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Climate Change Impacts on Mercury Cycling in Peatlands
by
Kristine Marie Haynes
A thesis submitted in conformity with the requirements for the degree of Doctor of Philosophy
Department of Geography and Planning University of Toronto
© Copyright by Kristine Marie Haynes, 2017
ii
Climate Change Impacts on Mercury Cycling in Peatlands
Kristine Marie Haynes
Doctor of Philosophy
Department of Geography and Planning
University of Toronto
2017
Abstract
Climate change, through hydrological impacts and shifts in vascular plant communities, may
significantly affect the strength of peatlands as sinks of inorganic mercury (Hg) and downstream
sources of neurotoxic methylmercury (MeHg). Four complementary studies were conducted at
the PEATcosm (Peatland Experiment at the Houghton Mesocosm Facility) and SPRUCE
(Spruce and Peatland Responses Under Climatic and Environmental change) sites to investigate
the anticipated effects on Hg cycling. Lowered, more variable water tables and the removal of
Ericaceae shrubs in the PEATcosm mesocosms significantly enhance Hg and MeHg mobility in
peat pore waters as well as export from the peat in runoff during the snowmelt period. Mercury
is mobilized in association with dissolved organic carbon leached from the peat as a result of
increased aerobic decomposition. Enhanced Hg and MeHg solid phase peat concentrations are
observed within the zone of the lowered, fluctuating water tables due to translocation within the
peat profile. This shift in peat concentrations does not correspond to any significant differences
in Hg methylation or MeHg demethylation. The strongest trend of Hg deposition to peat occurs
with increased sedge cover, possibly as a result of coincidental shuttling of Hg to the peat by
aerenchymous tissues. Different plant functional groups alter the pathway of gaseous Hg
exchange, with greater Hg sorption to leaf surfaces and accumulation in leaf tissues by stomatal
uptake for the Ericaceae shrubs as compared to the sedges. Total gaseous Hg fluxes increase
iii
marginally with deep soil warming. Isotopic analyses demonstrate differences in Hg depositional
sources to the two peatland sites, likely due to differences in snow accumulation, and provide
insight into the mechanisms governing Hg cycling in peatlands. Collectively, these studies
demonstrate that anticipated climate change may significantly affect peatland Hg mobility and
cycling within and among the terrestrial, atmospheric and hydrological pools.
iv
Acknowledgments
I would like to express my gratitude to all those who gave of their valuable time on my journey
to the completion of my research.
Thank you to my supervisor, Dr. Carl Mitchell, for his comprehensive, capable guidance and
support. I am also grateful for the professional development opportunities provided to me by Dr.
Mitchell.
Thank you to my committee members, Dr. George Arhonditsis and Dr. Bridget Bergquist, for
their advice with regard to data analysis and interpretation, and their constructive suggestions
over the course of this research.
I appreciate the financial assistance provided through the Natural Sciences and Engineering
Research Council of Canada (NSERC) Alexander Graham Bell Canada Graduate Scholarship
(CGS-Doctoral), an Ontario Graduate Scholarship, and the University of Toronto awards
program.
I am grateful to Dr. Evan Kane (Michigan Technological University), Lynette Potvin, Dr. Erik
Lilleskov (USDA Forest Service Northern Research Station, Houghton, MI) and Dr. Randy
Kolka (USDA Forest Service Northern Research Station, Grand Rapids, MN) for facilitating
access and collaboration with the PEATcosm experiment.
Thank you to Dr. Paul Hanson (Oak Ridge National Laboratory) and Dr. Randy Kolka for
granting access to the SPRUCE site, and Robert Nettles for assistance in field sampling. Thank
you to Deacon Kyllander at the Marcell Experimental Forest Northern Research Station (USDA
Forest Service), Minnesota for logistical field assistance.
Within the Mitchell lab group, thank you to Kevin Ng for hours of extensive field and laboratory
assistance and Planck Huang for considerable laboratory analysis. I also appreciate the
assistance of Raymond Co, Brent Perron, and Ilana Tavshunsky in the lab.
Thank you to the Bergquist lab for the use of their mercury isotope analytical equipment under
the capable guidance of Dr. Wang Zheng and the assistance of Laura Zimmermann.
v
Thank you to my external appraiser Dr. Scott Brooks, Oak Ridge National Laboratory and
external examiners Dr. Sarah Finkelstein and Dr. Igor Lehnherr.
Thanks also to Jessica Finlayson, Geography Graduate Program Administrator for her assistance
dealing with academic matters.
To my family, my sincere gratitude for their ongoing and unwavering support through these
academic years to achieve the end towards my new beginning.
vi
Table of Contents
Acknowledgments.......................................................................................................................... iv
Table of Contents ........................................................................................................................... vi
List of Tables ................................................................................................................................. xi
List of Figures .............................................................................................................................. xiii
List of Appendices .........................................................................................................................xx
Chapter 1 Introduction .....................................................................................................................1
1.1 Mercury Cycling in the Environment ..................................................................................1
1.1.1 Overview of the Mercury Cycle ..............................................................................1
1.1.2 Mercury Methylation and MeHg Demethylation in Wetlands ................................3
1.1.3 Mercury - Organic Matter Relationships .................................................................6
1.1.4 Soil-Air Hg Exchange ..............................................................................................7
1.1.5 Mercury Stable Isotope Fractionation ......................................................................9
1.2 Peatland Hydrology and Hydrogeology.............................................................................10
1.3 Peatland Ecosystems and Global Climate Change ............................................................12
1.3.1 Global Climate Change and Peatland Hydrological Processes..............................12
1.3.2 Peatland Ecological Effects due to Climate Change ..............................................14
1.3.3 Potential Impacts on Mercury Cycling ..................................................................17
1.4 Objectives, Research Questions and Study Sites ...............................................................20
1.5 Thesis Structure and Publication Information ...................................................................25
1.5.1 Chapter 1 ................................................................................................................25
1.5.2 Chapter 2 ................................................................................................................25
1.5.3 Chapter 3 ................................................................................................................26
1.5.4 Chapter 4 ................................................................................................................27
1.5.5 Chapter 5 ................................................................................................................27
vii
1.5.6 Chapter 6 ................................................................................................................28
1.6 References ..........................................................................................................................28
Chapter 2 Mobility and Transport of Mercury and Methylmercury in Peat as a Function of
Changes in Water Table Regime and Plant Functional Groups ................................................45
2.1 Abstract ..............................................................................................................................45
2.2 Introduction ........................................................................................................................45
2.3 Materials and Methods .......................................................................................................48
2.3.1 Study Site and Experimental Design .....................................................................48
2.3.2 Water Sampling .....................................................................................................51
2.3.3 Pre-Treatment Peat.................................................................................................52
2.3.4 Peat Decomposition Assays ...................................................................................52
2.3.5 Analytical Methods ................................................................................................53
2.3.6 Statistical Analyses ................................................................................................53
2.4 Results and Discussion ......................................................................................................54
2.5 Conclusions ........................................................................................................................63
2.6 Acknowledgements ............................................................................................................64
2.7 References ..........................................................................................................................65
Chapter 3 Gaseous Mercury Fluxes in Peatlands and the Potential Influence of Climate
Change.......................................................................................................................................72
3.1 Abstract ..............................................................................................................................72
3.2 Introduction ........................................................................................................................73
3.3 Methods..............................................................................................................................76
3.3.1 PEATcosm Site Description ..................................................................................76
3.3.2 SPRUCE Site Description......................................................................................77
3.3.3 Experimental design and mercury flux measurements ..........................................78
3.3.4 PEATcosm vegetation and surface peat Hg ...........................................................81
3.3.5 Analytical Methods ................................................................................................82
viii
3.3.6 Statistical Analyses ................................................................................................82
3.4 Results ................................................................................................................................83
3.4.1 PEATcosm – water table and plant functional groups influence ...........................83
3.4.2 SPRUCE bog TGM fluxes .....................................................................................89
3.5 Discussion ..........................................................................................................................92
3.5.1 PEATcosm peat monoliths – influence of vascular plant community on TGM
fluxes ......................................................................................................................92
3.5.2 SPRUCE bog fluxes – deep soil warming influence on TGM fluxes....................96
3.5.3 Peatland TGM fluxes compared to other wetland environments ...........................97
3.6 Conclusions ........................................................................................................................98
3.7 Acknowledgements ............................................................................................................98
3.8 References ..........................................................................................................................99
Chapter 4 Impacts of Experimental Alteration of Water Table Regime and Vascular Plant
Community Composition on Solid Phase Peat Mercury Profiles and Methylmercury
Production ...............................................................................................................................109
4.1 Abstract ............................................................................................................................109
4.2 Introduction ......................................................................................................................110
4.3 Methods............................................................................................................................113
4.3.1 Study Site and Experimental Design ...................................................................113
4.3.2 Peat Sampling ......................................................................................................114
4.3.3 Pore Water Sampling ...........................................................................................116
4.3.4 Analytical Methods ..............................................................................................117
4.3.5 Statistical Analyses ..............................................................................................118
4.4 Results ..............................................................................................................................119
4.4.1 Hg Methylation and MeHg Demethylation .........................................................122
4.4.2 Hg(II) and MeHg Partitioning .............................................................................125
4.5 Discussion ........................................................................................................................126
ix
4.5.1 Conclusions and Implications ..............................................................................130
4.6 Acknowledgements ..........................................................................................................132
4.7 References ........................................................................................................................132
Chapter 5 Contrasting Mercury Isotopic Compositions of Two Sub-boreal Peatlands ...............141
5.1 Abstract ............................................................................................................................141
5.2 Introduction ......................................................................................................................142
5.3 Methods............................................................................................................................145
5.3.1 S1 Peatland Site Description ................................................................................145
5.3.2 PEATcosm Site Description ................................................................................145
5.3.3 Peat Sampling and THg Analysis ........................................................................147
5.3.4 Total Gaseous Mercury Flux Measurements .......................................................148
5.3.5 Sample Preparation and Mercury Isotope Analysis .............................................149
5.3.6 Data Analyses ......................................................................................................150
5.4 Results and Discussion ....................................................................................................151
5.4.1 S1 Peat Isotopic Composition ..............................................................................151
5.4.2 S1 Peat MIF .........................................................................................................156
5.4.3 PEATcosm Peat Isotopic Signature .....................................................................157
5.4.4 PEATcosm Treatment Effects – Water Table and Plant Functional Groups .......159
5.4.5 PEATcosm Peat MIF ...........................................................................................161
5.4.6 Conclusions and Implications ..............................................................................162
5.5 Acknowledgements ..........................................................................................................163
5.6 References ........................................................................................................................164
Chapter 6 Summary and Synthesis ..............................................................................................172
6.1 Summary ..........................................................................................................................172
6.1.1 Hg and MeHg Mobility in Peat Pore Water and Runoff .....................................173
6.1.2 Solid Phase Hg and MeHg – Translocation and Net MeHg Production..............174
x
6.1.3 Vascular Plant Communities – Interactions with Atmospheric Hg Deposition ..176
6.1.4 Spatial Variability in Hg Isotopic Signatures ......................................................177
6.1.5 Impacts of Climate Change on Peatland Hg Cycling ..........................................177
6.2 Limitations and Future Research Directions ....................................................................178
6.3 References ........................................................................................................................181
Appendix A. Supplementary Information for Chapter 2 .............................................................182
Appendix B. Supplementary Information for Chapter 3 .............................................................194
Appendix C. Supplementary Information for Chapter 4 .............................................................197
Appendix D. Supplementary Information for Chapter 5 .............................................................200
xi
List of Tables
Table 2-1 PEATcosm full-factorial experimental design. N=4 mesocosm bins per crossed
treatment. ...................................................................................................................................... 49
Table 2-2 Mean ± standard deviation THg and MeHg concentrations (in ng g-1
) in solid-phase
peat prior to experimental manipulations (n = 8 samples per treatment and depth increment). ... 56
Table 5-1 Contributions of Hg(II) (fHg(II)) and Hg(0) (fHg(0)) to peat isotopic signatures for the
PEATcosm 2014 treatment peat and S1 bog 2014 peat. Results are the mean of 1000 iterations
of the isotopic mixing model with Monte Carlo simulation based on the Δ201
Hg isotopic values.
..................................................................................................................................................... 152
Table A-2-1 Volume of water drained (mean ± standard deviation, in L) from the mesocosm
bins during spring snowmelt 2014 to achieve the water table treatment positions (n=4 per
treatment). ................................................................................................................................... 193
Table D-5-1 Mercury isotope signatures of reference materials and standards. ........................ 204
Table D-5-2 Mercury isotope signatures of all peat samples. Each sample was analyzed twice
for isotopes. 2σ is the higher of either 2 standard deviation (2SD) of the JTBaker standard or 2
standard error (2SE) of the two replicate results of each sample. .............................................. 205
Table D-5-3 Mean and standard deviation values of Δ199
Hg, Δ200
Hg and Δ201
Hg (all expressed in
‰) for Hg(II) and Hg(0) used in the binary mixing model with Monte Carlo simulation. ........ 206
Table D-5-4 Contributions of Hg(II) (fHg(II)) and Hg(0) (fHg(0)) to peat isotopic signatures for the
PEATcosm 2014 treatment peat and S1 2014 peat. Results are the mean of 1000 iterations of the
isotopic mixing model with Monte Carlo simulation based on the Δ199
Hg isotopic values. ...... 207
xii
Table D-5-5 Contributions of Hg(II) (fHg(II)) and Hg(0) (fHg(0)) to peat isotopic signatures for the
PEATcosm 2014 treatment peat and S1 2014 peat. Results are the mean of 1000 iterations of the
isotopic mixing model with Monte Carlo simulation based on the Δ200
Hg isotopic values. ...... 208
xiii
List of Figures
Figure 1-1 Dominant aspects of the mercury cycle in a peatland system (Source: C. Mitchell). .. 2
Figure 1-2 Potential synergistic and antagonistic implications of climate-induced peatland
hydrologic changes and hydraulic parameters (white boxes) and their potential impacts on Hg
cycling and transport (grey boxes). Solid arrows represent direct effects and dashed arrows
represent indirect or potential effects. (Mercury implications added to climate-induced peatland
hydrological changes that are modelled after Whittington and Price 2006). ................................ 18
Figure 1-3 PEATcosm Mesocosm Facility in Houghton, Michigan with 24 cubic metre peat
mesocosm bins. ............................................................................................................................. 22
Figure 1-4 Climate-controlled tunnel beneath Mesocosm Facility allowing access to the 24
mesocosm bins. Each bin was equipped with overflow tubing and collection bin to allow for
runoff sampling from each mesocosm, particularly during the snowmelt period. ....................... 23
Figure 1-5 Aerial view (left) of the SPRUCE plot footprints and boardwalks in the S1 bog,
Marcell Experimental Forest, Minnesota (September 2013 aerial photo from Paul Hanson, Oak
Ridge National Laboratory (ORNL)). One SPRUCE experimental plot (right). ......................... 25
Figure 2-1 Mean water table positions (in cm below the peat surface) in the low and high WT
treatment mesocosms and precipitation (in mm) over the course of this study from 2013 to 2015.
Water table manipulations were conducted in the summer months, while water table levels were
left to stabilize during the winter months. Dashed lines around the low and high WT treatment
mean water table positions represent the 95% confidence interval. Pore water and snowmelt
sampling events are denoted. ........................................................................................................ 50
Figure 2-2 Total Hg and MeHg concentrations and %MeHg in pore water and snowmelt runoff
in relation to water table and vascular plant functional group manipulations (treatment means of
all depths and sampling events). (a) pore water THg, (b) pore water MeHg, (c) pore water
%MeHg, (d) snowmelt runoff THg, (e) snowmelt runoff MeHg, and (f) %MeHg in snowmelt
runoff. Letters denote statistically similar groups based on transformed data. No significant
differences were observed in snowmelt runoff %MeHg across treatments. Significance of
xiv
treatment effects both individual (water table “WT” and plant functional group “Veg”) and
interactive (“WT*Veg”) for THg, MeHg, and %MeHg for both pore waters and snowmelt runoff
are noted. n = 318 pore water samples total (n = 50–56 per treatment). ...................................... 55
Figure 2-3 Relationships between shallow pore water (20 and 40 cm below the peat surface) Hg
and mean % mass loss of cellulose decomposition assays in the top 40 cm of the peat. (a) THg
concentrations, (b) MeHg concentrations, (c) %MeHg. Error bars represent standard deviation.
Decomposition assays were harvested in 2014 and represent potential peat decomposition among
the treatments. Pore water Hg data were averaged by treatment from all five sampling events. . 59
Figure 2-4 Hg-DOC relationships in 2014 and 2015 snowmelt runoff for all 24 mesocosms (n =
48 total). (a) THg in relation to DOC concentrations and (b) MeHg in relation to DOC
concentrations. Data are plotted on log-transformed axes. ........................................................... 63
Figure 3-1 PEATcosm daily TGM fluxes (in ng m-2
d-1
) across the four crossed water table
(WT) and plant functional group treatments (8 mesocosms in total monitored). * signifies that
sedge vegetation was present beneath the footprint of the DFC placed on the peat surface. ....... 84
Figure 3-2 Relationship between daily TGM fluxes (ng m-2
d-1
) and the number of sedge stems
present beneath the DFC at both the PEATcosm (closed red circles) and SPRUCE (open black
circles) sites. .................................................................................................................................. 85
Figure 3-3 PEATcosm 24-h TGM flux measurements (in ng m-2
h-1
) for the four crossed water
table and plant functional group treatments (two replicates each) in relation to surface soil
temperature (in °C) and solar radiation (kW m-2
). ........................................................................ 86
Figure 3-4 Relationship between PEATcosm hourly TGM fluxes (ng m-2
h-1
) and a) soil
temperature (in °C) and b) solar radiation (kW m-2
) among the four crossed water table and
vascular plant functional group treatments. .................................................................................. 87
Figure 3-5 Total Hg concentrations of PEATcosm vegetation a) leaf rinses expressed per unit
leaf tissue mass (ng g-1
dry wt.), b) leaf rinses per unit leaf area (ng m-2
) and c) leaf tissue mass
(ng g-1
dry wt.). Letters denote statistically similar values. No significant differences between
THg concentrations expressed on a per leaf surface area basis. ................................................... 88
xv
Figure 3-6 PEATcosm daily TGM fluxes (in ng m-2
d-1
) in relation to a) estimates of total leaf
area (in cm2) of all Ericaceae vegetation present beneath the DFC footprint, b) estimates of THg
sorbed to the surfaces of Ericaceae leaves under each DFC and c) surface (0–10 cm) peat THg
concentrations. .............................................................................................................................. 89
Figure 3-7 SPRUCE daily Hg fluxes (in ng m-2
d-1
) for both of the two replicate DFCs across the
range of deep peat heating temperature treatment plots (+0, +4.5 and +9 °C) in May 2014,
August 2014 and June 2015. Flux data for Plots 4 and 6 in June 2015 not available. ................. 90
Figure 3-8 Relationship between SPRUCE daily TGM fluxes and mean daily surface peat
temperatures for the May 2014 (pre-warming-solid circles), August 2014 (deep warming
achieved at depth-open circles) and June 2015 (prolonged deep warming-stars) measurement
periods. Deep warming target temperature differential treatments (+0, +4.5, +9 °C) are colour-
coded in blue, orange and red, respectively. ................................................................................. 92
Figure 4-1 Mean peat MeHg and THg concentration (ng g-1
) profiles for the six crossed water
table and vascular plant functional group treatments throughout the PEATcosm experiment from
2011 to 2014 and associated mean June to October water table positions for the High and Low
WT prescriptions. For clarity, error bars for each 10 cm depth increment have been omitted. 120
Figure 4-2 Relationship between mean treatment peat MeHg concentrations (ng g-1
) at 40 cm
below the peat surface (mean of the 30-40 and 40-50 cm depth increments) and pore water
acetate concentrations at 40 cm depth. ....................................................................................... 122
Figure 4-3 Mean kmeth (d-1
) in the upper 15 cm of the peat profile and 35-50 cm below the peat
surface for the a) High WT and b) Low WT treatments. Mean water table levels for the month
prior to sampling are denoted by the dashed lines. Letters denote statistically similar depths for
each of the High and Low WT data. ........................................................................................... 123
Figure 4-4 Relationship between the fraction of THg as MeHg (%MeHg) in the peat and a) kmeth
(% d-1
), b) kdemeth (% d-1
). ............................................................................................................ 123
Figure 4-5 Mean kmeth (d-1
) among the six experimental treatments in the upper 15 cm of the peat
profile (left column) and 35-50 cm below the surface (situated beneath the water table of the
Low WT treatments) (right column). Letters denote statistically similar groups among the six
xvi
treatments in the 10-15 cm and 35-40 cm depth increments. No significant differences among
treatments for any other peat depths. .......................................................................................... 124
Figure 4-6 Soil-water partition coefficients (logKD) for MeHg at 20 and 40 cm below the peat
surface among the six crossed water table and plant functional group treatments in 2013 and
2014 (n = 4 per treatment). Letters denote statistically similar groups. .................................... 126
Figure 5-1 Relationship between Δ200
Hg, Δ201
Hg and the surface peat THg concentrations (in ng
g-1
) for the S1 peatland and PEATcosm sites. Note different scale for Δ201
Hg for S1 bog and
PEATcosm data. ......................................................................................................................... 154
Figure 5-2 Relationship between δ202
Hg values (in ‰) and the mean daily TGM fluxes (in ng m-
2 d
-1) for the a) S1 and b) PEATcosm peatlands. Note different δ
202Hg and mean daily TGM flux
scales for both plots. ................................................................................................................... 155
Figure 5-3 Relationship between a) Δ199
Hg and Δ201
Hg and b) Δ199
Hg and δ202
Hg (all expressed
in ‰) for the surface 0-10 cm peat from both the S1 and PEATcosm peatland sites. ............... 157
Figure 5-4 Δ199
Hg and Δ201
Hg values (in ‰) among the six crossed water table and plant
functional group treatments of the PEATcosm peat monoliths. Analyses were on pooled samples
from replicate treatments, thus error bars represent analytical precision, not sample replication
variability. ................................................................................................................................... 160
Figure 5-5 Relationship between Δ200
Hg and Δ201
Hg (expressed in ‰) for the surface 0-10 cm
peat from both the S1 and PEATcosm peatland sites. ................................................................ 162
Figure 6-1 Influence of water table lowering on the partitioning of Hg and MeHg from the solid
phase into the aqueous pore water phase, translocation down the peat profile and re-sorption to
the solid phase within the zone of lowered water table fluctuation. ........................................... 175
Figure A-2-1 Photos of mesocosm bins. .................................................................................... 184
xvii
Figure A-2-2 Runoff total Hg and MeHg loads across the six crossed water table and vascular
plant functional group treatments during the 2014 spring snowmelt period (n=4 per treatment). a)
Total Hg and b) MeHg. Letters denote statistically similar groups based on transformed data. 185
Figure A-2-3 Mean pore water MeHg concentrations (on log scale) during the five sampling
events considered for this study. a) through e) denote specifically sampling periods. Each box
represents the mean pore water concentration of all three sampling depths across the four
replicate mesocosms (n=10 - 12). ............................................................................................... 186
Figure A-2-4 Mean pore water THg concentrations (on log scale) during the five sampling
events considered for this study. a) through e) denote specifically sampling periods. Each box
represents the mean pore water concentrations of all three sampling depths across the four
replicate mesocosms (n=10 - 12). ............................................................................................... 187
Figure A-2-5 Pore water Hg-DOC relationships. a) THg (log scale) in relation to DOC (square
root transformed) concentration, and b) MeHg (square root transformed) in relation to DOC
(square root transformed) concentration. The data includes all three pore water depths from each
of the four replicates per each of the six treatments. All five sampling events are included. ... 188
Figure A-2-6 Relationships between shallow pore water (20 and 40 cm below the peat surface)
Hg and % mass loss of cellulose decomposition assays in the top 40 cm of the peat. a) THg
concentrations, b) MeHg concentrations, c) %MeHg. Decomposition assays were harvested in
2014 and represent potential peat decomposition among the treatments. From all five sampling
events, pore water Hg data collected at 20 and 40 cm depths was averaged by mesocosm. n=18-
20 per treatment. ......................................................................................................................... 189
Figure A-2-7 Locally-weighted scatterplot smoothing (LOWESS) relationships between shallow
pore water (20 and 40 cm below the peat surface) Hg and % mass loss of cellulose
decomposition assays in the top 40 cm of the peat. a) THg concentrations, b) MeHg
concentrations, c) %MeHg. Decomposition assays were harvested in 2014 and represent
potential peat decomposition among the treatments. From all five sampling events, pore water
Hg data collected at 20 and 40 cm depths was averaged by mesocosm. n=18-20 per treatment.
..................................................................................................................................................... 190
xviii
Figure A-2-8 Pore water Hg-total phenolics relationships. a) THg (log transformed) in relation
to total phenolics (square root transformed) concentrations, and b) MeHg (square root
transformed) in relation to total phenolics (square root transformed) concentrations. The data
includes all three pore water sampling depths from each of the four replicates per each of the six
treatments. All five sampling events are included. .................................................................... 191
Figure A-2-9 Sulfate concentrations among the six crossed WT and plant functional group
treatments including all three sampling depths for the 2013 and 2014 sampling events. Letters
denote statistically similar groups. .............................................................................................. 192
Figure B-3-1 a) Relationship between mean inlet TGM concentrations (ng m-3
) and
corresponding 20-min TGM fluxes measured during daylight hours across the PEATcosm
experimental treatments. b) Boxplots of mean inlet Hg concentrations across the four PEATcosm
treatments. Letters denote statistically similar values. ............................................................... 194
Figure B-3-2 Piecewise regressions of PEATcosm hourly TGM fluxes (in ng m-2
h-1
) and
surface soil temperatures (in °C) for the a) High WT Control, b) Low WT Control, c) Low WT
Sedge, and d) Low WT Ericaceae treatments. Temperature thresholds are noted for each
treatment by the vertical dashed lines. The equations of the relationships before and after each
threshold temperature are specified. ........................................................................................... 195
Figure B-3-3 a), c) and e) Boxplots of mean inlet TGM concentrations measured during daylight
hours across the SPRUCE plots for each of the three sampling events. Letters denote statistically
similar values. b), d) and f) Relationship between mean inlet TGM concentrations (ng m-3
) and
corresponding 20-min TGM fluxes (daylight only) across the SPRUCE plots during each of the
three samplings. .......................................................................................................................... 196
Figure C-4-1 Peat MeHg and THg stock (in ng cm-3
) profiles for the six combined water table
and vascular plant functional group treatments throughout the PEATcosm experiment from 2011
to 2014 and associated mean June to October water table positions for the High and Low WT
prescriptions. For clarity, error bars for each 10 cm depth increment have been omitted. ........ 197
xix
Figure C-4-2 Mean kdemeth (d-1
) in the upper 15 cm of the peat profile and 35-50 cm below the
peat surface for the a) High WT and b) Low WT treatments. Mean water table levels for the
month prior to sampling are denoted by the dashed lines. No statistically significant differences
among the depths for each of the High and Low WT data. ........................................................ 198
Figure C-4-3 Soil-water partition coefficients (logKD) for Hg(II) at 20 and 40 cm below the peat
surface among the six crossed water table and plant functional group treatments in 2013 and
2014 (n = 4 per treatment). Letters denote statistically similar groups. .................................... 199
Figure D-5-1a S1 peatland of the Marcell Experimental Forest in north-central Minnesota
showing one SPRUCE plot footprint. ......................................................................................... 200
Figure D-5-1b PEATcosm Mesocosm Facility, Houghton, Michigan, with the 24 peat monoliths
(top) and underlain by a climate-controlled tunnel (bottom)…………………………………...200
Figure D-5-2 Relationship between Δ200
Hg and Δ204
Hg (all expressed in ‰) for the surface 0-10
cm peat from both the S1 and PEATcosm peatland sites. .......................................................... 202
Figure D-5-3 Relationship between Δ200
Hg and Δ201
Hg (expressed in ‰) for the surface 0-10 cm
peat from both the S1 and PEATcosm peatland sites as compared to reported values of Hg(II)
deposition in the literature (Gratz et al. 2010; Sherman et al. 2012; Demers et al. 2013). ......... 203
xx
List of Appendices
Appendix A. Supplementary Information for Chapter 2.............................................................182
Appendix B. Supplementary Information for Chapter 3.............................................................194
Appendix C. Supplementary Information for Chapter 4.............................................................197
Appendix D. Supplementary Information for Chapter 5.............................................................200
1
Chapter 1 Introduction
1
1.1 Mercury Cycling in the Environment
1.1.1 Overview of the Mercury Cycle
Mercury (Hg) is a toxic global pollutant that poses a severe toxicological and teratological threat
to both wildlife and humans (Morel et al. 1998; Mergler et al. 2007; Scheuhammer et al. 2007).
This threat is evidenced by the extent of fish consumption advisories due to high Hg levels
across both Canada and the United States (Driscoll et al. 2007). In addition to natural sources
such as volcanic emissions, anthropogenic activities such as coal burning, artisanal gold mining
and waste incineration have contributed to global mercury pools in the atmospheric, terrestrial
and aquatic environments (Selin 2009). Mercury exists predominantly in three forms in the
environment: elemental mercury (Hg(0)), inorganic divalent mercury (Hg(II)) and organic
methylmercury (CH3Hg+; MeHg) (Zillioux et al. 1993). A potent neurotoxin, MeHg is known to
bioaccumulate in aquatic organisms and biomagnify in higher trophic levels of the food web.
Sinks, sources and fluxes involved in Hg biogeochemical cycling have been generally
established on a global scale (Mason et al. 1994). Novel techniques are emerging that are greatly
advancing the level of understanding of Hg cycling processes across a range of scales. Some of
these recent advancements include the determination of genes hgcAB involved in Hg
methylation (Parks et al. 2013; Gilmour et al. 2013), an important step in the direction of
developing a Hg gene probe, the role of nanoparticles in controlling the association of Hg with
dissolved organic matter (e.g. Graham et al. 2012) and the application of high-resolution mass
spectroscopy techniques to track natural stable isotope fractionation (e.g. Demers et al. 2013;
Kritee et al. 2013; Zheng et al. 2016) allowing further insight into the sources and processes
governing Hg cycling.
Following emission from either natural or anthropogenic sources, elemental Hg in the
atmosphere has an average residence time of 0.5 to 2 years; sufficient time for global-scale
transport from point sources to remote landscapes (Mason et al. 1994; Schroeder and Munthe
2
1998). This elemental Hg in the atmosphere may be oxidized by strong oxidants such as
halogens resulting in the production of reactive gaseous Hg(II) as well as particulate Hg(II); both
with shorter atmospheric residence times than that of elemental Hg(0) prior to deposition
(Lehnherr 2014). Atmospheric deposition, via either wet or dry deposition, represents the
dominant input of Hg pollution to watersheds (Fitzgerald et al. 1998; Driscoll et al. 2007), with a
large portion of this deposited Hg being incorporated into the soil and vegetation pools
(Hintelmann et al. 2002). However, while acting as sinks of Hg, terrestrial watersheds can also
be significant sources of MeHg to downstream aquatic ecosystems via runoff (Munthe et al.
2007; Harris et al. 2007). Wetlands as well as other aquatic environments such as lake-bottom
sediments are known to be sites of Hg methylation due to favourable redox and hydrological
conditions (Branfireun and Roulet 2002; Tjerngren et al. 2012; Figure 1-1). Deposited Hg is
subjected to numerous influences in soils following deposition which may incorporate it into the
soil and vegetation pools, including accumulation in organic peat soils (Figure 1-1) or this
deposited Hg may be subsequently re-emitted from soils to the atmosphere through reduction to
Hg(0). Inorganic Hg and MeHg may also be mobilized in pore waters, exported from peatlands
in runoff (Figure 1-1) and transported to downstream aquatic environments. In such
environments, Hg and MeHg may enter the aquatic food chain.
Figure 1-1 Dominant aspects of the mercury cycle in a peatland system (Source: C. Mitchell).
3
1.1.2 Mercury Methylation and MeHg Demethylation in Wetlands
The process of Hg methylation which transforms inorganic Hg into the organic form MeHg
occurs predominately via biological mediation. Chemical abiotic methylation of Hg is also
known to occur in sediments via biochemical pathways not involving microbial communities.
However, abiotic Hg methylation is relatively minor in terms of environmental significance as
compared to that of biological methylation (Berman and Bartha 1986). Abiotic methylation may
play an increasingly important role in certain environments such as wetlands due to the potential
role of dissolved organic matter, specifically recalcitrant humic substances, as has been
suggested by Ravichandran (2004). An important benchmark study by Compeau and Bartha
(1985) was first to identify sulfate-reducing bacteria (SRB) as the principal methylators of Hg(II)
in anoxic sediments. This critical finding was further supported by research conducted by
Gilmour et al. (1992) in which sulfate additions stimulated MeHg production while specific
inhibition of SRB using molybdate prevented methylation. Sulfate additions have also been
observed to enhance MeHg production at the wetland scale (Jeremiason et al. 2006). However,
the accumulation of sulfides may act to inhibit MeHg production by reducing the bioavailability
of Hg for methylating bacterial communities (Benoit et al. 1999). As well, research has found
that iron-reducing bacteria may play a role in Hg methylation (Warner et al. 2003; Mitchell and
Gilmour 2008). Early studies investigating the effect of sulfate amendments on the methylation
potential of SRB employed only Desulfovibrio desulfuricans in culture experiments as being
representative of all SRB species (Compeau and Bartha 1985). However, more recent studies
have revealed 19 genera of SRB, each of which have the potential to methylate Hg to varying
extents and rates (King et al. 2002), while some SRB do not methylate Hg (Benoit et al. 2003;
Gilmour et al. 2013). Despite the observation of methylation in artificial cultures, some bacteria
such as Desulfovibrio desulfuricans and Desulfobulbus propionicus are not prominent
methylators of Hg within the natural environment (Benoit et al. 2003). Similarly,
Desulfobacterium studied by King et al. (2000) required sulfate reduction to be occurring in
order to methylate Hg. In contrast, Desulfobulbus propionicus methylates Hg while growing
fermentatively (Benoit et al. 2003).
Not all sulfate- and iron-reducing bacteria have the ability to methylate Hg (Gilmour et al. 2011).
Recent novel research has determined the gene pair (hgcAB) necessary for MeHg production
(Parks et al. 2013). Further testing has confirmed that the presence of this gene pair predicts an
4
organism’s ability to methylate Hg (Gilmour et al. 2013). The study by Gilmour et al. (2013)
extends the knowledge of Hg methylators beyond the Deltaproteobacteria. In addition to sulfate-
and iron-reducers, Gilmour et al. (2013) determined that some methanogens and syntrophic,
acetogenic and fermentative Firmicutes that possess the hgcAB gene pair are capable of Hg
methylation. The experimental confirmation that other microbial communities than sulfate- and
iron-reducers can methylate Hg may have important implications for different habitats from
which the Hg-methylating strains were isolated; including northern peatlands (Gilmour et al.
2013). The linkage between Hg methylation and discrete gene markers in microbial communities
is an important step in order to develop gene probes for Hg methylation (Gilmour et al. 2013;
Parks et al. 2013).
Given the nature of Hg methylation as a reaction facilitated by anaerobic SRB and other
heterotrophic, redox-sensitive organisms, highly specific conditions are required in order for the
process to occur. The synthesis and stability of MeHg is favoured by negative redox potentials
(Compeau and Bartha 1984) and as such MeHg production is known to occur in anaerobic
environments such as lake-bottom sediments (Winfrey and Rudd 1990; Ramlal et al. 1993;
Miskimmin et al. 1992) and wetlands including peatlands (St. Louis et al. 1994; Tjerngren et al.
2012). Temperature is also a major contributing factor controlling Hg methylating activity.
Mercury methylation is limited by low temperatures; occurring optimally at a temperature of
35ºC (Winfrey and Rudd 1990). This influence of temperature on MeHg production suggests
seasonality in which methylation is highest in late summer (Winfrey and Rudd 1990), as well as
the potential for climate change-induced alteration of this transformation with increasing soil and
atmospheric temperatures. However, the influence of climate change on Hg methylation has not
been explored in the literature. The role of pH in Hg methylation is dependent upon the
conditions of the system under study. Some research reports increased methylation with
decreasing pH; while others conclude the opposite (i.e. Miskimmin et al. 1992; Ramlal et al.
1985). The pH level may play a role in the association of Hg with solid phases rendering it
unavailable for methylation; with greater partitioning to the liquid, mobile phase in acidic
conditions (Benoit et al. 2003; Haitzer et al. 2003).
Microbial communities also require an available labile carbon source in order to carry out the
methylation process (Mitchell et al. 2008). The source of available carbon could originate in
upland systems and be delivered to wetlands via runoff in natural environments (Mitchell et al.
5
2008), while enhanced organic matter decomposition in newly-flooded reservoirs may provide
this source of labile carbon (Winfrey and Rudd 1990; Heyes et al. 2000). As part of the
Experimental Lakes Area Reservoir Project (ELARP), Heyes et al. (2000) reported an increase in
net MeHg production as a result of impoundment of a boreal peatland. The observed spike in
MeHg production was attributed to the increasing areal extent of anoxia due to flooding as well
as the release of carbon and other nutrients due to increased decomposition of newly-inundated
vegetation. A field study in which MeHg production in a wetland decreased as a result of
vegetation removal further highlights the need for an organic carbon source in order for
methylation to occur (Windham-Myers et al. 2009). This study suggested that the exudates from
roots including acetate provide labile carbon for the SRB thereby stimulating Hg methylation in
the rhizosphere. Mitchell et al. (2008) amended peatland mesocosms with sulfate and an array of
carbon sources with the goal of simulating nutrient fluxes from the upland to the peatland that
would occur naturally and then assessed the resultant MeHg production. This mesocosm study
demonstrated that the addition of both a labile carbon source and sulfate in combination
stimulates the greatest net MeHg production. Ravichandran (2004) suggested that predominantly
labile and easily decomposed organic matter may stimulate microbial growth and therefore
promote Hg methylation. However, recalcitrant, refractory portions of dissolved organic matter
(including humic and fulvic acids) may contribute to abiotic Hg methylation (Ravichandran
2004).
