potential mercury methylation rates in prairie...
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POTENTIAL MERCURY METHYLATION RATES IN PRAIRIE WETLAND
SEDIMENT
A Thesis
Submitted to the Faculty of Graduate Studies and Research
In Partial Fulfillment of the Requirements
For the Degree of
Master of Science
in
Biology
University of Regina
By
Cameron Grant John Hoggarth
Regina, Saskatchewan
July, 2013
Copyright 2013: C.G.J. Hoggarth
UNIVERSITY OF REGINA
FACULTY OF GRADUATE STUDIES AND RESEARCH
SUPERVISORY AND EXAMINING COMMITTEE
Carmeron Grant John Hoggarth, candidate for the degree of Master of Science in Biology, has presented a thesis titled, Potential Mercury Methylation Rates in Prairie Wetland Sediment, in an oral examination held on April 30, 2013. The following committee members have found the thesis acceptable in form and content, and that the candidate demonstrated satisfactory knowledge of the subject material. External Examiner: Dr. Kyle Hodder, Department of Geography
Supervisor: Dr. Britt Hall, Department of Biology
Committee Member: *Dr. Carl Mitchell, Adjunct
Committee Member: Dr. Andrew Cameron, Department of Biology
Chair of Defense: Dr. Karen Meagher, Department of Mathematics & Statistics *participated via Video Conference
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ABSTRACT
Neurotoxic methylmercury (MeHg) biomagnifies in food webs and may harm
human and wildlife health. Methylmercury is produced by the methylation of mercury
(Hg), primarily by sulphate reducing bacteria and iron reducing bacteria in anaerobic
sediments of aquatic systems. Anthropogenic emissions of Hg may circulate globally
in the atmosphere and have increased deposition of Hg to aquatic systems remote from
the sources of Hg emissions. Deposited Hg adds to the pool of Hg available for
methylation. Wetlands in particular have been identified as sites of elevated MeHg
production because of the anaerobic nature of wetland sediments. North America’s
prairie pothole region contains millions of wetlands which provide waterfowl breeding
habitat, carbon storage, and groundwater recharge. Surface water methylmercury
concentrations in some prairie wetlands are elevated, suggesting the potential for
substantial production of MeHg within these wetlands. In summer 2011 sediment
cores from wetlands in Saskatchewan’s St. Denis National Wildlife Area were injected
with 201Hg and incubated within wetlands to measure the rate of formation of Me201Hg
from the injected isotope. In addition to measuring potential rates of MeHg
production, concentrations of MeHg and total mercury (THg) in sediment, porewater,
and surface water were also measured along other sediment and surface water
parameters. Potential rates of MeHg production and sediment MeHg and THg
concentrations were similar to those observed in other remote freshwater wetlands.
Sediment porosity was negatively correlated with MeHg production. Concentrations
of MeHg in wetland surface water were positively correlated with concentrations of
sediment MeHg and calculated MeHg diffusive flux was from porewater to the surface
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water in the majority of studied wetlands. This is the first study to report potential Hg
methylation rates in wetlands from the prairie pothole region.
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ACKNOWLEDGEMENTS
I would like to thank everyone who helped me with guidance, advice, funding, lab
and field work. I would like to thank my committee members, Drs. Britt Hall, Andrew
Cameron, and Carl Mitchell for their assistance.
Advice from Lara Bates, Dr. Igor Lehnherr, Dr. Gavin Simpson, and Vanessa
Swarbrick was much appreciated. I would also like to thank Dr. Vincent St. Louis and
Valery Bazira from the University of Alberta, Dr. Brian Branfireun, Michelle Collins,
and Robin Tiller from the University of Western Ontario, and Vincent Ignatiuk and the
University of Regina Limnology Lab for their assistance preparing and analyzing
samples. I would like to thank Stacy Boczulak, Nolan Hoggarth, Aleksandra Bugajski,
Korrieh Gurniak, Derek Wright, and Kyleen Pangracs for their help in the field. Thank
you to the SDNWA and Marc Loiselle for access to sample the wetlands. I would also
like to thank my family and friends for their support.
I would like to acknowledge the support provided the University of Regina Faculty
of Graduate Studies and Research for funding through a Graduate Research Award and
Graduate Teaching Assistantships, the Riegert Fund for assistance to present at the
2011 Amercian Geophysical Union conference, and Natural Sciences and Engineering
Research Council of Canada.
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Table of Contents
Abstract ............................................................................................................................ i
Acknowledgements ........................................................................................................iii
Table of Contents ........................................................................................................... iv
List of Tables..................................................................................................................vi
List of Figures ...............................................................................................................vii
List of Appendices.......................................................................................................... ix
CHAPTER ONE: General Introduction ........................................................................1 1 Introduction ................................................................................................ 1 2 Mercury in the Environment ...................................................................... 3
2.1 Wetlands and MeHg................................................................................ 3 2.2 Wetlands and THg................................................................................... 4 2.3 Sediment Mercury Concentrations .......................................................... 5
3 Methylmercury Production ........................................................................ 6 4 Factors Affecting Methylmercury Production ........................................... 7
4.1 Microbial Community ............................................................................. 7 4.2 Temperature............................................................................................. 7 4.3 pH ............................................................................................................ 8 4.4 Organic Matter ........................................................................................ 9 4.5 Redox Conditions .................................................................................. 10 4.6 Sulphate/Sulphide/Iron.......................................................................... 12
5 The Use of Mercury Isotopes in MeHg Production Studies .................... 13 6 Surface Water MeHg Concentrations ...................................................... 17 7 Biotic Uptake of Water Column Mercury................................................ 17 8 Mercury and Wetland Biota ..................................................................... 18 9 Prairie Pothole Region ............................................................................. 19 10 Objectives................................................................................................. 23
CHAPTER TWO: Methylmercury Production in Prairie Wetland Sediment ............25 1 Introduction .............................................................................................. 25 2 Methods.................................................................................................... 27
2.1 Study Site .............................................................................................. 27 2.2 Sample Collection ................................................................................. 32
2.2.1 Sediment MeHg and THg Concentrations and Methylation Potentials........................................................................................................... 32
2.2.2 Sediment Water Content, Porosity, and Organic Content................. 33 2.2.3 Porewater MeHg and THg Concentrations ....................................... 34 2.2.4 Surface Water MeHg and THg Concentrations ................................ 35
2.3 Sample Analysis .................................................................................... 36
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2.3.1 Sediment MeHg and THg Concentrations and Methylation Potentials........................................................................................................... 36
2.3.2 Surface and Porewater MeHg and THg Concentrations ................... 38 2.3.3 Sediment Water Content, Porosity, and Organic Content................. 39 2.3.4 Diffusive Flux of MeHg.................................................................... 40
2.4 Statistical Analysis ................................................................................ 41 3 Results and Discussion............................................................................. 41
3.1 Sediment MeHg and THg Concentrations ............................................ 41 3.2 Sediment Hg Methylation Potentials..................................................... 44 3.3 Factors Controlling km Values............................................................... 47 3.4 Porewater MeHg and THg Concentrations ........................................... 50 3.5 Factors Potentially Influencing Sediment km Values and MeHg and THg
Concentrations in Wetlands................................................................... 53 3.5.1 Organic Matter .................................................................................. 54 3.5.2 Divalent Cations ................................................................................ 55 3.5.3 Sulphur and Iron Cycling .................................................................. 58 3.5.4 pH ...................................................................................................... 60
3.6 Partitioning Coefficients........................................................................ 61 3.7 Surface Water MeHg and THg Concentrations .................................... 62 3.8 Diffusive Flux of MeHg........................................................................ 64 3.9 Conclusions ........................................................................................... 68
References: ..................................................................................................................69
Appendices: .................................................................................................................80
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List of Tables
Table 1: Latitude and longitude for wetlands sampled for sediment in 2008, 2010, and 2011. The (-) symbol indicates that the wetland was not sampled, (○) the wetland was sampled only for sediment total mercury (THg), methylmercury (MeHg), water content, and organic carbon, (●) the wetland was sampled for surface water and porewater in addition to sediment methylation potential, THg, MeHg, water content, organic carbon, and porosity, (D) the wetland was dry and if preceded by a sampling symbol indicates a dry sediment sample was taken from the wetland.
Table 2: Quality assurance and quality control for total mercury (THg) and
methylmercury (MeHg) analysis of water and sediment samples with recovery of standard reference materials (SRM) or spiked samples ± one standard deviation. Approximately 10% of samples were analyzed in duplicate. Water MeHg and October sediment MeHg spike recovery are reported for all spikes as well as only spikes that were greater than sample MeHg. For water THg and October sediment THg minimum detection limits were less than 0.3 ng L-1, for water MeHg and October sediment MeHg 0.02 ng L-1, and for July and August sediment were 1.13 ng g-1 for THg and 0.01056 ng g-1 for MeHg.
Table 3: Methylation potential (km (d-1)) and percent of total mercury as
methylmercury (%MeHg) for freshwater wetland, saltwater wetland, and marine sediment.
Table 4: Table 4: Surface water conductivity and pH, sediment water content,
porosity, organic carbon, log sediment/porewater total mercury (THg) and methylmercury (MeHg) partitioning coefficients (L kg-1), and MeHg diffusive flux from porewater to surface water (ng m-2 day-1) for wetlands sampled in 2011. Negative values for MeHg diffusive flux indicate diffusive flux from surface water to porewater. ND indicates sites with no data.
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List of Figures
Figure 1: Map of wetlands sampled within the St. Denis National Wildlife Area (SDNWA) and three additional sampled wetlands northeast of the SDNWA.
Fig. 2: (A) Sediment methylmercury (MeHg) concentrations, (B) total mercury (THg)
concentrations, and (C) percent THg that is MeHg (%MeHg) from nine wetlands sampled in 2011. Wetland OC 2 was dry in August and October 2011. Two 0-2 cm depth sediment samples were analyzed per wetland. Error bars indicate one standard error.
Fig. 3: Sediment mercury methylation potentials (d-1) for nine wetlands sampled in
2011. Two samples per wetland were analyzed. OC 2 was dry in August. Added 201Hg was below the detection limit in August in Pond 2, 130, and 139. Error bars indicate one standard error.
Fig. 4: Relationships between (A): July surface water methylmercury (MeHg) and
porosity. (B): 201Hg methylation potential (d-1) and porosity in July. (C): August surface water MeHg and porosity. (D): August 201Hg methylation potential (d-1) and porosity. Upland sites are indicated by triangles, organic sites by squares, and lowland sites by circles.
Fig. 5: (A) Porewater methylmercury (MeHg) concentrations, (B) total mercury (THg)
concentrations, and (C) percent of total mercury that is methylmercury (%MeHg) from twelve wetlands sampled in July and August 2011. Wetland OC 2 was dry in August 2011 and the filter was damaged during collection of the July MeHg sample from Pond 139. No error bars are reported because only one porewater sample was collected per analyte per site.
Fig. 6: Sediment methylmercury concentrations and organic content in August 2011.
Triangles represent upland wetlands, squares wetlands within organically farmed fields, and circles lowland wetlands.
Fig. 7: Surface water methylmercury (MeHg) concentrations, total mercury (THg)
concentrations, and percent of total mercury that is methylmercury (%MeHg) from twelve wetlands sampled in July and August 2011. Wetland OC 2 was dry in August 2011. No error bars are reported as one sample for MeHg and one sample for THg were collected per site.
Fig. 8: (A) Surface water methylmercury (MeHg) was correlated (p = 0.027) with ln 0-
2 cm section sediment MeHg concentrations in nine wetlands sampled in July 2011. (B) August 2011 surface water MeHg concentrations were not correlated with sediment MeHg (p = 0.069). Upland sites are indicated by triangles, organic sites by squares, and lowland sites by circles.
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Fig. 9: Calculated diffusive flux of MeHg between porewater and surface water. The upper two panels include the sites sampled within the St. Denis National Wildlife Area (SDNWA) while the lower two panels include wetlands surrounded by organically farmed fields northeast of the SDNWA. Red circles indicate movement of MeHg from porewater to surface water and yellow circles indicate MeHg transfer from surface water to porewater.
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List of Appendices
Appendix A: Surface water temperature, oxygen concentration, and time sample collected for wetlands sampled in July and August 2011.
Appendix B: Porewater temperature, oxygen concentration, conductivity, pH and time
sample collected for wetlands sampled in July and August 2011. Appendix C: Sediment temperature, water content in 2-4 cm and 4-8 cm sections, and
organic content in 2-4 cm and 4-8 cm sections from July and August 2011. Appendix D: Sediment porosity and bulk density in 0-2 cm, 2-4cm, and 4-8 cm
sections of sediment from July and August 2011. Appendix E: 2011 sediment organic content in 2-4 cm and 4-8 cm sections and
proportion of coarse (not passing through 2 mm sieve mesh), medium (did not pass through 63 µm sieve), and fine sediment (passed through both sieves), as well as vegetation that collected on the 2 mm sieve from the 0-2 cm sediment section.
Appendix F: Concentration of excess (due to the addition of 201Hg) Me201Hg and 201Hg
in the 0-2 cm section of incubated sediment cores from July and August 2011. Appendix G: Mean sediment water content, organic content, THg, MeHg, and number
of years the wetland was sampled in October of 2008, 2010, and 2011. Appendix H: Mean sediment water content and organic content from triplicate 0-2 cm
cores collected in May, June, July, and August 2012.
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CHAPTER 1: General Introduction
1. INTRODUCTION
Methylmercury (MeHg) is a neurotoxin produced by the methylation of inorganic
mercury (Hg) by microorganisms. Sulphate-reducing bacteria (SRB) and iron-
reducing bacteria (FeRB) are likely the main methylators of Hg to MeHg (Ullrich et
al., 2001) and, because of their anoxic requirements, Hg methylation occurs in low
oxygen and high carbon environments such as lake and wetland sediment.
Methylmercury produced in anaerobic sediment may be transferred to the water
column, by bioturbation, ebullition, and diffusive flux, and biomagnify in food webs
linked to aquatic systems, potentially harming the health of humans and wildlife
consuming aquatic organisms with elevated MeHg.
The natural mercury cycle involves atmospheric transport of geologically derived
Hg emitted via geothermal activity, deposition to terrestrial and aquatic ecosystems,
and revolatilisation to the atmosphere (Ullrich et al., 2001). Deep ocean sediment is
the ultimate sink for mercury (Selin, 2009). Anthropogenic activities including the
mining and industrial use of Hg, as well as emissions from combustion of fossil fuels,
have increased Hg emissions to the atmosphere above pre-industrial levels (Selin,
2009). The high vapour pressure of elemental Hg (Hg0) increases the mean
atmospheric residence time, thus allowing global dispersion of Hg from emissions
sources (Hintelmann, 2010). Long-range atmospheric transport of elemental Hg, and
its subsequent deposition, results in the potential for pollution of ecosystems distant
from the original sources of Hg emissions (Lindqvist et al., 1991).
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Anthropogenic emissions of Hg to the atmosphere are from a variety of sources.
Gold and silver mining during the second half of the 19th century led to substantial Hg
emissions and increasing artisanal gold mining and coal combustion from the middle
of the 20th century to the present have increased anthropogenic emissions of mercury to
the atmosphere (Streets et al., 2011). The major sources of anthropogenic emissions
recently include biomass burning (25%), power plant fuel combustion (18%), zinc
smelting (15%), transportation fuel combustion (12%), and industrial fuel combustion
(11%; data from 2006, Streets et al., 2009). Projections of anthropogenic Hg emissions
indicate that by 2050 emissions are likely to rise with a range of -4% to +96% (2390-
4860 Mg Hg) from 2006 emissions, based on scenarios from the Intergovernmental
Panel on Climate Change for future economic, demographic, and technological
changes, with much of the projected rise from increasing use of Hg-emitting coal
combustion for electricity generation in developing countries (Streets et al., 2009).
Implementation of Hg emission control technologies in coal-fired power plants, such
as activated carbon injection, could reduce projected 2050 emissions by approximately
30% to 1670-3480 Mg (Streets et al., 2009). Atmospheric Hg0 is oxidized by light or
chemical reactions to the more soluble Hg+2 which deposits to terrestrial and aquatic
ecosystems (Selin, 2009). Anthropogenic emissions have resulted in increased
sediment Hg concentrations in mid-continental lakes of the United States and Canada
by approximately 2.7 times background Hg concentrations (Fitzgerald et al., 1998).
