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A comparison of grazed and ungrazed sedge meadows in the Canadian High Arctic Elliott, Tammy Lynn 2009

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A Comparison of Grazed and Ungrazed Sedge Meadows in the Canadian High Arctic by Tammy Lynn Elliott  BSA, The University of Saskatchewan, 2000  A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF SCIENCE in The Faculty of Graduate Studies (Geography)  THE UNIVERSITY OF BRITISH COLUMBIA (Vancouver) May 2009 c Tammy Lynn Elliott 2009  Abstract The grazing optimization hypothesis predicts that net primary production (NPP) and nitrogen levels within vegetation will be highest with moderate grazing levels. In the Canadian High Arctic, muskoxen are one of two major herbivores; they prefer to graze in wet sedge meadow plant communities. To test the grazing optimization hypothesis in these plant communities, two studies were initiated in 2007. The first study spanned two years and compared grazed and ungrazed sedge meadows. The grazed meadows had higher belowground biomass in 2007 and graminoid net primary production was larger in 2008. The ungrazed meadows had greater quantities of dead biomass. Nitrogen concentrations in Carex aquatilis ssp. stans and Eriophorum angustifolium ssp. triste and soil ammonium availability were higher at the grazed site. In the second study, we created two experimental grids with clipping and litter removal treatments. Aboveground net primary production, ecosystem respiration, and shoot carbon concentrations decreased due to clipping. However, shoot nitrogen concentrations increased in C. membranacea and E. triste as clipping frequencies increased. Soil moisture levels also rose with clipping frequencies. Litter removal did not effect aboveground net primary production or soil moisture content. We conclude that the grazing optimization hypothesis applies to High Arctic wet sedge meadows because of the higher aboveground NPP and belowground biomass at the grazed site.  However, decreased aboveground NPP in the clipping experiment indicates that  muskoxen stimulate primary production in these plant communities by accelerating the ii  Abstract nitrogen cycle by the addition of nutrients to the soil from their excrement.  iii  Table of Contents Abstract  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  ii  Table of Contents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  iv  List of Tables  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  vii  List of Figures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  viii  Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  xi  Statement of Co-Authorship . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  xii  1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  1  1.1  1.2  Ecology of grazed ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . .  1  1.1.1  4  Grazing in the Arctic  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  6  Muskox grazing in the Arctic . . . . . . . . . . . . . . . . . . . . . .  7  Site descriptions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  10  1.3.1  Alexandra Fiord . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  10  1.3.2  Sverdrup Pass  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  11  Tables and Figures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  12  1.2.1 1.3  1.4  Effects of grazing on ecosystem processes . . . . . . . . . . . . . . .  References  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  18 iv  Table of Contents 2 Evidence of increased plant growth and nitrogen cycling as a result of muskox grazing in High Arctic sedge meadow plant communities. . . .  25  2.1  Introduction  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  25  2.2  Methods  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  28  2.2.1  Site descriptions . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  28  2.2.2  Aboveground NPP  . . . . . . . . . . . . . . . . . . . . . . . . . . .  29  2.2.3  Belowground biomass . . . . . . . . . . . . . . . . . . . . . . . . . .  31  2.2.4  Shoot nitrogen and carbon concentrations  . . . . . . . . . . . . . .  32  2.2.5  Soil N03 − and NH4 + availability . . . . . . . . . . . . . . . . . . . .  32  2.2.6  Ecosystem respiration . . . . . . . . . . . . . . . . . . . . . . . . . .  33  2.2.7  Statistical analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . .  33  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  34  2.3  Results 2.3.1  Aboveground NPP  . . . . . . . . . . . . . . . . . . . . . . . . . . .  34  2.3.2  Belowground biomass . . . . . . . . . . . . . . . . . . . . . . . . . .  35  2.3.3  Shoot nitrogen and carbon concentrations  . . . . . . . . . . . . . .  35  2.3.4  Soil N03 − and NH4 + availability . . . . . . . . . . . . . . . . . . . .  35  2.3.5  Ecosystem respiration . . . . . . . . . . . . . . . . . . . . . . . . . .  36  2.4  Discussion  2.5  Figures  References  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  36  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  41  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  49  3 Defoliation and litter removal effects on ungrazed wet sedge meadows in the High Arctic.  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  55  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  55  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  59  3.1  Introduction  3.2  Methods 3.2.1  Site description  . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  59 v  Table of Contents  3.3  3.2.2  Clipping experiments  3.2.3  Data analysis  Results  . . . . . . . . . . . . . . . . . . . . . . . . . .  60  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  65  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  66  3.3.1  Aboveground NPP  . . . . . . . . . . . . . . . . . . . . . . . . . . .  66  3.3.2  Ecosystem respiration . . . . . . . . . . . . . . . . . . . . . . . . . .  66  3.3.3  Soil NO3 − and NH4 + availability  66  3.3.4  Shoot nitrogen and carbon concentrations  3.3.5  Soil temperatures  3.3.6  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  67  . . . . . . . . . . . . . . . . . . . . . . . . . . . .  67  Soil moisture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  67  3.4  Discussion  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  68  3.5  Tables and Figures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  74  References  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  83  4 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . .  89  References  95  . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  vi  List of Tables 1.1  Results of selected studies of the response of aboveground NPP to grazing .  1.2  Previous studies conducted on ungrazed sedge meadow plant communities at  12  Alexandra Fiord and grazed sedge meadow plant communities at Sverdrup Pass . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1  13  Schedule for experimental clipping treatments and harvest dates at Alexandra Fiord in 2007 and 2008 . . . . . . . . . . . . . . . . . . . . . . . . . . .  74  vii  List of Figures 1.1  Grazing optimization curve (adapted from Belsky 1986). . . . . . . . . . . .  14  1.2  Map of Ellesmere Island and adjacent islands. . . . . . . . . . . . . . . . . .  15  1.3  Plant communities and the locations of the four study sites at Alexandra Fiord (adapted from Muc et al. 1989). . . . . . . . . . . . . . . . . . . . . .  1.4  Plant communities and the locations of the four study sites at Sverdrup Pass (adapted from Bergeron 1989). . . . . . . . . . . . . . . . . . . . . . . . . .  2.1  16  17  Aboveground NPP and graminoid NPP in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) in 2007 and 2008 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  2.2  41  Aboveground plant biomass in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) at the start of the growing season and at the peak of live aboveground biomass in 2007 and 2008 42  2.3  Non-green, dead, and total standing crop dry weights in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) in 2007 and 2008 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  2.4  43  Belowground standing crop in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) in 2007. . . . . . .  44  viii  List of Figures 2.5  Nitrogen concentrations in three plant species from ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) in 2008. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  2.6  45  Carbon concentrations in three plant species from ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) in 2008. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  2.7  46  Nitrogen supply rates in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) during peak production in 2008. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  2.8  47  Ecosystem respiration in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows Sverdrup Pass (SP) during peak production in 2008 48  3.1  Aboveground NPP from two wet sedge meadow sites at Alexandra Fiord in 2008. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  75  3.2  Aboveground NPP in Grid 2 at Alexandra Fiord in 2008. . . . . . . . . . .  76  3.3  Effect of clipping on ecosystem respiration in two wet sedge meadows at Alexandra Fiord during peak aboveground biomass in 2008. . . . . . . . . .  3.4  Nitrogen supply rates for control and 3x clipping treatment in two wet sedge meadows at Alexandra Fiord in 2008. . . . . . . . . . . . . . . . . . . . . .  3.5  79  Carbon concentrations in shoots of four plant species at Alexandra Fiord in 2008. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  3.7  78  Nitrogen concentrations in shoots of four plant species at Alexandra Fiord in 2008. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  3.6  77  80  Average soil volumetric water content from two clipping experiments at Alexandra Fiord in 2008.  . . . . . . . . . . . . . . . . . . . . . . . . . . . .  81  ix  List of Figures 3.8  Average soil volumetric water content in four treatments in Grid 2 at Alexandra Fiord in 2008. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .  82  x  Acknowledgements This M.Sc. research was supported by grants to my supervisor, Dr. Greg Henry, from the Natural Sciences and Engineering Research Council of Canada (NSERC) and the Government of Canada International Polar Year Program (IPY). The Northern Scientific Training Program (NSTP) of the Department of Indian and Northern Affairs Canada contributed funds for transportation and living costs in the High Arctic. Logistical support in the field was provided by the Polar Continental Shelf Project (PCSP) and the Royal Canadian Mounted Police (RCMP). I would like to extend my gratitude to the Government of Nunavut - Department of Environment and the Qikiqtani Inuit Association for permission to use their land. I acknowledge the grateful help of Jessica Baas, Jennifer Wurz, Adrian Leitch, Carolyn Churchland, Sarah Trefry, and William Brown in the field. The assistance of Michelle Sawka, Justin Lau, and Linda Quamme was appreciated sorting roots in the lab at UBC. Sorting was fun when accompanied by the right people and foreign radio station. Dr. Sarah Elmendorf, Dr. Calvin Winter, Frank deGagne, Carolyn Churchland, Laura Machial, and James Hudson provided comments on early editions of this document. I am confident that my memories of Alexandra Fiord will always remain with me.  xi  Statement of Co-Authorship GHR Henry co-authored Chapters 2 and 3 of this M.Sc. thesis. I completed the majority of the data collection, analyses, and preliminary chapter drafts. The idea and sampling design for this thesis was provided by GHR Henry and he provided guidance with data collection techniques, explanations of underlying processes, and the writing of the final documents.  xii  Chapter 1  Introduction Compared to other terrestrial biomes, few species can tolerate the environmental conditions of the High Arctic (Callaghan et al. 2004). This region is resource-poor because of low temperatures, reduced incoming solar radiation, low precipitation levels, and limited nutrient availabilities (Barry et al. 1981; Freedman et al. 1994; Edlund & Alt 1989; Tolvanen & Henry 2001). The plants that survive in Arctic regions have special adaptations that permit them to grow there, but primary productivity is restricted (Freedman et al. 1994; Gauthier et al. 1995). The lack of vegetative production restricts the number of animals that can survive in these regions (Tener 1958; Raillard & Svoboda 2000). Muskoxen (Ovibos moschatus) are one of two large mammalian herbivores that can endure conditions in the Canadian High Arctic and their preferred forage habitat is wet sedge meadow plant communities (Henry 1998; Raillard & Svoboda 2000). The effects of grazing on ecosystem structure and functioning are not understood as well in High Arctic ecosystems as in other natural systems. The overall objective of this study was to determine the effects of muskox grazing on High Arctic wet sedge meadow plant communities.  1.1  Ecology of grazed ecosystems  Grazing occurs when animal consumers remove tissues from plants. Animals in the wild have a strong tendency to maximize the ingestion of energy and essential nutrients per 1  1.1. Ecology of grazed ecosystems unit time or energy spent in feeding (Arnold 1987). The quality and quantity of plant food determines animal survivorship and the carrying capacity of a piece of land is determined by the availability of food resources (Crawley 1983). Plants and animals have co-evolved and several plant species have adaptations that aid in their ability to withstand the pressures exerted on them by herbivores (Notzold et al. 1998). The distribution of grazing animals at the landscape level depends on the configuration of forage resources, water locations, climate, threat of predators, and terrain constraints (Struth 1991). Adler et al. (2001) stated that the greatest requirement of animals is water; domestic and wild animal species exhibit exponentially less grazing pressure on the landscapes as the distance from water sources increases. Grazing animals often prefer the most productive communities on the landscape because these plant communities are the richest in forage resources. However, steep slopes and the threat of predators may exclude herbivores from otherwise suitable habitats (Adler et al. 2001). Grazing influences plant diversity by creating environmental heterogeneity at several different spatial scales (Adler et al. 2001). Large herbivores disperse themselves spatially on the landscape by favouring certain plant communities from which individual plants are grazed (Briske 1993). Herbivores influence plant community structure by modifying competitive interactions between plants and altering nutrient availability (Olofsson et al. 2001). Grazing animals remove various amounts of leaf areas from plants, which establishes the potential for differential growth rates following a defoliation event (Briske 1993). Defoliation by herbivores also affects individual plants. More frequent and intense defoliation leads to reduced light interception by photosynthetic tissue, depletion of metabolic reserves, reduced uptake of nutrients and water, damage to meristems, and depletion of seed reserves (Korte & Harris 1987). In some systems that have little or no grazing pressure, the build-up of litter is a common characteristic (Henry et al. 1990; Henry 1998). Litter build-up favours the growth of 2  1.1. Ecology of grazed ecosystems some plant species while suppressing that of others (Knapp & Seastedt 1986; Willms et al. 2002). Plant growth is often limited because the interception of light is blocked by old and dead plant tissues that have accumulated (Mulder 1999). The accumulation of dead plant material shades the soil surface and may reduce soil temperatures, which may shorten the growing season and delay plant emergence in the spring (Knapp & Seastedt 1986). In addition, soil moisture contents may decrease as dead foliage intercepts water, preventing it from reaching the ground (Knapp & Seastedt 1986). However, litter may conserve soil moisture by reducing evaporation in some grassland ecosystems (Willms et al. 2002). Grazing affects soil temperatures and moisture levels. Soil temperatures increase as grazing decreases litter cover and the extent of the plant canopy, reducing the quantity of insulative material and permitting more solar radiation to reach the soil surface (Bremer et al. 1998; Yates et al. 2000; Vermeire et al. 2005). However, mosses provide additional insulation to the soil surface and this alters soil temperature responses to grazing (van der Wal et al. 2001). Furthermore, the amount of moisture in the soil may rise or fall as a response to grazing. Defoliation can increase soil moisture levels because the removal of aboveground biomass decreases plant transpiration rates (Wan et al. 2002). In contrast, soil moisture levels may decline because of higher soil surface temperatures and evaporation rates (Wan et al. 2002). Trampling by herbivores compacts the soil, which reduces soil aeration (Bakker et al. 2004) and porosity (Yates et al. 2000). It also decreases rainfall infiltration and increases sediment loss in heavily grazed areas; however, these negative effects are less pronounced in areas that are subjected to light to moderate grazing regimes (Warren et al. 1986). There are many adaptations that plants have made in response to herbivory. Briske (1993) stated that leaves generally exhibit maximum photosynthetic rates at about the time of full expansion and this rate declines afterwards. Therefore, the leaves of defoliated plants may display increased rates of photosynthesis compared to non-defoliated plants because 3  1.1. Ecology of grazed ecosystems many of the leaves are chronologically younger and photosynthesize more efficiently (Briske 1993). For example, Peng et al. (2007) found that photosynthetic rates increased from non-grazed to lightly grazed plots in four species grassland plant species in Inner Mongolia. In addition, grazing resistant species tend to branch prolifically and they have a high ratio of vegetative to reproductive shoots (Korte & Harris 1987). Grazing has induced tillering in the tussock-forming Agropyron desertorum (Caldwell et al. 1981) and the rhizomatous A. dasystachyum (Zhang & Romo 1995). The removal of older transpiring leaf tissues may also improve the water balance of remaining plant parts (Mulder 1999).  