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Mercury Speciation and Distribution in Coastal Wetlands
and Tidal Mudflats: Relationships with Sulphur Speciation
and Organic Carbon
Nelson J. O’Driscoll &João Canário &
Nathan Crowell &Tim Webster
Received: 13 September 2010 / Accepted: 21 January 2011 / Published online: 12 February 2011
#Springer Science+Business Media B.V. 2011
Abstract Sediment cores were analysed from four
coastal wetland sites within the Minas Basin, Bay
of Fundy to compare mercury speciation and
sediment characteristics. The coastal wetland sedi-
ments were low in total mercury (mean=17.4±
9.9ngg
−1
); however, MeHg concentration was 92
times higher (mean of 249 pg g
−1
) than intertidal
mudflat sediment (mean of 2.7 pg g
−1
). Total
mercury concentrations in intertidal mudflat cores
were also low (0.5–23.7 ng g
−1
) and correlated
(Pearson correlation=0.98; p<0.01) with % organic
carbon; with low concentrations of MeHg present
only below depths of 6 cm (mean=2.7± 1.0 pg g
−1
).
Total mercury concentrations were negatively corre-
lated (correlation=0.56, p<0.05) with inorganic
sulphur (acid volatile sulphides (AVS) and pyrite)
while MeHg concentrations were inversely correlat-
ed (Pearson correlation = −0.68; p<0.05) with the
pyrite content but not with AVS. Methyl mercury
concentrations were not significantly correlated with
organic carbon content in the wetland sediments, and
mercury-in-biomass enrichment factors were lower
(totalmercurymean1.5±1.9andMeHgmean=3.6±
4.8) than published measurements from mercury
polluted sites. Modelling estimates found on average
4.4 times more total mercury mass in the intertidal
mudflat sediments relative to vegetated wetlands. A
negative relationship was observed between MeHg
concentrations (below 20 cm depth) and modelled
tidal inundation. The mineral fraction within wetland
sediments contained 96.2% of the total mercury
mass; however, the highest concentrations of mercu-
ry species were in root biomass. This research
confirms that vegetated coastal wetlands are key
areas for formation of bioavailable methyl mercury,
andmercurydistributionistiedtoorganiccarbon
and sulphur speciation.
Keywords Mercury .Methyl mercury .Sulphur .
Coastal wetlands .Salt marsh .Sediments .Mudflats .
Bay of Fundy
Water Air Soil Pollut (2011) 220:313–326
DOI 10.1007/s11270-011-0756-2
N. J. O’Driscoll
Department of Earth & Environmental Science,
Acadia University,
Wolfville, NS, Canada
N. J. O’Driscoll (*)
K.C. Irving Environmental Science Centre,
Room LL33 K.C,
Wolfville, NS, Canada B4P 2R6
e-mail: nelson.odriscoll@acadiau.ca
J. Canário
IPIMAR/INRB IP,
Av. Brasilia,
1449006 Lisbon, Portugal
N. Crowell :T. Webster
Applied Geomatics Research Group (AGRG),
Centre of Geographic Sciences (COGS),
50 Elliott Road, RR#1,
Lawrencetown, NS, Canada B0S 1M0
1 Introduction
Mercury is a globally dispersed pollutant which exists
in several different forms in aquatic ecosystems
(O’Driscoll et al. 2005). One of the most toxic forms
is methyl mercury (MeHg), a neurotoxin which may
bioaccumulate in aquatic organisms and biomagnify
in the food web to concentrations which result in
adverse effects on reproduction and neural develop-
ment in fish and mammals (Wang et al.1998; Burgess
and Meyer 2008; Drevnick and Sandheinrich 2003).
Anoxic wetland sediments are known to be key areas
of mercury methylation in freshwater ecosystems
(O’Driscoll et al. 2005b) due to the presence of
sulphate reducing bacteria and/or other mercury
methylating microorganisms (Compeau and Bartha
1985; Fleming et al. 2006). While mercury speciation
has been extensively studied in freshwater wetlands,
much less is known about the distribution and
speciation of mercury in coastal wetlands.
The Minas Basin, Nova Scotia is home to the
highest tidal amplitude in the world (with average
tidal amplitudes ranging between 13 and 16 m) and
therefore has a significant intertidal area within
Southern Bight (163.84 km
2
using tidal range of
−7.7 to 7.3 m based on current light detection and
ranging (LiDAR) survey; Wells et al. 2004; BOFEP
2008). The Minas Basin mudflats are highly produc-
tive areas that supply a large amount of biomass in the
form of mud shrimp (Corophium volutator)and
polychaete worms to migratory birds (Wells et al.
2004) and thus the potential for mercury accumula-
tion in areas such as the Bay of Fundy. The Southern
Bight portion of the Minas Basin (Fig. 1)also
contains substantial areas of intertidal coastal wetland
(14.84 km
2
based on a current LiDAR survey)
vegetated primarily with salt tolerant grasses (Spar-
tina alterniflora and Spartina patens; Hatcher et al.
1981). Several researchers have shown that mercury
concentrations and distribution in sediment and at the
base of the food chain (e.g. algae and invertebrates)
are key to modelling mercury fate in freshwater food
webs (Becker and Bigham 1995; Pickhardt et al.