Inorganic Hg deposited from the atmosphere to aquatic systems may be
methylated, typically by SRB or FeRB in anaerobic sediment, to MeHg. This MeHg
accumulates in organisms connected to aquatic food chains. Primary producers
3
passively take up MeHg from their environment (Mason et al., 1995). Dietary uptake
is the main source of MeHg to consumers with direct uptake from water a secondary
pathway (Hall et al., 2004; Hrenchuk et al., 2012). Biomagnification of MeHg occurs
in aquatic food chains (Chasar et al., 2009), with increasing MeHg concentration and
the proportion of total Hg that is MeHg (%MeHg) at higher trophic levels (Watras and
Bloom, 1992).
Uptake of either inorganic Hg or MeHg is harmful to animal and human health.
The target organ of inorganic Hg exposure tends to be the kidney, while MeHg targets
the central nervous system and leads to neurodevelopmental effects (Agency for Toxic
Substances and Disease Registry, 1994; Grandjean et al., 1997). Developmental
effects of MeHg have been observed in children related to increasing maternal hair Hg
concentrations below 10 µg g-1 (after exclusion of children of mothers with maternal
hair Hg concentrations above 10 µg g-1) (Grandjean et al., 1997). Historical
occurrences of large scale MeHg poisoning include Minamata and Niigata, Japan in
1950-1960s from consumption of fish which had accumulated MeHg discharged by
factories, and Iraq in 1972 from consumption of wheat, intended for use as seed,
treated with MeHg fungicide (Sanfeliu et al., 2003). Harm to the environment and
human health resulting from anthropogenic emissions of Hg has resulted in substantial
economic costs (Sundseth et al., 2010).
2. MERCURY IN THE ENVIRONMENT
2.1 Wetlands and MeHg
Freshwater wetlands are sites of high species richness for wildlife and provide
valuable environmental services, such as habitat, greenhouse gas regulation, water
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supply, groundwater recharge, flood control, waste treatment, and recreation (Costanza
et al., 1997). These services are often much more valuable than those obtained after
draining wetlands for agriculture (Millennium Ecosystem Assessment, 2005.
Ecosystems and Human Well-Being: Wetlands and Water, 2005). In addition,
wetlands may directly contribute to MeHg concentrations in lakes (St. Louis et al.,
1994) and streams (Brigham et al., 2009) via export of MeHg produced in situ, and
indirectly contribute to lake MeHg concentrations by exporting inorganic Hg, organic
matter, and sulfate (Watras et al., 2005) to downstream environments where inorganic
Hg could be methylated. Wetland sediments often support anaerobic conditions
promoting the activity of Hg methylating SRB (Gilmour et al., 1992), and therefore,
flux of MeHg to the water column (Holmes and Lean, 2006). Both the total amount, as
well as the partitioning (Hammerschmidt et al., 2008) and speciation (Benoit et al.,
1999), of inorganic Hg will influence the amount of MeHg produced. Spatial and
temporal variability of porewater %MeHg within peatlands suggests that methylation
rates may not be uniform within sediments from the same wetland. For example,
porewater %MeHg was higher at the peatland upland interface than in the interior of a
peatland within forested uplands, possibly due to delivery of upland sulphate and
carbon to the edge of the peatland (Mitchell et al., 2008a).
2.2 Wetlands and THg
Wetland sediments are often a net sink for THg (inorganic Hg plus MeHg) and
THg concentrations are influenced by Hg binding to organic carbon and inorganic
matter, sediment particle size, and organic content. Wetland sediment THg binds to
organic thiols (Benoit et al., 1999) and inorganic complexes such as FeS2 (Bower et al.,
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2008). Particle size also influences THg concentrations in sediment (Håkanson, 1977;
Kaplan et al., 2002) since the higher surface area to volume ratios of smaller particles
provide additional binding sites for Hg (Horowitz and Elrick, 1987). Sediment organic
content may also influence THg concentrations because decreased sediment Hg
absorption after removal of organic matter by treatment with H2O2 relative to untreated
sediment has been observed (Liao et al., 2009).
Sediment THg and MeHg are largely associated with sediment rather than
porewater. Partitioning coefficients (ratio of concentration of a chemical in sediment
to concentration of chemical in porewater) of Hg(II) range from log10 3.8 to 6.0 L kg-1
and log10 2.8 to 5.0 L kg-1 for MeHg (Lyon et al., 1997). Mercury tends to have a
higher sediment/porewater partitioning coefficient than other metals, although the
range of sediment/porewater partitioning coefficients was at least two orders of
magnitude for each metal considered in a literature survey (Allison and Allison, 2005).
For example, Hg sediment/porewater partitioning coefficients from Long Island Sound
marine sediments ranged from log10 3.18 to 4.92 L kg-1 and were positively correlated
with sediment organic content and negatively related to methylation potentials,
possibly as Hg bound to sediment is less available than Hg in porewater
(Hammerschmidt and Fitzgerald, 2004). Diffusive flux of THg from sediments with
high organic matter content to surface water is likely less significant than other
processes such as ebullition (Canario et al., 2009), as diffusive flux accounted for less
than 3% of measured THg flux to the water column of Quebec’s Lake Saint-Louis.
2.3 Sediment Mercury Concentrations
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There are very few studies examining Hg concentrations in sediments from
wetlands in the Prairie Pothole Region (PPR). A study from the Lostwood National
Wildlife Refuge (LNWR) in North Dakota examined freshwater wetlands for sediment
and surface water THg and MeHg concentrations, as well as surface water chemistry
(Sando et al., 2007). Concentrations of THg in studied LNWR wetland sediments
ranged from 6.77 to 99.0 ng g-1 (Sando et al., 2007). Concentrations of sediment
MeHg (<0.4 to 4.16 ng g-1) in the LNWR varied by year and wetland type with the
highest MeHg in seasonal wetlands, followed by those in temporary, semi-permanent,
and lake wetlands (Sando et al., 2007). Seasonal wetlands are described as having
longer hydroperiods than temporary wetlands, whereas semi-permanent wetlands are
flooded throughout growing season in most years and lake wetlands are permanently
flooded (Sando et al., 2007). Higher MeHg concentrations in LNWR seasonal
wetlands may be due to the more frequent wetting and drying cycles, which allow for
oxidation of electron acceptors in dry periods followed by flooding which could
stimulate the activity of Hg methylating anaerobic microbes (Sando et al., 2007).
3. METHYLMERCURY PRODUCTION
Sediment MeHg concentrations depend on net Hg methylation rates (Gilmour et
al., 1992) because, in the absence of MeHg flux to or from the sediment, changes in the
concentration of MeHg would depend on the rate of net Hg methylation. Net Hg
methylation rates are determined by the gross rate of Hg methylation less the gross rate
of the degradation of MeHg. Demethylation is mediated by a variety of microbes and
the rates of both methylation of Hg and demethylation of MeHg depend on a variety of
environmental factors. Methylation rates are a product of the bioavailability of Hg and
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the activity of methylating bacteria (Hintelmann, 2010). The forms of Hg most
bioavailable to methylating bacteria may be neutrally charged Hg-sulphur (Benoit et
al., 1999) and HgS nanoparticles (Graham et al., 2012) because they can diffuse across
the bacterial membrane. There is also evidence for facilitated uptake of Hg+2 (Golding
et al., 2002). Once mercury is present in the cell the activity of methylating bacteria
determines the rate of methylation.
4. FACTORS AFFECTING MEHG PRODUCTION:
Microbial activity is dependent on a number of factors such as temperature,
quantity and quality of organic matter available, redox conditions, and availability of
electron acceptors (Ullrich et al., 2001). Bioavailability of Hg is also dependent on a
similar set of factors including pH, organic matter, redox conditions, and sulphide
concentrations.
4.1. Microbial Community
Although SRB are thought to be the main microbial community contributing to Hg
methylation (Compeau and Bartha, 1985), groups other than the SRB may also
methylate Hg (Achá et al., 2011). Iron-reducing bacteria (FeRB) have been shown to
contribute to net methylation rates in freshwater systems by methylating Hg (Fleming
et al., 2006; Yu et al., 2012) and possibly by reducing rates of demethylation
(Avramescu et al., 2011). In lake periphyton, methanogens may be important Hg
methylators (Hamelin et al., 2011).
4.2. Temperature
Temperature influences Hg methylation rates because microbial activity increases
with increasing temperatures (Ullrich et al., 2001). Net methylation rates in arctic
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wetland sediments were found to increase with arctic summer temperatures (Loseto et
al., 2004). Similarly gross methylation rates were higher at increased temperatures in
marine sediment samples from Long Island Sound (Hammerschmidt and Fitzgerald,
2004). Seasonal influences on other factors complicate the relationship between
temperature and methylation rates. For example sulphate reduction, calculated from
changes in surface water sulphate concentrations, in prairie recharge wetlands may be
highest in the late spring (May-June) and decline over the summer (Heagle et al.,
2007), possibly due to spring flooding of wetland soils and depletion of electron
acceptors in anaerobic sediments until sulphate reduction is favoured (Reddy and
DeLaune, 2008). As surface water sulphate concentrations decline over the course of
the summer, sulphate reduction rates decreased (Heagle et al., 2007), while DOC
(Waiser, 2006) and temperature in prairie wetlands typically increase. Increasing
temperature may also promote greater rates of methylation relative to demethylation,
leading to higher net methylation (Ullrich et al., 2001).
4.3. pH
The influence of pH on MeHg production may be due to the impacts on Hg
speciation, as well as microbial Hg uptake and microbial community composition. The
effect of pH on methylation rates is likely partially due to differing stabilities of
various species of Hg under different pH conditions (Reddy and DeLaune, 2008).
Under reducing conditions uncharged Hg complexes such as Hg(HS)2 are favoured
when pH is less than six, while HgHS-2 and HgS2-2 are favoured with increasing pH
(Reddy and DeLaune, 2008). Aerobic conditions with pH less than six favours the
formation of HgCl2 and higher pH favours Hg(OH)2 (Reddy and DeLaune, 2008).
9
Uptake of Hg by bacteria was found to increase with even small changes (7.3-6.3) in
pH, indicating that pH should be considered a factor in the bioavailability of Hg for
biotic methylation (Kelly et al., 2003). As biotic Hg methylation is conducted inside
the controlled internal environment of bacteria, pH should not have a direct impact on
the metabolic reaction resulting in methylation (Kelly et al., 2003). Acidity may also
influence the composition of wetland sediment microbial communities (Hartman et al.,
2008) and, if the ratio of methylating to demethylating bacteria changes in response to
pH, net methylation could be influenced (Ullrich et al., 2001).
4.4. Organic Matter
Organic matter influences methylation rates by forming complexes with Hg,
changing its bioavailability, and providing a source of energy increasing microbial
activity (Hintelmann, 2010). Higher levels of organic carbon have been found to
influence production of MeHg in lake sediments (Gilmour and Riedel, 1995), marine
sediments (Hammerschmidt and Fitzgerald, 2004), and fresh water stream sediments
(Marvin-DiPasquale et al., 2003). In peatland sediments with methylation rates
originally limited by sulphate, addition of organic matter and sulphate increased net
methylation rates more than the addition of sulphate alone (Mitchell et al., 2008b). In
addition to the quantity of organic matter the quality also influences methylation rates.
Gross methylation rates were found to be positively related to the aromaticity of
organic carbon in Chesapeake Bay salt marsh sediment (Mitchell and Gilmour, 2008).
Surface water MeHg concentrations may be positively related to the concentration of
hydrophobic organic acids (Hall et al., 2008), which have been found to increase the
bioavailability of Hg by stabilizing HgS nanoparticles against aggregation allowing for
10
uptake by methylating bacteria (Graham et al., 2012). Gross methylation rates in
marine sediments from the Venice Lagoon were observed to increase with decreasing
C:N in sediment organic matter (Kim et al., 2011). Flooding of terrestrial organic
matter encourages microbial activity and increase MeHg concentrations in surface
water (Hall et al., 2005); surface water MeHg concentrations remain elevated for
longer if a greater quantity of organic matter is flooded (Hall et al., 2009b). The type
of organic matter flooded may also be a factor. In an enclosure experiment
decomposing jack pine (Pinus banksiana) needles resulted in lower production of
MeHg than birch (Betula papyrifera) leaves, possibly due to higher carbon in the birch
leaves enclosure increasing microbial activity (Hall et al., 2004). Removal of
vegetation from wetlands reduces MeHg production by removing a possible source of
organic carbon (acetate) delivered by plant roots to wetland sediment (Windham-
Myers et al., 2009). There may be an optimal amount of organic matter for peak
methylation rates, with a trade-off between reduced Hg bioavailability with increasing
amounts of organic matter and limitation of microbial activity with decreasing amounts
of organic matter (Hintelmann, 2010).
4.5. Redox Conditions
Redox conditions influence MeHg production as SRB, the primary methylators of
Hg, are strict anaerobes (Ullrich et al., 2001). Sulphate reduction by SRBs using
sulphate as an electron acceptor is usually found in sediments with redox potential less
than 100 mV (Reddy and DeLaune, 2008). Under aerobic conditions oxygen is the
primary electron acceptor, but as redox potential decreases, microbes (facultative
aerobes and various anaerobic microbes) sequentially use oxygen and oxides of
11
nitrogen, manganese, and iron as electron acceptors before reducing sulphate (Reddy
and DeLaune, 2008). MeHg production was observed in anaerobic-biotic incubations
(1 week incubation, daily sampling) of estuarine sediment slurry from the Adour River,
but not in aerobic-biotic, aerobic-abiotic, and anaerobic-abiotic sediment incubations,
although demethylation of 30-43% per week was observed in each of the four
conditions, supporting the role of anaerobic bacteria in MeHg production (Rodriguez
Martin-Doimeadios et al., 2004). In marine sediments with low oxygen, methylation
may be greatest in the surface sediments while in sediments with deeper oxygen
penetration the highest methylation rates occur below the surface sediments (Hollweg
et al., 2009). Under anaerobic conditions mercury uptake by facultatively anaerobic
bacteria increased relative to aerobic conditions, suggesting energy requiring facilitated
uptake of Hg, rather than passive diffusion (Golding et al., 2002).
Redox conditions influence methylation rates by influencing microbial activity and
bioavailability of Hg. Under aerobic conditions, demethylation occurs as oxidative
demethylation, producing CO2 and Hg+2 which is then bioavailable for methylation
(Gilmour et al., 1998). Demethylation rates may be reduced in wetland sediment under
anaerobic conditions (Goulet et al., 2007), because of the absence of oxygen.
Macrophtye roots in anaerobic sediment supply oxygen to the sediment, allowing for
the oxidation of sulphide to sulphate, increasing the bioavailability of Hg and
supplying SRB with a source of sulphate (Gilmour et al., 1998; Marvin-DiPasquale et
al., 2003). Increased methylation rates in estuarine sediment from the Gironde Estuary
may be linked to fluctuating oxygen availability stimulating activity of the microbial
community (Schäfer et al., 2010).
12
4.6. Sulphate/Sulphide/Iron
Since SRB are considered to be the main microbial community methylating Hg
(Compeau and Bartha, 1985), sulphate and sulphide concentrations are potential
factors influencing MeHg production because SRB use sulphate as an electron acceptor
(Hintelmann, 2010). Higher concentrations of sulphate have been observed to
stimulate sediment MeHg production in a variety of ecosystems including peatlands
(Branfireun et al., 1999; Mitchell et al., 2008b), reservoirs (Gilmour et al., 1992), and
wetlands (Gilmour et al., 1998; Harmon et al., 2004). Higher concentrations of
sulphide have been observed to reduce salt marsh sediment MeHg production (Mitchell
and Gilmour, 2008). Sulphide may reduce Hg methylation rates by reducing the
bioavailability of Hg+2 through the formation of unavailable solid HgS (cinnabar)
(Hintelmann, 2010), although recent research suggests that HgS nanoparticles
stabilized by DOM are available for uptake by methylating bacteria (Graham et al.,
2012), or by forming metal sulphides, such as pyrite, that bind Hg (Bower et al., 2008).
Lower concentrations of sulphide increases MeHg production by forming neutrally
charged complexes with Hg, which may diffuse across the membranes of bacterial
cells, increasing the bioavailability of Hg (Benoit et al., 1999; Ullrich et al., 2001).