1.1.1  Effects of grazing on ecosystem processes  Removing moderate amounts of plant material should either maintain or stimulate net primary production (NPP) and shoot N levels according to the grazing optimization hypothesis (McNaughton 1979; Leriche et al. 2003; Sjogersten et al. 2008). The grazing optimization curve in Figure 1.1 predicts higher yields of plant dry mass and N in shoots in moderately grazed ecosystems. Biondini et al. (1998) stated that this relationship applies to both aboveground and belowground NPP. The increase in NPP and shoot N concentrations at moderate grazing levels may be explained by the ability of plants to compensate in their photosynthetic rates (Caldwell et al. 1981; Detling & Painter 1983) and increase tillering and nutrient uptakes after defoliation (McNaughton 1979; Oesterheld & McNaughton 1988). However, compensatory growth does not occur in response to all combinations of environmental variables or in all plant species and systems (Briske 1993). Defoliation is most likely to increase NPP in ecosystems that have low primary production and long evolutionary histories of grazing (Milchunas & Lauenroth 1993; Klein et al. 2007). Studies have investigated the grazing optimization hypothesis in several different ecosystems. Table 1.1 is a summary of five studies that looked at the responses of aboveground NPP to grazing; three agree with the grazing optimization hypothesis and two refute it. 4  1.1. Ecology of grazed ecosystems Grazing stimulated aboveground NPP in the Serengeti plains (McNaughton 1979), nonforested meadows in Yellowstone National Park (Frank & McNaughton 1993), and coastal marshes of the Hudson Bay lowlands (Cargill & Jefferies 1984). However, aboveground NPP did not increase at moderate grazing levels in tundra heath (Olofsson et al. 2001). Responses of belowground biomass to grazing vary. Higher grazing frequencies can decrease belowground biomass levels by both reducing the size of the plant organs that assimilate carbon and by intensifying the re-translocation of root carbohydrates into the growing tips of shoots (Gao et al. 2008). However, grazing may cause belowground plant biomass to increase as higher proportions of roots and rhizomes compared to shoots increase the capacity of plants to tolerate environmental stresses and disturbances (Welker et al. 2004; Gao et al. 2007). Ferraro & Oesterheld (2002) conducted a meta-analysis of 53 studies that showed that root biomass was not significantly affected by different intensities and frequencies of defoliation. Plant respiration contributes to the majority of whole ecosystem respiration in some tundra plant communities (Johnson et al. 2000), whereas levels of microbial respiration are minor (Stark & Grellmann 2002). Therefore, the responses of plant growth and ecosystem respiration to grazing should correspond. For example, grazing did not affect shrub cover or ecosystem respiration in Finnish alpine tundra dominated by dwarf shrubs (Susiluoto et al. 2008). In addition, heavy grazing decreased both root biomass and ecosystem respiration in mixed prairie (Polley et al. 2008). Grazing can either reduce or enhance the rate of nitrogen-cycling (N-cycling). Soil nitrogen availabilities increase as soluble N is added to the soil from the urine and feces of herbivores (Hik et al. 1991; Olofsson et al. 2004). This process accelerates the return of nutrients to the soil, which can increase aboveground NPP (Hik et al. 1991). In addition, frequently grazed sites may have more NH4 + available in the soil because larger amounts of organic N are mineralized in them compared to sites that are not grazed as much (Green & 5  1.2. Grazing in the Arctic Detling 2000). Nitrogen in leaves often increases due to defoliation because roots enhance N uptake, which is preferentially allocated to shoots (Green & Detling 2000). If N in dead plant material increases due to herbivory, litter quality improves and this accelerates nutrient cycling because decomposition rates depend on initial N concentrations (Taylor et al. 1989). For example, reindeer grazing improved litter quality by increasing the N content and decreasing carbon to nitrogen (C/N) ratios in four plant species at a suboceanic tundra site in Norway (Olofsson & Oksanen 2002). In addition, trampling by herbivores accelerates decomposition by compacting litter and increasing its contact with microbe-rich soil (Willms et al. 2002). However, grazing by herbivores may reduce the rate of N-cycling by decreasing the abundance of species with tissues rich in N or by selectively consuming plant species with high quality litter (Ritchie et al. 1998; Bakker et al. 2004).  1.2  Grazing in the Arctic  The dominant grazers in the Arctic are mammals, but other herbivores include birds and insects. The main Arctic mammalian herbivores are caribou, muskoxen, and various rodents (Mulder 1999). Vertebrate herbivores in northern latitudes are generalists in their selection of forage species (Jefferies et al. 1994). The addition of biotic waste to Arctic plant communities by herbivores influences interspecific plant relationships because species vary in their responses to fertilization (Mulder 1999). In the Arctic, herbivores influence species richness by either increasing (Mulder 1999) or decreasing (Henry 1998) habitat heterogeneity. Individual plants in Arctic plant communities respond to grazing events differently. Tundra plants are generally long-lived organisms capable of growing at low temperatures during short growing seasons (Bliss 1997). Many graminoids in the Arctic rely on tillering for reproduction and this allows them to tolerate intense grazing because they can draw from the reserves of nearby ungrazed tillers (Raillard & Svoboda 1999). In addition, grazing 6  1.2. Grazing in the Arctic can affect the reproductive strategy of plants by reducing or eliminating seedlings and by causing tillers to reproduce earlier (Tolvanen & Henry 2000).  1.2.1  Muskox grazing in the Arctic  Muskoxen are endemic to Arctic regions with limited snow cover that permits access to winter forage (Lent 1988). There are large populations of muskoxen in Canada and eastern Greenland and introductions have been made to western Greenland, Norway, Sweden, Alaska, and Russia (Lent 1988; Laikre et al. 1997). Although hunters and explorers reduced the Canadian population of muskoxen in the late 1800s and the early 1900s, numbers of this animal have increased since the establishment of hunting restrictions in 1917 (Tener 1958). The population has increased from an estimated 5 000 animals in 1955 (Tener 1958) to a size of approximately 85 000 animals in 2009 (Canadian Wildlife Service and Canadian Wildlife Federation 2009). Muskoxen prefer to graze in wet sedge meadow plant communities that are dominated by graminoids of the Cyperaceae family such as Carex and Eriophorum. Approximately 3.6% of the circumpolar High Arctic is covered by small patches of sedge meadows (Bliss 1997; Walker et al. 2005) that occur along streams and below seepage slopes (Raillard & Svoboda 2000). Raillard (1992) found that muskoxen spent 83% of their time grazing sedge meadow plant communities, even though those plant communities only covered 31% of the study area. Defoliation of Carex and Eriophorum can result in higher N and phosphorous (P) levels in new leaf material (Chapin 1980). These graminoids are rhizotamous and have large belowground nutrient reserves, which leads to the rapid production of new shoots in response to defoliation. The plant material harvested by muskoxen cycles back into the soil and water in their urine and feces (Raillard & Svoboda 2000). This process increases N and organic matter in soils and it also increases total P in surface water (Henry 1998; Mulder 1999). Because of 7  1.2. Grazing in the Arctic this addition of nutrients, the primary productivity in grazed sedge meadows is significantly higher than that in sedge meadows in nearby sites that have remained ungrazed (Henry & Svoboda 1994; Raillard & Svoboda 2000). There are a limited number of studies of grazing effects by muskoxen. Thing et al. (1987) noted that a herd of 4 100 muskoxen in northeastern Greenland preferred to graze graminoids in the winter and willows in the summer. These authors also reported that the Greenlandic muskoxen utilized mineral licks to get more dietary sodium in the winter and that the relative quality of summer forage was four times greater than that of the winter forage. Other research has been conducted on a small herd of muskoxen in Norwegian alpine tundra. Dryas heath provides the best winter habitat for this herd because these areas have higher graminoid availabilities and shallower snow densities compared to surrounding plant communities (Nellemann 1998). Muskoxen are not ubiquitous in the Canadian High Arctic and studies have been conducted in both grazed and ungrazed sedge meadows. Table 1.2 is a summary of five studies that have been done at Sverdrup Pass (SP) and Alexandra Fiord (AF) on Ellesmere Island, Canada. The AF lowland is unique because it may be that the relative isolation and small size of the site prevents the establishment of large viable population of muskoxen (Henry et al. 1986a; Muc et al. 1994a). However, SP sustains a herd of between 25-75 muskoxen (Raillard & Svoboda 2000; Henry, unpublished data) because it is a much larger site with more available forage. Three of the five studies in Table 1.2 investigated grazing. The comparisons of sedge meadow plant communities at AF and SP by Henry (1998) and Henry & Svoboda (1994) indicated that the grazed meadows at SP had more root and rhizome biomass, higher NO3 − and cation exchange capacities in the soil, greater shoot densities of Carex aquatilis ssp. stans, and lower litter cover. Additionally, Raillard & Svoboda (1999) showed that plants compensated in growth and increased N concentrations in their shoots when defoliated 8  1.2. Grazing in the Arctic at SP. The ungrazed meadows at AF had large standing crops of dead biomass and net production of graminoids and forbs increased with high levels of fertilizer application in the meadows (Table 1.2). Grazing by muskoxen at SP decreases the accumulation of standing dead tissue and may increase available nutrients for plant growth (Raillard 1992; Henry & Svoboda 1994; Henry 1998). Comparing the grazed meadows at SP to the ungrazed meadows at AF presents an opportunity to expand the understanding of the effects of muskox grazing on soil parameters, net primary production, and nutrient cycling in High Arctic sedge meadow plant communities. There have been preliminary comparisons of the grazed and ungrazed sedge meadows at AF and SP (Henry & Svoboda 1994; Henry 1998); however, there is much scope for additional research on the details of the differences between the two locations. A clipping experiment in the ungrazed meadows at AF will allow us to determine the responses of the ungrazed sedge meadows to defoliation, which will complement the studies conducted in the grazed meadows at SP (Raillard 1992; Raillard & Svoboda 1999).  9  1.3. Site descriptions  1.3  Site descriptions  The two field sites chosen for this study are both located on Ellesmere Island, which is the northernmost island in the Canadian Arctic Archipelago (Figure 1.2). Alexandra Fiord (78◦ 53’N, 75◦ 55’W) and Sverdrup Pass (79◦ 08’N, 79◦ 38’W) are about 65 km apart and they have similar geological and climatic characteristics (Raillard 1992; Henry 1998). Both of these sites are polar oases that have warmer climates, higher species diversity, greater plant cover, and higher primary productivity than the surrounding polar desert (Freedman et al. 1994; Raillard & Svoboda 1999). Wet sedge meadow plant communities are on gleysolic cryosol soils at Alexandra Fiord and Sverdrup Pass (Bergeron 1989; Muc et al. 1994b).  1.3.1  Alexandra Fiord  The entire experimental study and part of the observational study was conducted at the Alexandra Fiord (AF) lowland, which is located on the north side of the Johan Peninsula on the east-central coast of Ellesmere Island (Figure 1.2). There are glaciers to the south and the water of Alexandra Fiord to the north of the lowland. In addition, polar desert uplands (500-700 m) are located to the east and west of this site. Mammalian grazers are rare along the north coast of the Johan Peninsula and populations have likely never been large in the AF lowland region. Hunting by explorers in the late 1800s and the early 1900s decimated the eastern Ellesmere Island muskox population and numbers have not recovered (Henry et al. 1986b; Freedman et al. 1994; Dick 2001). The absence of muskoxen at AF is likely because of this excessive hunting coupled with the geographical isolation and relative small size of the site (Henry et al. 1986b; Muc et al. 1989; Henry 1998). Wet sedge meadows cover approximately 20% of the AF lowland and they correspond to the S-CP-DS (Sedge-Cushion Plant-Dwarf Shrub) plant community in Figure 1.3. These wet meadows are found on flat sites that have impeded drainage and the dominant plants include 10  1.3. Site descriptions Eriophorum angustifolium, Carex aquatilis ssp. stans, C. membranacea, Dryas integrifolia, Vaccinium uliginosum, and Salix arctica (Henry et al. 1990; Muc et al. 1994a). The two black circles in Figure 1.3 indicate the sites for the clipping experiments and observational studies at AF. In addition to these sites, aboveground biomass was harvested from the two wet sedge meadows represented by squares for analysis of shoot nitrogen and carbon concentrations.  1.3.2  Sverdrup Pass  Sverdup Pass (SP) is a 75 km long valley orientated in a west-east direction that connects two fiords on Ellesmere Island (Figure 1.2). Sverdup Pass functions as a corridor for muskoxen as they move between the large grazing ranges on either side of the island (Raillard & Svoboda 2000). However, it is believed there is a resident population of muskoxen that remain in SP (Henry et al. 1986a; Raillard 1992). From 1986 to 1989, the valley had a population of about 50 muskoxen (Raillard & Svoboda 1999), but the population has varied between 22 and 72 animals since 1981 (Henry, pers. obs). The valley bottom of Sverdrup Pass is a river plain bordered by gentle seepage slopes, steep cliffs, and mountains up to 1200 metres in height (Raillard & Svoboda 2000). This geography provides suitable habitat for wet sedge meadows dominated by C. stans (Bergeron 1989). Bergeron (1989) classified these wet sedge meadows as C. aquatilis - Bryophyte plant communities. The black circles in Figure 1.4 indicate the locations of the two sites selected as grazed sedge meadows for the observational study. The black squares represent the two additional sites where aboveground plant biomass was harvested to determine shoot nitrogen and carbon concentrations.  11  1.4. Tables and Figures  1.4  Tables and Figures  Table 1.1: Results of selected studies of the response of aboveground NPP to grazing Authors and Year McNaughton (1979) Frank & McNaughton (1993) Cargill & Jefferies (1984)  Olofsson et. al (2001)  Ecosystem Serengeti-Maro region of Tanzania and Kenya - open Serengeti Plains; dominated by Andropogon greenwayi Nonforested habitat in Yellowstone, Wyoming Coastal marshes of the Hudson Bay lowlands, Canada (La P´erouse Bay) dominated by Puccinellia phryganodes and Carex subspathacea Tundra heath vegetation in northern Norway  Herbivores Wildebeest, zebra, Thomson’s gazelle Elk and bison  Results - moderate grazing stimulated productivity up to two times that of ungrazed control plots - productivity depended on moisture availability - aboveground production in grazed areas was 47% higher than in ungrazed areas  Lesser Snow Geese  - grazing significantly increased aboveground NPP  Reindeer  - highest NPP in a heavily grazed area - less NPP in moderately grazed areas  12  1.4. Tables and Figures  Table 1.2: Previous studies conducted on ungrazed sedge meadow plant communities at Alexandra Fiord and grazed sedge meadow plant communities at Sverdrup Pass Authors, Year, and Location Results - net N accumulation in biomass increased Raillard & Svoboda (1999) with increased frequency of clipping - compensation of biomass production -Sverdrup Pass occurred when C. aquatilis ssp. stans plants were clipped to 1.5 cm in height - the total standing crop in ungrazed sedge meadows Henry et al. (1990) was 1400 to 3200 g m−2 ; 50 to 80% of this was attached dead belowground biomass - Alexandra Fiord - average aboveground NPP was 28 g m−2 y−1 Henry (1998) - grazing increased moss cover - Alexandra Fiord cation exchange capacities, and and soil NO3 − , Sverdrup Pass but decreased litter cover - net production of graminoids and forbs Henry et al. (1986b) was greatest with high rates - Alexandra Fiord of fertilizer application Compared to ungrazed meadows, the grazed meadows had: Henry & Svoboda (1994) - more total soil N and organic matter - higher total P concentrations in surface waters - Alexandra Fiord - greater shoot density of C. stans and - greater concentration of P in the Sverdrup Pass leaves of the major sedge species  13  1.4. Tables and Figures  Figure 1.1: Grazing optimization curve (adapted from Belsky 1986).  