dating of cores collected in 1976 (Smith and Loring, 1981).
It is interesting to note that there is a fairly good correlation between the vertical distribution of particulate organic matter and mercury in the sediments
24 (e.g., at SAG-30/2000: ~=O.55, t(17)=2.70, p a) THg (ng g-1) 0 500 1000 1500 2000 2500 0 X •• • • :t:: X • • •• x •• :t:: • • , • :t:: X 20 • :t:: ) X • • x • :t::• K X :E 30 • c. Q) lx .1996 0 • • X:t:: .1999 40 ~ X X 2001 :t:: • 50 • :t::2002 /--, 60 b) 0 50 100 150 200 o ___~~~---I 10 20 Ê ~ .c:a. 30 oQ) 40 50 60 -'------' Figure 2: Geochemical characteristics and steady-state conditions of sediments recovered at SAG-30. a) Vertical distribution of THgs (ng g-1) on four sampling 25 1 ~, 1 dates over a 6-year period; b) Vertical distribution of THgd (ng L- ) at SAG-30 from a core recovered in 2001 (SAG-30/2001). (1997), the vertical distribution of pore water mercury (THgd) at SAG-30 (Figure 2) and elsewhere in the fjord is characterized by peaks at the oxic/anoxic sediment boundary and its covariance with THgs deeper in the core. The latter suggests that the [THgd1 distribution at depth in the SAG-30 core is mostly controlled by exchange equilibrium with THgs. 1 COrg (wt%) '1norg (w t%) THgs (n9 9- ) 0.00 2.00 4.00 0.00 0.05 0.10 0.15 1000 2000 3000 0 10 10 10 20 20 20 Ê ~ 30 30 30 aal Cl 40 40 40 50 50 50 60 60 60 a) b) c) Figure 3: Geochemical characteristics of pre-flood sediments at SAG-09 from a core recovered in 1995 (SAG-09/1995): a) Corg (wt%), b) Cinorg (wt%) and c) THgs (ng g-1) depth profiles. ln contrast, near the surface, the [THgd1 is likely controlled by the fate of its main carrier phases. Sediment-bound mercury is mostly associated with fresh particulate organic matter and authigenic iron oxides accumulating at or near the sediment-water interface (e.g., Farrah and Pickering, 1978; Gobeil and Cossa, 1993). Under suboxic and sulphate-reducing conditions, organic matter is degraded and the reactive iron oxides are reduced, releasing Fe(lI) and associated trace metals to the pore waters. The fate of mercury released to the pore waters under these conditions is determined by a number of competing processes. A fraction of the dissolved Hg may form organic matter complexes ,~. (Dyrssen and Wedborg, 1991; Guentzel et al., 1996; Stordal et al., 1996; Gagnon 26 c~ et al., 1997; Mason and Sullivan, 1998; Turner et al., 2001), as weil as 2 complexes with sulphides and polysulphides such as HgS2 - (Lu and Chen, 1977; Dyrssen, 1985; Paquette and Helz, 1995; Gagnon et al., 1997). The dissolved mercury can also be re-adsorbed by residual particulate organic matter, it can diffuse to the oxic zone and be adsorbed onto metal oxides (Gag non et al., 1995) and fresh particulate organic matter, be adsorbed/co-precipitated with sulphides (e.g., pyrite and AVS) in the sulphate reduction zone (Jean and Bancroft, 1986; Hyland et al., 1990; Huerta-Diaz and Morse, 1992) or be methylated (e.g., Gilmour and Henry, 1991; King et al., 2001; Hammerschmidt and Fitzgerald, 2004; Hammerschmidt et al., 2004). At station SAG-05 (Figure 4), a landslide that occurred at St. Jean Vianney (Figure 1) in 1971, resulted in the rapid deposition of Corg-poor and Hg poor post-Wisconsinian marine clays. Consequently, it modified the distribution of remobilizable elements such as iron and manganese. Whereas sorne of the reactive Fe migrated up to the new sediment-water interface, most of it was trapped at depth as acid-volatile sulphides (Mucci and Edenborn, 1992). Formed from the reaction between dissolved sulphides and ferrous iron accumulated in the sediment pore waters (Canfield, 1989), the distribution of the AVS is controlled by the amount and reactivity of organic matter and the abundance of dissolved sulphate and ferrous iron (Berner, 1984). Conditions conducive to AVS precipitation could be found at the lower interface between the indigenous sediments and the St. Jean-Vianney landslide material. Based on the covariance between the THgs and AVS profiles (~=0.55, t(14)=2.44, p<0.025; see Appendices 1.3, 2.3), the mercury, remobilized as a result of the degradation of organic matter and the dissolution of authigenic Fe-oxides accumulated at the former sediment-water interface, appears to have been adsorbed/co-precipitated with these sulphides. Further evidence of this remobilization is seen in the elevated [THgd1 within/near the landslide layer as reported by Gagnon et al. (1997; see their Figure 5). 27 c,norg (w t%) lH9. (ng g'1) 0.0 1.0 2.0 3.0 4.0 0.0 0.5 1.0 0 500 1000 1500 0 10 10 10 20 20 20 Ê ~ ~ 30 30 30 oQ) 40 40 40 50 50 50 60.1..------' 60 60 a) b) c) (Umolg,1) AVS (Umolg,1) FeS2 o 100 200 0 50 100 0 10 10 20 20 30 30 40 40 50 50 60 -'------' 60 d) e) Figure 4: Geochemical characteristics of pre-flood sediments at SAG-05 from a core recovered in 1991 (SAG-05/1991): a) C org (wt%) , b) Cinorg (wt%) , c) THgs (ng g-1), d) AVS (IJmol g-1) and e) pyrite (FeS2; IJmol g-1) depth profiles, The short-dashed lines mark the approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide deposit. Flood Material The physical and chemical properties of the material deposited following the 1996 flash flood were described in detail in other publications (e,g., Deflandre et al., 2002; Mucci et al., 2003; Tremblay et al., 2003). In addition to its geotechnical properties (e.g., shear strength; Tremblay et al., 2003), it can be 28 ~. distinguished in the cores collected at SAG-05 and SAG-09 by its light brown grey color, that contrasts with the darker indigenous sediment, and by its higher CinOrg content (Figures 5 and 6). Consequently, the distribution of Cinorg is used to identify the upper and lower boundaries of the flood layer. The variable thickness of the deposited layer reflects the mode of deposition of the material over the bottom topography of the area (Crémer et al., 2002). The flood sediments have low THgs concentrations, ranging between 45 and 144 ng/g at SAG-09 and between 90 and 172 ng/g at SAG-05 (Figures 5 and 9 respectively). Given the jagged vertical distribution of total mercury in the immediate vicinity of the lower boundary of the flood deposit, physical mixing with the contaminated indigenous sediments likely occurred during deposition. Post-flood Geochemistry Following the 1996 flood and the rapid burial of the indigenous sediments, / .~. dissolved oxygen trapped by the deposit was consumed and the oxygen penetration depth migrated to within a few millimeters of the new sediment-water interface (SWI) in less than three weeks (LeFrançois, 1998; Deflandre et al., 2002). Under the prevailing conditions, the authigenic Fe-oxides that had accumulated at the original SWI were progressively dissolved and the associated mercury released to the pore waters. Evidence of this remobilization can be seen throughout the cores, in particular in the elevated pore water Hg concentrations (Figures 7, 8, 11 and 12), in the correlation of the THgs and authigenic iron sulphide distributions at depth and the association of Hg with pyrite (Le., chemical analysis of extracted pyrite) (Figures 13 and 14), as weil as in the production of methyl mercury in anoxic sediments (Figures 16 and 17). Distribution and phase correlations of sediment THg: The vertical distribution of total mercury in the solids sampled from cores collected at SAG-05 and SAG-09 are presented in Figures 6 and 10. Surface sediment concentrations range from 50 to 300 ng g-1 at both stations. At depth, 29 ~, maximum [THgs) generally reach 500 to 600 ng g-1 at SAG-05 and 200 to 400 ng g-1 at SAG-09 but peaks as high as 800 to 900 ng g-1 were observed in sorne cores. Typical profiles are characterized by slightly elevated concentrations in the surficial sediments and lower concentrations in the flood sediments. The remobilization of mercury after the catastrophic depositional event is controlled by its affinity for organic matter and sulphides and their distribution within the sedimentary column after the flood. Like after the st. Jean-Vianney landslide, much of the Fe(lI) released upon the reductive dissolution of authigenic Fe-oxides that had accumulated at the former SWI, was trapped as authigenic sulphides such as AVS (Le., amorphous FeSx, mackinawite, poorly-crystallized greigite) and pyrite following the 