Macrophytes also have an influence on sulphide concentrations with an increase in
porewater sulphide observed after the flowering of aquatic plants (Harmon et al.,
2004).
Iron reducing bacteria (FeRB) may also be involved in MeHg in freshwater
wetlands since devegetation reduced Hg methylation and Fe(OH)3 reduction, but not
sulphate reduction, indicating involvement of FeRB in MeHg production (Windham-
13
Myers et al., 2009). Both SRB and FeRB were likely to methylate Hg in sediments at
some sites along a Hg contaminated freshwater river (Yu et al., 2012). Bioavailability
of Hg to methylating bacteria and MeHg production are reduced if Hg+2 binds to FeS
(Rothenberg et al., 2008; Xiong et al., 2009).
5. THE USE OF MERCURY ISOTOPES IN MEHG PRODUCTION STUDIES
Stable isotopes of Hg include: 196Hg, 198Hg, 199Hg, 200Hg, 201Hg, 202Hg, and 204Hg
(Ridley and Stetson, 2006). In addition to stable isotopes, radioisotopes of Hg,
including 203Hg, are also found; however these radioisotopes do not occur naturally
(Ridley and Stetson, 2006). Mass dependent fractionation of Hg has been observed
under both light and dark conditions with preferential reduction of lighter isotopes by
photoreduction under light conditions and reactions with organic matter under dark
conditions (Bergquist and Blum, 2007). Mass dependent fractionation may also result
from the preferential methylation of lighter Hg isotopes by sulphate-reducing bacteria
(Desulfobulbus propionicus) resulting in MeHg enriched in lighter isotopes relative to
the pool of Hg available for methylation (Rodríguez-González et al., 2009). Although
Hg stable isotopes have a higher atomic weight than stable isotopes of other elements
such as C, H, N, O, and S that are used in ecology, fractionation of Hg stable isotopes
may be large enough to identify sources of Hg (Perrot et al., 2010; Sherman et al.,
2012). Stable isotopes of Hg have been used to determine the source of Hg in fish,
with an increasing fraction of heavy Hg isotopes at higher trophic levels in fish from
uncontaminated lakes (fish preferentially excrete lighter Hg isotopes; Bergquist and
Blum, 2007), while no fractionation was observed in a nearby reservoir contaminated
with Hg from chlor-alkali production (Perrot et al., 2010). Stable isotopes of Hg in
14
precipitation may also be used to identify Hg from large local point sources, such as a
coal-fired power plant, although the isotopic composition of the Hg released depends
on the isotopic composition of the coal combusted and air pollution control devices in
place (Sherman et al., 2012). However, in ecological studies Hg stable isotopes are
mostly used to measure rates of Hg methylation and MeHg demethylation.
Addition of stable isotopes (Hintelmann et al., 1995) and radioisotopes (Gilmour
and Riedel, 1995) to sediment, followed by incubation, has been used to measure
sediment rates of Hg methylation and MeHg demethylation. The radioisotopes 203Hg
and 14C have been employed to measure methylation rates of Hg and demethylation of
MeHg (Furutani and Rudd, 1980; Gilmour and Riedel, 1995; Marvin-DiPasquale and
Oremland, 1998), respectively, although it is now more common to use stable isotopes
to measure methylation rates because commercially available 203Hg has low specific
activity (Gilmour and Riedel, 1995) and therefore the high concentrations used in
incubations are not environmentally relevant (Hintelmann et al., 1995). Stable isotopes
allow for measurement of methylation rates using additions of an inorganic mercury
stable isotope close to ambient levels of Hg (Rodriguez Martin-Doimeadios et al.,
2004). Unlike radioisotopes, different Hg stable isotopes can be measured allowing for
the measurement of sediment methylation and demethylation in the same sample
(Hintelmann et al., 1995). Conversely, measurements of radioisotopes require
different processing for 203Hg (Marvin-DiPasquale et al., 2003) and 14C (Marvin-
DiPasquale and Oremland, 1998).
Equilibration of the radioisotope or stable isotope with surface (Mitchell and
Gilmour, 2008) or porewater (Gilmour and Riedel, 1995) from the site before addition
15
of the isotope may allow the added tracer to better reflect the bioavailability of ambient
MeHg. However, care must be taken to consider methylation rates from isotope
additions as potential rates because added isotope is likely more bioavailable than
ambient Hg (Jonsson et al., 2012; Mitchell and Gilmour, 2008). Aqueous Hg tracers
are much more available than metacinnabar and cinnabar, with Hg bound to
mackinawite and natural organic matter intermediate in availability for methylation
(Jonsson et al., 2012).
Rates of net methylation are very difficult to determine because Hg isotopes used
to measure rates may be more available than ambient Hg and the net rate depends on
both methylation and demethylation. Past methods of determining potential rates of
both methylation and demethylation were made by injecting of a radioisotopic 203Hg
into sediment cores and measuring the resulting 203Hg and Me203Hg by scintillation
counting (Marvin-DiPasquale et al., 2003). However, because the specific activity of
radioisotopic Hg is quite low, these incubations could not be performed at
environmentally relevant concentrations. More recently, rates of methylation and
demethylation have been obtained using stable Hg isotopes in incubated sediment and
therefore allowing for the calculation of gross rates of Hg methylation and MeHg
demethylation. Studies using stable isotopes of Hg and 203Hg have shown that
sediment MeHg concentrations in aquatic ecosystems may be related to Hg
methylation rates (Hammerschmidt and Fitzgerald, 2006), although demethylation
rates also influence ambient MeHg concentrations (Tjerngren et al., 2012a).
Demethylation of MeHg may be biotic or abiotic (Hintelmann, 2010). Microbial
organisms, such as SRB and methanogens, demethylate through the cometabolism of
16
MeHg in oxidative demethylation (Marvin-DiPasquale and Oremland, 1998) and
production of organomercurial lyase (Robinson and Tuovinen, 1984). Reductive
demethylation of MeHg may produce Hg0 and CH4 (Hintelmann, 2010). Abiotic
photodegradation of MeHg to Hg+2 and Hg0 occurs in light-receiving surface waters
(Hintelmann, 2010) and can be a significant sink of MeHg in aquatic systems.
Photodegradation has been observed in the water column of lakes (Sellers et al., 1996),
wetlands (Lehnherr et al., 2012a; Naftz et al., 2011), and in rainwater (Bittrich et al.,
2011). Photodegradation of MeHg in surface water of Lake 240 at the Experimental
Lakes Area in northwestern Ontario was related to photosynthetically active radiation
and equivalent to nearly double the external inputs of MeHg into the lake from
streamflow and runoff (Sellers et al., 1996). In wetlands, photodegradation of MeHg
can cause diurnal variation of surface water MeHg concentrations (Naftz et al., 2011).
Studies using Hg isotopes to measure rates of methylation in sediments have been
conducted in freshwater and saltwater wetlands, streams, lakes, and marine sediment.
Rates of MeHg production are usually reported as rate constants (d-1) and range from
approximately 0 (Langer et al., 2001) to 0.37 (d-1) (Windham-Myers et al., 2009). In
freshwater wetlands MeHg production has been measured in sediment of wetlands
from the Arctic (Lehnherr et al., 2012b), the Florida Everglades (Gilmour et al., 1998),
California’s Yolo Bypass (Windham-Myers et al., 2009), and boreal Sweden
(Tjerngren et al., 2012a). Studies in freshwater wetlands have found a positive
correlation between MeHg production rates and MeHg concentrations or %MeHg
(Gilmour et al., 1998; Lehnherr et al., 2012b), although sediment %MeHg may follow
17
ratios of MeHg production and demethylation rates more closely than MeHg
production rates alone (Tjerngren et al., 2012a).
6. SURFACE WATER MEHG CONCENTRATIONS
Concentrations of MeHg in wetland water columns may be related to sediment
MeHg concentrations due to flux of MeHg from sites of MeHg production to the
wetland water column, therefore providing a link between sediments and the water
column where aquatic biota may be exposed to MeHg (Holmes and Lean, 2006). Flux
of MeHg from sediment porewater to the water column has been observed in
freshwater and marine systems (Holmes and Lean, 2006; Rothenberg et al., 2008).
Diffusive flux may be a small fraction of total flux as ebullition (Canario et al., 2009)
and bioturbation (Benoit et al., 2009) both contribute to total MeHg flux to the water
column. Flux of MeHg from wetland sediment to surface water is a much greater
source of MeHg to the water column than other sources, such as MeHg from
precipitation (Lehnherr et al., 2012a). Concentrations of whole water MeHg in lakes
and wetlands in the Experimental Lake Area of northwestern Ontario are typically
0.04-0.25 ng L-1 (Hall et al., 2009b). Whole water MeHg concentrations from
wetlands and lakes in the PPR of Saskatchewan were 0.02-4.21 ng L-1, with higher
MeHg concentrations in wetlands than lakes (Hall et al., 2009a). In the LNWR the
middle 80% of wetland whole water THg samples were 1.60-8.71 ng L-1 and 0.11-1.62
ng L-1 for MeHg (Sando et al., 2007).
7. BIOTIC UPTAKE OF WATER COLUMN MERCURY
Mercury concentrations in aquatic biota, such as zooplankton (Hall et al., 2009b)
and fish (Chasar et al., 2009), may be positively correlated to water column THg and
18
MeHg concentrations. Uptake of MeHg from water by phytoplankton (enrichment of
105.5 fold) accounts for much of the enrichment of MeHg between water and fish
(enrichment of 106.5 fold; Mason et al., 1995). Methylmercury in the phytoplankton
Thalassiosira weissflogii was largely associated with the phytoplankton’s cytoplasm
(63%) rather than the membrane (37%) in contrast to inorganic Hg (9% associated with
cytoplasm and 91% with the membrane; Mason et al., 1995). A study of four
freshwater phytoplankton species found similar results with 59-64% of MeHg and 9-
16% of inorganic Hg associated with the cytoplasm (Pickhardt and Fisher, 2007).
Uptake of MeHg by phytoplankton is active since only 4.1% of MeHg was associated
with cytoplasm rather than the membrane in dead cells exposed to MeHg (Pickhardt
and Fisher, 2007).
Zooplankton preferentially digest cytoplasm rather than the membrane, thus
increasing their dietary uptake of MeHg relative to inorganic Hg (Mason et al., 1995).
Fish primarily take up MeHg through their food instead of directly from water (Hall et
al., 1997). Fish MeHg levels have been shown to decline in response to decreased
atmospheric deposition of Hg (Harris et al., 2007), although fish MeHg concentrations
in aquatic systems receiving Hg from adjacent uplands are likely to require more time
to decline in response to decreased Hg emissions as Hg transport from uplands is
relatively slow (Harris et al., 2007).
8. MERCURY AND WETLAND BIOTA
Methylmercury biomagnifies within wetland food webs. Methylmercury
concentrations of wetland invertebrates, which are an important food source for
vertebrates, vary by trophic level. Lower MeHg concentrations are reported in
19
Gastropoda compared to omnivorous and predatory invertebrates (Bates and Hall,
2012). Other biota connected to the wetland foodweb are also exposed to wetland
MeHg. Tree swallow (Tachycineta bicolor) eggs from seasonal wetlands at the LNWR
were observed to have higher Hg concentrations than eggs from semi-permanent
wetlands or lakes (Custer et al., 2008), similar to the pattern of sediment MeHg
concentrations by wetland type (Sando et al., 2007). Big brown bats (Eptesicus fuscus)
from near a Hg contaminated river in Virginia had significantly higher blood and fur
Hg than bats from an upstream reference site (Wada et al., 2010). Diet and trophic
position also influence MeHg accumulation in biota with prairie waterfowl total
mercury concentrations increasing with higher proportions of dietary animal matter
(Hall et al., 2009a). Primary producers may have a role in phytoremediation as
wetland macrophytes take up mercury from the wetland water column or sediments
(Patra and Sharma, 2000) and transgenic plants can reduce mercuric ions and
demethylate MeHg, respectively (Czarkó et al., 2006).
9. PRAIRIE POTHOLE REGION
The Prairie Pothole Region (PPR) in central North America covers an area of
approximately 850,000 square kilometres (Johnson et al., 2010) and contains 5-8
million wetlands (Voldseth et al., 2009). Wetlands in the PPR provide carbon storage
(Euliss et al., 2006) and valuable habitat for waterfowl populations (Niemuth and
Solberg, 2003). Wetland soils in Saskatchewan are often Gleysolic with soil parent
material containing less than 30% organic content and greater evapotranspiration than
precipitation (Bedard-Haughn, 2010). Prairie Gleysolic soils are usually from parent
material rich in iron and manganese (Bedard-Haughn, 2010). Clay-rich glacial tills
20
greatly limit hydraulic conductivity in the PPR, especially at depths greater than 4-5
metres (Van der Kamp and Hayashi, 2009). However, just below the soil surface
hydraulic conductivity is up to 1000 metres year-1, in contrast to less than 0.1 metres
year-1 in deeper soil (Van der Kamp and Hayashi, 2009). Groundwater flow in
relatively high hydraulic conductivity surface soil is important for wetland water
balance while groundwater flow in deeper low conductivity soil influences wetland
salinity (Van der Kamp and Hayashi, 2009).
Variability in the water depth of PPR wetlands controls wetland vegetation and
hydroperiod with depths of less than 0.5 metres in seasonal wetlands and 1.5-2 metres
in semi-permanent wetlands (Van der Valk, 2005). Snowmelt runoff is an important
source of water to wetlands, as there is little infiltration of water into frozen soil,
although small depressions can trap much of this runoff within the wetland catchment
(Hayashi et al., 2003). Wind can redistribute snow to wetlands after it is deposited
(Donald et al., 2011). As potential evaporation is usually higher than precipitation,
wetlands in the PPR receive little runoff from uplands during the growing season,
resulting in declining water levels during the summer (Van der Kamp and Hayashi,
2009). Wetlands may lose water through infiltration to the adjacent shoreline,
particularly in smaller wetlands which have a higher shoreline to pond ratio (Van der
Kamp and Hayashi, 2009). Changes in land use affect wetland water balance with
conversion of cropland to grassland decreasing the amount of blown snow delivered to
wetlands and reducing runoff to the wetland through increased infiltration of runoff
into the soil (Van der Kamp et al., 2003).
21
The hydrologic connectivity of a wetland in the landscape may influence its ion
concentration. Wetlands higher in the landscape potentially have a lower
concentration of ions, as a result of flow from the wetland to groundwater removing
solute from recharge wetlands, relative to discharge wetlands lower in the landscape
that receive groundwater flow (Van der Kamp and Hayashi, 2009). Ion concentrations
also change seasonally in response to the relative significance of infiltration to the
wetland margin and evaporation in the hydrology of the wetland (Waiser, 2006).
Dissolved organic carbon (DOC) concentrations in surface water from wetland ponds
increase from spring to fall and are often higher in wetlands with higher conductivity
(Waiser, 2006).
Sulphate is likely the dominant ion in northern PPR discharge wetlands (Heagle,
2008) and its cycling has the potential to influence MeHg cycling in wetlands. In
recharge wetlands, sulphate reduction rates may decline after spring, with sulphate
reduction removing more sulphate than infiltration from recharge wetlands (Heagle et
al., 2007). Two shallow prairie pothole lakes in North Dakota’s Cottonwood Lake
study area had maximum porewater sulphide and depleted porewater sulphate by
approximately 5 cm and 10 cm depths, respectively, suggesting that sulphate reducing
bacteria would be most active in the upper layers of sediment (Zeng et al., 2013). At
the same sites the proportion of reduced organic sulphur in sediment decreased
between April and June, which may release Hg from the sediment (Zeng et al., 2013).
High wetland productivity, sulphate concentrations, seasonally fluctuating redox
conditions, and flooding of terrestrial organic matter are potential factors that may
promote the production of methylmercury in Saskatchewan prairie wetlands, leading to
22
accumulation of Hg by aquatic invertebrates (Bates and Hall, 2012). Both THg and
MeHg have been measured in surface water, sediment, and biota from wetlands in the
PPR in two recent studies (Hall et al., 2009a; Sando et al., 2007). Surface water from
prairie wetland ponds in Saskatchewan were observed to have higher THg and MeHg
concentrations than prairie lakes (Hall et al., 2009a). Waterfowl from the PPR were
found to have a range of THg concentrations that varied with diet, although THg
concentrations in sampled tissue did not exceed mercury guidelines for human
consumption of fish (Hall et al., 2009a). Invertebrates from wetlands with organic
cultivation upland land use had higher mercury concentrations than invertebrates from
wetlands surrounded by grasslands or conventional cultivation (Bates and Hall, 2012).