14  1.4. Tables and Figures  Figure 1.2: Map of Ellesmere Island and adjacent islands.  15  1.4. Tables and Figures  Figure 1.3: Plant communities and the locations of the four study sites at Alexandra Fiord (from Muc et al. 1989). Circles indicate the locations of the clipping experiments and observational studies. The two squares indicate the additional sites where plants were harvested to measure shoot carbon and nitrogen concentrations.  16  1.4. Tables and Figures  Figure 1.4: Plant communities and the locations of the four study sites at Sverdrup Pass (modified from Bergeron 1989). Circles indicate the locations of the meadows where plant biomass samples were obtained for the observational studies. The two squares indicate the additional sites where plants were harvested to measure shoot carbon and nitrogen concentrations.  17  References Adler, PB, Raff, DA, & Lauenroth, WK. 2001. The effect of grazing on the spatial heterogeneity of vegetation. Oecologia, 128(4), 465–479. Arnold, GW. 1987. Grazing behaviour. Pages 129–136 of: Snaydon, RW (ed), Ecosystems of the world, 17B. Managed grasslands: analytical studies. Amsterdam: Elsevier Science Publishers B.V. Bakker, ES, Olff, H, Boekhoff, M, Gleichman, JM, & Berendse, F. 2004. Impact of herbivores on nitrogen cycling: contrasting effects of small and large species. Oecologia, 138(1), 91– 101. Barry, RG, Courtin, GM, & Labine, C. 1981. Tundra climates. 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Freedman, B, Svoboda, J, & Henry, GHR. 1994. Alexandra Fiord - an ecological oasis in the polar desert. Pages 1–9 of: Svoboda, J, & Freedman, B (eds), Ecology of a polar oasis: Alexandra Fiord, Ellesmere Island, Canada. Toronto: Captus Press Inc. Gao, Y, Luo, P, Wu, N, Yi, S, & Chen, H. 2007. Biomass and nitrogen responses to grazing intensity in an alpine meadow on the eastern Tibetan Plateau. Polish Journal of Ecology, 55(3), 469–479. Gao, YZ, Giese, M, Lin, S, Sattelmacher, B, Zhao, Y, & Brueck, H. 2008. Belowground net primary productivity and biomass allocation of a grassland in Inner Mongolia is affected by grazing intensity. Plant and Soil, 307(1-2), 41–50. 19  Chapter 1. References Gauthier, G, Hughes, RJ, Reed, A, Beaulieu, J, & Rochefort, L. 1995. Effect of grazing by greater snow geese on the production of graminoids at an arctic site (Bylot Island, NWT, Canada). Journal of Ecology, 83(4), 653–664. Green, RA, & Detling, JK. 2000. 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High grazing impact, selectivity, and local density of muskoxen in Cental Ellesmere Island, Canadian High Arctic. Arctic Antarctic and Alpine Research, 32(3), 278–285. Ritchie, ME, Tilman, D, & Knops, JMH. 1998. Herbivore effects on plant and nitrogen dynamics in oak savanna. Ecology, 79(1), 165–177. Sjogersten, S, van der Wal, R, & Woodin, SJ. 2008. Habitat type determines herbivory controls over CO2 fluxes in a warmer arctic. Ecology, 89(8), 2103–2116. Stark, S, & Grellmann, D. 2002. Soil microbial responses to herbivory in an arctic tundra heath at two levels of nutrient availability. Ecology, 83(10), 2736–2744. Struth, JW. 1991. Foraging behavior. Pages 65–83 of: Struth, JW (ed), Grazing management: an ecological perspective. Portland, Oregon: Timber Press Inc. Susiluoto, Sanna, Rasilo, Terhi, Pumpanen, Jukka, & Berninger, Frank. 2008. Effects of grazing on the vegetation structure and carbon dioxide exchange of a fennoscandian fell ecosystem. 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Alpine grassland CO2 exchange and nitrogen cycling: Grazing history effects, medicine bow range, Wyoming, USA. Arctic Antarctic and Alpine Research, 36(1), 11–20. Willms, WD, Dormaar, JF, Adams, BW, & Douwes, HE. 2002. Response of the mixed prairie to protection from grazing. Journal of Range Management, 55(3), 210–216. Yates, CJ, Norton, DA, & Hobbs, RJ. 2000. Grazing effects on plant cover, soil and microclimate in fragmented woodlands in south-western Australia: implications for restoration. Austral Ecology, 25(1), 36–47. 23  Chapter 1. References Zhang, J, & Romo, JT. 1995. Impacts of defoliation on tiller production and survival in northern wheatgrass. Journal of Range Management, 48(2), 115–120.  24  Chapter 2 Evidence of increased plant growth and nitrogen cycling as a result of muskox grazing in High Arctic sedge meadow plant communities. 2.1  1  Introduction  Few plant species can tolerate the growing conditions of the High Arctic because of low temperatures, precipitation, and solar radiation levels (Barry et al. 1981; Edlund & Alt 1989; Freedman et al. 1994; Tolvanen & Henry 2001). In addition, nutrient availabilities are restricted because nitrogen (N) deposition and fixation rates are low and the cold soils of this region have slow decomposition and mineralization rates (Jefferies et al. 1994; Hobbie et al. 2002). Because of these factors, net primary production (NPP) in Arctic ecosystems is low compared to other terrestrial systems (Callaghan et al. 2004). Herbivory can alter ecosystem function (Leriche et al. 2001). Defecation and urination by grazing animals accelerates the return of nutrients to the soil, which can increase above1  A version of this chapter will be submitted for publication. Elliott, TE, & Henry, GHR. 2009. Evidence of increased plant growth and nitrogen cycling as a result of muskox grazing in High Arctic sedge meadow plant communities.  25  2.1. Introduction ground NPP (Hik et al. 1991). In addition, there may be more ammonium available in the soil of frequently grazed sites because larger amounts of organic N are mineralized in them compared to sites that are less frequently grazed (Green & Detling 2000). Herbivory changes the quantity and quality of plant litter, which affects decomposition processes (Olofsson et al. 2004). The concentration of carbon-based defensive compounds in leaves can decrease as soil nutrient availabilities increase (Dormann 2003). Additionally, carbon (C) may be diluted in the fresh biomass of leaves during early regrowth following grazing or by plant growth throughout the growing season (Chapin et al. 1986; deVisser et al. 1997). Grazed tundra systems can have higher nutrient turnover rates in the rooting zone of plants (Jefferies et al. 1994), but often there is a net immobilization of N in soils during the growing season by microbes (Schimel & Chapin 1996). Carbon dioxide (C02 ) flux may increase because of higher nutrient availabilities in grazed sites than ungrazed sites (Sjogersten et al. 2008); however, heavy grazing may decrease both root biomass and ecosystem respiration (Polley et al. 2008). The grazing optimization hypothesis predicts that removing moderate amounts of plant material should either maintain or increase NPP and nitrogen (N) in vegetation (McNaughton 1979; Leriche et al. 2003; Sjogersten et al. 2008). These patterns have been observed in communities dominated by coral, algae, and terrestrial graminoids (Hik & Jefferies 1990). This relation applies to both aboveground and belowground NPP (Biondini et al. 1998). However, compensatory growth does not occur in all plant species and ecosystems (Briske 1993). For example, defoliation is most likely to increase NPP in ecosystems that have low primary production and long evolutionary histories of grazing (Milchunas & Lauenroth 1993; Klein et al. 2007). The underlying process of compensatory growth may involve grazing tolerant species increasing their nutrient uptake, relative growth rates, and tissue nutrient concentrations in response to grazing (Ritchie et al. 1998). However, nutrient cycling may be reduced as herbivores preferentially select plants with nutrient-rich 26  2.1. Introduction tissues, which increases the abundance of plants with slowly decomposing nutrient-poor tissues (Vitousek & Howarth 1991; Ritchie et al. 1998). Evidence in support of the grazing optimization hypothesis is mixed. Grazing stimulated aboveground NPP in grasslands (McNaughton 1979; Frank & McNaughton 1993) and coastal marshes (Cargill & Jefferies 1984b). However, NPP did not increase at moderate grazing levels in tundra heath (Olofsson et al. 2001) and oak savannas (Ritchie et al. 1998). Belowground plant production in Serengeti grasslands was reduced because of a pronounced diversion of carbohydrates away from roots following defoliation (McNaughton 1979). In the High Arctic, evidence exists of herbivory aiding nutrient cycling (Henry 1998) and stimulating NPP in wet sedge meadows (Henry & Svoboda 1994). Muskoxen (Ovibos moschatus) are one of two large mammalian herbivores that can endure conditions in the Canadian High Arctic. Muskoxen prefer to graze in wet sedge meadow plant communities that are dominated by graminoids of the Cyperaceae family such as Carex and Eriophorum. Approximately 3.6% of the circumpolar High Arctic is covered by small patches of sedge meadows (Bliss 1997; Walker et al. 2005), which typically occur along streams and below seepage slopes (Muc 1977; Sheard & Geale 1983; Raillard & Svoboda 2000). Muskoxen spent 83% of their time grazing sedge meadow plant communities during an intensive study conducted on Ellesmere Island, even though those plant communities only covered 31% of the study area (Raillard 1992). Research has also examined the effects of grazing on some Arctic plant species. Increased defoliation of Carex aquatilis ssp. aquatilis and Eriophorum vaginatum resulted in higher N levels in new leaf material and lower shoot C:N (Chapin 1980; Jefferies et al. 1994). Large belowground nutrient reserves accumulated and production of new shoots increased in these two rhizomatous graminoid species in response to leaf removal (Chapin 1980). Nutrients from the plant material that is harvested by muskoxen cycle back into the soil and water in their urine and feces (Raillard & Svoboda 2000). This process increases N and 27  2.2. Methods organic matter in soils and it also increases total phosphorous (P) in surface water (Henry 1998; Mulder 1999). Because of this addition of nutrients, the productivity in grazed sedge meadows can be significantly higher than that in sedge meadows in nearby lowlands that have remained ungrazed (Henry & Svoboda 1994; Henry 1998; Raillard & Svoboda 2000). Nitrate (N03 − ) and ammonium (NH4 + ) are valuable forms of N for tundra plants (Cargill & Jefferies 1984a), but Arctic plants can also absorb organic N directly (Schimel & Chapin 1996; Kielland et al. 2007). Ammonium is immobilized by microbes in Arctic soils so that plants have restricted access to this nutrient (Schimel & Chapin 1996). Slow rates of N mineralization due to unfavourable environmental conditions in Arctic ecosystems cause the accumulation of organic N (Henry & Jefferies 2002). The objective of this study was to test the grazing optimization hypothesis in High Arctic sedge meadow plant communities. We hypothesized that moderate levels of grazing would increase NPP and shoot N concentrations in these low productivity, resource-poor plant communities. To test this hypothesis, we measured plant biomass, ecosystem respiration, shoot N and C concentrations, and soil N availability in moderately grazed sedge meadows and meadows that have remained ungrazed due to historical and geographical constraints.  2.2 2.2.1  Methods Site descriptions  The two field sites chosen for this study were both located on Ellesmere Island, which is the northernmost island in the Canadian Arctic Archipelago. Alexandra Fiord (AF) (78◦ 53’N, 75◦ 55’W) and Sverdrup Pass (SP) (79◦ 08’N, 79◦ 38’W) are about 65 km apart and they have similar geological and climatic characteristics (Raillard 1992; Henry 1998). Both of these sites are polar oases that have warmer climates, higher species diversity, greater plant cover, and higher primary productivity than the surrounding areas (Freedman et al. 1994; Raillard 28  2.2. Methods & Svoboda 1999). Wet sedge meadow plant communities occur on gleysolic cryosol soils at both sites (Bergeron 1989; Henry et al. 1990; Muc et al. 1994b). These wet meadows are found on flat areas with impeded drainage at the base of seepage slopes at AF and along river plains at both sites (Muc et al. 1994a; Raillard & Svoboda 2000). From 1986 to 1989, SP had a population of about 50 muskoxen (Raillard & Svoboda 1999) and the population has varied between 22 and 72 animals since 1981 (Henry, pers. obs). However, mammalian grazers are not a significant component of the AF lowland and have not been for at least 100 years. Hunting by explorers in the late 1800s and the early 1900s decimated the Ellesmere Island muskox population and numbers have not recovered (Henry et al. 1986b; Freedman et al. 1994; Dick 2001). The absence of muskoxen at AF is likely because of this excessive hunting coupled with the geographical isolation and relative small size of the site (Henry et al. 1986a; Muc et al. 1989; Henry 1998).  2.2.2  Aboveground NPP  In 2007 and 2008, plant shoots and reproductive structures were harvested to measure aboveground NPP. Relatively homogeneous areas within two sedge meadows were selected at AF in 2007. To capture variability in production, ten 20 cm x 50 cm quadrats were tossed haphazardly into each of these areas and all plant material within each quadrat was cut to the ground or moss surface and placed into a paper bag. Directly after snowmelt both years (June 12-14), overwinter green plant material was harvested to determine the amount of plant biomass carried-over from the previous year. Following the methods of Henry et al. (1990), annual NPP was calculated by subtracting these overwinter green weights from the peak production values at both AF and SP. Due to logistical constraints, overwinter green biomass could not be measured at SP. Therefore, it is assumed that the overwinter green values were similar for AF and SP. Aboveground plant biomass was harvested at the peak of live aboveground biomass 29  2.2. Methods (peak season), which is the last week of July to the first week of August in High Arctic plant communities (Muc 1977). At the peak of live aboveground biomass in 2007, all aboveground vegetation was harvested from two grazed sedge meadows at SP and two of the ungrazed meadows at AF. A similar sampling procedure was followed in 2008; however, sedge meadows at SP were chosen that had not been grazed throughout the growing season. This was determined by selecting sedge meadows at SP that had plants with attached dead plant material from previous growing seasons. In the field lab, all plant material within each quadrat was sorted to species, dried, and then sent to the lab at UBC. In the lab, the samples were dried again for 24 to 30 hours at 60◦ C before weighing. The following categories were then formed by combining sample weights: • green - all photosynthetic plant tissue and new stem growth in shrubs; • non-green - secondary stem growth from shrubs; • dead - litter and dead plant material attached to new growth; and • standing crop - sum of green, non-green, and dead plant material. All values were converted to g m−2 prior to analysis. The net production of dwarf shrubs was estimated using the following adjustments adapted from Henry et al. (1990): • for Salix arctica and Vaccinium uliginosum: 100% of peak season green biomass + 25% of this value for stem diametre growth; • for Cassiope tetragona: 25% of peak season green biomass because green leaves function for four years; and • for Dryas integrifolia and Saxifraga oppositifolia: 50% of green peak season biomass because leaves function for two years. 30  2.2. Methods  2.2.3  Belowground biomass  In 2007, plant roots and rhizomes were harvested to calculate belowground biomass. Immediately after the peak season harvest in 2007, one 6-cm diametre soil core was dug to a depth of 20 cm in the approximate centre of each of the aboveground quadrats at both sites. Loose soil was removed by gently washing the samples in water before rinsing them over a 2 mm sieve. The belowground biomass was collected in plastic bags and maintained in coolers. In the lab, the belowground material was sorted according to the following categories after Henry et al. (1990): • live roots - white to tannish, with lateral roots, turgid; • dead roots - brown, with lateral roots, not turgid; • live rhizomes - glossy, brown, turgid, no lateral roots; • dead rhizomes - not glossy, dark, not turgid, no lateral roots; • aboveground material - leaves; and • wood. Large material was first removed from each sample, sorted according to the above categories, and the remaining sample was divided into four visually equal subsamples. The belowground material from one of the four subsamples was then sorted into the six aforementioned categories and weighed. To calculate belowground biomass, the proportion of each of the six categories to the subsample was first determined. These proportions were then multiplied to the weight of the unsorted sample and the values for the large material were added. Belowground biomass in 2007 was determined by combining live root and rhizome weights, whereas belowground standing crop was calculated by also considering the mass of dead roots and rhizomes. Logistical constraints prevented the measurement of belowground biomass in 2008.  