1996 flood (Mucci et al., 2003). The abundance of metabolizable organic matter at the former SWI increased sulphate reduction rates which, in the presence of pore water Fe(lI) or reactive Fe(lIl) phases in the sediment, promotes AVS precipitation (Berner, 1984; Canfield, 1989; Gagnon et al., 1995). Sulphide minerais are strong scavengers of metals such as mercury (e.g., Jean and Bancroft, 1986; Hyland et al., 1990; Huerta-Diaz and Morse, 1992; Morse and Arakaki, 1993; Gagnon et al., 1997) and, thus, sorne of the mercury remobilized following these catastrophic events was immobilized by these solids in the vicinity of the lower boundary of the deposits (Figures 7, 8, 11 and 12). At SAG-09, the covariance between [THgs) , organic carbon (1996: t(10)=6.24, p<0.005; 1997: t(16)=5.82, p<0.005; and 1999(2): t(14)=3.94, p<0.005; 2000: t(15)=4.79, p<0.005; 2002: t(15)=3.66, p<0.005) and iron sulphides (A VS) (AVS: 1998: t(15)=4.30, p<0.005; 2000: t(17)=2.55, p<0.01; see Appendices 1.2, 2.2) contents confirms that these are the main carrier phases or sinks for this element (Figures 7 and 9). [THgs1 peaks usually occur below the AVS maxima. This probably reflects the fact that AVS serve as a sink for remobilized mercury and, thus, inhibit its diffusion up the sedimentary column. 30 However, as will be seen below, the geochemical behavior of mercury at this station is made more complex by mercury methylation in the flood layer. COll! (w t%) c,nOIl! (w t%) THg. (ng g.1) 0.0 1.0 2.0 3.0 4.0 0.00 0.20 0.40 0.60 0.80 0 100 200 300 400 0 10 10 10 20 20 20 Ê ~ :5 30 30 30 Co (1) 0 40 40 40 50 50 50 60 60 60 a) b) c) Figure 5: Geochemical characteristics of sediments at SAG-09 from a core recovered three weeks after the 1996 flood (SAG-09/1996post-flood): a) C org (wt%) , b) Cinorg (wt%) and c) THgs (ng g-1). The stippled line marks the approximate location of the contact between the flood material and indigenous sediment. 31 ) ) ) lHg. (ng g-l) lHg. (ng g-l) lHg. (ng g-l) lHg. (ng g-l) 500 1000 100 200 300 100 200 300 400 200 400 ------, o 1 .. 0 1997 1:1:10 10 10 10 ] ]~ 20 f 20 20 Ê ~ ...... 30 ~ 30 30 :5 30 ~ ""- C- G) 0 40 '\ 40 • 40 40 50 50 50 50 60 60 60 60~1------J a) b) c) d) lHg. (ng g-1) lHg. (ng g-l) lHg. (ng g-l) o 500 1000 200 400 600 50 100 150 200 01. o 1 ... 10 2000 10 10 20 20 20 30 30 40 40 40 50 50 50 60~1------~ 60~1------~ 60~1------~ e) f) g) Figure 6: THgs depth profiles at SAG-09 from cores recovered in a) 1997 (SAG-09/1997), b) 1998 (SAG-09/1998), c) 1999 (SAG- 09/1999(1», d) 1999 (SAG-09/1999(2), e) 2000 (SAG-09/2000), f) 2001 (SAG-09/2001) and g) 2002 (SAG-09/2002)_ The stippled lines mark the approximate location of the upper and lower boundaries of the flood deposit. 32 AVS (urrol g-l) nt. (ng g-l) ntd (ng L-l) 0 50 100 150 200 400 600 0 10 20 30 40 0 0 10 10 10 20 20 20 Ê ~ J:: 30 30 30 "El- Q) 0 40 40 40 50 50 50 60 60 60 a) b) c) Figure 7: Geochemical characteristics of sediments at SAG-09 from a core recovered in 2001 (SAG-09/2001): a) AVS (~mol g-\ b) THgs (ng g-1) and c) THgd (ng L-1). The stippled lines mark the approximate location of the upper and lower boundaries of the flood deposit. AVS (urrol g-l) nt. (n9 g-l) o 50 100 150 50 100 150 200 20 40 60 80 10 10 10 20 20 20 Ê ~ 30 30 30 '8.Q) 0 40 40 40 50 50 50 60 L-____--I 60 L-____---l 60 a) b) c) Figure 8: Geochemical characteristics of sediments at SAG-09 from a core recovered in 2002 (SAG-09/2002): a) AVS (~mol g-1), b) THg (ng g-1) and c) THg (ng L-1). The stippled lines mark the approximate location of the upper and lower boundaries of the flood deposit. 33 At SAG-05, positive correlations between THgs, organic carbon (e.g., 1996: t(23)=3.81, p<0.005; 1997: t(13)=3.41, p<0.005; 1998: (t(15)=3.97, p<0.005; and 1999(2): t(19)=3.65, p<0.005) and iron sulphides (e.g., for pyrite, 1996: t(25)=2.86, p<0.005; 1997: t(13)=4.83, p<0.005; and 1998: t(15)=3.77, p<0.005; for AVS, 1998: t(16)=5.61, p<0.005; and 1999: t(13)=12.62, p<0.005; 2001: t(17)=6.86, p<0.005; 2002: t(20)=7.39, p<0.005; see Appendices 1.3, 2.3) are also observed in a number of cores. Again, limited remobilization is indicated by a variable distribution of AVS and mercury as weil as apparent scavenging of mercury by organic matter within the flood layer. Mercury levels deeper in the indigenous sediments, however, appear similar to pre-flood conditions. Peak solid mercury concentrations are seen 10 to 20 cm below the base of the 1996 mass flow layer and appear to be associated with high AVS and THgd (Figures 11 and 12). These AVS peaks were formed following the 1971 St. Jean-Vianney landslide (Mucci and Edenborn, 1992) and the associated mercury appears to have remained immobile since then and after the 1996 flood. corg (wt%) Gnorg (wt%) THg. (ng g.l) 0.00 1.00 2.00 3.00 4.00 0.000 0.100 0.200 0.300 0.400 a 200 400 600 800 a Or-~~~~ 60 60 60 .1....-___-----1 ~ ~ ~ Figure 9: Geochemical characteristics of sediments at SAG-05 from a core recovered three weeks after the 1996 flood (SAG-05/1996post-flood): a) Corg (wt%) , b) Cinorg (wt%) and c) THgs (ng g-1) depth profiles. The short-dashed line marks the approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide deposit whereas the stippled line corresponds to the contact between the 1996 flood deposit and the indigenous sediment. 34 ) ) 1 1 1 TItI. (n9 gl) TItI. (n9 9- ) TItI. (n9 9- ) TItI. (n9 9- ) o 200 400 600 o 100 200 300 o 500 1000 o 200 400 600 800 01 .... 01 , ------19<]9------i999 10 10 10 1) 20 20 20 20 Ê ~ .r::. 30 30 30 30 ë. ID C 40 40 40 40 50 50 50 50 60 ______J kl 60~1------J 60~1------~ 60i------~ a) b) c) d) 1 TItI. (n9 gl) TItI. (n9 9- ) 200 400 600 200 400 600 2002 10 10 20 20 30 40 40 50 50 il 60LI------~ 60 ------J e) f) Figure 10: Geochemical characteristics of sediments at SAG-05 from cores recovered in: a) 1997 (SAG-05/1997), b) 1998 (SAG- 05/1998), c) 1999 (SAG-05/1999(1), d) 1999 (SAG-05/1999(2», e) 2001 (SAG-05/2001) and f) 2002 (SAG-05/2002). The short-dashed and stippled lines mark, respectively, the approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide and 1996 flood deposits. 35 AVS (umol g-l) lH9d (ng L-l) o 100 200 300 400 200 400 600 o 50 100 150 10 10 10 20 20 20 Ê ~ oC 30 15. ~ 40 40 40 50 50 50 60.1....------' 60.1...------' 60.1...------' a) b) c) Figure 11: Geochemical characteristics of sediments at SAG-05 from a core recovered in 2001 (SAG-05/2001): a) AVS (~mol g-1), b) THgs (ng g-1) and 1 c) THgd (ng L- ). The approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide and 1996 flood deposits are marked as in Figure 10. AVS (umol g-l) lH9. (ng g-l) lH9d (ng L-l) 0 100 200 300 400 200 400 600 0 50 100 150 0 0 ~-,,-,,-,,-,,_ .. 10 10 10 20 20 20 Ê ~ oC 30 30 30 15. Q) 0 40 40 40 50 50 50 60 60 60 a) b) c) Figure 12: Geochemical characteristics of sediments at SAG-05 fram a core recovered in 2002 (SAG-05/2002): a) AVS (~mol g-1), b) THgs (ng g-1) and 1 c) THgd (ng L- ) depth profiles. The approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide and 1996 flood deposits are marked as in Figure 10. 36 Pore water THg distribution: Depth profiles of total dissolved mercury (THgd) in sediment cores recovered from SAG-05 and SAG-09 between 2000 and 2002 are presented in Figures 7, 8, 11 and 12. Concentrations range from 5.2 to 107 ng L-1 at SAG-05 and from 1.5 to 72 ng L-1 at SAG-09. They are relatively low in the oxidized surface sediments but increase with THgs levels at depth. Elevated [THgd] are found just below the oxic layer in some cores (Figures 7, 11), a result of organic matter degradation and the dissolution of Fe oxides with burial. Variable and elevated [THgd] within the flood layer as weil as in the older indigenous sediments may reflect the remobilization of mercury following the dissolution of authigenic Fe oxides and the degradationof organic matter from the older, Hg-Iaden indigenous sediments, remobilization driven by the concentration gradient in the Hg-poor flood sediments and, to a smaller degree, compaction of the high porosity flood deposit (Maurice and Locat, 2000; Tremblay et al., 2003). Combined with the high affinity of mercury for organic matter and sulphides, this results in a high degree of covariance between [THg d], [THg s], Corg and AVS throughout the cores. Extraction of pyrite-associated mercury Authigenic iron sulphides (Le., AVS and pyrite) are important sinks for adsorbed/co-precipitated trace metals (Huerta-Diaz and Morse, 1992; Morse and Arakaki, 1993; Morse, 1994). The apparent association (Le., co-variance) of mercury with acid-volatile sulphides in the sediments of the Saguenay Fjord was noted above. AVS are metastable and precursor phases of authigenic pyrite in marine sediments (Berner 1970, 1984; Goldhaber and Kaplan, 1980; Rickard, 1975; Gagnon et al., 1995). The trace elements associated with AVS are released during its conversion to pyrite and can be re-adsorbed by other solid phases (Elderfield et al., 1979). Conversion rates of AVS to pyrite appear to be controlled by the availability of reductants (e.g., organic carbon; Berner 1970, 1984), by sulphide concentrations (Canfield and Raiswell, 1991) as weil as by the time elapsed since burial (Middleburg, 1991; Gagnon et al., 1995). 