In natural systems the concentration of MeHg present depends upon the balance between the
competing processes of methylation and demethylation. Methylmercury degradation may occur
via a number of both abiotic and biotic pathways (Marvin-Dipasquale et al. 2000). Abiotic
pathways include photodegradation of MeHg (Sellers et al. 1996) as well as the production of
dimethylmercury through a reaction with sulfides, particularly in marine waters (Jensen and
Jernelöv 1969; Mason et al. 1995; Lehnherr 2014; Jonsson et al. 2016). The most widely studied
biotic degradation pathways involve the investigation of microbes present in mercury-
contaminated systems that have developed a resistance to mercury with the evolution of genes
encoded by the mer operon (Marvin-Dipasquale et al. 2000). The mer-A gene results in the
production of the mercuric reductase enzyme which reduces inorganic Hg2+
to volatile elemental
Hg. This occurs following the cleavage of MeHg into CH4 and Hg2+
by the mer-B gene which
encodes for the organomercurial-lyase enzyme (Barkay et al. 1991; Marvin-Dipasquale et al.
6
2000; Parks et al. 2009). The relative contributions of oxidative degradation (OD) and reductive
demethylation pathways including mer-mediated and non-mer-mediated methylmercury
degradation were suggested by Marvin-Dipasquale et al. (2000) to be a function of the level of
site Hg contamination; with OD dominating in less-contaminated sites. Further research is
required to decipher the controls on demethylation and how this process influences net MeHg
production. This may be particularly important for wetland environments, as this antagonistic
process influences the amount of MeHg available for export.
1.1.3 Mercury - Organic Matter Relationships
Organic matter is important in both the transport of Hg among different watershed
compartments such as from upland soils to wetlands and water bodies (i.e. Mierle and Ingram
1991), as well as in the retention of Hg in soils (Skyllberg et al. 2000). Several studies have
found mercury fluxes during high-flow events exhibit strong correlations to dissolved organic
carbon (DOC) fluxes (i.e. Bishop et al. 1995; Dittman et al. 2010; Schuster et al. 2008). Mercury
has a tendency to complex with dissolved organic matter (DOM) (e.g. Driscoll et al. 1995;
Schuster 1991) and as a result Hg mobilization may be enhanced during high-flow events due to
this complexation.
The extent of Hg binding to organic matter depends upon the available functional groups. Thiol
and other reduced sulfur groups (including those with either one oxygen or nitrogen atom) in
humic and fulvic acids in both soil and fresh waters have been observed to exhibit high affinity
for inorganic mercury (Hg2+
) as well as MeHg (Hintelmann et al. 1997; Karlsson and Skyllberg
2003; Skyllberg et al. 2000). It has been suggested that the affinity of Hg to complex with DOC
may act to limit its bioavailability for transformation processes such as Hg methylation
facilitated by microbial communities capable of performing this reaction (Graham et al. 2012;
2013). Graham et al. (2012) highlights the potential role of metal nanoparticles, such as HgS,
from preventing the binding of Hg with organic matter in sulfidic environments. By inhibiting
binding with organic matter these small HgS particles remain bioavailable to Hg methylating
bacteria (Graham et al. 2012). The presence of such particles may account for the propensity of
organic-rich environments such as wetlands and lake sediments to facilitate the production of
MeHg.
7
1.1.4 Soil-Air Hg Exchange
Depending upon the speciation of atmospheric Hg (elemental Hg(0), reactive gaseous Hg and
particulate-bound ionic Hg(II)), residence times of Hg in the atmosphere may range on the order
of hours to days for particulate-bound and reactive gaseous Hg to between six months to two
years for Hg(0) (Schroeder and Munthe 1998; Driscoll et al. 2013). These variable atmospheric
residence times account for both observed hot spots of Hg contamination near emission sources
as well as the propensity of Hg to be a global pollutant (Lindberg et al. 2007). Mercury can be
deposited from the atmosphere by both wet and dry processes. Local or regional deposition via
both wet and dry depositional processes is most common with both particulate and reactive Hg
forms whereas elemental Hg may be transported long distances prior to being either deposited
directly via dry deposition or oxidized and deposited through either wet or dry mechanisms to
the soil or water surface. Wet deposition involves scavenging of gas-phase and aerosol-phase Hg
by precipitation and is delivered to the soil surface (Zhang et al. 2009). In addition to the direct
net flux of gaseous Hg to the soil surface, dry deposition may involve plant uptake of
atmospheric Hg followed by deposition to soil via litterfall or deposition of Hg to vegetation
surfaces and delivery to soil via washing of foliage during precipitation events (Rea et al. 2000;
Laacouri et al. 2013; Driscoll et al. 2013).
Soils may act both as a sink or source of Hg to the atmosphere. Once deposited, Hg enters
various soil pools and undergoes numerous reactions which affect the speciation and therefore
determine the fate of deposited Hg. Following deposition, Hg particularly in the divalent form,
has the potential to be retained in the soil due to its ability to bind to organic and mineral particle
surfaces (Gabriel and Williamson 2004). However, both biotic and abiotic processes can act to
reduce this soil-bound Hg for re-emission to the atmosphere (e.g. Xin et al. 2007; Lin et al.
2010). Adsorption to organic, particularly humic, material and to a lesser extent inorganic
material may be important in indirectly controlling the potential reduction and re-emission of
deposited Hg. The strong tendency of Hg to bind to reduced sulfur functional groups (e.g. thiol
groups) in organic matter (Graham et al. 2012; Khwaja et al. 2006; Ravichandran 2004;
Skyllberg et al. 2006) may act to reduce the pool of Hg in soil available for re-emission.
However, processes which enhance desorption may counteract such adsorption and shift the
balance of whether a soil acts as a source or sink of Hg.
8
Both biotic and abiotic processes have been observed to influence desorption and subsequent
emission from a variety of soils (Zhang and Lindberg 1999). Biologically-mediated reduction of
Hg(II) may act to increase Hg(0) evasion from soils. Soil temperature has been observed to be a
significant control on Hg fluxes to the atmosphere with enhanced fluxes from warmer soils
(Edwards et al. 2001), as increased temperatures effectively lower the activation energy required
for Hg reduction (Carpi and Lindberg 1998). As well, ultraviolet radiation is a strong influence
on the enhancement of evasion from soils (Carpi and Lindberg 1998; Xin et al. 2007). Strong
diel variations in Hg fluxes are often observed with enhanced evasion during midday and
diminished evasion or net uptake reported during the overnight hours. Soil moisture may also
play an important role in influencing the adsorption and desorption from soil particle surfaces
(Gustin and Stamenkovic 2005; Song and Van Heyst 2005). Soils, particularly mineral soils,
often show a greater affinity for water molecules than Hg. As such, adsorbed Hg may be
dislodged from soil surfaces and emitted to the atmosphere. This process has been observed with
enhanced Hg fluxes following precipitation events (Song and Van Heyst 2005), which
subsequently diminish as Hg diffusion is suppressed in saturated soils (Bahlmann et al. 2004;
Gustin and Stamenkovic 2005; Song and Van Heyst 2005). This concept requires further
investigation as it has been suggested that antecedent soil moisture conditions may be a
contributing influence (Gustin and Stamenkovic 2005). Other factors which may influence the
reduction of divalent Hg in terrestrial environments and therefore play a role in controlling
emissions include soil organic matter content and soil Hg concentrations (Lin et al. 2010). The
individual controls of certain processes influencing Hg evasion from soils warrant further
investigation particularly with controlled experimental treatments or at the laboratory scale as
possible synergisms may exist between the factors governing soil-air Hg fluxes.
Recent research has examined atmospheric Hg fluxes from forest and wetland ecosystems
including northern boreal peatlands using both chamber methods (Kyllönen et al. 2012; Fritsche
et al. 2014) and relaxed eddy accumulation techniques (Osterwalder et al. 2016). In order to
effectively model the potential global Hg sources to the atmosphere and how the evasion
processes may be impacted by climatic and other environmental changes, further monitoring of
soil-air Hg fluxes is warranted across a range of environments. Mercury fluxes and the processes
that govern Hg evasion from climate change-sensitive northern peatlands are understudied at the
9
present time despite the large stocks of Hg stored in these systems that represent a potentially
significant source of Hg to the atmosphere (Kolka et al. 2001; Grigal 2003).
1.1.5 Mercury Stable Isotope Fractionation
The fractionation of stable isotopes of light elements such as carbon, oxygen and nitrogen has
been applied as a means of tracing the transformations involved in the biogeochemical cycles of
these elements. With the advent of multiple collector inductively coupled plasma mass
spectrometry (MC-ICP-MS), the natural fractionation of stable isotopes of heavier elements can
be utilized to elucidate the processes controlling the cycling of Hg through the environment
(Bergquist and Blum 2007; Yin et al. 2010). Mercury has seven stable isotopes: 196
Hg, 198
Hg,
199Hg,
200Hg,
201Hg,
202Hg and
204Hg with a mass difference of 4% (Bergquist and Blum 2009).
Mercury also has an active redox chemistry, a volatile form (Hg(0)), and a tendency to form
covalent bonds thereby facilitating ample opportunities for isotopic fractionation to occur
(Bergquist and Blum 2007). Recent and on-going research is investigating the natural
fractionation of Hg stable isotopes in order to trace Hg sources (due to anthropogenic pollution,
for example) and pathways involved in Hg biogeochemical cycling (i.e. Hg methylation, Hg2+
reduction to elemental Hg) (Yin et al. 2010; Donovan et al. 2014). Atmospheric Hg reaction
mechanisms are an area of study to which Hg isotopic analyses, investigating the atmospheric
processes that fractionate Hg, are making significant strides (e.g. Gratz et al. 2010; Chen et al.
2012; Sun et al. 2016).
Mercury isotopes are useful in tracking and characterizing Hg transformations due to the
significant variations observed in the isotope ratios in the natural environment. In studying these
isotopic variations of Hg in nature it is difficult to discern the signature of the source of Hg from
that resulting from an array of processes which transform Hg from one species to another
(Bergquist and Blum 2009). Several processes, both biotic and abiotic, have been found to cause
mass-dependent fractionation (MDF), while mass-independent fractionation (MIF) has been
determined for only a limited number of processes such as photochemical reduction (Bergquist
and Blum 2007; Bergquist and Blum 2009; Zheng et al. 2016). It is important to understand the
mechanisms that govern the natural fractionation of Hg isotopes in order to accurately trace Hg
fluxes through the environment.
10
Recent studies have begun to investigate Hg isotopic signatures in peat soils (Jiskra et al. 2015;
Enrico et al. 2016). These studies have taken an important first step in modelling relative Hg(0)
and Hg(II) contributions to peatlands as well as discerning the mechanisms controlling peat Hg
cycling including non-photochemical abiotic reduction of Hg facilitated by natural organic
matter reduction as well as photochemical reduction (Jiskra et al. 2015). Further research
applying Hg isotopic analyses in peatland ecosystems is required to better understand the sources
of Hg deposition to these systems. Isotopic analyses may also provide insight into how Hg
cycling in peatlands may be affected in a changing climate.
1.2 Peatland Hydrology and Hydrogeology
Peatland ecosystems are predominantly located in boreal and subarctic regions globally and
cover approximately 12% of the Canadian landscape (Tarnocai 2009). Through their unique
biogeochemical and hydrological characteristics these ecosystems play an important role in
global climate; acting as seemingly perpetual sinks of atmospheric carbon dioxide (CO2)
throughout the Holocene epoch via accumulation of immense quantities of organic matter
(Gorham 1991). Accumulation of organic matter in thick peat soils occurs in peatlands due to a
net imbalance between plant net primary production and microbial decomposition. Hydrological
processes are a significant control on maintaining peatland water tables and therefore the delicate
balance between carbon accumulation through the maintenance of anoxia and microbial
mineralization (Whittington and Price 2006). The slow rates of organic matter decomposition are
primarily attributed to waterlogged, anoxic soils, cold temperatures and chemically-recalcitrant
organic matter substrate for microbial decomposers (Moore and Basiliko 2006). Despite their
small global land area, covering only approximately 3%, these ecosystems store between 10%
(Gorham 1991) to 30% (Turunen et al. 2002) of the global soil carbon pool. Northern peatlands
generally comprise a spectrum from nutrient-poor, ombrotrophic bogs to rich, minerotrophic fens
(Glaser et al. 1997). Ombrotrophic bogs are generally isolated from groundwater inputs as well
as from significant surface water inputs. As a result, the hydrology of these systems is
predominantly controlled by the balance between precipitation and evapotranspiration. In
contrast, minerotrophic fens are connected to groundwater inputs and consequently receive
mineral-rich groundwater resulting in higher pH values than bog systems (Glaser et al. 1997). As
11
a result of the disparity between bogs and fens in terms of hydrological inputs and nutrient status,
the vegetation communities present in these ecosystems differ with fens dominated by vascular
plant communities (such as sedges) while acidophilic Sphagnum mosses are predominant in bogs
(Glaser et al. 1997).
The structure of accumulated peat in these systems plays an important role in controlling
flowpaths through peatlands. The upper peat layers, known as the acrotelm, consist of fresh,
minimally-decomposed organic matter and have higher hydraulic conductivities than that of the
deeper peat, known as the catotelm. The catotelm consists of well-humified organic material
with low porosity and hydraulic conductivity (Clymo 1984). Due to the highly decomposed and
compressed peat in the deeper catotelm layer, the majority of subsurface flow is considered to
occur through the acrotelm. However, research by Chason and Siegel (1986) suggests that the
connection between surface water and groundwater may be significant. Field and laboratory
measurements of vertical and horizontal conductivity demonstrated the ability of peat at various
stages of humification to transmit groundwater quickly (Chason and Siegel 1986). Such a result
calls into question the idea that well-humified peat does not effectively transmit groundwater due
to low hydraulic conductivity. In large patterned peatland systems such as the Glacial Lake
Agassiz Peatlands in Minnesota, groundwater contributions have been observed to play an
important role in the water balance of the ecosystem as well as the pore water chemistry and
consequently the dominant vegetation communities (i.e. Siegel and Glaser 1987).
Although peatlands are important zones of groundwater recharge in the landscape they are also
considerable discharge zones (i.e. Siegel and Glaser 1987; Siegel et al. 1995). By monitoring
groundwater observation wells during the winter in a number of peatlands, Nichols and Verry
(2001) calculated the rate of water table decline without any additional recharge due to
precipitation. Deviations from these background winter water table recession rates facilitate the
determination of groundwater recharge throughout the rest of the year. Nichols and Verry (2001)
concluded that approximately 40% of total water yield in these watersheds was the result of
groundwater recharge and that this recharge varied linearly with precipitation. Nichols and Verry
(2001) suggested that the water table depths in the S1 peatland did not exhibit a predictable
decline during the winter season such as those recorded for nearby watersheds in the Marcell
Experimental Forest because S1 is an area of groundwater discharge. In some cases groundwater
flow reversals have been observed to occur as a result of drought conditions; resulting in
12
groundwater upwelling due to a change in hydraulic gradient (i.e. Romanowicz et al. 1993;
Siegel et al. 1995).
1.3 Peatland Ecosystems and Global Climate Change
Global climate change has the potential to significantly impact the hydrology of mid- to high-
latitude northern peatland ecosystems due to increasing temperatures and changes in
precipitation patterns (Bridgham et al. 2008). Within Canada, peatlands in the Boreal and
Subarctic regions are forecasted to exhibit severe to extremely severe sensitivity to climatic
changes (Tarnocai 2006). The susceptibility of peatlands to climatic changes may depend upon
the relative contributions of groundwater inputs versus those of precipitation; with precipitation-
dominated wetlands exhibiting particular vulnerability to climate change (Winter 2000).
Climate-induced alterations in hydrology may have subsequent impacts on vegetation
communities present across peatland types as a result of fluctuating water tables (Weltzin et al.
2000; Strack et al. 2006). Soil microbial communities present in these peatland ecosystems may
in turn be affected as a result of the hydrological and redox conditions as well as potential
changes in available organic matter substrate for decomposition (Peltoniemi et al. 2009). Due to
the coupled relationship between the cycling of Hg and DOC (Driscoll et al. 1995; Skyllberg et
al. 2000), the cycling of Hg and transport in peatlands will likely be impacted by alterations in
carbon cycling and redox conditions resulting from climate-induced changes in hydrological
processes.
1.3.1 Global Climate Change and Peatland Hydrological Processes
At mid- to high-latitudes in Canada, mean atmospheric temperatures are forecasted to increase 3-
4°C by 2020 and 5-10°C by 2050; relative to the 1960-1990 period (Tarnocai 2009). These
predicted temperature increases in conjunction with altered precipitation patterns with potentially
less frequent, more intense events (Groisman et al. 2012; Janssen et al. 2014; Yu et al. 2016)
may affect peatland water tables. Evapotranspiration may significantly increase as a result of
enhanced temperatures. Consequently, peatland water tables are predicted to decline due to
enhanced water losses via evapotranspiration exceeding precipitation inputs; particularly in
13
ombrotrophic bog systems due to the dominance of precipitation in maintaining water levels
(Winter 2000). Depending upon the severity and duration of water table drawdown, the carbon
storage function of peatlands may be lost as a result of feedback mechanisms facilitating the
release of carbon to the atmosphere as CO2 due to enhanced peat aeration (Ise et al. 2008),
becoming weaker carbon sinks under drier conditions (Chimner et al. 2016). This decline in
peatland water tables due to enhanced evapotranspiration may be further exacerbated by the
predicted shift in vegetation communities present in peatlands from Sphagnum- to vascular-
dominated as a result of prolonged climatic changes. Modelling of evapotranspiration by
Admiral and Lafleur (2007) suggested that vascular plants contribute the majority of the
observed total latent heat flux (~60-80%) when compared to both hummock and hollow
Sphagnum mosses. However, Sphagnum moss present in peatlands may be able to control energy
and water fluxes resulting in a potential cooling effect (Blok et al. 2011), through their ability to
blanch during periods of drought which in turn increases their albedo. The amount of energy
available for evapotranspiration by vascular plants may be affected and therefore Sphagnum
moss may provide a self-regulating feedback mechanism in the maintenance of higher water
table levels. However, this resiliency mechanism may be lost if the changes in peatland
hydrology result in the successful recruitment of vascular plants at the expense of moss species
(Waddington et al. 2015).
Climate change may also dramatically impact both surface and groundwater flowpaths and rates
in peatland ecosystems as a result of alterations in important hydraulic parameters which govern
the nature of hydrological processes and subsequent feedback mechanisms. With deeper water
tables as a result of climate change-induced hydrological variations, the character of the peat
may change due to effects on the bulk density, hydraulic conductivity and specific yield
(Waddington et al. 2015; Whittington and Price 2006). Due to its compressibility and large
potential capacity for water storage, long-term or seasonal changes in peatland water table
position may affect the storage capacity of peatlands. As the water pressure in the surface peat
soils decreases due to water table recession, subsidence occurs as a result of enhanced
compression and decomposition due to increased oxidation (Whittington and Price 2006);
potentially providing a shallower relative water table. Whittington and Price (2006) observed a
significant increase in peat bulk density as a result of drawdown-induced peat compression. This
result was observed with a concurrent decrease in hydraulic conductivity across the studied
14
peatland microforms (ridge, mat, lawn). These changes in hydraulic parameters of peat may
consequently affect the magnitude, direction and rates of groundwater and surface water flows
(Siegel et al. 1995; Ferone and Devito 2004). Peatlands that experience prolonged periods of
drawdown also exhibit considerable fluctuations in water table position upon precipitation inputs
(Whittington and Price 2006). As peat becomes more compressed due to water table drawdown it
loses its ability to swell and subside with changes in water table position as the reduction of pore
water pressure results in irreversible compression (Shantz and Price 2006). Consequently, water
table variations are amplified relative to the surface due to this increase in peat rigidity
(Waddington et al. 2015; Whittington and Price 2006). These amplifications in peatland water
table and the increase in episodic periods of peat aeration may have subsequent impacts on
oxidation-reduction reactions that are controlled by water table fluctuations (such as carbon
mineralization and Hg methylation) as well as the mobility and transport of dissolved organic
matter (Fenner et al. 2007) and contaminants (including inorganic Hg and MeHg) from peatlands
to downstream aquatic ecosystems.
1.3.2 Peatland Ecological Effects due to Climate Change
In addition to the potential effects on hydrological parameters and processes, climate-induced
changes in hydrology may also significantly impact the ecology of peatland ecosystems with
regard to vegetation community composition and structure as well as the types and activity of
microbial populations responsible for organic matter decomposition. Although the potential
development of shallower relative water tables as a result of peat decomposition and
compression is possible with climate change, the dramatic fluctuations in water levels with
precipitation inputs following prolonged dry periods may exert considerable water stress on
Sphagnum mosses (Weltzin et al. 2000; Waddington et al. 2015). Vascular plants may have a
competitive advantage over Sphagnum moss species in such situations of prolonged water stress
given their physiological attributes (Chimner et al. 2016). Due to their vascular structures with
extensive rooting systems, vascular plants can access deeper stores of water whereas bryophytes
must be in close contact with water for survival. The rooting systems of vascular plants such as
sedges may also act to shuttle oxygen to the rhizosphere, increasing the zone of aeration, and act
as a conduit for gaseous carbon emissions providing a positive feedback mechanism to enhanced
15
water table recession (Bridgham et al. 2008; Strack et al. 2006; Waddington et al. 2015; Weltzin
et al. 2000). Although mosses may be able to regulate available water for vascular plants by
energy balance feedback mechanisms, increased shading by vascular plants over Sphagnum
mosses may further exacerbate the stress placed upon moss species and reduce their resilience
(Turetsky et al. 2012). The timing and coverage of such changes in vegetation communities may
be governed by spatial heterogeneity as a result of natural peatland microforms such as
hummocks, lawns and hollows (Strack et al. 2006). In currently wet pool locations in peatlands it
has been hypothesized that the forecasted drawdown in water table levels may act to increase
carbon accumulation in these microforms over time as a result of ecological succession and
colonization (Strack et al. 2006). Fenner et al. (2007) observed that increasing temperatures and
enhanced atmospheric CO2 concentrations exhibited an additive effect on vascular plant
dominance as a result of water table drawdown and CO2 fertilization which favoured the
colonization and growth of vascular plants due to their ability to access deeper water stores
during periods of water stress as compared to Sphagnum species. Both graminoids, including
sedges, and Ericaceae shrubs may increase in productivity and abundance under drier conditions
depending upon the level of water stress (e.g. Dieleman et al. 2015; Potvin et al. 2015).
Succession of plant communities may also play an important role in carbon biogeochemical and
hydrological processes under future climate scenarios. Upon senescence and through root
exudation, vascular plants may provide more labile organic matter to the peat surface and
subsurface as compared to Sphagnum plant material; which is more chemically-recalcitrant for
microbial mineralization (Scheffer et al. 2001; Haynes et al. 2015). The resiliency mechanisms
inherent in peatland ecosystems which act to maintain slow rates of carbon mineralization such
as the recalcitrant Sphagnum-derived organic matter substrate and the anoxic soils resulting from
high water tables may be significantly impacted if there were to be a shift in the dominant
peatland plant communities from Sphagnum-to vascular-dominated. In addition to potentially
enhancing fluxes of CO2 to the atmosphere, the delivery of organic matter derived from vascular
plants to the underlying peat may alter the hydrological characteristics of the peat over time.
Changes in vegetation type over the long-term may affect seasonal water storage as a result of
the reduced compressibility of peat containing woody roots or rhizomes (Whittington et al.
2007). The reduced compressibility of vascular-derived peat may therefore also yield higher
hydraulic conductivities as compared to Sphagnum-derived peat (Limpens et al. 2008).
16
Consequently, hydrological flowpaths may be affected and flow rates increased, which may in
turn impact the redox conditions through enhanced aeration. As a result, these hydrological
changes could have important positive feedbacks on nutrient and contaminant mobility and
transport from these systems as well as impacting transformations of these constituents within
the peat by potentially reducing the residence time of nutrients and contaminants in the peat
soils.
Climate-induced changes in peatland hydrology significantly impact microbial community
structure, which subsequently affects peat decomposition and the ability of peatlands to sequester
carbon (Nunes et al. 2015; Peltoniemi et al. 2015). Microbial (bacterial, fungal and archaeal)
populations present in peatlands may also be altered due to the organic matter quality of the
substrate in both the surface and subsurface. Changes in temperature and ambient CO2
concentrations may govern the activity and community composition of microbial populations
given their dependence on temperature for optimal functioning. Peltoniemi et al. (2009)
determined that over the long-term, fungal and actinobacterial communities become more similar
between peatland types of different nutrient status and hydrological conditions. Actinobacteria
appear to be less sensitive to hydrological changes, such as water table drawdown (Peltoniemi et
al. 2009). With regard to the mineralization of organic matter substrate, there may be minimal
change in microbial usage of different quality substrates due to the potential observed pattern of
functional redundancy (Strickland and Rousk 2010), suggesting no change in carbon
mineralization and CO2 production with changing vegetation across peatland types. Recent
research suggests that microbial communities can rapidly adapt to utilizing the available organic
matter substrate (Haynes et al. 2015).
The shift toward the dominance of vascular plants at the expense of Sphagnum in peatlands over
time due to climate change may lead to enhanced litter decomposition, increased heterotrophic
respiration and greater hydraulic conductivity of the upper surface peat layers (Limpens et al.
2008). The increase in vascular plants is expected to enhance carbon mineralization through the
provision of a more labile carbon source, as compared to Sphagnum, to the peat surface and to
the subsurface through root exudates (Ise et al. 2008; Waddington et al. 2015). Increasing
microbial activity with the concurrent increasing temperatures and greater peat aeration through
water table fluctuations will also contribute to this greater expected carbon turnover. The loss of
negative feedback mechanisms that facilitate the carbon-accumulating nature of peatland
17
ecosystems may be both the direct and indirect results of the impacts of a changing climate on
the hydrological processes and ecological function of these systems (Limpens et al. 2008;
Waddington et al. 2015).
1.3.3 Potential Impacts on Mercury Cycling
The hydrological and redox conditions of peatlands facilitate the production of MeHg (Tjerngren
et al. 2012). In concert with the potential hydro-climatic impacts of global climate change, the
process of Hg methylation and the export of MeHg to downstream aquatic ecosystems may be
significantly impacted (Krabbenhoft and Sunderland 2013). Multiple synergistic and antagonistic
processes impacted by a changing climate may differentially affect Hg cycling in peatlands
(Figure 1-2). Therefore, forecasting the potential direction and magnitude of such hydrological,
biogeochemical and ecological changes on Hg cycling is challenging, but warranted given that
peatlands accumulate on the order of 7 to 44 µg total Hg m-2
y-1
including litterfall (Kolka et al.
2011).
18
Figure 1-2 Potential synergistic and antagonistic implications of climate-induced peatland
hydrologic changes and hydraulic parameters (white boxes) and their potential impacts on Hg
cycling and transport (grey boxes). Solid arrows represent direct effects and dashed arrows
represent indirect or potential effects. (Mercury implications added to climate-induced peatland
hydrological changes that are modelled after Whittington and Price 2006).
Increased water table drawdown has the potential to reduce the zone of suitable anoxic
conditions for Hg methylation as a result of increased peat aeration. However, self-regulating
negative feedback mechanisms which alter peat hydraulic properties to minimize the impact of
external forces on peatlands, such as peat surface subsidence, may act to lessen the degree of
water table recession (Waddington et al. 2015) and therefore the zone of aerated peat. Lowered,
fluctuating water tables may in fact stimulate MeHg production due to the regeneration of
terminal electron acceptors, such as sulfate, required for Hg methylating microbial communities
(Coleman Wasik et al. 2015). Increased atmospheric and soil temperatures may also facilitate
greater production of MeHg in redox-suitable environments within the peat through the
stimulation of microbial activity (Benoit et al. 2003). The increasing contribution of vascular
plant material to the peat surface and subsurface over time due to a shift in plant communities
19
may act to increase MeHg production by providing a more labile source of carbon for microbial
methylators; which is required for methylation to occur (Mitchell et al. 2008). The increasing
prevalence of vascular vegetation may also positively feedback to cause further water table
recession as a result of oxygen transport to the rhizosphere through vascular tissues and
enhanced evapotranspiration (Waddington et al. 2015). Vascular plants with aerenchymous
tissues, such as sedges, may also prime aerobic organic matter decomposition as well as
anaerobic respiration by regenerating terminal electron acceptors via root oxygen loss (Mueller
et al. 2016) and in turn enhance MeHg production. Although likely to be strongly influenced by
the potential loss or interrupted periods of anoxia, these changes in the physical and chemical
conditions of the peat may be important in influencing the manner in which the process of Hg
methylation is impacted. Another consideration, although minimally explored in the literature, is
whether bacterial communities capable of Hg methylation may be impacted by the anticipated
climatic changes of increased temperatures and enhanced CO2 concentrations. New, cutting-
edge research in the field of microbial ecology by Parks et al. (2013) and Gilmour et al. (2013)
signals the potential future development of gene probe techniques to target Hg methylating
bacterial activity.
Greater fluctuations in water table levels due to changes in the frequency and intensity of
precipitation events as well as those resulting from climate-induced alterations in the hydraulic
characteristics of the peat, (e.g. hydraulic conductivity, bulk density and storage volume)
(Whittington and Price 2006) may considerably affect the mobility and transport of Hg and
MeHg to receiving systems downstream of peatland-dominated watersheds. Mercury mobility is
known to be enhanced from initially drier soils upon input of large precipitation events in both
terrestrial upland (Haynes and Mitchell 2015) and peatland environments (Grondin et al. 1995;
Gustin et al. 2006). Given the tightly coupled relationship between Hg and DOC (Driscoll et al.
1995; Skyllberg et al. 2000), Hg fluxes may be enhanced due to the increased export of DOC
(Fenner et al. 2007). Greater amplitude fluctuating water tables and the increase in episodic
periods of peat aeration may promote greater decomposition of organic matter (Ise et al. 2008;
Whittington and Price 2006). More labile organic matter as a result of enhanced decomposition
as well as that delivered from vascular root exudates may facilitate Hg partitioning from the solid
to dissolved phase. Collectively, these influences may lead to enhanced mobility and flushing of
Hg from peatlands (Figure 1-2).
20
Increasing temperatures and fluctuating water tables may act to increase soil-air Hg emissions
from peatlands due to increased soil drying, as has been observed from mineral soils (Gustin and
Stamenkovic 2005). Enhanced organic matter decomposition may also contribute to greater
exchange of Hg from the soil surface to the atmosphere (Figure 1-2). Shifting vegetation
communities may be instrumental in promoting Hg fluxes from the soil surface due to the
influence of vascular physiology. However, it is possible that the stomata of the vascular
vegetation, as well as non-stomatal mechanisms such as sorption to leaf surfaces, may influence
the exchange of Hg with the atmosphere (Lee et al., 2000; Rea et al., 2002; Ericksen et al., 2003;
Marsik et al., 2005; Stamenkovic and Gustin, 2009; Laacouri et al., 2013). Stomatal uptake and
other non-stomatal pathways may represent a competing mechanism resulting in potentially
greater Hg deposition from the atmosphere.
Climate change-induced hydrological impacts and the anticipated corresponding shifts in plant
community composition could significantly affect all aspects of the Hg cycle in peatlands;
including MeHg production, soil-air Hg evasion as well as Hg transport and mobility. Both the
hydrological and subsequent ecological impacts could influence peatland Hg cycling in different,
and potentially competing, manners in terms of magnitude and direction. Despite the
vulnerability of peatland ecosystems to climatic change and the significant stores of Hg in the
accumulated peat soils, very little field or experimental research in the literature is focused upon
the consequences of these changes on Hg cycling. This research will contribute in beginning to
close this considerable gap in the literature.
1.4 Objectives, Research Questions and Study Sites
The overall objective of this research is to assess the potential impacts of climate change;
specifically water table drawdown, shifting plant functional groups to more vascular-dominated
communities and increasing soil temperatures on Hg cycling in peatlands through mesocosm-
and field-scale experimental research. To address this objective, research was conducted at two
peatland sites at the southern extent of the boreal ecosystem. The primary location was the
Mesocosm Facility, termed PEATcosm (Peatland Experiment at the Houghton Mesocosm),
established in Houghton, Michigan by the United States Department of Agriculture (USDA)
Forest Service in association with Michigan Technological University. Research was also
21
conducted in a spruce-Sphagnum bog (locally known as S1) at the Marcell Experimental Forest
in north-central Minnesota, which is now part of the SPRUCE (Spruce and Peatland Responses
Under Climatic and Environmental change) experiment; funded and managed by Oak Ridge
National Laboratory (ORNL), the United States Department of Energy, Office of Science,
Biological and Environmental Research and the USDA Forest Service.
In this thesis, I set out to address the following questions/objectives:
1) How will the accumulation of MeHg and inorganic Hg, within peatland pore waters and
export from peatlands during spring snowmelt, be affected by the combined effects of, and
interaction between, altered plant functional groups and greater amplitude fluctuating water
tables as are anticipated with climate change?
2) Characterize total gaseous Hg fluxes in an ombrotrophic bog peatland. How will soil-air Hg
fluxes in a peatland be impacted by deep (~2 m) peat warming across a range of temperatures up
to +9°C?
3) How will soil-air Hg fluxes be affected by altered plant functional groups and lowered
peatland water tables as are anticipated with climate change?
4) What role will different plant functional groups play in controlling the direction and
magnitude of total gaseous Hg (TGM) exchange between peatlands and the atmosphere?
5) How will solid peat MeHg and inorganic Hg concentrations, throughout the peat profile, be
impacted by lowered water tables, altered plant functional groups as well as the interactive
effects of such hydrological and ecological changes?
6) How will climate change-induced fluctuations in water table regime and the resultant shift in
vascular vegetation impact peat Hg methylation and MeHg demethylation potentials?
7) Characterize and compare the Hg isotopic signatures in surface peat from two sub-boreal
peatlands.
8) What processes govern Hg deposition and re-emission in two contrasting sub-boreal
peatlands?
22
Chapter 2 addresses Question #1.
Chapter 3 addresses Questions #2, 3, 4.
Chapter 4 addresses Questions #5, 6.
Chapter 5 addresses Questions #7, 8.
All four research chapters are based on data collected from the PEATcosm climate change
experiment. The PEATcosm experiment involved manipulations of water table position and
vascular plant communities to simulate anticipated climate change impacts on twenty-four cubic
metre (1 m x 1 m x 1 m) peat monoliths. The peat was harvested from a bog peatland in
Meadowlands, Minnesota in May 2010 and placed in Teflon-lined stainless steel mesocosm bins
in Houghton, Michigan (Figure 1-3). Each bin was insulated and the Mesocosm Facility was
underlain by a climate-controlled tunnel which facilitated the maintenance of natural peat
temperature profiles (Figure 1-4).
Figure 1-3 PEATcosm Mesocosm Facility in Houghton, Michigan with 24 cubic metre peat
mesocosm bins.
23
Figure 1-4 Climate-controlled tunnel beneath Mesocosm Facility allowing access to the 24
mesocosm bins. Each bin was equipped with overflow tubing and collection bin to allow for
runoff sampling from each mesocosm, particularly during the snowmelt period.
The PEATcosm study comprises a full-factorial experimental design with two water table (WT)
prescriptions crossed with three plant functional group treatments in a randomized complete
block design with four replicates per treatment combination (24 experimental units). The two
WT treatments include:
1) low variability, relatively high WT position (referred to as ‘high WT’ and used as the
control)
2) high variability, low WT position (referred to as ‘low WT’).
The water table treatments were based on long-term (approximately 50 years) data from the
Marcell Experimental Forest, with the two target water table seasonal profiles modelled after
typical low variability, average WT years (‘high WT’) and typical high variability, low water
table years (‘low WT’). These target WT profiles were maintained through a combination of
artificial precipitation additions (formulated to mimic rainwater chemistry at the nearby National
Atmospheric Deposition Program monitoring site at Chassell, MI), rain-exclusion covers and
regulated outflow from each of the bins during the spring snowmelt periods from the
24
approximate acrotelm-catotelm boundary. A piezometer nest was installed in each of the 24 bins
to allow for water sampling at three depths – 20, 40 and 70 cm below the peat surface.
The three plant functional group treatments were as follows:
1) all Ericaceae removed (referred to as ‘sedge only’ treatment)
2) all sedge removed (‘Ericaceae only’ treatment)
3) both sedge and Ericaceae present (Unmanipulated Control)
Vegetation manipulations were achieved and maintained by clipping stems of the excluded
species as deeply into the peat as possible while attempting to minimize peat disturbance. The
PEATcosm vascular plant functional group manipulations were initiated in 2011 and water table
experimental treatments were implemented in 2012. The experiment concluded in 2015 with the
destructive harvest of the peat monoliths.
Chapters 3 and 5 also include work conducted at the SPRUCE experiment in the Marcell
Experimental Forest in north-central Minnesota (Figure 1-5). During this research, an array of
large open-top enclosures, approximately 12 m in diameter, were installed in the S1 peatland
(Figure 1-5). In June 2014 through to August 2015, a preliminary study of deep soil warming,
approximately 2 m below the peat surface, was conducted within each enclosure plot.
Temperatures were manipulated in ten enclosures according to a regression design (duplicates of
+0, +2.25, +4.5, +6.75, +9°C above ambient temperatures). This deep warming experiment was
a pre-cursor to the full SPRUCE experiment, which now involves whole-ecosystem warming and
elevated atmospheric CO2 concentrations to simulate climate change within the bog.
25
Figure 1-5 Aerial view (left) of the SPRUCE plot footprints and boardwalks in the S1 bog,
Marcell Experimental Forest, Minnesota (September 2013 aerial photo from Paul Hanson, Oak
Ridge National Laboratory (ORNL)). One SPRUCE experimental plot (right).