31  2.2. Methods  2.2.4  Shoot nitrogen and carbon concentrations  Shoot nitrogen and carbon concentrations were measured in Carex aquatilis ssp. stans, Eriophorum angustifolium ssp. triste, and Salix arctica from four wet sedge meadows at both AF and SP in 2008. During peak season, aboveground biomass was harvested from ten 20 cm x 50 cm quadrats at each site and sorted to species. Samples were then created from plant material that was selected from five plots at each site because all three species were not present in every quadrat. In addition, data from only three sites (n=3) were used for C. stans and two sites (n=2) for S. arctica at both AF and SP because of the lack of these species in all four wet sedge meadows. Therefore, there were 15 samples of C. stans, 20 samples of E. triste, and 10 samples of S. arctica from AF and SP. For the three plant species, samples were taken from each of the selected plots from approximately 50% of the shoot material harvested at peak season. A Wiley Mill with mesh size #4 was used to grind larger shoot material from the samples into a powder, whereas a mortar and pestle with liquid N was utilized for smaller quantities. Nitrogen and carbon concentrations were then measured by dry combustion with a Carlo-Erba NA-1500 NCS analyzer at the Analytical Chemistry Laboratory of the BC Ministry of Forests and Range in Victoria.  2.2.5  Soil N03 − and NH4 + availability  During peak season in 2008, soil N03 − and NH4 + availability at AF and SP was measured with Plant Root Simulators (PRS). These ion exchange membranes show ion flux in the soil and other heterogeneous media by simulating nutrient absorption by plant roots (Western Ag Innovations 2008). In both the grazed and ungrazed sites, four visually homogeneous sedge meadows of similar size were chosen and a transect was created through the centre of each area. Within a four day period, ten PRS were inserted vertically approximately 7 m apart along each transect and left untouched for about 24 hours. The probes were washed 32  2.2. Methods with deionized water and scrubbed with a coarse brush to remove residue shortly after retrieval. They were then sent to Western Ag Innovations in Saskatoon for ion extraction and analysis.  2.2.6  Ecosystem respiration  Ecosystem respiration was measured with a closed flow infrared gas analyzer (IRGA; Li840, Li-Cor Inc., Lincoln, NE, USA) in the two meadows from which aboveground NPP was measured at both AF and SP. During the period of maximum aboveground biomass in 2008, twelve 10.16-cm diameter x 7.62-cm high polyvinyl chloride (PVC) collars were placed approximately seven metres apart along a transect established in a representative area of each of the sedge meadows at SP (July 22) and AF (July 24). After placement, the collars were left untouched for approximately 24 hours so that soil microbial respiration would return to normal levels after disturbance. The IRGA chamber was then placed over each PVC collar and air was mixed within the chamber during measurement (Jassal et al. 2007; Myers-Smith et al. 2007). After steady state was achieved within the chamber, two consecutive measurements of CO2 were taken in each plot over a 120 second period. The chamber was aerated between each of the two measurements. These two values were then averaged to calculate ecosystem respiration. Temperatures inside the chamber were recorded every second and factored into flux calculations. All measurements were conducted during the day between 10:00 and 19:00 and sampling along a transect took about 1.5 hours. Weather conditions were consistent during the sampling period.  2.2.7  Statistical analysis  Analysis of variance (ANOVA) was preformed using PROC GLM in SAS (version 9.1) to test for differences in ecosystem variables between the grazed meadows at SP and the ungrazed meadows at AF. This statistical technique fits general linear models by using least squares 33  2.3. Results regression to analyze variance (SAS Institute Inc. 2009). In all of the models that tested for differences in variables between SP and AF, site (grazed or ungrazed) was considered a fixed effect and meadow nested within site a random effect. Box-Cox transformations (Marazzi & Yohai 2006) in SAS were used to recommend the best possible transformation for the analysis of belowground biomass. To compare NPP and ecosystem respiration between the grazed and ungrazed meadows, data were normalized by first ranking the data with PROC RANK and then normal scores for these ranks were computed with the BLOM procedure (SAS Institute Inc. 2009). In addition, ranks were computed and data normalized to compare non-green, dead, and standing crop plant material between the grazed and ungrazed meadows. Analysis of variance was also utilized to test for differences in aboveground plant biomass between 2007 and 2008. In those models, year was considered a fixed effect and meadow a random effect. No data transformations were required to test for differences between the two years.  2.3 2.3.1  Results Aboveground NPP  Aboveground NPP values were similar at AF and SP in 2007 (P = 0.68) (Figure 2.1). However, there was marginally more aboveground NPP at SP in 2008 (P = 0.09). Overwinter green biomass at AF (P = 0.70), aboveground NPP at AF (P = 0.41), and aboveground NPP at SP (P = 0.15) did not differ significantly when compared between the two years (Figure 2.2). In both study years, NPP was predominately from graminoids at AF and SP (Figure 2.1). Graminoid production was similar at both sites in 2007 (P = 0.67). In contrast, the sedge meadows at SP had marginally higher graminoid NPP values in 2008 (P = 0.09). 34  2.3. Results There was marginally more total standing crop (P= 0.06) and dead plant material (P = 0.07) in the ungrazed meadows of AF in 2007 (Figure 2.3). However, there was no significant difference in non-green plant material between the two sites in 2007 (P = 0.70). Total standing crop (P = 0.26), non-green (P = 0.14), and dead plant material (P = 0.10) were similar in the grazed meadows at SP and the ungrazed meadows at AF in 2008.  2.3.2  Belowground biomass  Belowground standing crop and live biomass were higher in the grazed meadows at SP (Figure 2.4). Belowground live biomass was 3.5 times higher in the grazed sites compared to that of the ungrazed sites (P = 0.03). In addition, belowground standing crop at SP (Mean ± S.E.) was 1260.4 ± 104.8 g m−2 compared to 491.4 ± 88.2 g m−2 at AF (P = 0.04).  2.3.3  Shoot nitrogen and carbon concentrations  There were higher concentrations of N in the shoots of C. stans (P = 0.02) and E. triste (P = 0.05) from the grazed site than in the ungrazed site at AF (Figure 2.5). However, there was no significant difference in N concentrations in S. arctica between the two sites (P = 0.91). Carbon concentrations were similar between the two sites in C. stans (P = 0.14), E. triste (P = 0.70), and S. arctica (P = 0.62).  2.3.4  Soil N03 − and NH4 + availability  Nitrogen supply rates over a 24 hour period indicated very low levels of soil N03 − at both SP and AF in 2008 (Figure 2.7). The two meadows had similar supply rates of N03 − (P = 0.77). However, supply rates of NH4 + to the ion exchange membranes were much higher (Figure 2.7) and the values were significantly larger at SP compared to AF (P < 0.01). 35  2.4. Discussion  2.3.5  Ecosystem respiration  Mean ecosystem respiration during peak season was 0.6 µmol m−2 s−1 lower in ungrazed meadows (Figure 2.8). However, this difference was not significant (P = 0.17).  2.4  Discussion  In this study, the grazing optimization hypothesis was supported by increased belowground production, greater aboveground NPP in 2008, higher N concentrations in C. stans and E. triste shoots, and greater soil NH4 + levels in the grazed sedge meadows at SP compared to the ungrazed meadows at AF. The observed difference in belowground biomass between the grazed and ungrazed meadows supports the grazing optimization hypothesis. The results from this study confirm those from research in the early 1980’s where the ungrazed meadows at AF had less root and rhizome biomass compared to grazed meadows at SP (Henry & Svoboda 1994). Similarly, van der Maarel & Titlyanova (1989) reported that grazed steppes in Sweden had more belowground standing crop compared to ungrazed Siberian steppes. Results from other studies on belowground production have shown decreases in belowground biomass in response to grazing. Root biomass was lower in grazed shortgrass, midgrass, and tallgrass prairies, but the difference was only significant in the midgrass prairie (Derner et al. 2006). In dry tropical savannas in India, belowground NPP decreased in grazed sites by 25 - 65% (Pandey & Singh 1992). McNaughton et al. (1998) remarked that long term belowground productivity in Serengeti grasslands was not affected by grazing intensity. Relative to the ecosystems mentioned above, shoot to root ratios in wet tundra plant communities are low because of restricted nutrient availabilities (Bliss et al. 1973; Wielgolaski et al. 1981). Because of this, grazing may enhance compensatory belowground production in these systems as plants allocate more growth to roots in order 36  2.4. Discussion to seek more nutrients. Aboveground NPP was marginally higher at SP than AF in 2008, which suggests overcompensation in plant growth due to grazing. This result is supported by studies in Low Arctic coastal salt marsh plant communities (Cargill & Jefferies 1984b), open meadows in Yellowstone National Park (Frank & McNaughton 1993), and in Serengeti grasslands (McNaughton 1979). However, muskox grazing decreased aboveground NPP in E. triste compared to ungrazed treatments on Banks Island, Northwest Territories (Smith 1996). In our study, there was more aboveground biomass at SP in 2008 compared to 2007; however, this difference was not significant. In contrast, aboveground NPP was similar at SP and AF in 2007. It was sunnier and warmer in 2007 compared to 2008 at both sites during our study. These different growing conditions did not affect peak season aboveground plant biomass at AF, contrasting with the results from SP. Larter & Nagy (2003) found differences in late-August live standing crop between years in areas grazed by muskoxen on Banks Island. They attributed this difference to annual differences in grazing pressures affecting the removal of plant material and annual variations in environmental conditions. We attempted to select meadows at SP that had not yet been grazed each year, but annual aboveground biomass in 2007 may have been reduced by a grazing event earlier in the growing season since it was not possible to observe muskox grazing patterns throughout the entire year. Graminoid NPP was higher in the grazed sedge meadow, where over 80% of NPP was from graminoids, which is within the range reported for these plant communities (Henry et al. 1990). Graminoid-dominated xeric grasslands in the eastern USA grazed by elk (Stewart et al. 2006) and sub-arctic salt marshes dominated by Puccinellia phryganodes and C. subspathacea grazed by lesser snow geese (Cargill & Jefferies 1984b) also had higher NPP values than ungrazed areas. In contrast, graminoid NPP was lower in grasslands dominated by Agropyron smithii that were colonized by prairie dogs compared to areas 37  2.4. Discussion protected from grazing by exclosures (Detling & Painter 1983). According to the results of our study, grazing increased NH4 + availability in High Arctic sedge meadow plant communities. Higher levels of soil NH4 + may be explained by the deposition of wastes from the relatively high density of muskoxen at SP (Raillard & Svoboda 1999). Murray (1991) noted that muskox feces contributed 0.07 g m−2 yr−1 of total N to wet meadow plant communities at SP; however, most NO3 − was leached out of the system within 48 hours of deposition. Along with significantly higher levels of soil NH4 + in the grazed meadows, N concentrations in shoots of C. stans and E. triste were higher at SP relative to the ungrazed meadows at AF. Increases in soil NH4 + and plant N in new shoot growth have also been correlated in previous studies on reindeer grazing (Olofsson et al. 2004). In contrast, N concentrations in S. arctica leaves were similar at both sites. Levels of carbon-based secondary metabolites in deciduous shrubs can rise due to herbivory causing declines in leaf N concentrations in nutrient-poor environments (Tuomi et al. 1984). There were marginally higher amounts of dead plant material and total standing crop at AF in 2007. Slowly decomposing unattached and attached plant litter accumulates due to the lack of herbivory at this site (Henry 1998). In experiments conducted in the mixed prairie, litter accumulated in areas protected from grazing by exclosures (Willms et al. 2002). Similarly, grazing prevented the accumulation of litter in mountain snowbeds in Finland (Virtanen 2000). Ecosystem respiration was nonsignificantly higher in the grazed meadows at SP than in the ungrazed meadows at AF. Plant respiration contributed to the majority of whole ecosystem respiration in wet sedge meadows in other tundra studies (Johnson et al. 2000), whereas levels of microbial respiration have been reported to be minor (Stark & Grellmann 2002). For example, grazing did not affect shrub cover or ecosystem respiration in Finnish alpine tundra dominated by dwarf shrubs (Susiluoto et al. 2008). Furthermore, heavy grazing decreased both root biomass and ecosystem respiration in mixed prairie sites (Polley 38  2.4. Discussion et al. 2008). Soil respiration and belowground biomass in ungrazed exclosures were higher than in grazed areas in the steppes of Inner Mongolia, suggesting that root respiration was the primary contributer to soil respiration (Jia et al. 2007). In our study, the higher amounts of plant biomass in the grazed meadows at SP possibly contributed to increased rates of ecosystem respiration compared to the ungrazed meadows at AF. However, our ecosystem respiration data has limited power because of the low number of sites sampled and measurements were conducted on only one day at both AF and SP. Higher aboveground NPP and belowground biomass values can be related to increased soil N availability in the grazed meadows at SP compared to those at AF. Nitrogen is often a limiting factor in tundra ecosystems (Cargill & Jefferies 1984b; Chapin et al. 1995; Henry 1998; van der Wal et al. 2004), but the addition of urine and feces to the soil by grazing muskoxen increases the availability of N in the soil (Raillard & Svoboda 1999). Therefore, plant growth is less restricted by N limitation and primary production increases. These interpretations must consider possible shortcomings related to the observational nature of this study. Even though an attempt was made to sample within a period of a few days, SP and AF are approximately 65 km apart. Abiotic factors such as air temperature, incoming radiation, and soil types are similar, but not the same at these two sites. Other limitations in this study are limited sampling opportunities at SP and the lack of sampling sites. Overwinter green biomass was only measured at AF since it was not possible to visit SP at that point of the growing season due to logistical constraints. This study consisted of only one grazed and one ungrazed site. This restricts the nature of the conclusions that can be made. High Arctic sedge meadow plant communities may respond to grazing by overcompensation in growth as indicated by higher belowground production and aboveground NPP at the grazed site. To elucidate the underlying ecosystem processes involved with these patterns, future research is required in these plant communities. At the global scale, it is necessary 39  2.4. Discussion that we discern the effects of herbivory on C balances as the Earth’s climate continues to change.  40  2.5. Figures  2.5  Figures  Figure 2.1: Aboveground NPP (annual plant growth on a dry weight basis) and graminoid NPP in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) in 2007 and 2008. Vertical bars indicate standard errors and significantly different means are denoted by ∗∗ (P < 0.05) and ∗ (P < 0.10).  41  2.5. Figures  Figure 2.2: Aboveground plant biomass in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) at the start of the growing season and at the peak of live aboveground biomass in 2007 and 2008. Vertical bars indicate standard errors and significantly different means are denoted by ∗∗ (P < 0.05) and ∗ (P < 0.10). Two pooled samples of aboveground shoot material were made from 10 subsamples at each site.  42  2.5. Figures  Figure 2.3: Non-green (a), dead (b), and total standing crop (c) dry weights in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) in 2007 and 2008. Vertical bars indicate standard errors and significantly different means are denoted by ∗∗ (P < 0.05) and ∗ (P < 0.10).  43  2.5. Figures  Figure 2.4: Belowground standing crop in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) in 2007. Vertical bars indicate standard errors.  44  2.5. Figures  Figure 2.5: Nitrogen concentrations in three plant species from ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) in 2008. Vertical bars represent standard errors and significantly different means are denoted by ∗∗ (P < 0.05) and ∗ (P < 0.10). Pooled samples of aboveground shoot material were made from five subsamples of Carex aquatilis ssp. stans (n=3), Eriophorum angustifolium ssp. triste (n=4), and Salix arctica (n=2).  