Concentrations of MeHg were found to generally higher in seasonal wetlands, possibly
due to cycles of wetting and drying, and at intermediate sulphate concentrations (Sando
et al., 2007). Although sediment THg and MeHg concentrations in PPR wetlands have
been reported from prairie wetlands with varying permanence (Sando et al., 2007),
sediment Hg methylation potentials have not.
The measurement of potential Hg methylation rates in prairie wetland sediment is
important in determining factors that impact mercury accumulation in prairie wildlife.
These measurements would add to, and allow comparison with, existing research on
Hg cycling in aquatic systems. Depending on which factors are associated with MeHg
production it may be possible to reduce MeHg production and possibly wildlife
exposure to MeHg in managed wetlands. For example concentrations of anions, such
as sulphate, and divalent cations, which are known to influence MeHg production, may
vary greatly between recharge and discharge wetlands. If moderate concentrations of
23
sulphate promote MeHg production, avoiding the application of sulphur fertilizers
directly onto or, depending on the mobility of the sulphur fertilizer applied, adjacent to
wetlands with low sulphate concentrations may reduce MeHg production. If land use
influences MeHg production it may be possible to choose land uses to reduce MeHg
production in wetlands. Varying water levels in managed wetlands to cause vegetation
cover cycles valuable for waterfowl production (Murkin et al., 1997) may also promote
MeHg production due to wet-dry cycles and flooding of terrestrial organic matter.
Wetland managers may be able to balance management practices to meet wetland
habitat objectives and to limit wildlife exposure to MeHg.
10. OBJECTIVES
The study of prairie wetlands possessing different hydrology and land use may allow
us to identify factors associated with MeHg production and bioaccumulation. Our lab
has been examining THg and MeHg concentrations in surface water and sediments
from wetlands in the PPR of Saskatchewan. However, quantification of the
contribution of sediments to MeHg production and diffusive flux of MeHg to surface
water has not yet been attempted. The focus of this thesis is to examine mercury
cycling in prairie wetland sediments with emphasis on summer Hg methylation rate
potentials and wetland sediment THg and MeHg concentrations. Wetland
sediment/porewater partitioning coefficients and diffusive flux of MeHg between
porewater and surface water will also be examined. Additional variables including
water chemistry, sediment carbon content, and sediment water content may provide
context for the Hg cycling results. Primary questions to be addressed are: 1) What are
the rates of MeHg production in prairie wetland sediments? 2) Do rates of MeHg
24
production vary among prairie wetlands and, if so, are other measured variables
associated with the rates of MeHg production?
25
CHAPTER 2: Methylmercury Production in Prairie Wetland
Sediment
1. INTRODUCTION
Methylmercury (MeHg) is a neurotoxin that biomagnifies with trophic level,
resulting in deleterious effects in humans and wildlife that consume large amounts of
fish. Anthropogenic mercury (Hg) emissions, mainly via the burning of coal, have
substantially increased the amount of Hg cycling in the environment (Selin, 2009) and
emissions are not projected to decline by 2050 (Streets et al., 2009). Anthropogenic
Hg emissions in the form of Hg0 may be transported long distances in the atmosphere
and, after photooxidation, mercuric ions (HgII) may be deposited in systems remote
from initial sources (Phillips et al., 2011). Sediments of freshwater systems are a
substantial source of MeHg to surface water (Lehnherr et al., 2012a; Sellers et al.,
1996), where aquatic organisms may accumulate MeHg (Chasar et al., 2009; Edmunds
et al., 2012). Methylmercury concentrations in sediment depend on both the net rate of
Hg methylation (rate of Hg methylation less the rate of MeHg demethylation), as well
as the flux of MeHg to and from sediment. Sulphate reducing bacteria (SRB) and iron
reducing bacteria have been identified as Hg methylators in anaerobic freshwater
sediments (Fleming et al., 2006; Yu et al., 2012) and the rate of MeHg production in a
system depends in large part on both the availability of Hg to, and the metabolic
activity of, the methylating bacteria (Hintelmann, 2010). Thus, factors that influence
the bioavailability of Hg, such as pH (Kelly et al., 2003) and sulphur speciation (Benoit
et al., 1999), and the activity of methylating bacteria, such as temperature (Loseto et
al., 2004) and organic matter (Mitchell et al., 2008b), control MeHg production rates.
26
Wetlands have been identified as important sources of MeHg (St. Louis et al.,
1994), exporting substantially more MeHg than an equivalent area of upland. In
freshwater streams MeHg concentrations increased with increasing proportion of
wetlands in the stream catchment (Brigham et al., 2009). Mass balance budgets of
MeHg in Arctic ponds (Lehnherr et al., 2012a) and of MeHg exports from Swedish
boreal wetlands (Tjerngren et al., 2012b) generally indicate that wetlands are a source
of MeHg. Flooding of terrestrial organic carbon promotes MeHg production when
reservoirs are created (Hall et al., 2005), and cycles of wetting and drying may be
responsible for high MeHg concentrations in seasonal and semipermanent wetlands
(Sando et al., 2007). Cycles of wetting and drying have been hypothesized to allow for
oxidation of sulphur to sulphate as the wetland dries, providing a supply of sulphate to
anaerobic methylating bacteria when the wetland is later flooded (Sando et al., 2007).
Inundation of terrestrial organic carbon when wetlands are flooded could also promote
MeHg production.
The Prairie Pothole Region (PPR) of the North American Great Plains extends over
850 000 km2 (Johnson et al., 2010). This region is rich in small, depressional
wetlands, by some accounts 5-8 million wetlands, that provide carbon storage,
groundwater recharge, and valuable breeding habitat for North American waterfowl
(Euliss et al., 2006; Niemuth and Solberg, 2003; Voldseth et al., 2009). These wetland
systems have high hydrologic variability, with water levels that are typically highest
after snowmelt and decline over the growing season until the wetland dries up (Van der
Kamp and Hayashi, 2009). High hydrologic and chemical variability, combined with
shallow ponds and warm temperatures, may result in high rates of MeHg production.
27
In fact, concentrations of MeHg and %MeHg are high in the surface water and
sediment of some prairie wetlands in North Dakota and Saskatchewan (Hall et al.,
2009a; Sando et al., 2007), but rates of MeHg production have not yet been measured
in prairie wetlands.
The primary objective of this study was to measure potential Hg methylation rates
in prairie wetland sediment using in situ incubations of sediment cores injected with
enriched Hg stable isotope. These measurements have previously been made in
sediments from a variety of aquatic systems including freshwater (Lehnherr et al.,
2012b) and saltwater wetlands (Mitchell and Gilmour, 2008), streams (Marvin-
DiPasquale et al., 2009b), lakes (Gentès et al., 2013), and marine sediment
(Hammerschmidt and Fitzgerald, 2004). Studies from freshwater wetlands have
generally shown that there is potential for high MeHg production rates and
concentrations of MeHg in these systems (Lehnherr et al., 2012b; Tjerngren et al.,
2012a; Windham-Myers et al., 2009). The second objective of this study was to
examine the influence of environmental parameters, such as pH and conductivity and
sediment THg concentrations, organic content, and porosity, on rates of MeHg
production. We also measured THg and MeHg concentrations in surface water and
porewater in order to calculate diffusive flux of MeHg to wetland surface water. This
study is the first to report potential Hg methylation rates in prairie pothole region
wetland sediment, a process that must be understood to adequately evaluate the risk of
Hg accumulation in wildlife.
2. METHODS
2.1 Study Site
28
The 12 wetlands studied were within and close to the St. Denis National Wildlife
Area (SDNWA; 52° 12'N, 106° 5'W; Fig. 1). The 361 hectare SDNWA is located ~40
km east of Saskatoon, SK within the prairie pothole region near the boundary between
the Aspen Parkland and Moist-Mixed Grassland eco-regions (Environment Canada -
Canadian Wildlife Service, 2012). When the SDNWA was established in 1961, 68%
of the total area was cultivated with annual crops but since then, planting to perennial
grass has reduced the cultivated area to 34% of total area. Native and tame grasslands,
woodlands (Populus tremuloides, Salix spp.), and wetlands account for the remainder
of the habitat in the SDNWA (Environment Canada - Canadian Wildlife Service,
2012). Daily mean temperature from 1971-2000 at Saskatoon was -17.0°C in January,
18.2°C in July, and 2.2°C annually, with daily mean temperatures greater than 0°C
from April to October (Environment Canada, 2012). Over the same period, mean
annual precipitation was 350 mm with 265.2 mm as rainfall and 97.2 cm as snowfall
(Environment Canada, 2012). Inter-annual variation of precipitation results in
wetlands covering 1-22% of the SDNWA (Environment Canada - Canadian Wildlife
Service, 2012). Approximately 20 larger wetlands form a ring at the base of an upland
area 5 to 10 m above the lowland wetlands (Pennock et al., 2010). The upland area
contains ~40 smaller wetlands. In general, lowland wetlands are surrounded by
hayfields and grasses and upland wetlands are within the cultivated and grass area.
Topographic position may influence the hydrology of wetlands. Upland wetlands are
more likely to be groundwater recharge wetlands, whereas lowland wetlands are more
likely to be groundwater discharge wetlands (Lissey, 1968). The hydrology of
wetlands in the SDNWA is also influenced by the hydraulic conductivity of glacial till
29
Fig. 1: Map of wetlands sampled within the St. Denis National Wildlife Area (SDNWA) and three additional sampled wetlands northeast of the SDNWA.
1. Pond 100 2. Pond 103 3. Pond 110 4 Pond 113 5. Pond 118 6. OC 1 7. OC 2 8. OC 4 9. Pond 2 10. Pond 3 11. Pond 130 12. Pond 139
Canada
0
SDNWA Region
2
SDNWA
11
10
4 1 5 3
12
9
8
6
7
N
Scale
km
30
in the SDNWA which decreases for depths greater than >5 m resulting in slow deep
groundwater flow (Hayashi et al., 1998; van der Kamp and Hayashi, 2009).
Depending on annual and summer precipitation, timing of spring snow melt, and
temperature patterns, it is typical that these wetland ponds dry completely by the end of
the growing season; most likely due to infiltration to shallow groundwater and
evapotranspiration (Hayashi et al., 1998). During our sampling period, a number of
wetlands dried up including OC (organic cultivation) 4 and Pond 130 in October 2008,
Pond 100 and 110 in October 2010, and OC 2 in August and October 2011 (Table 1).
Sampling sites included nine wetlands within the SDNWA and three additional
wetlands less than 10 km northeast of the SDNWA. Five wetlands were located in
upland areas (Pond 100, Pond 103, Pond 110, Pond 113, Pond 118) and four in
lowland areas (Pond 2, Pond 3, Pond 130, Pond 139) within the SDNWA. Three
“Organic” wetlands (OC 1, OC 2, OC 4) were located in organically farmed fields
northeast of the SDNWA (Fig. 1). All wetlands, except OC 2, were classified as
shallow marshes (Type III ponds; Bates and Hall, 2012) which are seasonally flooded
(Stewart and Kantrud, 1971). Pond OC 2 was classified as an ephemeral basin (Type I
pond), which is temporarily or intermittently flooded early in the growing season
(Bates and Hall, 2012; Stewart and Kantrud, 1971).
Wetlands were sampled in October 2008, 2010, and 2011 for sediment THg and
MeHg concentrations, water content, and organic content (Table 1). In the summer of
2011, porewater and sediment samples were also taken from all twelve wetlands in
July and from eleven wetlands in August (OC 2 was dry). Surface water was sampled
at the end of June and August of each year. As well, sediment samples for THg and
31
Table 1: Latitude and longitude for wetlands sampled for sediment in 2008, 2010, and 2011. The (-) symbol indicates that the wetland was not sampled, (○) the wetland was sampled only for sediment total mercury (THg), methylmercury (MeHg), water content, and organic carbon, (●) the wetland was sampled for surface water and porewater in addition to sediment methylation potential, THg, MeHg, water content, organic carbon, and porosity, (D) the wetland was dry and if preceded by a sampling symbol indicates a dry sediment sample was taken from the wetland.
Sampling Trip
Position 2008 2010 2011
Pond Latitude Longitude Oct Oct July Aug Oct
100 52°12.605’ 106°04.852’ - ○D ● ● ○ 103 52°12.567’ 106°05.010’ - ○ ● ● ○ 110 52°12.486’ 106°05.278’ - ○D ● ● ○ 113 52°12.634’ 106°05.112’ - ○ ● ● ○ 118 52°12.587’ 106°05.044’ - ○ ● ● ○
OC 1 52°17.040’ 106°03.244’ ○ ○ ● ● ○ OC 2 52°17.642’ 106°04.014’ - ○ ● D ○D OC 4 52°17.511’ 106°04.521’ ○D ○ ● ● ○
2 52°12.883’ 106°05.271’ ○ ○ ● ● ○ 3 52°12.869’ 106°05.401’ - ○ ● ● ○
130 52°12.724’ 106°04.688’ ○D ○ ● ● ○ 139 52°12.859’ 106°05.036’ - ○ ● ● ○
32
MeHg concentrations and potential 201Hg methylation rates were analyzed for nine of
the wetlands sampled in July 2011 and eight of the wetlands sampled in August 2011.
2.2 Sample Collection
2.2.1 Sediment MeHg and THg Concentrations and Methylation Potentials
Wetland sediments collected in July and August 2011 were sampled for THg and
MeHg concentrations, potential methylation rates, water content, porosity, organic
content, and particle size. All cores sampled for THg and MeHg concentrations and
Hg methylation potential rates were collected in 30 cm long, 5.08 cm diameter acrylic
cylinders (Maljohn Plastics) with a beveled bottom edge and silicone septa spaced 1
cm apart in a row down the side of the cylinder. Cores were sealed with a #11 rubber
stopper. For the determination of potential methylation rates, three cores at each
wetland were taken from sediment submerged in 5-25 cm of water. These cores had an
organic sediment layer (excluding plant litter, sand, and clay) of at least 9 cm. Cores
with thick roots or more than 3 cm of sediment compression were discarded. Patches
of dense vegetation, as well as disturbed areas, were avoided. Methylation potential
rate cores were injected with 100 µl of a solution of porewater (see collection methods
below) and 201Hg stock solution prepared approximately 1-2 hours before injection.
The solution was composed of 5 mL of filtered porewater and 20 to 140 µL of 200 µg
mL-1 201Hg stock solution. The amount of solution injected into cores was determined
using THg concentrations and water contents of sediments sampled in 2010 with a
target of adding an amount of spike similar to ambient concentration. Injections of the
201Hg/porewater solution were 2-108% of ambient sediment THg (results from three
cores with very low additions are not reported). The 2010 cores contained vegetation
33
rich surface sediment, excluded from methylation potential rate cores, which lead to an
underestimate of the bulk density of the sediment and therefore the amount of isotope
required to match the ambient Hg concentrations. Starting from the surface layer of
sediment within the core, 100 µl of solution was injected into at least the top 10 cm of
sediment below the surface layer using a 100 µl borosilicate glass syringe and stainless
steel needle (Hamilton part #: 84859). The solution was injected over three separate
paths every 1 cm by inserting the needle into the sediment core and injecting
approximately one third of the solution each of the three times the needle was being
withdrawn from the core. As newly added 201Hg is likely more available for
methylation than ambient Hg, the resulting values represent potential methylation rates
rather than absolute rates of MeHg produced (Mitchell and Gilmour, 2008).
Cores were then returned to the wetland and positioned with the surface of the
water overlying the sediment inside the core approximately at or below the surface of
the water in the wetland. Incubation times of between 3:55 to 4:22 hours began after
the last injection and finished when the cores were sectioned. Near the end of the
incubation period, the incubated cores were collected from the wetland. Cores were
then sectioned into 0-2, 2-4, and 4-8 cm sections using an extruding stand and bread
knife, placed in 120 ml polypropylene containers, immediately frozen on dry ice, and
subsequently freeze dried. Freeze dried samples to be measured for potential
methylation rates were then homogenized using an acid-washed mortar and pestle,
transferred to 20 ml acid washed glass vials, and stored at room temperature until
analysis of isotopic and ambient MeHg and THg concentrations.