37 Whereas the association of a number of metals with AVS has been documented (e.g., Huerta-Diaz et al., 1998), the determination of mercury associated with these solids has proven to be difficult due to its propensity to re adsorb onto other solid phases upon dissolution of the AVS. On the other hand, measurements of mercury associated with pyrite [py-Hg] (Figures 13a and 15a) were obtained following the isolation of pyrite using the method of Lord (1982), dissolution of the solid in concentrated nitric acid, and CVAFS analysis. The efficiency of Hg sequestration by pyrite can be investigated further by identifying the loci of authigenic pyrite formation, here identified by the degree of pyritization (DOP; Berner, 1970). The %DOP is defined as %DOP = [py-Fe]/([py-Fe]+[reactive-Fe])*100, where py-Fe is the pyrite concentration and reactive-Fe corresponds the 1N HCI extractable Fe (Raiswell et al., 1994). • llig. (ng g-l) pyrite (IJmol g-l) %OOP 0 1000 2000 0 2 4 6 8 10 12 0 0.5 1.5 0 0 10 10 10 20 20 20 Ê ~ .r. 30 30 30 ë.. Q) Cl 40 40 40 50 50 50 _ py-Hg (ng g-1 py) 60 60 60 a) b) c) Figure 13: Pyrite-associated mercury in the sediments of a core taken at SAG-30 in 2002 (SAG-30/2002): a) THRs (ng g-1) and pyrite-associated mercury (py-Hg; ng g-1 pyrite); b) pyrite (IJmol g- ); c) % DOP. 38 01 01 • llig. (n9 9 ) • pyrite (Ilmol 9 ) %OOP 0.00 5.00 10.00 15.00 o 200 400 600 020406080 0~~-~---1 10 20 Ê ~ oC 30 15. ~ 40 40 _._._._._._._. 01 50 50 _ py-Hg (ng 9 py) 60.l..-----....J 60.1..-..-----' 60.1..------1 a) b) c) Figure 14: Pyrite-associated mercury in the sediments of a core taken at SAG-05 in 2002 (SAG-05/2002): a) THgs ~ng g-1) and pyrite-associated mercury (py-Hg; ng g-1 pyrite); b) pyrite (tJmol g-) and AVS (tJmol g-1*0.01); c) % DOP. The approximate location of the upper and lower boundaries of the 1971 St. Jean Vianney landslide and 1996 flood deposits are marked as in Figure 10. A high proportion of the mercury is associated with pyrite at both SAG-30 and SAG-05 (Figures 13a and 14a). At SAG-30 (Figure 13), the DOP is relatively constant at approximately 0.8%. The amount of mercury associated with pyrite increases with depth and follows the THgs depth profile and, thus, the historic input of mercury to the sediments. At SAG-05 (Figure 14), however, the DOP increases with depth to values averaging 2% and up to 12%, whereas py Hg:THgs ratios remain relatively constant, ranging between 0.02 and 0.09. Maximum pyritization and pyrite concentrations occur below the AVS maximum within the older indigenous sediments. Preferential association of mercury with pyrite has been previously identified even at these low DOP levels and in diverse sedimentary environments (e.g., Huerta-Diaz and Morse, 1992). ;--'. 39 Mercury methylation: Methyl mercury concentrations in sediments reflect a balance between methylation and demethylation rates (Compeau and Bartha, 1985). The microbial methylation of mercury is generally favoured by the presence of high concentrations of organic matter (biodegradable carbon and other nutrients) and reducing conditions (Compeau and Bartha, 1985; Gilmour and Henry, 1991). It is, however, inhibited by high pore water sulphide concentrations as they lead to the formation of stable HgHS+ complexes and the precipitation of HgS and iron sulphides such as AVS and pyrite that adsorb/co-precipitate Hg(lI) and, thus, limit the amount of mercury available to methylation (Bartlett and Craig, 1981; Compeau and Bartha, 1985; Gilmour and Henry, 1991). On the other hand, in environments with abundant reactive iron such as the Saguenay Fjord sediments, the precipitation of iron sulphides inhibits the build-up of pore water IH2S, which promotes Hg methylation (e.g., Gagnon et al., 1995). In the presence of these two competing processes, methylation can still be favoured if Hg is sufficiently abundant. Low methylation rates and/or demethylation are promoted by aerobic, high-salinity conditions (Compeau and Bartha, 1985). The methyl mercury can be absorbed onto organic matter and other solid phases (Bartlett and Craig, 1981; Compeau and Bartha, 1985; Gilmour and Henry, 1991), or complexed by dissolved and colloidal compounds (Guentzel et al., 1996; Gagnon et al., 1997). Vertical profiles of dissolved and solid methyl mercury (MeHg) in sediment cores recovered along the main axis of the fjord in 1992, four years prior to the flood, were published in Gagnon et al. (1996). In the Bras Nord (Le., SAG-05, SAG-06; Figure 1), the solid MeHg profiles were characterized by peaks below the oxic layer and within or near the lower boundary of a landslide layer at depths corresponding to maximum THgs concentrations (Gagnon et al., 1997). Similarly, throughout the fjord, higher pore water methyl mercury (MeHgd) concentrations were found below the oxic/anoxic boundary and within/near landslide layers; they are also strongly correlated with [Fed1 (Gag non et al., 1996). High sulphate 40 reduction rates and low [rH2S] at the oxic/anoxic boundary are conducive to the buildup of high [MeHg] (Watras et al., 1995; Gagnon et al., 1996). The presence of reactive and reducible Fe phases and low [rH2S] within the landslide deposits also appear to favour mercury methylation (Gag non et al., 1996). Dissolved Me Hg profiles in cores taken at SAG-30 between 2000 and 2001 are strongly correlated to the distribution of [THgd] and are similar to the one measured by Gagnon et al. (1996) in 1992. The MeHgd produced at the oxic/anoxic boundary (Le., ... 5 mm) is scavenged by solid phases as it diffuses upwards through the oxic sediments, resulting in a [MeHgs] peak near the sediment-water interface in certain cores. At depth, both solid and pore water MeHg follow the THgs distribution (Figure 15). At SAG-09, MeHgs accounts for variable proportions of the sediment total mercury concentration at each depth, ranging from 0.01 to 0.21 % (0.0268 to 0.216 ng Hg g-1) in 1999 and 0.02 to 0.62% (0.014 to 0.457 ng Hg g-1) in 2001. MeHgd, on the other hand, makes up between 0.96 and 91 % of the [THgd], with values ranging from 6.66 to 57.4 ng Hg L-1 in 1999 and from 0.31 to 85.8 ng Hg L-1 in 2001. Elevated [MeHgd] throughout the cores are evidence of mercury remobilization and diffusion towards the new SWI. Whereas the 1999 core shows a [MeHgs] maximum below the oxic layer, the 2001 core displays elevated concentrations within the surface oxic layer. This observation contrasts with those of Gagnon et al. (1996) who proposed that oxic sediments serve as a geochemical barrier to the diffusion of MeHgd to the water column through demethylation. Our results suggest that MeHgd may be adsorbed to solid phases such as fresh particulate organic matter or authigenic metal oxides in the oxic layer. 41 1 1 1999 IIt1eHg. (n9 Hg 9- ) THg. (n9 9- ) ~. 0.00 0.25 0.50 o 1000 2000 10 10 20 20 .eÊ a 30 30 a> C 40 40 50 50 60 60.1..------.1 a) 1 IIt1eHgd (n9 Hg L-l) 1 2000 IIt1eHg. (n9 Hg 9- ) THg. (n9 9- ) 0.00 0.50 1.00 0 10 20 1000 2000 3000 o 100 200 300 0 0 1) 10 10 10 20 20 20 20 Ê /--- .e 30 30 30 30 ~a> C 40 40 40 40 50 50 50 50 60 60 60 60 b) 1 1 IIt1eHg. (n9 Hg 9- ) IIt1eHgd (n9 Hg L-l) THg. (n9 9- ) THgd (n9 L-l) 2001 0.00 0.10 0.20 0.30 0 50 100 1000 2000 3000 0 50 100 150 200 o +--"'":--=*"'-----1 0 0 10 10 10 10 20 20 20 20 Ê ~ ..c: 30 30 30 30 0- -a> C 40 40 40 40 50 50 50 50 60.1.