Effects of changes in climate and UV radiation levels on structure of arctic ecosystems in the short and long term

This is Section 7.4.1 of the Arctic Climate Impact Assessment
Lead Author: Terry V. Callaghan; Contributing Authors: Lars Olof Björn, F. Stuart Chapin III,Yuri Chernov,Torben R. Christensen, Brian Huntley, Rolf Ims, Margareta Johansson, Dyanna Jolly Riedlinger, Sven Jonasson, Nadya Matveyeva,Walter Oechel, Nicolai Panikov, Gus Shaver; Consulting Authors: Josef Elster, Heikki Henttonen, Ingibjörg S. Jónsdóttir, Kari Laine, Sibyll Schaphoff, Stephen Sitch, Erja Taulavuori, Kari Taulavuori, Christoph Zöckler

Local and latitudinal variation (7.4.1.1)

The Arctic is characterized by ecosystems that lack trees. There is a broad diversity in ecosystem structure among these northern treeless ecosystems that follows a latitudinal gradient from the treeline to the polar deserts. Typical communities for a particular latitude are called "zonal", but local variation at the landscape level occurs and these "intrazonal communities" are frequently associated with variations in soil moisture and snow accumulation[4].

 

caption Fig. 7.16. Forest tundra vegetation represented by the Fennoscandian mountain birch forest, Abisko, northern Sweden. (Source: Photograph by T.V. Callaghan)
caption Fig. 7.17. Zonal tussock tundra near Toolik Lake, Alaska, with large shrubs/small trees of ''Salix'' in moist sheltered depressions. (Source: Photograph by T.V. Callaghan)

 

According to Bliss and Matveyeva[5], zonal communities south of the arctic boundary near the mean July isotherms of 10 to 12 °C consist of taiga (i.e., the northern edge of the boreal forest). This is characterized by a closed-canopy forest of northern coniferous trees with mires in poorly drained areas. To the north of this transition zone is the forest tundra. It is characterized by white spruce (Picea glauca) in Alaska, mountain birch (Betula pubescens ssp. czerepanovii) in Fennoscandia (Fig. 7.16), by birch and Norway spruce in the European Russian Arctic (Kola Peninsula and the Pechora lowlands), by Dahurian larch (Larix dahurica) in central and eastern Siberia, and by evergreen coniferous trees in Canada[6]. The vegetation of the forest tundra is characterized by sparse, low-growing trees with thickets of shrubs. North of this zone is the low Arctic, which is characterized by tundra vegetation in the strict sense (Fig. 7.17), consisting of communities of low, thicket-forming shrubs with sedges, tussock-forming sedges with dwarf shrubs, and mires in poorly drained areas. To the north of this zone is the high Arctic, which consists of polar semi-desert communities (Fig. 7.18) in the south, characterized by cryptogam–herb, cushion plant–cryptogam, and, to a limited extent, mire communities. To the extreme north is the polar desert where only about 5% of the ground surface is covered by herb–cryptogam communities (Fig. 7.19). In this zone, the mean July temperature is below 2 °C and precipitation, which falls mainly as snow, is about 50 millimeters (mm) per year.

 

caption Fig. 7.18. Polar semi-desert dominated by mountain avens (''Dryas octopetala''), Ny Ålesund, Svalbard. (Source: Photograph by T.V. Callaghan)
caption Fig. 7.19. Polar desert, Cornwallis Island, Northwest Territories, Canada. (Source: Photograph by J. Svoboda)

 

The tundra zone can be further subdivided into three subzones: the southern tundra with shrub–sedge, tussock–dwarf shrub, and mire communities; the typical tundra with sedge–dwarf shrub and polygonal mire communities (Fig. 7.20); and the northern arctic tundra that consists of dwarf shrub–herb communities. The northern end of the latitudinal gradient, occurring primarily on islands and on the mainland only at Cape Chelyuskin (Taymir Peninsula), is occupied by polar deserts where woody plants are absent, and forbs and grasses with mosses and lichens are the main components of plant communities[7].