2005). However, there are few data available for
methyl mercury concentrations in coastal wetlands.
While there is little research on mercury in
unpolluted coastal areas, the available data from
contaminated sites in Portugal suggest that mercury
methylation may be particularly significant in areas
colonized by vegetation. Canário et al.(2007a)
working in coastal sediments in three Portuguese salt
marshes with different degree of Hg contamination
found that sediments colonized by vegetation had 70
times more MeHg (up to 18% of the total mercury
concentration) than sediments without vegetation. The
authors found enrichment factors of nine times for
total mercury and 44 times for MeHg in belowground
biomass compared to non-vegetated sediment. They
also measured enrichment factors of 400 for total
mercury and 4,700 for MeHg in the belowground
biomass compared to surface vegetation, suggesting
low translocation of mercury species from roots to
surface vegetation.
The increase of MeHg concentration in salt marsh
sediments may be also related to sulphur chemistry.
Salt marsh plants release oxidants through their roots
into anoxic sediments (Hines et al. 1989; Luther et al.
1991) creating microenvironments with different
chemical properties than the bulk sediment (Vale et
al. 1990; Hines 1991). This process changes the
sulphur fractionation in the solids and speciation in
pore waters (Lord and Church 1983) involving the
anaerobic bacterial reduction of sulphate to hydrogen
sulphide and aerobic reoxidation of H
2
StoSO
42−
(Luther and Church 1992; Thamdrup et al. 1994).
These processes are particularly key in coastal wet-
lands due to the frequent input of sulphate during tidal
inundation and the presence of plant respiration.
These processes affecting sulphur speciation may also
affect mercury partitioning and speciation, particular-
ly MeHg which is formed by sulphate reducing
bacteria during sulphate reduction to sulphide
(Madureira et al. 1997; Ullrich et al. 2001).
Another factor that has been shown to be critical to
mercury speciation in surface intertidal sediments is
the exposure to solar radiation. Research on contam-
inated intertidal mudflat surface sediments in Lisbon,
Portugal indicated that exposure to solar radiation
resulted in significant losses of total mercury (39–
43% of total mercury) over a 6–8-h period (Canário
and Vale 2004). The authors attributed the loss to
photo-reduction of inorganic mercury and volatiliza-
tion similar to what has been observed in open ocean
waters from the North Atlantic (Qureshi et al. 2010).
The purpose of this research is to quantify the
vertical distribution of total mercury and methyl
mercury species in vegetated sediment cores from
four coastal wetlands and one intertidal mudflat in the
314 Water Air Soil Pollut (2011) 220:313–326
Minas Basin, Nova Scotia. The relationships between
mercury speciation, sediment depth, organic carbon
content and sulphur speciation are examined and
geographic information system (GIS) modelling is
used to quantify a spatially integrated calculation of
mercury mass in surface sediments. We hypothesize
that (a) sediments colonized by vegetation will be
significantly higher in total mercury and methyl
mercury concentrations than intertidal mudflats due
to higher concentrations of organic carbon-based
ligands and subsequently increased microbial methyl-
ation, (b) total mercury concentrations will be
correlated with inorganic sulphur in sediments and
sulphur speciation may control the availability of
inorganic mercury for methylation and (c) intertidal
mudflats will contain the majority of total mercury in
the Bay of Fundy due to the large intertidal area in the
ecosystem.
2 Material and Methods
2.1 Sample Collection and Preparation
Sediment cores were collected from four coastal
wetland sites (Wolfville, Hantsport, Kingsport and
Fig. 1 LiDAR digital elevation model of the Minas Basin with intertidal wetland areas highlighted in red and the locations of coastal
wetland sampling sites in Southern Bight
Water Air Soil Pollut (2011) 220:313–326 315
Windsor) as well as from one intertidal mud flat
(Kingsport) in the Minas Basin (Fig. 1) between the
dates of September 10–21, 2007. Cores were taken at
all sites near in the leading edge of the wetland ~10 m
from the vegetation line. A transect of sediment cores
was collected at low tide from the Kingsport site
along the intertidal mud flats (cores taken at 100, 300,
600 and 900 m from the marsh edge towards low tide
line; Fig. 2). A 45-cm polycarbonate core tube with a
serrated edge was pushed into the soft sediment and
dug out for removal of the sediment cores. The core
tubes were then sealed in plastic bags and stored in a
dark cooler prior to extruding and processing.
Once in the lab, the sediment cores were
sectioned into 1-cm slices for the first 10 cm, 2-
cm slices from 10 to 20 cm and 5-cm slices from
20 to 30 cm. The sediment slices were sieved in
carbon-free double deionized water (d.d. water;
Milli-Q system with UV organic carbon removal)
to separate vegetation from the non-vegetated
sediment as described by Canário et al. (2007a).
The suspended sediment was allowed to settle in
high-density polyethylene containers, collected in
polypropylene boats, weighed and dried with the
other processed samples in a clean oven at 40°C over
a 5-day period. The vegetation samples were washed
multiple times with d.d. water, weighed and dried in
a sample oven using the same procedure.