------1 60 60 60 c) Figure 15: Solid and pore water methyl mercury (MeHg) and total mercury (THg) profiles in sediment cores recovered at SAG-30 in a) 1999, b) 2000 and c) 2001. 42 r--' SAG-09: MeHg. (ng Hg g-l) THg. (ng g-l) 1999 0.00 0.10 0.20 0.30 20 40 60 80 200 400 O+---~_-_t 10 10 20 20 Ê ~ 30 ______.s:: 30 30 ! 40 40 50 50 50 60 -'------' 60 -'------' 60 -'------' a) MeHg. (ng Hg g-l) 10 20 30 40 2001 0.00 0.20 0.40 0.60 50 100 200 400 600 10 10 10 10 20 20 20 20 30 30 30 40 40 40 40 50 50 50 50 60.1.-..------1 60.1...------' 60 -'------' 60 -'------' b) Figure 16: Solid and pore water methyl mercury (MeHg) and total mercury profiles (THg) in sediment cores recovered at SAG-09 in a) 1999 and b) 2001_ The stippled lines mark the approximate location of the upper and lower boundaries of the flood deposit. 43 At depth, the 2001 core shows elevated [MeHgsl throughout most of the flood layer, resulting in a negative correlation between [MeHgs1 and [THgs1, in contrast to the general observations made for the other cores (e.g., for Figure 15), and possibly indicating local conditions conducive to the methylation and sorption of the remobilized mercury. Figure 17 shows the vertical distributions of MeHgs and MeHgd at SAG-05 from cores recovered in 1999 and 2001. MeHgs makes up between 0.01 % and 0.40% of the total mercury at each depth interval. The concentrations are much lower than those measured before the flood, ranging from 0.034 to 0.600 ng Hg g-1, including the surface concentrations that are on the order of 0.1 - 0.2 ng Hg g-1, compared to concentrations that exceeded 2.9 ng-Hg g-1 before the flood 1 (Gagnon et al., 1996). [MeHgd1vary between 0 and 45 ng Hg L- and account for 1% to 74% of the total dissolved mercury in the 2001 core. At both SAG-05 and SAG-09, high [THg d1 and reducing conditions promote bacterial methylation near the oxic/anoxic boundary, resulting in a MeHgd maximum at this depth (Le., -3.5 cm) seen in Figures 16 and 17. On the other hand, upward diffusion of MeHgd through the oxic sediments (Le., top 5 mm) appears to be inhibited through a yet unconfirmed process, either by adsorption onto solid phases such as particulate organic matter or by demethylation (Gag non et al., 1996). As mentioned above, the first process could in particular explain the behaviour observed in the 2001 core at SAG-09. As observed in other environments, organic matter plays a major role in controlling sediment-water partitioning of both MeHg and Hg(lI) (e.g., Hammerschmidt et al., 2004). The second process, demethylation, could occur through a number of pathways, such as oxidative demethylation, a mechanism promoted by high sulphate and metal oxide concentrations (Jackson, 1989; Oremland et al., 1991) or aerobic biotic degradation, a detoxification response by bacteria possessing genes of the mer-operon (e.g., Tsai and Oison, 1990). 44 SAG-05: 1999 IVeHg. (ng Hg g-l) IVeHgd (ng Hg L-l) llig. (ng g-l) 0.000 0.500 1.000 0 10 20 30 40 50 0 200 400 600 800 0 0 0 10 10 10 20 20 20 Ê ~ :5 30 30 30 C- a> 0 40 40 40 50 50 50 60 60 60 a) 2001 IVeHg. (ng Hg g-l) THg. (ng g-l) 0.00 0.50 1.00 1.50 o 20 40 60 o 200 400 600 50 100 150 0 0_ ...... -~--1 10 10 10 10 20 20 20 20 Ê ~ .r: 30 ë. a> 0 40 40 40 40 50 50 50 50 60.L..--___--l 60 60.l...-----....J 60.l...-----...J b) Figure 17: Solid and pore water methyl mercury (MeHg) and total mercury profiles (THg) in sediment cores recovered at SAG-05 in a) 1999 and b) 2001. The approximate location of the upper and lower boundaries of the 1971 St. Jean-Vianney landslide and 1996 flood deposits are marked as in Figure 10. 45 The Hg(lI) released upon demethylation would be adsorbed by the organic matter and authigenic Fe-oxides in the oxic zone (Farrah and Pickering, 1978; Gobeil and Cossa, 1993). The vertical distributions of [MeHgd1 and [MeHgs1 appear to be mostly controlled by exchange reactions with the solid components of the sediments such as organic matter and acid-volatile sulphides (Figures 16 and 17). As mentioned above, in addition to the availability of dissolved sulphate, major factors controlling AVS formation are the amount and reactivity of organic matter and the availability of reducible iron minerais in the sediments (Berner, 1984; Gagnon et al., 1995). Following the deposition of the flood layer, the bacterial degradation of fresh organic matter and the reductive dissolution of authigenic iron oxides lead to AVS precipitation near the former SWI. In turn, by preventing the build-up of rH2S in the pore waters, AVS precipitation can promote mercury methylation. Sorption by AVS would provide a sink for both mercury and the methyl mercury formed below the flood deposit. Finally, AVS peaks formed under similar conditions following the deposition of the 1971 St. Jean-Vianney landslide material appear to have been unaffected by the 1996 event and have remained immobile. The mercury remobilized from the buried, contaminated sediments at this location, co-precipitated with authigenic iron sulphides, was mostly sequestered below the original SWI, whereas the Hg moved up higher and was sequestered within the flood layer at SAG-09. The indigenous sediments at SAG-05 are richer in organic carbon and, thus, more reducing, resulting in a more rapid establishment of anoxic conditions following the mass flow event and a more rapid and localized precipitation of authigenic sulphides. 46 .~~. Summary and Conclusions Prior to the 1996 Saguenay flood and the rapid deposition of post-Wisconsinian sediments at the head of the fjord, the vertical distribution of total mercury in the sediment (THgs) of the Baie des Ha!Ha! reflected the progressive burial of sediment-bound mercury under steady state conditions and recorded a fairly accu rate history of mercury discharge to the Saguenay Fjord. The concentration of pore water mercury (THgd) in surface sediments was controlled by the fate of its main carrier phases, particulate organic matter and iron oxides, and by an exchange equilibrium with THgs at depth. As described in Mucci and Edenborn (1992) and Mucci et al. (2003), a former landslide event (Le., 1971 St. Jean-Vianney landslide), recorded in the sediments at SAG-05 in the North Arm of the fjord, modified the distribution of redox-sensitive elements such as iron and manganese, resulting in the immobilization of the reactive Fe as acid-volatile sulphides (AVS) and pyrite in the vicinity of the former sediment-water interface. The mercury, remobilized as a result of the degradation of organic matter and the dissolution of Fe-oxides, appears to have been adsorbed/co-precipitated with these sulphides. Burial of the contaminated indigenous sediments by mercury-poor postglacial deltaic sediments during the 1996 mass flow event resulted in the rapid consumption of the dissolved oxygen contained in the water trapped with the deposit and the migration of the oxygen penetration depth to within a few millimeters of the new sediment-water interface (SWI) in less than three weeks (Deflandre et al., 2002). Upon the establishment of suboxic/anoxic conditions in the flood and former surficial indigenous sediments, mercury was remobilized following the reductive dissolution of authigenic Fe-oxides and the degradation of reactive organic matter. 