45  2.5. Figures  Figure 2.6: Carbon concentrations in three plant species from ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) in 2008. Vertical bars represent standard errors and significantly different means are denoted by ∗∗ (P < 0.05) and ∗ (P < 0.10). Pooled samples of aboveground shoot material were made from five subsamples of Carex aquatilis ssp. stans (n=3), Eriophorum angustifolium ssp. triste (n=4), and Salix arctica (n=2).  46  2.5. Figures  Figure 2.7: Nitrogen supply rates in ungrazed wet sedge meadows at Alexandra Fiord (AF) and grazed meadows at Sverdrup Pass (SP) during peak production in 2008. Values shown are the means of four samples. Vertical bars indicate standard errors. 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Canadian Journal of Botany, 64(11), 2502– 2507. Henry, GHR, Svoboda, J, & Freedman, B. 1990. Standing crop and net production of sedge meadows of an ungrazed polar desert oasis. Canadian Journal of Botany, 68(12), 2660–2667. Henry, HAL, & Jefferies, RL. 2002. Free amino acid, ammonium and nitrate concentrations in soil solutions of a grazed coastal marsh in relation to plant growth. Plant Cell and Environment, 25(5), 665–675. Hik, DS, & Jefferies, RL. 1990. Increases in the net aboveground primary production of a salt-marsh forage grass - a test of the predictions of the herbivore-optimization model. Journal of Ecology, 78(1), 180–195. Hik, DS, Sadul, HA, & Jefferies, RL. 1991. Effects of the timing of multiple grazings by geese on net aboveground primary production of swards of Puccinellia phryganodes. Journal Of Ecology, 79(3), 715–730. Hobbie, SE, Nadelhoffer, KJ, & Hogberg, P. 2002. A synthesis: The role of nutrients as constraints on carbon balances in boreal and arctic regions. Plant and Soil, 242(1), 163–170. Jassal, RS, Black, TA, Cai, T, Morgenstern, K, Li, Z, Gaumont-Guay, D, & Nesic, Z. 2007. Components of ecosystem respiration and an estimate of net primary productivity of an intermediate-aged Douglas-fir stand. Agricultural and Forest Meteorology, 144(1-2), 44–57. Jefferies, RL, Klein, DR, & Shaver, GR. 1994. Vertebrate herbivores and northern plantcommunities: reciprocal influences and responses. Oikos, 71(2), 193–206. Jia, B, Zhou, G, Wang, F, Wang, Y, & Weng, E. 2007. Effects of grazing on soil respiration on Leymus chinensis steppe. Climatic Change, 82, 211–223. Johnson, LC, Shaver, GR, Cades, DH, Rastetter, E, Nadelhoffer, K, Giblin, A, Laundre, J, & Stanley, A. 2000. Plant carbon-nutrient interactions control CO2 exchange in Alaskan wet sedge tundra ecosystems. Ecology, 81(2), 453–469. Kielland, K., McFarland, J. W., Ruess, R. W., & Olson, K. 2007. Rapid cycling of organic nitrogen in taiga forest ecosystems. Ecosystems, 10(3), 360–368. Klein, Julia A., Harte, John, & Zhao, Xin-Quan. 2007. Experimental warming, not grazing, decreases rangeland quality on the Tibetan Plateau. Ecological Applications, 17(2), 541– 557. 51  Chapter 2. References Larter, NC, & Nagy, JA. 2003. Wet sedge meadow habitat of southern Banks Island excluded from grazing by large herbivores for five years: effects on above ground standing crop. Manuscript Report 150. Department of Resources, Wildlife, and Economic Development, Government of the Northwest Territories, Inuvik, NT. Leriche, H, LeRoux, X, Gignoux, J, Tuzet, A, Fritz, H, Abbadie, L, & Loreau, M. 2001. Which functional processes control the short-term effect of grazing on net primary production in grasslands? Oecologia, 129(1), 114–124. Leriche, H, LeRoux, X, Desnoyers, F, Benest, D, Simioni, G, & Abbadie, L. 2003. Grass response to clipping in an African savanna: testing the grazing optimization hypothesis. Ecological Applications, 13, 1346–1354. Marazzi, A, & Yohai, VJ. 2006. Robust Box-Cox transformations based on minimum residual autocorrelation. Computational Statistics & Data Analysis, 50(10), 2752–2768. McNaughton, S. 1979. Grazing as an optimization process: grass ungulate relationships in the Serengeti. American Naturalist, 113(5), 691–703. McNaughton, SJ, Banyikwa, FF, & McNaughton, MM. 1998. Root biomass and productivity in a grazing ecosystem: The Serengeti. Ecology, 79(2), 587–592. Milchunas, DG, & Lauenroth, WK. 1993. Quantitative effects of grazing on vegetation and soils over a global range of environments. Ecological Monographs, 63(4), 327–366. Muc, M. 1977. Ecology and primary production of sedge-moss meadow communities, Truelove Lowland. In: Truelove Lowland, Devon Island, Canada. The University of Alberta Press. Muc, M, Freedman, B, & Svoboda, J. 1989. Vascular plant-communities of a polar oasis at Alexandra Fiord (79-Degrees-N), Ellesmere Island, Canada. Canadian Journal Of Botany, 67(4), 1126–1136. Muc, M, Svoboda, J, & Freedman, B. 1994a. Aboveground standing crop in plant communities of a polar desert oasis, Alexandra Fiord, Ellesmere Island. Pages 65–74 of: Svoboda, J, & Freedman, B (eds), Ecology of a polar oasis: Alexandra Fiord, Ellesmere Island, Canada. Toronto: Captus Press Inc. Muc, M, Svoboda, J, & Freedman, B. 1994b. Soils of an extensively vegetated polar desert oasis. Pages 41–50 of: Svoboda, J, & Freedman, B (eds), Ecology of a polar oasis: Alexandra Fiord, Ellesmere Island, Canada. Toronto: Captus Press Inc. Mulder, CPH. 1999. Vertebrate herbivores and plants in the arctic and subarctic: effects on individuals, populations, communities and ecosystems. Perspectives in Plant Ecology, Evolution and Systematics, 2, 29–55. 52  Chapter 2. References Murray, JL. 1991. Biomass distribution and nutrient pool and dynamics in the major muskox-grazed communities at Sverdrup Pass (79-Degrees-N) Ellesmere Island, N.W.T., Canada. M.Phil. thesis, University of Toronto. Myers-Smith, IH, McGuire, AD, Harden, JW, & Chapin, III, FS. 2007. Influence of disturbance on carbon exchange in a permafrost collapse and adjacent burned forest. Journal of Geophysical Research-Biogeosciences, 112(G4). Olofsson, J, Kitti, H, Rautiainen, P, Stark, S, & Oksanen, L. 2001. Effects of summer grazing by reindeer on composition of vegetation, productivity and nitrogen cycling. Ecography, 24(1), 13–24. Olofsson, J, Stark, S, & Oksanen, L. 2004. Reindeer influence on ecosystem processes in the tundra. Oikos, 105(2), 386–396. Pandey, CB, & Singh, JS. 1992. Rainfall and grazing effects on net primary productivity in a tropical savanna, India. Ecology, 73(6), 2007–2021. Polley, H. Wayne, Frank, Albert B., Sanabria, Joaquin, & Phillips, Rebecca L. 2008. Interannual variability in carbon dioxide fluxes and flux-climate relationships on grazed and ungrazed northern mixed-grass prairie. Global Change Biology, 14(7), 1620–1632. Raillard, M. 1992. Influence of muskox grazing on plant communities of Sverdrup Pass (79 N), Ellesmere Island, N.W.T. Canada. Ph.D. thesis, University of Toronto. Raillard, M, & Svoboda, J. 1999. Exact growth and increased nitrogen compensation by the arctic sedge Carex aquatilis var. stans after simulated grazing. Arctic, Antarctic, and Alpine Research, 31, 21–26. Raillard, M, & Svoboda, J. 2000. High grazing impact, selectivity, and local density of muskoxen in Cental Ellesmere Island, Canadian High Arctic. Arctic Antarctic and Alpine Research, 32(3), 278–285. Ritchie, ME, Tilman, D, & Knops, JMH. 1998. Herbivore effects on plant and nitrogen dynamics in oak savanna. Ecology, 79(1), 165–177. SAS Institute Inc. 2009 (January). SAS Online Doc 9.1.3 (http://supportr.sas.com/). Schimel, JP, & Chapin, FS. 1996. Tundra plant uptake of amino acid and NH4+ nitrogen in situ: Plants compete well for amino acid N. Ecology, 77(7), 2142–2147. Sheard, JW, & Geale, DW. 1983. Vegetation studies at Polar Bear Pass, Bathurst Island, NWT .1. Classification of plant communities. Canadian Journal of Botany-Revue Canadienne de Botanique, 61(6), 1618–1636. Sjogersten, S, van der Wal, R, & Woodin, SJ. 2008. Habitat type determines herbivory controls over CO2 fluxes in a warmer arctic. Ecology, 89(8), 2103–2116. 53  Chapter 2. References Smith, DL. 1996. Muskoxen / Sedge Meadow Interactions, North-central Banks Island, Northwest Territories, Canada. Ph.D. thesis, University of Saskatchewan. Stark, S, & Grellmann, D. 2002. Soil microbial responses to herbivory in an arctic tundra heath at two levels of nutrient availability. Ecology, 83(10), 2736–2744. Stewart, Kelley M., Bowyer, R. Terry, Ruess, Roger W., Dick, Brian L., & Kie, John G. 2006. Herbivore optimization by north American elk: Consequences for theory and management. Wildlife Monographs, DEC, 1–24. Susiluoto, Sanna, Rasilo, Terhi, Pumpanen, Jukka, & Berninger, Frank. 2008. Effects of grazing on the vegetation structure and carbon dioxide exchange of a fennoscandian fell ecosystem. Arctic Antarctic and Alpine Research, 40(2), 422–431. Tolvanen, A, & Henry, GHR. 2001. Responses of carbon and nitrogen concentrations in high arctic plants to experimental warming. Canadian Journal of Botany-Revue Canadienne de Botanique, 79(6), 711–718. Tuomi, J, Niemela, P A, Haukioja, E, Siren, S, & Neuvonen, S. 1984. Nutrient stress - an explanation for plant anti-herbivore responses to defoliation. Oecologia, 61(2), 208–210. van der Maarel, E, & Titlyanova, A. 1989. Above-ground and below-ground biomass relations in steppes under different grazing conditions. Oikos, 56, 364–370. van der Wal, R, Bardgett, RD, Harrison, KA, & Stien, A. 2004. Vertebrate herbivores and ecosystem control: cascading effects of faeces on tundra ecosystems. Ecography, 27(2), 242–252. Virtanen, R. 2000. Effects of grazing on above-ground biomass on a mountain snowbed, NW Finland. Oikos, 90(2), 295–300. Vitousek, P, & Howarth, RW. 1991. Nitrogen limitation on land and in the sea - how can it occur. Biogeochemistry, 13(2), 87–115. Walker, DA, Raynolds, MK, Daniels, FJA, Einarsson, E, Elvebakk, A, Gould, WA, Katenin, AE, Kholod, SS, Markon, CJ, Melnikov, ES, Moskalenko, NG, Talbot, SS, Yurtsev, BA, & CAVM Team. 2005. The Circumpolar Arctic vegetation map. Journal of Vegetation Science, 16(3), 267–282. Western Ag Innovations. 2008 (March). Western Ag Innovations Inc. About PRS-TM-probes (http://www.westernag.ca/innov). Wielgolaski, FE, Bliss, LC, Svoboda, J, & G, Doyle. 1981. Primary production of tundra. Chap. Tundra climates., pages 81–114 of: Bliss, LC, Heal, OW, & Moore, JJ (eds), Tunda ecosystems: a comparative analysis. Cambridge: Cambridge University Press. Willms, WD, Dormaar, JF, Adams, BW, & Douwes, HE. 2002. Response of the mixed prairie to protection from grazing. Journal of Range Management, 55(3), 210–216. 54  Chapter 3 Defoliation and litter removal effects on ungrazed wet sedge meadows in the High Arctic. 3.1  2  Introduction  The grazing optimization hypothesis predicts that yields of plant dry mass and nitrogen (N) in vegetative material will be higher in moderately grazed ecosystems compared to ungrazed or overgrazed natural systems (McNaughton 1979; Leriche et al. 2003; Sjogersten et al. 2008). However, compensation (similar net annual aboveground primary production in defoliated and non-defoliated plants at the individual, community, or ecosystem level) does not occur in response to all combinations of environmental variables or in all plant species and systems (Briske 1993). Overcompensation is possible in moderately grazed ecosystems because photosynthate that would normally be used to support the growth of roots and reproductive organs is used to replace lost leaf area (Turner et al. 1993). Moderate levels of grazing have increased aboveground net primary production in communities dominated by coral, algae, and terrestrial graminoids (Hik & Jefferies 1990). In graminoid-dominated plant commu2 A version of this chapter will be submitted for publication. Elliott, TE, & Henry, GHR. 2009. Defoliation and litter removal effects on ungrazed wet sedge meadows in the High Arctic.  55  3.1. Introduction nities, evidence of increased aboveground primary production in response to grazing has been documented in salt marshes (Cargill & Jefferies 1984; Hik & Jefferies 1990), savannas of the Serengeti (McNaughton 1979), and western North American grasslands (Frank & McNaughton 1993). In addition, growth of potted plants has increased in response to moderate frequencies of experimental defoliation (McNaughton et al. 1983; Georgiadis et al. 1989). Unlike other graminoid-dominated plant communities, little knowledge exists on the effects of grazing on High Arctic wet sedge meadows. Several factors determine the response of plants to grazing. The competitive effects of neighboring plants, the phenology of plant growth, and the history of grazing at a site can influence compensation abilities (Hik et al. 1991; Paige 1992). In addition, plant production is often N-limited (Belsky 1987; Bakker et al. 2004) and herbivores can either enhance or reduce nutrient cycling (McNaughton et al. 1983; Bakker et al. 2004). Nutrient cycling is altered because herbivory reduces litter accumulation and affects the quality of plant litter (Bakker et al. 2004). Younger leaves may have lower carbon (C) concentrations because they have less cellulose in their cell walls (Whaley et al. 1952). Furthermore, the concentration of defensive carbon-based compounds in leaves may decrease as soil nutrient availabilities increase (Dormann 2003). Defoliated plants can also compensate for nutrient loss by increasing their nutrient uptake (McNaughton 1979; Jaramillo & Detling 1988; Oesterheld & McNaughton 1988). The nutrients are then allocated preferentially to plant shoots (Green & Detling 2000). If defoliation causes N concentrations to increase in plant shoots, nutrient cycling accelerates because decomposition rates are faster with higher initial N concentrations (Taylor et al. 1989). Higher N contents in leaves also correlates with higher rates of photosynthesis (Welker et al. 2004). Therefore, higher N contents in leaves in response to defoliation may be a response to repair damaged tissues (Jaramillo & Detling 1988). Nutrient-limitation may deter plant growth following defoliation, but studies indicate 56  3.1. Introduction that fertilizing the soil stimulates compensatory plant growth (Leriche et al. 2003). The majority of unfertilized clipping experiments have resulted in undercompensation in production (Belsky 1986). Nutrient addition through animal excrement can increase soil NH4 + and NO3 − availability (Bazely & Jefferies 1985; Henry & Svoboda 1994). In certain grazed systems, aboveground net primary production (NPP) has increased due to the addition of animal excrement (Hik et al. 1991). Ecosystem respiration is measured as emissions of carbon dioxide (CO2 ) to the atmosphere from plants and soil microorganisms (Oberbauer et al. 2007). In Arctic wet sedge tundra, plant respiration contributes to the majority of whole ecosystem respiration (Johnson et al. 2000), whereas levels of microbial respiration can be minor (Stark & Grellmann 2002). In addition, defoliation can affect plant respiration. For example, excised roots of Eriophorum vaginatum plants that were defoliated had higher respiration rates compared to those extracted from control plots in Alaska (Chapin 1980). Other abiotic variables are affected by experimental defoliation. In tallgrass prairie, defoliation increased mean soil temperature (Wan et al. 2002; Zhou & Luo 2006). However, soil moisture either decreased (Wan et al. 2002) or did not change due to experimental defoliation (Zhou & Luo 2006). Plant litter affects ecosystem function by changing abiotic conditions. Restricting the access of grazers to forage with exclosures results in the accumulation of litter at the soil surface. This build-up of litter alters light reception, soil temperatures, soil water contents, and nutrient dynamics (Facelli & Pickett 1991). Litter can change interactions between plant species, ameliorating stressful growing conditions in harsh environments (Foster & Gross 1998). Soil moisture is conserved and evaporation is reduced in some grassland ecosystems because of the build-up of litter (Willms et al. 2002). Responses in net production to litter removal also depend on the abiotic conditions of the ecosystem (Willms et al. 1986). For instance, removing standing litter and mulch decreased production in mixed prairies, but 57  3.1. Introduction increased aboveground NPP in fescue prairies in Alberta. Mixed prairies are more xeric than fescue prairies and removing litter induced a soil moisture deficit (Willms et al. 1986). Plants in the High Arctic grow in resource-limited conditions. Their growth and distribution is restricted by short growing seasons, low temperatures, reduced solar radiation, low precipitation levels, and limited nutrient availabilities (Barry et al. 1981; Edlund & Alt 1989; Freedman et al. 1994; Tolvanen & Henry 2001; Brooker & van der Wal 2003). High Arctic lowlands with favourable soil moisture levels and microclimates support wet sedge meadows, which are the most productive plant communities in this region (Henry et al. 1990). The relatively high primary productivity of these plant communities is permitted by the physiological and morphological adaptations of the dominant graminoid species to cool soil temperatures (Henry et al. 1990). Because of this high primary productivity, High Arctic wet sedge meadows are important sources of forage for several herbivores such as muskoxen (Ovibos moschatus) and Peary caribou (Rangifer tarandus pearyi ) (Henry et al. 1990; Henry 1998). In the High Arctic, low decomposition rates exist because of cold and dry environments that result in the immobilization of nutrients in plant litter (Brathen & Odasz-Albrigtsen 2000). Decomposition rates in wet sedge meadows are further reduced because of low oxygen concentrations and diffusion rates (Jonasson & Shaver 1999). In addition to low decomposition rates, litter accumulations in ungrazed High Arctic sedge meadows are higher than those in grazed meadows because of the absence of large herbivores (Henry et al. 1990; Henry 1998). Several ecological studies in the last 30 years have focused on the response of plants to grazing; however, research has primarily been conducted on temperate and sub-tropical grass-dominated systems. In High Arctic wet sedge meadows, Henry & Svoboda (1994), Henry (1998), and Raillard & Svoboda (1999) have studied the effects of muskox grazing. Henry & Svoboda (1994) found that grazed sites had higher aboveground NPP than un58  3.2. Methods grazed sites. In addition, Raillard & Svoboda (1999) investigated the effects of experimental defoliation on Carex aquatilis ssp. stans, a common graminoid in sedge meadow plant communities. The effects of defoliation and litter removal on ecosystem function in High Arctic sedge meadow plant communities had not been previously studied and this was the focus of our study. Our study tested the following predictions for High Arctic sedge meadows: • Moderate clipping frequencies increase aboveground NPP, • Moderate clipping frequencies increase ecosystem respiration, • Soil N (nitrate and ammonium) remains constant regardless of clipping frequency, • Moderate clipping frequencies increase shoot N in C. stans, E. angustifolium ssp. triste, C. membranacea, and Salix arctica, • Clipping decreases shoot C in C. stans, E. triste, C. membranacea, and Salix arctica, • Clipping increases soil moisture and temperatures, and • Litter removal increases aboveground NPP and soil temperatures, but decreases soil moisture. To test these predictions, we established an artificial defoliation experiment with combinations of clipping and litter removal treatments in ungrazed High Arctic sedge meadows at a site on eastern Ellesmere Island.  