1.5 Thesis Structure and Publication Information
1.5.1 Chapter 1
Chapter 1 provides background information on the aspects of the mercury cycle investigated in
this thesis in relation to climate change impacts. A brief overview on peatland hydrology is
given and the anticipated effects of climate change on peatland hydrology and ecology are
discussed. The potential implications of these climate-induced changes on mercury cycling are
hypothesized. Information on the publication status of each of the thesis chapters is also
provided.
Contributions: Written by K.M.H. with comprehensive review by C.P.J.M.
1.5.2 Chapter 2
Chapter 2 focuses on the anticipated climate change impacts of lowered water tables and shifting
vascular plant functional groups on Hg and MeHg mobility in peat pore waters and transport in
snowmelt runoff. This paper is currently in press at Global Biogeochemical Cycles (31, DOI:
10.1002/2016GB005471). The co-authors on this paper are Evan S. Kane (USDA Forest Service
Northern Research Station Houghton, MI and Michigan Technological University), Lynette
26
Potvin, Erik A. Lilleskov (USDA Forest Service Northern Research Station Houghton, MI),
Randall K. Kolka (USDA Forest Service Northern Research Station Grand Rapids, MN) and
Carl P. J. Mitchell.
Contributions: The PEATcosm experiment was designed and implemented by E.A.L., E.S.K.,
L.P. and R.K.K. Research questions and experimental approach for this chapter were articulated
by K.M.H. with input from C.P.J.M. Pore water and snowmelt runoff samples were collected by
L.P. with assistance from E.S.K. and K.M.H. All mercury laboratory analyses were performed
by K.M.H. Sulfate and phenolics pore water data and cellulose decomposition assay data were
provided by L.P. and E.S.K. Data analyses and data interpretation were performed by K.M.H.
with input from C.P.J.M. The manuscript was prepared by K.M.H. with comprehensive review
by C.P.J.M. All co-authors L.P., E.S.K., E.A.L. and R.K.K. provided revisions to the manuscript
prior to submission to Global Biogeochemical Cycles.
1.5.3 Chapter 3
Chapter 3 investigates the potential impacts of deep peat warming as well as lowered water
tables and altered vascular plant communities on total gaseous Hg exchange with peat soils. This
paper is published in Atmospheric Environment (2017, Vol. 154, pg. 247-259,
http://dx.doi.org/10.1016/j.atmosenv.2017.01.049). The co-authors on this paper are Evan S.
Kane (USDA Forest Service Northern Research Station Houghton, MI and Michigan
Technological University), Lynette Potvin, Erik A. Lilleskov (USDA Forest Service Northern
Research Station Houghton, MI), Randall K. Kolka (USDA Forest Service Northern Research
Station Grand Rapids, MN) and Carl P. J. Mitchell.
Contributions: The PEATcosm experiment was designed and implemented by E.A.L., E.S.K.,
L.P. and R.K.K. Access to the SPRUCE site was facilitated by R.K.K. Research questions and
experimental approach for this chapter were articulated by K.M.H. with input from C.P.J.M.
Total gaseous mercury flux measurements at PEATcosm were performed by K.M.H. with
assistance from L.P. and E.S.K. SPRUCE flux measurements were performed by K.M.H.
Vegetation samples from PEATcosm were collected by L.P. All mercury laboratory analyses
were performed by K.M.H. Data analyses and data interpretation were performed by K.M.H.
27
with input from C.P.J.M. The manuscript was prepared by K.M.H. with comprehensive review
by C.P.J.M. All co-authors L.P., E.S.K., E.A.L. and R.K.K. provided revisions to the manuscript
prior to submission to Atmospheric Environment.
1.5.4 Chapter 4
Chapter 4 examines the impacts of lowered, fluctuating water tables and shifting vascular plant
communities on solid phase peat Hg and MeHg including partitioning to pore water as well as
methylation and demethylation rate potentials. This paper has been submitted to Biogeochemistry
and is currently under review.. The co-authors on this paper will be Evan S. Kane (USDA Forest
Service Northern Research Station Houghton, MI and Michigan Technological University),
Lynette Potvin, Erik A. Lilleskov (USDA Forest Service Northern Research Station Houghton,
MI), Randall K. Kolka (USDA Forest Service Northern Research Station Grand Rapids, MN)
and Carl P. J. Mitchell.
Contributions: The PEATcosm experiment was designed and implemented by E.A.L., E.S.K.,
L.P. and R.K.K. Research questions and experimental approach for this chapter were articulated
by K.M.H. with input from C.P.J.M. Peat samples were collected by E.A.L., L.P. and E.S.K.
Peat coring and isotopic incubations for methylation-demethylation assays were performed by
K.M.H. with input from C.P.J.M. All mercury laboratory analyses were performed by K.M.H.
Data analyses and data interpretation were performed by K.M.H. with input from C.P.J.M. The
manuscript was prepared by K.M.H. with comprehensive review by C.P.J.M. All co-authors
L.P., E.S.K., E.A.L. and R.K.K. provided revisions to the manuscript prior to submission to
Biogeochemistry.
1.5.5 Chapter 5
Chapter 5 applies Hg isotopic analyses to compare the sources of Hg to the two sub-boreal
peatlands study sites and the processes influencing the fractionation of natural abundance stable
Hg isotopes in the surface peat. This paper is currently being prepared for submission to Global
Biogeochemical Cycles. The co-authors on this paper will be Evan S. Kane (USDA Forest
28
Service Northern Research Station Houghton, MI and Michigan Technological University),
Lynette Potvin, Erik A. Lilleskov (USDA Forest Service Northern Research Station Houghton,
MI), Randall K. Kolka (USDA Forest Service Northern Research Station Grand Rapids, MN),
Wang Zheng (University of Toronto), Bridget A. Bergquist (University of Toronto) and Carl P.
J. Mitchell.
Contributions: The PEATcosm experiment was designed and implemented by E.A.L., E.S.K.,
L.P. and R.K.K. Access to the S1 (SPRUCE) site was facilitated by R.K.K. Research questions
and experimental approach for this chapter were articulated by K.M.H. with input from C.P.J.M.
Peat samples were collected by E.A.L., L.P. and E.S.K. at PEATcosm and K.M.H. at the S1
(SPRUCE) peatland. Isotopic analyses were performed in B.A.B.’s laboratory at the University
of Toronto. All mercury laboratory analyses were performed by W.Z. and K.M.H. Data
analyses were performed by W.Z. and data interpretation was performed by K.M.H. with input
from C.P.J.M. and B.A.B. The manuscript was prepared by K.M.H. with comprehensive review
by C.P.J.M.
1.5.6 Chapter 6
Chapter 6 summarizes and synthesizes the results of the thesis. Areas of future research to
further explore the impacts of climate change on peatland Hg cycling are also addressed.
Contributions: Written by K.M.H. with comprehensive review by C.P.J.M.
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wetland ecosystems. Environmental Toxicology and Chemistry, 12, 2245-2264.
45
Chapter 2 Mobility and Transport of Mercury and Methylmercury in Peat as a Function of Changes in Water Table Regime and Plant Functional
Groups
2
2.1 Abstract
Climate change is likely to significantly affect the hydrology, ecology and ecosystem function of
peatlands, with potentially important but unclear impacts on mercury mobility within and
transport from peatlands. Using a full-factorial mesocosm approach we investigated the potential
impacts on mercury mobility of water table regime changes (high and low) and vegetation
community shifts (sedge-dominated, Ericaceae-dominated or unmanipulated control) in peat
monoliths at the PEATcosm mesocosm facility in Houghton, Michigan. Lower and more
variable water table regimes and the loss of Ericaceae shrubs act significantly and independently
to increase both total Hg (THg) and methylmercury (MeHg) concentrations in peat pore water
and in spring snowmelt runoff. These differences are related to enhanced peat decomposition and
internal regeneration of electron acceptors which are more strongly related to water table regime
than to plant community changes. Loss of Ericaceae shrubs and an increase in sedge cover may
also affect Hg concentrations and mobility via oxygen shuttling and/or the provision of labile
root exudates. Altered hydrological regimes and shifting vegetation communities, as a result of
global climate change, are likely to enhance Hg transport from peatlands to downstream aquatic
ecosystems.
2.2 Introduction
Boreal peatlands store vast amounts of carbon, are critical to hydrological cycling, and are
landscape scale “hot spots” for the storage and biogeochemical transformations of pollutants
such as mercury (Hg) in the landscape (Gorham, 1991; Branfireun et al., 1998; Limpens et al.,
2008). Peatland systems act as strong sinks of atmospherically deposited inorganic Hg (Kolka et
al., 2001; Grigal, 2003). With predominantly saturated, anoxic, organic peat soils and the
availability of electron acceptors such as sulfate, the hydrological and redox conditions of these
46
systems also provide an optimal setting for the production of methylmercury (MeHg), a potent
neurotoxin which bioaccumulates and biomagnifies up the food chain (Branfireun et al., 1996;
Scheuhammer et al., 2007; Tjerngren et al., 2012). Considerable research has focused on the
influence of enhanced sulfate deposition on the process of Hg methylation in peatland
ecosystems as sulfate-reducing bacteria are known to actively methylate Hg (Gilmour et al.,
1992; Bergman et al., 2012; Coleman Wasik et al., 2015). Recent research has confirmed that in
addition to sulfate- and iron-reducing bacteria, some methanogens and syntrophic, acetogenic
and fermentative Firmicutes are capable of Hg methylation, including those that have been
isolated from northern peatland ecosystems (Gilmour et al., 2013).
Peatlands, particularly those located at high latitudes, are sensitive to increasing temperatures
and altered precipitation regimes resulting from global climate change (Bridgham et al., 1995;
Ise et al., 2008; Limpens et al., 2008). The susceptibility of peatlands to climatic change may
depend upon the relative contributions of groundwater versus precipitation inputs. Bogs,
wherein water table position and peat accumulation capacity are governed strongly by the
balance between precipitation and evapotranspiration, are particularly vulnerable to climate
change (Winter, 2000). Climate-induced changes in the hydrological regimes of peatlands
significantly influence microbial community structure, which may have profound effects on peat
decomposition and carbon sequestration (Nunes et al., 2015; Peltoniemi et al., 2015). Changes in
bacterial community structure also have important links to methylmercury production in
peatlands (Strickman et al., 2016). Alterations to the carbon storage abilities of peatlands due to
climate-induced changes in the hydrology and ecology of these systems may affect the
biogeochemical cycling and mobility of Hg within, and export from, these wetland systems.
With increased variability in precipitation patterns predicted to occur in the northern continental
United States, prolonged periods of water table recession particularly during the summer months
may have important ecological implications (Kunkel et al., 2003; Groisman et al., 2005;
Thomson et al., 2005). For example, these hydrological changes may result in a shift in peatland
vegetation communities toward vascular-dominated functional groups (Weltzin et al., 2003;
Strack et al., 2006; Breeuwer et al., 2009; Dieleman et al., 2015). Different plant functional
groups including graminoids and Ericaceae shrubs may be more suited to the hydrological
conditions resulting from climate change. Although the development of shallower relative water
tables as a result of peat decomposition and compression is possible, dramatic fluctuations in
47
water levels when precipitation inputs follow prolonged dry periods may exert considerable
water stress on bryophytes including Sphagnum mosses (Weltzin et al., 2000; Whittington and
Price, 2006; Waddington et al., 2014). Vascular plants may have a competitive advantage over
Sphagnum moss species in such situations of prolonged water stress because of their ability to
access deeper stores of water through their more extensive rooting systems (Dieleman et al.,
2015). Vascular plants can also regulate water loss through their stomata, while Sphagnum
mosses do not possess stomata (Breeuwer et al., 2009). The ecophysiological traits of the
established plant functional groups may influence the biogeochemistry of the peat and the
regulation of ecosystem carbon dynamics. The rooting systems of some vascular plants such as
sedges may also act to shuttle oxygen to the rhizosphere, increasing the zone of peat aeration and
acting as a conduit for gaseous emissions (Weltzin et al., 2000; Strack et al., 2006; Bridgham et
al., 2008; Waddington et al., 2014). Shallowly-rooted ericaceous shrubs lack the adaptive
structures to survive in flooded conditions (Chapin et al., 1996) and therefore may thrive under
drier conditions. However, growth of shrubs may be hindered during periods of prolonged water
stress with significant water table recession, particularly in association with increasing
temperatures (Armstrong et al., 1991; Weltzin et al., 2003). In contrast, the deep root systems of
sedges may allow this plant functional group to survive and out-compete Ericaceae shrubs during
prolonged drought periods (Dieleman et al., 2015). The complex synergistic and antagonistic
feedback mechanisms limit the ability to predict the response of peatland vegetation
communities to climate change-induced water table recession. Therefore, both the sedge and
Ericaceae plant functional groups should be considered, both individually and coexisting, when
investigating the effects of shifting plant communities.
Water table recession and fluctuation, increased peat aeration, and shifting plant communities
may significantly affect Hg cycling and the process of Hg methylation due to alterations in redox
conditions of the peat and the internal recycling of sulfur species (Coleman Wasik et al., 2015).
The potential increased liberation of dissolved organic matter (DOM) may act to augment Hg
methylation through the provision of labile carbon substrates needed as electron donors for
methylating microbes (Mitchell et al., 2008a; Graham et al., 2013). The establishment of deeply-
rooted vascular plants such as sedges may also contribute an additional labile carbon source
through root exudates, stimulating methylating microbial communities (Bridgham et al., 1995;
Windham-Myers et al., 2009). Increased Hg mobilization via peat decomposition, partitioning,
48
and/or desorption into the pore water phase as well as Hg complex formation with DOM may
impact downstream aquatic ecosystems with the potential for enhanced transport and loading
particularly during high-flow events such as spring snowmelt (Bishop et al., 1995; Mitchell et al.,
2008c; Haynes and Mitchell, 2012). Moreover, the complexation of Hg with DOM, particularly
on thiol functional groups, increases Hg solubility and mobility, and possibly the bioavailability
of Hg for methylation (Skyllberg et al., 2000; Graham et al., 2013). Despite the propensity of
peatland ecosystems to methylate and export Hg to downstream ecosystems, minimal research
has been conducted to explore how Hg mobility and transport from peatlands may be affected by
climate change.
The purpose of this study was to investigate the potential effects of climate change-induced shifts
in water table regime and plant community composition on the mobility/movement of Hg within
and transport from peatland ecosystems. A novel peatland mesocosm experiment known as
PEATcosm (Peatland Experiment at the Houghton Mesocosm Facility) allowed for the
manipulation of water tables and plant community composition to simulate some of the potential
effects of climate change on peatland ecosystem functioning, including Hg cycling. Investigation
at the mesocosm scale provides a process-level understanding, which may be applied to the
landscape-scale, of the influences of hydrological and vegetation changes on the movement of
Hg within and export from peatlands.
2.3 Materials and Methods
2.3.1 Study Site and Experimental Design
The PEATcosm Mesocosm Facility is located in Houghton, Michigan, USA (47.11469° N,
88.54787° W) on the property of the United States Department of Agriculture (USDA) Forest
Service Northern Research Station - Forestry Sciences Laboratory. The regional climate is humid
continental with typical annual precipitation of approximately 870 mm at the site. Mean
temperatures in this area range from -13°C in January to 24°C in July. The average growing
season lasts approximately 132 days (30 year means at Houghton County Airport; NOAA
National Climatic Data Center) (Potvin et al., 2015).
Twenty-four intact 1-m3 (1 m x 1 m x 1 m) peat monoliths were extracted from an ombrotrophic
peatland located in Meadowlands, Minnesota, USA in May 2010. Monoliths were placed into
49
individual mesocosm bins, transported to and subsequently installed in the PEATcosm facility.
The stainless steel interior of each bin was Teflon-coated to prevent direct contact and potential
metal transfer between the bin and the peat. The top of each bin was open, exposing the peat to
ambient climate conditions. The mesocosm bins were inserted into a climate-controlled tunnel
and insulated on the sides, which allowed belowground access to one face of each of the bins as
well as to facilitate a vertical temperature gradient, as would be observed in a natural peatland
profile. Potvin et al. (2015) provides a detailed and comprehensive overview of the PEATcosm
experiment including the peat harvest, experimental design and treatment maintenance. The
PEATcosm study was a full-factorial experimental design with two water table (WT)
prescriptions crossed with three vascular plant functional group treatments in a randomized
complete block design, with four replicates per treatment combination, resulting in a total of 24
experimental units (Table 2-1). The plant functional group treatments were designed to
distinguish between effects of Ericaceae (non-aerenchymous shallowly-rooted shrubs with
enzymatically competent mycorrhizal root symbionts) and sedges (aerenchymous herbaceous
graminoids with non-mycorrhizal roots that penetrate deeper into the saturated peat).
Table 2-1 PEATcosm full-factorial experimental design. N=4 mesocosm bins per crossed
treatment.
WATER TABLE (WT) POSITION
PLANT
FUNCTIONAL
GROUP
HIGH WT LOW WT
SEDGE ONLY High WT Sedge Low WT Sedge
ERICACEAE ONLY High WT Ericaceae Low WT Ericaceae
CONTROL
(UNMANIPULATED) High WT Control Low WT Control
The water table treatments were based on long-term (approximately 50 years) data from the
Marcell Experimental Forest in north-central Minnesota (47.51907° N, 93.45966° W), located
50
near the peat monolith harvest site. The two target water table seasonal profiles were modelled
after typical variability, average water table years (‘high WT’) and typical high variability, low
water table years (‘low WT’). The mean difference between the high and low WT positions was
approximately 20 cm throughout the experiment. These target water table profiles were
maintained via a combination of artificial precipitation additions, rain-exclusion covers, and
regulated outflow (spring-only, from ~25 cm depth, roughly at the acrotelm-catotelm boundary)
from each of the bins (Potvin et al., 2015). Water table manipulations were performed only
during the summer months and water table positions were allowed to stabilize throughout the
winter months. Figure 2-1 displays the mean water table positions for the low and high WT
treatments as well as the precipitation throughout the course of this study from 2013 to 2015.
Figure 2-1 Mean water table positions (in cm below the peat surface) in the low and high WT
treatment mesocosms and precipitation (in mm) over the course of this study from 2013 to 2015.
Water table manipulations were conducted in the summer months, while water table levels were
left to stabilize during the winter months. Dashed lines around the low and high WT treatment
mean water table positions represent the 95% confidence interval. Pore water and snowmelt
sampling events are denoted.
51
The three vascular plant functional group treatments simulated anticipated climate change-
induced community composition (Chapin et al., 1996; Weltzin et al., 2000; Strack et al., 2006)
and included 1) sedge only (all Ericaceae removed), 2) Ericaceae only (all sedge removed), and
3) unmanipulated control (includes both sedge and Ericaceae) (see Figure A-2-1 for
photographs). The plant functional group treatments were initiated in June 2011 and maintained
weekly by clipping the stems of the excluded species. The dominant sedge species present was
Carex oligosperma Michx., while the dominant ericaceous shrubs included Chamaedaphne
calyculata (L.) Moench., Kalmia polifolia Wangenh., and Vaccinium oxycoccos L.. The mosses
Sphagnum rubellum Wilson, S. magellanicum Brid., S. fuscum (Schimp.) Klinggr and
Polytrichum strictum Brid. comprised the dominant bryophyte species in all of the 24
mesocosms. To a lesser extent, P. commune Hedw., Eriophorum vaginatum L., Andromeda
polifolia L. var. glaucophylla (Link) DC., Rhododendron groenlandicum (Oeder) Kron and Judd,
and Drosera rotundifolia L. were also present.
2.3.2 Water Sampling
Pore waters were sampled repeatedly throughout the 2013, 2014 and early 2015 growing seasons
to monitor the influence of the simulated climate change effects on Hg and MeHg mobility. Pore
waters were collected from a micro-piezometer nest (ultra-high-density polyethylene casing with
Teflon tubing) installed in each of the 24 mesocosms at three depths – 20 cm, 40 cm and 70 cm
below the peat surface (see Romanowicz et al. (2015) for details on piezometer construction).
Sampling was conducted in June, August and November 2013, May, July and September 2014,
and in May 2015 prior to the mid-summer destructive harvest of the mesocosms at the
conclusion of the experiment. Due to low mid-summer water tables preventing the collection of
samples for analysis from the 20 and 40 cm depths in the Low WT bins, only the spring and fall
data is considered throughout this analysis (see Text A-2-1 in Appendix A for full details).
Spring snowmelt runoff was also collected in 2014 and 2015 to examine impacts on the potential
export of Hg from peatlands to downstream aquatic ecosystems. Outflow or runoff from each
mesocosm was initiated by sufficiently high water tables sensed via a pressure transducer and
controlled via an outflow line located at about 25 cm depth below the peat surface (approximate
acrotelm-catotelm boundary), draining into plastic tubs in the belowground tunnel. Large rain
events and spring snowmelt triggered the production of runoff. However, only runoff produced
as a result of snowmelt was collected for the purposes of this study.
52
Pore water samples were collected from the micro-piezometers using a portable peristaltic pump
equipped with a Teflon line rinsed with dilute hydrochloric acid (HCl) prior to sampling. Acid-
cleaned Teflon in-line filtering units (Savillex, Eden Prairie, MN) were attached to the pump line
to filter all water samples to a level of 0.7 μm using ashed glass fiber filters (Lewis and Brigham,
2004; Shanley et al., 2008). Samples were collected in new polyethylene terephthalate glycol
(PETG) bottles, acidified to 0.5% by volume with trace metal grade HCl and stored at 4°C in
darkness until analysis. Method blanks were collected by running deionized water through the
cleaned pump line and filter units during each sampling event to ensure the cleanliness of the
sampling equipment and sample handling. Mesocosm runoff was similarly collected during the
snowmelt period from each bin outflow line, preserved and stored until analysis. Ultra-clean
trace metal techniques were followed during sample collection, laboratory handling and analyses
(Shanley et al., 2008).
Pore waters from 20, 40 and 70 cm depths and mesocosm runoff were individually analyzed for
both total Hg (THg) and MeHg concentrations. Ancillary chemical analyses including dissolved
organic carbon (DOC), chloride, sulfate and total phenolic concentrations were determined for
pore water and runoff samples collected at the same time as those for Hg analyses.
2.3.3 Pre-Treatment Peat
Peat was collected from each of the 24 mesocosms in 2011 prior to the initiation of the water
table and plant functional group prescriptions to assess any pre-treatment trends in solid phase
THg and MeHg concentrations among the monoliths. Peat cores were collected with a stainless
steel corer from the surface to 60 cm depth and were sectioned in 10 cm increments.
2.3.4 Peat Decomposition Assays
To assess peat decomposition potential for each of the crossed water table and plant functional
group treatments, wooden dowels with strips of cellulose filter paper (pre-dried at 55°C and
weighed) attached at 10 cm increments using heat shrink tubing and encased in Nylon mesh
netting (20x20, 67% open area) were installed to a depth of 80 cm. Two rods were deployed in
each of the peat monoliths for the duration of the 2014 growing season (n=48 total). They were
removed and kept frozen until processed. Peat adhering to the rods upon removal was gently
53
rinsed off. The cellulose strips were removed, dried at 55°C, and weighed to determine the
percentage mass lost in relation to the initial weight.
2.3.5 Analytical Methods
Total Hg concentrations of the pore water and runoff samples were determined with a Tekran
Model 2600 automated Total Mercury Analyzer using cold vapor atomic fluorescence
spectroscopy (CVAFS) according to US EPA Method 1631 (US EPA Method 1631, 2002).
Recovery of a THg spike was 94.5 ± 7.9% (mean ± standard deviation, n=33), replication of
duplicates was 1.7 ± 1.4% (n=32) and the detection limit, calculated as three standard deviations
of matrix blanks, was 0.25 ng L-1
(n=140). Freeze-dried peat samples were digested in hot nitric
acid and diluted digestates were similarly analyzed by CVAFS. Methylmercury analysis was
conducted by isotope dilution-gas chromatography-inductively coupled plasma mass
spectrometry (ID-GC-ICPMS; Hintelmann and Evans, 1997) on samples, both water and solid
phase peat, that were distilled in Teflon vessels according to US EPA Method 1630 (US EPA
Method 1630, 1998). During the distillation process a trace amount of enriched stable Me199Hg
isotope was added to each sample as an internal standard (Hintelmann and Evans, 1997).
Recovery of standard reference material (estuarine sediment ERM CC580) was 99.8 ± 8.6%
(n=50), replication of duplicates was 5.6 ± 4.9% (n=45) and the MeHg detection limit was
calculated (n=48 matrix blanks) to be 0.04 ng L-1
. Field blanks were below detection limits for
both THg and MeHg concentrations.
Sulfate and chloride were determined with an ICS-2000 ion chromatograph with an IonPac AS11
separator column (Dionex Corporation, Bannockburn, IL, USA). Sulfate and chloride pore water
analyses were conducted only for the 2013 and 2014 growing seasons. Dissolved organic carbon
was determined from piezometer samples filtered to 0.45 μm, acidified to pH 2, and analyzed
with a TOC-V Analyzer (Shimadzu Scientific Instruments, Columbia, MD, USA). Pore water
collection and analysis for total phenolics by microplate technique (with tannic acid standard)
were as previously described in Romanowicz et al. (2015).
2.3.6 Statistical Analyses
All statistical analyses were performed using R statistical software (R Development Core Team,
2014) with α = 0.05. The influence of water table regime and vascular plant functional group on
54
THg and MeHg concentrations, as well as %MeHg over the three sampling depths of pore waters
collected over the course of the experiment, was assessed using repeated measures analyses of
variance (ANOVA); with water table, plant functional group and sampling depth treated as main
factors and sampling event as the repeated factor. A repeated measures ANOVA was similarly
performed for the snowmelt runoff Hg concentration data collected from the bins over the two
sampling years. One-way ANOVAs were performed on the pre-treatment solid phase peat THg
and MeHg concentrations individually at 10-30 cm and 30-50 cm to assess any significant
differences among assigned treatment bins prior to implementation of the treatments. These peat
depths were selected for statistical analysis to coincide with the pore water sampling depths of 20
and 40 cm. Additional details regarding statistical analyses are provided in the Supplementary
Information (Text A-2-1).
2.4 Results and Discussion
Both THg and MeHg concentrations in pore waters and snowmelt runoff were significantly
affected by the experimental manipulation of water table (both p < 0.0001) and vascular plant
functional group (both p < 0.0001), with no significant interaction (p = 0.11 to 0.35; Figure 2-2)
between these factors indicating that changes in either are likely to have an impact on Hg
accumulation in pore waters and subsequently, for mobility to downstream regions. Specifically,
lower, more variable water table regimes significantly increased pore water and snowmelt runoff
THg and MeHg concentrations, as did the change to a sedge-dominated vascular plant
community composition with no Ericaceae shrubs present. The combined low WT – sedge only
treatment resulted in the highest pore water and runoff concentrations of both THg and MeHg
(Figure 2-2). One-time mass transfer loads were determined for the 2014 spring snowmelt period
by multiplying the THg and MeHg concentrations of the runoff by the amount of water drained
from the mesocosms acrotelm-catotelm boundaries (Table A-2-1). The trends in THg and MeHg
loads among the six crossed WT and vascular plant functional group treatments (Figure A-2-2)
were consistent with those observed for THg and MeHg pore water and runoff concentrations
(Figure 2-2). No significant, systematic differences were observed in solid phase peat THg or
MeHg concentrations at either the 10 – 30 cm or 30 – 50 cm depth intervals as these values were
quite variable among the six assigned experimental treatment combinations prior to the
experiment (Table 2-2). The differences observed in pore water and runoff THg and MeHg
concentrations can therefore be confidently attributed to the experimental manipulation.
55
Figure 2-2 Total Hg and MeHg concentrations and %MeHg in pore water and snowmelt runoff
in relation to water table and vascular plant functional group manipulations (treatment means of
all depths and sampling events). (a) pore water THg, (b) pore water MeHg, (c) pore water
%MeHg, (d) snowmelt runoff THg, (e) snowmelt runoff MeHg, and (f) %MeHg in snowmelt
runoff. Letters denote statistically similar groups based on transformed data. No significant
differences were observed in snowmelt runoff %MeHg across treatments. Significance of
treatment effects both individual (water table “WT” and plant functional group “Veg”) and
interactive (“WT*Veg”) for THg, MeHg, and %MeHg for both pore waters and snowmelt runoff
are noted. n = 318 pore water samples total (n = 50–56 per treatment).
56
Table 2-2 Mean ± standard deviation THg and MeHg concentrations (in ng g-1
) in solid-phase
peat prior to experimental manipulations (n = 8 samples per treatment and depth increment).
THg Concentration (ng g-1
) MeHg Concentration (ng g-1
)
10 – 30 cm 30 – 50 cm 10 – 30 cm 30 – 50 cm
High WT
Ericaceae 91.8 ± 63.4 87.8 ± 38.4 5.4 ± 6.6 5.9 ± 5.7
High WT
Sedge 82.8 ± 27.0 104.7 ± 40.5 5.0 ± 3.6 8.1 ± 6.7
High WT
Control 80.9 ± 33.3 91.3 ± 21.3 4.5 ± 6.0 5.8 ± 4.1
Low WT
Ericaceae 65.3 ± 15.5 103.3 ± 28.2 1.0 ± 0.7 6.6 ± 6.4
Low WT
Sedge 92.0 ± 33.3 83.2 ± 22.6 5.6 ± 3.9 3.5 ± 1.8
Low WT
Control 75.9 ± 23.8 100.2 ± 28.9 2.8 ± 3.0 6.8 ± 5.9
For both THg and MeHg pore water concentrations, the effect of water table significantly
interacted with both the depth (p = 0.0002 to 0.008) and the timing of the sampling event (p <
0.0001 to 0.002). Total Hg and MeHg concentrations decreased from the 20 cm through to the 70
cm sampling depths in the low WT bins whereas THg and MeHg concentrations at the three
sampling depths were not significantly different from one another in the high WT bins. Greater
fluctuations in water table position in the low WT bins (mean position approximately 35 cm
below the peat surface with a standard deviation of 12 cm), likely resulted in enhanced leaching
and mobility of Hg into the pore waters due to greater aerobic decomposition of the surface peat
layers, and may account for the observed differences in Hg concentrations with depth. These
differences are not the result of evaporative concentration of Hg under low water table conditions
as there was no significant difference in chloride pore water concentrations between the two
water table treatments (p = 0.76). In contrast, the comparatively minimal water table variability
in the high WT bins (mean position approximately 14 cm below the peat surface with a standard
deviation of 5 cm) reduced this mobilization, likely as a result of reduced leaching, yielding both
57
lower and more temporally consistent pore water THg, MeHg and DOC concentrations at all
sampling depths. In association with the effect of WT position, THg and MeHg concentrations
were slightly higher during the early growing season sampling events as compared to the fall
samplings. This may be due to considerable water table variability as a result of draining at the
beginning of each growing season to re-establish the water table treatments following spring
snowmelt. Despite this seasonal variability, a consistent trend in mean THg and MeHg
concentrations was observed among the six crossed WT and vascular plant functional group
treatments in each of the five sampling events from 2013 through to 2015 (Figures A-2-3 and A-
2-4). A marginally significant interaction (p = 0.02) was observed for THg concentrations only
between the vascular plant functional group and the sampling depth. The influence of vegetation
on Hg mobility appears to be more evident nearer to the rooting zone; particularly with the
removal of Ericaceae shrubs.
Water table position (p < 0.0001) and vascular plant functional group (p < 0.0001) also
significantly affected the percentage of THg found as MeHg in pore waters (Figure 2-2c). The
observed differences in %MeHg may be the result of the greater solubility of MeHg in water as
compared to inorganic Hg (Morel et al., 1998). However, the significant differences in pore
water %MeHg among treatments suggest that differences in Hg methylation may also be caused
by the water table and plant functional group treatments. In contrast to the pore waters, no
significant effect of either water table (p = 0.62) or vascular plant functional group treatment (p =
0.77) was observed for the %MeHg in mesocosm runoff collected during the spring snowmelt
periods. This trend was also consistent over the two sampling periods as no significant direct or
interactive influence of time was observed. There were no significant differences in runoff
%MeHg between any of the six crossed treatments combinations (p = 0.87) (Figure 2-2f). This
suggests that a similar process is responsible for the transport of both THg and MeHg to runoff
during spring snowmelt.
One contributing factor controlling the mobilization and accumulation of THg and MeHg into
peat pore waters as well as transport in snowmelt runoff may be the degradation of peat through
decomposition. Pore water THg and MeHg concentrations were both positively and significantly,
but weakly correlated with DOC concentrations (r = 0.32 for THg, r = 0.30 for MeHg; both p <
0.0001) (Figure A-2-5), which would be expected to similarly increase as peat decomposes. Pore
water DOC concentrations significantly increased in the low WT treatment bins throughout the
58
sampling period (p < 0.001), while those in the high WT treatments did not significantly change
over time. Mean THg and MeHg concentrations as well as %MeHg in the shallow pore waters
(20 and 40 cm below the peat surface) were positively correlated with the amount of cellulose
decomposition (expressed as percent loss) in the top 40 cm of peat, near the mean water table
position in the low WT mesocosms (Figure 2-3). These relationships between pore water THg
and MeHg concentrations and the percentage of mass lost from the cellulose decomposition
assays are significant when all pore water sampling events are considered (Figure A-2-6). The
summary Figure 2-3 is included to demonstrate the role of Ericaceae removal on pore water THg
and MeHg concentrations beyond the influence of peat decomposition. The decomposition data
could only be used in a correlative manner, as it is to be published separately as part of the
PEATcosm experiment. Although the decomposition data was only from 2014, the percentage
of mass lost during this time should still be representative of potential peat decomposition as a
result of the water table and plant functional group treatment effects. The low WT – sedge only
treatment which yielded the highest concentrations of THg and MeHg in pore waters also
exhibited the greatest potential for peat decomposition through the cellulose decomposition
assays. The observed relationships are similar when examining the trends fitted using locally-
weighted scatterplot smoothing (LOWESS; Figure A-2-7). Elevated pore water THg and MeHg
in the sedge treatments under both water table prescriptions resulted in a curved relationship with
potential peat decomposition. This illustrates that the removal of Ericaceae shrubs may further
enhance THg and MeHg accumulation in pore waters beyond that predicted by peat
decomposition alone.
59
Figure 2-3 Relationships between shallow pore water (20 and 40 cm below the peat surface) Hg
and mean % mass loss of cellulose decomposition assays in the top 40 cm of the peat. (a) THg
concentrations, (b) MeHg concentrations, (c) %MeHg. Error bars represent standard deviation.
Decomposition assays were harvested in 2014 and represent potential peat decomposition among
the treatments. Pore water Hg data were averaged by treatment from all five sampling events.
The prolonged periods of water table drawdown sustained in the low WT treatments likely
caused the enhanced observed peat decomposition by increasing the zone of aerated peat,
resulting in greater aerobic respiration as compared to the relatively high water table position in
60
the high WT treatments (Whittington and Price, 2006; Ise et al., 2008; Waddington et al., 2014).
The degradation of soil organic matter leads to increased DOC mobilization and accumulation in
pore waters as reflected by higher concentrations in the mesocosm pore waters, with repeated
fluctuations in water table position resulting from precipitation inputs likely leaching carbon
from the peat into the pore waters following periods of drawdown (Fenner et al., 2007). This
may also account for the greater mobilization of THg and MeHg at the upper depths of the peat
monoliths, with higher pore water THg and MeHg concentrations in the 20 and 40 cm samples
within the zone of fluctuation of the low WT treatments. Significant, but relatively weak,
positive correlations with total phenolics in pore water were observed for MeHg (r = 0.34 p <
0.0001) and THg (r = 0.32; p < 0.0001) (Figure A-2-8). Romanowicz et al. (2015) found a
positive correlation between pore water total phenolics and peroxidase enzyme activity.
Oxidative enzymes including peroxidase and phenol oxidase contribute to the degradation of
lignin and polymeric complexes (Romanowicz et al. 2015). As pore water concentrations of
THg and MeHg increased with total phenolics, which were positively correlated with peroxidase
enzyme activity in the PEATcosm pore waters (Romanowicz et al. 2015), organic matter
decomposition was likely the source of increased pore water Hg mobility (Figure A-2-8).
Mercury forms complexes with organic matter (Driscoll et al., 1995) predominantly with thiol
moieties (Graham et al., 2012). Therefore, enhanced peat decomposition and degradation are
likely responsible for much of the observed increase of both THg and MeHg, as well as DOC
accumulation in the mobile pore water phase. Given that solid phase peat THg and MeHg
concentrations among bins were not significantly different prior to initiation of the experiment,
and THg concentrations in particular are not expected to change significantly over time, the
observed pore water THg and MeHg trends are a clear indication that Hg mobility has increased.
In addition to the peat decomposition effect on Hg and MeHg concentrations in pore water, the
water table manipulations may also act to enhance MeHg production in the peat. Prolonged
periods of water table drawdown followed by rewetting of the peat through precipitation input
may act to reset the oxidation-reduction conditions in the zone of aeration (Coleman Wasik et al.,
2015). Mercury methylation is a redox-sensitive process, facilitated by a diverse set of anaerobic
bacteria and Archaea (Gilmour et al., 2013). The periodic oxidation of peat through water table
variability may stimulate MeHg production via regeneration of the oxidized forms of terminal
electron acceptors, such as the oxidation of sulfide to regenerate sulfate (Coleman Wasik et al.,
61
2015). The highest sulfate concentrations in pore waters were indeed observed in the low WT
treatment mesocosms (Figure A-2-9).
In addition to the clear influence of the hydrological treatments, the dominant vascular plant
functional group also significantly affects the accumulation of Hg in pore waters. The
correlations between Hg concentrations and decomposition demonstrate the likely influence of
vascular plant functional groups affecting Hg mobility by mechanisms other than decomposition.