47 At SAG-09, most of the reduced iron released to the pore waters was sequestered as acid-volatile sulphides and pyrite in the vicinity of the former SWI and within the lower section of the flood deposit, whereas much of the remobilized Hg appears to have co-precipitated/adsorbed onto these sulphides. At SAG-05, the AVS peaks formed following the 1971 St Jean-Vianney landslide (Mucci and Edenborn, 1992) and the associated Hg, remained immobile and unaffected by the later 1996 event. The mercury remobilized from the buried, contaminated sediments at this location also seems to have been sequestered by authigenic iron sulphides. In contrast to SAG-09, the precipitation of authigenic iron sulphides and the sequestration of Hg occur mostly below the original SWI. The indigenous sediments at SAG-05 are richer in organic carbon and, thus, more reducing than at SAG-09, resulting in a more rapid establishment of anoxie conditions upon the deposition of the flood material and a more localized precipitation of authigenic sulphides which, in turn, limits Hg migration. A similar behaviour was observed for arsenic in these sa me sediments (Mucci et al., 2003). Our results therefore support Mucci et a/.'s (200:~) conclusion that metals forming distinct insoluble sulphide minerais or that co-precipitate with Fe sulphides would be immobilized and trapped in the vicinity of the former sediment-water interface. Results of our pyrite-associated mercury extractions and, in particular, a comparison of the py-Hg:THg ratio with DOP confirms the strong affinity of this trace metal for pyrite, a stable and important sink of Hg in these sediments. Nevertheless, a fraction of the Hg(lI) released from the buried, contaminated sediments migrated upwards into the flood sediments, to be likely scavenged by residual particulate organic matter. Following the release of Hg(II), bacterial methylation of mercury in the vicinity of the former SWI resulted from two competing processes: the precipitation of acid-volatile sulphides favoured mercury methylation by inhibiting the build-up of pore water rH2S, whereas mercury adsorption onto these 48 .~ sulphides limited the quantity of mercury available to methylating bacteria. In this environment, the bio-available mercury was methylated and most of the MeHg appears to be either sorbed onto the freshly precipitated iron sulphides or to the residual organic matter found in the vicinity of the former SWI. Throughout the sediment column, sediment-water partitioning of MeHg as weil as Hg(II) seems to be controlled in great part by the residual organic matter content, resulting in covariant distributions of Hg(II) and MeHg with organic carbon. Mercury methylation could be further enhanced by bioturbation, by advecting labile organic substrates to depth, or by making inorganic Hg bio available to methylating bacteria and, thus, increase its availability to benthic organisms. However, in areas such as at SAG-Q9, the Hg is trapped below the depth at which most benthic organisms borrow, thus limiting its availability to the benthic fauna. Finally, the persistence of high Hg concentrations at the new SWI can be explained by the resumption of the normal sedimentation regime and the delivery of Hg-Iaden particles originating from the erosion of contaminated soils in the drainage basin. 49 CHAPTER3 SUMMARY AND CONCLUSIONS Acting as both sink and source of toxic contaminants, sediments represent one of the most significant risk factors to water quality and benthic organisms, and ultimately to ecosystem and human health (US Environmental Protection Agency, 1997). Remediation technologies established for soil mercury are often not applicable to sediment due to the latter's higher percentage of clay, silt and organic matter. Methods such as natural attenuation and containment by burial under the natural sedimentation regime or capping seem to be the only effective ways of reclaiming these contaminated aquatic environments. Evaluation of the environmental risks posed by trace metal contamination is a determining factor in selecting remediation actions and requires a clear understanding of the geochemical behaviour of the metal, its mobility, the identity and efficiency of its permanent sinks, as weil as its pathways into the biosphere. ln the Saguenay region, the fjord sediments were left to slowly decontaminate with time since removal of the source of mercury in 1976. The vertical distribution of total mercury in sediments (THgs) of the Baie des Ha!Ha! and downstream, prior to 1996, reflected the progressive burial of sediment bound mercury under steady state conditions and recorded a fairly accu rate history of mercury discharge to the Saguenay Fjord (Smith and Loring, 1981). At SAG-05, in the North Arm of the fjord, a previous landslide event (i.e., the 1971 St. Jean-Vianney landslide) modified the distribution of redox-sensitive elements such as iron and manganese, resulting in the immobilization of the reactive Fe as acid-volatile sulphides (AVS) and pyrite in the vicinity of the former sediment water interface. The mercury, remobilized as a result of the degradation of organic matter and the dissolution of Fe-oxides, appeared to have been adsorbed/co-precipitated with these sulphides (Gag non et al., 1997). 50 ~.. The postglacial marine clays deposited during the 1996 flash flood in the upper reaches of the Saguenay Fjord and the Baie des Ha!Ha! contained much lower levels of mercury and other trace metals (Pelletier et al., 1999; this study) than the indigenous sediments. Rapid consumption of the dissolved oxygen contained in the water trapped with the deposit and migration of the oxygen penetration depth to within -5 mm of the new sediment-water interface, (Le, less than 3 weeks; LeFrançois, 1998) led to the establishment of suboxic/anoxic conditions in the flood and former surficial indigenous sediments. Under these reducing conditions, authigenic Fe and Mn oxides that had accumulated at the original sediment-water interface (SWI) were progressively dissolved. The reductive dissolution of these oxides as weil as the degradation of organic matter in the older, Hg-Iaden indigenous sediments released the associated mercury to the pore waters. ln determining the speciation and phase associations of sediment mercury after the flood, observed correlations between sediment total mercury and acid volatile sulphide distributions at both SAG-09 and SAG-05 suggest that most of the mercury remobilized from the buried contaminated sediments was sequestered by the AVS. Direct measurements of AVS-associated mercury could not be obtained as the mercury released upon the selective dissolution of the AVS is rapidly re-adsorbed onto other solid phases. Whereas, at SAG-09, part of the pore water mercury diffused upwards in the flood deposit before being immobilized, the sequestration of Hg at SAG-05 occurred mostly below the original SWI. The indigenous sediments at SAG-05 are richer in organic carbon and, thus, more reducing, resulting in a more rapid establishment of anoxie conditions upon the deposition of the flood material and a more localized precipitation of authigenic sulphides which, in turn, limited Hg migration. Determinations of pyrite and reactive iron in the sediments of the fjord gave DOP values of approximately 0.8% for SAG-30 and as high as 12% for 51 /'""" SAG-05. The results obtained from pyrite-associated mercury extractions and, in particular, a comparison of the py-Hg:THg ratio with DOP, confirms the strong affinity of this trace metal for pyrite, a stable and important sink of Hg in these sediments. At SAG-05, maximum pyritization and pyrite concentrations occur within the St. Jean-Vianney deposit and below the AVS maximum formed after the landslide event within the indigenous sediments deposited thereafter. A preferential association of mercury with pyrite has been previously identified in diverse sedimentary environments, even at the low DOP levels encountered at SAG-30 (e.g., Hurta-Diaz and Morse, 1992). Our results support Mucci et a/.'s (2003) conclusion that metals that form distinct insoluble sulphide minerais (e.g., Zn) or co-precipitate with Fe sulphides (e.g., As) will be immobilized and trapped in the vicinity of the former sediment water interface (SWI) in these organic-rich sediments. Beyond this important sink near the former SWI, the variable and elevated pore water Hg(lI) concentrations measured within the flood layer as weil as in the older indigenous sediments may reflect the remobilization of mercury released from the buried, contaminated sediments, and its enhanced solubility in the presence of high pore water dissolved organic carbon concentrations (Deflandre et al., 2002). The remobilization of Hg through the sedimentary column is driven by the concentration gradient in the Hg-poor flood sediments and, to a smaller degree, compaction of the high porosity flood deposit (Maurice and Locat, 2000; Tremblay et al., 2003). Methyl mercury concentrations in sediments reflect the balance between methylation and demethylation rates (Compeau and Bartha, 1985). In the Saguenay Fjord, this results in MeHgs that accounts for 0.01 to 0.62% of the sediment total mercury content at SAG-09, whereas MeHgd makes up between 0.96 and 91 % of the [THgd]. Similarly, at SAG-05, MeHgs makes up