3.2 3.2.1  Methods Site description  The study was conducted at a coastal lowland adjacent to Alexandra Fiord (AF) (78◦ 53’N, 75◦ 55’W) on the central east coast of Ellesmere Island. Alexandra Fiord is a polar oasis 59  3.2. Methods that has a warmer climate, higher species diversity, greater plant cover, and higher primary productivity than the surrounding area (Freedman et al. 1994; Raillard & Svoboda 1999). Wet sedge meadow plant communities occur on gleysolic cryosol soils on flat sites with impeded drainage (Muc et al. 1994b). The plant communities are dominated by E. triste, C. stans, C. membranacea, Dryas integrifolia, Vaccinium uliginosum, and Salix arctica (Henry et al. 1990; Muc et al. 1994a; Henry 1998). More detailed descriptions of AF are outlined in Svoboda & Freedman (1994). The lowland is approximately 8 km2 in area (Henry et al. 1986; Svoboda & Freedman 1994); however, the nearest population of muskoxen is approximately 15-20 km to the north on the Bache Peninsula (Henry et al. 1986). Mammalian grazers are rare along the north coast of the Johan Peninsula and populations have likely never been large in the AF lowland region (Henry, pers. obs.). The geographical isolation and relative small size of AF makes it unsuitable to sustain populations of large mammalian grazers such as muskoxen (Henry et al. 1986; Muc et al. 1989). The accumulation of litter is considerably higher in the ungrazed meadows of AF compared to grazed polar oases because of this lack of large herbivores (Muc et al. 1994a; Henry 1998).  3.2.2  Clipping experiments  Two clipping experiments, Grid 1 and Grid 2, were established at AF in visually homogeneous wet sedge meadows in 2007. These two meadows were representative of the dominant wet sedge meadows in the AF lowland (Muc et al. 1989; Henry et al. 1990). In these experiments, treatments were applied to 50 cm x 50 cm plots that were separated by 1.5 m. Plots were randomly assigned to one of six treatments that were maintained throughout the 2007 and 2008 growing seasons. This was a two factor experiment with combinations of litter removal and clipping levels as the treatments. There were six repetitions per treatment for a total of 30 plots in Grid 1 and 36 plots in Grid 2. Litter removal treatments 60  3.2. Methods were conducted by removing the unattached dead plant material within each plot. Clipping treatments were applied to each plot by cutting and collecting all plant material over 1.5 cm in height. This height was chosen because muskoxen leave the lower sections of grazed sedge shoots intact since they do not graze to ground level (Raillard 1992; Raillard & Svoboda 1999). The treatments included: i. control: no clipping, no litter removal; ii. no clipping with litter removal (LR); iii. one clipping / season, no litter removal (1x); iv. two clippings / season, no litter removal (2x); v. three clippings / season, no litter removal (3x); and vi. one clipping / season, litter removal (1x + LR). All of these treatments were applied to Grid 2; however, the 1x + LR treatment was not conducted at Grid 1 because there was much less plant litter at that site. Litter was removed from selected treatments in mid-June in 2007 and 2008. Table 3.1 shows the clipping schedule at AF during the two study years. Clipping dates varied between the two years because it was necessary to complete all 2008 clipping treatments before the harvest of aboveground plant material. Harvests were conducted at the peak of live aboveground biomass (peak season), which typically occurs the last week of July to the first week of August in High Arctic plant communities (Henry et al. 1990). Aboveground NPP During the 2007 and 2008 growing seasons, plant material removed from each of the clipping treatments was placed into a paper bag, sorted to species, and dried. In addition, all vascular 61  3.2. Methods vegetation was harvested from 20 cm x 50 cm quadrats during peak season by placing these quadrats in the centre of each 50 cm x 50 cm plot. All aboveground plant material within each 20 cm x 50 cm quadrat was then cut at the ground surface and placed into a paper bag, sorted to species, dried, and then sent to the lab at UBC. In the lab, the samples were dried again for 24 to 30 hours at 60◦ C before weighing. Sample weights were combined to form the following categories: • green - all plant material from the current year’s growth including reproductive tissues; • non-green - stem growth in shrubs; • dead - unattached litter and dead plant material attached to new growth; and • standing crop - sum of green, non-green, and dead plant material. All values were converted to g m−2 prior to analysis and the weights of the 2008 clipped plant material were factored into the calculations for each plot. Aboveground NPP was calculated as the difference between peak season green biomass and the average of overwinter green material harvested from ten quadrats (20 x 50 cm) just after snow melt from both sites. The net production of dwarf shrubs was estimated using the following adjustments adapted from Henry et al. (1990): • for Salix arctica and Vaccinium uliginosum: 100% of peak season green biomass + 25% of this value for stem diametre growth; • for Cassiope tetragona: 25% of peak season green biomass because green leaves function for four years; and • for Dryas integrifolia and Saxifraga oppositifolia: 50% of green peak season biomass because leaves function for two years. 62  3.2. Methods Ecosystem respiration A closed flow infrared gas analyzer, (IRGA; LI-COR 840, Lincoln, NE, USA) was used to measure ecosystem respiration on August 1, 2008. In each grid, one 10.16 cm diameter x 7.62 cm high polyvinyl chloride (PVC) collar was placed in each control and 3x plot (n=6 per treatment). The collars were left untouched for approximately 24 hours so that soil microbial respiration would return to normal levels after disturbance. The IRGA chamber was then placed over each PVC collar and air was mixed by a small fan within the chamber during measurement. Pressure differences between the inside and outside of the chamber were minimized with a 3 mm vent inside the chamber. After steady state was achieved within the chamber, two consecutive measurements of CO2 were taken in each plot over a 120 second period. The chamber was aerated between each of the two measurements. These two values were then averaged to calculate ecosystem respiration (µmol CO2 m−2 s−1 ). Winds were light with broken cloud cover and the temperature remained constant during measurements. Soil N03 − and NH4 + availability Plant Root SimulatorsT M (PRS) were used to measure soil NO3 − and NH4 + availability in each plot. These probes are ion exchange membranes that show ion flux in the soil by simulating nutrient absorption by plant roots (Western Ag Innovations 2008). Logistical problems prevented the arrival of the PRS membranes at the research site for the 2007 field season; however, four cation and four anion probes were inserted vertically into each plot on June 20, 2008 and removed on August 16, 2008. The PRST M probes were washed with deionized water and scrubbed with a coarse brush to remove residue immediately after removal from each plot. The probes were combined for each plot and they were sent to Western Ag Innovations in Saskatoon for ion extraction and analysis 63  3.2. Methods Shoot nitrogen and carbon concentrations To measure shoot N and C concentrations, subsamples were taken from plant material that was used to measure aboveground NPP. Subsamples from the control, 1x, 2x, and 3x clipping frequencies were obtained from samples of C. membranacea (Grid 1) and C. stans (Grid 2). In addition, subsamples of E. triste (Grid 1) and S. arctica (Grid 2) were taken from the control and 3x treatment. Different species were chosen for each grid because there was insufficient aboveground biomass for a single species from both Grid 1 and 2. A Wiley Mill with mesh size #4 was used to grind larger shoot material from the subsamples into a powder, while a mortar and pestle with liquid N was utilized for smaller quantities. Nitrogen and carbon concentrations were then measured by dry combustion with a Carlo-Erba NA-1500 NCS analyzer at the Analytical Chemistry Laboratory of the BC Ministry of Forests and Range in Victoria.  Soil temperatures In both study years, thermistors were placed to a depth of 10 cm in eight plots in each of the clipping grids on June 20 and removed on August 1. The thermistors were programmed to read belowground temperatures every ten minutes; this information was stored in data loggers and downloaded at the end of the field season. The number of thermistors set-up was limited by their availability at the research site. The temperature data were gathered from three control, two LR, and two 3x treatments. In 2007, all temperature data was lost in Grid 2 because of equipment failure. In 2008, two thermistor cables in Grid 1 were damaged by wildlife preventing the collection of data from two plots.  Soil moisture Soil moisture was measured with a time domain reflectometry (TDR) probe (Hydrosense, Campbell Scientific Inc., Edmonton) throughout the 2008 growing season in both grids on 64  3.2. Methods July 24, August 2, and August 16. Average soil water contents for the summer were then calculated from these measurements. The TDR probe was vertically inserted to a depth of 12 cm for each measurement.  3.2.3  Data analysis  All statistical analyses were completed using SAS (version 9.1) (SAS Institute Inc. 2009). The assumptions of normality and constant variance of the residuals were met in all statistical tests except for one. The data for aboveground NPP were Ln transformed. Analysis of covariance (ANCOVA) was used to test the effects of increased clipping frequencies on aboveground NPP and soil moisture. The grids were the class variable and clipping frequencies were the covariate in these models. Linear regression analysis was used to measure the response of C and N in the shoots of four plant species to increased clipping frequencies. Analysis of variance (ANOVA) was utilized to test if litter removal affected aboveground NPP and soil moisture levels compared to the control. In addition, the effects of the 3x clipping treatment on ecosystem respiration and the availability of soil N03 − and NH4 + were analyzed with ANOVA models. In cases when there were data for both grids, variables were first tested to see if there was an effect of grid x treatment. If no interaction was found, the data for the two grids was pooled together for analysis. The 1x + LR treatment was only applied to Grid 2. For this reason, it was possible to use 2-way ANOVA models to test if litter removal and clipping treatments influenced aboveground NPP and soil moisture in that Grid.  65  3.3. Results  3.3 3.3.1  Results Aboveground NPP  Clipping decreased aboveground NPP, but litter removal treatments had no effect on it (Figure 3.1). Aboveground NPP decreased marginally with increased clipping frequency (slope = -0.08, P = 0.07, ANCOVA). In contrast, litter removal did not affect aboveground NPP (P = 0.16, ANOVA). There was no site x treatment interaction so both clipping experiments were analyzed together. In Grid 2, clipping and litter removal decreased aboveground NPP (Figure 3.2). Aboveground NPP decreased significantly as clipping frequencies increased (P < 0.01, 2-way ANOVA), but declined marginally with litter removal treatments (P = 0.07, 2-way ANOVA). There was no clipping x litter removal interaction (P = 0.27, 2-way ANOVA).  3.3.2  Ecosystem respiration  Results from both sites were combined together because there was no site x treatment interaction. Figure 3.3 indicates that ecosystem respiration (Mean ± S.E.) was significantly greater in the control plots (1.05 ± 0.06 µmol CO2 m−2 s−1 ) than the 3x clipping treatment (0.70 ± 0.03 µmol CO2 m−2 s−1 ) (P < 0.01).  3.3.3  Soil NO3 − and NH4 + availability  The control and 3x plots had similar NO3 − (P = 0.98) and NH4 + (P = 0.64) availabilities in the soil (Figure 3.4). However, the availability of soil NH4 + in the control was 59% greater than in the 3x treatment and 57% higher than soil NO3 − availability. There was no site x treatment interaction so data from both sites were combined together for analysis.  66  3.3. Results  3.3.4  Shoot nitrogen and carbon concentrations  Concentrations of N in the shoots of in four plant species from AF responded differently to clipping. Figure 3.5 shows that increased clipping frequencies resulted in higher shoot N concentrations in C. membranacea (slope = 0.12, P < 0.01) and E. triste (slope = 0.08, P < 0.01). However, higher clipping frequencies decreased S. arctica shoot N concentrations (slope = -0.09, P = 0.04). Concentrations of N in C. stans shoots did not change as clipping frequency increased (P = 0.12). Decreases in the C concentrations in three of the four plant species at AF occurred as a response to clipping (Figure 3.6). Carbon concentrations declined in the shoots of C. membranacea (slope = -0.25, P < 0.01), C. stans (slope = -0.43, P < 0.01), and S. arctica (slope = -1.11, P < 0.01) that were clipped more frequently. However, concentrations of C in shoots of E. triste remained similar regardless of clipping frequency (P = 0.36).  3.3.5  Soil temperatures  Soil temperature data for Grid 1 and Grid 2 could not be pooled together because of missing data. Grid 1 data indicated that clipping and litter removal treatments did not affect average soil temperatures for the 2007 growing season (P = 0.12, ANOVA). Similarly, soil temperature were not affected by clipping and litter removal treatments in Grid 2 for the 2008 growing season (P = 0.99, ANOVA).  3.3.6  Soil moisture  Soil moisture responded to increased clipping frequencies and litter removal treatments (Figure 3.7). With data from both sites combined, soil moisture increased with clipping frequency (slope =0.85, P < 0.01, ANCOVA). In addition, removing litter marginally increased soil moisture contents (P = 0.09, ANOVA). 67  3.4. Discussion Litter removal and clipping treatments did not influence soil moisture in Grid 2 (Figure 3.8). There was no interaction of clipping x litter removal (P = 0.90, 2-way ANOVA). Soil moisture levels remained similar in Grid 2 despite applying clipping (P = 0.22, 2-way ANOVA) and litter removal (P = 0.33, 2-way ANOVA) treatments.  3.4  Discussion  According to the grazing optimization hypothesis, net primary production and nitrogen levels in plants increase with moderate levels of grazing. However, compensatory growth does not occur in all plant species and ecosystems (Briske 1993). If the grazing optimization hypothesis was supported in High Arctic sedge meadow plant communities, defoliation should have increased NPP and shoot N concentrations in aboveground plant material. However, not all of the results from our study supported the grazing optimization hypothesis. Aboveground NPP declined with increased clipping frequencies, and this does not support the hypothesis. In contrast, the hypothesis was supported by higher shoot N concentrations in the aboveground plant material in two of four species subjected to increased clipping frequencies. Studies in the High Arctic do not support the grazing optimization hypothesis. For example, in our study aboveground NPP decreased marginally with increased clipping frequencies. Additionally, Smith (1996) found that aboveground NPP in E. triste decreased with more intense clipping treatments on Banks Island. However, Raillard & Svoboda (1999) found thatC. stans compensated in growth in response to clipping at Sverdrup Pass, Ellesmere Island. It should be noted that Raillard & Svoboda (1999) conducted their study by erecting exclosures in wet sedge meadows that had been heavily grazed; therefore, the grazing history of the sites must be considered when making comparisons between these studies. 