This is exemplified by the increased explanation of variability in the relationships when the
sedge only treatments for both water table prescriptions are removed (r = 0.96 for THg, r = 0.83
for MeHg, r = 0.49 for %MeHg). The removal of Ericaceae and the establishment of sedges may
lead to greater Hg methylation and the amount of Hg mobilized in pore waters, as sedge roots
exude labile carbon compounds and oxygen in the rhizosphere (Crow and Wieder 2005;
Bridgham et al. 2013). Under both water table treatments, the sedge-dominated peat monoliths
had the highest concentrations of THg and MeHg in pore water (Figure 2-2a, b), higher than may
be explained by potential peat decomposition and leaching into pore waters (Figure 2-3). Sedges
increase peat aeration and subsequent oxidation via their aerenchyma tissues (Bridgham et al.,
2008; Waddington et al., 2014) and may, similarly to the water table fluctuation, prime Hg
methylation by stimulating the regeneration of the terminal electron acceptors required for
methylating microbial communities (Coleman Wasik et al. 2015). The leakage of oxygen from
the roots of salt marsh sedges has been observed to stimulate aerobic respiration by acting as a
terminal electron acceptor as well as facilitating anaerobic respiration by regenerating alternative
electron acceptors including nitrate, ferric iron and sulfate (Mueller et al., 2016). Pore water
sulfate concentrations were slightly higher for the sedge-dominated treatments subjected to both
water table prescriptions (Figure A-2-9). Thus, in addition to supplying the necessary terminal
electron acceptors for soil organic matter decomposition (Mueller et al. 2016), sedges may prime
MeHg production through the provision of oxygen and labile organic compounds via root
exudation to the rhizosphere, which are required for methylation to occur (Mitchell et al. 2008a).
Conversely, the presence of Ericaceae shrub roots in the Ericaceae only and unmanipulated
control treatment bins (both of which had lower Hg concentrations than the sedge treatment) may
compete with the heterotrophs for available oxygen (Romanowicz et al., 2015), thereby reducing
sulfate regeneration (slightly, although not significantly; Figure A-2-9) and suppressing MeHg
production. Ericaceous shrubs form a symbiotic relationship with mycorrhizal fungi, which may
62
mediate changes in soil microbial communities, compete for available oxygen thereby limiting
terminal electron acceptor regeneration and decrease peat decomposition through the suppression
of heterotrophs. This has been demonstrated in the PEATcosm experimental monoliths wherein a
negative correlation was observed between Ericaceae biomass and peroxidase as well as β-
glucosidase enzyme activity (Romanowicz et al., 2015). Inhibition of MeHg production in the
upper rooting zone depths of the Ericaceae only as well as the unmanipulated control mesocosms
as compared to the sedge only monoliths may contribute to the peat pore water Hg concentration
differences observed.
The trends in THg and MeHg concentrations in snowmelt runoff are similar to those observed in
the pore waters (Figure 2-2). Snowmelt runoff THg and MeHg concentrations both exhibit a
significant (p < 0.001), positive correlation with DOC concentrations (r = 0.68 for THg, r = 0.66
for MeHg; Figure 2-4). Increased Hg export in association with organic carbon has similarly
been observed in other systems (St. Louis et al., 1994; Driscoll et al., 1995). The statistically
similar %MeHg across the treatments along with the positive correlation with DOC
concentrations provides further support for peat degradation being the principal driving factor
controlling not only the accumulation of Hg in pore waters, but also its transport within and
potential export from peatlands. The concentrations of THg and MeHg observed in snowmelt are
likely the result of peat decomposition liberating Hg in association with organic matter, which
has accumulated in pore waters throughout the growing season and flushed from the peat during
the spring snowmelt period.
63
Figure 2-4 Hg-DOC relationships in 2014 and 2015 snowmelt runoff for all 24 mesocosms (n =
48 total). (a) THg in relation to DOC concentrations and (b) MeHg in relation to DOC
concentrations. Data are plotted on log-transformed axes.
2.5 Conclusions
Climate-induced changes in the hydrological regime and vegetation communities of peatland
systems have the potential to significantly enhance both THg and MeHg accumulation in peat
pore water and transport downstream during spring snowmelt. Although peatlands are strong
sinks of inorganic Hg in the landscape, peatlands are also well-documented sources of MeHg (St.
Louis et al., 1994). By potentially increasing the amount of MeHg exported to aquatic
ecosystems for uptake into the food chain, as well as the amount of inorganic Hg available for in-
lake methylation, wildlife and human populations which rely upon fish consumption for
subsistence may be exposed to increased Hg levels (Scheuhammer et al., 2007; Mergler et al.,
64
2007). Given that peatlands cover approximately 3% of the global land area and are located
predominantly in the subarctic and boreal regions which are vulnerable to the effects of climatic
change (Bridgham et al., 2008), significant stores of Hg sequestered with vast stocks of carbon
(Grigal, 2003) may be liberated from either impacts on water table regime, plant community
changes, or a combination of both.
The small scale of the peat monoliths in this study facilitated a direct assessment of the influence
of water table position and vascular plant functional groups on Hg mobility by removing
potential confounding factors such as heterogeneities in microtopography and hydrological flow
paths and sources. At larger scales, spatial heterogeneities in Hg methylation and accumulation
have been observed, including MeHg hot spots at the outer edges of peatlands, which tend to
increase MeHg export from peatland systems (Mitchell et al., 2008b). The results of this study
are most applicable to the predominant interior peatland area and suggest that higher MeHg
concentrations observed at the outer edges of peatlands may be even further enhanced by climate
change-driven increases in Hg methylation and/or mobility from the peatland interior. Certainly,
results from this experimental work may be scaled to larger peatland ecosystems and may aid in
developing modeling approaches to forecast the response of peatlands under future climate
change scenarios. Overall, the results of this study suggest that deeper, more fluctuating water
tables and shifts toward sedge-dominated plant functional groups will measurably enhance Hg
and MeHg concentrations in peatland pore waters and likely lead to greater Hg and MeHg export
to downstream ecosystems during spring snowmelt.
2.6 Acknowledgements
We would like to acknowledge the laboratory assistance of P. Huang, K. Ng, R. Co and B.
Perron as well as the field assistance of K. Ng, S. Rao, K. Griffith, and E. Grupido. We thank T.
Veverica for pore water carbon chemistry data. Funding was provided through a Natural
Sciences and Engineering Research Council of Canada (NSERC) Alexander Graham Bell
Canada Graduate Scholarship (CGS-Doctoral) to K.M.H and a NSERC Discovery Grant to
C.P.J.M. The PEATcosm experiment was funded by the USDA Forest Service Northern
Research Station Climate Change Program and the National Science Foundation (DEB-
1146149).
65
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72
Chapter 3 Gaseous Mercury Fluxes in Peatlands and the Potential Influence
of Climate Change
3
3.1 Abstract
Climate change has the potential to significantly impact the stability of large stocks of mercury
(Hg) stored in peatland systems due to increasing temperatures, altered water table regimes and
subsequent shifts in vascular plant communities. However, the Hg exchange dynamics between
the atmosphere and peatlands are not well understood. At the PEATcosm Mesocosm Facility in
Houghton, Michigan, total gaseous Hg (TGM) fluxes were monitored in a subset of 1-m3 peat
monoliths with altered water table positions (high and low) and vascular plant functional groups
(sedge only, Ericaceae only or unmanipulated control) above the Sphagnum moss layer. At the
SPRUCE bog in north-central Minnesota, TGM fluxes were measured from plots subjected to
deep peat soil warming (up to +9 °C above ambient at a depth of 2 m). At PEATcosm, the
strongest depositional trend was observed with the Low WT – sedge only treatment mesocosms
with a mean TGM flux of −73.7 ± 6.3 ng m-2
d-1
, likely due to shuttling of Hg to the peat at depth
by aerenchymous tissues. The highest total leaf surface and tissue Hg concentrations were
observed with the Ericaceae shrubs. A negative correlation between TGM flux and Ericaceae
total leaf surface area suggests an influence of shrubs in controlling Hg exchange through
stomatal uptake, surface sorption and potentially, peat shading. Surface peat total Hg
concentrations are highest in treatments with greatest deposition suggesting deposition controls
Hg accumulation in surface peat. Fluxes in the SPRUCE plots ranged from −45.9 ± 93.8 ng m-2
d-1
prior to the implementation of the deep warming treatments to −1.41 ± 27.1 ng m-2
d-1
once
warming targets were achieved at depth and +10.2 ± 44.6 ng m-2
d-1
following prolonged deep
soil warming. While these intervals did not differ significantly, a significant positive increase in
the slope of the regression between flux and surface temperature was observed across the pre-
treatment and warming periods. Shifts in vascular vegetation cover and peat warming as a result
73
of climate change may significantly affect the dynamics of TGM fluxes between peatlands and
the atmosphere.
3.2 Introduction
In addition to natural sources such as volcanic emissions, anthropogenic sources including coal
burning, gold mining, and waste incineration have contributed to significantly increase mercury
(Hg) in global terrestrial, aquatic, and atmospheric pools (Selin, 2009). Atmospheric deposition
via either wet or dry processes represents the dominant input of Hg to the Earth's surface
(Fitzgerald et al., 1998; Driscoll et al., 2007). Wetlands are particularly important sinks of
atmospherically-deposited Hg in the landscape (Wallschläger et al., 2000; Zillioux et al., 1993).
Following deposition, Hg is subjected to numerous hydrological and biogeochemical influences,
which might either incorporate it into soil and vegetation stocks (Hintelmann et al., 2002), or
lead to its re-emission from soil, water and/or vegetation back to the atmosphere following
reduction to its elemental form, Hg(0) (Poissant et al., 2004a). Total gaseous mercury (including
all forms of gaseous Hg, the majority of which is typically Hg(0)) fluxes have been monitored in
freshwater wetlands (Poissant et al., 2004a, b; Zhang et al., 2006; Lindberg and Zhang, 2000;
Marsik et al., 2005). However, only a few more recent studies have focused on soil-air exchange
dynamics in boreal peatlands (Kyllönen et al., 2012; Fritsche et al., 2014; Osterwalder et al.,
2016) despite the large amounts of Hg stored in these systems (Grigal, 2003).
Peatlands are immense stores of organic matter in the landscape (Gorham, 1991) and may
represent a significant source of Hg to the atmosphere (Grigal, 2003). Climate change has the
potential to significantly impact the hydrology, biogeochemistry and ecology of peatland
ecosystems due to increasing temperatures and changes in precipitation patterns (Kunkel et al.,
2003; Groisman et al., 2005; Bridgham et al., 2008). Enhanced variability in precipitation, with
less frequent and more intense events anticipated across the northern continental United States
(Kunkel et al., 2003; Groisman et al., 2005), may have direct impacts on the hydrology of
peatlands with the potential for prolonged periods of water table recession and greater variability
throughout the summer months (Thomson et al., 2005; Whittington and Price, 2006). Subsequent
changes in the dominant vascular plant functional groups may occur as a result of prolonged dry
periods and warming temperatures in peatlands (Weltzin et al., 2003; Strack et al., 2006;
Breeuwer et al., 2009; Dieleman et al., 2015). Due to their ability to reach deeper water stores,
74
regulate water loss through stomata and oxygenate the rooting zone by shuttling oxygen via
aerenchymous tissue, sedges may have a competitive advantage over shrub vegetation and
Sphagnum mosses during periods of water stress (Dieleman et al., 2015). However, several
studies (i.e. Weltzin et al., 2000; Breeuwer et al., 2009) including the PEATcosm mesocosm
experiment (Potvin et al., 2015) have observed a shift towards increasing Ericaceae shrub
abundance and productivity under drier conditions. Therefore, plant communities in both bogs
and fens are likely to change with the increasing temperatures and enhanced water stress
resulting from climate change, although the direction of this change is not clear (Weltzin et al.,
2003). Changes in plant community composition in concert with alterations to the hydrology and
redox condition of the peat may have profound effects on peat decomposition and therefore the
carbon storage abilities of peatlands (Waddington et al., 2015). Given the strong relationship
between Hg and organic matter, the potential changes in water table recession and variability,
dominant vascular plant communities, peat mineralization and shifting carbon balances resulting
from hydrological, ecological and biogeochemical changes may affect the dynamic exchange of
total gaseous Hg (TGM) fluxes between the peat and the atmosphere.
Total gaseous Hg exchange with background soils is often well correlated with meteorological
variables (Gustin et al., 2006; Lin et al., 2010). Air and soil temperatures are important
thermodynamic controls on Hg reduction and emission from soils (Edwards et al., 2001), as is
solar radiation because of the importance of photoreduction processes in controlling Hg
emissions (Moore and Carpi, 2005). These factors act to influence Hg emissions by effectively
lowering the activation energy required for Hg reduction (Carpi and Lindberg, 1998). Changes in
soil moisture significantly impact Hg fluxes from mineral soils (Gustin and Stamenkovic, 2005;
Song and Van Heyst, 2005). Wetting of the soil surface by precipitation increases Hg emissions
from initially dry soils, with spikes in fluxes following the rain event and decreasing fluxes as
the soil dries (Gustin and Stamenkovic, 2005; Mazur et al., 2015). The mechanisms governing
this increase with wetting may involve desorption of Hg from the soil particles and subsequent
displacement of the Hg from the soil pores. However, emissions from saturated soils are often
suppressed, potentially due to the very slow diffusion of Hg(0) through water (Bahlmann et al.,
2004; Gustin and Stamenkovic, 2005; Song and Van Heyst, 2005). Enhanced evaporation
coincident with increasing atmospheric temperatures may also affect the dynamic exchange of
Hg by transporting Hg to the soil surface via capillary action (Briggs and Gustin, 2013).
75
Therefore, altered water table position and variability in precipitation patterns may have a direct
impact on seasonal Hg emission and deposition patterns in peatlands resulting from changes in
soil moisture and redox conditions. Assuming similar mechanisms to those observed in mineral
soils (Bahlmann et al., 2004; Gustin and Stamenkovic, 2005; Song and Van Heyst, 2005),
episodic re-wetting of the peat following prolonged water table drawdown may enhance TGM
fluxes from peatlands. Once saturated however, significant emission of Hg to the atmosphere
may be inhibited. Periods of wetting may provide the necessary redox conditions to promote
Hg(II) reduction (Moore and Castro, 2012) to Hg(0) and subsequent release from the peat. Taken
together, the direct and indirect influences of increasing temperatures and altered precipitation
regimes on peatland TGM fluxes clearly warrant further investigation.
The shifting composition of the vascular plant community may also affect the exchange of Hg
through stomata and the potential liberation of Hg from soil via the roots (Rea et al., 2002;
Leonard et al., 1998a, b). Sedges with deep roots and aerenchymous tissues (Breeuwer et al.,
2009) may affect the magnitude and direction of TGM fluxes depending upon root depth and Hg
concentrations within the soil profiles, as has been observed with other wetland vegetation
(Lindberg et al., 2005). Lindberg et al. (2005) determined that the source of TGM emissions
from aquatic macrophytes was the rhizosphere. However, translocation of Hg between the roots
and the foliage of forest canopies is minimal (Ericksen et al., 2003). Although several studies of
soil-air Hg dynamics have been conducted over bare soil (e.g. Carpi and Lindberg, 1997;
Poissant et al., 2004b; Miller et al., 2011), vegetation in both forests and wetland systems plays
an important role in controlling the exchange of Hg with the atmosphere via both stomatal and
non-stomatal pathways including leaf surface sorption (Lee et al., 2000; Rea et al., 2002;
Ericksen et al., 2003; Marsik et al., 2005; Stamenkovic and Gustin, 2009; Laacouri et al., 2013).
Evaporation of morning dew from the surface of vegetation has also been identified as a
potential source of Hg to the atmosphere (Marsik et al., 2005). Stomatal exchange was observed
to be an instrumental process governing net TGM emission from the canopy of a mixed
sawgrass-cattail wetland in the Everglades as Hg emission was coincident with stomatal uptake
of carbon dioxide (Marsik et al., 2005). The manner in which climate change-induced shifts in
vascular plant community composition affects the direction and magnitude of TGM fluxes
between peatlands and the atmosphere remains unclear and therefore requires further study.
76
The overall objectives of this study were to characterize diurnal TGM fluxes in peatlands and to
investigate impacts on TGM fluxes due to changes in vascular plant community composition,
water table position, and peat warming, using two field-scale peatland climate change simulation
experiments. Diurnal fluxes were measured in a subset of the PEATcosm (Peatland Experiment
at the Houghton Mesocosm Facility) peat monoliths, manipulated to simulate climate-induced
changes in vascular plant functional groups and water table position and variability. To assess
the potential effects of soil warming, TGM fluxes were also measured in the SPRUCE (Spruce
and Peatland Responses Under Climatic and Environmental Change) experimental peatland
prior to and throughout a large-scale manipulation of deep (2 m) peat temperatures. It was
hypothesized that 1) peatland plant functional groups would differentially affect TGM fluxes due
to differences in vascular structures 2) increased sedge cover would increase TGM deposition to
peatlands due to incidental shuttling of Hg via aerenchymous tissue and 3) increased soil
temperatures would enhance TGM fluxes from the peat to the atmosphere.
3.3 Methods
3.3.1 PEATcosm Site Description
The PEATcosm experiment was located at the United States Forest Service Houghton Mesocosm
Facility, at the Forestry Sciences Laboratory in Houghton, Michigan, USA (47.11469° N,
88.54787° W). The regional climate is humid continental with typical annual precipitation of
approximately 870 mm. Mean temperatures range from −13 °C in January to +24 °C in July (30
year means at Houghton County Airport; Potvin et al., 2015).
The Mesocosm Facility houses 24 intact 1 m3 (1 m × 1 m × 1 m) peat monoliths open to the air
on top and inserted into a climate-controlled tunnel which allows belowground access to each of
the bins and the simulation of natural peat temperature gradients with depth. The monoliths were
harvested from an ombrotrophic peatland located in Meadowlands, MN, USA in May 2010 and
transferred into individual Teflon-coated, stainless steel mesocosm bins, preventing metal
transfer to the peat. Vegetation manipulations were initiated in 2011, and full water table
manipulations were initiated in 2012. At the conclusion of the experiment in July 2015, the
monoliths were destructively harvested. The full description of the PEATcosm experiment can
be found in Potvin et al. (2015), including peat monolith harvest, facility details and
experimental design.
77
The PEATcosm study comprises a full-factorial experimental design, with two water table (WT)
prescriptions crossed with three different plant functional group treatments, simulating potential
climate change outcomes. Each treatment combination is replicated across four mesocosms, in a
randomized complete block design, for a total of 24 experimental units. The water table
treatments were based on long-term (approximately 50 years) data from the Marcell
Experimental Forest in north-central Minnesota (47.51907° N, 93.45966° W), located near the
peat monolith harvest site. The two WT treatments were modelled after years with: 1) typical
variability and average WT position (referred to as ‘High WT’) and 2) comparably high
variability and low WT position (referred to as ‘Low WT’). These target WT profiles were
maintained through a combination of artificial precipitation additions, rain-exclusion covers and
regulated outflow from approximately the acrotelm-catotelm boundary during the snowmelt
period from each of the bins (see Potvin et al., 2015 for details). The three plant functional group
treatments, simulating likely community composition alterations resulting from climate change
(Strack et al., 2006; Weltzin et al., 2000; Chapin et al., 1996), are as follows: 1) all Ericaceae
removed (referred to as ‘sedge only’ treatment), 2) all sedge removed (‘Ericaceae only’
treatment) and 3) both sedge and Ericaceae present (‘unmanipulated control’ treatment). The
dominant sedge species present in the bins was Carex oligosperma Michx., while the dominant
Ericaceae shrubs included Chamaedaphne calyculata (L.) Moench., Kalmia polifolia Wangenh.,
and Vaccinium oxycoccos L. The dominant moss species in the peat monoliths were Sphagnum
rubellum Wilson, S. magellanicum Brid., S. fuscum (Schimp.) Klinggr and Polytrichum strictum
Brid. Polytrichum commune Hedw., Eriophorum vaginatum L., Andromeda polifolia L. var.
glaucophylla (Link) DC., Rhododendron groenlandicum (Oeder) Kron and Judd, and Drosera
rotundifolia L. were also present in the mesocosms.
3.3.2 SPRUCE Site Description
The SPRUCE experimental peatland is an 8.1 ha ombrotrophic, spruce-Sphagnum bog (known
locally as “S1”), located within the Marcell Experimental Forest (MEF) in north-central
Minnesota (47.51907° N, 93.45966° W). Typical annual precipitation for the MEF is
approximately 780 mm. Mean temperatures at the MEF range from −15 °C in January to +19 °C
in July (Sebestyen et al., 2011). The overstory is dominated by two tree species: Picea mariana
(Mill.) B.S.P and Larix laricina (Du Roi) K. Koch. The understory consists primarily of
Ericaceae shrubs including Chamaedaphne calyculata, Kalmia polifolia, Vaccinium
78
angustifolium Aiton, Vaccinium oxycoccos, and Rhododendron groenlandicum as well as sedges
Eriophorum spp. and the lily Maianthemum trifolium (L.) Sloboda. The dominant bryophyte
present on hummocks is Sphagnum magellanicum, while hollows are mainly colonized by S.
angustifolium (C.E.O. Jensen ex Russow).
The SPRUCE study (http://mnspruce.ornl.gov/) is a climate change field experiment that
involves the long-term (10-year) manipulation of both soil and air temperature, and atmospheric
carbon dioxide (CO2) concentrations within open-top, approximately 12 m diameter, enclosures
along three transects within the S1 bog . Prior to the full initiation of the SPRUCE experiment in
2016, an approximate one-year, deep (2 m below the peat surface) soil warming experiment was
conducted from June 2014 through the end of July 2015. This study describes TGM fluxes
before (May 2014), during (August 2014), and following prolonged deep peat warming (June
2015). The heating infrastructure utilized to warm the peat at depth within the experimental
enclosures is similar to that described in Hanson et al. (2011).
3.3.3 Experimental design and mercury flux measurements
Subsets of the total number of experimental treatments in each experiment were monitored for
TGM flux using dynamic flux chambers (DFCs). Despite their limitations (Wallschläger et al.,
1999; Gillis and Miller, 2000), DFCs were the only appropriate method given the size and scale
of the experimental mesocosms and plots. These measurements were taken over individual 24-h
periods. This was necessary as similar meteorological conditions (i.e. no rain, with similar cloud
cover and incoming solar radiation) could only be achieved within relatively short (4–6 days)
timeframes. Similar ambient conditions were necessary to discern potential treatment effects. At
PEATcosm, TGM fluxes were monitored on a subset of 8 of the mesocosm bins in July 2014 at
the peak of the growing season. Simultaneous measurements were taken on two mesocosms in
each 24-h period. In all, duplicate mesocosms of the following experimental treatment pairings
were monitored for TGM fluxes: High WT – control vegetation, Low WT – control vegetation,
Low WT – Ericaceae only, and Low WT – sedge only. These treatments span the range of
vascular plant functional group treatments with the Low WT prescription representing the
greatest anticipated hydrological changes in peatlands due to climate change. The High WT –
control vegetation treatment acts as the unmanipulated, control scenario. The duplicate
mesocosms for each treatment were not measured during the same 24-h period. Mean water table
79
positions during the flux measurement period were 37 ± 5 cm below the peat surface (mean ±
standard deviation) for the Low WT treatments and 15 ± 0.1 cm below the peat surface for the
High WT treatment.
At the SPRUCE site, TGM fluxes were measured in six experimental enclosures, focusing on
duplicates of deep soil warming treatments with target temperature differentials of +0, +4.5 and
+9 °C above ambient soil temperatures at 2 m depth. These target differentials were achieved at
2 m depth within 60 days of the June 2014 initiation, but were not similarly observed nearer the
surface. Due to the considerable distance between SPRUCE plots and equipment availability,
two replicate locations within each experimental plot were measured simultaneously. Flux
measurements at SPRUCE were performed in May 2014 prior to the initiation of the deep soil
warming treatments, August 2014 once soil warming target temperatures were achieved at depth
and June 2015 following prolonged deep soil warming. Due to an equipment malfunction, one
+0 and one +4.5 °C plot (Plots 4 and 6, respectively) could not be monitored in June 2015.
Ambient TGM fluxes were measured using two transparent Teflon DFCs (0.036 m2 footprint, 6.5
cm height and 2.0 L volume with 1 cm diameter inlet holes every 2.5 cm around the perimeter
and a 5 cm wide bottom flange as per Eckley et al., 2010) placed on the peat surface and
connected to a Tekran 2537A Gaseous Mercury Analyzer. The DFCs were placed on hummock
microtopographic forms within each plot and mesocosm. Standing vegetation was present
beneath the footprint of each DFC. Weighted rings (clear plastic tubing filled with small ball
bearings) were placed on the DFC bottom flange to ensure a tight seal to the peat surface. The
sampling flow rate of the analyzer was 1.5 L min-1
. A soda lime trap was connected in the
sampling inlet line just prior to introduction into the analyzer to prevent passivation of the gold
traps by acid aerosols (Ericksen et al., 2006; Miller et al., 2011). A four-port sampling
manifold/switching unit (Tekran 1115 Synchronized Multi-Port Sampling System) connected to
the analyzer facilitated the alternation of air accumulation among four Teflon lines, an inlet and
outlet for each of the two DFCs, each fitted with a 0.2 μm disposable particulate syringe filter
(PTFE-membrane in a polypropylene housing, Cole-Parmer). The manifold allowed nearly
simultaneous duplication of measurements (Carpi and Lindberg, 1997). One complete flux
measurement for one DFC was achieved every 20 min. Fluxes were calculated using the
equation:
80
𝐹 = 𝑄 ∗ (𝐶𝑜 − 𝐶𝑖)/𝐴 (3.1)
where F is the total flux (ng m-2
h-1
), Q is the chamber flushing flow rate (m3 h
-1), Co is the Hg
concentration measured at the chamber outlet (ng m-3
), Ci is the Hg concentration measured
adjacent to one of the chamber inlet holes (ng m-3
), and A is the footprint area of the chamber
(m2). Quality control of the calculated fluxes was based on the determination of the inter-
cartridge concentration difference for each of the TGM fluxes measured using the Tekran 2537A
according to the method described by Eckley et al. (2011). Calculation of daily TGM fluxes was
achieved by integrating the area under the 24-h curve of all valid measurements for each DFC.
Positive flux values represent emission from the soil surface to the atmosphere, while negative
values signify TGM deposition to the soil.
To ensure consistent conditions in each DFC and its inlet and outlet lines while not being
measured, a secondary bypass vacuum pump was attached to the four-port manifold to draw
ambient air through the lines at the 1.5 L min-1
sampling rate (Mazur et al., 2014). This flushing
flow rate was applied to prevent artificial concentration gradients over the soil and vegetation
surfaces within the DFCs and as a consequence induce higher fluxes due to exceedingly high
flow rates, erring on the side of a lower flow rate as suggested by Eckley et al. (2010). A bubble
flow calibrator (mini-Buck Calibrator M-5) was connected to each of the inlet and outlet lines
prior to each sampling campaign to ensure the rate of air flow facilitated by the secondary pump.
The analyzer was calibrated prior to each 24-h cycle measurement using the internal mercury
permeation source. Each of the two DFCs were soaked in a 10% hydrochloric acid bath for 24 h
and rinsed thoroughly with deionized water prior to sampling. To ensure the cleanliness of the
DFCs prior to and throughout each monitoring campaign, blank measurements were performed
in the field for approximately 24-h by placing each DFC on clean polycarbonate sheeting on the
ground surface adjacent to the plots or mesocosms (0.04 ± 0.16 ng m-2
h-1
, mean ± standard
deviation, n = 123 20-min fluxes) and were not subtracted from flux measurements as is
consistent with other published work (e.g. Eckley et al., 2010; Mazur et al., 2014).
Condensation occurred on the inside of the DFCs during the experiments, which could not be
mitigated through an increase in DFC flushing flow rate. To assess if Hg(0) accumulation
occurred in the condensation, samples were collected by rinsing each DFC following the
measurement cycle using a known volume of deionized water (18 MΩ cm-1
) as has been
81
previously done by Briggs et al. (2012). The collected rinse water was acidified with 0.5% trace-
metal grade HCl and stored in darkness at 4 °C until total Hg analysis. Mean THg concentrations
of these unfiltered DFC condensation rinses were low, representing approximately 1.2 ± 1.2%
(mean ± standard deviation) of the mean daily TGM fluxes.
Incoming solar radiation as well as air and surface peat temperature were recorded at 5-min
intervals on a Campbell Scientific CR1000 data logger during the Hg flux measurements. A
pyranometer (Kipp and Zonen SP-Lite 2 Silicon Pyranometer) was placed mid-way between the
two DFCs being monitored for each 24-h cycle. Surface peat temperature adjacent to each of the
monitored DFCs (to prevent disturbance under the DFC) was measured using temperature probes
(109 Temperature Probes) inserted approximately 2 cm below the peat surface. Ambient air
temperature outside of the chambers was also measured using a shielded thermocouple
throughout the flux measurements.
3.3.4 PEATcosm vegetation and surface peat Hg
To determine the relative influences of TGM uptake through the stomata of vascular vegetation
as compared to sorption to leaf surfaces, foliar rinsing was performed similar to the methods
conducted by Rea et al. (2000), Ericksen et al. (2003) and Laacouri et al. (2013). Following the
completion of the PEATcosm experiment in July 2015, vascular vegetation was collected from
each of the mesocosm bins that were monitored for TGM fluxes. The vegetation from each bin
was separated according to species: Carex oligosperma (n = 5), Eriophorum vaginatum (n = 3),
Kalmia polifolia (n = 6), Chamaedaphne calyculata (n = 6). The leaves of each species collected
from each bin (ranging from approx. 170–500 leaves) were placed in a known volume of
deionized water and gently shaken for five minutes using a horizontal shaker. The rinse solution
was decanted into a new polyethylene terephthalate glycol (PETG) bottle and a second DI water
rinse was sequentially performed on each leaf sample. These rinses were followed by two
sequential rinses using a dilute (pH = 4) trace-metal grade nitric acid solution to ensure all
surface-deposited Hg was removed. All samples were stored in darkness at 4 °C until THg
analysis the following day. Surface Hg loads are expressed both by dry weight of leaves washed
(ng g-1
dry wt.) and by one-sided surface area (ng m-2
) similar to Rea et al. (2000) and Laacouri
et al. (2013). An average leaf surface area for each species was determined by measuring the
length and width of a subset of leaves and assuming the leaves to be oval in shape (one-sided
82
surface area = (½ length)*(½ width)*π). This average was multiplied by the number of leaves
counted for each species to estimate a total area of leaves rinsed. To estimate the amount of Hg
sorbed to the vascular vegetation under each DFC, the number of leaves and an average leaf
length and width for each species present under the footprint of each DFC was counted in the
field following the TGM flux measurement period. The coverage of vascular vegetation was not
consistent across the bins and therefore by counting the leaves under each DFC and applying a
mean leaf area, an estimate of total leaf area (in cm2) could be made. After rinsing, the leaf
samples were frozen and subsequently lyophilized. The dried samples were ground prior to THg
analysis to determine the amount of Hg found within the leaf tissue following surface rinsing as
compared to that resulting from the leaf rinses and therefore sorbed to the leaf surface. To assess
the possible relationship between TGM fluxes and peat Hg concentrations, surface peat (0–10
cm) samples were collected from each of the eight monitored mesocosms using a clean stainless
steel corer. Peat samples were immediately frozen and lyophilized prior to analysis for THg
concentrations.
3.3.5 Analytical Methods
Total Hg concentrations of the DFC condensation rinses, foliar rinses, lyophilized vegetation and
peat were determined using a Tekran Model 2600 automated Total Mercury Analyzer, using cold
vapor atomic fluorescence spectroscopy (CVAFS) according to US EPA Method 1631. Freeze-
dried vegetation and peat samples were digested in hot nitric acid. All Hg in the condensation
and foliar rinses as well as the diluted peat and vegetation digestates was oxidized by reaction
with bromine monochloride (BrCl) overnight prior to analysis. Recovery of a THg spike was 96
± 5% (mean ± standard deviation, n = 6), replication of duplicates was 2.0 ± 1.6% (n = 6) and the
detection limit for water samples, calculated as three standard deviations of matrix blanks, was
0.16 ng L-1
(n = 23). Recovery of standard reference material (MESS-3) following digestion was
97 ± 2% (n = 2). The detection limit for solid material was 0.01 ng g-1
(n = 6).
3.3.6 Statistical Analyses
All statistical analyses were performed using R statistical software (R Development Core Team,
2014) with α = 0.05. For the PEATcosm measurements, only a qualitative assessment of TGM
flux trends among the four crossed water table and plant functional group treatments were
considered since only two replicate mesocosms per treatment were measured and statistical
83
power of duplicates is low. Correlative relationships using Pearson correlation were investigated
between hourly TGM fluxes and incoming solar radiation as well as surface peat soil
temperatures among the four treatments. Only non-zero solar radiation values (i.e. daylight) were
included in these correlative relationships. Further examination of the relationships between
TGM fluxes and surface peat temperatures was performed using piecewise regressions. The
correlative relationships between daily TGM fluxes and surface peat THg concentrations, total
leaf area beneath each DFC, and Hg sorbed to the vegetation surfaces in each of the eight
monitored mesocosms were also examined. The relationship between the number of sedge stems
present beneath each DFC and the daily TGM fluxes was investigated collectively for the
PEATcosm and SPRUCE sites. Significant differences in THg concentrations on the surface of
the different vascular plant species, as well as that within the vascular plant tissues were assessed
using a one-way analysis of variance (ANOVA), followed by post-hoc testing with Tukey
adjustments. Differences in daily TGM fluxes among SPRUCE treatment plots throughout the
course of the deep soil warming experiment were assessed using regression analysis with mean
daily surface peat temperatures for each of the three sampling campaigns. Significant changes in
the slope of this regression over the three sampling events of the deep warming experiment were
assessed using an analysis of covariance (ANCOVA) with the surface peat temperature as the
covariate.
3.4 Results
3.4.1 PEATcosm – water table and plant functional groups influence
The crossed water table position and vascular plant functional group treatments influenced the
dynamic exchange of TGM between the peat monoliths and the atmosphere (Figure 3-1).
Consistent evasion of TGM from the peat was observed in both of the High WT control
mesocosms over the course of the 24-h periods with a mean (±standard deviation) daily flux of
+35.2 ± 10 ng m-2
d-1
. In contrast, strong deposition of TGM to the peat was observed in the Low
WT sedge only bins, with a mean daily TGM flux of −73.7 ± 6.3 ng m-2
d-1
. Total gaseous Hg
was deposited to one of the Low WT Ericaceae only bins (−69.1 ng m-2
d-1
), while slight
emission of TGM was observed from the other bin (+4.2 ng m-2
d-1
). Two contrasting flux
directions were also observed for the Low WT control mesocosms with one strongly emitting
TGM to the atmosphere (+144.5 ng m-2
d-1
) while TGM deposition occurred in the other
84
mesocosm (−35.2 ng m-2
d-1
; Figure 3-1). There was a significant, positive, but weak correlation
(r = 0.32, p < 0.001) between the mean inlet TGM concentrations and the corresponding 20-min
TGM fluxes measured during daylight hours across the treatments (Figure B-3-1).
Figure 3-1 PEATcosm daily TGM fluxes (in ng m-2
d-1
) across the four crossed water table
(WT) and plant functional group treatments (8 mesocosms in total monitored). * signifies that
sedge vegetation was present beneath the footprint of the DFC placed on the peat surface.
Sedges were present beneath the DFC placed on the peat surface of the Low WT – control
monolith exhibiting deposition throughout the 24-h monitoring cycle. Although sedges were
present in the other monitored Low WT – control mesocosm, which emitted TGM to the
atmosphere over a 24-h period, no sedge stems were located beneath the footprint of the DFC
(Figure 3-1). No sedges were present beneath the footprints of the DFCs for either of the High
WT – control mesocosms. A significant negative correlation (r = - 0.91, p < 0.001) was observed
between the TGM fluxes measured at both the PEATcosm and SPRUCE sites and the number of
sedge stems present beneath the DFC (Figure 3-2). Stronger depositional fluxes were observed as
the number of sedges under the DFC increased across the plots at both peatland sites.
85
Figure 3-2 Relationship between daily TGM fluxes (ng m-2
d-1
) and the number of sedge stems
present beneath the DFC at both the PEATcosm (closed red circles) and SPRUCE (open black
circles) sites.
The trend and magnitude of incoming solar radiation were similar across the four measurement
days (Figure 3-3), with no significant differences in radiation between the treatments (p = 0.92).
Therefore, the trend in TGM fluxes among the treatments is not the result of differences in
incoming solar radiation. Diel variation was observed across the treatments with early-afternoon
peaks in TGM fluxes coincident with peaks in solar radiation and surface soil temperatures
(Figure 3-3). Daylight solar radiation is strongly correlated with TGM flux, although to varying
degrees depending upon the treatment (Figure 3-4b). Similarly, positive, second-order
polynomial correlations were also observed between the hourly TGM fluxes and surface peat
temperatures (Figure 3-4a). These relationships may be suggestive of temperature thresholds in
each treatment, beyond which TGM fluxes increased (Figure B-3-2).
86
Figure 3-3 PEATcosm 24-h TGM flux measurements (in ng m-2
h-1
) for the four crossed water
table and plant functional group treatments (two replicates each) in relation to surface soil
temperature (in °C) and solar radiation (kW m-2
).
87
Figure 3-4 Relationship between PEATcosm hourly TGM fluxes (ng m-2
h-1
) and a) soil
temperature (in °C) and b) solar radiation (kW m-2
) among the four crossed water table and
vascular plant functional group treatments.
When considering the foliar rinses of the Ericaceae and sedge vegetation, the highest surface
THg concentrations were washed from the Chamaedaphne calyculata (leatherleaf) leaves. This
shrub had significantly higher THg sorbed to the leaf surfaces on a dry mass basis than Kalmia
polifolia (bog laurel) as well as both sedges, Carex oligosperma and Eriophorum vaginatum
(Figure 3-5a). Carex oligosperma had the lowest surface THg concentrations; significantly lower
than either of the Ericaceae shrubs. These patterns were also consistent, although not significant
88
on a per area basis, wherein C. calyculata and K. polifolia had slightly higher THg
concentrations than both of the sedges (Figure 3-5b). When considering the amount of THg in
the leaf tissue, both C. calyculata and K. polifolia had significantly higher tissue THg
concentrations than either of the sedges (Figure 3-5c).