between 0.01 % and 0.40% of the sediment total mercury content whereas [MeHgd1 vary between 0 and 45 ng-Hg/L and account for 1% to 74% of the total dissolved 52 ~.. mercury. With only limited remobilization of mercury occurring upwards through the Hg-poor flood sediments, MeHgd is now present at much lower concentrations (ranging from 0.034 to 0.600 ng-Hg g-1) than those measured before the flood (in excess of 2.9 ng-Hg g-1; Gagnon etaI., 1996). The high sulphate reduction rates but low dissolved [I:H2S] at the oxic/anoxic boundary are conducive to mercury methylation and resulted in the elevated solid and pore water methyl mercury concentrations, particularly at Station 09. The abundance of reactive Fe phases in the flood deposit (Mucci et aL, 2003) buffers dissolved I:H2S to low concentrations and favours mercury methylation (Gag non et al., 1996). Following the release of Hg(lI) upon the reductive dissolution of authigenic Fe-oxides and the degradation of organic matter, bacterial methylation of mercury in the vicinity of the former SWI resulted from two competing processes: the precipitation of acid-volatile sulphides favoured mercury methylation by inhibiting the build-up of pore water I:H2S, whereas mercury adsorption onto these sulphides limited the quantity of mercury available to methylating bacteria. In this environment, the bio-available mercury was methylated and most of the Me Hg appears to be either sorbed onto the freshly precipitated iron sulphides or to the residual organic matter found in the vicinity of the former SWI. Throughout the sediment column, sediment-water partitioning of Me Hg as weil as Hg(lI) seems to be controlled in great part by the residual organic matter content and associated organic ligands, a process proposed to explain partitioning of these compounds in sediments of Long Island Sound (Hammerschmidt et al., 2004) and Lavaca Bay (Bloom et al., 1999), resulting in covariant distributions of Hg(lI) and MeHg with organic carbon. Mercury methylation may be further enhanced by bioturbation: by advecting labile organic substrates to depth, removing toxic by-products, or by making inorganic Hg bio-available to methylating bacteria and, thus, increasing its availability to benthic organisms. Over long time scales, the feeding habits and subsurface migration of benthic infauna can be felt to depths greater than 53 /""" one meter in sediment (Benninger et al., 1979) but, in the Saguenay area, most of the benthic organisms (Le., polychaetes) that recolonized the sediment do not burrow much below 15 cm in depth (Michaud et al., 2003). In areas such as at SAG-09, where the Hg is trapped below this depth, it is unlikely that they will be able to access it. Finally, the persistence of high Hg concentrations at the new SWI, with surface sediment concentrations ranging from 50 to 300 ng g-1, can be explained by the resumption of the normal sedimentation regime and the delivery of Hg laden particles originating from the erosion of contaminated soils in the drainage basin. RECOMMENDA TIONS FOR FUTURE WORK This study addressed and resolved a number of questions regarding the behaviour of mercury in the Saguenay Fjord sediments, in particular regarding its relationship with iron sulphides and the effect of the 1996 flood event on its mobility and, to sorne extent, bio-availability. It has, thus, augmented our extensive and growing knowledge base of that area, while having an important effect in the management of the Saguenay system and its resources. Nevertheless, certain unresolved issues still merit further inquiry. Our findings indicated that mercury was effectively sequestered but further, episodic, observations may be necessary to monitor the progress of the natural decontamination, to verify the efficiency of this capping mechanism over time and ensure that mercury remains isolated from the water column and unavailable to the benthic fauna, particularly in areas where the thickness of the flood deposit would permit disturbance by burrowing organisms or maritime activities (e,g" dragging of ship anchors). Such activities could oxidize the metastable AVS with which mercury is associated and release it back to the pore 54 ~ or overlying waters where it would become available to organisms. Future landslide events, or dredging activities near harbour installations, also constitute possible disturbances to this capping mechanism. Further investigation of the present and future biotic MeHg burden of organisms such as polychaetes and bivalves would be necessary for a complete evaluation of the bio-availability of mercury in this system. Disturbances to the fjord sediments, or benthic mixing or irrigation, may enhance mercury methylation by redistributing the inorganic Hg to zones where it is available to methylating bacteria. It may therefore be prudent to further evaluate the amount of mercury actually available to methylating bacteria in the following years, and the link between sedimentary production of MeHg, mobilization and bioaccumulation within the Saguenay Fjord sediments. ln addition, an elucidation of certain aspects of the mercury speciation and phase associations in the Saguenay Fjord sediment would prove useful. This study provided interesting insights on the relationship of mercury with both AVS and pyrite in the Saguenay sediments. Nevertheless, direct measurements of AVS-associated mercury could not be obtained, due to its rapid re-adsorption onto other solid phases upon the AVS dissolution. If possible, elaboration of extraction protocols for, and precise measurements of, the AVS-associated mercury would confirm this association and its mechanism (e.g., adsorption or co-precipitation) as weil as the role of sedimentary sulphides as sinks of Hg. Whereas the pyrite extractions mainly provided information on the loci of authigenic pyrite formation, and the role of Fe in Iimiting the precipitation of pyrite in the sediment column, a temporal evolution of the DOP compared to py-Hg:Hg ratios would shed light on the importance and nature of this association. Such a study would greatly benefit from refinements to the extraction and analytical methodologies. 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Benoit (1997) High resolution simulation of the severe precipitation event over the Saguenay, Quebec region in July 1996. Geophysical Research Letters, 24, 1951-1954. Zelewski, L.M., G. Benoit, and D.E. Armstrong (2001) Mercury dynamics in Tivoli South Bay, a freshwater tidal mudflat wetland in the Hudson River. Biogeochemistry, 52, 93-112. 71 APPENDIX 1: Distribution and Concentration of Total Mercury Appendix 1.1: SAG-30 (Inner Basin Saguenay Fjord) SAG-30: Inner Basin Saguenay Fjord 1996 1999 2000 2001 2002 .1 Depth (cm) THgs (ng gO') Depth (cm) THgs (ng gO') Depth (cm) THg. (ng gO') Depth (cm) THgs (ng g-') THgd (ng L-') Depth (cm) THgs (ng g-') TH9d (n9 L-') 0.25 305 0.25 286 0.25 277 0.25 298 6.8 0.25 288 72 0.75 347 0.75 473 0.75 375 0.75 430 19 0.75 275 49 1.5 554 1.5 456 1 398 1.5 398 12 1.5 326 52 2.5 513 2.5 531 3 547 2.5 479 6.9 2.5 369 60 3.5 422 3.5 502 4 521 3.5 714 106 3.5 446 66 4.5 491 4.5 499 6 699 4.5 413 19 4.5 466 83 5.5 504 6 648 8 667 6 423 26 6 461 46 6.5 584 8 523 10 957 8 475 13 8 554 55 7.5 759 10 700 12.5 826 10 629 19 10 531 42 8.5 913 12 834 15.5 1233 12.5 654 8.4 12.5 523 50 9.5 860 14 1770 21.5 1910 15.5 888 162 15.5 1664 55 10.5 1110 16 1540 25 1084 18.5 1910 27 18.5 1728 20 11.5 1177 19.5 970 28 544 21.5 1770 19 22.5 837 17 12.5 1066 22.5 621 32 406 25 952 20 27.5 325 8.6 13.5 1299 25.5 350 36 101 29 423 9.3 32.5 141 13 14.5 1333 28.5 198 40 63 33 197 6.1 37.5 265 8.2 15.5 1498 31.5 152 44 54 37 198 15 42.5 60 10 16.5 1743 34.5 91 48 56 42 206 11 47.5 51 12 17.5 1561 38 121 18.5 1225 42 95 19.5 1238 46 99 20.5 1357 50 87 21.5 1368 22.5 1344 23.5 1191 24.5 863 ---- 72 \ ) Appendix 1.2: SAG-09 (Baie des Ha! Ha!) 1.5 49 0.75 76 0.75 97 8.6 6.4 2.5 215 57 1.5 1.5 93 70 1.5 242 9.1 1.5 85 8.6 3.5 233 11.5 82 57 2.5 2.5 54 70 2.5 148 8.4 2.5 73 7.6 4.5 375 20 74 2.5 43 3.5 62 4.0 108 4.00 37 3.5 88 13.4 3.5 38 9.3 5.5 386 21.25 61 3.5 58 4.5 58 6.0 79 6.00 35 4.5 69 16.7 4.5 49 8.7 6.5 295 23 135 4.5 53 5.5 47 8.0 53 8.00 36 6 74 19.4 6 46 10.7 7.5 313 25 128 5.5 53 6.5 55 10.5 46 10.00 30 8 109 11.3 8.5 