The dried sediment samples were homogenized
with a mortar and pestle that was cleaned between
samples with 20% HCl, followed by d.d. water and
ultrapure ethanol. Aboveground vegetation samples
were also collected at each salt marsh site, and these
were cut into short strips, dried and ground into a
powder. All homogenized and dry samples were
stored in 50 mL polypropylene tubes until analysis.
2.2 Analytical Methods
2.2.1 Total Mercury Analysis
All samples were analysed by thermal degradation
atomic absorbance analysis using a Nippon MA-2000
analyzer with auto sampler (USEPA 2007). Detection
limits using this technique are 0.002 ng (or
0.02 ng g
−1
for a 100-mg sample). Quality assur-
ance/quality control included standard calibration,
check standards (<5% error), reagent blanks and
triplicate sample analyses (RSD <10%, n=45) within
each sample run. Certified reference materials for fish
tissue (DORM-2, DOLT-3) and marine sediment
(MESS-3) were obtained from the National Research
Council of Canada and recoveries were within 1
standard deviation of certified values. Internal stan-
dard reference materials for crushed geological
samples (QUA-1) were analysed at Acadia and were
within 1 standard deviation of accepted mean values
(Stanley et al. 2010).
2.2.2 Methyl Mercury in Sediments
Methyl mercury was determined in dry sediments by
alkaline digestion (KOH/MeOH), organic extraction
with dichloromethane (DCM) pre-concentration in
aqueous sulphide solution, back-extraction into DCM
and quantification by GC-AFS using an Agilent
Chromatograph coupled with a pyrolizer unit and
a PSA fluorescence detector (Canário et al. 2004).
Recoveries and the possible MeHg artefact forma-
tion were evaluated by spiking several samples with
Hg(II) and MeHg standard solutions with different
concentrations. Recoveries varied between 97% and
103%, and no artefact MeHg formation was
observed during the procedure. Precision, expressed
as relative standard deviation of four replicate
samples, was less than 4% (p<0.05). International
certified standards (IAEA-405 and BCR-580) were
used to ensure the accuracy of the procedure, and
obtained and certified values were not statistically
different (p<0.05).
2.2.3 Methyl Mercury in Vegetation
For MeHg determinations in vegetation, a modified
method based on Canário et al. (2006) was used.
Dried tissues were digested with a concentrated HBr
(Merck suprapur) solution saturated with CuSO
4
.
Methyl mercury in the digested solution was then
extracted in a dichloromethane (DCM) solution pre-
concentrated in a slight alkaline H
2
S solution, back
extracted into DCM and quantified by GC-AFS in an
Agilent chromatograph coupled with a pyrolysis unit
and a PSA fluorescence detector. Sediment analysis
recoveries and MeHg artefact formation were evalu-
ated by spiking several samples with Hg(II) and
MeHg standard solutions with different concentra-
tions. Recoveries varied between 96% and 104% for
all plant tissues investigated and no artefact MeHg
316 Water Air Soil Pollut (2011) 220:313–326
formation was observed. In all the analysis, precision
was expressed as relative standard deviation of three
replicate samples and was less than 2.5% (p< 0.05).
International certified standard AIEA-140/TM (Fucus
sea plant homogenate) was used to ensure the
accuracy of the procedure, and measured Hg and
MeHg concentrations were not statistically different
(p<0.05) from the certified values.
2.2.4 Total Al, Si, Fe and Mn
Total determinations of Al, Si, Fe and Mn were
performed by mineralization of the sediment samples
with a mixture of acids (HF, HNO
3
and HCl)
according to the method described by Rantala and
Loring (1975). Metal concentrations were obtained by
flame-AAS (Perkin-Elmer AAnalyst 100) using direct
aspiration into N
2
O-acetylene flame (Al and Si) or
air-acetylene flame (Fe and Mn). International certi-
fied standards (MESS-3, PACS-2, BCSS-1) from
NRCC-Canada were used to ensure the accuracy of
our procedure.
2.2.5 Total Sulphur and Total Carbon
Total sulphur and carbon sediment content were
measured in homogenized and dried sediments, using
a CHN Fissons NA 1500 Analyzer, calibrated with
sulphanilamide standards. Procedural blanks were
obtained by running several empty ashed tin capsules.
Organic carbon was estimated by difference between
total carbon and inorganic carbon after heating
samples at 450°C for 2 h in order to remove the
organic carbon from the sediment.
2.2.6 Sulphur Speciation
Total inorganic sulphur was determined in 100-mg
sediment sample using the chromium reduced
sulphur (CRS) method described by Canfield et
al. (1986). This method is specific for inorganic
sulphur (AVS +FeS
2
+S
0
)withanaccuracythatisnot
affected by the presence of organic sulphur. After
determination of the total inorganic sulphur, inor-
ganic sulphur species were determined individually.
Before analysis, elemental sulphur (S
0
) was extracted
from100mgofsedimentsampleby16-hstirring
with 20 cm
3
of acetone followed by centrifugation at
3,000 rpm for 10 min and filtration through 0.45-μm
Millipore® membranes. The residue was then placed
in the reaction vessel with 10 cm
3
HCl 1 M and
purged with N
2
for 20 min to release AVS. Finally,
the CRS method was used in the remaining residue
to analyse the pyrite content. Elemental sulphur was
determined using the same CRS method in the
acetone extracts. The measurements of the released
H
2
S were made by differential pulse polarography
(DPP) using a Metrohm apparatus equipped with a
693VAProcessoranda694VAStand.Inanother
aliquot, acid volatile sulphides (AVS; mainly amor-
phous iron sulphides and poorly crystallized Fe
oxides) were extracted with 1 M HCl (Henneke et
al. 1991; Luther et al. 1985). Sulphide was trapped in
a de-aerated NaOH solution and analysed by DPP.