Figure 3-5 Total Hg concentrations of PEATcosm vegetation a) leaf rinses expressed per unit
leaf tissue mass (ng g-1
dry wt.), b) leaf rinses per unit leaf area (ng m-2
) and c) leaf tissue mass
(ng g-1
dry wt.). Letters denote statistically similar values. No significant differences between
THg concentrations expressed on a per leaf surface area basis.
89
Among leaf area, the amount of Hg sorbed to the vegetation under each DFC, and THg in
surface peat, only peat THg concentrations were significantly correlated with mean daily TGM
fluxes (Figure 3-6). A negative, but not significant correlation was observed between the
measured TGM daily fluxes and the estimated Ericaceae total leaf area present beneath each
DFC (Figure 3-6a; r = - 0.66, p = 0.15). A marginally significant negative correlation was
observed between the estimated amount of leaf-sorbed Hg on Ericaceae shrubs under each DFC
and measured TGM fluxes (Figure 3-6b; r = - 0.74, p = 0.09). A significant negative correlation
(r = - 0.90, p < 0.01) was observed between the PEATcosm TGM fluxes and the surface (0–10
cm) peat THg concentrations (Figure 3-6c).
Figure 3-6 PEATcosm daily TGM fluxes (in ng m-2
d-1
) in relation to a) estimates of total leaf
area (in cm2) of all Ericaceae vegetation present beneath the DFC footprint, b) estimates of THg
sorbed to the surfaces of Ericaceae leaves under each DFC and c) surface (0–10 cm) peat THg
concentrations.
3.4.2 SPRUCE bog TGM fluxes
In May 2014 before the onset of experimental warming, 24-h TGM fluxes measured across all
plots had an average daily flux of −45.9 ± 93.8 ng m-2
d-1
(mean ± standard deviation). These
fluxes ranged from deposition at an average rate of −158.8 ng m-2
d-1
to evasion of +199.1 ng m-2
d-1
(Figure 3-7). In August 2014 once the deep soil warming treatments had been achieved at
depth, the minimum TGM flux recorded was −38.5 ng m-2
d-1
while the maximum TGM flux
90
observed was +59.2 ng m-2
d-1
with a mean TGM daily flux rate of −1.41 ± 27.1 ng m-2
d-1
.
Following a prolonged period under deep soil warming conditions, the mean daily TGM flux
measured in June 2015 was +10.2 ± 44.6 ng m-2
d-1
, ranging from deposition at a rate of −42.6 ng
m-2
d-1
to evasion at a rate of +107.4 ng m-2
d-1
(Figure 3-7). Significant, positive, but relatively
weak correlations were observed between the mean TGM inlet concentrations and the
corresponding TGM fluxes measured during daylight hours during the May 2014 (r = 0.47, p <
0.0001) and June 2015 (r = 0.58, p < 0.0001) samplings (Figure B-3-3).
Figure 3-7 SPRUCE daily Hg fluxes (in ng m-2
d-1
) for both of the two replicate DFCs across the
range of deep peat heating temperature treatment plots (+0, +4.5 and +9 °C) in May 2014,
August 2014 and June 2015. Flux data for Plots 4 and 6 in June 2015 not available.
3.4.2.1 Influence of deep peat warming
In May 2014, prior to the implementation of the deep warming treatments, no significant
correlation was observed between the mean TGM fluxes and the mean daily surface peat
91
temperatures (p = 0.23). Similarly, no relationships were observed in August 2014 (p = 0.55)
once the warming targets in the deep peat layers had been achieved nor in June 2015 (p = 0.69)
following prolonged deep warming. No significant correlations were observed between the TGM
fluxes and the peat temperatures at a depth of 2 m for the August 2014 (p = 0.78) or June 2015 (p
= 0.46) samplings, both of which achieved the target temperature differentials.
When comparing the slopes for the relationships between surface peat temperature and TGM
flux, the slope of the May 2014 data collected prior to deep soil warming is significantly
different from those collected in August 2014 and June 2015 under warming conditions (Figure
3-8; p = 0.04). Deep warming may therefore influence the amount of Hg emitted from the peat as
there is a transition in the slope of the relationship from strongly negative in May 2014 prior to
the initiation of the deep warming treatments, to moderately negative once warming target
differentials were achieved at depth in August 2014, and to slightly positive in June 2015
following prolonged deep warming (Figure 3-8). This trend in TGM flux may not be strongly
reflected in relation to surface peat temperature as target temperature differentials were not
achieved nearer the surface of the peat profile as air warming was not initiated for this one-year
experiment. However, this progressive change in slope throughout the deep warming experiment
under warming conditions suggests some influence of deep soil warming on TGM flux from the
peatland.
92
Figure 3-8 Relationship between SPRUCE daily TGM fluxes and mean daily surface peat
temperatures for the May 2014 (pre-warming-solid circles), August 2014 (deep warming
achieved at depth-open circles) and June 2015 (prolonged deep warming-stars) measurement
periods. Deep warming target temperature differential treatments (+0, +4.5, +9 °C) are colour-
coded in blue, orange and red, respectively.
3.5 Discussion
3.5.1 PEATcosm peat monoliths – influence of vascular plant community on TGM fluxes
The relatively strong depositional trends observed when sedges were present under DFCs, both
as the sole surface vascular cover and with other vegetation present (Figures 3-1 and 3-2),
suggest a significant role for vegetation type in controlling TGM fluxes. Foliage may interact
with atmospheric Hg in multiple ways including exchange of Hg through stomata (Ericksen and
Gustin, 2004; Hanson et al., 1995; Lindberg et al., 1998), wet and dry deposition of Hg onto
foliar surfaces (Stamenkovic and Gustin, 2009) as well as uptake of Hg into vascular vegetation
from soil water transported to the foliage via transpiration (Graydon et al., 2006; Rea et al.,
2002). Rooting structures of vascular vegetation may also be important conduits for the dynamic
exchange of Hg between the atmosphere and soil (Leonard et al., 1998a, b). Since the sedge
species (C. oligosperma and E. vaginatum) present in the mesocosm bins had generally both the
lowest surface-sorbed and leaf tissue THg concentrations, it is unlikely that sorption to sedges or
uptake by sedges is the mechanism by which they enhance TGM deposition. Rather, the impact
93
is more likely a function of increased shuttling of TGM from the atmosphere, coincident with
oxygen transport to the peat via aerenchymous tissues (Breeuwer et al., 2009) under low water
table conditions.
Relative to the role of sedges, Ericaceae shrubs appear to influence gaseous Hg exchange
dynamics through sorption onto foliar surfaces and/or uptake through stomata and accumulation
in leaf tissue. Greater sorption of Hg to the foliar surfaces of Ericaceae shrubs as well as stomatal
uptake into the shrub leaf tissue is associated, albeit weakly, with reduced emission and some
depositional fluxes (Figure 3-6a and b). Mercury sorbed to the surface of Ericaceae leaves,
particularly that which is exposed to incoming solar radiation on the top of the leaves, may likely
be photoreactive (Laacouri et al., 2013). Once deposited, this surface-sorbed Hg may undergo
considerable re-emission to the atmosphere and contribute to the predominantly emissive fluxes
from the Ericaceae-dominated treatments. Increased leaf area will also result in more stomata
present and therefore the potential for greater Hg uptake, likely in the form of gaseous Hg(0),
while stomata are open during daylight hours for the process of photosynthesis (Lindberg et al.,
2002; Marsik et al., 2005; Laacouri et al., 2013). Increased uptake of Hg through stomata may
account for the increased accumulation of Hg within the leaf tissues of the Ericaceae shrubs, as
translocation of Hg from the roots to the leaves is often observed to be negligible in vascular
plants (Ericksen et al., 2003) and therefore contribute to the diminished emissions from the
monoliths with greater Ericaceae leaf area coverage.
The relative importance of surface Hg sorption for different plant functional groups may depend
upon such factors as surface roughness (Rea et al., 2000) as well as any potential chemical
compounds present on the vegetation or in the atmosphere which may promote TGM sorption
and the subsequent potential for Hg reduction and re-emission back to the atmosphere (Hanson et
al., 1995; Marsik et al., 2005). Sedge leaves are sleek and glabrous as compared to the scabrous
and leathery surfaces of C. calyculata and K. polifolia leaves. These differences may influence
the strength of Hg deposition and the potential re-emission from the plant surface. In addition to
exposure to solar radiation, leaf surface texture may contribute to the observed differences in
TGM flux dynamics between treatments as well as the leaf surface Hg concentrations between
vascular functional groups. Rea et al. (2000) suggested that the rough surface of birch leaves
with small hairs may more efficiently trap dry deposition as compared to the smooth surface of
maple leaves. The form of Hg deposited to the leaf surfaces may also be influenced by the
94
texture of the leaves. Reactive and water soluble Hg(II) and aerosol Hg were the likely forms of
Hg rinsed from foliar surfaces in the study by Rea et al. (2000). Comparatively, gaseous
elemental Hg(0) is more likely to be quickly re-emitted to the atmosphere following deposition
to foliar surfaces due to its minimal solubility in water, but may also be taken up through stomata
or oxidized to soluble Hg(II) (Hanson et al., 1995; Rea et al., 2000). Leaves are considered
dynamic surfaces for Hg exchange acting as either sites of deposition or emission depending
upon the prevailing conditions such as temperature and moisture (Hanson et al., 1995).
Therefore, leaves may contribute to the predominant emission observed in the Ericaceae-
dominated PEATcosm peat mesocosms, although this flux may vary depending upon the amount
of leaf area.
The observed differences in the strength of the relationships between TGM fluxes and solar
radiation and soil temperature (Figures 3-4 and B-3-2) among the treatments may be due to
differences in vegetative coverage among the plant functional groups, which results in
differential peat shading. Reduced emissions occurred in mesocosms with greater cover of
Ericaceae shrubs which, in addition to the influence of TGM sorption to and potential re-
emission from foliar surfaces, may be the result of diminished solar radiation reaching the peat
surface because of higher leaf area (Figure 3-6a). In forested environments, canopy cover
influences Hg fluxes from the forest floor due to shading and shade impacts on solar radiation
and soil temperatures; with soil temperature becoming an increasingly stronger control as canopy
cover inhibits solar radiation penetration to the forest floor (Choi and Holsen, 2009a). Among the
PEATcosm treatments, peat temperature thresholds between 20.2°C to 24.5°C were observed,
beyond which fluxes increased (Figure B-3-2). A reduction of incoming solar radiation reaching
the peat surface may reduce the amount of photoreduction of Hg(II) to Hg(0) available for re-
emission as seen in a harvested forest system with varying degrees of biomass removal (Mazur et
al., 2014). Shading variations as a result of seasonal progression from leaf-out to leaf-off periods
in a forested system significantly affect the magnitude of Hg flux with the highest fluxes
occurring prior to leaf-out (Choi and Holsen, 2009a). Therefore, shading by the Ericaceae
vegetation may also contribute to the decreased emissions of Hg from the peat surface by
potentially inhibiting photoreduction and Hg re-emission. Shading by shrub cover in bog systems
is therefore an important potential control on TGM fluxes and may influence the relative
contributions of soil temperature and solar radiation in governing Hg exchange dynamics.
95
The trend in TGM fluxes across the PEATcosm treatments is reflected in the THg concentrations
in the surface 0–10 cm of peat, with higher peat THg in monoliths with stronger deposition,
suggesting that the deposited Hg is retained in the surface peat. This is a rather novel finding
since higher surface soil concentrations typically correspond to higher TGM emissions,
particularly in polluted sites (Gustin et al., 2003; Miller et al., 2011), but also less distinctively in
background soils (Gustin et al., 2006; Kuiken et al., 2008). In some low Hg soils (i.e. less than
100 ng g-1
) no relationship between soil Hg concentrations and TGM fluxes may be observed
(Ericksen et al., 2006). The observed negative relationship between flux and surface peat Hg
provides further support for the shuttling of Hg from the atmosphere and its deposition to the
peat by the sedge vascular vegetation. The strength of the correlation may also suggest that there
is minimal re-emission of Hg from the peat surface once deposited.
3.5.1.1 Potential climate change implications for TGM fluxes
Global climate change has the potential to significantly affect the dominant vascular plant
functional groups present in peatlands due to changes in hydrology (Weltzin et al., 2003; Strack
et al., 2006; Breeuwer et al., 2009; Dieleman et al., 2015; Potvin et al., 2015). Regardless of the
direction of shifting plant communities, the exchange of TGM with the atmosphere will likely be
affected. Dominance of Ericaceae shrubs may result in a greater proportion of atmospheric Hg
being sorbed to the leaf surfaces of these plants. This pool of Hg sorbed to the surface of
vegetation may be susceptible to reduction mechanisms including photochemical and subsequent
re-emission to the atmosphere, although Ericaceae leaf coverage will likely reduce emissions
from the peat due to shading. If increased temperatures and elevated atmospheric CO2
concentrations in concert with prolonged periods of water table recession do indeed promote the
establishment of sedges in peatlands (Fenner et al., 2007; Dieleman et al., 2015), enhanced
deposition of TGM to the peat may occur. This deposited Hg may then be bound to the soil
organic matter or contributed to pore water due to the favourable redox environment of the soil,
particularly with aerated surface peat as a result of lowered, fluctuating water tables
(Waddington et al., 2015), promoting the oxidation of gaseous Hg(0) to Hg(II) (Moore and
Castro, 2012). Increased provision of Hg to peatlands by sedge vegetation may have important
implications in terms of both availability for inorganic Hg methylation and inorganic Hg and
methylmercury (MeHg) mobility. Peatlands are ideal environments for MeHg production with
predominantly saturated soils and favourable redox conditions for sulfate and iron-reducing as
96
well as methanogenic bacteria to carry out this reaction (Branfireun et al., 1996; Tjerngren et al.,
2012). Previous research at the PEATcosm experimental mesocosms demonstrated that the
highest THg and MeHg concentrations in pore water and snowmelt runoff were consistently in
the Low WT – sedge only monoliths (Haynes et al., 2017). With greater Hg deposition in these
sedge-dominated wetlands, more Hg is available for the conversion to MeHg and subsequent
export to downstream aquatic ecosystems. The observation of temperature thresholds controlling
TGM fluxes may have important implications for the increasing temperatures associated with
climate change. Prolonged periods above these threshold temperatures have the potential to
increase TGM fluxes from peat, regardless of the dominant vascular plant community. Although
each of the four treatment combinations were only monitored for TGM fluxes on two separate
24-h periods, the mechanisms postulated to account for the observed magnitude and direction of
these fluxes provide an important first look at how these influences simulating climate change
may impact TGM fluxes within peatland ecosystems.
3.5.2 SPRUCE bog fluxes – deep soil warming influence on TGM fluxes
Warming of peat soils at depth influenced TGM fluxes in the SPRUCE bog with a statistically
significant increase in the slope of the surface peat temperature-TGM flux relationship in
response to warming conditions (Figure 3-8). Although consistent temperature differentials of
+4.5 and +9 °C above ambient throughout the peat profile could not be achieved in this initial
SPRUCE deep warming manipulation experiment due to the lack of added air warming,
differences were observed in the surface peat temperatures among the treatment plots (Figure 3-
8). Gustin et al. (1997) concluded that heating of the air above the soil had a greater impact on
Hg flux than heating of the soil itself in contaminated tailings soil cores. Even without added air
warming to the enclosures, a slight warming of the surface peat does appear to result in an
increase in TGM emission to the atmosphere. Previous studies have observed correlative
relationships between both air and soil temperatures and TGM fluxes from background soils
including wetlands (e.g. Poissant and Casimir, 1998; Edwards et al., 2001; Marsik et al., 2005).
Increased temperatures result in enhanced TGM fluxes due to a lowering of the activation energy
for Hg reduction (Carpi and Lindberg, 1998; Kim et al., 2012; Moore and Castro, 2012). Biotic
influences on TGM emissions including microbially-mediated reduction of Hg may also be
accelerated by warming influences by augmenting microbial activity, although they exert less
control than abiotic factors (Choi and Holsen, 2009b). Deep warming within the peat profile does
97
indeed result in enhanced TGM fluxes from the peatland, suggesting either that the slight
warming observed at the surface enhanced reduction processes or that greater warming at depth
resulted in Hg reduction at depth, with only small proportions able to diffuse to the surface for
evasion. With increasing air and soil temperatures as a result of global climate change, TGM
fluxes from peatland ecosystems may be enhanced, but confidence in this conclusion would be
enhanced by further, longer term research at the SPRUCE project, including surface peat
warming.
3.5.3 Peatland TGM fluxes compared to other wetland environments
Only two published studies have previously examined soil-air Hg dynamics in boreal peatland
systems using DFCs (Kyllönen et al., 2012; Fritsche et al., 2014). However, the measurements in
these studies were not conducted continuously throughout 24-h diurnal cycles, which integrate
both daytime and nighttime fluxes. The magnitude of the fluxes recorded for the ombrotrophic,
spruce-Sphagnum, SPRUCE bog is similar to that of other studies of boreal peatlands as well as
other freshwater wetlands. For example, Kyllönen et al. (2012) observed mean hourly fluxes of
0.2 ng m-2
h-1
with values ranging from −0.3 to +0.6 ng m-2
h-1
in a boreal Sphagnum and grass-
dominated wetland. Measured in August under shaded plots within a mixed acid peatland,
Fritsche et al. (2014) observed TGM fluxes 1.39 ± 2.3 ng m-2
h-1
. In a marsh with vegetation
removed from beneath the DFCs, TGM fluxes ranged from 0.35 ± 0.15 ng m-2
h-1
measured in
cloudy conditions to 1.03 ± 0.79 ng m-2
h-1
recorded in sunny conditions (Zhang et al., 2006).
Using a relaxed eddy accumulation system, Osterwalder et al. (2016) observed mean fluxes of
3.0 ± 3.8 ng m-2
h-1
from a boreal peatland during the spring snowmelt period. Collectively, these
studies demonstrate that peatlands and some freshwater wetlands represent nearly neutral fluxes.
The diel pattern in TGM fluxes observed for both the PEATcosm (Figure 3-3) and SPRUCE
peatland sites, with peak fluxes occurring in concert with peak solar radiation and surface soil
temperatures shortly after midday due to radiative heating, is similar to the trends observed over
both mineral soil and open water sites (Poissant and Casimir, 1998) as well as boreal forest soils
(Kyllönen et al., 2012). This trend has similarly been observed for most soils with both
contaminated and background Hg levels (e.g. Poissant and Casimir, 1998), as well as in a boreal
wetland environment (Kyllönen et al., 2012). Despite the moderate TGM fluxes observed in
peatlands as compared to upland, mineral soils, the congruence in diurnal trends suggests that
98
influences such as solar radiation and soil and air temperatures may similarly affect peat soils.
Other meteorological variables such as precipitation patterns and soil moisture (Gustin and
Stamenkovic, 2005) as well as wind speed and direction (Marsik et al., 2005), which are
important controls on TGM exchange dynamics in other environments, may also play key roles
in peatland TGM flux. Given the likely impacts of climate change on the hydrology, ecology and
biogeochemistry of peatland systems, further study on the influence of these factors is warranted.
3.6 Conclusions
Combined changes in water table position and variability, and vascular plant functional groups
may significantly alter TGM deposition to peatlands. Aboveground and belowground biomass of
vascular vegetation is instrumental in controlling TGM dynamics via stomatal exchange, surface
sorption and vascular aerenchyma tissues. Increased Ericaceae shrub abundance with climate
change (Potvin et al., 2015) may alter the dynamic exchange of gaseous Hg via leaf surface
sorption and reduction, leading to more positive (surface to air) emission fluxes. A shift in
dominant vascular plant functional group towards sedges under a changing climate may result in
greater deposition of gaseous Hg to these systems via vascular shuttling under low water table
conditions. This deposited Hg may sorb to the organic matter increasing the amount of Hg
available for methylation or be mobilized in peat pore waters (Haynes et al., 2017). From this
study, the potential impact of increasing peat temperatures on TGM flux is still equivocal. Inter-
site variability was too great to make a strong conclusion about temperature impacts. The deep-
peat-only warming in the current experiment is unlikely to adequately represent future peat
temperature scenarios, particularly at the surface. Still, the significantly shifting trend between
flux and soil temperatures throughout the deep-peat warming experiment is suggestive of a
developing temperature influence on increasing TGM flux. Given the potential antagonistic
impacts of climate change-induced increases in soil and air temperatures combined with altered
water table regimes and shifts in plant functional groups on TGM fluxes, further integrated
investigation into how these factors may, in combination, affect the exchange of gaseous Hg
between the atmosphere and peatland systems is warranted.
3.7 Acknowledgements
We would like to acknowledge the field assistance of K. Ng and S. Rao and the laboratory
assistance of P. Huang. We would like to thank P. Hanson, W.R. Nettles, Oak Ridge National
99
Laboratory, U.S. Department of Energy, and U.S. Forest Service for access to the SPRUCE site.
We thank E.B. Swain for use of his Tekran 2537A Gaseous Mercury Analyzer. Funding was
provided through a Natural Sciences and Engineering Research Council of Canada Alexander
Graham Bell Canada Graduate Scholarship (CGS-Doctoral) to K.M.H and a NSERC Discovery
Grant (Fund #355866) to C.P.J.M. The PEATcosm experiment was funded by the USDA Forest
Service Northern Research Station and the National Science Foundation (DEB-1146149).
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Chapter 4 Impacts of Experimental Alteration of Water Table Regime and Vascular Plant Community Composition on Solid Phase Peat
Mercury Profiles and Methylmercury Production
4
4.1 Abstract
Climate change has the potential to significantly alter the hydrology and vascular plant
community composition in peatland ecosystems. These changes are likely to have as yet
unexplored impacts on mercury (Hg) cycling and net methylmercury (MeHg) production in the
solid phase peat. In this study, peat was collected from the PEATcosm peat monoliths, an
outdoor, controlled experiment where water table positions (High and Low) and vascular plant
functional groups (sedge-dominated, Ericaceae-dominated and unmanipulated control) were
manipulated to simulate potential climate change effects. Potential inorganic Hg methylation
and MeHg demethylation rate constants using enriched stable isotope incubations were assessed
in 2015, and ambient peat THg and MeHg concentrations were tracked annually from 2011
(prior to manipulation) through to the end of the PEATcosm study in 2015. Water table position
significantly affects THg, MeHg and the percentage of THg as MeHg (%MeHg), with elevated
concentrations in the lowered water table treatments within the zone of water table fluctuation.
In mesocosms where Ericaceae shrubs were removed and sedges became the dominant vascular
plant, MeHg concentrations and %MeHg within the rooting zone become significantly elevated.
At approximately 40 cm below the peat surface, where the lowered water tables were located
during the experiment, mean MeHg concentrations are positively and strongly correlated
(r=0.96) with pore water acetate concentrations, which are derived from aerobic peat
decomposition under a lowered water table. Potential Hg methylation rate constants in the
depths coinciding with the lowered water table treatment are similar under both water table
treatments. No significant treatment effects on demethylation are observed. Enhanced
partitioning of inorganic Hg and MeHg from the solid phase peat into pore waters occur with a
lowered, fluctuating water table and sedge vegetation. Lowered fluctuating water tables and
110
sedge-dominated vascular plant communities with a changing climate may enhance inorganic Hg
and MeHg accumulation within the zone of water table fluctuation via partitioning and
translocation, more so than net MeHg production.
4.2 Introduction
Boreal peatland ecosystems store large amounts of atmospherically-deposited mercury (Hg) in
accumulated organic soils (Kolka et al. 2001; Grigal 2003), on the order of 7 to 44 µg total Hg
m-2
y-1
including litterfall (Kolka et al. 2011). These systems are also considerable sources of
methylmercury (MeHg) to downstream aquatic ecosystems (Heyes et al. 2000), with wetland-
dominated catchments exporting approximately 26 to 79 times higher loads than upland
environments (St. Louis et al. 1994). A potent neurotoxin, MeHg poses a significant threat to
vulnerable wildlife and human populations as it biomagnifies and accumulates to significant
concentrations at higher trophic levels (Mergler et al. 2007; Scheuhammer et al. 2007). With
predominantly saturated, anoxic soils, peatland systems are optimal sites for the production of
MeHg (Tjerngren et al. 2012); a process mediated by a relatively wide array of anaerobic
microbes such as sulfate- and iron-reducing bacteria and fermentative archaea (Gilmour et al.
2013). Sulfate-reducing and methanogenic microbes are also known to facilitate oxidative
demethylation in freshwater ecosystems (Oremland et al. 1991; Marvin-DiPasquale and
Oremland 1998; Marvin-DiPasquale et al. 2000). Demethylation may also occur via abiotic
mechanisms including photo-decomposition (Sellers et al. 2001; Lehnherr and St. Louis 2009;
Lehnherr et al. 2012a), which may be influenced by the presence of dissolved organic matter and
solutes such as ferric iron and nitrate (Chen et al. 2003; Hammerschmidt and Fitzgerald 2010;
Kim and Zoh 2013). Overall, MeHg concentrations in natural systems reflect the net balance
between methylation and demethylation processes.
Climate change has the potential to significantly affect the ecosystem function and carbon
storage abilities of peatlands through changes in hydrology (Ise et al. 2008) and as a
consequence, significantly impact peatland mercury cycling. Increased evaporation associated
with increasing temperatures, as well as enhanced variability in precipitation patterns in the
northern continental United States (Groisman et al. 2012; Janssen et al. 2014; Yu et al. 2016)
will likely result in prolonged water table drawdown and considerable fluctuations in water table
position particularly during the summer months (Thomson et al. 2005; Whittington and Price
111
2006). In peatlands, greater fluctuations in water table position may drastically alter
biogeochemical cycling and transformations of several elements (Limpens et al. 2008), including
Hg (Coleman Wasik et al. 2015), as well as peatland vegetation community structure (Dieleman
et al. 2015; Weltzin et al. 2003; Breeuwer et al. 2009). Redox-sensitive processes such as Hg
methylation are likely to be affected by longer sustained peat aeration as a result of periodic
water table drawdown and fluctuation (Coleman Wasik et al. 2015). Regeneration of terminal
electron acceptors, such as the oxidation of sulfide to sulfate during periods of peat aeration,
enhance MeHg production (Coleman Wasik et al. 2015) and MeHg mobilization in peatland pore
waters (Haynes et al., 2017a). However, factors that affect Hg methylation may also impact the
process of MeHg demethylation, thereby influencing net MeHg production (Lehnherr 2014). For
example, microbial demethylation is likely to exceed methylation under oxidative conditions in
aquatic environments (Ullrich et al. 2001). Warmer temperatures may enhance both Hg
methylation (Yang et al. 2016a) and MeHg demethylation (Matilainen et al. 1991) by stimulating
microbial activity. In Arctic soils where warming temperatures are resulting in permafrost thaw,
Hg methylation under anoxic conditions was strongly associated with methane (CH4) fluxes
(Yang et al. 2016b). Both methylation and methanogenesis were enhanced due to accelerated
degradation of labile soil organic matter (Yang et al. 2016b). The hydrological controls on the
balance between Hg methylation and MeHg demethylation in peatlands have not been previously
investigated. Therefore, the effect of more drastically fluctuating water tables, as may be
observed due to global climate change, on the THg and MeHg storage capacity in peatlands is
unclear.
With prolonged periods of water stress due to less frequent but more intense precipitation events
(Groisman et al. 2012; Janssen et al. 2014; Yu et al. 2016), the dominant vascular plant
functional groups present in peatlands may shift (Weltzin et al. 2003; Strack et al. 2006;
Breeuwer et al. 2009; Dieleman et al. 2015; Potvin et al. 2015). The degree to which plant
communities change may depend upon the severity of water stress and the adaptive abilities of
the different plant functional groups to survive in such conditions (Potvin et al. 2015). Sedges
may thrive under conditions where the water table position becomes more variable. Typically
located on hummocks and lawns due to the inability of shallowly-rooted Ericaceae shrubs to
survive in flooded conditions, shrub coverage and productivity are anticipated to increase with a
warmer and drier climate (Potvin et al. 2015; Strack et al. 2006; Weltzin et al. 2003). Prolonged
112
water stress however, may suppress Ericaceae shrub growth (Potvin et al. 2015). The
physiological features of sedges, such as deeper roots compared to other vascular vegetation,
allow sedges to access deeper water stores during periods of water deprivation (Strack et al.
2006; Dieleman et al. 2015). Sedges also regulate water losses via stomata (Breeuwer et al.
2009) and expand the rooting zone by actively shuttling oxygen via aerenchymous tissues in
waterlogged conditions (Armstrong et al. 1991). Therefore, altered plant communities may
influence both Hg methylation and demethylation processes. A potential shift toward sedge-
dominated plant communities may enhance MeHg production by altering the peat redox potential
through oxygen shuttling and the availability of terminal electron acceptors within the rooting
zone (Mueller et al. 2016). The provision of labile sources of carbon, such as acetate via root
exudates, may stimulate sulfate and iron reduction, and thus also augment methylation
(Windham-Myers et al. 2009; 2014). In contrast, Ericaceae shrubs, the roots of which form a
symbiotic relationship with mycorrhizal fungi, may affect the soil microbial communities within
the rooting zone. The presence of Ericaceae shrubs may influence the redox potential of the peat
by outcompeting heterotrophs for available oxygen (Romanowicz et al. 2015) and therefore may
reduce peat net MeHg production. The potential effect of the dominance of either sedges or
Ericaceae shrubs under warmer and drier conditions on peat MeHg production has not been
previously examined.
Given the large stocks of Hg in peatlands, the connection of these wetlands to downstream
aquatic ecosystems and the vulnerability of these systems to a changing climate, it is important to
understand how combined changes in hydrology and plant communities may affect the net
production of MeHg in peatlands. In this study, peat at several depths was collected throughout
the course of the PEATcosm (Peatland Experiment at the Houghton Mesocosm Facility)
experiment, annually from 2011 through 2015, to assess the influence of simulated climate
change effects on solid phase peat THg and MeHg. PEATcosm is an outdoor, controlled
experiment that manipulated water table positions and vascular plant functional groups in peat
monoliths to simulate potential climate change impacts. Pore waters were also collected to
determine the effect of the experimental treatments on Hg and MeHg partitioning. Enriched Hg
isotope-based measurements of Hg methylation and MeHg demethylation potential rate constants
were completed at the conclusion of the overall PEATcosm study, once larger-scale destructive
sampling was possible. The main hypotheses of this experiment are: 1) The water table and
113
plant functional group treatments will increase solid phase peat THg and MeHg accumulation
throughout the peat profile , 2) Hg methylation will be enhanced within the zone of the lowered
water tables due to the regeneration of terminal electron acceptors (i.e. sulfate) and increased
labile carbon availability as a result of peat decomposition and 3) MeHg and inorganic Hg
partitioning from the solid phase into the peat pore waters will be enhanced with a lowered,
fluctuating water table and dominant sedge vegetation.
4.3 Methods
4.3.1 Study Site and Experimental Design
The PEATcosm experiment is located at the United States Forest Service Mesocosm Facility at
the Forestry Sciences Laboratory in Houghton, Michigan, USA (47.11469° N, 88.54787° W).
The regional climate is humid continental with typical annual precipitation of approximately 870
mm, approximately 50% of this falling as snow. Mean temperatures in this area range from -
13°C in January to +24°C in July (30 year means at Houghton County Airport; Potvin et al.
2015).
Twenty-four intact 1-m3 (1 m x 1 m x 1 m) peat monoliths were harvested from an ombrotrophic
peatland located in Meadowlands, MN, USA in May 2010 and transferred into individual
mesocosm bins. The interior of each stainless steel bin was coated with Teflon to prevent
potential metal transfer between the bin and the peat. The top of each mesocosm bin was open,
exposing the peat to the ambient climate. The bins were insulated and installed in a climate-
controlled tunnel, allowing belowground access to each of the bins as well as the simulation of a
natural vertical temperature gradient. Potvin et al. (2015) provides a detailed and comprehensive
explanation of the peat monolith harvest and experimental set-up. The experiment concluded in
July 2015 with the destructive harvest of the peat monoliths.
The PEATcosm study comprises a full-factorial experimental design with two water table (WT)
prescriptions crossed with three different plant functional group treatments simulating potential
climate change outcomes. Each treatment combination is replicated across four mesocosms, in a
randomized complete block design, for a total of 24 experimental units. Vascular plant
functional group manipulations were initiated in 2011 and water table experimental treatments
were implemented in 2012. The water table treatments were based on long-term (approximately
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50 years) data from the Marcell Experimental Forest in north-central Minnesota (47.51907° N,
93.45966° W), located near the peat monolith harvest site. The two water table treatments were
modelled after years with: 1) typical variability and average water table position (referred to as
‘High WT’) and 2) comparably high variability and lower average water table position (referred
to as ‘Low WT’). These target water table positions were maintained through a combination of
artificial precipitation additions, rain-exclusion covers and regulated outflow from approximately
the acrotelm-catotelm boundary (Potvin et al. 2015). The three plant functional group treatments
simulating likely community composition alterations resulting from climate change (Strack et al.
2006; Weltzin et al. 2000; Chapin et al. 1996), were as follows: 1) all Ericaceae removed
(referred to as the ‘sedge only’ treatment), 2) all sedge removed (‘Ericaceae only’ treatment) and
3) both sedge and Ericaceae present (‘unmanipulated control’ treatment). The dominant sedge
species present in the bins was Carex oligosperma Michx., while the dominant Ericaceae shrubs
included Chamaedaphne calyculata (L.) Moench., Kalmia polifolia Wangenh., and Vaccinium
oxycoccos L. The dominant moss species in the peat monoliths were Sphagnum rubellum
Wilson, S. magellanicum Brid., S. fuscum (Schimp.) Klinggr and Polytrichum strictum Brid.
Polytrichum commune Hedw., Eriophorum vaginatum L., Andromeda polifolia L. var.
glaucophylla (Link) DC., Rhododendron groenlandicum (Oeder) Kron and Judd, and Drosera
rotundifolia L. were also present in the mesocosms.
In 2011 mean water table positions between the beginning of June and the end of October were 7
± 5 cm (mean ± standard deviation) below the peat surface for the High WT treatments and 11 ±
4 cm for the Low WT bins. Mean 2012 water table positions were 11 ± 4 cm for the High WT
mesocosms and 19 ± 8 cm for the Low WT treatments. In 2013 a mean differential of
approximately 20 cm was imposed between the High WT (15 ± 5 cm) and the Low WT
treatments (35 ± 11 cm). Similarly in 2014 the High WT (12 ± 5 cm) and Low WT bins (33 ± 12
cm) had an approximate mean differential of 20 cm from June through to the end of October.
4.3.2 Peat Sampling
Peat samples were collected for ambient THg and MeHg analyses from each of the 24
mesocosms in August 2011, 2012, 2013 and 2014. Cores were also collected in July 2015 when
the mesocosms were destructively harvested and THg and MeHg concentrations were obtained
simultaneously with methylation and demethylation potential assays (see explanation next
115
paragraph). Cores were collected using a stainless steel corer from which the peat was extruded
and divided in 10 cm increments. Core lengths were typically 60 cm. The corer was rinsed with
deionized water between sampling each mesocosm. Peat samples were immediately frozen
following sub-sectioning. In 2012, peat samples were only analyzed down to 40 cm below the
peat surface as the water table treatments were only initiated during this growing season and
were not dramatically drawn down.
To determine potential Hg methylation and MeHg demethylation rate constants (kmeth and kdemeth,
respectively), cores collected from each mesocosm in July 2015 were injected with enriched,
stable inorganic Hg and MeHg isotopes (Hintelmann et al. 1995; Hintelmann and Ogrinc 2003).
Surface peat (0-20 cm) and deeper peat cores (20-60 cm) were sampled using clear
polycarbonate corers (4.8 cm internal diameter) with silicone septa (1/16”) spaced at 1 cm
intervals. Once extracted from the peat monolith, the core tubes were capped immediately at
both ends. Capped cores remained upright and stored at ambient peat temperatures in an
incubator until the isotopic injections were completed (within approximately 2 hours). Potential
mercury methylation rate constants were assessed using spike additions of enriched 200
Hg
(94.3% abundance), while potential demethylation rate constants were measured using additions
of enriched Me201
Hg (84.7% abundance). Spike solutions were made by diluting the 200
HgCl+
and Me201
HgCl+ stock solutions with filtered pore water collected from each peat mesocosm on
the same day as peat collection. The solutions were allowed to equilibrate for approximately one
hour prior to injection into the peat cores, with the assumption that enriched isotopes would
equilibrate with the dissolved organic carbon present in the pore waters. Through each of the 1
cm-spaced silicone septa from 0-15 cm and 35-50 cm, 100 µL of the prepared spike solution (~2
µg mL-1
200
Hg and ~0.07 µg mL-1
Me201
Hg) was injected using a gas-tight borosilicate glass
syringe into the peat. As the added isotopic tracers are likely more bioavailable than ambient
inorganic Hg and MeHg, the values of kmeth and kdemeth represent potential methylation and
demethylation rate constants, respectively (Mitchell and Gilmour 2008). Following the spike
additions the cores were incubated in the dark in an incubator for six hours at 18°C (ambient peat
temperature measured at approximately 5 cm below the peat surface in the mesocosms). After
the six-hour interval, the peat was extruded from the tubes and sectioned into 5 cm increments.
Each peat sample was homogenized using a stainless steel hand blender, which was thoroughly
116
washed with deionized water between each sample. Samples were then immediately frozen to
terminate the methylation and demethylation incubations.
4.3.3 Pore Water Sampling
Pore waters were collected from micro-piezometer nests; ultra-high-density polyethylene casings
with Teflon tubing, located at 20, 40 and 70 cm below the peat surface in each of the 24
mesocosms (see Romanowicz et al. 2015 for complete piezometer construction details). Both
THg and MeHg analyses were performed on pore waters collected in June, August and
November 2013, May, July and September 2014 and in May 2015 prior to the destructive harvest
of the mesocosms at the conclusion of the experiment. Detailed description of the pore water
sampling and analytical methods, including quality control, can be found in Haynes et al.