32 14.0 8.5 279 26.5 257 6.5 46 7.5 54 13.5 39 12.50 27 11.5 194 13.7 12 30 22.1 9.5 135 27.5 28B 8 51 9 66 17.0 66 15.50 22 14 81 6.5 16 40 12.8 11 824 28.5 348 10 74 11 100 21.0 46 18.50 39 18 36 15.6 19.5 39 13.8 13.5 742 30 253 12 44 13 52 25.0 65 21.50 46 22 92 10.4 22.5 45 7.5 16.5 1230 32.5 350 14 101 15 60 27.75 181 24.50 25 26 65 5.9 25.5 41 71.9 19.5 2635 36 337 16 94 17 63 30.0 43 27.50 31 29.5 50 1.5 28.5 56 20.4 23.5 2336 45 201 17.5 76 19.5 69 32.25 51 30.50 24 32.5 81 9.8 31.5 80 10.5 28.5 391 19 70 24.5 112 34.5 58 34.00 158 35.5 124 6.2 34.5 137 34.2 33.5 279 21 89 29 199 37.5 121 36.50 265 38.5 130 7.2 37.5 104 28.1 38.35 116 25 163 34 247 40.5 231 39.50 304 41.5 382 32.3 40.5 179 29.1 43.2 26 29.5 473 39 229 43.5 319 45 342 15.3 44 91 12.2 48.3 3.0 35 661 46.5 308 49 394 48 54 17.2 52.95 4.0 41 810 Appendix 1.3: SAG-05 (North Arm Saguenay Fjord) 0.75 0.75 90 0.75 135 0.75 144 0.75 19.4 1.5 1.6 139 1.25 129 1.5 121 70 1.5 82 1.5 25.9 9.8 2.5 410 1.75 76.9 2.7 58 1.75 129 2.5 81 93 2.5 88 2.5 71.7 2.5 12.8 3.5 500 2.25 55 3.7 44 2.75 98 3.5 96 69 3.5 84 3.5 77.3 3.5 9.6 4.5 690 2.75 144 4.7 38 4 84 4.5 68 42 4.5 87 4.5 53.8 4.5 14.8 6 535 3.5 45 5.7 71 5 81 6 88 8 54 6 73 8.5 41.1 6 12.1 8 415 4.5 79.1 7.2 46 6 109 8 83 10 73 8.5 60 11.5 77.2 8.5 11.2 10 320 5.5 71 9.2 52 7 101 10 63 12 58 11.5 68 15 75.9 11.5 9.7 12.5 400 6.5 103 11.2 68 8 44 12.5 72 14.5 54 15 155 19 61.8 14.5 105 15.9 15.5 545 7.5 91 13.2 105 9 62 15.5 63 17.5 53 19 168 23 47.3 17.5 139 25.0 18.5 800 8.5 84 15.2 62 10 46 18.5 141 20.5 132 23 442 27 107 20.5 186 18.4 22 215 9.5 94 19.2 182 11.5 75 21 156 23 129 26 269 31 80.1 24 239 20.6 27 190 10.5 134 23.2 211 13.5 81 23 177 29.5 613 28 521 35 52.0 28.0 525 45.5 33 155 11.5 193 25.2 156 15.5 78 25 226 32 429 30 349 39 88.7 31.5 269 33.7 39 1000 12.5 198 27.2 215 17.5 42 27.5 402 35 365 32 424 43 44.4 34.5 383 56.4 13.5 243 32.2 569 19.5 63 30.5 530 39 434 34 317 37 319 56.4 14.5 201 37.2 214 23 279 33.5 228 43 332 36 278 39 193 35.4 16 248 28 257 36.5 640 38 275 41 217 35.4 18 228 33 180 39.5 360 40 208 43 362 92.3 20.5 671 43 335 42 344 45 304 92.3 23.5 247 47 369 44 353 27.5 383 51 301 35 302 45 383 73 ) '\) APPENDIX 2: Distribution and Concentration of Geochemically Important Elements with Depth Appendix 2.1: SAG-30 (Inner Basin Saguenay Fjord) 1996 1999 May21 1999 May22 AVS Dep1h C,,,, Dep1h C .... FeCHen Dep1h c_ C"" FaS. C_ MeHg, 1 (cm) (%Wii (~:)(%wll (cm) (%wt) (%wt) (~molg") (~molg") (cm) (%wt) (%wtl (ng Hg g") 0.25 0.074 2.53 3.61 0.25 0.09 2.60 1.81 4.03 0.25 0.08 2.59 0.08 0.75 0.081 2.71 4.50 0.75 0.09 2.55 1.62 2.33 0.75 0.08 2.64 0.14 1.5 0.077 2.8 3.49 1.5 0.10 2.88 1.08 2.72 1.5 0.08 2.6 0.08 2.5 0.081 2.88 3.46 2.5 0.10 2.69 2.45 3.38 2.5 0.08 2.59 0.08 3.5 0.092 2.89 3.43 3.5 0.09 2.64 3.84 2.64 3.5 0.08 2.65 0.19 4.5 0.084 2.79 3.37 4.5 0.09 2.58 7.87 2.21 4.5 0.08 2.68 0.23 5.5 0.079 2.72 3.49 6 0.08 2.61 12.6 2.18 6 0.08 2.67 0.15 6.5 0.083 2.71 3.57 8 0.08 2.62 24.2 2.28 8 0.08 2.64 0.14 7.5 0.077 2.62 3.53 10 0.08 2.64 28.4 2.38 10 0.08 2.59 0.14 8.5 0.077 2.58 3.59 12 0.08 2.56 35.9 2.24 12 0.08 2.67 0.21 9.5 0.069 2.57 3.69 14 0.08 2.50 39.5 2.25 14 0.08 2.59 0.31 10.5 0.070 2.53 3.65 16 0.08 2.56 37.5 3.20 16 0.08 2.57 0.26 11.5 0.074 2.47 3.73 19.5 0.08 2.51 59.6 2.41 19.5 0.08 2.44 0.08 12.5 0.082 2.57 3.69 22.5 0.08 2.47 56.1 2.59 22.5 0.07 2.42 0.10 13.5 0.076 2.56 3.83 25.5 0.07 2.41 65.8 2.48 25.5 0.07 2.29 0.16 14.5 0.070 2.55 3.71 28.5 0.07 2.18 56.1 2.54 28.5 0.07 2.12 0.07 15.5 0.077 2.47 3.71 31.5 0.08 2.06 55.3 2.21 31.5 0.08 2.04 0.12 16.5 0.079 2.5 3.84 34.5 0.08 1.96 55.3 2.20 34.5 0.07 1.97 0.08 17.5 0.065 2.49 3.72 38 0.07 1.88 71.5 2.51 38 0.07 1.9 0.05 18.5 0.072 2.48 3.63 42 0.07 1.74 50.8 2.28 42 0.07 1.81 0.06 19.5 0.071 2.46 3.52 46 0.07 1.74 82.5 2.35 46 0.08 1.83 0.44 20.5 0.073 2.46 3.69 50 64.0 2.03 50 0.08 1.82 0.04 21.5 0.077 2.38 3.45 22.5 0.073 2.33 3.50 23.5 0.072 2.33 3.45 24.5 0.068 2.21 3.13 74 ) '\ ) :lendix 201: SAG-30 (Inner Basin Saguenay Fjord) 2000 2001 2002 lpth C inorq AVS MeHg, MeHg Depth C inora Cora Fe(HCn AVS MeHg, MeHg Depth inora Cora Fe(HCn AVS Fe~ Hg-py C"'" d dO C gO, :m~ ~%wt) !%wt~(limai gO') (ng Hg go')I(ng Hg LO') (cm) (%wt) (%wt) (%wt) (limai gO') (ng Hg go')I(ng Hg L ') (cm) (%wt) (%wt) (%wt) (limai gO') (limai gO') (ng py) .25 0.10 2.71 1.97 0.72 0.58 0.25 2.64 4.34 2.20 0.18 1.64 0.25 0.08 2.66 4.69 2.22 4.91 108 .75 0.11 2.68 1.94 0.91 1.24 0.75 0.09 2.6 4.25 0.18 0.75 0.08 2.77 5.05 2.17 5.69 101 .5 0.10 2.73 1.96 0.47 6.43 1.5 0.09 2.72 4.33 1.64 0.11 1.41 1.5 0.08 2.59 4.46 5.94 6.11 85 !.5 0.09 2.68 2.01 0.12 NID 2.5 0.08 2.62 4.02 3.48 0.11 1.97 2.5 0.08 2.75 4.83 8.85 7.79 140 4 0.10 2.76 2.80 0.46 3.12 3.5 0.07 2.82 4.01 5.44 0.12 3.5 0.09 2.66 5.04 13.0 7.48 234 6 0.10 2.69 24.5 0.31 2.57 4.5 0.07 2.7 4.06 12.7 3.03 4.5 0.08 2.66 5.19 28.7 7.38 241 8 0.09 2.64 38.5 0.35 2.55 6 0.07 2.63 3.84 0.12 81.1 6 0.08 2.71 5.31 7.4 232 10 0.10 2.53 39.4 0.38 8 0.07 2.67 3.65 11.9 0.12 1.62 8 0.08 2.67 5.11 7.42 232 2.5 0.10 2.52 57.0 0.34 1.55 10 0.07 2.58 3.87 36.6 0.07 1.02 10 0.08 2.66 4.78 30.8 7.97 165 5.5 0.09 2.61 60.1 0.60 4.49 12.5 0.07 2.53 3.78 34.9 0.11 12.5 0.07 2.61 5.02 29.1 7.52 200 B.5 0.10 2.61 71.1 0.30 2.04 15.5 0.08 2.51 3.73 46.2 0.13 2.09 15.5 0.08 2.55 4.97 63.0 8.22 715 1.5 0.09 2.61 54.5 0.59 2.47 18.5 0.07 2.52 3.85 52.4 0.22 1.21 18.5 0.08 2.55 4.83 87.0 7.66 661 4.5 0.08 2.51 79.1 0.31 1.71 21.5 0.09 2.43 3070 54.5 0.01 22.5 0.07 2.37 4.43 73.4 7.95 310 18 0.09 2.21 49.9 0.21 2.04 25 0.07 2.32 3.54 49.0 0.00 3.76 27.5 0.07 2.21 4.38 66.5 6.58 129 12 0.08 2.09 43.9 0.24 1.36 29 0.07 2.13 3.68 60.5 0.02 4.93 32.5 0.07 2.08 4.57 88.8 9.9 84 16 0.09 1.99 127.6 0.23 0.82 33 0.08 2.12 3.74 57.8 0.00 0.57 37.5 0.08 1.84 4.86 76.4 7.55 62 10 0.09 1.84 120.5 0.20 1.18 37 0.08 1.85 3.75 78.9 0.89 42.5 0.08 1.79 4.27 88.7 8.38 70 14 0.09 1.80 67.8 0.18 42 0.09 1.75 3.74 42.2 0.08 47.5 0.08 1.69 4.14 73.0 8.56 49 18 0.08 1.74 189.7 0.33 75 ') ) ! Appendix 2.2: SAG-09 (Baie des Ha! Ha!) 1995 1996 1997 Depth C inolll C 0111 Depth Cino", C OIlI FeS2 Depth Cino", Co", FeCHel) FeS2 (cm) (%wt) (%wt) (cm) (%wt) (%wt) (~molg") (cm) (%wt) (%wt) (%wt) (~molg") 0.5 0.10 3.62 0.5 0.48 1.41 21.4 0.25 0.25 1.55 2.41 9.69 1.5 0.07 2.85 1.5 0.59 1.48 23.9 0.75 0.32 1.26 3.72 10.5 2.5 0.09 2.61 2.5 0.57 1.39 24.6 1.25 0.50 0.99 1.69 17.0 3.5 0.09 2.56 11.5 0.48 1.2 20.1 1.75 0.52 1.16 1.90 18.2 4.5 0.08 2.39 20 0.32 1.14 20 2.5 0.53 1.17 1.92 19.5 5.5 0.09 2.25 21.25 0.32 1.05 17.4 3.5 0.51 1.27 2.01 24.3 6.5 0.09 2.12 23 0.25 1.12 19.3 4.5 0.55 1.21 1.98 23.3 7.5 0.10 1.77 25 0.20 1.5 11.7 5.5 0.50 1.27 1.88 19.5 8.5 0.09 1.61 26.5 0.08 2.54 3.9 6.5 0.52 1.07 1.76 19.6 9.5 0.09 1.26 27.5 0.08 2.45 6.09 8 0.47 0.82 1.36 16.4 11 0.07 2.83 28.5 0.10 2.34 13.3 10 0.46 1.11 1.69 18.7 13.5 0.07 2.99 30 0.09 1.61 12.5 12 0.42 1.08 1.52 13.3 16.5 0.07 2.75 32.5 0.08 2.9 3.77 14 0.40 1.01 1.30 14.6 19.5 0.07 3.03 36 0.08 3.09 4.35 16 0.34 1.03 1.08 17.5 23.5 0.06 2.88 45 0.07 2.42 3.31 17.5 0.25 0.2 0.64 7.23 28.5 0.03 2.57 19 0.17 0.65 0.98 7.58 33.5 0.07 2.24 21 0.14 1.19 1.88 9.34 38.35 0.07 1.65 25 0.09 2.16 3.23 8.18 43.2 0.07 1.45 29.5 0.06 2.89 2.91 7.12 48.3 0.08 1.46 35 0.08 2.62 2.83 4.87 52.95 0.10 1.38 41 0.06 2.5 2.50 6.43 76 ') '\ / Appendix 2.2: SAG-09 (Baie des Ha! Ha!) 1998 1999 (May) Depth C inorg C Org FeCHCO AVS FeS2 Depth C inorg C Org FeCHCI) AVS FeS2 1 1 1 1 (cm) (%wt) (%wt) (%wt) (IJmoI9- ) (IJmoI9- ) (cm) (%wt) (%wt) (%wt) (IJmoI9- ) (IJmoI9- ) 0.25 0.13 1.91 2.72 1.51 4.39 0.25 0.11 2.06 2.85 0.90 4.01 0.75 0.21 1.61 2.53 1.37 7.02 0.75 0.10 1.93 3.32 1.16 4.59 1.5 0.39 1.32 2.23 1.9 11 1.5 0.10 2.03 3.15 1.28 4.54 2.5 0.46 1-24 2.72 1.83 10.5 2.5 0.36 1.52 2.49 5.53 16.6 3.5 0.57 1.16 1.94 1.73 20.8 4 0.44 1.58 1.98 12.4 21.9 4.5 0.57 1.31 2.05 1.56 19.5 6 0.44 1.17 1.53 5.14 12.4 5.5 0.59 1.23 2.03 2.36 20.5 8 0.59 1.63 2.14 7.67 19.2 6.5 0.58 1.21 1.97 3.42 19.6 10.5 0.54 1.14 1.67 5.45 23.7 7.5 0.53 1.29 1.47 3.85 20.5 13.5 0.51 1.3 1.59 12.1 14.0 9 0.56 1.21 1.93 3.42 18.7 17 0.48 1.05 1.50 8.97 12.4 11 0.48 0.83 1.4 3.2 11.6 21 0.45 0.87 1.36 6.43 11.6 13 0.50 1.14 1.61 4.88 15.7 25 0.49 1.57 1.64 35.0 15.4 15 0.46 1.1 1.43 5.97 14.7 27.75 0.30 8.47 1.31 65.0 28.6 17 0.45 1.08 1.43 5.58 12.9 30 0.40 1.42 1.12 20.6 