Recovery of standard sulphide solutions in all
analysed species was 97%. The detection limit for
sulphide was 0.01 μmol g
−1
, and error was less than
5%. The sum of the individual sulphur species was
not statistically different (p<0.05) from the analysed
total inorganic sulphur. Total organic sulphur was
determined for each sample by subtracting the total
inorganic sulphur concentrations from the total
sulphur content (Likens et al. 2002).
2.3 Inundation Modelling
A GIS approach was used to determine both (a) the
intertidal zone of the Southern Bight study area and
(b) the area of coastal wetland which was within the
established intertidal zone. In order to model the
intertidal zone, a raster-based high-resolution digital
elevation model (DEM) was required to accurately
establish terrain elevations which could be compared
against tidal elevations. The DEM was constructed
using the results of a LiDAR survey which was
performed by the Applied Geomatics Research Group
(Middleton, Nova Scotia) over the Southern Bight
study area. The survey was comprised of two
missions. The first mission used a Mark I LiDAR
system from Terra Remote Sensing Inc. to measure
the western block (Blomidon, Wolfville) in 2003. The
second mission used an ALTM 3100 LiDAR system
to measure the eastern block (Hantsport and Windsor)
in 2007. The survey recorded elevation data spanning
10.04× 10
3
km
2
(30 cm vertical accuracy). Point
elevation data were classified into ground hits, using
a ground/non-ground separation algorithm within the
TerraScan software suite. Ground classified LiDAR
Water Air Soil Pollut (2011) 220:313–326 317
points were interpolated to construct the DEM of the
study area (4 m
2
ground resolution).
Tidal data were retrieved from hourly predic-
tions for 2009 (from Flater and Pentcheff 2008)for
Hantsport, Nova Scotia. Tidal water elevations were
converted from local chart datum (Hantsport) to
orthometric heights above Canadian Geodetic Vertical
Datum 1928 (CGVD28) to be in a common with
DEM elevation. Converted predictions were used
to establish a tidal range (−7.3to7.7m),which
was divided into 0.10 m increments for inundation
modelling. The model was run within the ArcGIS
interface using a Visual Basic script developed by
AGRG to simulate water distribution over costal
terrain based on elevation differences. Terrain cells
were flooded if they met two criteria: (a) The cell
was of a lower elevation than the modelled tidal
elevation and (b) the cell was connected to the
source of inundation by a series of cells which met
the first criterion. Tidal elevations were modelled
at 0.10 m elevation increments for the full 2009
tidal range (15.0 m, n= 150). The result was an
accurate depiction of the intertidal zone within the
Southern Bight study area.
Wetland coverage data were retrieved from the
Nova Scotia Department of Natural Resources
provincial forestry layer. Wetland areas were
classified as intertidal if their spatial extent was
within the previously described intertidal zone.
Corresponding areas of intertidal wetland were
calculated using simple GIS techniques within
ArcGIS.
3 Results and Discussion
3.1 Mercury Speciation in Intertidal Sediment
The non-vegetated intertidal mudflat cores taken near
Kingsport, Nova Scotia were found to have total
mercury concentrations ranging from 0.5 to
23.7 ng g
−1
in dry sediment (Fig. 2). This is lower
than the data published by Leermakers et al. (1993)
who measured total mercury concentrations ranging
between 30 and 1,756 ng g
−1
intertidal mudflat
sediments taken from the Scheldt estuary. Methyl
mercury in the non-vegetated intertidal mudflat cores
was very low in concentration generally <2.0 pg g
−1
below a depth of 6 cm in all cores. Above 6-cm depth,
there was little variation in MeHg concentrations
(mean=2.7 pg g
−1
; standard deviation=1.0 pg g
−1
).
These total mercury data are slightly lower but in
the same range as total mercury in surface sediment
on the North Shore of the Bay of Fundy ranging from
7to79ngg
−1
as measured by Hung and Chmura
(2006). Leermakers et al. (1993) found that total
mercury concentrations in the intertidal mudflats of
the Scheldt estuary were strongly correlated with
organic matter content and sulphide concentrations.
They also observed in incubation experiments that
~1% to 2% of mercury was converted to methyl
mercury. This is much larger than the 0.008% to
0.14% of total mercury present as methyl mercury in
the samples in this study. The results found in the
Minas Basin results correspond with Canário et al.