2.2.2 Sediment Water Content, Porosity, and Organic Content
34
An additional core for water content, porosity, and organic content was collected
from approximately the centre of the area that the three potential methylation rate cores
were taken from. This core was collected using the same methods as the potential
methylation rate cores except for the injection of 201Hg, incubation, and was
refrigerated instead of frozen after sectioning. We also collected sediment cores for
only THg and MeHg concentrations, water content, and organic content in October
2008, 2010, and 2011. These cores were collected using a 10 cm diameter corer and
the surface 0-3 cm section was kept for analysis in 2008, while in 2010 and 2011 the
surface 0-2 cm was kept for analysis. Sites with dense vegetation and disturbance were
avoided. For each wetland sampled, a wet core was taken from within the flooded
portion of the wetland and a dry core from exposed sediment that had been flooded
earlier in the growing season. No wet cores were collected from wetlands that were
entirely dry by October (Table 1). Sediment samples were bagged in polyethylene
zipper sealable bags, placed on ice in the field, and stored frozen until analysis.
2.2.3 Porewater MeHg and THg Concentrations
Porewater was used as a base of 201Hg solution injected into methylation cores. We
also determined THg and MeHg concentrations, temperature, oxygen, pH, and
conductivity in porewater. The day before the incubations a ten cm diameter acrylic
corer was used to create 10-20 cm deep holes slightly inland from the wetland.
Sampling holes were left overnight to allow water to accumulate and sediment
particles to settle. At that point, porewater samples were taken using a Series I
Geopump Peristaltic Pump (Geotech) with Teflon and silicone tubing from
approximately 1-2 cm below the water surface into pre-cleaned 125 mL glass bottles
35
with fluoropolymer resin lined lids. Using ultraclean handling techniques, a
perfluoroalkoxy filter cartridge containing a Whatman quartz fibre filter, muffled
overnight at 500°C, was assembled to filter porewater collected for THg and MeHg
concentrations. Five mL of filtered porewater was also collected in a 50 mL centrifuge
tube and used to make the 201Hg solution injected into the potential methylation rate
sediment cores. Samples for porewater THg concentration analysis were preserved
with trace metal grade HCl to 0.2% by volume and MeHg samples were preserved with
trace metal grade HCl to 0.4%; preserved samples were refrigerated until analysis.
Results for Pond 139 in July for MeHg concentration and %MeHg are not reported as
the filter was damaged while collecting the MeHg sample. Porewater temperature,
conductivity, pH, and dissolved oxygen were measured using a YSI 556 multiprobe
meter.
2.2.4 Surface Water MeHg and THg Concentrations
Surface water was sampled for THg and MeHg concentrations, temperature,
oxygen, pH, and conductivity. Surface water samples were collected by wading into
the wetland and sampling water from within the open water section approximately 5-
10 cm beneath the surface. Wetland surface water THg and MeHg samples were
collected in pre-cleaned glass bottles with fluoropolymer resin lined lids using ultra-
clean field sampling techniques (St. Louis et al., 1994). Care was taken to avoid large
particles in the sample. Surface water THg samples were preserved to 0.2% with trace
metal grade HCl and MeHg samples were preserved to 0.4% with trace metal grade
HCl and refrigerated until analysis. Surface water temperature, oxygen, pH, and
36
conductivity were measured using a YSI 556 meter at approximately the same site as
the surface water samples were collected.
2.3 Sample Analysis
2.3.1 Sediment MeHg and THg Concentrations and Methylation Potentials
Freeze dried sediment samples collected in 2011 for THg and MeHg concentrations
and methylation rate potentials were shipped to the Biogeochemical Analytical Service
Laboratory at the University of Alberta. Although triplicate sediment cores for THg
and MeHg concentrations and Hg methylation rates were collected and sectioned into
0-2, 2-4, and 4-8 cm depths, budgetary constraints limited analysis to duplicate 0-2 cm
sections from up to three wetlands selected from each of the upland, organic, and
lowland groups of wetlands. Freeze dried sediment was digested with nitric acid,
sulphuric acid, and BrCl before THg analysis. The solution was analyzed using a
Tekran 2600 connected to a PerkinElmer Elan DRC-e inductively coupled plasma
mass spectrometer (ICP-MS). Ambient THg concentration was measured using 202Hg
as the ratio of 202Hg to other stable isotopes of Hg in the spike and under ambient
conditions were known (Hintelmann and Evans, 1997). Sediment MeHg samples were
distilled, ethylated with sodium tetraethyl borate, and analyzed on a Tekran 2700
methylmercury analyzer connected to a PerkinElmer Elan DRC-e ICP-MS. Ambient
MeHg concentration was measured similarly to ambient THg concentrations, using
Me202Hg and Me199Hg as an internal standard. Detection limits were 1.13 ng g-1 for
THg and 0.011 ng g-1 for MeHg, for additional QA/QC data see Table 2 (Table 2).
Sediment THg and MeHg concentrations from samples collected in October 2008,
2010, and 2011 were analyzed using similar methods, except that the Tekran 2600 and
37
Table 2: Quality assurance and quality control for total mercury (THg) and methylmercury (MeHg) analysis of water and sediment samples with recovery of standard reference materials (SRM) or spiked samples ± one standard deviation. Approximately 10% of samples for all parameters were analyzed in duplicate. Water MeHg and October sediment MeHg spike recovery are reported only for spikes with a higher concentration than the sample. For water THg and October sediment THg minimum detection limits were less than 0.3 ng L-1, for water MeHg and October sediment MeHg 0.02 ng L-1, and for July and August sediment were 1.13 ng g-1 for THg and 0.01056 ng g-1 for MeHg.
Matrix and Analyte SRM %Recovery Spike
%Recovery
Summer sediment THg MESS-3 95 ± 2% 85 ± 4% Summer sediment MeHg IAEA-405 99.7 ± 0.1% October Sediment THg MESS-3 124 ± 8% 112 ± 10%
October Sediment MeHg 105 ±15% Surface and porewater THg MESS-3 95 ± 2% 92 ± 13%
Surface and porewater MeHg 107 ± 10%
38
2700 were not connected to an ICP-MS, with analysis for THg concentrations at the
University of Regina and MeHg concentrations at the University of Western Ontario.
Methylation potentials were calculated as the amount of spike 201Hg methylated to
Me201Hg over the duration of the incubation (Hintelmann et al., 2000):
201Hg methylation potential = -ln(1 - (Me201Hg / 201Hg)) / (incubation time * (1 day /
24 hours))
where Me201Hg is the excess Me201Hg (concentration of spiked stable isotope above
the ambient concentration) in ng g-1, 201Hg is the excess 201Hg in ng g-1 from the added
spike, and incubation time is the length of the incubation in hours. For three sites in
August (Pond 2, Pond 130, and Pond 139) the spike added to the cores was lower than
the detection limit. As a result, results for these three sites are not reported.
2.3.2 Surface and Porewater MeHg and THg Concentrations
Surface water and porewater THg concentrations were analyzed at the University
of Regina by CVAFS using a Tekran 2600 following EPA Method 1631 (U.S. EPA,
2002). All forms of Hg were oxidized to Hg(II) with BrCl, reduced with SnCl to
Hg(0), collected on gold traps, and thermally desorbed into argon before detection by
CVAFS. Surface and porewater MeHg concentrations were analyzed following EPA
Method 1630 (U.S. EPA, 2001). Surface and porewater samples were distilled at
127°C at the University of Regina on a Tekran 2750 after the addition of ammonium 1-
pyrrolidinecarbodithiolate and HCl to sample water. Samples were distilled until ~45
mL of sample had collected in the receiving vial. Distilled samples were shipped
overnight in a cooler with ice to the Biotron at the University of Western Ontario
where they were analyzed by CVAFS on a Tekran 2700 MeHg analyzer after
39
ethylation with sodium tetraethyl borate (U.S. EPA, 2001). The August surface water
sample from Pond 103 was below the MeHg detection limit and was assigned a value
of half the detection limit (0.01 ng L-1). Detection limits for THg and MeHg in water
samples were 0.3 ng L-1 and 0.02 ng L-1, respectively (Table 2).
2.3.3 Sediment Water Content, Porosity, and Organic Content
Sediment sections for water content, porosity, and organic content were processed
from a single sediment core. The entire sediment section was split into three
subsamples and each subsample was measured for wet weight, dry weight after drying
the samples to constant weight in an oven at 100°C, and loss on ignition after heating
the samples to 450°C in a muffle furnace for 8 hours. Water content was calculated as
the average %water content of the three subsamples of sediment:
%water content = ((WW – DW100) / WW) * 100
where WW is the wet weight (g) of the sediment and DW100 is the dry weight (g) of the
sediment. Organic content was calculated as the average %organic content of each of
the three sediment subsamples (Heiri et al., 2001):
%organic content = ((DW100 – DW450) / DW100) * 100
where DW100 is the dry weight (g) of the sediment after heating at 100°C, and DW450 is
the dry weight (g) after heating at 450°C. Porosity was calculated as the volumetric
water content of the sediment samples:
Porosity = (WW – DW100) / V)
where WW is the wet weight (g), DW100 is the dry weight (g) after heating at 100°C,
and V is the volume (cm3) of the sediment section.
40
2.3.4 Diffusive Flux of MeHg
Sediment porosity and concentrations of MeHg in surface and porewater were used
to determine the direction and magnitude of diffusive flux of MeHg between sediment
porewater and surface water. Diffusive flux of MeHg from porewater to surface water
was calculated as:
where Dw is the diffusion coefficient of MeHg in water at 25°C (1.2 * 10-5 cm-2 s-1)
(Rothenberg et al., 2008), φ is sediment porosity, θ is sediment tortuosity, Cw is the
concentration of MeHg in the surface water, Cpw is the concentration of MeHg in the
porewater, and ∆x is the distance in cm between surface and porewater (Holmes and
Lean, 2006). Sediment porosity was determined from the porosity of the 0-2 cm
section of sediment and ∆x was assigned a value of 1 cm. Sediment tortuosity was
calculated from sediment porosity (Boudreau, 1996) using:
θ2 = 1 – ln(φ2).
Partitioning coefficients (Kd), indicating the relative concentrations of THg and
MeHg in porewater and sediment, were calculated using THg and MeHg
concentrations in porewater and sediment. Log partitioning coefficients (log Kd) in L
kg-1 for THg concentrations in porewater and sediment were calculated as:
THg log Kd = log10 (THgsediment / THgporewater)
J = ( ) Dw * φ (Cw – Cpw) θ2 ∆x
41
where THgsediment is the concentration (ng kg-1) of THg in sediment and THgporewater is
the concentration (ng L-1) of THg in porewater. The calculation for the sediment
porewater MeHg partitioning coefficient is similar.
2.4 Statistical Analysis
Data with non-normal distributions were identified using Kolmogorov-Smirnov
tests and transformed when necessary. Associations between variables were quantified
through regression analysis. One-way analysis of variance (ANOVA) was used to
compare variables by wetland group; if the variable did not meet the assumptions of
normality or equal variance a Kruskal-Wallis one-way ANOVA by ranks was
performed. If a Kruskal-Wallis test indicated a significant difference Dunn’s method
was subsequently performed. Statistical significance was α = 0.05. SigmaPlot 10 was
used for statistical tests. Power of statistical tests was generally low so there is a
higher chance of type II error. Generally, variability around the mean is presented as
standard error.
3. RESULTS AND DISCUSSION
3.1 Sediment MeHg and THg Concentrations
Methylmercury and THg concentrations, as well as the proportion of THg that is
MeHg (%MeHg), in sediment may offer an indication of the amount of MeHg
produced in wetland sediments (Lehnherr et al., 2012b; Tjerngren et al., 2012a). The
range of MeHg concentrations in sediment cores from SDNWA wetlands was
generally similar in all sampling periods (0.08 to 2.99 ng g-1, p = 0.571; Fig. 2), but
differed among wetland types. Mean MeHg concentrations in upland wetland
sediments were generally lower than in organic and lowland wetlands and this trend
42
Fig. 2: (A) Sediment methylmercury (MeHg) concentrations, (B) total mercury (THg) concentrations, and (C) percent THg that is MeHg (%MeHg) from nine wetlands sampled in 2011. Wetland OC 2 was dry in August and October 2011. Two 0-2 cm depth sediment samples were analyzed per wetland. Error bars indicate one standard error.
Sediment MeHg (ng/g)
0
1
2
3
4
Sediment THg (ng/g)
0
20
40
60
80
100
120
July
August
October
Pond 103
Pond 113
Pond 118
OC 1
OC 2
OC 4
Pond 2
Pond 130
Pond 139
Sediment %MeHg (%)
0
2
4
6
8
10
12
14
A
B
C
Upland Organic Lowland
43
was consistent in July and August. Sediment MeHg concentrations of less than 1 ng g-
1 were found at all upland sites, whereas concentrations ranged from less than 1 ng g-1
to nearly 3 ng g-1 at organic and lowland sites, and were highest at OC 2 in July (Fig.
2). Although we did not see a temporal trend in MeHg concentrations, THg
concentrations were higher in October compared to other months, ranging from 18.3 to
96.6 ng g-1 with a mean of 37.9 ± 19.4 ng g-1 (p = 0.012; Fig. 2). Higher THg
concentrations in October samples were likely due to the inclusion of surface
vegetation, resulting in higher organic carbon content (see below), in cores taken in
October. Since there were no differences in sediment THg concentrations among the
wetlands studied we can assume that deposition of Hg was probably similar at all of
our sites. Sediment %MeHg in 2011 ranged from 0.18% to 10.3% and was lower in
October compared to July with intermediate %MeHg in August (p = 0.039; Fig. 2).
Concentrations of MeHg in SDNWA wetland sediments were generally similar to,
or lower than, those in other freshwater ecosystems. Our MeHg concentrations were
slightly lower than those in sediments from prairie wetlands in North Dakota’s
Lostwood National Wildlife Refuge (LNWR) which ranged from <0.4 to 4.16 ng g-1
(Sando et al., 2007), as well as those in wetlands in the Arctic (Ellesmere Island; 0.4 to
3.4 ng g-1; (Lehnherr et al., 2012b), northern boreal (Sweden; 3.5 to 21 ng g-1;
Tjerngren et al., 2012a), and near the St. Lawrence river (0.08 to 12.8 ng g-1; Holmes
and Lean, 2006). Concentrations of sediment MeHg were similar to sediments from a
riverine wetland in Quebec (~0.9 to ~1.7 ng g-1; Goulet et al., 2007) and freshwater
wetlands in Maine (0.14 to 2.22 ng g-1; Bank et al., 2007).
44
Summer sediment THg concentrations were lower than those from prairie pothole
wetlands in LNWR (6.77 to 99.0 ng g-1; Sando et al., 2007) and other freshwater
wetlands (Holmes & Lean 2006; Bank et al. 2007; Marvin-DiPasquale et al. 2009;
Tjerngren et al. 2012) and similar to those on Ellesmere Island (9.6 to 52 ng g-1;
(Lehnherr et al., 2012b). In October, THg concentrations were similar to those
sampled from the LNWR wetlands.
The %MeHg in sediments can be an indication of the potential for mercury
methylation (Lehnherr et al., 2012b; Tjerngren et al., 2012a). In our wetlands,
sediment %MeHg values were similar to those from wetlands at LNWR (0.3 to 8.4%;
Sando et al., 2007) and Ellesmere Island (1.2 to 12%; (Lehnherr et al., 2012b).
Wetlands from SDNWA typically had higher %MeHg than Florida Everglades
wetlands (0.1 to 1.7%; Gilmour et al., 1998) and Cornwallis Island Arctic wetlands
(0.01 to 1.04%; Loseto et al., 2004). The %MeHg was lower in SDNWA wetlands
than Swedish boreal wetlands (2.3 to 17%; (Tjerngren et al., 2012a). Sediment
%MeHg from some of the organic and lowland wetlands sampled was higher than the
range of %MeHg in marine sediment and sediment from saltwater wetlands (Table 3).
3.2 Sediment Hg Methylation Potentials
Potential rates of gross methylation of 201Hg added to wetland sediment cores
incubated in situ suggest that MeHg production in our systems is significant and
comparable to other freshwater wetlands. Sediment 201Hg methylation potentials (km)
ranged from 0.016 to 0.18 d-1 (Fig. 3). Methylation potentials did not significantly
differ either by month (p = 0.389) or by type of wetland (p = 0.259 and 0.628 for July
45
Tab
le 3
: M
ethy
lati
on p
oten
tial
(k m
(d-1
)) a
nd p
erce
nt o
f to
tal
mer
cury
as
met
hylm
ercu
ry (
%M
eHg)
for
fre
shw
ater
wet
land
, sa
ltw
ater
w
etla
nd, a
nd m
arin
e se
dim
ent.