(2017a). For other chemical analyses, including sulfate and acetate concentrations, pore waters
were collected monthly throughout the growing season of each study year. For this experiment,
only pore water acetate concentrations collected at 40 cm below the peat surface were considered
in association with solid phase peat THg and MeHg concentrations at corresponding depths (30-
40 and 40-50 cm).
Soil-water partition coefficients (KD, expressed in units of L kg-1
) were determined according to
the following equation:
𝐾𝐷 = [𝐻𝑔]𝑆
[𝐻𝑔]𝑃𝑊 (4.1)
where [Hg]S is the ambient solid phase peat Hg (either inorganic Hg or MeHg) concentration (in
ng kg-1
) and [Hg]PW is the ambient pore water Hg (inorganic Hg or MeHg) concentration (in ng
L-1
). Values of KD are expressed in log form. Partition coefficients were calculated for both
experimental years of 2013 and 2014 at 20 cm and 40 cm below the peat surface to coincide with
the zone of greatest water table variability. Mean solid phase peat concentrations at 10-20 cm
and 20-30 cm, as well as 30-40 cm and 40-50 cm were determined to correspond with the pore
waters collected at 20 cm and 40 cm below the peat surface, respectively. Mean annual pore
water Hg concentrations for each of 2013 and 2014 were used to determine the KD coefficients
since pore waters were not collected at the same time as the solid phase peat.
117
4.3.4 Analytical Methods
All frozen peat samples were freeze-dried prior to analysis. For MeHg analysis, both ambient
and tracer incubated peat samples were distilled in Teflon vessels according to US EPA Method
1630 (US EPA Method 1630, 1998). Prior to distillation a known, trace amount of enriched
stable Me199
Hg isotope was added to each sample as an internal standard (Hintelmann and
Evans, 1997). Both ambient and excess MeHg concentrations were assessed by gas
chromatography-inductively coupled plasma mass spectrometry (GC-ICP-MS) using an Agilent
Technologies 7700x ICP-MS for Hg isotope detection. Ambient and excess tracer
concentrations were calculated as per the isotope dilution calculations of Hintelmann and Ogrinc
(2003). Recovery of standard reference material (estuarine sediment ERM CC580) was 100 ±
6% (n=42), replication of duplicates was 3.9 ± 3.3% (n=37) and the MeHg detection limit was
calculated (n=42 matrix blanks) to be 0.05 ng g-1
. For THg analysis, the peat samples were
digested in hot nitric acid and the diluted digestates were analyzed by cold vapor atomic
fluorescence spectroscopy (CVAFS) according to US EPA Method 1631 (US EPA Method
1631, 2002) using a Tekran 2600 automated Total Mercury Analyzer. For detection of the added
inorganic 200
Hg tracer in the 2015 incubated peat samples, the Tekran 2600 analyzer was directly
hyphenated to the ICP-MS. Recovery of a THg spike was 101 ± 4% (mean ± standard deviation,
n=38), replication of duplicates was 3.4 ± 2.8% (n=41) and the detection limit was 0.04 ng g-1
(n=40 matrix blanks). Recovery of standard reference material (MESS-3) following digestion
was 101 ± 5% (n=42).
Potential rate constants for inorganic Hg methylation (kmeth) were calculated using the
concentration of the isotopic tracer solution that was methylated over the course of the six hour
incubation period with respect to the abundance of the isotope tracer (Hintelmann et al. 1995;
Hintelmann and Ogrinc 2003). Time zero cores for Hg methylation were not collected, which is
consistent with other studies of methylation rate potential determination (e.g. Mitchell and
Gilmour 2008). Potential MeHg demethylation rate constants (kdemeth) were determined
assuming first-order reaction kinetics according to Lehnherr et al. (2012b). The initial
concentrations of excess Me201
Hg were assumed to be those of excess total 201
Hg at the end of
the incubation as time zero peat cores were not collected. This approach facilitated the
determination of MeHg degradation over the course of the incubation period, as the Me201
Hg
spike solution contained no 201
Hg prior to injection. Detection limits for both kmeth and kdemeth
118
potential rate constants were determined based on the error associated with isotope ratio
measurements, ambient peat MeHg and inorganic Hg concentrations and the tracer spike levels,
as calculated by Mitchell and Gilmour (2008). The detection limit for kmeth ranged from 2 x 10-5
to 0.028 d-1
(mean 0.0028 d-1
). The detection limit for kdemeth ranged from 0.09 to 0.16 d-1
(mean
0.12 d-1
).
Pore water acetate concentrations were determined using a Dionex ICS 2000 ion chromatograph
(Dionex Corp.). Analytical methods and quality control information for pore water THg and
MeHg concentration determination are provided in Haynes et al. (2017a).
4.3.5 Statistical Analyses
All statistical analyses were performed using R statistical software (R Development Core Team,
2014) with α = 0.05. All data were tested for normality (Shapiro-Wilk W test) and
heteroscedasticity and were successfully log-transformed to achieve normality when parametric
assumptions were not met. Repeated measures analyses of variance (ANOVA) were performed
with water table, plant functional group and depth as the main factors to explore the direct and
interactive treatment effects on THg and MeHg concentrations and %MeHg from the initial
sampling in 2011, following preliminary exposure to the imposed treatments in 2012, through to
the full experimental treatment years of 2013 and 2014. In addition to concentrations of THg
and MeHg expressed in units of ng g-1
, solid phase concentrations in units of ng cm-3
were also
considered for this analysis to account for the changes in peat bulk density that occurred with
time throughout the experiment and that naturally occur down the peat profile. The relationship
between mean peat MeHg concentrations and mean pore water acetate concentrations at 40 cm
below the peat surface was assessed using a Pearson correlation.
For the kmeth and kdemeth data from the 2015 peat core incubations, two-way ANOVAs were
performed for each individual isotopically-spiked 5 cm depth increment from 0-15cm and 35-
50cm to assess significant influences of water table position, plant functional groups and
potential interactions between these factors at each spiked depth. When no significant
interactions between the two treatment factors were observed, one-way ANOVAs were also
performed to determine significant differences in potential methylation and demethylation rate
constants among the six crossed water table and vascular plant functional group treatments. The
119
correlative relationships between kmeth, kdemeth and the percentage of THg present as MeHg in the
peat were investigated using Pearson correlations.
Significant differences in soil-water partition KD coefficients for each sampling depth and year
were assessed individually with two-way ANOVAs to determine any significant direct and
interactive influences of the water table and plant functional group treatments. As no interactive
effects were observed at any depth in both 2013 and 2014, one-way ANOVAs were also applied
to the KD values to determine significant differences among the six crossed treatments. The
correlative relationship between pore water acetate concentrations collected at 40 cm depth and
mean peat MeHg concentrations from 30-40 and 40-50 cm below the peat surface for the six
treatments was examined in 2014 using Pearson correlation. These depths were selected as the
mean lowered water tables were located around 40 cm below the peat surface and the peak peat
MeHg concentrations in the Low WT mesocosms were observed at this depth.
4.4 Results
Over the course of the PEATcosm experiment both peat MeHg concentrations and the
percentage of THg as MeHg (%MeHg) were significantly affected by the altered water table
positions (p = 0.02 and < 0.01, respectively). Peat MeHg concentrations and %MeHg
significantly changed with depth throughout the peat profile (both p < 0.0001). The differences
observed among the treatments primarily occurred within the zone of water table fluctuation of
the lowered water table treatments with elevated MeHg concentrations at these depths in the Low
WT monoliths as compared to the High WT treatments (Figure 4-1). A significant interaction
between peat depth and water table treatment was observed for peat MeHg concentrations and
%MeHg (p ≤ 0.01), suggesting that the depth at which peak MeHg concentrations were located
changed in response to the implementation of a lowered water table. The depth at which the
peak MeHg concentrations occurred within the peat profile changed over time (significant
interaction between sampling depth and year; p < 0.0001), as the depth profile of the peat MeHg
concentrations became increasingly similar with prolonged exposure to the water table and
vascular plant functional group treatments (Figure 4-1). The influence of water table position on
MeHg concentrations was not significant at 30-40 and 40-50 cm below the peat surface in 2013
(p = 0.07-0.08). In 2014 MeHg concentrations and %MeHg were significantly elevated at 30-40
cm below the surface in the Low WT treatment peat (p < 0.01 and 0.05, respectively). The
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observed changes in the peat THg and MeHg concentrations throughout the peat profile can be
attributed to the imposed treatments as no systematic trends with depth were observed prior to
the experimental manipulations. In 2011, prior to the implementation of the plant functional
group treatments and with a mean difference of approximately 4 cm between the water table
positions of the High and Low WT bins, the profiles of THg concentrations were very similar
among the mesocosms assigned to each of the six treatment combinations (Figure 4-1).
Methylmercury concentrations also did not have any systematic pattern among the assigned
treatments in 2011 (Figure 4-1). In 2012 with the initiation of the water table treatments and a
mean growing season differential of approximately 8 cm between the Low and High WT
prescriptions, all treatments had similar THg and MeHg concentrations down to the water table
level regardless of the treatment prescription (Figure 4-1), likely as a function of the maintenance
of the water table positions. Similar trends were observed in MeHg and THg stocks, which
accounted for changes in solid phase bulk density over time, expressed in ng cm-3
, throughout
the peat profile (Figure C-4-1).
Figure 4-1 Mean peat MeHg and THg concentration (ng g-1
) profiles for the six crossed water
table and vascular plant functional group treatments throughout the PEATcosm experiment from
2011 to 2014 and associated mean June to October water table positions for the High and Low
WT prescriptions. For clarity, error bars for each 10 cm depth increment have been omitted.
121
Solid phase THg concentrations significantly changed with peat depth (p < 0.0001). A
significant interaction was observed between the water table treatments and the depth of peat
sampled (p < 0.0001), suggesting that the influence of the imposed water table treatments
changed throughout the peat profile. Similar to the peat MeHg concentrations, the depth at
which the peak THg concentrations occurred in the peat changed over the course of the
experiment (p < 0.0001), with a deepening of the peak THg concentrations from pre-treatment to
experimental conditions (Figure 4-1). The impact of the water table treatments became apparent
over the course of the experiment (significant interaction between water table treatment and the
sampling year; p < 0.01), with the peat within the zone of water table fluctuation of the Low WT
treatment increasing in THg concentrations with time as compared to the High WT treatment.
Peat THg concentrations at both 30-40 and 40-50 cm below the peat surface in 2013 were
significantly affected by the water table treatments (both p < 0.05), with higher peat THg
concentrations in the Low WT bins at these depths in 2013. In 2014, water table position
significantly affected THg concentrations in the peat 30-40 cm below the surface (p < 0.01), with
greater concentrations near the mean water table positions in the Low WT prescription
mesocosms (Figure 4-1).
The vascular plant functional group treatments also significantly affected the peat MeHg
concentrations and %MeHg (both p < 0.001). A significant interaction between the plant
functional groups and the depth of the peat sampled for both MeHg concentrations and %MeHg
(p < 0.01 and p = 0.03, respectively) suggests that only certain depths within the peat profile
were influenced by the vegetation communities. Within the top 30 cm of the peat profile, MeHg
concentrations and %MeHg were significantly enhanced in the sedge only treatments as
compared to the Ericaceae only and control vegetation bins (Figure 4-1). No significant impact
of vascular plant functional group was observed directly on peat THg concentrations (p = 0.31)
or in association with peat depth (p = 0.07).
A strong, positive correlation was observed between pore water acetate concentrations and mean
peat MeHg concentrations around the depth of the mean water table (r = 0.96, p = 0.002; Figure
4-2). This depth of 40 cm below the peat surface was selected to correspond with the zone of
lowered water table fluctuation where the greatest influence of the water table treatments was
observed on peat MeHg and THg concentrations (Figure 4-1).
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Figure 4-2 Relationship between mean treatment peat MeHg concentrations (ng g-1
) at 40 cm
below the peat surface (mean of the 30-40 and 40-50 cm depth increments) and pore water
acetate concentrations at 40 cm depth.
4.4.1 Hg Methylation and MeHg Demethylation
Based on the isotopic incubations of peat collected in 2015, both water table and plant functional
group treatments significantly influenced potential Hg methylation rate constants at depths
throughout the peat profile (Figure 4-3). The potential Hg methylation rate constants in the High
WT treatment bins were relatively consistent down the peat profile, with the kmeth of the aerated
upper 10 cm of the peat profile significantly lower than that in the 35-40 cm peat (Figure 4-3).
In the Low WT treatment mesocosms, the kmeth values in the aerated upper 0-15cm peat were
significantly lower than all depth increments below the water table at 35-50 cm below the peat
surface (Figure 4-3). In terms of MeHg demethylation, no significant differences were observed
with depth for either the High WT or Low WT treatments (Figure C-4-2). No strongly
significant influences of either water table position or vascular plant functional group on
demethylation were observed throughout the individual 5 cm increments located 0-15 cm and
35-50 cm below the peat surface. A significant (p < 0.0001), albeit weak (r = 0.50) positive
correlation was observed between the Hg methylation potential and the percentage of THg
present as MeHg (%MeHg) in the peat (Figure 4-4a). No relationship was observed between the
MeHg demethylation potential and %MeHg (p = 0.55; Figure 4-4b).
123
Figure 4-3 Mean kmeth (d-1
) in the upper 15 cm of the peat profile and 35-50 cm below the peat
surface for the a) High WT and b) Low WT treatments. Mean water table levels for the month
prior to sampling are denoted by the dashed lines. Letters denote statistically similar depths for
each of the High and Low WT data.
Figure 4-4 Relationship between the fraction of THg as MeHg (%MeHg) in the peat and a) kmeth
(% d-1
), b) kdemeth (% d-1
).
124
When looking at the individual 5 cm peat increments, a significant influence of water table
position on kmeth values was observed at both 5-10 cm and 10-15 cm below the peat surface, with
higher potential methylation rate constants in the High WT monoliths (p ≤ 0.01; Figure 4-5). No
significant effect of either water table or vascular plant functional group was observed in the
surface 0-5 cm as both treatment water table positions were maintained lower than 5 cm (Figure
4-5).
Figure 4-5 Mean kmeth (d-1
) among the six experimental treatments in the upper 15 cm of the peat
profile (left column) and 35-50 cm below the surface (situated beneath the water table of the
Low WT treatments) (right column). Letters denote statistically similar groups among the six
treatments in the 10-15 cm and 35-40 cm depth increments. No significant differences among
treatments for any other peat depths.
125
Once below the water table level of the Low WT treatment, a significant influence of the
vascular plant functional group was apparent (p < 0.01) and this effect was dependent upon the
water table treatment, with a significant interaction between water table and plant functional
group (p = 0.02; Figure 4-5). Both the sedge only and Ericaceae only vegetation treatments with
high water tables and the Low WT Ericaceae had significantly lower kmeth values than the High
WT control vegetation treatment (Figure 4-5). No significant differences in kmeth among the
treatments were observed below 40 cm.
4.4.2 Hg(II) and MeHg Partitioning
The partitioning of MeHg and inorganic Hg between the peat and pore water was significantly
affected by the variations in water table position and shifts in plant functional group (Figures 4-6
and C-4-3). At a depth of 20 cm below the peat surface, the water table treatment effect was
significant in both 2013 and 2014 (p < 0.01), with lower KD values in the Low WT monoliths
(Figures 4-6c and C-4-3c). Lower KD values in the Low WT treatment mesocosms (Figure 4-6a)
suggest that enhanced partitioning of Hg into the mobile pore water phase occurred with
lowered, more variable water table position. No significant water table or plant functional group
effects on the KD values were observed at a depth of 40 cm in 2013 (Figure 4-6b). Inorganic Hg
partitioning at both 20 and 40 cm were not significantly affected by either the water table or the
plant functional group treatments in 2013 (Figure C-4-3a, b). The vascular plant functional
groups present in the mesocosms significantly affected solid phase-pore water partitioning of
both MeHg and inorganic Hg at a depth of 40 cm in 2014 (p = 0.02 for MeHg, Figure 4-6d; p <
0.01 for THg, Figure C-4-3d).
126
Figure 4-6 Soil-water partition coefficients (logKD) for MeHg at 20 and 40 cm below the peat
surface among the six crossed water table and plant functional group treatments in 2013 and
2014 (n = 4 per treatment). Letters denote statistically similar groups.
4.5 Discussion
Both the imposed water table treatments as well as the shifts in the vascular plant functional
groups significantly affected the accumulation of THg and MeHg within the peat profile. The
THg profile of the peat installed in the mesocosms prior to any experimental manipulations of
water table and vascular vegetation communities was similar to that observed by Rydberg et al.
(2010) in a Swedish mire and likely reflects the historical record of past Hg atmospheric
deposition (Biester et al. 2007). The observed significant influence of water table position on
peat MeHg and THg concentrations and the shift in peak concentrations to approach the zone of
lowered mean water table fluctuation under the experimental conditions may suggest that
lowered, fluctuating water tables may, over time, enhance MeHg and THg accumulation in the
127
solid phase peat. Significant differences in MeHg concentrations and %MeHg in the upper 30
cm of peat as a result of the plant functional group treatments were also observed. This depth
coincides with the location of the majority of the shallowly-rooted Ericaceae shrub roots in the
Ericaceae only and control vegetation mesocosms. Beyond the shallow depths, sedge root
density typically increased in the PEATcosm monoliths (Romanowicz et al. 2015).
Multiple mechanisms may be responsible for the observed increased accumulation of MeHg
within the zone of water table fluctuation, which coincides with the Ericaceae shrub rooting
zone. One such mechanism is the balance between the competing processes of Hg methylation
and MeHg demethylation. The strong positive relationship between %MeHg and kmeth and
overall lack of a significant relationship between %MeHg and kdemeth may collectively suggest
that Hg methylation was a stronger control on net MeHg accumulation in the peat monoliths than
the opposing process of MeHg demethylation (Mitchell and Gilmour 2008; Lehnherr et al.
2012b). Enhanced Hg methylation has been observed to occur in peat subject to water table
fluctuations, due to the regeneration of the terminal electron acceptors, such as sulfate, required
for Hg methylating microbial communities (Coleman Wasik et al. 2015). The observed
significant influence of water table position and variability on %MeHg may be an indication that
Hg methylation may be affected by the changes in redox conditions within the zone of water
table fluctuation.
However, potential Hg methylation rate constants were not significantly influenced by water
table position and variability at the depth of the lowered water table treatments. Within the peat
depths near the lowered water table position in the Low WT mesocosms, kmeth values were not
significantly different between the Low and High WT treatments. This is surprising given that
the mean water table position for the Low WT treatments was approximately 40 cm below the
peat surface during the peat collection and isotopic incubation. The lack of difference in kmeth at
the lowered water table position may suggest that the fluctuating water table position may be
suitable for Hg methylation to occur and this process is not suppressed by periodic oxic
conditions. In the PEATcosm mesocosms, Haynes et al. (2017a) determined that the pore water
sulfate concentrations were significantly influenced by the water table position with elevated
pore water sulfate concentrations in the Low WT mesocosms. These pore water sulfate
concentrations were typically highest in the 40 cm depth pore waters of the Low WT bins, which
corresponds with the zone of water table fluctuation and the peak peat MeHg concentrations.
128
Therefore, increased regeneration and availability of sulfate under lowered and variable water
tables may prime the production of MeHg in the solid phase peat and result in similar
methylation potential as is observed under flooded conditions.
Increased availability of labile carbon under Low WT conditions may also contribute to the
enhanced MeHg concentrations in the solid phase peat in the zone of water table fluctuation and
the similarity in kmeth between the Low and High WT treatments. Available labile forms of
carbon, including acetate, act as an energy source for methylating microbial communities
including iron- and sulfate-reducing bacteria and methanogens (Mitchell et al. 2008; Yang et al.
2016b). King et al. (2000) demonstrated that acetate-consuming sulfate-reducing bacteria
actively methylate Hg. The strong correlation between pore water acetate concentrations at 40
cm depth and the mean solid phase peat MeHg concentrations in the corresponding depths across
the six treatment combinations suggest that an available source of labile carbon for methylating
microbial communities may promote methylation in these monoliths. This effect has similarly
been observed in association with added sulfate which stimulates microbial Hg methylation
(Mitchell et al. 2008), and combined with sulfate and other terminal electron acceptor
regeneration due to fluctuating water levels in the PEATcosm mesocosms, MeHg production
may be augmented over time.
The vascular plant functional groups may also influence the balance contributing to net MeHg
production in the peat. Within the rooting zone of the Ericaceae shrubs, oxygen may be taken up
by the roots and associated mycorrhizal fungi of these plants thereby limiting its availability to
heterotrophs (Romanowicz et al. 2015). The presence of Ericaceae shrubs has been observed to
suppress heterotrophic peat decomposition and Hg mobility in pore water due to oxygen
availability (Haynes et al., 2017a). Limited availability of labile carbon sources due to
heterotrophic suppression of decomposition with the presence of Ericaceae shrubs may
contribute to the lower observed MeHg concentrations as a result of reduced Hg methylation.
The consumption of oxygen by heterotrophic bacteria may also suppress terminal electron
acceptor regeneration, thereby limiting methylation potential. Therefore, the removal of
Ericaceae shrubs and their mycorrhizal fungi may result in elevated MeHg under both high water
table and dry conditions. The establishment of sedges may also influence net MeHg production
in the peat. The shuttling of oxygen by the aerenchymous tissues of the sedges may influence
Hg methylation in the peat due to more oxic conditions within the rooting zone (Crow and
129
Wieder 2005). The presence of both sedge and Ericaceae vascular vegetation in the control
treatments may balance the availability of oxygen and, in conjunction with the peat water table
position, provide an optimal redox state for methylation to occur in the solid phase. This may
account for the elevated potential Hg methylation rate constants in the control vegetation
treatments observed just below the water table.
Given that collectively the observed kmeth across the treatments cannot solely account for the
shifts in peak MeHg and THg concentrations in the peat profile, partitioning of inorganic Hg and
MeHg from the solid phase into the mobile pore waters may be a stronger control. The
mechanism by which accumulation at the observed depths may occur likely includes
translocation of Hg within the peat profile (Canario et al. 2008; Franzen et al. 2004).
Mobilization of Hg from decomposition of peat soil and subsequent re-sorption in the underlying
mineral layers has previously been observed to account for natural internal enrichment within the
soil profile (Franzen et al. 2004). The similarity between the THg and MeHg peat profiles over
the course of the experiment provides support for this mechanism as it may suggest that a
common process is influencing both inorganic Hg and MeHg. The observed changes in the peat
profiles over time are not likely a function of changes in peat bulk density as a result of
decomposition, resulting in elevated inorganic Hg and MeHg concentrations at certain depths.
The peat profile inorganic Hg and MeHg concentrations are similar to the inorganic Hg and
MeHg stock profiles, which account for changes in bulk density (see Figure C-4-1). The trends
in KD coefficients reflect the observed patterns in pore water MeHg and THg concentrations
(reported in Haynes et al., 2017a), which may suggest that changes in pore water inorganic Hg
and MeHg, as opposed to those in the solid phase peat, may be driving the observed partitioning.
Leaching of Hg in association with dissolved organic matter from the solid phase peat in the Low
WT mesocosms was primarily attributed to enhanced decomposition in the aerobic upper depths
(Haynes et al., 2017a), which may also account for the enhanced partitioning of both inorganic
Hg and MeHg in the shallow peat depths under Low WT conditions (Figures 4-6 and C-4-3).
The deepening of soil oxygenation with water table drawdown has been observed to stimulate
microbial decomposition in peat (Bragazza et al. 2016) resulting in enhanced concentrations of
dissolved organic carbon liberated in peatland pore waters (Liu et al. 2016), including labile
carbon sources such as acetate (Yang et al. 2016b). The positive correlation between pore water
acetate concentrations and MeHg concentrations may also indicate that, as a result of aerobic
130
decomposition, Hg may be liberated from the solid phase in association with small, mobile
organic carbon molecules. Leaching of inorganic Hg and MeHg from the peat into the pore
waters in association with dissolved organic carbon has previously been observed to occur in the
Low WT treatment mesocosms (Haynes et al. 2017a). Translocation of Hg within the peat
profile to the depths associated with the lowered water table position may be facilitated by
dissolved organic carbon movement. This mobile inorganic Hg and MeHg may subsequently be
transported by fluctuating water levels within the peat profile and re-sorbed to the solid phase in
the depths coinciding with the greatest variation in water table position; potentially binding to the
reduced sulfur groups in the organic matter (Skyllberg et al. 2000). This mechanism may play an
important role in contributing to the shifting peak peat THg and MeHg concentrations toward the
position of the lowered water table.
The presence of different plant functional types have also been observed to affect dissolved
organic matter in peat pore waters, with vascular plants such as graminoids and Ericaceae shrubs
responsible for the destabilization of organic matter and the facilitation of carbon losses
(Robroek et al. 2016). Aeration of the peat in the rooting zone via oxygen leakage from the
aerenchymous tissues of the sedge vegetation (Greenup et al. 2000; Crow and Wieder 2005;
Waddington et al. 2015) and the lack of competition for available oxygen by Ericaceae shrubs
(Romanowicz et al. 2015) may contribute to peat degradation deeper within the peat profile and
further augment Hg partitioning to pore water (Figures 4-6d and C-4-3d). In addition to
affecting peat decomposition, the removal of different plant functional types has also been
observed to result in a shift of the microbial community composition (Robroek et al. 2015) and
therefore influence such processes as methanogenesis (Hodgkins et al. 2014). Removal of
sedges, and the subsequent loss of a labile source of carbon from root exudation, may limit the
amount of labile carbon available for microbial decomposer communities; thereby affecting peat
decomposition (Robroek et al. 2016) and potential MeHg and inorganic Hg translocation within
the peat profile.
4.5.1 Conclusions and Implications
This study demonstrates that inorganic Hg and MeHg partitioning, translocation and re-sorption
to the solid phase may contribute to the shifts in the depth of peak concentrations as a result of
changes in hydrology and plant community. Inorganic Hg and MeHg will likely be partitioned
131
into pore waters as a result of aerobic peat decomposition above the lowered water table level
and mobilized within the peat in association with dissolved organic carbon. Some of this Hg
may be translocated and re-sorbed to the solid phase peat within the zone of the lowered water
tables. Alterations in peatland water table positions may also affect potential Hg methylation
rate constants, with comparable kmeth values under both Low and High WT conditions in the peat
depths subjected to water table fluctuations. The comparable MeHg production under aerated
conditions to that observed in anoxic peat around the lowered water table depth may be
facilitated by residual pockets of anoxia and the regeneration of terminal electron acceptors such
as sulfate. Root oxygen loss from sedges may also contribute to terminal electron acceptor
regeneration. The availability of a labile carbon source for methylating communities, such as
acetate, liberated from enhanced peat decomposition under aerobic conditions and via sedge root
exudation may also be a contributing factor leading to greater MeHg production. Climate-
induced shifts in water table position and vascular plant functional groups have the potential to
enhance the transport and export of MeHg from peatlands to downstream aquatic ecosystems
(Haynes et al., 2017a), which may harm vulnerable wildlife and human populations which rely
upon aquatic life for sustenance (Mergler et al. 2007; Scheuhammer et al. 2007). Additionally, if
a changing climate favors the establishment of sedges and the removal of Ericaceae shrubs,
inorganic Hg may be delivered to the peat from the atmosphere by the sedge aerenchyma
(Haynes et al., 2017b) resulting in elevated peat THg concentrations, which may be available for
methylation. Combined changes in both water table levels and vascular plant communities have
the potential to significantly affect the cycling of inorganic Hg and MeHg, both speciation and
mobility, within the solid phase peat and the partitioning of Hg species into peatland pore waters.
Further research is required to examine the role of Hg translocation within the profile of peatland
systems, as this mechanism of Hg accumulation is understudied in the literature.
132
4.6 Acknowledgements
We would like to acknowledge the laboratory assistance of P. Huang, K. Ng, R. Co and B.
Perron as well as the field assistance of K. Ng. We thank L. Jamie Lamit for ambient peat
collection. Funding was provided through a Natural Sciences and Engineering Research Council
of Canada (NSERC) Alexander Graham Bell Canada Graduate Scholarship (CGS-Doctoral) to
K.M.H and a NSERC Discovery Grant to C.P.J.M. The PEATcosm experiment was funded by
the USDA Forest Service Northern Research Station Climate Change Program and the National
Science Foundation (DEB-1146149).
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Chapter 5 Contrasting Mercury Isotopic Compositions of Two Sub-boreal
Peatlands
5
5.1 Abstract
Peatland ecosystems store large amounts of mercury (Hg) in organic peat soils, which are
vulnerable to hydrological changes and shifts in vascular plant communities likely to result from
climate change. Stable isotope fractionation of Hg aids in identifying sources and the processes
controlling Hg cycling, but this approach has not been widely applied in peatland ecosystems.
To compare different peatland systems and to examine the potential influence of climate change
impacts on peat Hg cycling, surface (0-10 cm) peat was collected for Hg isotopic analysis from a
peatland in northern Minnesota and a plot-scale climate change experiment (PEATcosm) in
northern Michigan where water table positions and vascular plant functional groups were
manipulated. Based on Δ201
Hg mass balance mixing models, gaseous Hg(0) deposition accounts
for 57 ± 9% of the peat Hg concentrations at the northern Minnesota site, whereas Hg(II)
deposition dominated (74 ± 12%) the Hg stores at the northern Michigan site, potentially as a
result of significant differences in seasonal snow accumulation. Higher Hg(II) contributions are
observed in peat where higher water tables are maintained through greater precipitation and
snowmelt inputs. A strong relationship is observed between mass dependent fractionation
(MDF) and total gaseous Hg fluxes measured at the PEATcosm experiment, but not at the
northern Minnesota site, potentially due to the presence of a tree canopy, which is not present in
the PEATcosm site and which likely affects boundary layer stability. Odd isotope mass
independent fractionation (MIF) of peat at the northern Minnesota site is indicative of
photochemical reduction. In contrast, the PEATcosm peat exhibits significant positive Δ201
Hg
and Δ200
Hg anomalies with only a minimal or no corresponding shift in Δ199
Hg and Δ204
Hg,
respectively, which does not align with any previously identified MIF mechanisms. This study
demonstrates that Hg cycling in peatlands varies by site and isotopic analyses may aid in
identifying how cycling mechanisms may be altered with a changing climate. Further research
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in peatland ecosystems is required to determine the processes responsible for MDF in peat soil,
the isotopic anomalies observed in the PEATcosm peatland and the influence of snow Hg
accumulation processes in peat.
5.2 Introduction
Peatland ecosystems cover approximately 3% of the global landscape and are predominantly
located in subarctic, boreal and temperate regions globally (Gorham 1991). Despite their small
global land area, these ecosystems store between 10% (Gorham 1991) to 30% (Turunen et al.
2002) of the global soil carbon pool. With their unique biogeochemical and hydrological
characteristics these ecosystems play an important role in global climate; acting as sinks of
atmospheric carbon dioxide (CO2) throughout the Holocene epoch via accumulation of immense
quantities of organic matter (Gorham 1991). Accumulation of organic matter in thick peat soils
occurs in peatlands due to a net imbalance between plant net primary production and microbial
decomposition (Clymo 1984; Moore and Basiliko 2006). Hydrological processes that maintain
peatland water tables are significant controls on the delicate balance between carbon
accumulation and carbon losses via the microbial mineralization of organic matter (Whittington
and Price 2006; Waddington et al. 2015).
Climate change threatens to alter the hydrological budget in peatlands in the boreal and sub-
boreal regions (Bridgham et al. 1995) with both changes in precipitation patterns and greater
evapotranspiration expected in association with increasing temperatures (Kunkel et al. 2003;
Groisman et al. 2005; Thomson et al. 2005). Changes in hydrology may further result in a shift
in plant community composition towards vascular-dominated vegetation including graminoids
and Ericaceae shrubs, as compared to Sphagnum moss-dominated systems (Weltzin et al. 2003;
Strack et al. 2006; Dieleman et al. 2015), with potential feedbacks on peatland carbon cycling
(Bridgham et al. 1995). Therefore, the carbon storage capacity of peatlands may be greatly
reduced and ecosystem feedbacks may result in considerable re-release of sequestered carbon to
the atmosphere (Trettin et al. 2006).
In addition to storing immense stocks of carbon, peatland ecosystems also act as strong sinks of
inorganic mercury (Hg), sequestering more Hg than associated upland systems (Kolka et al.
2001; Grigal 2003). Atmospheric deposition is the principal process delivering Hg to terrestrial
ecosystems from local, regional and global sources (Driscoll et al. 2007). Gaseous elemental
143
mercury, Hg(0), may be directly deposited to terrestrial surfaces as dry deposition (Driscoll et al.
2013). Vascular vegetation has also been observed to influence the exchange of gaseous Hg with
the atmosphere via uptake by stomata and accumulation in leaf tissues as well as by sorption to
leaf surfaces (Marsik et al. 2005; Laacouri et al. 2013; Haynes et al. 2017). This accumulated
Hg in vegetation may subsequently be contributed to the soil in litter (Demers et al. 2013).
Atmospheric Hg(0) may also be oxidized to Hg(II) and deposited either in association with
precipitation or directly to soils (Driscoll et al. 2013). Once deposited, peatlands provide an
optimal setting, in terms of hydrology and redox potential, for the process of Hg methylation to
occur (St. Louis et al. 1994; Tjerngren et al. 2012), which is mediated by numerous anaerobic
microbes (Gilmour et al. 2013). Mercury in peatlands may also be reduced to Hg(0) and re-
emitted to the atmosphere (Kyllönen et al. 2012; Fritsche et al. 2014). Solar radiation, soil and
air temperatures, soil moisture and redox potential all influence the dynamic exchange between
the atmosphere and background soils (Edwards et al. 2001; Gustin and Stamenkovic 2005;
Moore and Carpi 2005; Moore and Castro 2012). Processes by which Hg may be reduced
include photochemical reduction (Bergquist and Blum 2007), microbial reduction (Kritee et al.
2007) and natural organic matter reduction (Zheng and Hintelmann 2010), as evidenced by
studies of Hg isotope fractionation. However, the relative influences of deposition and reduction
processes, as well as the controls governing these processes in peatland ecosystems, are not well
understood.
Mercury stable isotope fractionation can be utilized to elucidate the processes controlling the
cycling of Hg through the environment and to attribute sources of Hg at a particular site (Yin et
al. 2010). Mercury has seven stable isotopes: 196
Hg, 198
Hg, 199
Hg, 200
Hg, 201
Hg, 202
Hg and 204
Hg
with a mass difference of up to 4%. Both mass-dependent (MDF) and mass-independent (MIF)
isotope fractionation has been observed during transformations and reactions (Bergquist and
Blum 2009). Mass-dependent fractionation is most often represented by δ202
Hg, while MIF is
most commonly observed for odd-mass-number isotopes (MIFodd, Δ199
Hg and Δ201
Hg)
(Bergquist and Blum 2009). Even mass number MIF (MIFeven, Δ200
Hg and Δ204
Hg) is a less
common isotopic anomaly, but has been observed in atmospheric Hg samples likely due to gas
phase atmospheric processes (Gratz et al. 2010; Sherman et al. 2012; Cai and Chen 2016) and is
reflected in large positive Δ200
Hg precipitation Hg(II) isotopic signatures (Chen et al. 2012).
However, MIFeven has not been similarly observed in terrestrial environments.
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Recent studies have begun to investigate natural abundance stable Hg isotopic signatures in
peatland ecosystems to distinguish Hg sources to these landscapes and to determine the
mechanisms influencing Hg cycling within peatlands, including deposition and re-emission
pathways (Jiskra et al. 2015; Enrico et al. 2016). Jiskra et al. (2015) concluded that the majority
of Hg found in a boreal forest peat soil was the result of litter deposition, with some contribution
from wet deposition via precipitation. Similarly, Enrico et al. (2016) determined in a French
Pyrenees bog that predominantly gaseous elemental Hg, with negative Δ199
Hg and near zero
Δ200
Hg, accumulated via dry deposition in the living Sphagnum moss and recently accumulated
peat. In terms of the controls on soil-air exchange dynamics in peatlands, Jiskra et al. (2015)
attributed significant Hg re-emission from the peat soil to non-photochemical abiotic reduction
of Hg facilitated by natural organic matter reduction, while Enrico et al. (2016) observed
evidence of photochemical re-emission. These two studies take an important first step in
identifying Hg sources and cycling processes in peatland ecosystems using Hg isotope analysis.
Given the vulnerability of peatland ecosystems to climate change, it is important to better
understand the processes governing Hg deposition, reduction and re-emission to the atmosphere
from these systems, particularly across different sites.
The purpose of this study was to compare the Hg isotopic signatures of two sub-boreal
peatlands. Surface peat samples were collected for isotopic analysis from the PEATcosm
(Peatland Experiment at the Houghton Mesocosm Facility) experimental peat monoliths located
in Houghton, Michigan both prior to (2011) and during (2014) the experimental manipulations.
PEATcosm is an experiment wherein water table levels and vascular vegetation communities
were manipulated to simulate potential climate change impacts. Surface peat samples were also
collected from the S1 bog of the Marcell Experimental Forest in north-central Minnesota. The
main objectives of this study were 1) to compare the Hg isotopic signatures of surface peat from
both sites, which may differ because of different amounts of snow during the winter months and
different vascular vegetation canopy structures, 2) to investigate the processes governing peat Hg
cycling, particularly as it relates to the total gaseous Hg fluxes measured at these sites in a
previous study (Haynes et al. 2017) and 3) to discern any treatments effects of the PEATcosm
experimental water table and plant functional group treatments on the Hg isotopic signature of
surface peat.