(2005a) who observed that MeHg accounted for
0.02% to 0.4% of total mercury in intertidal mudflat
sediments in the Tagus estuary in Portugal. Similarly,
in Vigo Ria in northwest Spain MeHg was found to
range between 0.01% and 0.1% of total mercury
concentration (Canário et al. 2007b). These results
suggest there are similar and low rates of mercury
methylation in the intertidal mudflat areas studied in
Portugal, Vigo Ria and the Minas Basin. The low
concentrations of MeHg in the Minas Basin are likely
due to both the low levels of total mercury present
and a low methylation rate. The low levels of total
mercury and MeHg may be a result of the low organic
carbon content (0.44–4.66% organic carbon) and low
root biomass observed in the intertidal mudflat
samples as opposed to coastal wetland sites with
Fig. 2 Total mercury concentration with depth in intertidal
mudflat sediments. Line represents spline curve fit to data.
Graph shows decreasing total mercury in upper 10 cm of
sediment with increasing distance from shoreline
318 Water Air Soil Pollut (2011) 220:313–326
large amounts of root biomass. This is supported by
Canário et al. (2007a) who found that coastal wetland
sediments colonized by vegetation contained 70 times
more methyl mercury than non-vegetated sediments
in three Portuguese estuaries. The authors suggested
that the network of roots, distributed heterogeneously
throughout the sediment, creates a correspondingly
heterogeneous pattern of sediment redox properties.
This shifting of redox conditions appears to be ideal
for high microbial activity, which when coupled with
the reservoirs for bacterial respiration (e.g. organic
carbon and sulphate) favours Hg methylation.
While total mercury and MeHg concentrations
were low in the intertidal mudflat sediments, it was
observed that there was a general decrease in total
mercury in surface sediment (0–5 cm) from 9.8 to
1.4 ng g
−1
(n=20) with increasing distance (100–
900 m) from the coastal wetland (Figs. 2and 3). With
increasing distance from the shoreline, the mudflat
sediments are increasingly covered by water and have
less exposure to solar radiation (Fig. 3). Canário and
Vale (2004) found that exposure of intertidal mudflat
sediment to solar radiation resulted in significant
losses of total mercury (39–43% of total mercury)
over a 6–8-h period. The primary loss mechanism was
assumed to be mercury release to pore water followed
by photo-reduction and volatilization to the air. The
opposite trend is observed in this field study where
increased total mercury in sediment with increased
solar exposure is observed. The mercury available at
each sediment site may relate to the amount of
organic carbon available to retain mercury in the
sediments (decreasing from 4.66% to 0.44% between
100 and 900 m distance from the coastal wetland;
Table 1). Mean total mercury in the top 15 cm of the
intertidal mudflats was found to be highly correlated
with percentage organic carbon content (Pearson
correlation=0.98; p<0.01). Since organic carbon is a
key ligand for mercury species (Ravichandran 2004),
this may explain the increased mercury retention
closer to the coastal wetland.
While grain size distribution was not determined in
this study, Si concentrations normalized to Al content
(Si/Al ratios) are a good proxy of the grain size
distribution (Loring 1991). Higher Si/Al ratios indi-
cate the presence of coarser material (enriched in
Fig. 3 Location of sampling plots (yellow dots) for intertidal sediment cores at Kingsport, Nova Scotia and the elevation profile of site
for CGVD28 (Canadian Geodetic Vertical Datum, 1928)
Water Air Soil Pollut (2011) 220:313–326 319
silica) while lower Si/Al ratios indicate the presence
of finer material mainly composed of aluminosilicates
with a higher potential for trace metal and mercury
binding. In this research, a small increase in the
Si/Al rations is observed from 2.63 at 100 m to
3.21 at 900 m indicating the potential for increas-
ing size of particles with distance from the
shoreline. As such, this may suggest that a portion
of the observed decrease in mercury concentration
from100to900mfromtheshorelinemaybe
attributed to difference in grain size distribution.
However, the small difference in the Si/Al ratios
suggests that this may be a minor factor affecting
mercury distribution.
3.2 Mercury Speciation in Coastal Wetland
Non-vegetated Sediment
In comparison with intertidal sediment samples, the
coastal wetland mineral sediment samples displayed a
similar range of total mercury concentrations ranging
from 1.2 to 50 ng g
−1
in dried sediment (mean=
17.4 ng g
−1
; standard deviation=9.9 ng g
−1
;n=95)
and slightly lower organic carbon content (0.29–1.39%
organic carbon) likely due to the removal of all root
biomass (Fig. 4; Table 1). The MeHg concentration in
the mineral sediment portion of the coastal wetland
ranged between 14 and 715 pg g
−1
(mean= 249 pg g
−1
;
standard deviation= 203 pg g
−1
;n=70);thisison
average 92 times higher in concentration than MeHg in
the non-vegetated intertidal sediment. The percentage
of total mercury present as methyl mercury in the mineral
portion of wetland sediment cores ranged between 0%
and 2.9% (mean= 0.9; standard deviation= 0.8; n=70)
Al (%) Si/Al C
org
(%) Fe/Al Mn/Al
Kingsport (mudflats
100 m)
9.6± 1.3 2.43 ±1.21 4.66± 1.81 0.59 ±0.12 63± 9
Kingsport (mudflats
300 m)
8.6± 2.1 2.63 ±1.34 3.43± 1.91 0.63 ±0.21 58± 5
Kingsport (mudflats
600 m)
7.6± 1.1 3.02 ±2.02 1.05± 1.28 0.61 ±0.18 62± 7
Kingsport (mudflats
900 m)
8.2± 2.0 3.87 ±1.38 0.44± 0.29 0.59 ±0.17 60± 8
Windsor (wetland) 6.3± 1.3 2.87 ±1.12 0.89± 0.08 0.94 ±0.54 71± 4
Wolfville (wetland) 6.8± 2.3 3.01 ±1.12 0.59± 0.06 1.02 ±0.21 59± 10
Hantsport (wetland) 7.1 ± 1.4 2.54±0.21 0.29 ± 0.22 1.00±0.19 73 ± 12
Kingsport (wetland) 5.5±0.8 4.21 ± 0.34 1.39±0.33 0.89 ± 0.12 67±5
Kingsport (wetland) 6.3±0.4 3.98 ± 0.87 0.91±0.06 1.07 ± 0.14 74±9
Table 1 Range of Al con-
centrations and Si/Al, Fe/Al
and Mn/Al ratios in the
solid non-vegetated fraction
of the sediments (n=8 for
each site)
Note that coastal wetland
sites had all vegetation re-
moved prior to analyses
Fig. 4 Sediment profiles showing total mercury and methyl
mercury concentrations in both mineral sediment (closed
markers) and vegetation samples (open markers) for each
sampling site with line representing spline curve
320 Water Air Soil Pollut (2011) 220:313–326
which is similar to that observed by Canário et al.