68  3.4. Discussion Plant phenological stage usually influences responses to clipping (Hik et al. 1991). In our study, clipping for the single clipping treatment (1x) was conducted during the middle of the growing season on July 10, 2007 and July 6, 2008; aboveground NPP was lower in this treatment compared to control plots. Results from other Arctic defoliation studies that considered plant phenology varied depending on the time of aboveground vegetation removal (Archer & Tieszen 1983; Hik et al. 1991; Raillard & Svoboda 1999). In salt marsh plant communities subjected to grazing by lesser snow geese, Hik et al. (1991) found that aboveground NPP was lower in Puccinellia phryganodes swards grazed later in the growing season compared to those grazed earlier. It was suggested that P. phryganodes needed a period free of grazing at the end of the summer to acquire nutrients and carbohydrate reserves for the winter and upcoming growing season (Hik et al. 1991). In contrast, Archer & Tieszen (1983) found that defoliating E. vaginatum plants early in the growing season resulted in less aboveground shoot growth when compared to plants that were either defoliated later in the season or not at all. However, the long-term results of their experiment indicated that increasing the number of defoliation events per year decreased production in E. vaginatum more than applying early clipping treatments (Archer & Tieszen 1983). This suggests that there were less photosynthetic inputs to individual plants that were clipped more frequently (Archer & Tieszen 1983). In High Arctic wet sedge meadows, Raillard & Svoboda (1999) found that aboveground NPP in C. stans was similar in all treatments regardless of the time and frequency of defoliation events. The results from our study suggest that the two year duration of the clipping treatments may have affected aboveground NPP in High Arctic wet sedge meadows more than the timing of the defoliation events. This is similar to the results from Archer & Tieszen (1983). The photosynthetic input to plants in plots that were subjected to higher clipping frequencies may have been reduced, decreasing their ability to produce new aboveground vegetation. It must be noted that our experiment was set-up to test the effects of defoliation 69  3.4. Discussion at the plot level whereas Archer & Tieszen (1983) focused primarily on responses at the plant level. Similar to aboveground NPP, ecosystem respiration was greater in unclipped plots compared to plots that were clipped at a high frequency (3x - three times over two growing seasons). Due to the lack of equipment at the study site, we were not able to examine if the intermediate clipping frequencies (1x and 2x clipping treatments) led to higher rates of ecosystem respiration. It has been found that plant respiration accounts for the majority of ecosystem respiration in tundra systems (Johnson et al. 2000). Sites protected from reindeer grazing in Finnish alpine tundra had similar ecosystem respiration levels compared to grazed sites (Susiluoto et al. 2008). Grazing did not effect dwarf shrub cover in the study, which was correlated with ecosystem respiration (Susiluoto et al. 2008). Hence, lower aboveground NPP in the 3x clipping treatment at AF may have contributed to decreased plant respiration rates compared to the control plots. Soil NH4 + and NO3 − availabilities were similar in both defoliated and undefoliated plots, supporting the predictions that soil N remains constant regardless of clipping frequency. Marion et al. (1991) found similar results in the Chaparral of the United States. Nutrient addition in the form of fertilizer or animal excrement may be necessary to increase soil NH4 + and NO3 − availability (Bazely & Jefferies 1985; Henry & Svoboda 1994). Without the addition of N in the form of animal excrement or fertilizer, aboveground NPP possibly decreased as a response to clipping in wet sedge meadows at AF. Compensatory regrowth occurred in C. stans plants in response to defoliation in High Arctic wet sedge meadows protected from grazing by exclosures at SP (Raillard & Svoboda 1999). In the sedge meadows at SP, soil NH4 + availability was higher than in the ungrazed sites at AF. Higher soil NH4 + availability may have lead to increases in aboveground NPP and belowground biomass at SP. The importance of increased available N was also noted in another Arctic clipping experiment where fertilized P. phryganodes plants that were clipped 70  3.4. Discussion had higher aboveground NPP compared to those plants that were clipped, but not fertilized (Hik et al. 1991). These results suggest that higher available soil N in the grazed and fertilized sites aided the growth of vegetation because plants did not have to rely only on the slow mineralization of N in plant litter (Ruess & McNaughton 1987; Henry 1998; Knapp et al. 1999). Our study did provide some support for the grazing optimization hypothesis. Higher clipping frequencies resulted in increased shoot N in C. membranacea and E. triste compared to the control plots. In Arctic Alaska, N concentrations in leaves after defoliation of E. vaginatum and C. aquatilis were higher than in leaves that were not defoliated and concentrations rose as clipping frequencies increased (Chapin 1980). In addition, clipping increased shoot N in plants in wet meadow sedges in the Arctic (Ouellet et al. 1994; Raillard & Svoboda 1999), mixed prairies (Green & Detling 2000), and tallgrass prairies (Turner et al. 1993). Defoliation also increased shoot N concentrations in Kyllinga nervosa in the Serengeti plains (Ruess et al. 1983), but decreased those in the mangrove Kandelia candel in Hong Kong (Tong et al. 2003). Graminoids use their belowground nutrient reserves to support the growth of nutritious shoots in response to grazing (Bryant et al. 1983). It is possible that roots enhanced N uptake and allocated it preferentially to shoots following defoliation (Green & Detling 2000). High concentrations of N in leaves may equate to increased photosynthetic rates leading to compensatory growth (Welker et al. 2004). In contrast to higher N concentrations in graminoid shoots, concentrations of N in S. arctica shoots decreased with clipping. Chapin (1980) found that deciduous shrubs allocate maximum nutrients to their shoot early in the summer. In our experiment it was possible that N-rich plant material in S. arctica was removed in the first clipping event in the 3x treatment and levels of this nutrient did not recuperate by harvest time. In contrast, S. arctica plants in the control had higher shoot N-levels throughout the growing season possibly because no shoot material was removed early in the season. 71  3.4. Discussion Carbon concentrations in C. membranacea, C. stans, and S. arctica declined as clipping frequencies increased, supporting the predictions of our study. Ouellet et al. (1994) measured C as total non-structural carbohydrates, a surrogate for C concentrations, in sedges and S. lanata in meadows on Southampton Island. They found that total non-structural carbohydrates decreased in plants that were clipped more frequently. In our study at AF, plants that were clipped more frequently had younger leaves. The newer cells in these leaves may have had lower proportions of cellulose in their walls because they were younger (Whaley et al. 1952). We predicted that soil temperatures would increase in plots that were clipped. However, soil temperatures did not increase as we predicted. The removal of aboveground plant material by defoliation opens the plant canopy, increasing the amount of incoming solar radiation that reaches and subsequently heats the soil surface (Frank & McNaughton 1993; Semmartin & Oesterheld 1996; Bremer et al. 1998). In our study, soil temperatures were similar in control and 3x treatments in both Grids in 2007 and 2008. Soil temperatures also remained similar in control and litter removal plots in both Grids; they did not increase when litter was removed as we predicted. The statistical power required to detect differences in this variable was limited by equipment limitations and failures that reduced sample sizes. Soil moisture levels rose as clipping frequencies increased, which was predicted. Soil water may be conserved when plants are defoliated because areas of leaf transpiring surfaces are reduced (Archer & Detling 1986; Wraith et al. 1987). It was predicted that litter removal would increase aboveground NPP in our study. Large standing crops of litter relative to living plant biomass may block photosynthetically active radiation as it passes through the litter dominated canopy, reducing the amount of light available for primary production (Bazely & Jefferies 1986; Knapp & Seastedt 1986). However, the combined data from both Grids indicated that aboveground NPP did not change when litter was removed from plots. 72  3.4. Discussion Contrasting to our predictions, litter removal caused soil moisture contents to rise marginally. This indicates that a soil moisture deficit was not induced as it was in xeric mixed prairies in Alberta (Willms et al. 1986). Unlike xeric mixed prairies, sedge meadows at AF are more hydric and evapotranspiration rates are lower because of cooler conditions. Litter removal may have increased the amount of incoming solar radiation that reached the soil surface; however, our data did not indicate any differences in soil temperatures. Since energy balance at the soil surface considers sensible heat fluxes to the atmosphere and soil along with latent heat flux (Oke 1978), the increased energy available at the soil surface may have been used in evaporation, drawing additional soil moisture into the plot. As with any field experiment, there are shortcomings and strengths associated with this study. Experimental clipping defoliates plants, but it does not completely simulate grazing. For example, grazers trample litter and this process increases the contact of dead plant material with soil microbes, enhancing decomposition rates (Willms et al. 2002). A second shortcoming is that in clipping experiments, nutrient-cycling is hindered because nutrients often are not returned to the soil in the form of animal excrement. As in most ecological field experiments, increasing the number of replicates of each treatment would have increased the statistical power of the experiment. Overall, this study was effective because it improved our understanding of the connection between N-cycling and NPP in High Arctic wet sedge meadow plant communities. The grazing optimization hypothesis was supported because N concentrations in plant shoots increased with clipping; however, decreases in aboveground NPP with higher clipping frequencies did not agree with the hypothesis.  73  3.5. Tables and Figures  3.5  Tables and Figures  Table 3.1: Schedule for experimental clipping treatments and harvest dates at Alexandra Fiord in 2007 and 2008 Year Grid Treatment Clipping Date(s) Harvest Date 2007 1 1x July 10 August 2, 2008 2007 2 1x July 11 August 1, 2008 2008 1 1x July 6 August 2, 2008 2008 2 1x July 6 August 1, 2008 2007 2 1x + LR July 11 August 1, 2008 2008 2 1x + LR July 6 August 1, 2008 2007 1 2x June 26 & July 25 August 2, 2008 2007 2 2x June 25 & July 25 August 1, 2008 2008 1 2x June 27 & July 16 August 2, 2008 2008 2 2x June 27 & July 16 August 1, 2008 2007 1 3x June 19, July 10, & August 6 August 2, 2008 2007 2 3x June 19, July 11, & August 6 August 1, 2008 2008 1 3x June 20, July 6, & July 26 August 2, 2008 2008 2 3x June 20, July 6, & July 26 August 1, 2008  74  3.5. Tables and Figures  Figure 3.1: Aboveground NPP from two wet sedge meadow sites at Alexandra Fiord in 2008. In (a), the clipping sites have been combined and the values are n=12 per clipping frequency. The litter removal treatment (- Litter) and control are shown in (b). Treatments consisted of six replications and values shown in (b) are the means of data pooled for the two grids. Vertical bars indicate standard errors.  75  3.5. Tables and Figures  Figure 3.2: Aboveground NPP in Grid 2 at Alexandra Fiord in 2008. Treatments consisted of six replications. Vertical bars indicate standard error and means with the same letter are not significantly different than the control (P > 0.05). There was no LR and 1x interaction.  76  3.5. Tables and Figures  Figure 3.3: Effect of clipping on ecosystem respiration in two wet sedge meadows at Alexandra Fiord during peak aboveground biomass in 2008. Treatments consisted of six replications and values shown are the means for the data pooled from the two grids. Statistically significant means are denoted by ∗∗ (P < 0.05) and ∗ (P < 0.10). Vertical bars indicate standard error.  77  3.5. Tables and Figures  Figure 3.4: Nitrogen supply rates for the control and 3x treatment in two wet sedge meadows at Alexandra Fiord in 2008. There were six replications per treatment and the values for both grids were combined. Vertical bars indicate standard errors.  78  3.5. Tables and Figures  Figure 3.5: Nitrogen concentrations in shoots of four plant species at Alexandra Fiord in 2008. Carex membranacea (24 samples) and Eriophorum angustifolium ssp. triste (12 samples) were analyzed from Grid 1. Carex aquatilis ssp. stans (24 samples) and Salix arctica (12 samples) were analyzed from Grid 2. Adjusted R2 values are indicated.  79  3.5. Tables and Figures  Figure 3.6: Carbon concentrations in shoots of four plant species at Alexandra Fiord in 2008. Carex membranacea (24 samples) and Eriophorum angustifolium ssp. triste (12 samples) were analyzed from Grid 1. Carex aquatilis ssp. stans (24 samples) and Salix arctica (12 samples) were analyzed from Grid 2. Adjusted R2 values are indicated.  80  3.5. Tables and Figures  Figure 3.7: Average soil volumetric water content from two clipping experiments at Alexandra Fiord in 2008. The relation between soil volumetric water content and clipping frequency in 48 samples is shown in (a). The effect of the litter removal treatment (- Litter) and control are shown in (b). Treatments consisted of six replications and values shown in (b) are the means of data pooled for the two grids. Vertical bars indicate standard errors and statistically significant means are denoted by ∗∗ (P < 0.05) and ∗ (P < 0.10).  81  3.5. Tables and Figures  Figure 3.8: Average soil volumetric water content in four treatments in Grid 2 at Alexandra Fiord in 2008. Treatments consisted of six replications. 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References McNaughton, SJ, Wallace, LL, & Coughenour, MB. 1983. Plant adaptation in an ecosystem context - effects of defoliation, nitrogen, and water on growth of an African C-4 sedge. Ecology, 64(2), 307–318. Muc, M, Freedman, B, & Svoboda, J. 1989. Vascular plant-communities of a polar oasis at Alexandra Fiord (79-Degrees-N), Ellesmere Island, Canada. Canadian Journal Of Botany, 67(4), 1126–1136. Muc, M, Svoboda, J, & Freedman, B. 1994a. Aboveground standing crop in plant communities of a polar desert oasis, Alexandra Fiord, Ellesmere Island. Pages 65–74 of: Svoboda, J, & Freedman, B (eds), Ecology of a polar oasis: Alexandra Fiord, Ellesmere Island, Canada. Toronto: Captus Press Inc. Muc, M, Svoboda, J, & Freedman, B. 1994b. Soils of an extensively vegetated polar desert oasis. Pages 41–50 of: Svoboda, J, & Freedman, B (eds), Ecology of a polar oasis: Alexandra Fiord, Ellesmere Island, Canada. Toronto: Captus Press Inc. Oberbauer, SF, Tweedie, CE, Welker, JM, Fahnestock, JT., Henry, GHR, Webber, PJ, Hollister, RD, Walker, MD, Kuchy, A, Elmore, E, & Starr, G. 2007. Tundra CO2 fluxes in response to experimental warming across latitudinal and moisture gradients. Ecological Monographs, 77(2), 221–238. Oesterheld, M, & McNaughton, SJ. 1988. Intraspecific variation in the response of themedatriandra to defoliation - the effect of time of recovery and growth-rates on compensatory growth. Oecologia, 77(2), 181–186. Oke, TR. 1978. Boundary Layer Climates. London: Mehuen & Co Ltd. Ouellet, JP, Boutin, S, & Heard, DC. 1994. Responses to simulated grazing and browsing of vegetation available to caribou in the arctic. Canadian Journal of Zoology-Revue Canadienne de Zoologie, 72(8), 1426–1435. Paige, KN. 1992. Overcompensation in response to mammalian herbivory - from mutualistic to antagonistic interactions. Ecology, 73(6), 2076–2085. Raillard, M. 1992. Influence of muskox grazing on plant communities of Sverdrup Pass (79 N), Ellesmere Island, N.W.T. Canada. Ph.D. thesis, University of Toronto. Raillard, M, & Svoboda, J. 1999. Exact growth and increased nitrogen compensation by the arctic sedge Carex aquatilis var. stans after simulated grazing. Arctic, Antarctic, and Alpine Research, 31, 21–26. Ruess, RW, & McNaughton, SJ. 1987. Grazing and the dynamics of nutrient and energy regulated microbial processes in the Serengeti grasslands. Oikos, 49(1), 101–110. 