Site
Type
km (d-1)
%MeH
g
Reference
SD
NW
A S
K
Fre
shw
ater
0.
016-
0.18
0.
6-10
T
his
stud
y L
NW
R N
D
Fre
shw
ater
n
o da
ta
0.3-
8.4
San
do e
t al
. 200
7 A
rcti
c F
resh
wat
er
0.04
-0.1
6 1.
2-12
L
ehnh
err
et a
l. 2
012
Flo
rida
Eve
rgla
des
Fre
shw
ater
0-
0.12
0.
1-1.
7 G
ilm
our
et a
l. 1
998
Yol
o B
ypas
s C
A
Fre
shw
ater
0.
075-
0.36
0.
8-0.
9 W
indh
am-M
eyer
s et
al.
200
9 B
orea
l S
wed
en
Fre
shw
ater
0.
011-
0.05
7 2.
3-17
T
jern
gren
et
al. 2
012
Pea
tlan
ds M
N
Fre
shw
ater
~
0-0.
05
No
data
B
ranf
ireu
n an
d K
rabb
enho
ft, a
s ci
ted
in M
itch
ell
and
Gil
mou
r 20
08
Con
nect
icut
S
altw
ater
~
0.00
1-0.
2
Lan
ger
et a
l. 2
001
San
Pab
lo B
ay
Sal
twat
er
0.01
4 ~
1.8
Mar
vin-
Dip
asqu
ale
et a
l. 2
003
Che
sape
ake
Bay
S
altw
ater
0.
002-
0.07
0.
2-4.
6 M
itch
ell
and
Gil
mou
r 20
08
San
Fra
ncis
co
Sal
twat
er
0.00
2-0.
122
0.1-
1.4
Win
dham
-Mey
ers
et a
l. 2
009
Lon
g Is
land
Sou
nd
Mar
ine
0.01
4-0.
082
0.4-
1.1
Ham
mer
schm
idt
and
Fit
zger
ald
2004
N
ew E
ngla
nd c
oast
al s
helf
M
arin
e 0.
02-0
.21
0.4-
1 H
amm
ersc
hmid
t an
d F
itzg
eral
d 20
06
Che
sape
ake
Bay
M
arin
e 0.
007-
0.04
5 0.
3-1.
6 H
ollw
eg e
t al
. 200
9
46
Fig. 3: Sediment mercury methylation potentials (d-1) for nine wetlands sampled in 2011. Two samples per wetland were analyzed. OC 2 was dry in August. Added 201Hg was below the detection limit in August in Pond 2, 130, and 139. Error bars indicate one standard error.
Pond 103
Pond 113
Pond 118
OC 1
OC 2
OC 4
Pond 2
Pond 130
Pond 139
Methylation Potential (d-1)
0.00
0.05
0.10
0.15
0.20
0.25
0.30
July
August
Upland Organic Lowland
47
and August, respectively). The range of km values from SDNWA wetland sediment
was generally similar to km values from other freshwater wetlands, saltwater wetlands,
and marine sediment (Table 3). Mean km values from SDNWA wetlands were most
similar to the values from surface sediments of freshwater Arctic wetlands (0.071 ±
0.060 d-1; Lehnherr et al., 2012b) and were greater than those in wetlands in the Florida
Everglades (0.02 d-1; Gilmour et al., 1998) and boreal wetlands in Sweden (Tjerngren
et al., 2012a). Values from wetlands at SDNWA were lower than freshwater wetlands
in California (0.15 ± 0.12 d-1; Windham-Myers et al., 2009). Methylation potentials
from freshwater SDNWA wetlands were generally greater than those from salt marsh
and marine sediment (with the exception of New England coastal shelf sediment
Hammerschmidt and Fitzgerald, 2006: Table 3). Comparison of Hg methylation rates
may be limited by the use of different methods or conditions including the amount of
Hg isotope added, temperature, length of incubation, and speciation of spike Hg
(Lehnherr et al., 2012b). Compared to other freshwater km studies, our methods were
most similar to those employed in the Florida Everglades (Gilmour et al., 1998) and
Arctic wetlands (Lehnherr et al., 2012b) with injection of Hg isotope after equilibration
with filtered porewater/pond water into intact cores with approximately four hour
incubations at in situ temperatures. Studies in wetlands from Sweden and California
incubated subsamples of sediment in a laboratory and in the Swedish study a 48 hour
incubation was used (Tjerngren et al., 2012a; Windham-Myers et al., 2009).
3.3 Factors Controlling km Values
Methylation potentials were not correlated with sediment MeHg concentrations;
nor were they correlated with most measured factors likely to influence methylation,
48
such as porewater and sediment THg concentrations, surface water conductivity,
surface water pH, and sediment organic content (see below). The absence of a
relationship between measured km values and sediment MeHg concentrations may be
due to variation in demethylation rates which were not measured. For example km
values were highest in Pond 113, but MeHg concentrations in Pond 113 were not
elevated, suggesting that the high km value observed may have been countered by
similarly high rates of demethylation. In Swedish freshwater boreal wetlands, the
mean ratio of sediment methylation and demethylation rate constants was 0.50 ± 0.29,
and the ratios were more indicative of %MeHg than km values alone because some
wetlands with high km values also had high demethylation potentials (Tjerngren et al.,
2012a). In our wetlands sediment %MeHg, like MeHg concentrations, were not
related to km values suggesting a role for demethylation in controlling %MeHg.
There was a negative relationship between km values and sediment porosity,
although it was only significant in July (p = 0.010) compared to August (p = 0.243;
Fig. 4). The negative relationship of km and sediment porosity may be a result of
deeper oxygen penetration in higher porosity sediment (Cai and Sayles, 1996)
inhibiting activity of anaerobic Hg methylating microbes. Methylation potentials in
organic sediment of Chesapeake Bay were highest in surface sediments as oxygen
penetration was limited to several millimetres, while in sandy sediments of the mid-
Atlantic continental shelf oxygen penetration was several centimetres and methylation
potentials were higher below the oxic surface sediment (Hollweg et al., 2009).
Porosity of our sediments was generally greater than in sandy marine sediments
(Hollweg et al., 2009). The organic content of wetland sediment from the SDNWA
49
Fig
. 4:
Rel
atio
nshi
ps b
etw
een
(A):
Jul
y su
rfac
e w
ater
met
hylm
ercu
ry (
MeH
g) a
nd p
oros
ity.
(B
): A
ugus
t ln
sur
face
wat
er M
eHg
and
poro
sity
. (C
): 20
1 Hg
met
hyla
tion
pot
enti
al (
d-1)
and
poro
sity
in
July
. (D
): A
ugus
t 20
1 Hg
met
hyla
tion
pot
enti
al (
d-1)
and
poro
sity
. U
plan
d si
tes
are
indi
cate
d by
tri
angl
es, o
rgan
ic s
ites
by
squa
res,
and
low
land
sit
es b
y ci
rcle
s.
Sediment Porosity
0.4
0.5
0.6
0.7
0.8
0.9
Surface MeHg (ng L-1)
0123456
Upland
Organic
Lowland
Sediment Porosity
0.4
0.5
0.6
0.7
0.8
0.9
Hg Methylation Potential (d-1)
0.00
0.02
0.04
0.06
0.08
0.10
0.12
0.14
0.16
0.18
Upland
Organic
Lowland
Sediment Porosity
0.56
0.58
0.60
0.62
0.64
0.66
0.68
0.70
0.72
0.74
0.76
Hg Methylation Potential (d-1)
0.00
0.02
0.04
0.06
0.08
0.10
0.12
0.14
0.16
0.18
0.20
Upland
Organic
Sediment Porosity
0.50
0.55
0.60
0.65
0.70
0.75
ln Surface MeHg (ng L-1)
-5-4-3-2-101
Upland
Organic
Lowland
A
B
C
D
r2 = 0
.35
r2 = 0
.64
50
(7.4 to 29.6%; Table 4) suggests high oxygen demand as carbon may be available to
promote microbial activity, reducing the oxygen penetration depth, although epipelic
phytoplankton may cause diel variation in the depth of oxygen penetration (Carlton
and Wetzel, 1987).
3.4 Porewater MeHg and THg Concentrations
Porewater MeHg and THg concentrations from our wetlands were generally within the
range of concentrations observed in other studies, although three wetlands sampled in
July had relatively high MeHg and THg concentrations (Fig. 5). Porewater MeHg
concentrations ranged from 0.19 to 55.6 ng L-1 (Fig. 5) and did not differ significantly
between months (p = 0.088). THg concentrations (range = 2.17 to 50.6 ng L-1) and
%MeHg (range = 4.60% - 110%) were significantly higher in July samples compared
to August (p = 0.020 for THg and 0.036 for %MeHg; Fig. 5). Concentrations of MeHg
were much higher in July in wetlands Pond 2, Pond 130, and OC 1, compared to the
other wetlands sampled (Fig. 5). The high concentrations of Hg and MeHg in the
porewater of some wetlands may have been related to a seasonal decline in the
proportion of sediment reduced organic sulphur, potentially releasing sediment Hg and
MeHg to the porewater. This was observed over the spring-summer transition in
shallow prairie lakes in North Dakota’s Cottonwood Lake study area (Zeng et al.,
2013) with sulphate concentrations similar to or higher than our discharge wetlands.
Porewater MeHg concentrations in Pond 2 (55.6 ng L-1) and Pond 130 (30.2 ng L-1)
exceeded the highest porewater MeHg concentration (10.2 ng L-1) measured from four
extensively surveyed Minnesota peatlands (Mitchell et al., 2008a). Other SDNWA
wetlands sampled in July were within the range of porewater MeHg concentrations
51
Tab
le 4
: S
urfa
ce w
ater
con
duct
ivit
y an
d pH
, sed
imen
t w
ater
con
tent
, por
osit
y, o
rgan
ic c
arbo
n, l
og s
edim
ent/
pore
wat
er t
otal
mer
cury
(T
Hg) a
nd m
ethy
lmer
cury
(M
eHg)
par
titi
onin
g co
effi
cien
ts (
L k
g-1),
and
MeH
g di
ffus
ive
flux
fro
m p
orew
ater
to
surf
ace
wat
er (
ng m
-2
day-1
) fo
r w
etla
nds
sam
pled
in
2011
. N
egat
ive
valu
es f
or M
eHg
diff
usiv
e fl
ux i
ndic
ate
diff
usiv
e fl
ux f
rom
sur
face
wat
er t
o po
rew
ater
. N
D i
ndic
ates
sit
es w
ith
no d
ata.
Pond
Conductivity
(mS cm
-1)
pH
Water
Content (%
) Porosity
Organic
Content (%
) THg K
d (L
kg-1)
MeHg K
d (L
kg-1)
MeHg diffusive
flux (ng m
-2 d
-1)
July
Aug
July
Aug
July
Aug
July
Aug
July
Aug
July
Aug
July
Aug
July
Aug
100
436
595
7.27
7.35
39.0
40.3
0.46
0.59
7.4
12.1
ND
ND
ND
ND
0.227
0.089
103
232
340
6.98
6.95
40.7
45.1
0.64
0.61
12.6
11.2
3.12
3.62
1.75
2.40
1.93
0.318
110
300
505
6.58
6.53
59.0
47.8
0.66
0.58
16.7
9.5
ND
ND
ND
ND
0.510
0.447
113
273
570
7.26
6.41
42.9
39.7
0.51
0.60
12.0
9.2
3.65
4.06
2.69
3.21 0.0937
0.0255
118
291
471
6.86
6.82
48.5
44.9
0.70
0.59
16.3
10.4
3.37
3.79
2.07
2.22
0.386
0.195
OC 1
827
1272
7.77
8.17
54.1
48.9
0.86
0.74
17.8
11.5
3.50
3.78
1.46
2.52
5.27
0.255
OC 2
2077
Dry
7.68
Dry
42.4
Dry
0.69
Dry
9.3
Dry
3.88
Dry
3.50
Dry
-0.940
Dry
OC 4
480
527
7.36
8.26
53.0
52.7
0.61
0.61
11.7
13.2
3.71
4.15
2.77
3.21
-0.482
-0.359
2
3888 4397
8.16
8.33
47.7
42.6
0.71
0.65
12.5
9.0
2.73
3.61 0.882 2.78
24.0
0.0705
3
2577 3152
7.24
7.73
55.8
53.7
0.70
0.71
17.3
18.6
ND
ND
ND
ND
-0.871
-0.314
130
750
1005
7.41
7.67
64.6
65.2
0.74
0.68
29.6
25.5
3.18
3.92
1.85
3.23
11.9
0.00036
139
1304 1808
7.10
7.79
59.2
35.8
0.73
0.52
17.5
5.5
3.03
3.38
1.21
2.54
21.4
0.089
52
Fig. 5: (A) Porewater methylmercury (MeHg) concentrations, (B) total mercury (THg) concentrations, and (C) percent of total mercury that is methylmercury (%MeHg) from twelve wetlands sampled in July and August 2011. Wetland OC 2 was dry in August 2011 and the filter was damaged during collection of the July MeHg sample from Pond 139. No error bars are reported because only one porewater sample was collected per analyte per site.
Pond 100
Pond 103
Pond 110
Pond 113
Pond 118
OC 1
OC 2
OC 4
Pond 2
Pond 3
Pond 130
Pond 139
Porewater %MeHg (%)
0
5
10
15
20
25
30
35
40
100
120
Porewater THg (ng/L)
0
5
10
15
20
25
30
50July
August
Porewater MeHg (ng/L)
0
1
2
3
4
5
6
7
8
9
10
11
12
30
40
50
Upland Organic Lowland
A
B
C
53
observed in the same Minnesota wetlands (Mitchell et al., 2008a) and similar to
brackish wetlands from southern Louisiana (2.15 ± 0.97 ng L-1; Hall et al., 2008).
August porewater MeHg concentrations were less variable among wetlands (Fig. 5)
and similar to southern Louisiana freshwater wetlands (Hall et al., 2008) and boreal
peatlands (Branfireun et al., 1999; Coleman Wasik et al., 2012; Mitchell et al., 2008a).
July THg concentrations and %MeHg were within the range of those observed in
Minnesota peatlands (Mitchell et al., 2008a). Concentrations of THg in porewater
samples from August were similar to brackish wetlands from southern Louisiana (5.58
± 0.82 ng L-1) and Minnesota peatlands (Mitchell et al., 2008a).
Porewater THg concentrations and %MeHg may be related to MeHg production as
concentrations of THg in porewater may be correlated with km values in marine
sediment (Hammerschmidt and Fitzgerald, 2004) and %MeHg has been used as an
indicator of methylation hotspots in peatlands (Mitchell et al., 2008a). However, our
porewater THg concentrations were not correlated with Hg methylation potentials.
Since Hg bioavailability depends on speciation (Benoit et al., 1999) it may be possible
that the lack of a correlation is due to differing speciation, and therefore availability, of
porewater THg among our sites. Porewater %MeHg concentrations were not
correlated with km values and the sites with the highest %MeHg in July had neither
high km values nor high August %MeHg suggesting that factors other than km values
may have been responsible for the high concentrations of MeHg in those wetlands.
3.5 Factors Potentially Influencing Sediment km Values and MeHg and THg
Concentrations in Wetlands
Variation of sediment MeHg concentrations and %MeHg in wetland sediments
may be explained in a large part by differing rates of Hg methylation, which are
54
dependent on both the activity level of Hg methylating microbes and the amount of Hg
available for methylation (Hintelmann, 2010). These are, in turn, influenced by
environmental factors including temperature, pH, redox potential, and the quantity and
quality of organic matter (Ullrich et al., 2001). Although we did not measure all these
parameters in the sediment, we have data (Table 4) that, taken together with past
research in the SDNWA, suggests that our wetlands span a wide range of values for
some of these environmental factors which may influence km values by altering the
availability of Hg(II) for methylation and the activity of methylating bacteria.