(2007a).
It was observed that the Windsor coastal wetland
had substantially higher concentrations in the top 0–
15 cm of mineral sediment than any of the other
wetlands sampled (ranging from 142 to 716 pg g
−1
in
Windsor and 14 to 269 pg g
−1
at all other sites;
Fig. 5). However, below 15 cm depth, concentrations
in the Windsor coastal wetland decrease to levels
similar to that observed at all other sites (Figs. 4and
5). No significant correlations were observed with
organic carbon content of the mineral sediment at the
wetland sites. However, below 20 cm depth, concen-
trations were found to be linked with elevation
(ranging from 125 pg g
−1
at 4.94 m elevation in
Kingsport wetland near water to 498 pg g
−1
at 5.79 m
elevation in Windsor wetland near upland; Fig. 3).
This trend became more prominent as depth increased
to 30 cm where Kingsport near upland (6.44 m
elevation) was found to be the most heavily concen-
trated site (521 pg g
−1
). This trend may be due to
higher rates of tidal flooding resulting in mercury
dissolution and removal at low elevations.
The Windsor salt marsh is unique in its develop-
ment as it was formed as a result of increased
sedimentation due to the construction of a causeway
across the Avon River in Nova Scotia in 1968–1970.
Colonization of the sampling site by S. alterniflora is
estimated to have begun in the early 1990s (BOFEP
2008). As such, this is a newly formed coastal
wetland (<20 years old). Recent research by St Louis
et al. 2004 suggests that newly flooded peat lands are
most efficient at mercury methylation in the first
2 years post-flooding, followed by substantial
decreases in methylation efficiency. While there are
currently no data on coastal wetland formation and
mercury methylation, this research suggests that
wetland age could be a factor in the increased MeHg
in the Windsor site.
3.3 Mercury in Root Mass and Aboveground
Vegetation in Coastal Wetlands
The concentration of total mercury in belowground
root biomass ranged between 0.7 and 223 ng g
−1
(mean=17.4 ng g
−1
; standard deviation=10.0 ng g
−1
;
n=95), and methyl mercury ranged between 0 and
957 pg g
−1
(mean=371.4 pg g
−1
; standard deviation=
184.8 pg g
−1
;n=66). A low enrichment factor for
total mercury (mean 1.5; standard deviation =1.9; n=
68) and for MeHg (mean=3.6; standard deviation=
4.8; n=66) was observed between mineral sediment
and belowground root biomass in the Minas Basin
wetlands suggesting that roots were not actively
bioaccumulating mercury. This is much less than the
nine times enrichment in concentration for total
mercury and 44 for MeHg observed by Canário et
al. (2007a) in industrially impacted sediments.
Aboveground vegetation in the sites sampled were
found to contain total mercury concentrations ranging
Fig. 5 Methyl mercury
concentrations (dried
weight) in near water and
near upland sites for both
Windsor and Kingsport
coastal wetlands
Water Air Soil Pollut (2011) 220:313–326 321
from13.7to44.9ngg
−1
(mean=23.1ngg
−1
;
standard deviation=10.7 ng g
−1
;n=8) and MeHg
concentrations ranging from detection limit to
3.2 pg g
−1
(mean=2.6 pg g
−1
; standard deviation=
0.5 pg g
−1
;n=4). This resulted in an average
enrichment factor between roots and aboveground
plants of 0.2 for total mercury.
Canário et al. (2007a) measured much higher
mercury concentrations in below and aboveground
vegetation in industrially impacted Portuguese estu-
aries. Canário et al. (2007a) found total mercury
concentrations of root biomass ranging between 170
and 13,000 ng g
−1
which was 2 orders of magnitude
higher than total mercury concentrations in above-
ground plants (18–220 ng g
−1
). Canário et al. (2007a)
also observed MeHg concentrations in root biomass
as high as 941 ng g
−1
which is 1,000 times more
concentrated than MeHg concentrations found in the
Minas Basin. This corresponded to a percentage of
total mercury as MeHg as high as 18% in vegetated
sediments in Portugal, with no significant differences
between sediments with varying plant species.