86  Chapter 3. References Ruess, RW, McNaughton, SJ, & Coughenour, MB. 1983. The effects of clipping, nitrogensource and nitrogen concentration on the growth-responses and nitrogen uptake of an east-african sedge. Oecologia, 59(2-3), 253–261. SAS Institute Inc. 2009 (January). SAS Online Doc 9.1.3 (http://supportr.sas.com/). Semmartin, M, & Oesterheld, M. 1996. Effect of grazing pattern on primary productivity. Oikos, 75(3), 431–436. Sjogersten, S, van der Wal, R, & Woodin, SJ. 2008. Habitat type determines herbivory controls over CO2 fluxes in a warmer arctic. Ecology, 89(8), 2103–2116. Smith, DL. 1996. Muskoxen / Sedge Meadow Interactions, North-central Banks Island, Northwest Territories, Canada. Ph.D. thesis, University of Saskatchewan. Stark, S, & Grellmann, D. 2002. Soil microbial responses to herbivory in an arctic tundra heath at two levels of nutrient availability. Ecology, 83(10), 2736–2744. Susiluoto, Sanna, Rasilo, Terhi, Pumpanen, Jukka, & Berninger, Frank. 2008. Effects of grazing on the vegetation structure and carbon dioxide exchange of a fennoscandian fell ecosystem. Arctic Antarctic and Alpine Research, 40(2), 422–431. Svoboda, Josef, & Freedman, Bill (eds). 1994. Ecology of a polar oasis: Alexandra Fiord, Ellesmere Island, Canada. Captus Press Inc., Toronto. Taylor, BR, Parkinson, D, & Parsons, WFJ. 1989. Nitrogen and lignin content as predictors of litter decay-rates - a microcosm test. Ecology, 70(1), 97–104. Tolvanen, A, & Henry, GHR. 2001. Responses of carbon and nitrogen concentrations in high arctic plants to experimental warming. Canadian Journal of Botany-Revue Canadienne de Botanique, 79(6), 711–718. Tong, YF, Lee, SY, & Moqon, B. 2003. Effects of artificial defoliation on growth, reproduction and leaf chemistry of the mangrove Kandelia candel . Journal Of Tropical Ecology, 19(Part 4), 397–406. Turner, C, Seastedt, TR, & Dyer, MI. 1993. Maximization of aboveground grassland production - the role of defoliation frequency, intensity, and history. Ecological Applications, 3(1), 175–186. Wan, S, Luo, Y, & Wallace, LL. 2002. Changes in microclimate induced by experimental warming and clipping in tallgrass prairie. Global Change Biology, 8(8), 754–768. Welker, JM, Fahnestock, JT, Povirk, KL, Bilbrough, CJ, & Piper, RE. 2004. Alpine grassland CO2 exchange and nitrogen cycling: Grazing history effects, medicine bow range, Wyoming, USA. Arctic Antarctic and Alpine Research, 36(1), 11–20. 87  Chapter 3. References Western Ag Innovations. 2008 (March). Western Ag Innovations Inc. About PRS-TM-probes (http://www.westernag.ca/innov). Whaley, WG, Mericle, LW, & Heimsch, C. 1952. The wall of the meristematic cell. American Journal of Botany, 39(1), 20–26. Willms, WD, Smoliak, S, & Bailey, AW. 1986. Herbage production following litter removal on Alberta native grasslands. Journal of Range Management, 39(6), 536–540. Willms, WD, Dormaar, JF, Adams, BW, & Douwes, HE. 2002. Response of the mixed prairie to protection from grazing. Journal of Range Management, 55(3), 210–216. Wraith, JM, Johnson, DA, Hanks, RJ, & Sisson, DV. 1987. Soil and plant water relations in a crested wheatgrass pasture - response to spring grazing by cattle. Oecologia, 73(4), 573–578. Zhou, X., R. A. Sherry Y. An L. L. Wallace, & Luo, Y. 2006. Main and interactive effects of warming, clipping, and doubled precipitation on soil CO2 efflux in a grassland ecosystem. Global Biogeochemical Cycles, 20, GB1003, doi:10.1029/2005GB002526.  88  Chapter 4  Summary and Conclusions We conducted two studies to investigate the effects of grazing on High Arctic wet sedge meadow plant communities: an observational study comparing sites with contrasting grazing regimes and histories; and an experimental study examining the responses of ungrazed meadows to clipping and litter removal. Alexandra Fiord (AF) and Sverdrup Pass (SP) on Ellesmere Island were chosen to test the grazing optimization hypothesis and other predictions related to the effects of muskox grazing on these plant communities. Alexandra Fiord and Sverdrup Pass are both polar oases that have similar geological and climatic characteristics (Raillard 1992; Henry 1998). Mammalian grazers are not a significant component of the AF lowland (Raillard 1992; Henry 1998), whereas SP is heavily grazed by muskoxen and functions as a corridor for muskoxen as they move between the large grazing ranges on either side of the island (Raillard & Svoboda 2000). At AF, slowly decomposing unattached and attached plant litter accumulates due to the lack of herbivory (Henry 1998). According to the grazing optimization hypothesis, net primary production (NPP) and nitrogen (N) levels in plants increase with moderate levels of grazing and this typically occurs in ecosystems dominated by graminoids. Therefore, moderate frequencies of defoliation should increase NPP and shoot N concentrations in aboveground plant material in High Arctic sedge meadow plant communities. In our observational study, the grazing optimization hypothesis was supported by more belowground plant biomass, larger 2008 aboveground NPP and graminoid NPP levels, and higher N concentrations in shoots of Carex aquatilis ssp. stans and Eriophorum angustifolium ssp. triste in the grazed sedge 89  Chapter 4. Summary and Conclusions meadows at SP compared to the ungrazed meadows. However, lower aboveground production values in clipped plots in the two experimental sites suggested that compensation in plant production is N-dependent in these plant communities. Similarly, defoliation without N addition reduced aboveground NPP in subarctic salt marsh ecosystems (Hik et al. 1991) and the Serengeti plains (Ruess et al. 1983). Raillard & Svoboda (1999) found that clipped plants compensated for the loss of aboveground biomass; however, those meadows had been grazed in previous years. Plants were defoliated by clipping in our experiment, but this treatment did not simulate all processes associated with grazing. For instance, N was not added to the clipping treatments at AF unlike the grazed meadows at SP that received nutrients from muskox urine and feces. Essentially, muskoxen short-circuited the nutrient cycle. The grazed meadows at SP had higher aboveground NPP in 2008. In contrast to the results from the observational study, aboveground NPP decreased with higher clipping frequencies in wet sedge meadows at the ungrazed site (AF). In addition, NPP was similar regardless of clipping frequency in areas protected by exclosures at SP (Raillard & Svoboda 1999). Possible explanations for these differences include soil NH4 + availability and belowground root production. Soil NH4 + availability was higher at SP than in the ungrazed meadows at AF; however, the availability of this nutrient was similar in the control and 3x treatment in the clipping experiments at AF. This pattern suggests that the compensatory growth in wet sedge meadows at the grazed site (SP) was stimulated by increased N availability at least partly in the form of soil NH4 + . For example, nutrient addition ameliorated the negative effects of clipping on shoot growth in subarctic salt marshes (Hik et al. 1991). Another factor that could have resulted in compensation in aboveground NPP in the grazed meadows at SP was the greater biomass of roots and rhizomes, which increased N uptake and provided a source of carbon for regrowth. Concentrations of N in plant shoots were higher at SP than in the ungrazed meadows 90  Chapter 4. Summary and Conclusions at AF. This indicates that the larger biomass of roots and rhizomes in the plants at SP increased N uptake and translocation, supporting the growth of more nutritious shoots (Bryant et al. 1983). Higher soil NH4 + availability could have also contributed to increased shoot N as this nutrient was more available for uptake by the grazed plants. In addition, clipped plants in the ungrazed meadows at AF had higher N concentrations in their shoots compared those that were not clipped. In areas protected by exclosures at SP, Raillard & Svoboda (1999) found that plants that were clipped more frequently had higher N concentrations in their vegetative regrowth. As aboveground plant material was removed by grazing and experimental clipping in the studies mentioned above, plants appeared to have responded by increasing N uptake and allocating it more to their shoots (Green & Detling 2000). Increases in shoot N correlate with higher rates of photosynthesis (Welker et al. 2004). Therefore, higher N contents in leaves in response to defoliation may be an evolutionary response to repair damaged tissues (Jaramillo & Detling 1988). Plant responses to defoliation at AF differed depending on growth form. In contrast to graminoids, N concentrations in the new growth of Salix arctica were lower in the clipped plots relative to controls. However, the concentrations of N in S. arctica were similar in the grazed and ungrazed meadows. Most of the nutrient allocation in deciduous shrubs is to new growth in the spring (Chapin 1980). It is possible that the S. arctica shoots that were removed in the clipping experiments early in the growing season contained high concentrations of N that were not replaced in the plant biomass by harvest later in the season. Clipping at AF decreased the concentration of carbon in the shoots of C. membranacea, C. stans, and S. arctica. However, there was no difference in carbon concentrations in the shoots of four plant species in the grazed and ungrazed meadows. Whaley et al. (1952) found that younger leaves had lower carbon concentrations because they had less cellulose in their cell walls. In the experimental study at AF, plants that were clipped more frequently 91  Chapter 4. Summary and Conclusions had younger leaves. Therefore, the newer cells in the younger leaves may have had lower proportions of cellulose in their walls. Additionally, carbon may have been diluted in the fresh biomass of leaves during early regrowth following defoliation (deVisser et al. 1997). Ecosystem respiration was similar in the grazed and ungrazed meadows, although the trend indicated that it was higher in the grazed meadows at SP. At AF, ecosystem respiration was higher in control plots compared to the 3x clipping treatment. According to Johnson et al. (2000), plants are the main source of ecosystem respiration in wet sedge meadows. Possibly, higher ecosystem respiration rates in the grazed meadows at SP and in the control plots at AF can be explained by increased levels of NPP in those areas. There was more live plant material to contribute to ecosystem respiration. Soil moisture was only measured in the experimental study and its levels increased with higher clipping frequencies. This suggests that evapotranspiration rates declined because surface areas of leaf transpiring surfaces were reduced (Archer & Detling 1986; Wraith et al. 1987). The clipping experiment was conducted in ungrazed meadows at AF where there were large accumulations of plant litter. Litter removal treatments were only applied to Site 2 because Site 1 did not have a large build-up of litter. Plant growth is often limited because the interception of light is blocked by old and dead plant tissues that have accumulated (Mulder 1999). However, in our study aboveground NPP decreased in response to litter removal in the ungrazed meadows at AF. Also, litter removal caused soil volumetric water contents to rise marginally. Our data did not indicate any differences in soil temperatures; therefore, air temperatures most likely were the same over both the control and litter removal plots. Since energy balance at the soil surface considers sensible heat fluxes to the atmosphere and soil along with latent heat flux (Oke 1978), the increased energy available at the soil surface may have been used in evaporation, drawing additional soil water into the plot. 92  Chapter 4. Summary and Conclusions There are both strengths and weaknesses associated with the research conducted in the two studies presented in this thesis. Comparing the grazed meadows at SP to the ungrazed meadows at AF allowed us to substitute space for time. We conducted our experiments over only two years, even though SP was grazed and AF remained ungrazed for probably hundreds of years. However, the analysis of annual changes in aboveground primary production was possible because the studies were conducted over a two year period. Combining the results of the observational component with those of the clipping experiment clarified some of the connections between nutrient cycling and NPP in High Arctic sedge meadow plant communities. The results of the experimental study isolated the effects of defoliation on these plant communities. Weaknesses in the experimental study include the low number of sites and few replications within treatments; although this is a common weakness in ecological field experiments. Statistical power could have been increased by adding more clipping grids with larger numbers of replications. Also, the 1x + LR treatment should have been applied to both clipping sites to aid in data interpretation. In the observational study comparing grazed and ungrazed meadows, it would have been best to have sampled at more sites with fewer subsamples per sample. For the most part, logistical constraints determined the design of the clipping grids and the observational study. The results from the comparison of the grazed and ungrazed meadows and the experimental clipping grids present future research opportunities. The effects of clipping and litter removal on different plant species could be investigated to evaluate if individual species respond differently to these treatments. To learn more about responses to grazing in different biomes, sedge meadows dominated by C. aquatilis ssp. aquatilis common in more southern climes could be subjected to similar treatments. Also, a better understanding of the differences in C cycling in grazed and ungrazed system could be gained by measuring net ecosystem exchange. The effects of grazing on ecosystem function have been studied in detail in grazed ecosys93  Chapter 4. Summary and Conclusions tems that occur in more southern latitudes. In contrast, little research has focused on grazing in High Arctic sedge meadows. The research outlined in this thesis increases our understanding of the relation between grazers and grazed ecosystems and elucidates connections between primary productivity and nutrient cycling in High Arctic sedge meadow plant communities.  94  References Archer, S, & Detling, JK. 1986. Evaluation of potential herbivore mediation of plant water status in a North-American mixed-grass prairie. Oikos, 47(3), 287–291. Bryant, JP, Chapin, FS, & Klein, DR. 1983. Carbon nutrient balance of boreal plants in relation to vertebrate herbivory. Oikos, 40(3), 357–368. Chapin, FS. 1980. Nutrient allocation and responses to defoliation in tundra plants. Arctic and Alpine Research, 12(4), 553–563. deVisser, R, Vianden, H, & Schnyder, H. 1997. Kinetics and relative significance of remobilized and current C and N incorporation in leaf and root growth zones of Lolium perenne after defoliation: Assessment by C-13 and N-15 steady-state labelling. Plant Cell and Environment, 20(1), 37–46. Green, RA, & Detling, JK. 2000. Defoliation-induced enhancement of total aboveground nitrogen yield of grasses. Oikos, 91(2), 280–284. Henry, GHR. 1998. Environmental influences on the structure of sedge meadows in the Canadian High Arctic. Plant Ecology, 134(1), 119–129. Hik, DS, Sadul, HA, & Jefferies, RL. 1991. Effects of the timing of multiple grazings by geese on net aboveground primary production of swards of Puccinellia phryganodes. Journal Of Ecology, 79(3), 715–730. Jaramillo, VJ, & Detling, JK. 1988. Grazing history, defoliation, and competition - effects on shortgrass production and nitrogen accumulation. Ecology, 69(5), 1599–1608. Johnson, LC, Shaver, GR, Cades, DH, Rastetter, E, Nadelhoffer, K, Giblin, A, Laundre, J, & Stanley, A. 2000. Plant carbon-nutrient interactions control CO2 exchange in Alaskan wet sedge tundra ecosystems. Ecology, 81(2), 453–469. Mulder, CPH. 1999. Vertebrate herbivores and plants in the arctic and subarctic: effects on individuals, populations, communities and ecosystems. Perspectives in Plant Ecology, Evolution and Systematics, 2, 29–55. Oke, TR. 1978. Boundary Layer Climates. London: Mehuen & Co Ltd. Raillard, M. 1992. Influence of muskox grazing on plant communities of Sverdrup Pass (79 N), Ellesmere Island, N.W.T. Canada. Ph.D. thesis, University of Toronto. 95  Chapter 4. References Raillard, M, & Svoboda, J. 1999. Exact growth and increased nitrogen compensation by the arctic sedge Carex aquatilis var. stans after simulated grazing. Arctic, Antarctic, and Alpine Research, 31, 21–26. Raillard, M, & Svoboda, J. 2000. High grazing impact, selectivity, and local density of muskoxen in Cental Ellesmere Island, Canadian High Arctic. Arctic Antarctic and Alpine Research, 32(3), 278–285. Ruess, RW, McNaughton, SJ, & Coughenour, MB. 1983. The effects of clipping, nitrogensource and nitrogen concentration on the growth-responses and nitrogen uptake of an east-african sedge. Oecologia, 59(2-3), 253–261. Welker, JM, Fahnestock, JT, Povirk, KL, Bilbrough, CJ, & Piper, RE. 2004. Alpine grassland CO2 exchange and nitrogen cycling: Grazing history effects, medicine bow range, Wyoming, USA. Arctic Antarctic and Alpine Research, 36(1), 11–20. Whaley, WG, Mericle, LW, & Heimsch, C. 1952. The wall of the meristematic cell. American Journal of Botany, 39(1), 20–26. Wraith, JM, Johnson, DA, Hanks, RJ, & Sisson, DV. 1987. Soil and plant water relations in a crested wheatgrass pasture - response to spring grazing by cattle. Oecologia, 73(4), 573–578.  96  

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