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5.3 Methods
5.3.1 S1 Peatland Site Description
The S1 peatland is an 8.1 ha ombrotrophic, spruce-Sphagnum bog located within the Marcell
Experimental Forest (MEF) in north-central Minnesota (47.51907° N, 93.45966° W). Typical
annual precipitation for the MEF is approximately 780 mm, with approximately one-third of this
occurring as snow. Mean temperatures at the MEF range from -15°C in January to 19°C in July
(Sebestyen et al. 2011). The overstory is dominated by two tree species: Picea mariana (Mill.)
B.S.P and Larix laricina (Du Roi) K. Koch. The understory consists primarily of Ericaceae
shrubs including Chamaedaphne calyculata, Kalmia polifolia, Vaccinium angustifolium Aiton,
Vaccinium oxycoccos, and Rhododendron groenlandicum as well as sedges Eriophorum spp. and
the lily Maianthemum trifolium (L.) Sloboda. The dominant bryophyte present on hummocks is
Sphagnum magellanicum, while hollows are mainly colonized by S. angustifolium (C.E.O.
Jensen ex Russow).
Surface peat (0-10 cm) from this site was collected and analyzed for isotopic signatures to serve
as an in situ peat contrast to the experimental conditions of the PEATcosm mesocosms. The S1
peatland also receives considerably less snow accumulation than the PEATcosm site located in
the Upper Peninsula of Michigan.
The S1 peatland has now become the site of the SPRUCE (Spruce and Peatland Responses
Under Climatic and Environmental Change) experiment. The SPRUCE study is a climate
change field experiment that involves a ten year manipulation of both soil and air temperature
(up to 9°C above ambient temperatures in a regression design), and atmospheric carbon dioxide
(CO2; elevated to approx. 800 ppm) concentrations within open-top, approximately 12 m
diameter enclosures along three transects within the S1 bog (Hanson et al. 2017; see Figure D-5-
1a for photo of a SPRUCE plot). This study describes the isotopic signatures of surface peat
samples collected in May 2014 prior to the initiation of any manipulations.
5.3.2 PEATcosm Site Description
The PEATcosm Mesocosm Facility is located in Houghton, Michigan, USA (47.11469° N,
88.54787° W) at the United States Department of Agriculture (USDA) Forest Service Northern
Research Station - Forestry Sciences Laboratory. Typical annual precipitation at this site is
146
approximately 870 mm, with approximately 50% of this precipitation in the form of snow. As
compared to the approximately 275 mm of snow water equivalent received at the S1 peatland,
the PEATcosm site receives more than 1.5 times the amount of snow water equivalent, with an
annual average of 435 mm. Mean temperatures in this area range from -13°C in January to 24°C
in July (Potvin et al. 2015). This facility housed twenty-four intact 1 m3 (1 m x 1 m x 1 m) peat
monoliths that were harvested from an ombrotrophic peatland in Meadowlands, Minnesota, USA
in May 2010 and placed in individual Teflon-coated stainless steel mesocosm bins. The bins
were open on top, exposing the peat to ambient climate conditions and inserted into a climate-
controlled tunnel allowing belowground access to one face of each of the bins. The mesocosms
were insulated on the sides which facilitated the maintenance of a natural vertical temperature
gradient to simulate that of a natural peatland (see Figure D-5-1b for photos of the Facility).
Potvin et al. (2015) provides a comprehensive and detailed explanation of the peat harvest and
experimental set-up. The PEATcosm experiment concluded in July 2015 with the destructive
harvest of each of the monoliths.
The PEATcosm experiment was designed to assess the potential effects of anticipated shifts due
to global climate change in vascular plant communities and water table regimes on numerous
aspects of peatland ecosystem functioning, including Hg cycling. This study was a full-factorial
experimental design with two water table (WT) prescriptions crossed with three vascular plant
functional group treatments in a randomized complete block design. There were four replicate
mesocosms per treatment combination for a total of 24 experimental units. The water table
treatments were based on long-term (approximately 50 years) data from the Marcell
Experimental Forest in north-central Minnesota (47.51907° N, 93.45966° W), located north of
the peat monolith harvest site. The two target water table treatment profiles were modelled after
1) typical variability, average water table position (referred to as ‘High WT’) and 2) comparably
high variability, low water table position (referred to as ‘Low WT’). The mean difference
between the High and Low WT positions was approximately 20 cm throughout the experiment.
These target water table profiles were maintained via a combination of artificial precipitation
additions, rain-exclusion covers and regulated outflow during the spring months from
approximately the acrotelm-catotelm boundary from each of the bins. The three vascular plant
functional group treatments simulated anticipated climate change-induced community
composition (Chapin et al. 1996; Weltzin et al. 2000; Strack et al. 2006). The treatments
147
included 1) all Ericaceae removed (referred to as ‘sedge only’ treatment), 2) all sedge removed
(‘Ericaceae only’ treatment) and 3) both sedge and Ericaceae present (‘unmanipulated control’
treatment). The plant functional group treatments were initiated in June 2011 and maintained
weekly by the clipping of stems of the excluded species. The dominant sedge species present
was Carex oligosperma Michx., while the dominant ericaceous shrubs included Chamaedaphne
calyculata (L.) Moench., Kalmia polifolia Wangenh., and Vaccinium oxycoccos L.. The mosses
Sphagnum rubellum Wilson, S. magellanicum Brid., S. fuscum (Schimp.) Klinggr and
Polytrichum strictum Brid. comprised the dominant bryophyte species in all of the 24
mesocosms. To a lesser extent, P. commune Hedw., Eriophorum vaginatum L., Andromeda
polifolia L. var. glaucophylla (Link) DC., Rhododendron groenlandicum (Oeder) Kron and Judd,
and Drosera rotundifolia L. were also present.
5.3.3 Peat Sampling and THg Analysis
In June 2014 two peat cores were collected from hollow microforms within each enclosure plot
in the S1 bog. Surface peat (0-10 cm only for this study) was collected using a serrated knife,
which was rinsed with deionized water between samples. Peat from the two cores was combined
in a large sealable plastic bag and homogenized by hand. Samples were taken from the bag
using clean, gloved hands. The samples were placed on ice and frozen following collection.
At the PEATcosm study, peat analyzed for isotopes was collected from each of the 24 mesocosm
bins in 2011 prior to the initiation of the experimental treatments and in 2014 during the
experimental manipulations. Cores were collected using a stainless steel corer from which the
peat was extruded. The corer was rinsed with deionized water between each mesocosm. Peat
samples were immediately frozen following sub-sectioning. As one objective of this study is to
relate trends in total gaseous Hg (TGM) fluxes to Hg isotopic signatures, only the surface 0-10
cm samples were analyzed for isotopes. Due to low bulk density and relatively low THg
concentrations (mean ± standard deviation [THg] = 31 ± 7 ng g-1
) of these surface peat samples
as well as the limited amount of material available within the overall PEATcosm framework, the
four replicate peat samples for each of the six crossed water table and plant functional group
treatments were pooled to ensure enough Hg for isotopic analytical detection. To identify the
potential influence of the PEATcosm treatments on the Hg isotopic signatures of the peat, only
three of the 2011 pre-treatment samples from both water table treatments and two of the three
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plant functional groups (Low WT sedge only, Low WT Ericaceae only and High WT sedge only)
were analyzed for isotopes. Comparing the pre-treatment to the treatment peat Hg signatures
also facilitated the assessment of the potential impact of the pre-treatment climate in the harvest
site of Meadowlands, MN to the experimental site climate in Houghton, MI.
All peat samples were freeze-dried and homogenized prior to digestion in hot nitric acid.
Digestates were diluted and bromine monochloride (BrCl) was added to oxidize all Hg in the
samples overnight prior to analysis. Total Hg analysis was performed using a Tekran 2600 Total
Mercury Analyzer by cold vapor atomic fluorescence spectroscopy (CVAFS) according to US
EPA Method 1631. Recovery of a THg spike was 104 ± 1 % (mean ± standard deviation, n=4),
replication of duplicates was 1.2 ± 0.1 % (n=4) and the detection limit, calculated as three
standard deviations of matrix blanks, was 0.01 ng g-1
(n=4). Recovery of standard reference
material (MESS-3) following digestion was 99 ± 3% (n=4).
5.3.4 Total Gaseous Mercury Flux Measurements
Ambient TGM fluxes were measured using Teflon dynamic flux chambers (DFCs) placed on the
peat surface and connected to a Tekran 2537A Gaseous Mercury Analyzer at both the
PEATcosm and S1 peatland sites. Full details of the measurement procedures and equipment are
provided in Haynes et al. (2017). Briefly, subsets of the total number of experimental treatments
in each experiment were monitored for TGM fluxes over individual 24-hour measurement
periods. In order to discern potential treatment effects, similar ambient conditions were required.
At PEATcosm, TGM fluxes were monitored on 8 of the 24 mesocosm bins in July 2014 at the
peak of the growing season. These measurements were performed approximately one month
prior to the peat sampling. Duplicate mesocosms of the following experimental treatment
pairings were monitored for TGM fluxes: High WT – control vegetation, Low WT – control
vegetation, Low WT – Ericaceae only, and Low WT – sedge only. These treatments span the
range of vascular plant functional group treatments with the Low WT prescription representing
the greatest anticipated hydrological changes in peatlands due to climate change. The High WT
– control vegetation treatment acted as the unmanipulated, control scenario.
At the S1 peatland twelve replicate TGM flux measurements were conducted in May-June 2014.
These fluxes were measured within the footprints of six experimental enclosures installed for the
SPRUCE experiment. Due to the considerable distance between plots and equipment
149
availability, two replicate locations within each experimental plot were measured simultaneously
during 24-hour measurement periods. These duplicate flux measurements within each plot were
averaged to relate TGM flux to the peat isotopic signatures.
5.3.5 Sample Preparation and Mercury Isotope Analysis
Total Hg in the peat samples was trapped in a 0.2% potassium permanganate (KMnO4) (w/w) in
5% sulfuric acid (H2SO4) (v/v) trapping solution using a thermal combustion method with a
modified Teledyne Leeman Labs Hydra IIc Direct Mercury Analyzer (Zheng et al. 2016). Once
trapped, an aliquot of the trapping solution was neutralized with hydroxylamine hydrochloride
(NH2OH∙HCl) and analyzed by CVAFS to determine the Hg concentration in each trap. Mean
trap recovery for all peat samples was 95 ± 14% (mean ± standard deviation). Method blanks
had an average concentration of 0.005 ng g-1
(range 0.0044 to 0.0049 ng g-1
, n=2), which
represented on average less than 0.5% of the concentration of peat sample traps. Standard
reference materials CCRMP TILL-1 (Ontario soils), NIST 1944 (New York/New Jersey
waterway sediment) and NIST 1646a (estuarine sediment) as well as the procedural standard
NIST SRM 3133 were also combusted in order to ensure the performance of the trapping method
as well as the accuracy and precision of isotopic measurements. Trap recovery of CCRMP
TILL-1 was 86 ± 5% (n=9), NIST 1944 was 107 ± 24% (n=2), NIST 1646a was 80% (n=1) and
the procedural standard NIST SRM 3133 was 91 ± 3% (n=7).
Mercury isotopic composition was determined by cold vapor multi-collector inductively coupled
plasma mass spectrometry (CV-MC-ICP-MS, Neptune Plus, Thermo Finnigan). Trap solutions
were neutralized with NH2OH∙HCl to reduce the KMnO4, and then diluted to 1 ng g-1
using a
pre-neutralized 0.2% KMnO4 solution. The diluted trap solutions were introduced into the
instrument using stannous chloride reduction and Hg(0) vapor separation. Instrumental mass
bias was corrected using an internal Tl standard (NIST 997), introduced as a desolvated aerosol.
Standard-sample-standard bracketing was conducted with NIST 3133 Hg standard. The
bracketing standard was prepared using the same matrix solution as samples and the
concentration was matched to samples such that the difference in signal intensities was within
10%. Isobaric interference from 204
Pb was corrected by measuring 206
Pb, but was consistently
negligible. On-peak zero corrections were applied to all Hg and Pb masses. Mercury isotope
compositions are reported using δ notation defined by the following equation:
150
𝛿𝑥𝑥𝑥𝐻𝑔 (‰) = [( 𝐻𝑔/ 𝐻𝑔198𝑥𝑥𝑥 )
𝑠𝑎𝑚𝑝𝑙𝑒
( 𝐻𝑔/ 𝐻𝑔198𝑥𝑥𝑥 )𝑁𝐼𝑆𝑇3133
− 1] × 1000 (5.1)
where x is the mass number of each Hg isotope from 199
Hg to 204
Hg. We used δ202
Hg to assess
MDF. Mass independent fractionation is defined as the difference between measured δxxx
Hg
values and the theoretical values determined based on kinetic mass dependent fractionation law
and is reported as Δxxx
Hg (Blum and Bergquist, 2007):
Δxx𝑥𝐻𝑔 = 𝛿xx𝑥𝐻𝑔 – 𝛽 × 𝛿202𝐻𝑔 (5.2)
where x is the mass number of each Hg isotope 199, 200, 201 and 204. β is a scaling constant to
estimate the theoretical kinetic MDF, which is 0.2520, 0.5024, 0.7520 and 1.4930 for 199
Hg,
200Hg,
201Hg and
204Hg, respectively. Each sample was measured twice and an in-house
secondary Hg standard (JTBaker) was measured at least 4 times in each analytical session to
monitor the performance of the instrument. The analytical uncertainties are reported as 2SD of
the JTBaker standard. The 2SD (n=13) for JTBaker was 0.05‰ for δ202
Hg, 0.04‰ for Δ199
Hg,
0.04‰ for Δ201
Hg, 0.02‰ for Δ200
Hg and 0.07‰ for Δ204
Hg (Table D-5-1).
5.3.6 Data Analyses
All statistical analyses were performed using R statistical software (R Development Core Team,
2014) with α = 0.05. All isotope data for all peat samples analyzed are provided in Table D-5-2.
Regression analyses using Pearson correlation coefficients were performed on the relationships
between δ202
Hg and Δ199
Hg, Δ199
Hg and Δ201
Hg, Δ200
Hg and Δ201
Hg, and Δ200
Hg and Δ204
Hg for
both the S1 and PEATcosm peat individually. For both peatland sites 24-hr TGM flux
measurements from a previous study (Haynes et al. 2017) were related to peat MDF signatures
(δ202
Hg). The relationships between the surface 0-10 cm peat THg concentrations and isotopic
values (Δ199
Hg, Δ200
Hg and Δ201
Hg) were also examined for both peatlands. As the surface
PEATcosm peat samples had to be pooled from all of the four replicate mesocosms for each of
the six treatment combinations in order to achieve enough Hg for isotopic detection, only a
qualitative assessment of the combined water table and plant functional group treatment effects
on isotopic signatures was performed.
The fraction of Hg(II) and Hg(0) contributions to the peat of both the S1 and PEATcosm sites
were estimated using Δ199
Hg, Δ201
Hg and Δ200
Hg values in a binary mixing model, with Hg(II)
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and Hg(0) as the two end members, similar to the approach used by Enrico et al. (2016). The
isotopic values of the end members were selected based on the mean (± standard deviation) of
literature values and the appropriate post-depositional offsets were applied similar to the methods
of Zheng et al. (2016). The Δ199
Hg, Δ200
Hg and Δ201
Hg values for the Hg(0) and Hg(II) end
members are presented in Table D-5-3. To estimate the uncertainty associated with the
calculated Hg(II) and Hg(0) contributions, a Monte Carlo simulation was conducted with 1000
iterations, based on the uncertainties (standard deviation) in the selected end members, for each
isotopic value (Δ199
Hg, Δ200
Hg and Δ201
Hg) of each peat sample. The number of iterations was
determined to be sufficient to provide a stable solution with an uncertainty of approximately 1%
of the mean and has been applied in previous studies using Monte Carlo simulations (e.g. Zheng
et al. 2016).
5.4 Results and Discussion
5.4.1 S1 Peat Isotopic Composition
The Hg isotopic signatures of surface peat (top 0-10 cm) collected from the S1 bog had
significantly negative odd isotope anomalies (Δ199
Hg = -0.13 ± 0.10‰, Δ201
Hg = -0.15 ± 0.07‰;
mean ± standard deviation), positive Δ200
Hg anomalies (+0.07 ± 0.03‰) and near zero Δ204
Hg
values (+0.01 ± 0.02‰; see Table D-5-2). These values are similar to those recorded in living
Sphagnum moss (Δ199
Hg = -0.11 ± 0.09‰, Δ200
Hg = +0.03 ± 0.02‰, 1σ) and in recently
accumulated peat (Δ199
Hg = -0.22 ± 0.06‰, Δ200
Hg = 0.00 ± 0.04‰, 1σ) collected in a French
Pyrenees peat bog (Enrico et al. 2016). In contrast to that peatland site, wherein the majority
(~79%) of the Hg present was sourced from Hg(0) deposition, the S1 peat isotopic values
suggest that Hg(0) contributions from the atmosphere are approximately equal to Hg(II)
contributions to surface peat Hg concentrations. The binary mixing models suggest that Hg(0)
contributions range from 46 ± 9% based on the conservative Δ200
Hg values, up to 57 ± 9% based
on Δ201
Hg data (Table 5-1 for Δ201
Hg model, Tables D-5-4 and D-5-5 for Δ199
Hg and Δ200
Hg
models, respectively). One limitation to the model results may be the values selected to
represent the Hg(0) and Hg(II) end members. The isotopic signatures of Hg(0) and Hg(II) at
each of the two peatland sites were not measured, but rather synthesized from literature values
collected at multiple disparate sites. Therefore, given the range of values observed for Hg(0) and
Hg(II) in other studies, the mean of values reported in the literature may not be representative of
152
the isotopic signatures contributing to this peatland. Potential shifts in the isotopic values due to
post-depositional processes were accounted for by adding the associated offsets to the mean
literature values as has been performed in studies in forested environments (Zheng et al. 2016).
However, despite this caveat we believe that the model results indicating nearly equal Hg(0) and
Hg(II) contributions, with increasing Hg(0) contributions in some samples, is valid given the
trends in the isotopic data.
Table 5-1 Contributions of Hg(II) (fHg(II)) and Hg(0) (fHg(0)) to peat isotopic signatures for the
PEATcosm 2014 treatment peat and S1 bog 2014 peat. Results are the mean of 1000 iterations
of the isotopic mixing model with Monte Carlo simulation based on the Δ201
Hg isotopic values.
Δ201
Hg fHg(II) fHg(0)= 1 – fHg(II)
Mean Standard Deviation Mean Standard Deviation
PEATcosm Experimental Peat - 2014
Low WT – Sedge 0.64 0.12 0.36 0.12
Low WT – Ericaceae 0.59 0.12 0.41 0.12
Low WT – Control 0.69 0.13 0.31 0.13
High WT – Sedge 0.80 0.15 0.20 0.15
High WT – Ericaceae 0.82 0.16 0.18 0.16
High WT - Control 0.91 0.17 0.09 0.17
S1 Peat – 2014
Plot 4 0.34 0.08 0.66 0.08
Plot 6 0.31 0.08 0.69 0.08
Plot 10 0.48 0.10 0.52 0.10
Plot 13 0.52 0.10 0.48 0.10
Plot 17 0.50 0.10 0.50 0.10
Plot 19 0.42 0.09 0.58 0.09
Significant negative MDF anomalies, represented by the δ202
Hg values, were also observed in the
peat, ranging from -1.78 to -1.36‰. These values are higher than the mean δ202
Hg values of -
2.4‰ for Sphagnum moss observed by Enrico et al. (2016). Negative shifts in MDF have been
observed in pine needles, heather leaves and lichens in peat bogs, with δ202
Hg values ranging
from -2.3 to -2.8‰ (Enrico et al. 2016). Jiskra et al. (2015) also measured negative MDF values
153
(-2.59 ± 0.25‰) in boreal litter samples. Previous research in forest and bog ecosystems has
attributed this negative shift in δ202
Hg to MDF during stomatal uptake of Hg into the foliage of
trees and vascular vegetation and delivered to the soil via litterfall (Demers et al. 2013; Enrico et
al. 2016; Zheng et al. 2016). Some contribution of litter deposition from the vascular vegetation
to the S1 peat surface may occur but litterfall is not likely to be a dominant pathway for Hg
deposition to the peat. The higher δ202
Hg values recorded in the S1 peat as compared to those
observed in forest ecosystems may suggest an alternative mechanism is responsible for this
MDF. As Sphagnum mosses do not possess stomata it may be speculated that direct deposition
of Hg(0) via sorption to the moss surface may be the dominant source of Hg(0) in the surface
peat. Although MDF anomalies during direct Hg(0) deposition and oxidation in soil are
unconstrained, it has been suggested that a positive shift in MDF may occur due to preferential
deposition of heavier isotopes (Zheng et al. 2016). Therefore, the S1 δ202
Hg values likely
encompass both direct adsorption of Hg(0) to the Sphagnum peat (positive MDF) as well as
likely minor contributions from litter deposition of vascular vegetation, which accumulates Hg
via stomatal uptake (negative MDF). However, this remains a hypothesis as the mechanisms
governing MDF in Sphagnum peat soils with vascular vegetation cover have not been
experimentally determined in the literature.
The relationships between the isotopic signatures and the THg concentrations in the peat are also
indicative of Hg(0) contributions at the S1 site. A strong, significant, negative correlation was
observed between S1 peat THg concentrations and Δ201
Hg (r = - 0.94, p = 0.005; Figure 5-1b).
A similar trend was observed between THg concentrations and Δ199
Hg (r = - 0.90, p = 0.01),
with more negative MIFodd related to greater THg concentrations. There was no similar
relationship between the peat THg concentrations and Δ200
Hg (p = 0.24; Figure 5-1a) suggesting
that the S1 surface peat THg concentrations are not strongly reflective of Hg(II) isotopic
signatures.
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Figure 5-1 Relationship between Δ200
Hg, Δ201
Hg and the surface peat THg concentrations (in ng
g-1
) for the S1 peatland and PEATcosm sites. Note different scale for Δ201
Hg for S1 bog and
PEATcosm data.
Collectively this may provide support for the increased contributions from Hg(0), rather than
Hg(II) deposition to the surface peat of the S1 peatland wherein higher THg concentrations were
observed. When investigating isotopic data in relation to peat TGM fluxes at the S1 site, there
was no relationship with MDF (p = 0.36; Figure 5-2a). Similarly, the relationship between TGM
fluxes from the peat and Δ201
Hg (p = 0.11) was not significant, while the relationship with
Δ199
Hg (r = - 0.79, p = 0.06) was marginally significant.
155
Figure 5-2 Relationship between δ202
Hg values (in ‰) and the mean daily TGM fluxes (in ng m-
2 d
-1) for the a) S1 and b) PEATcosm peatlands. Note different δ
202Hg and mean daily TGM flux
scales for both plots.
Total gaseous Hg exchange between soils and the atmosphere is a dynamic process influenced by
such factors as soil and air temperature, soil moisture, solar radiation and atmospheric Hg
concentrations over short timeframes (Edwards et al. 2001; Gustin and Stamenkovic 2005;
Moore and Carpi 2005; Moore and Castro 2012). Exchange of Hg(0) between standing vascular
vegetation via both stomatal and non-stomatal pathways including surface sorption may also
occur on short timescales affecting the magnitude and direction of TGM fluxes (Rea et al. 2002;
Stamenkovic and Gustin 2009). Gaseous Hg(0) is more likely to be quickly re-emitted to the
156
atmosphere following deposition due to its minimal solubility in water, but may also be
accumulated in leaf tissue through stomata where it may be oxidized to soluble Hg(II) (Hanson
et al. 1995; Rea et al. 2000). As litterfall contributions to peat are limited by slow rates of net
primary production of aboveground biomass (Frolking et al. 2001), gaseous Hg(0) exchange with
vascular vegetation may not be evident in the solid phase peat isotopic composition. Therefore,
as the isotopic signature of the peat is integrated over a longer time period, short term, dynamic
exchange in TGM fluxes dominated by Hg(0) may not be reflected in the isotopic data.
5.4.2 S1 Peat MIF
The mechanism for MIFodd in the surface 0-10 cm of peat from the S1 bog is likely
photochemical reduction. There was a strong, significantly positive linear relationship (r = 0.95,
p = 0.004) between Δ199
Hg and Δ201
Hg for the S1 peat with a slope of 1.35 ± 0.22 (Figure 5-3)
and no relationship in MIFeven (Figure D-5-2). The slope of the MIFodd relationship for the S1
peat was higher than the slope of 1.0 observed in aqueous solutions as a result of photochemical
reduction in the controlled experiments conducted by Bergquist and Blum (2007). This
relationship is also steeper for the S1 peat as compared to those of other peat soils (Jiskra et al.
2015; Enrico et al 2016) and those of other soils and plants (Demers et al. 2013; Zheng et al.
2016), which observed Δ199
Hg – Δ201
Hg slopes close to 1.0. A curved relationship between
Δ199
Hg and δ202
Hg was observed in the S1 peat (r = 0.87, p < 0.05 for second-order polynomial
fit), which is indicative of the influence of photochemical reduction (Bergquist and Blum 2007).
The potential source of this significant MIFodd may be a combination of both atmospheric
processes before Hg is deposited to the peat and mechanisms acting on the cycling of Hg in the
peat itself. Photo-oxidation of Hg(0) in the gas phase has previously been observed to result in a
slope of 1.6 to 1.9 for the relationship between Δ199
Hg and Δ201
Hg (Sun et al. 2016). This may
be occurring in the atmosphere before the Hg is deposited to the peat and is therefore reflected in
the peat isotopic signature. In addition to photo-oxidation of Hg(0) prior to deposition, abiotic
non-photochemical reduction of Hg(II) in the presence of organic matter, which produces a slope
of 1.5 to 1.6 (Zheng and Hintelmann 2010), may also contribute to the higher than expected
slope than would be observed due to photochemical reduction alone. It is reasonable that the
organic matter-rich peat may influence the reduction of Hg(II). Without direct experimental
evidence under controlled conditions however, the mechanisms governing MIFodd in addition to
photochemical reduction cannot be conclusively stated in this study.
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Figure 5-3 Relationship between a) Δ199
Hg and Δ201
Hg and b) Δ199
Hg and δ202
Hg (all expressed
in ‰) for the surface 0-10 cm peat from both the S1 and PEATcosm peatland sites.
5.4.3 PEATcosm Peat Isotopic Signature
In contrast to the peat from the S1 bog, the control PEATcosm peat collected in 2011 prior to
experimental manipulations had near zero Δ199
Hg values (0.01 ± 0.02‰) and variable Δ201
Hg
values ranging from -0.04 to +0.23‰. Significant positive Δ200
Hg anomalies were also observed
in these samples (0.20 ± 0.14‰; Table D-5-2). The peat collected in 2014 from the PEATcosm
mesocosms, subjected to water table manipulations and altered plant communities, had higher
positive odd-isotope anomalies (Δ199
Hg = 0.10 ± 0.04‰, Δ201
Hg = 0.11 ± 0.10; mean ± standard
deviation) than both the pre-treatment peat and that from the S1 site. Similar to the S1 peat, the
PEATcosm peat had positive Δ200
Hg values (0.07 ± 0.02‰) with near-zero Δ204
Hg anomalies
(0.01 ± 0.03‰). Positive odd-mass isotopes have been previously observed in peat (Shi et al.
2011). Shi et al. (2011) observed enriched odd-mass isotopes in peat collected from a bog in
northern China, which was attributed to increasing anthropogenic Hg pollution with
industrialization. Similar to the S1 peat, negative MDF in the range of -1.47 to -1.04‰ was
observed in the PEATcosm peat. This again may be the result of direct deposition of Hg to the
peat via surface sorption as well as some deposition of vascular leaf litter in which Hg is
accumulated via stomatal uptake, which indirectly incorporates Hg into the peat (Zheng et al.
2016).
With MIFodd anomalies similar to that of Hg(II) in atmospheric samples (Gratz et al. 2010;
Sherman et al. 2012; Demers et al. 2013), the PEATcosm isotopic signatures suggest that the Hg
in the peat was predominantly the result of Hg(II) deposition. The mixing model using odd-
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mass-number isotopes suggests that between 68-82% (74 ± 5%, mean ± standard deviation,
based on Δ199
Hg; Table D-5-4) to 59-91% (74 ± 12%, based on Δ201
Hg; Table 5-1) of the Hg in
the surface 10 cm of the PEATcosm peat was deposited as Hg(II). The fraction of Hg(II)
deposition is considerably lower when using the Δ200
Hg data, with only 53 ± 9% of Hg(II)
contributing to the peat Hg concentrations (Table D-5-5). The reason for this discrepancy in
mixing model results using the different isotopic values is unclear. Considerably more Hg(II)
deposition was reflected in the surface PEATcosm peat as compared to that collected from the
S1 bog based on the MIFodd model results. One potential factor contributing to this difference in
Hg(II) deposition may be the amount of snow accumulation during the winter months at these
two sites. Approximately 70% more snow falls at the PEATcosm site in Houghton, Michigan
(snow-water equivalent (SWE) of 435 mm per year) versus the Marcell Experimental Forest in
northern Minnesota (SWE of 260 mm per year) (Potvin et al. 2015; Sebestyen et al. 2011). Due
to the overstory tree canopy of spruce and tamarack within the S1 bog, the amount of snow
accumulation on the peatland is likely further diminished by canopy interception (Hedmond and
Pomeroy 1998; Sebestyen et al. 2011). Chen et al. (2012) observed predominantly positive
values of both Δ199
Hg (-0.29 to +0.73) and Δ200
Hg (+0.31 to +1.24) in non-Arctic snow collected
immediately after deposition. It was suggested that photochemically-driven Hg(0) oxidation
facilitated by snow crystals in the tropopause may account for the positive anomalies in Δ200
Hg,
while photoreduction on snow crystals may result in odd isotope MIF (Chen et al. 2012).
Lalonde et al. (2001) also observed photo-oxidation of Hg(0) in water and snow prior to
deposition. Ambient temperature may also influence the magnitude of Δ200
Hg anomalies, as
Chen et al. (2012) observed higher Δ200
Hg values in the winter months. Therefore, the positive
Δ200
Hg, Δ199
Hg and Δ201
Hg isotopic anomalies in the PEATcosm peat may be the result of
considerably greater snow accumulation that occurs at this site during the winter months as
compared to that received by the S1 peatland. Differences in snow accumulation between the
pre-treatment site of Meadowlands, MN (located southeast of the Marcell Experimental Forest)
and the experimental site of Houghton, MI may also account for the observed shift in the peat
isotopic signature between the pre-treatment (2011) and experimental (2014) peat.
Dominant contributions of oxidized Hg(II) to the PEATcosm peat was also supported by the
relationships among Δ201
Hg, Δ200
Hg and THg concentrations. Significant negative relationships
were observed not only between Δ201
Hg and peat THg concentrations (r = - 0.90, p = 0.02)
159
similar to the S1 peat, but also Δ200
Hg and surface peat THg (r = - 0.83, p = 0.02; Figure 5-1).
These relationships may suggest that the PEATcosm mesocosms each received similar amounts
of Hg(II), which supports the idea that snow accumulation may significantly influence the
isotopic signature of the surface peat. Smaller Δ201
Hg and Δ200
Hg anomalies with higher peat
THg concentrations may suggest the influence of enhanced Hg(0) contributions across the
treatments. Differences in Hg(0) deposition across the treatments may be reflected in a dilution
of the Hg(II) deposition signal. These influences were also reflected in the strong relationship
between MDF and TGM fluxes measured from the PEATcosm mesocosms (Figure 5-2b). The
mesocosms in which strong depositional fluxes were observed yielded the highest peat MDF
signatures, supporting the idea of direct Hg deposition via sorption to the surface peat. The
strong influence of Hg(II) deposition at the PEATcosm site as compared to the S1 peatland may
contribute to the observed differences in the relationship between fluxes and MDF at the two
sites (Figure 5-2). Previous research in forested systems suggested that Hg(II) may be more
readily available to be reduced upon deposition and subsequently re-emitted quickly to the
atmosphere (Graydon et al. 2009; Graydon et al. 2012). The control vegetation mesocosms
under both water table prescriptions which had the highest modelled Hg(II) contributions also
had the highest TGM emissions. Therefore, re-volatilization of Hg(II) deposited to the
PEATcosm peat surface may also occur quickly in peatlands and contribute to enhanced
emissions. Finally, the difference in the structure of the overstory at the two sites may also affect
the deposition patterns due to the stability of the atmospheric boundary layer above the peat
surface (Kyllönen et al. 2012; Mazur et al. 2014). The spruce and tamarack overstory tree
canopy above the vascular shrub cover at the S1 peatland may facilitate enhanced Hg(0)
deposition due to increased boundary layer stability as compared to the short Ericaceae shrub
cover and lack of overstory tree cover in the PEATcosm mesocosms. Further study under
controlled conditions is required to identify the role these factors may play in controlling Hg
cycling and the subsequent isotopic fractionation pattern.
5.4.4 PEATcosm Treatment Effects – Water Table and Plant Functional Groups
Effects of the PEATcosm water table and plant functional group treatments were reflected in the
isotopic signatures of the surface peat. The highest MIFodd anomalies and therefore the highest
modelled contributions of Hg(II) were observed in the High WT treatment peat, which received
160
the most precipitation to maintain the shallow water table positions (Figure 5-4). During the
spring months following the melting of the snowpack, runoff from the mesocosms was drained
to re-establish the treatment water table positions. Considerably more snowmelt runoff was
released from the Low WT treatments whereas less snowmelt was drained from the High WT
bins to maintain the relatively high water table position. Therefore, this distinction in Hg(II)
contributions between the water table treatments provides further support for the role of snow
and snowmelt water as the source of Hg(II) in the PEATcosm peat.
Figure 5-4 Δ199
Hg and Δ201
Hg values (in ‰) among the six crossed water table and plant
functional group treatments of the PEATcosm peat monoliths. Analyses were on pooled samples
from replicate treatments, thus error bars represent analytical precision, not sample replication
variability.
Under saturated conditions diffusion of Hg through soil pores is suppressed (Bahlmann et al.
2004; Gustin and Stamenkovic 2005; Song and Van Heyst 2005) and therefore gaseous Hg(0)
deposition into peat pores may be reduced. However, under lowered water tables, greater
gaseous Hg(0) may enter aerated peat pores where it may be oxidized and retained in the peat,
resulting in higher peat THg concentrations (Figure 5-1). Greater Hg(II) deposition in the High
161
WT treatments may be susceptible to reduction and re-emission from the peat (Graydon et al.
2009; Graydon et al. 2012). Mercury sorbed to the surfaces of leaves, particularly those of
vascular shrubs in the Ericaceae only and Control vegetation treatments, has been observed to
influence TGM fluxes (Haynes et al. 2017). Photo-reduction of carboxylic-bound Hg(II), which
is typically the dominant species on leaf surfaces, has been demonstrated to produce positive
MIF in the residual Hg(II) pool (Bergquist and Blum 2007; Zheng and Hintelmann 2009; Zheng
and Hintelmann 2010), which may contribute to the observed positive isotopic anomalies.
Smaller positive Δ199
Hg and Δ201
Hg anomalies for the Low WT treatments may also be due to
enhanced stomatal conductance of Hg(0) by vascular vegetation under drier conditions due to
increased productivity of the PEATcosm treatment aboveground biomass (Potvin et al. 2015),
which has been observed in forest soils under wetter conditions (Zheng et al. 2016).
5.4.5 PEATcosm Peat MIF
Unlike the clear relationship between Δ199
Hg and Δ201
Hg for the S1 peat, there was no similar
relationship for the experimental PEATcosm peat (p = 0.17; Figure 5-3). There was also no
significant relationship between Δ199
Hg and δ202
Hg (Figure 5-3). This trend of a significant
Δ201
Hg anomaly with no corresponding Δ199
Hg shift or MDF does not resemble any previously
constrained process such as abiotic non-photochemical Hg(II) reduction in the presence of
organic matter (Zheng and Hintelmann 2010), aqueous-phase photochemical reduction of Hg(II)
and photochemical degradation of methylmercury (Bergquist and Blum 2007), all of which may
be expected to play an important role in peatland Hg cycling (Jiskra et al. 2015). Interestingly,
there was a significant positive relationship between Δ201
Hg and Δ200
Hg for the PEATcosm peat
(r = 0.93, p = 0.01; Figure 5-5), while no similar relationship was observed for the S1 peat (p =
0.32).
162
Figure 5-5 Relationship between Δ200
Hg and Δ201
Hg (expressed in ‰) for the surface 0-10 cm
peat from both the S1 and PEATcosm peatland sites.
A positive relationship between Δ201
Hg and Δ200
Hg is also evident when collectively looking at
Hg(II) in precipitation collected by Gratz et al. (2010), Sherman et al. (2012) and Demers et al.
(2013), with the same slope of 0.18 (Figure D-5-3). This provides support to the binary mixing
model results demonstrating that the Hg in the surface PEATcosm peat is predominantly Hg(II),
with minimal Hg(0) contributions. However, no MIFeven relationship between Δ200
Hg and
Δ204
Hg was observed (Figure D-5-2), which is typically associated with atmospheric gas phase
processes (Gratz et al. 2010; Chen et al. 2012; Cai and Chen 2016) and has been documented in
forest soils (Zheng et al. 2016). To the best of our knowledge, this is the first documented case
of significant Δ201
Hg MIF with only a slight response in Δ199
Hg as well as a significant Δ200
Hg
anomaly without a corresponding shift in Δ204
Hg. As deposition to the PEATcosm peat was
occurring predominantly in the form of Hg(II), it is likely that the process producing this
fractionation pattern is associated with Hg(II). However, the mechanisms responsible for such
trends in the PEATcosm peat samples are not clear and require further investigation to assess the
processes that fractionate Hg isotopes in this manner.
5.4.6 Conclusions and Implications
This study demonstrates the contrasting Hg isotopic signatures in peat from two sub-boreal
peatlands. The observed peat odd- and even-mass isotopic anomalies reflect different Hg
sources as well as different processes, which act to influence Hg cycling in these peatlands.