The average mass of root biomass in each core was
used to roughly estimate the total amounts of mercury
in the root biomass and non-vegetated portions of
sediments in the Southern Bight area of the Minas
Basin using Eq. 1.
MTRoots ¼MRoots DCore AWetlands
ðÞ
ACore
ð1Þ
where M
TRoots
is the total mass of roots in the top
30 cm of Southern Bight, Minas Basin sediments;
M
Roots
is the average mass of roots 1 cm section of
core (3.06 gcm
−1
; standard deviation=3.0; n=96);
D
Core
is the depth of core (30 cm); A
wetlands
is the total
area of vegetated wetlands in the Southern Bight area
of the Minas Basin (1.48×10
13
cm
2
) based on LiDAR
data (Fig. 1)andA
Core
is the surface area per core
(78.5 cm
2
).
MTRoots ¼3:06 gcm130 cm 1:48 1013 cm2
=78:5cm
2
Using this equation, the total estimated mass of
roots in the Southern Bight area of the Minas
Basin is 1.74×10
13
g (standard deviation= 2.35 ×
10
13
g). Therefore, using the mean value for mercury
in root biomass Hg
section avg
(37.3ngg
−1
;standard
deviation=41.8; n= 95), we can calculate the total
mass of mercury in root biomass in the Southern
Bight area of the Minas Basin as:
HgTRoots ¼MTRoots Hgsection avg
HgTRoots ¼6:47 1014ng
ð2Þ
Therefore, on average, there is 6.47×10
2
kg (stan-
dard deviation= 1.43 × 10
3
kg)ofmercuryinthetop
30 cm of root biomass of the Southern Bight area of
the Minas Basin based on data from the sites
sampled and the standard deviation for mercury
concentrations and root mass in the calculations.
Using similar calculations for the mineral portion of
the sediment, we calculate that there is 1.65×10
4
kg
(standard deviation= 9.40 ×10
3
kg)ofmercuryinthe
top 30 cm of mineral sediment in the wetlands of the
Southern Bight area of the Minas Basin (for a
combined mean total of 1.7×10
4
kg of mercury).
Similarly there is 7.47× 10
4
kg (standard deviation=
6.08×10
4
kg) of mercury in the top 30 cm of
intertidal mudflat sediment.
There is significantly more (on average 4.4 times
as much) mercury retained in the top 30 cm of
intertidal sediments as there is in the coastal wetland
sediments and biomass combined (ttest, p< 0.05).
This is primarily due to the large area of intertidal
sediment in the Southern Bight of the Minas Basin
(ten times the area of vegetated coastal wetland).
Within the wetland areas, the majority of the total
mercury mass is retained in the mineral sediments
(96.2% on average). However, there are higher concen-
trations in the root biomass (mean of 37 ng g
−1
;n=95)
as compared to the mineral sediment (17 ng g
−1
;n=
95). This higher concentration of mercury in roots may
be due to favourable redox conditions, binding to
organic matter in the coastal wetlands and uptake into
root material.
3.4 Sulphur Speciation and Mercury Distribution
Total sulphur concentrations and sulphur species
levels varied between intertidal mudflat and coastal
wetland sites. Concentrations of total sulphur were
higher in all coastal wetland sediments compared to
the ones in non-vegetated intertidal areas (Table 2).
These results were in line with Madureira et al. (1997)
who observed higher concentrations of sulphur
species in salt marsh sediments compared to non-
322 Water Air Soil Pollut (2011) 220:313–326
vegetated ones for the Tagus marshes (Portugal).
Sulphur speciation in vegetation-colonized areas is
influenced by the complex chemistry in the wetlands.
It is documented that plants release oxidants through
their roots into anoxic sediments (Hines et al. 1989;
Luther et al. 1991) and create microenvironments
with different chemical properties of the bulk sedi-
ment (Vale et al. 1990; Hines 1991). The oxidants
also change the sulphur fractionation in the solids and
speciation in pore waters (Lord and Church 1983).
Recent studies of sulphur transformations in sedi-
ments have emphasized that rather than a simple
cycle, composed of anaerobic bacterial reduction of
sulphate to hydrogen sulphide and aerobic reoxidation
of H
2
StoSO
42−
, products can have oxidation states
between −2 to +6 (Luther and Church 1992;
Thamdrup et al. 1994).
With the exception of the Hantsport coastal
wetland sediments, the concentrations of pyrite are
consistently higher than the AVS levels. The same
pattern was observed in non-vegetated areas of the
Minas Basin. It is well-known that pyrite formation
results from the reaction of FeS with H
2
SorS
0
in the
suboxic/anoxic environment frequently associated
with the oxidation of organic matter by sulphate
reducing bacteria (Shippers and Jorgenssen 2002;
Rickard and Morse 2005). In general, mercury species
in both dissolved and solid phases are primarily
bound to organic matter, with a minor fraction
recycled with the Mn and/or Fe oxides near the redox
boundary or co-precipitated with acid volatile sul-
phides (Drobner et al.1990; Morse and Luther 1999).