15.3 19.5 0.43 1.13 14.1 13.3 32.25 0.26 0.33 0.60 39.5 8.38 24.5 0.27 0.5 0.68 7.78 7.14 34.5 0.26 0.57 0.60 47.1 9.01 29 0.14 2.18 2.49 0.89 4.99 37.5 0.27 2.99 1.10 97.7 23.7 34 0.09 1.32 1.67 36.7 10.4 40.5 0.10 2.87 3.12 21.5 5.64 39 0.11 1.44 2.18 47.9 4.65 43.5 0.08 2.92 2.31 45.0 17.5 46.5 0.08 2.09 2.16 41.0 13.6 77 ') ') ) pendix 2.2: SAG-09 (Baie des Ha! Ha!) 1 (Augustl 2000 2001 epth C inora C oro MeHg. MeHgd Depth C lnoro C oro FeCHCIl AVS Depth C inoro C oro AVS MeHg. MeHgd ::ml (%wt) ~%wtl(ng Hg g-')I(ng Hg L-') (cm) (%wt) (%wt) (%wt) (IJmolg-') (cm) I%wti I%wt) (Il mol g-') (ng Hg g-')I(ng Hg L-') .25 0.13 1.88 7.09 0.03 0.25 0.11 2.01 2.98 1.67 0.25 0.13 1.89 2.27 0.46 4.58 .75 0.14 1.87 6.66 0.06 0.75 0.10 2.01 3.26 1.88 0.75 0.11 1.90 2.02 0.38 5.49 1.5 0.30 1.68 7.92 0.07 1.5 0.11 1.93 3.34 1.56 1.5 0.20 1.88 5.80 0.12 ~.5 0.36 1.8 21.9 0.10 2.5 0.20 1.63 2.89 3.68 2.5 0.32 1.91 10.6 0.15 2.09 4 0.38 1.26 57.4 0.03 4 0.49 1.32 1.98 5.69 3.5 0.40 1.42 6.02 0.05 4.08 6 0.49 1.35 57.4 0.03 6 0.50 1.26 1.82 11.1 4.5 0.48 1.53 0.09 3.29 3.5 0.42 0.83 47 0.06 8 0.54 1.13 1.92 10.7 6 0.51 1.14 13.5 0.01 12 0.39 0.28 47 0.04 10 0.51 0.79 1.60 8.34 8 0.40 1.74 17.6 0.03 16 0.41 0.51 11.6 0.04 12.5 0.43 0.77 1.26 3.54 11.5 0.56 1.25 14.5 0.05 6.25 20 0.33 1.16 19.1 0.04 15.5 0.46 0.65 1.47 5.5 14 0.55 1.26 25.6 0.25 3.55 :3.5 0.33 0.56 30.3 0.10 18.5 0.46 1.27 1.82 14.6 18 0.75 13.6 0.22 26 0.42 1.33 30.3 0.05 21.5 0.36 1.58 1.52 44.4 22 0.42 1.28 23.5 28 0.25 0.36 30.3 0.06 24.5 0.33 0.78 1.04 6.1 26 0.34 1.62 54.8 0.27 31 0.24 1.26 30.3 0.09 27.5 0.29 0.70 0.94 22 29.5 0.39 1.32 59 0.27 0.85 :4.5 0.09 2.88 8.8 0.16 30.5 0.25 0.22 0.63 12.0 32.5 0.27 2.51 90.1 0.17 1.08 :7.5 0.09 2.33 6.14 0.22 33.5 0.10 2.48 3.27 15.4 35.5 0.28 3.45 75.2 0.25 1.12 ·0.5 0.08 2.27 15 0.12 36.5 0.09 2.43 2.26 44.4 38.5 0.21 2.10 18.3 0.32 1.19 .3.5 0.08 2.3 15 0.13 39.5 0.07 2.68 2_96 49.0 41.5 0.15 2.12 22.6 0.15 0.31 ·6.5 0.08 2.43 15 1.08 42.5 0.08 2.89 3.15 31.6 45 0.19 2.26 22.6 0.10 ----- 49 0.08 2.97 50.3 0.07 2002 epth C inorQ C oro FeCHCIl AVS FeS2 Hg-py MeHg. cm) (%wt) ~%wt) (%wt) (IJmol g-') (IJmol g-') (ng g-' py) (ng Hg g-') 1.25 0.10 1.85 3.62 1.71 0.70 72 4.42 1.75 0.10 1.95 4.27 1.74 0.82 107 4.9 1.5 0.11 2.2 3.92 1.66 0.96 67 0.06 2.5 0.18 2.02 3.63 6.82 1.44 61 3.05 3.5 0.38 1.02 2.51 5.08 2.î7 43 4.61 4.5 0.54 1.4 3.49 8.30 2.35 58 2.64 6 0.53 1.24 2.94 14.2 2.50 138 5.82 B.5 0.47 0.83 2.53 18.0 2.04 49 6.09 12 0.51 1.12 2.98 18.7 2.43 76 2.83 16 0.47 1.23 2.61 36.6 2.41 58 8.28 9.5 0.42 1.09 2.45 39.1 2.26 138 3.86 :2.5 0.43 1.27 2.41 61.8 2.23 49 3.21 :5.5 0.37 2.11 2.23 47.1 2.91 76 2.09 :8.5 0.35 1.25 1.90 34.0 2.51 58 0.8 Il.5 0.29 4.1 1.68 106 4.60 42 0.42 14.5 0.38 5.7 1.90 65.1 4.16 70 7.01 17.5 0.27 4.29 1.59 62.8 4.02 46 7.01 10.5 0.27 3.18 2.00 60.9 3.52 59 7.01 44 0.19 0.39 1.19 30.8 1.56 97 5.22 48 0.11 2.41 3.60 41.8 2.94 113 0.08 78 'j \ \ ! Appendix 2.3: SAG-05 (North Arm Saguenay Fjord) 1991 1 1996 Depth Cinara Cora FeCHCO AVS FeS2 1 Depth Cinara Cora FeS2 FeCHCI) (cm) (%wt) (%wt) (%wt) (IJmol g"') (IJmol g"') (cm) (%wt) (%wt) (IJmol g"') (%wt) 0.25 0.07 1.97 2.22 2.7 18.7 0.75 0.07 2.34 2.98 5.6 5.4 1.5 0.07 2.30 2.34 7.3 16.2 2.5 0.07 1.84 1.92 20.1 56.3 3.5 0.08 1.65 1.83 46.5 29.2 4.5 0.09 1.78 1.85 121 23.3 6 0.09 2.02 1.91 147 8 0.08 2.81 2.06 125 65.7 10 0.10 2.60 1.98 87.8 85.4 12.5 0.08 2.15 1.92 136 34.7 15.5 0.07 2.13 2.00 152 21.3 18.5 0.24 0.79 2.07 24.4 66 22 0.66 0.20 2.29 6.8 33.7 27 0.66 0.16 2.34 6.9 45.7 33 0.72 0.13 2.28 16.6 41.2 39 0.05 3.69 2.25 121 30.6 50.5 0.09 2.27 29.6 79 ) ') Appendix 2.3: SAG-05 (North Arm Saguenay Fjord) 1997 1998 Depth Cino"" CO"" FeS2 Fe(HCIl Depth Cino"" CO"" AVS FeS2 Fe(HCI) (cm) (%wt) (%wt) (IJmol g.') (%wt) (cm) (%wt) (%wt) (IJmol gO,) (IJmol gO') (%wt) 0.25 0.09 1.24 6.15 1.38 0.25 0.07 1.76 0.77 5.95 1.79 0.75 0.08 1.66 5.72 1.79 0.75 0.07 1.80 0.76 5.99 1.93 1.6 0.09 1.53 5.94 1.63 1.25 0.06 1.37 0.81 5.74 1.53 2.7 0.05 1.06 7.34 1.27 1.75 0.01 1.69 0.84 6.00 1.68 3.7 0.22 0.72 8.95 1.47 2.75 0.10 1.61 0.53 6.40 1.70 4.7 0.21 1.01 8.57 1.50 4 0.13 1.07 3.68 7.26 1.41 5.7 0.22 1.30 8.21 1.65 5 0.19 0.79 2.02 7.45 1.45 7.2 0.22 1.36 7.03 1.93 6 0.20 1.15 2.95 8.50 1.45 9.2 0.20 1.36 6.27 1.70 7 0.18 1.40 2.86 8.67 1.52 11.2 0.18 1.57 5.39 1.45 8 0.21 1.36 2.90 6.53 1.73 13.2 0.18 3.14 5.51 1.71 9 0.23 1.35 2.61 6.75 1.84 15.2 0.17 1.89 4.46 1.35 10 0.22 1.46 3.29 6.25 1.77 19.2 0.05 2.28 5.88 2.03 11.5 0.20 1.91 5.60 5.67 1.91 23.2 0.05 2.60 6.58 1.78 13.5 0.21 2.09 6.29 6.12 1.74 25.2 0.05 2.18 7.65 1.81 15.5 0.23 1.93 4.26 5.14 1.78 27.2 0.06 2.36 21.7 1.88 17.5 0.33 0.75 2.24 3.81 2.18 32.2 0.08 2.93 26.5 1.81 19.5 0.27 1.00 19.1 3.66 1.87 37.2 0.09 2.67 9.65 2.01 23 0.09 2.20 22.3 6.57 1.81 28 0.08 2.48 109 12.6 1.93 33 0.08 2.55 112 19.9 1.79 80 ) 'j 'J pendix 2.3: SAG-05 (North Arm Saguenay Fjord) 9 (Ma~19) 1999 (May 20) Depth Cinol"!! COI"!! AVS FeS2 FeCHeil Depth Cinol"!! COI"!! MeHg. (cm) (%wt) (%wt) (J.lmol g") (J.lmol g") (%wt) (cm) (%wt) (%wt) (ng Hg g") 0.25 0.06 1.61 1.08 4.08 1.65 0.25 0.06 1.64 0.12 0.75 0.06 1.67 1.03 4.55 1.78 0.75 0.06 1.50 0.17 1.5 0.07 1.53 1.34 5.18 1.77 1.5 0.07 1.68 0.13 2.5 0.10 1.41 1.58 5.76 1.69 2.5 0.08 1.43 0.06 3.5 0.14 1.38 2.84 7.33 1.64 3.5 0.15 1.22 0.18 4.5 0.16 1.37 2.57 6.88 1.60 4.5 0.17 1.28 0.16 6 0.17 1.36 4.83 5.93 1.83 6 0.18 1.29 0.13 8 0.16 1.42 5.16 6.47 1.71 8 0.20 1.33 0.12 10 0.16 1.84 9.73 5.05 1.75 10 0.16 1.78 0.11 12.5 0.19 1.49 5.67 4.56 1.89 12.5 0.21 1.72 0.09 15.5 0.16 1.36 21.1 3.86 1.91 15.5 0.30 1.68 0.08 18.5 0.07 2.11 30.4 5.26 1.81 18.5 0.06 1.98 0.56 21 0.06 2.23 31.7 6.43 1.81 21 0.05 2.35 0.06 23 0.07 2.18 74.1 8.65 1.98 23 0.06 2.21 0.08 25 0.07 1.93 115 22.8 1.70 25 0.05 2.36 0.22 27.5 0.05 2.29 95.7 18.6 1.63 27.5 0.07 2.29 0.18 30.5 0.06 2.31 212 21.0 1.83 30.5 0.08 2.19 0.06 33.5 0.05 3.10 267 27.7 1.95 33.5 0.05 2.52 0.60 36.5 0.05 2.74 155 21.6 1.98 36.5 0.08 2.18 0.57 39.5 0.06 2.88 0.58 43 0.06 3.12 0.15 47 0.07 3.17 0.07 51 0.10 2.36 0.06 81 ) 'î ") Appendix 2.3: SAG-05 (North Arm Saguenay Fjord) 2000 ! 2001 Depth CinorQ COrQ AVS Fe(HCll Depth CinorQ COrQ AVS Fe(HCll Me Hg. MeHgd Depth (cm) (%wt) (%wt) (,",moi gO,) (%wt). (cm) (%wt) (%wt) (,",moi gO') (%wt) (ng Hg gO') (ng Hg L-') (cm) 0.25 0.07 1.75 1.50 2.05 0.25 0.06 1.61 1.83 2.16 0.19 5.39 0.25 0.75 0.08 1.67 1.38 2.04 0.75 0.06 1.75 1.10 2.44 0.06 14.3 0.75 1.5 0.08 1.67 1.52 2.03 1.5 0.Q7 1.71 1.00 2.43 0.03 12.0 1.5 2.5 0.08 1.58 1.75 2.08 2.5 0.07 1.64 1.03 2.31 0.04 16.2 2.5 4 0.10 1.32 1.58 2.00 3.5 0.08 1.65 1.29 1.99 0.07 44.6 3.5 6 0.14 1.31 2.27 1.93 4.5 0.11 1.49 1.71 2.20 0.10 2.3 4.5 8 0.18 1.44 4.79 2.01 6 0.09 1.44 7.35 2.33 0.04 3.61 6 10 0.19 1.38 10.7 2.12 8.5 0.15 1.46 9.79 2.28 0.11 14.8 8.5 12 0.20 1.84 17.0 2.23 11.5 0.17 1.58 8.97 2.48 0.08 9.9 11.5 14.5 0.23 1.81 18.5 2.29 15 0.20 2.06 16.8 2.31 0.09 9.74 15 17.5 0.33 0.97 9.15 2.70 19 0.06 2.21 27.7 2.33 0.15 5.57 19 20.5 0.08 1.94 42.7 2.17 23 0.07 1.74 280 2.11 0.26 5.20 23 23 0.07 2.41 38.2 2.15 26 0.05 2.87 207 2.33 1.74 27 25.5 0.08 1.91 86.0 1.93 28 0.07 2.32 293 2.21 0.27 4.29 31 28.5 0.07 2.11 270 2.13 30 0.06 2.76 175 2.25 0.27 1.39 35 31.5 0.07 2.53 262 2.29 32 0.05 2.98 90 2.10 0.23 0.88 39 35 0.07 2.73 429 2.10 34 0.05 3.32 90 2.26 0.17 0.55 43 39 0.07 2.09 207 2.16 36 0.05 3.23 238 2.35 0.05 42.5 0.07 3.13 267 2.09 38 0.05 3.42 129 2.35 0.02 40 0.08 2.53 190 2.03 0.11 42 0.08 3.42 163 2.40 0.14 li nnR ? "!.4 ? 1? nn"! 1 .. 82 \ ) '1 ') Appendix 2.3: SAG-05 (North Arm Saguenay Fjord) 2002 Depth CinOIll COIll Fe(Hel) AVS FeS2 Hg-py (cm) (%wt) (%wt) (%wt) (IJmol g-') (IJmol g-') (ng g-' py) 0.25 0.10 1.82 2.98 1.25 1.26 43 0.75 0.11 1.74 2.98 0.92 1.18 45 1.5 0.06 1.9 2.87 0.87 1.40 38 2.5 0.06 1.7 2.83 0.78 1.15 29 3.5 0.06 1.64 2.63 1.17 1.88 60 4.5 0.08 1.5 2.69 1.62 1.32 52 6 0.11 1.53 4.73 1.37 40 8.5 0.13 1.5 2.79 8.34 1.78 28 11.5 0.16 1.68 2.79 13.4 1.48 34 14.5 0.09 1.73 2.84 34.9 1.36 73 17.5 0.07 2.02 2.79 29.1 1.61 93 20.5 0.06 2.28 2.79 32.2 1.98 153 24 0.05 2.36 2.87 367 2.41 183 28.0 0.08 2.14 2.79 134 4.90 322 31.5 0.06 2.86 2.74 146 6.98 211 34.5 3.21 2.82 207 6.42 303 37 0.06 3.25 2.70 288 5.60 254 39 0.07 2.87 2.83 243 4.91 228 41 0.08 2.34 2.48 217 5.72 165 43 0.09 3.3 2.71 268 4.36 378 45 0.08 2.25 2.49 4.55 272 -- --_.. _------ 83