163
Significant disparities in the fraction of Hg(0) and Hg(II) contributions were observed between
the S1 peatland in northern Minnesota and the PEATcosm peat in the Upper Peninsula of
Michigan. Differences in snow accumulation and subsequent snowmelt runoff between these
two sites may account for the observed peat isotopic compositions, with significant positive
Δ201
Hg and Δ200
Hg in the PEATcosm peat which receives nearly ~70% more snow (as SWE)
annually. Interception of snow by the spruce and tamarack cover in the S1 peatland as compared
to the lack of tree canopy at PEATcosm may further exacerbate the observed isotopic trends.
The influence of snow in contributing Hg to soils including peat warrants further study,
particularly for landscapes wherein snow accounts for a significant portion of the annual
hydrological budget. As climate change will likely affect snow accumulation in these regions
(Garris et al. 2015), understanding the contribution of snow Hg in controlling peat Hg
accumulation is crucial.
The strong relationship between the peat MDF and the TGM fluxes in the PEATcosm
mesocosms, and the lack of correlation for the S1 peat, may also be influenced by canopy cover
and the stability of the near-surface boundary layer. Deposition of predominantly Hg(II) in the
PEATcosm mesocosms and subsequent reduction, potentially on the leaf surfaces of the vascular
vegetation, may influence TGM fluxes from the peat. Whereas photochemical reduction appears
to be an important controlling influence on Hg cycling in the S1 peat, the mechanism responsible
for the MIFodd and MIFeven anomalies in the PEATcosm peat is unclear. The isotopic patterns as
a result of atmospheric processes require further study including the examination of non-arctic
snow. As the large stocks of Hg stored in peatlands (Kolka et al. 2001; Grigal 2003) are
vulnerable to the effects of climate change, understanding the mechanisms controlling Hg(0) and
Hg(II) contributions to peat, the mechanisms influencing Hg exchange with the peat, including
oxidation-reduction processes and the influence of vegetation, and therefore overall peat Hg
accumulation is warranted.
5.5 Acknowledgements
We would like to acknowledge the laboratory assistance of L. Zimmermann and M. Lee as well
as the field assistance of K. Ng and S. Rao. We thank L. Jamie Lamit for ambient peat collection
at the PEATcosm experiment. We would like to thank P. Hanson and W.R. Nettles of Oak
164
Ridge National Laboratory and the U.S. Forest Service for access to the SPRUCE plots in the S1
peatland of the Marcell Experimental Forest. Funding was provided through a Natural Sciences
and Engineering Research Council of Canada (NSERC) Alexander Graham Bell Canada
Graduate Scholarship (CGS-Doctoral) to K.M.H and a NSERC Discovery Grant (Fund #355866)
to C.P.J.M. The PEATcosm experiment was funded by the USDA Forest Service Northern
Research Station Climate Change Program and the National Science Foundation (DEB-
1146149).
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Chapter 6 Summary and Synthesis
6
6.1 Summary
The overarching objective of this research was to identify the potential impacts of climatic
change on Hg cycling in peatlands using mesocosm- and field-scale approaches. In order to
address the objectives and questions set forth at the beginning of this thesis, this research used
two novel experiments designed to explore different anticipated climate change effects in sub-
boreal peatland ecosystems. Sampling was conducted primarily at the mesocosm-scale at the
PEATcosm experiment, which investigated the effects of altered water table positions and
shifting vascular plant communities, anticipated to occur with climate change. Field-scale work
was also conducted in the SPRUCE (S1) peatland of the Marcell Experimental Forest, which
underwent a deep peat warming experiment beginning in June 2014 and is now the site of the
whole ecosystem warming and elevated atmospheric carbon dioxide experiment. Total Hg and
MeHg concentrations in pore water and snowmelt runoff as well as in the solid phase peat were
monitored in the PEATcosm peat monoliths over several years. Experimental research
examining gaseous Hg fluxes in natural environments is lacking in the literature, likely due to the
numerous challenges associated with field flux measurements. In this research, great care was
taken to ensure consistent meteorological conditions during sampling events and to account for
variations in such factors as plant community coverage and hydrology. By mitigating these
challenges, climate change impacts of hydrology and vascular plant communities at the
PEATcosm experiment and peat warming at depth at the SPRUCE experiment on total gaseous
Hg fluxes could be assessed. Natural abundance Hg isotopic analyses were applied to
investigate relative contributions of Hg(0) and Hg(II) deposition to surface peat of the two
peatlands as well as to gain insight into the mechanisms potentially controlling peat Hg cycling.
The effects of the PEATcosm climate change manipulations on Hg cycling processes in the
surface peat were also explored using isotopic analyses. The following subsections are meant
173
not only to summarize conclusions from individual thesis chapters, but also to progressively
synthesize across chapter findings within each section.
6.1.1 Hg and MeHg Mobility in Peat Pore Water and Runoff
Peatland ecosystems are known to be significant sources of MeHg to downstream aquatic
ecosystems. Lowered, fluctuating water tables and the removal of Ericaceae shrub vegetation
resulted in enhanced Hg and MeHg concentrations in the peat pore waters (Chapter 2). This
treatment yielded the highest concentrations of Hg and MeHg in snowmelt runoff exported from
the peat monoliths. A significant positive correlation was observed between Hg and dissolved
organic carbon (DOC) concentrations in both the peat pore waters and snowmelt runoff. This
suggests that one factor driving the enhanced leaching of Hg and MeHg from the solid peat into
the mobile phase with simulated climate change effects is increased peat decomposition.
Significant positive correlations between the amount of mass lost from cellulose decomposition
assays, representing potential peat decomposition, and both THg and MeHg concentrations in
pore waters are in agreement with peat decomposition influencing the mobilization and
accumulation of inorganic Hg and MeHg in pore waters. In addition to the influence of water
table position, plant functional groups affected Hg and MeHg mobility with increased
concentrations when Ericaceae shrubs were removed under both water table treatments.
Although not measured, it is likely that root exudation of oxygen and labile organic carbon by
sedges contribute to augmented pore water Hg and MeHg accumulation by enhancing peat
decomposition as well as potentially priming the process of methylation. Oxygenation within the
sedge rooting zone due to root oxygen loss may contribute to both aerobic peat mineralization
and regeneration of alternative electron acceptors. Within the PEATcosm pore waters, sulfate
concentrations were enhanced for both the lowered water table treatments and the sedge-
dominated peat monoliths, providing support for these conditions to regenerate terminal electron
acceptors and enhance Hg methylation.
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6.1.2 Solid Phase Hg and MeHg – Translocation and Net MeHg Production
Over the course of the PEATcosm experiment, inorganic Hg and MeHg partitioning from the
peat to the pore waters was significantly influenced by the water table and plant functional group
treatments (Chapter 4). Greater partitioning into the aqueous phase was observed with lowered
and more variable water table positions in the upper depths of peat affected by increased peat
aeration. Changes in vascular vegetation cover also impacted inorganic Hg and MeHg
partitioning to pore water below the rooting zone, with enhanced mobilization in pore waters
when Ericaceae shrubs were removed. These results are in good agreement with the Chapter 2
findings in the peat pore waters.
In Chapter 4, significant changes in the peat THg and MeHg depth profiles were observed
throughout the course of the experiment, with the peak in THg and MeHg concentrations
occurring approximately 30-40 cm beneath the peat coinciding with the depth of the lowered
water tables in 2014. Within the rooting zone, MeHg was also significantly elevated when
Ericaceae shrubs were removed. Surprisingly, no significant differences in methylation or
demethylation were observed among either the water table or plant functional group treatments.
Collectively, this lends support to peat decomposition under aerobic conditions, subsequent
leaching of DOC and partitioning of both inorganic Hg and MeHg from the solid phase peat into
the aqueous phase as the processes instrumental in controlling the accumulation in pore waters.
Mercury partitioned into the pore waters may be transported to the depths near the lowered tables
and re-sorbed to the solid phase, potentially due to the affinity of Hg for organic matter (Figure
6-1). Translocation of inorganic Hg and MeHg within the peat profile appears to be a stronger
influence in governing the observed PEATcosm experimental peat profiles than net MeHg
production.
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Figure 6-1 Influence of water table lowering on the partitioning of Hg and MeHg from the solid
phase into the aqueous pore water phase, translocation down the peat profile and re-sorption to
the solid phase within the zone of lowered water table fluctuation.
The fact that methylation potentials were similar between the two water table treatments at a
depth of 35-40 cm below the peat surface, where the lowered water tables were maintained at the
time of incubation, suggests that methylation may be stimulated despite increased oxygenation of
the peat. Methylmercury production is known to occur under anoxic conditions suitable for
sulfate- and iron-reduction, which accounts for the observed methylation potentials under
saturated conditions. Under lowered water table conditions, however, Hg methylation may be
comparable to the saturated conditions as a result of terminal electron acceptor regeneration
priming the process. The elevated sulfate concentrations observed in the pore waters of the
lowered water table treatments in Chapter 2, particularly at a depth of 40 cm below the peat
surface suggest that terminal electron acceptor regeneration may be a contributing factor to this
finding of similar methylation under drier conditions. Available terminal electron acceptors, as
well as a supply of labile carbon from both peat mineralization and sedge root exudation, may
collectively act to stimulate Hg methylation within the zone of lowered, fluctuating water tables
similar to that produced by the anoxia alone under saturated conditions. Despite the anticipated
loss of sustained anoxic conditions within the surface peat depths with a changing climate, the
hydrological conditions and subsequent shifts in plant communities may augment the mobility of
inorganic Hg and MeHg, both within the peat profile as well as partitioning into the peat pore
waters available for export from the systems during high-flow events such as spring snowmelt.
176
6.1.3 Vascular Plant Communities – Interactions with Atmospheric Hg Deposition
The dominant vascular plant communities present in the peatland may also significantly affect
the pathway by which Hg is deposited to the peat. In Chapter 3, the strongest trends of total
gaseous Hg (TGM) deposition in the PEATcosm peat monoliths were observed when sedges
were present. In comparison, Ericaceae-dominated peat resulted in predominantly emissive
fluxes, the strength of which was correlated with the amount of Ericaceae leaf coverage. As the
sedges had significantly lower surface-sorbed and tissue Hg concentrations than the Ericaceae
shrubs, TGM may be shuttled to the peat in the rooting zone, incidentally with oxygen, via the
aerenchymous tissues of the sedges. Shuttling of TGM to the roots of the sedges may contribute
to the elevated Hg concentrations in the solid phase at these depths observed in Chapter 4.
Strong TGM deposition in the sedge-dominated peat is evident in the Hg isotope data (Chapter
5). The highest mass-dependent fractionation (MDF) among the PEATcosm treatments was
indeed in the surface peat with sedge cover.
In contrast to the sedges, Ericaceae shrub leaves had higher surface as well as leaf tissue Hg
concentrations (Chapter 3). The character of the shrub leaves may promote surface sorption.
Stomatal uptake of gaseous Hg likely contributes to elevated tissue Hg concentrations. The
different pathways, both stomatal and non-stomatal, by which gaseous Hg is deposited to the
Ericaceae shrub cover, are reflected in the peat isotopic signatures measured in Chapter 5. More
negative MDF in the Ericaceae only and unmanipulated vegetation treatments of the PEATcosm
experiment suggest that stomatal uptake of gaseous Hg by the shrubs, and subsequent deposition
of litter of the leaves during senescence, is reflected in the surface peat. Reduction mechanisms
may contribute to the observed MDF anomalies, but further controlled study of such processes is
required to identify those active in the peat soils. The considerable TGM emissions observed
with the Ericaceae-dominated peat monoliths (Chapter 3) was evident in the peat isotopic data
(Chapter 5). Positive mass-independent fractionation (MIF) anomalies in the peat as well as the
observed relationship between TGM fluxes and MDF may be indicative of photo-reduction of
Hg(II), which is typically the dominant form of Hg found on leaf surfaces. The agreement
between the observed TGM fluxes among the PEATcosm plant community treatments in
Chapter 3 and the Hg isotopic signatures in Chapter 5 suggests that shifting plant functional
177
groups in a changing climate will alter the mechanisms and pathways by which gaseous Hg is
deposited to peatland systems.
6.1.4 Spatial Variability in Hg Isotopic Signatures
In addition to vegetation influences on Hg deposition, the geographical location and canopy
structure of peatlands influence the sources and cycling of Hg (Chapter 5). Dramatic differences
in peat isotopic signatures between the northern Minnesota (S1 peatland) and northern Michigan
(PEATcosm) peatland sites highlight considerable differences in Hg(0) and Hg(II) deposition.
The isotopic evidence suggests that significant differences in the amount of snow water
equivalent received by the two sites during spring snowmelt period accounts for the isotopic
disparities between the two peatlands. The increasing contributions of Hg(II) from snow at the
PEATcosm site was confirmed by the differences in MIFodd values between the water table
treatments; with higher Δ199
Hg and Δ201
Hg values in the wetter peat monoliths. The presence of
a tree canopy at the S1 (SPRUCE) peatland increased the contributions of Hg(0) to the peat,
possibly due to boundary layer stability, as compared to the PEATcosm site. The previously
undocumented MIFodd and MIFeven isotopic anomalies observed in the PEATcosm peat samples
require further research to assess the sources and processes that may fractionate Hg in this
manner.
6.1.5 Impacts of Climate Change on Peatland Hg Cycling
The two experimental sites, PEATcosm and SPRUCE, were established to simulate anticipated
climate change impacts on peatlands. In this thesis, the effects of water table variations, different
vascular plant functional group assemblages and peat warming at depth on Hg cycling were
investigated. The vascular plant functional groups examined in the PEATcosm experiment
interact with atmospheric Hg in different ways and thereby affect the pathways by which Hg is
deposited to peatlands. Shuttling of Hg to peat soils via aerenchymous tissues may act to lessen
atmospheric Hg burdens. However, enhanced deposition to the peat may increase the amount of
Hg available to be methylated. On the other hand, Ericaceae shrubs accumulate Hg in the leaf
tissues and are subsequently deposited to peat as litter upon senescence. Sorption of Hg to
178
Ericaceae leaf surfaces has the potential to be a pathway of significant Hg re-emission to the
atmosphere. Therefore, as the direction in which vascular plant communities will shift with
climate change (dominance of sedges vs. Ericaceae shrubs) is still unclear, the manner by which
gaseous Hg exchange between peatlands and the atmosphere will progress cannot yet be
conclusively stated. The SPRUCE peat warming at depth may be an indication that increasing
peat temperatures enhance gaseous Hg fluxes from the peat to the atmosphere. This effect may
be amplified when air temperatures are also increased as part of the SPRUCE experiment, which
will result in a more even warming of the peat. Partitioning of inorganic Hg and MeHg from the
solid phase to the mobile pore waters increased as a result of enhanced aerobic decomposition
with lowered water tables and the removal of Ericaceae shrub cover. Lowered and more variable
water table positions, and to a lesser extent the removal of Ericaceae shrubs, has the potential to
increase loading of Hg and MeHg to downstream aquatic ecosystems. The considerable
influence of snow meltwater in contributing to the peat Hg stocks was an unexpected result of
this research. In regions where snow accumulation is expected to increase with climate change,
greater deposition of Hg(II) to vegetation surfaces may delay and reduce the incorporation of Hg
into the soil pool and potentially result in enhanced gaseous Hg emissions to the atmosphere, as
Hg(II) on ground vegetation is susceptible to reduction and re-emission. The differences in
Hg(0) and Hg(II) contributions due to peatland canopy structure as well as deposition via snow
may significantly impact Hg cycling in these systems.
6.2 Limitations and Future Research Directions
Conducting peatland research at the mesocosm-scale and at the field plot-scale allows for the
direct and largely controlled manipulation of the factors under study. This approach effectively
removes the influences of potential confounding factors such as groundwater contributions,
which cannot be easily accounted for at larger scales in a natural system. These studies were
conducted in peatlands located at the southern extent of the boreal ecosystem. Given that this
research examined the direct influences of hydrology, plant communities and deep soil warming
on Hg cycling, the observations made in these sites are transferable to peatlands located in
temperate, northern boreal and arctic ecosystems. This research demonstrates how peatland Hg
cycling may be mechanistically impacted by the imposed climate change manipulations of
179
lowered water tables, different vascular plant community assemblages and increasing peat
temperatures. Depending upon how climate change may influence these factors within a certain
region, the potential effects on Hg cycling could be interpreted accordingly. An important next
step in this research could involve the investigation of these factors on Hg cycling across
latitudinal gradients. The observed significance in snow accumulation in contributing to peat Hg
stocks, for example, lends itself to further analysis. A review of the current literature examining
Hg isotope analyses across a gradient in snow accumulation could be conducted to further
examine the mechanistic trends observed in this thesis.
An exciting international collaborative approach, termed PeatDataHub, has been recently
initiated to address global-scale research questions in peatland science (Young et al. 2016). By
sharing data sets collected from long-term peatland monitoring sites as well as short-term
measurements collected around the world, peatland researchers can synthesize observations
across a range of scales and topics. The aggregation of peatland data would, for example,
advance the understanding of changes in hydrology and ecology anticipated with climate change
on the global peatland response in terms of carbon sink-source functions and feedback
mechanisms. Investigation of other potential issues of climate change including the cycling and
mobility of Hg in peatlands would benefit from global collaborative efforts such as the
PeatDataHub. Currently in the literature, few studies examine the impacts of climate change on
Hg cycling processes in peatlands. These studies are typically restricted to boreal landscapes in
North America, such as this research, and the arctic and boreal regions of Fennoscandia (e.g.
Rydberg et al. 2010; Fritsche et al. 2014). Expansion of peatland Hg research in relation to
global climate change would be facilitated by a collaborative global network, including the
potential for synthesizing results across latitudinal and other gradients. In addition to the impacts
of a changing climate, the effects of anthropogenic land use change, such as drainage of
peatlands and harvesting of peat, on the cycling of both Hg and carbon could also be an avenue
for extensive research facilitated by such a network. One major opportunity presented by the
PeatDataHub network is the potential to influence and inform policy and peatland management
strategy practices (Young et al. 2016). Collaborative global research could incorporate not only
peatland hydrology and carbon storage capabilities, but also strategies to mitigate potential
effects on Hg cycling, export from peatlands, and availability for uptake into vulnerable aquatic
food chains.
180
Planned future research with the PEATcosm data involves a collaborative, interdisciplinary
effort to examine the hgcAB gene pair in the peat using the metagenomic and metatranscriptomic
data on bacteria and archaea analyzed in association with the Joint Genome Institute (JGI)
project. This data will facilitate the assessment of the effects of the water table and plant
functional group treatments on Hg methylation gene presence and activity. These results could
then be related to the trends in methylation and demethylation rate potentials (Chapter 4) to
determine whether the genetic potential for methylation is either supportive of or decoupled from
kmeth and kdemeth.
The next steps in examining the potential effects of climate change on Hg cycling in peatlands
should involve experimental manipulation of water tables and vascular plant functional groups at
the peatland-scale. Long-term peatland water table manipulations, such as the poor fen sites at
the Seney National Wildlife Refuge in the Upper Peninsula of Michigan examining carbon
cycling (Chimner et al. 2016), would be ideal sites to examine the subsequent impacts on Hg
cycling. Free air carbon enrichment studies, such as the SPRUCE experiment now underway
with whole ecosystem warming, will provide insight into how plant communities may shift and
peatland water tables may react naturally in response to warming temperatures and elevated
atmospheric carbon dioxide concentrations. A targeted approach could then be taken to aid in
modelling the subsequent effects on Hg cycling.
Extensive further research should be carried out using natural abundance stable Hg isotope
fractionation patterns. Given the observed significance of geographic location on Hg cycling in
the two study sites in this thesis, further isotopic work is required to examine the potential spatial
variation in peatlands globally. Controlled experiments of Hg cycling mechanisms, including
deposition, reduction and atmospheric cycling mechanisms, would aid in identifying the isotopic
patterns observed in natural samples: peat, pore waters, vegetation, bulk gaseous samples and
precipitation including snow. Understanding the processes controlling Hg cycling will greatly
enhance the ability to model how peatland Hg cycling may be affected by the anticipated
changes in global climate.
181
6.3 References
Chimner, R. A., T. G. Pypker, J. A. Hribljan, P. A. Moore, and J. M. Waddington (2016), Multi-
decadal changes in water table levels alter peatland carbon cycling. Ecosystems,
DOI:10.1007/s10021-016-0092-x.
Fritsche, J., S. Osterwalder, M. B. Nilsson, J. Sagerfors, S. Åkerblom, K. Bishop, and C. Alewell
(2014), Evasion of elemental mercury from a boreal peatland suppressed by long-term
sulfate addition. Environmental Science and Technology Letters, 1, 421-425.
Rydberg, J., J. Klaminder, P. Rosén, and R. Bindler (2010), Climate driven release of carbon and
mercury from permafrost mires increases mercury loading to sub-arctic lakes. Science of
the Total Environment, 408, 4778-4783.
Young, D. M., P. J. Morris, and J. Holden (2016), Upscaling peatland science through
collaborative big data. Eos, 97, https://doi.org/10.1029/2016EO06125.
182
Appendix A. Supplementary Information for Chapter 2
Text A-2-1
All data were tested for normality (Shapiro-Wilk W test) and heteroscedasticity and were
successfully log-transformed or square root-transformed to achieve normality when parametric
assumptions were not met. The data from the midsummer pore water collections (August 2013
and July 2014) were removed from statistical analyses as only a small percentage (50% for
August 2013, 65% for July 2014) of the samples could be obtained. Due to low midsummer
water table levels insufficient water was available for extraction from the peat monoliths,
particularly at the 20 and 40 cm depths for the low WT treatment mesocosms, which inhibited a
direct statistical comparison only during this time. As the purpose of this study was to assess the
treatment effects on Hg mobilization in pore water which could subsequently be transported out
of the peatland during high-flow events, we feel that the inclusion of these partial midsummer
sampling events would in any case not address our objectives, because very little water is
exported from peatlands in streamflow during mid-summer low water periods. For all of the
other collection events, more than 80% of possible pore water samples were collected. Missing
samples, due to the lack of available sample for analysis during these collections, were treated as
‘not available’ for statistical analyses.
The main influences of water table position and vascular plant functional group on pore water
THg concentrations, MeHg concentrations and the percentage of THg as MeHg (%MeHg) over
the three sampling depths were assessed using repeated measures analyses of variance
(ANOVA), with water table, plant functional group and sampling depth as the main factors,
taking into consideration the repeated sampling events of the pore waters over time. No
significant interactive effects of water table position and vascular plant functional group were
observed for pore water Hg concentrations. Therefore, one-way ANOVAs with post-hoc Tukey
tests were performed to discern significant differences in THg and MeHg concentrations, and
%MeHg between the six crossed treatment combinations. Water table and plant community
effects as well as any interactive effect between these factors on THg and MeHg concentrations,
and %MeHg in snowmelt runoff were analyzed using three-way repeated measures ANOVAs to
183
account for any potential differences in the main factors between the two sampling years. One-
way ANOVAs with post-hoc Tukey tests were performed to assess significant differences in
THg concentrations, MeHg concentrations and %MeHg in snowmelt between the six crossed
treatment combinations given no interactive effects of the main experimental factors.
Correlation matrices were created using Pearson correlation coefficients in order to assess any
significant relationships between pore water Hg concentrations and dissolved organic carbon
(DOC), total phenolics concentrations as well as Hg and DOC concentrations in snowmelt
runoff.
184
Figure A-2-1 Photos of mesocosm bins.
Sedge Only
Ericaceae Only
Unmanipulated Control
185
Figure A-2-2 Runoff total Hg and MeHg loads across the six crossed water table and vascular
plant functional group treatments during the 2014 spring snowmelt period (n=4 per treatment). a)
Total Hg and b) MeHg. Letters denote statistically similar groups based on transformed data.
186
Figure A-2-3 Mean pore water MeHg concentrations (on log scale) during the five sampling
events considered for this study. a) through e) denote specifically sampling periods. Each box
represents the mean pore water concentration of all three sampling depths across the four
replicate mesocosms (n=10 - 12).
187
Figure A-2-4 Mean pore water THg concentrations (on log scale) during the five sampling
events considered for this study. a) through e) denote specifically sampling periods. Each box
represents the mean pore water concentrations of all three sampling depths across the four
replicate mesocosms (n=10 - 12).
188
Figure A-2-5 Pore water Hg-DOC relationships. a) THg (log scale) in relation to DOC (square
root transformed) concentration, and b) MeHg (square root transformed) in relation to DOC
(square root transformed) concentration. The data includes all three pore water depths from each
of the four replicates per each of the six treatments. All five sampling events are included.
189
Figure A-2-6 Relationships between shallow pore water (20 and 40 cm below the peat surface)
Hg and % mass loss of cellulose decomposition assays in the top 40 cm of the peat. a) THg
concentrations, b) MeHg concentrations, c) %MeHg. Decomposition assays were harvested in
2014 and represent potential peat decomposition among the treatments. From all five sampling
events, pore water Hg data collected at 20 and 40 cm depths was averaged by mesocosm. n=18-
20 per treatment.
190
Figure A-2-7 Locally-weighted scatterplot smoothing (LOWESS) relationships between shallow
pore water (20 and 40 cm below the peat surface) Hg and % mass loss of cellulose
decomposition assays in the top 40 cm of the peat. a) THg concentrations, b) MeHg
concentrations, c) %MeHg. Decomposition assays were harvested in 2014 and represent
potential peat decomposition among the treatments. From all five sampling events, pore water
Hg data collected at 20 and 40 cm depths was averaged by mesocosm. n=18-20 per treatment.
191
Figure A-2-8 Pore water Hg-total phenolics relationships. a) THg (log transformed) in relation
to total phenolics (square root transformed) concentrations, and b) MeHg (square root
transformed) in relation to total phenolics (square root transformed) concentrations. The data
includes all three pore water sampling depths from each of the four replicates per each of the six
treatments. All five sampling events are included.
192
Figure A-2-9 Sulfate concentrations among the six crossed WT and plant functional group
treatments including all three sampling depths for the 2013 and 2014 sampling events. Letters
denote statistically similar groups.
193
Table A-2-1 Volume of water drained (mean ± standard deviation, in L) from the mesocosm
bins during spring snowmelt 2014 to achieve the water table treatment positions (n=4 per
treatment).
Mean ± Standard Deviation
Volume of Water Drained (L)
High WT Ericaceae 68 ± 31
High WT Sedge 89 ± 35
High WT Control 59 ± 34
Low WT Ericaceae 119 ± 25
Low WT Sedge 165 ± 7
Low WT Control 108 ± 16
194
Appendix B. Supplementary Information for Chapter 3
Figure B-3-1 a) Relationship between mean inlet TGM concentrations (ng m-3
) and
corresponding 20-min TGM fluxes measured during daylight hours across the PEATcosm
experimental treatments. b) Boxplots of mean inlet Hg concentrations across the four PEATcosm
treatments. Letters denote statistically similar values.
195
Figure B-3-2 Piecewise regressions of PEATcosm hourly TGM fluxes (in ng m-2
h-1
) and
surface soil temperatures (in °C) for the a) High WT Control, b) Low WT Control, c) Low WT
Sedge, and d) Low WT Ericaceae treatments. Temperature thresholds are noted for each
treatment by the vertical dashed lines. The equations of the relationships before and after each
threshold temperature are specified.
196
Figure B-3-3 a), c) and e) Boxplots of mean inlet TGM concentrations measured during daylight
hours across the SPRUCE plots for each of the three sampling events. Letters denote statistically
similar values. b), d) and f) Relationship between mean inlet TGM concentrations (ng m-3
) and
corresponding 20-min TGM fluxes (daylight only) across the SPRUCE plots during each of the
three samplings.
197
Appendix C. Supplementary Information for Chapter 4
Figure C-4-1 Peat MeHg and THg stock (in ng cm-3
) profiles for the six combined water table
and vascular plant functional group treatments throughout the PEATcosm experiment from 2011
to 2014 and associated mean June to October water table positions for the High and Low WT
prescriptions. For clarity, error bars for each 10 cm depth increment have been omitted.
198
Figure C-4-2 Mean kdemeth (d-1
) in the upper 15 cm of the peat profile and 35-50 cm below the
peat surface for the a) High WT and b) Low WT treatments. Mean water table levels for the
month prior to sampling are denoted by the dashed lines. No statistically significant differences
among the depths for each of the High and Low WT data.
199
Figure C-4-3 Soil-water partition coefficients (logKD) for Hg(II) at 20 and 40 cm below the peat
surface among the six crossed water table and plant functional group treatments in 2013 and
2014 (n = 4 per treatment). Letters denote statistically similar groups.
200
Appendix D. Supplementary Information for Chapter 5
Figure D-5-1a S1 peatland of the Marcell Experimental Forest in north-central Minnesota
showing one SPRUCE plot footprint.
201
Figure D-5-1b PEATcosm Mesocosm Facility, Houghton, Michigan, with the 24 peat monoliths
(top) and underlain by a climate-controlled tunnel (bottom).
202
Figure D-5-2 Relationship between Δ200
Hg and Δ204
Hg (all expressed in ‰) for the surface 0-10
cm peat from both the S1 and PEATcosm peatland sites.
203
Figure D-5-3 Relationship between Δ200
Hg and Δ201
Hg (expressed in ‰) for the surface 0-10 cm
peat from both the S1 and PEATcosm peatland sites as compared to reported values of Hg(II)
deposition in the literature (Gratz et al. 2010; Sherman et al. 2012; Demers et al. 2013).
204
Table D-5-1 Mercury isotope signatures of reference materials and standards.
Sample ID n Recovery SD
δ204Hg
(‰) 2σ
δ202Hg
(‰) 2σ
δ201Hg
(‰) 2σ
δ200Hg
(‰) 2σ
δ199Hg
(‰) 2σ
Δ204Hg
(‰) 2σ
Δ201Hg
(‰) 2σ
Δ200Hg
(‰) 2σ
Δ199Hg
(‰) 2σ
NIST SRM 1944 2 107% 24% -0.58 0.06 -0.39 0.05 -0.30 0.06 -0.20 0.03 -0.08 0.04 0.01 0.07 -0.01 0.04 0.00 0.02 0.02 0.04
NIST SRM 1646a 1 80% N/A -1.40 0.18 -0.97 0.08 -0.68 0.14 -0.44 0.04 -0.10 0.04 0.04 0.07 0.05 0.08 0.05 0.02 0.14 0.04
CCRMP TILL-1 9 86% 5% -1.69 0.06 -1.13 0.05 -0.97 0.06 -0.55 0.03 -0.38 0.04 -0.01 0.07 -0.12 0.04 0.01 0.02 -0.09 0.04
NIST SRM 3133 7 91% 3% -0.05 0.06 -0.01 0.05 0.00 0.06 -0.01 0.03 0.00 0.04 -0.04 0.07 0.01 0.06 0.00 0.02 0.00 0.04
JTBaker 22 -0.87 0.06 -0.58 0.05 -0.43 0.06 -0.28 0.03 -0.12 0.04 -0.01 0.07 0.00 0.04 0.01 0.02 0.03 0.04
Note: 2σ is the higher of either 2 standard deviation (2SD) of the JTBaker standard or 2 standard error (2SE) of the mean.
205
Table D-5-2 Mercury isotope signatures of all peat samples. Each sample was analyzed twice for isotopes. 2σ is the higher of either 2
standard deviation (2SD) of the JTBaker standard or 2 standard error (2SE) of the two replicate results of each sample.
Sample ID
THg
(ng g-1)
δ204Hg
(‰) 2σ
δ202Hg
(‰) 2σ
δ201Hg
(‰) 2σ
δ200Hg
(‰) 2σ
δ199Hg
(‰) 2σ
Δ204Hg
(‰) 2σ
Δ201Hg
(‰) 2σ
Δ200Hg
(‰) 2σ
Δ199Hg
(‰) 2σ
PEATcosm Experimental Peat - 2014
Low WT – Sedge 39 -1.64 0.06 -1.07 0.05 -0.78 0.07 -0.49 0.07 -0.18 0.08 -0.04 0.07 0.03 0.04 0.05 0.05 0.09 0.07
Low WT – Ericaceae 38 -1.89 0.06 -1.27 0.05 -0.96 0.06 -0.60 0.05 -0.27 0.04 0.01 0.07 -0.01 0.04 0.04 0.04 0.05 0.04
Low WT – Control 30 -2.17 0.06 -1.47 0.05 -1.03 0.06 -0.67 0.04 -0.29 0.04 0.02 0.07 0.07 0.04 0.07 0.02 0.08 0.04
High WT – Sedge 32 -1.54 0.09 -1.04 0.05 -0.63 0.06 -0.44 0.03 -0.09 0.04 0.01 0.07 0.16 0.04 0.08 0.03 0.17 0.04
High WT – Ericaceae 24 -1.64 0.06 -1.12 0.05 -0.66 0.06 -0.49 0.03 -0.17 0.04 0.03 0.07 0.18 0.04 0.08 0.02 0.11 0.04
High WT - Control 24 -1.93 0.06 -1.33 0.05 -0.75 0.06 -0.58 0.03 -0.22 0.04 0.05 0.07 0.25 0.04 0.08 0.03 0.11 0.04
PEATcosm Pre-Treatment Peat – 2011
Low WT – Sedge 52 -1.88 0.20 -1.23 0.26 -0.95 0.19 -0.52 0.16 -0.31 0.18 -0.05 0.18 -0.03 0.04 0.09 0.03 -0.01 0.11
Low WT – Ericaceae 55 -2.11 0.11 -1.34 0.10 -0.78 0.06 -0.31 0.09 -0.30 0.04 -0.10 0.07 0.23 0.04 0.36 0.04 0.03 0.04
High WT – Sedge 54 -1.42 0.06 -1.01 0.05 -0.80 0.06 -0.35 0.07 -0.24 0.04 0.09 0.07 -0.04 0.05 0.15 0.07 0.01 0.04
S1 Peat – 2014
Plot 4 69 -2.60 0.06 -1.74 0.05 -1.53 0.06 -0.83 0.04 -0.72 0.04 -0.01 0.07 -0.22 0.04 0.04 0.03 -0.28 0.04
Plot 6 76 -2.61 0.08 -1.78 0.05 -1.58 0.07 -0.83 0.03 -0.66 0.04 0.04 0.07 -0.24 0.05 0.07 0.02 -0.21 0.04
Plot 10 42 -2.48 0.09 -1.66 0.05 -1.35 0.06 -0.80 0.03 -0.49 0.04 0.00 0.07 -0.10 0.04 0.04 0.03 -0.07 0.04
Plot 13 38 -2.02 0.06 -1.36 0.05 -1.09 0.06 -0.58 0.03 -0.36 0.04 0.01 0.07 -0.07 0.04 0.10 0.02 -0.02 0.04
Plot 17 29 -2.34 0.06 -1.55 0.05 -1.26 0.06 -0.68 0.03 -0.44 0.04 -0.02 0.07 -0.09 0.04 0.10 0.02 -0.05 0.04
Plot 19 63 -2.58 0.12 -1.74 0.11 -1.46 0.12 -0.82 0.09 -0.58 0.08 0.01 0.07 -0.15 0.04 0.05 0.03 -0.15 0.05
206
Table D-5-3 Mean and standard deviation values of Δ199
Hg, Δ200
Hg and Δ201
Hg (all expressed in
‰) for Hg(II) and Hg(0) used in the binary mixing model with Monte Carlo simulation.
Hg(II) Hg(0)
Mean Standard Deviation Mean Standard Deviation
Δ199
Hg (‰) 0.36 0.28 -0.55 0.07
Δ200
Hg (‰) 0.26 0.25 -0.05 0.05
Δ201
Hg (‰) 0.35 0.26 -0.50 0.07
207
Table D-5-4 Contributions of Hg(II) (fHg(II)) and Hg(0) (fHg(0)) to peat isotopic signatures for the
PEATcosm 2014 treatment peat and S1 2014 peat. Results are the mean of 1000 iterations of the
isotopic mixing model with Monte Carlo simulation based on the Δ199
Hg isotopic values.
Δ199
Hg fHg(II) fHg(0)= 1 – fHg(II)
Mean Standard Deviation Mean Standard Deviation
PEATcosm Experimental Peat - 2014
Low WT – Sedge 0.73 0.13 0.27 0.13
Low WT – Ericaceae 0.68 0.13 0.32 0.13
Low WT – Control 0.72 0.13 0.28 0.13
High WT – Sedge 0.82 0.15 0.18 0.15
High WT – Ericaceae 0.75 0.14 0.25 0.14
High WT - Control 0.75 0.14 0.25 0.14
S1 Peat – 2014
Plot 4 0.30 0.07 0.70 0.07
Plot 6 0.38 0.08 0.62 0.08
Plot 10 0.54 0.10 0.46 0.10
Plot 13 0.60 0.11 0.40 0.11
Plot 17 0.57 0.11 0.43 0.11
Plot 19 0.45 0.09 0.55 0.09
208
Table D-5-5 Contributions of Hg(II) (fHg(II)) and Hg(0) (fHg(0)) to peat isotopic signatures for the
PEATcosm 2014 treatment peat and S1 2014 peat. Results are the mean of 1000 iterations of the
isotopic mixing model with Monte Carlo simulation based on the Δ200
Hg isotopic values.
Δ200
Hg fHg(II) fHg(0) = 1 – fHg(II)
Mean Standard Deviation Mean Standard Deviation
PEATcosm Experimental Peat - 2014
Low WT – Sedge 0.44 0.39 0.56 0.39
Low WT – Ericaceae 0.41 0.36 0.59 0.36
Low WT – Control 0.53 0.49 0.47 0.49
High WT – Sedge 0.62 0.61 0.38 0.61
High WT – Ericaceae 0.58 0.56 0.42 0.56
High WT - Control 0.62 0.61 0.38 0.61
S1 Peat – 2014
Plot 4 0.41 0.36 0.59 0.36
Plot 6 0.54 0.50 0.46 0.50
Plot 10 0.40 0.35 0.60 0.35
Plot 13 0.71 0.72 0.29 0.72
Plot 17 0.70 0.70 0.30 0.70
Plot 19 0.46 0.42 0.54 0.42