It is also well documented that the degree of trace
metal pyritization for mercury is one of the highest for
all trace metals (Morse and Luther 1999).
As such, since the Minas Basin sediments have
high mercury concentrations in areas with low
inorganic sulphur (AVS and pyrite) content, this
suggests that the accumulation of total mercury in
these sediments is negatively related to sulphur levels.
In contrast, MeHg concentrations in sediment were
found to be inversely correlated (Pearson correlation =
0.68; p<0.05) with the pyrite content but not with
AVS. Reduced sulphur in AVS is known to be redox
sensitive being easily converted to sulphate; on the
contrary, pyrite is relatively stable in anoxic/suboxic
sediments (e.g. Rickard and Morse 2005). As previ-
ously mentioned, pyrite is a good scavenger for
mercury and therefore may limit the amount of
divalent mercury available for methylation. This
may explain the inverse relationship between MeHg
concentrations and pyrite content observed in our
results.
In sediments colonized by salt marsh plants, no
significant correlations were observed between Hg,
MeHg and AVS or pyrite content. This may be
explained by the complex redox processes occurring
in the rhizosphere. The network of roots, distributed
heterogeneously throughout the sediment, creates a
correspondingly heterogeneous pattern of sediment
Table 2 Range of total, inorganic and organic sulphur concentrations as well as S
inorg
speciation in the solid fraction of the sediments
(n=4 for each site)
S
tot
(μmol g
−1
)
S
inorg
(μmol g
−1
)
S
org
*
(μmol g
−1
)
AV S
(μmol g
−1
)
Pyrite
(μmol g
−1
)
S
0
(μmol g
−1
)
Kingsport (mudflats
100 m)
46± 5 32± 3 14 8.2± 1.1 21± 6 0.71 ±0.08
Kingsport (mudflats
300 m)
113± 8 102± 8 11 46± 12 56± 12 1.40 ±0.04
Kingsport (mudflats
600 m)
221± 11 214± 10 7 75± 9.7 142 ±9 2.18±0.08
Kingsport (mudflats
900 m)
383± 9 378±7 5 65± 7.9 301± 11 6.12±0.31
Windsor (wetland) 824±23 801±13 23 231± 53 564±14 5.51± 0.25
Wolfville (wetland) 642±18 621±12 21 301± 41 331±7 7.46±0.32
Hantsport (wetland) 628± 21 620± 9 8 401±32 220 ±15 5.84±0.41
Kingsport (wetland) 423±17 421 ± 6 2 154± 19 236± 13 4.65± 0.23
Kingsport (wetland) 399±26 384 ± 7 15 124±14 168±17 2.12±0.45
S
org
*=S
tot
−S
inorg
Water Air Soil Pollut (2011) 220:313–326 323
redox properties (Sundby et al. 2003). The shifting of
redox conditions appears to be ideal for high
microbial activity coupled with the necessary ingre-
dients for bacterial respiration (e.g. organic carbon
and sulphate; Canário et al. 2005b,2003). The high
bacteria respiration greatly favours reduction and/or
oxidation of sulphur compounds and consequently
mercury partitioning and methylation (Canário et al.
2005a,2003).
4 Conclusions
In this paper, we present the first results of
mercury speciation in the Southern Bay of Fundy
and find that coastal wetlands are highly dynamic
areas in terms of mercury speciation and concen-
tration. In relation to our original hypotheses, this
work concludes the following: (a) MeHg concen-
trations in the non-vegetated portion of coastal
wetland sediment were on average 92 times higher
in concentration than intertidal mudflat sediment.
In addition, total mercury in intertidal mudflats is
highly correlated with organic carbon content;
however, methyl mercury concentrations were not
significantly correlated with organic carbon content
in the non-vegetated wetland samples. (b) Total
mercury concentrations were observed to be in-
versely correlated with inorganic sulphur (pyrite
and AVS) concentrations while MeHg concentra-
tions were inversely correlated with pyrite but not
AVS concentrations. (c) Modelling estimates of
mercury spatial distribution suggested that on
average there is 4.4 times more total mercury
mass retained in the top 30 cm of intertidal
mudflat sediments as compared to the top 30 cm
of vegetated coastal wetland areas in the Southern
Bight region of the Minas Basin, Bay of Fundy
(due to the large intertidal area). The majority of
the mercury (96.2% on average) is found to be
associated with mineral sediment in the vegetated
wetlands; however, the highest concentrations of total
mercury and methyl mercury are in root biomass which
highlights the importance of vegetation in these coastal
wetlands to mercury distribution dynamics. It was also
observed that there is a significant negative relationship
between MeHg concentrations below 20 cm depth and
modelled tidal inundation, suggesting potential effects
of solar degradation on MeHg in intertidal sediment.
Acknowledgements Dr. O’Driscoll was grateful for the
funding provided by the Canada Research Chairs program,
the Canadian Foundation for Innovation, the National
Science and Engineering Research Council and Maritime
Northeast Pipeline. Dr. Canário was thankful for the
funding provided by the Fundação para a Ciência e
Tecnologia, Portugal. Thanks are also due to Marta
Nogueira, Brendan McNeil, Nicole Oliver, Katherine Dugas
and Scott Ryan for the technical assistance in sample
collection, processing and mercury and carbon analysis.
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