3.5.1 Organic Matter
Organic matter may influence the cycling and production of MeHg by supporting
microbial activity (Gilmour and Riedel, 1995) and through binding of Hg and MeHg to
organic matter (Ullrich et al., 2001). For example, Hg methylation in the sediment of
some freshwater lakes with sandy sediments containing little (<0.5%) organic matter
was limited by organic matter rather than sulphate (Gilmour and Riedel, 1995).
Devegetation of wetlands has been reported to reduce concentrations of porewater
acetate and potential Hg methylation rates suggesting that organic carbon, as acetate,
supplied by the roots of aquatic plants may stimulate microbial activity and MeHg
production (Windham-Myers et al., 2009).
In addition to its role as a carbon source, dissolved organic matter, particularly
from terrestrial sources, has been observed to stabilize HgS against aggregation into
less available particles increasing their bioavailability to sulphate reducing bacteria
(Graham et al., 2012). Sediment organic content was high (5.6 to 60.1%) in our
wetlands. If HgS is available to Hg methylating bacteria in prairie wetland sediment,
55
high concentrations of organic carbon may promote greater production of MeHg than
would otherwise be expected. This might be particularly evident in wetlands with high
sulphate concentrations where sulphide produced via sulphate reduction could limit Hg
methylation. Surface water dissolved organic carbon concentrations ranged between
14.5 to 179 mg L-1 from samples taken from SDNWA wetlands during 2006-2009
(Hall et al. in prep). Higher concentrations of DOC were observed in wetlands with
higher conductivity (Waiser, 2006).
Positive correlations observed between sediment THg and organic content in
wetland (Marvin-DiPasquale et al., 2009a) and stream sediment (Marvin-DiPasquale et
al., 2009b) suggest that organic matter may bind Hg (Liao et al., 2009). Our sediment
THg concentrations were positively correlated with sediment organic content
suggesting that sediment THg in our wetlands is bound to organic carbon. We also
suspect that lower THg concentrations in our cores taken in summer 2011 compared to
October 2008, 2010, and 2011 were due to the exclusion of surface sediments with
excessive vegetation in summer cores. Indeed, the resulting mean sediment organic
content of cores taken in summer 2011 (13.8%) was lower than those taken in October
of 2008, 2010, and 2011 (32.8%). Organic content may influence sediment MeHg
concentrations (Hammerschmidt and Fitzgerald, 2006; Rothenberg et al., 2008).
Sediment MeHg concentrations were positively correlated with sediment organic
content in August 2011 (r2 = 0.470, p = 0.036; Fig. 6). However, this positive
correlation was dependent on a single data point, the high sediment MeHg
concentration and organic content in the wetland Pond 130 (Fig 2, Table 4).
3.5.2 Divalent Cations
56
Fig. 6: Sediment methylmercury concentrations and organic content in August 2011. Triangles represent upland wetlands, squares wetlands within organically farmed fields, and circles lowland wetlands.
Sediment Organic Content (%)
0 5 10 15 20 25 30
Sediment MeHg (ng g-1)
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
Upland
Organic
Lowland
r2 = 0.55
57
Increasing concentrations of divalent cations reduce uptake of Hg(II) by E. coli, a
gram-negative bacteria capable of Hg methylation (Daguené et al., 2012). If divalent
cations also reduce Hg(II) uptake by gram-negative sulphate reducing bacteria, km
values may decline at higher concentrations of divalent cations with Hg+2 accumulation
in E. coli declining by more than 60% with the addition of either approximately 2.2 mg
L-1 of Ca+2 or 2.4 mg L-1 of Mg+2 (Daguené et al., 2012). Surface water Ca+2 and Mg+2
concentrations from an upland recharge wetland (Ca+2: 28 to 52 mg L-1; Mg+2: 11 to 19
mg L-1) were lower than in a lowland discharge pond (Ca+2: 8 to 431 mg L-1; Mg+2: 10
to 1049 mg L-1) at the SDNWA (Ponds 109 and 50, respectively; Heagle et al., 2007;
Heagle, 2008). Although concentrations of divalent cations were not measured in our
wetlands, surface water conductivity may be positively correlated with concentrations
of anions and divalent cations in prairie wetlands. Mean surface water conductivity
during sediment sampling trips in summer 2011 was 401 ± 82 µS cm-1 for upland
wetlands, 1210 ± 799 µS cm-1 for organic wetlands, and 2360 ± 1446 µS cm-1 for
lowland wetlands. In our wetlands surface water and porewater conductivity were
positively correlated in both months with r2 values greater than 0.8 and p-values <
0.001.
If Hg uptake by the methylating community in prairie wetland sediments was
reduced with increasing concentrations of divalent cations as observed with E. coli
(Daguené et al., 2012) then there is potential for substantial inhibition of Hg+2 uptake.
However, our km values did not differ by wetland type and showed no relation to
surface water conductivity. Conductivities, and thus, cation concentrations are high
compared to values used in Daguene et al (2012), so it is possible that the
58
concentrations of divalent cations in all of our wetlands were sufficient to inhibit
uptake of Hg+2 resulting in similar km values by wetland type. An additional possibility
is that other factors, such as sulphur cycling and organic matter, countered the potential
effect of divalent cations on km values (see below).
3.5.3 Sulphur and Iron Cycling
Since sulphate is required as an electron acceptor by SRB (Gilmour et al., 1992),
high concentrations of sulphide may inhibit Hg methylation (Gilmour et al., 1998), and
reduced organic sulphur may bind Hg in the sediment thus reducing its bioavailability
to methylating organisms (Zeng et al., 2013). Sufficient sulphate concentrations must
be present to support sulphate reduction, and thus Hg methylation, without the
accumulation of high sulphide concentrations (Gilmour et al., 1998). As sulphide
concentrations increase, the bioavailabilty of Hg may decline because the
concentration of the bioavailable neutrally charged HgS0 complex declines with
increasing sulphide concentration (Benoit et al., 1999). Studies in salt marsh sediments
with porewater sulphide concentrations of ~0 to ~4 mg L-1 have shown that high
sulphide can be negatively correlated with methylation potential (Mitchell and
Gilmour, 2008). In shallow prairie lakes, sulphate may be depleted within the surface
10 cm of sediment because of SRB activity while sulphide concentrations, which
ranged from ~0 to ~77 mg L-1, may reach their maximum after the first 5 cm of surface
sediment due to accumulation of sulphide produced by the reduction of sulphate (Zeng
et al., 2013).
Surface water sulphate concentrations in ponds at the SDNWA from 2006 to 2009
ranged from 0.04 to 2.40 mg L-1 in upland wetlands, 9.45 to 1010 mg L-1 in organic
59
wetlands, and 2.50 to 2080 mg L-1 in lowland wetlands and were comparable to surface
water sulphate concentrations from the discharge (164 to 5952 mg L-1) and recharge
(6.08 to 36 mg L-1) wetlands studied by Heagle et al. (2007, 2008). Porewater sulphate
concentrations in the discharge wetlands were also high (1287 to 5484 mg L-1) in the
surface 0-5 cm sediment section and appeared to increase with depth (Heagle, 2008).
Surface water conductivity was not correlated with km values in our wetlands, although
past sulphate concentrations from SDNWA suggest that low sulphate concentrations
may be limiting for km values in upland recharge wetlands and sulphide accumulation
may reduce km values in some organic and lowland discharge wetlands with high
sulphate concentrations. Surface water sulphate concentrations were not correlated
with km values in Ellesmere Island wetlands, possibly as most of the studied wetlands
were within a range of optimal sulphate concentrations for MeHg production (Lehnherr
et al., 2012b). Similarly, correlations between sulphate and km values have not been
reported in other freshwater wetlands (Gilmour et al., 1998; Tjerngren et al., 2012a;
Windham-Myers et al., 2009).
The annual water balance of prairie wetlands, with maximum water levels
following snowmelt and loss of water over the growing season to infiltration and
evaporation (Hayashi et al., 1998), may also influence sulphur cycling and MeHg
production in prairie wetlands through changes to redox conditions. Heagle et al.
(2007) found that sulphate reduction was greatest in May and early June in a recharge
wetland at SDNWA and that sulphide remains in the sediment until oxidation to
sulphate as the wetland dries late in the growing season; in spring the sulphate may be
dissolved when the wetland floods. This suggests that wetlands with wet/dry cycles
60
are likely to have sulphate available to Hg methylating SRB at the beginning of the
growing season while sulphide may limit Hg methylation in wetlands with relatively
static water levels. At the LNWR higher MeHg concentrations were observed in
seasonal and semi-permanent prairie wetlands that had shorter hydroperiods,
suggesting that wet/dry cycles were impacting methylation rates. Maximum surface
water MeHg concentrations at the LNWR occurred at sulphate concentrations from 5
to 280 mg L-1 with lower MeHg in wetlands with higher and lower sulphate
concentrations (Sando et al., 2007). Since our wetlands, with the exception of OC 2,
are all seasonal wetlands wet/dry cycles may influence MeHg production in our
wetlands.
Although we did not attempt to identify the roles of SRB and iron reducing bacteria
in MeHg production, other research has indicated that iron reducing bacteria may be
responsible for a substantial proportion of Hg methylation in freshwater systems
(Fleming et al., 2006; Yu et al., 2012). If iron reducing bacteria make a substantial
contribution to km values, MeHg production could be influenced by iron concentrations
which may reduce the availability of Hg via iron sulphide immobilization of Hg in
sediment (Xiong et al., 2009).
3.5.4 pH
Uptake of Hg(II) by bacteria in freshwater systems increases at lower pH (Kelly et
al., 2003), suggesting that km values could be negatively correlated with pH. In 2011,
mean surface water pH was 6.90 ± 0.29 in upland sites, 7.82 ± 0.15 in organic sites,
and 7.68 ± 0.38 in lowland sites, which suggests that pH may be a factor in differing km
values between upland sites compared to organic and lowland sites and that pH was
61
generally circumneutral to alkaline. In our wetlands km values were not found to be
correlated with pH, although the pH range in our wetlands may have been too narrow
to observe significant effects.
3.6 Partitioning Coefficients
Partitioning coefficients (KD) are an indication of the amount of Hg sorbed onto
soil or sediment particles compared to that dissolved in porewater. Lower
sediment/porewater KD indicate a relatively higher proportion of Hg in porewater than
in sediment. Sediment/porewater (KD) values may be negatively correlated with km
values because Hg present in porewater may be more bioavailable than Hg bound to
sediment (Hammerschmidt and Fitzgerald, 2004). Sediment/porewater KD values for
MeHg and THg ranged from log10 0.88 to 3.50 L kg-1 and 2.73 to 4.15 L kg-1 (Table 4)
and were significantly lower in July compared to August for both MeHg (p = 0.037)
and THg (p = 0.025). The lower KD for both MeHg and THg in July compared to
August is likely due to the higher MeHg and THg values in July porewater since
sediment MeHg and THg concentrations were relatively constant between months.
Partitioning coefficients from SDNWA were generally lower than those in other
freshwater systems indicating a relatively higher proportion of MeHg and THg in
porewater than in sediment in our sites. Partitioning coefficients were lower than in
freshwater streams in Oregon, Wisconsin, and Florida that had varying degrees of
basin wetland cover and urban influence (Marvin-DiPasquale et al., 2009b) and
generally lower concentrations of sediment and porewater THg and MeHg. Our values
were also lower than freshwater sediment from Lake 239 at the Experimental Lakes
Area in northwestern Ontario, Canada (Hintelmann and Harris, 2004), even though the
62
lake sediment had relatively low (3.8%) organic content. Partitioning coefficients were
not correlated with km values in SDNWA wetlands. As concentrations of bioavailable
Hg(II) may decline with increasing sulphide while THg concentrations remain
relatively constant (Benoit et al., 1999), the lack of a correlation between THg
concentrations and km values may be due to THg concentrations not accurately
reflecting the quantity of Hg bioavailable for methylation.
3.7 Surface Water MeHg and THg Concentrations
Methylmercury produced in sediment may be a significant source of MeHg to
surface water and the resulting concentrations in surface water MeHg can predict
concentrations of MeHg in aquatic biota (Chasar et al., 2009). Surface water MeHg
concentrations in our wetland ponds ranged from below the detection limit to 5.72 ng
L-1 which was high compared to some freshwater systems (Hintelmann, 2010), but
similar to other freshwater wetlands. Total Hg concentrations and %MeHg ranged
from 0.61 to 6.82 ng L-1and from below detection limit to 113%, respectively (Fig. 7).
There were significantly higher MeHg (p = 0.004) and THg (p < 0.001) concentrations
in surface water collected in July compared to August, although %MeHg was not
significantly different (p = 0.059). Organic and lowland wetlands appeared to have
higher MeHg concentrations than upland wetlands (Fig. 7). Surface water THg was
not significantly different by wetland type. Organic and lowland wetlands had
significantly higher surface water %MeHg than upland wetlands. Wetland type could
influence surface water MeHg concentrations and %MeHg as differences in
topographic position may influence hydrology, in turn leading to differing
concentrations of solutes, such as divalent cations and sulphate, between recharge
63
Fig. 7: (A) Surface water methylmercury (MeHg) concentrations, (B) total mercury (THg) concentrations, and (C) percent of total mercury that is methylmercury (%MeHg) from twelve wetlands sampled in July and August 2011. Wetland OC 2 was dry in August 2011. No error bars are reported as one sample for MeHg and one sample for THg were collected per site.
Pond 100
Pond 103
Pond 110
Pond 113
Pond 118
OC 1
OC 2
OC 4
Pond 2
Pond 3
Pond 130
Pond 139
Surface Water %MeHg
0
20
40
60
80
100
120
Surface Water MeHg (ng/L)
0
1
2
3
4
5
6
7
Surface Water THg (ng/L)
0
2
4
6
8
July
August
Upland Organic Lowland
A
B
C
64
and discharge wetlands (Heagle, 2008; Heagle et al., 2007), although other factors such
as land use, vegetation, and wetland size may be responsible for the influence of
wetland type. Surface water MeHg was positively correlated with sediment porosity in
July (p = 0.042), but not August (p = 0.881; Fig. 4). Surface water MeHg was also
positively correlated with sediment MeHg in July (p = 0.014), but not with sediment
MeHg in August (p = 0.114; Fig. 8). These correlations suggest that sediment MeHg
may be transported to the water column as higher sediment porosity would lead to a
greater rate of MeHg diffusive flux.
Both MeHg (<0.02 to 9.56 ng L-1) and THg (0.61 to 6.82 ng L-1) concentrations
from surface water in SDNWA ponds were within the range of concentrations from the
LNWR (Sando et al., 2007) and similar to other freshwater wetlands, including those
in Saskatchewan (Hall et al., 2009a), southern Louisiana (Hall et al., 2008),
northeastern North America (Bank et al., 2007; Edmunds et al., 2012; Holmes and
Lean, 2006), and the high Arctic (Lehnherr et al., 2012b). Values of %MeHg were
also similar to Saskatchewan wetlands (1.44 to 64.5%; (Hall et al., 2009a), Louisiana
wetlands (6.4 to 84.7%; Hall et al., 2008), seasonal wetlands in the LNWR (0.06 to
55.6%; Sando et al., 2007), and Arctic wetlands (<1.5 to 52.6%; (Lehnherr et al.,
2012b). The %MeHg from SDNWA wetlands was higher than freshwater wetlands in
northeastern North America (Bank et al., 2007; Edmunds et al., 2012).
3.8 Diffusive Flux of MeHg
Methylmercury produced in wetland sediment may influence MeHg concentrations in
the overlying water via flux to the water column. The positive correlation between
65
Fig. 8: (A) Surface water methylmercury (MeHg) was correlated (p = 0.014) with ln 0-2 cm section sediment MeHg concentrations in nine wetlands sampled in July 2011. (B) ln August 2011 surface water MeHg concentrations were not correlated with sediment MeHg (p = 0.114). Upland sites are indicated by triangles, organic sites by squares, and lowland sites by circles.
-2.0 -1.5 -1.0 -0.5 0.0 0.5 1.0
Surface Water MeHg (ng L-1)
0
1
2
3
4
5
6
Upland
Organic
Lowland
ln Sediment MeHg (ng g-1)
-2.0 -1.5 -1.0 -0.5 0.0 0.5 1.0
ln Surface Water MeHg (ng L-1)
-5
-4
-3
-2
-1
0
1
Upland
Organic
Lowland
A
B
r2 = 0.60
66
wetland sediment MeHg and surface water in July may be due to transfer of MeHg
produced in wetland sediment to surface water. Mass balances of MeHg have also
indicated that sediments are a source of MeHg to the water column in a northwestern
Ontario freshwater lake (Sellers et al., 1996) and Arctic wetlands (Lehnherr et al.,
2012a). In fact, diffusive flux of MeHg, indicated by positive flux values, was a source
of MeHg to surface water in the majority of our wetlands (Fig. 9). Diffusive fluxes of
MeHg from porewater to surface water in our sites ranged from -0.94 to 24.0 ng m-2
day-1. July and August MeHg diffusive fluxes were not significantly different (p =
0.067). Mean diffusive flux was elevated, relative to our other wetlands, due to the
high porewater MeHg concentrations from wetlands OC 1, Pond 2, and Pond 130 in
July. Although diffusive fluxes were negative in OC 2, OC 4, and Pond 3 in July and
OC 4 and Pond 3 in August (Table 4) due to a combination of relatively high surface
water MeHg and low porewater MeHg, the calculated diffusive fluxes from all of our
wetlands suggest that porewater is usually a source of MeHg to prairie wetland surface
water. Additionally, diffusive flux of MeHg may be a small fraction of total MeHg
flux and bioirrigation may increase total flux of MeHg to the water column (Benoit et
al., 2009) and so, even in our wetlands where a negative MeHg diffusive flux was
calculated, the sediments may be a source of MeHg to the water column if total flux of
MeHg is considered.
Flux of sediment produced MeHg to surface water is likely countered by
degradation of MeHg in the surface water. Photodemethylation of surface water MeHg
is a significant factor in the mass balance of MeHg within wetlands (Lehnherr et al.,
2012a). Demethylation may also occur within sediment and studies of freshwater
67
Fig. 9: Calculated diffusive flux of MeHg between porewater and surface water. The upper two panels include the sites sampled within the St. Denis National Wildlife Area (SDNWA) while the lower two panels include wetlands surrounded by organically farmed fields northeast of the SDNWA. Red circles indicate movement of MeHg from porewater to surface water and yellow circles indicate MeHg transfer from surface water to porewater.
>0 >10 >100 >1000
Diffusive Flux pg/m2/day
July August
Red = Flux to surf. water
68
wetland sediment have measured potential rates of Hg methylation and demethylation
of MeHg (Lehnherr et al., 2012a; Tjerngren et al., 2012a). Since net Hg methylation
depends on the gross rates of Hg methylation and MeHg demethylation, both rates of
methylation and demethylation may influence the proportion of MeHg in sediment
(Tjerngren et al., 2012a). If rates of demethylation are relatively similar, %MeHg may
be related to Hg potential methylation rates (Lehnherr et al., 2012b).
3.9 Conclusion
Our research on MeHg production and cycling within prairie wetlands, as well
exploration of environmental parameters that potentially control the production and
cycling of MeHg, shows that rates of MeHg production are considerable in prairie
wetland sediment and that sediment porosity may be negatively correlated with rates of
MeHg production. Concentrations of MeHg and THg in the sediments of our wetlands
were similar to those measured in other prairie wetlands in the LNWR. Sediments
were typically a source of MeHg to surface water with a positive correlation between
sediment and surface water MeHg concentrations and diffusive flux of MeHg to the
water column in the majority of our wetlands. Future research could study potential
impacts of managing wetland water levels to promote vegetation cover cycles and the
use of sulphur fertilizers on MeHg production in prairie wetlands. Future research in
prairie wetlands on removal of MeHg through demethylation in the sediments may aid
in understanding net methylation and the factors leading to differences in MeHg
concentrations among wetlands. Studying photodemethylation of MeHg in the water
column of prairie wetlands, particularly if done as part of a mass-balance approach,
would also further understanding of mercury cycling in prairie wetlands.
69
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80
Appendices:
Appendix A: Surface water temperature, oxygen concentration, and time sample collected for wetlands sampled in July and August 2011. OC 2 was dry in August.
Pond Temperature (°C) Oxygen (mg L-1) Time
July August July August July August
100 13.87 15.68 10.63 8.54 13:51 12:50
103 11.96 13.73 11.9 4.5 14:13 12:15
110 11.4 13.24 5.28 5.11 15:17 11:30
113 16.52 13.16 11.55 5.22 14:32 11:45
118 12.39 14.65 9.12 3.56 14:55 12:30
OC 1 14.42 21.76 7.49 9.27 16:37 14:15
OC 2 18.79 Dry 9.44 Dry 17:07 Dry
OC 4 16.77 19.7 6.24 9.63 17:38 14:45
2 14.5 24.17 11.54 9.87 12:08 15:50
3 12.57 19.94 6.42 9.09 11:23 15:20
130 14.74 20.38 8.22 7.33 13:29 16:20
139 14.27 20.71 7.71 7.59 12:40 16:00
81
App
endi
x B
: P
orew
ater
tem
pera
ture
, oxy
gen
conc
entr
atio
n, c
ondu
ctiv
ity,
pH
and
tim
e sa
mpl
e co
llec
ted
for
wet
land
s sa
mpl
ed i
n Ju
ly
and
Aug
ust
2011
.
Pond
Temperature
(°C)
Oxygen
(mg L
-1)
Conductivity
(mS cm
-1)
pH
Tim
e
July
August
July
August
July
August
July
August
July
August
100
16.94
16.03
3.44
3.54
1.042
0.721
6.97
6.89
8:30
9:10
103
19.26
19.45
4.98
4.17
1.465
0.953
6.77
6.84
17:32
18:01
110
23.03
14.93
6.42
6.07
0.989
0.653
6.72
6.99
16:45
8:10
113
16.68
16.58
3.21
5.09
0.968
0.997
6.83
6.76
9:16
17:45
118
18.3
15.62
5.12
4.24
2.065
0.759
6.83
6.85
18:08
8:50
OC 1
21.76
17.85
3.44
3.43
1.275
1.509
6.8
6.99
20:55
12:00
OC 2
19.87
Dry
3.35
Dry
2.912
Dry
6.88
Dry
15:43
Dry
OC 4
21.57
16.63
4.37
3.4
1.649
0.757
6.72
6.95
20:12
11:35
2
16.94
16.47
3.64
4.81
4.17
3.327
7.05
7.01
10:20
9:45
3
18.63
15.13
1.97
2.98
2.894
4.547
7.24
7.17
10:48
10:05
130
20.59
17.99
4.64
3.16
1.918
1.059
6.73
7.34
20:49
10:50
139
20.31
15.28
5.23
3.17
2.078
1.869
6.67
7.11
21:28
10:26
82
App
endi
x C
: S
edim
ent
tem
pera
ture
, wat
er c
onte
nt i
n 2-
4 cm
and
4-8
cm
sec
tion
s, a
nd o
rgan
ic c
onte
nt i
n 2-
4 cm
and
4-8
cm
sec
tion
s fr
om J
uly
and
Aug
ust
2011
.
Pond
Temperature (°C
) Water Content
2-4 cm (%)
Water Content
4-8 cm (%)
Organic Content
2-4 cm (%)
Organic Content
4-8 cm (%)
July
August
July
August
July
August
July
August
July
August
100
19
15
37.9
32.5
43.5
30.5
8.0
8.1
10.5
7.5
103
17
14
40.8
43.9
39.4
45.3
11.0
11.1
9.7
11.6
110
17
13
37.4
34.9
34.2
30.3
7.0
6.3
6.5
5.4
113
20
13
34.8
51.3
32.2
45.0
7.8
14.0
6.5
10.6
118
19
15
35.4
42.8
32.7
37.0
8.1
10.0
8.1
8.6
OC 1
19
18
42.2
40.7
34.0
38.1
11.3
8.7
7.7
8.7
OC 2
18
Dry
37.2
Dry
35.2
Dry
8.8
Dry
9.4
Dry
OC 4
19
17
52.8
47.8
52.4
53.5
12.8
12.5
14.4
14.3
2
22
17
46.3
38.3
35.3
36.7
12.4
8.2
7.7
9.2
3
21
16
53.0
53.9
50.6
55.0
15.8
19.5
14.4
19.5
130
20
16
62.5
58.2
62.2
54.3
28.2
19.6
24.7
17.3
139
21
17
50.4
40.0
44.6
59.8
15.3
6.8
13.3
22.0
83
App
endi
x D
: S
edim
ent
poro
sity
and
bul
k de
nsit
y in
0-2
cm
, 2-4
cm, a
nd 4
-8 c
m s
ecti
ons
of s
edim
ent
from
Jul
y an
d A
ugus
t 20
11.
Pond
Porosity 0-2
cm
Porosity 2-4
cm
Porosity 4-8
cm
Bulk Density 0-2
cm (g cm
-3)
Bulk Density 2-4
cm (g cm
-3)
Bulk Density 4-8
cm (g cm
-3)
July August
July August
July August
July
August
July
August
July
August
100
0.46
0.59
0.58
0.32
0.67
0.51
0.72
0.88
0.95
0.66
0.86
1.15
103
0.64
0.61
0.59
0.59
0.58
0.71
0.93
0.74
0.86
0.76
0.90
0.85
110
0.66
0.58
0.54
0.50
0.48
0.54
0.43
0.64
0.91
0.93
0.93
1.25
113
0.51
0.60
0.43
0.62
0.53
0.69
0.68
0.91
0.80
0.58
1.13
0.85
118
0.70
0.59
0.55
0.52
0.57
0.52
0.74
0.72
1.01
0.70
1.18
0.89
OC 1
0.86
0.74
0.62
0.57
0.51
0.64
0.76
0.77
0.84
0.83
1.00
1.05
OC 2
0.69
Dry
0.56
Dry
0.65
Dry
0.93
Dry
0.94
Dry
1.19
Dry
OC 4
0.61
0.61
0.67
0.59
0.71
0.72
0.54
0.54
0.59
0.64
0.64
0.63
2
0.71
0.65
0.61
0.61
0.52
0.59
0.76
0.88
0.71
0.98
0.95
1.01
3
0.70
0.71
0.60
0.68
0.71
0.80
0.55
0.62
0.53
0.58
0.68
0.65
130
0.74
0.68
0.61
0.81
0.76
0.75
0.41
0.37
0.36
0.58
0.46
0.64
139
0.73
0.52
0.70
0.58
0.58
0.79
0.51
0.93
0.69
0.88
0.73
0.53
84
App
endi
x E
: 20
11 s
edim
ent
orga
nic
cont
ent
in 2
-4 c
m a
nd 4
-8 c
m s
ecti
ons
and
prop
orti
on o
f co
arse
(no
t pa
ssin
g th
roug
h 2
mm
sie
ve
mes
h), m
ediu
m (
did
not
pass
thr
ough
63
µm
sie
ve),
and
fin
e se
dim
ent
(pas
sed
thro
ugh
both
sie
ves)
, as
wel
l as
veg
etat
ion
that
co
llec
ted
on t
he 2
mm
sie
ve f
rom
the
0-2
cm
sed
imen
t se
ctio
n.
Pond
Organic
content 2-4
cm (%)
Organic
content 4-8
cm (%)
Coarse
fraction 0-2
cm (%)
Medium fraction
0-2 cm (%)
Fine fraction 0-2
cm (%)
Vegetation 0-2
cm (%)
July August
July August
July
August
July
August
July
August
July
August
100
8.0
8.1
10.5
7.5
0.2
1.4
58.1
43.4
41.7
55.2
0.9
0.6
103
11.0
11.1
9.7
11.6
0.0
0.0
34.1
51.5
65.9
48.5
0.5
3.0
110
7.0
6.3
6.5
5.4
0.3
0.1
42.2
49.7
57.5
50.3
0.7
0.5
113
7.8
14.0
6.5
10.6
0.1
0.0
38.1
34.1
61.8
65.9
0.7
1.2
118
8.1
10.0
8.1
8.6
0.0
0.0
38.7
49.5
61.3
50.5
1.1
3.1
OC 1
11.3
8.7
7.7
8.7
0.0
1.6
44.5
56.4
55.5
42.1
2.3
3.5
OC 2
8.8
Dry
9.4
Dry
0.4
Dry
28.7
Dry
70.9
Dry
0.6
Dry
OC 4
12.8
12.5
14.4
14.3
0.5
0.0
57.5
52.8
42.0
47.2
0.7
1.9
2
12.4
8.2
7.7
9.2
11.1
4.4
39.9
52.4
49.0
43.2
0.2
0.2
3
15.8
19.5
14.4
19.5
0.0
0.2
65.4
59.4
34.6
40.4
1.1
0.4
130
28.2
19.6
24.7
17.3
0.0
0.0
87.0
67.2
13.0
32.8
2.8
1.3
139
15.3
6.8
13.3
22.0
1.0
3.2
38.5
47.9
60.5
48.9
2.7
0.7
85
Appendix F: Concentration of excess (due to the addition of 201Hg) Me201Hg and 201Hg in the 0-2 cm section of incubated sediment cores from July and August 2011.
Pond
Excess Me
201Hg
(ng g-1)
Excess Me
201Hg
(ng g-1)
Excess 201Hg
(ng g-1)
Excess 201Hg
(ng g-1)
July August July August
103 0.056 0.054 5.897 4.143
113 0.060 0.093 2.130 3.257
118 0.064 0.031 3.976 2.979
OC 1 0.087 0.066 34.694 22.189
OC 2 0.068 Dry 9.977 Dry
OC 4 0.127 0.486 9.567 23.535
2 0.107 0.012 7.697 0.482
130 0.225 0.042 21.859 1.118
139 0.062 0.016 19.950 1.116
86
App
endi
x G
: M
ean
sedi
men
t w
ater
con
tent
, org
anic
con
tent
, TH
g, M
eHg,
and
num
ber
of y
ears
the
wet
land
was
sam
pled
in
Oct
ober
of
2008
, 201
0, a
nd 2
011.
Pond
Water
content
(%)
Organic
content
(%)
THg (ng g
-1)
MeHg (ng g
-1)
Wetlands
sampled
Wet
Dry
Wet Dry
Wet
Dry
Wet
Dry
Wet
Dry
100
82.8 56.9 51.5 28.1 86.87
55.00
0.58
0.50
1
2
103
76.0 74.9 28.8 42.6 53.15
66.33
0.36
0.32
2
2
110
78.7 67.4 41.7 44.8 52.85
66.45
0.41
0.86
1
2
113
79.6 81.3 39.4 44.4 52.11
61.86
0.68
0.58
2
2
118
85.7 74.2 57.5 34.4 83.44
54.43
0.50
0.47
2
2
OC 1
60.7 66.3 15.3 20.6 51.65
50.88
1.00
3.13
3
3
OC 2
48.4 40.1
7.9 10.9 26.53
29.18
1.05
0.95
1
2
OC 4
68.6 53.8 19.5 15.5 46.06
39.55
1.25
1.19
2
3
2
74.2 60.4 26.3 19.7 51.32
38.47
0.47
1.42
3
3
3
78.4 82.8 36.6 54.0 55.09
79.20
1.01
5.09
2
2
130
75.3 65.8 36.4 30.7 62.60
43.48
2.40
0.61
2
3
139
73.8 49.8 32.3 14.9 63.47
32.14
0.34
3.22
2
2
87
Appendix H: Mean sediment water content and organic content from triplicate 0-2 cm cores collected in May, June, July, and August 2012. Pond Water content (%) Organic content (%)
May June July August May June July August
100 40.9 41.5 36.5 37.2 8.5 9.3 8.2 7.6
103 47.2 53.2 40.2 43.3 11.2 15.4 9.1 10.3
110 55.0 44.8 40.5 41.4 14.8 9.4 8.2 8.3
113 54.7 44.9 44.4 46.7 16.1 9.8 9.8 11.9
118 53.1 50.6 43.9 45.0 16.2 13.1 10.6 10.6
OC 1 55.0 48.4 54.2 57.7 14.4 11.3 13.3 15.8
OC 2 43.8 37.6 44.0 50.0 10.8 11.0 12.1 12.0
OC 4 57.4 53.2 57.3 53.4 15.4 15.1 15.8 12.9
2 49.0 49.4 50.5 60.7 12.1 12.4 14.3 15.0
3 56.9 51.3 51.8 55.7 17.6 15.4 16.0 17.9
130 59.3 57.1 65.0 54.7 22.0 20.5 27.6 17.7
139 53.7 51.7 51.0 58.4 17.5 16.9 17.2 15.7