This vegetation classification has geographic connotations and cannot be applied easily to reconstructions of past vegetation throughout the circumpolar Arctic[8]. A recent classification of tundra vegetation at the biome level[9] has been proposed by Kaplan et al.[10] (Table 7.8, Fig. 7.2).

Within the biomes or zonal vegetation types, there are intrazonal habitats that are frequently associated with variations in soil moisture and snow accumulation, and that have a microclimate that deviates from the general macroclimate associated with flat surfaces. The intrazonal habitats form a mosaic of communities. Each of these tend to have fewer species than the "plakor," or zonal, communities. For example, poorly drained areas are often dominated by sedges with an understory of mosses and liverworts, but lack fruticose lichens[11]. Although each intrazonal community has relatively few species, together they are more differentiated and diverse than zonal ones, and are responsible for about 80% of total species diversity in the regional flora and fauna. Disturbances, particularly freeze–thaw cycles and thermokarstIrregular surace landforms resulting from the thawing of ice-rich permafrost (Fig. 7.21) that form patterned ground, also create landscape mosaics (Fig. 7.20). Diversity "focal points/hot spots"[12] and "oases"[13] enrich landscapes by possessing a larger number of species, including those of more southerly distribution. Examples include dense willow thickets two meters in height in sheltered valleys at 75° N in Taymir and stands of balsam poplar north of the treeline in the northern foothills of the Brooks Range, Alaska, that are likely to respond rapidly to warming. There are numerous other types of plant communities, such as the moss-dominated tundra of Iceland (Fig. 7.22).

caption Fig. 7.20. Polygonal wet tundra near Prudhoe Bay, Alaska (photo:T.V. Callaghan).
caption Fig. 7.21. Thermokarst in the Russian tundra, New Siberian Islands. (Source: Photograph by T.V. Callaghan)

 

The vertical structure of arctic ecosystems is as important as horizontal structure in explaining their current and future functioning. This structure is most pronounced in low-arctic shrub communities, where there is a well-developed shrub canopy and an understory of mosses, similar to the vertical structure of boreal forests. Vertical structure is also pronounced below ground, with mosses and lichens lacking roots, some species rooted in the moss layer, others rooted just beneath the mosses, and a few species rooted more deeply.

The most striking latitudinal trend in plant functional types is the decrease in height of woody plants (from trees to tall shrubs, to low and prostrate shrubs, to dwarf shrubs, and eventually the loss of woody plants with increasing latitude). These functional types often occur in low abundance in zones north of their main areas of dominance, suggesting that they are likely to expand rapidly in response to warming through vegetative reproduction[14] and sexual reproduction[15], although range expansion will depend on geographic barriers such as mountains and seas (section 7.6, Table 7.14). Recent warming in Alaska has caused a substantial increase in shrub density and size in the moist tundra of northern Alaska[16]. In areas where shrubs are absent, shrubs are likely to exhibit time lags in migrating to new habitats[17]. Shrubs colonize most effectively in association with disturbances such as flooding in riparian zones, thermokarst, and frost boils (patterned ground formation caused by soil heave) throughout their latitudinal range, so migration may be strongly influenced by climate- or human-induced changes in the disturbance regime. Woody species affect ecosystem structure and function because of their potential to dominate the canopy and reduce light availability to understory species[18] and to reduce overall litter quality[19] and rates of nutrient cycling.

Table 7.8. Circumpolar tundra biome classification.[20]
Biome Definition Typical taxa
Low- and high-shrub tundra Continuous shrubland, 50 cm to 2 m tall, deciduous or evergreen, sometimes with tussock-forming graminoids and true mossesa bog mosses, and lichens Alnus, Betula, Salix, Pinus pumila (in eastern Siberia), Eriophorum, Sphagnum
Erect dwarf-shrub tundra Continuous shrubland 2 to 50 cm tall, deciduous or evergreen, with graminoids, true mossesa, and lichens Betula, Cassiope, Empetrum, Salix, Vaccinium, Poaceae, Cyperaceae
Prostrate dwarf-shrub tundra Discontinuous “shrubland” of prostrate deciduous dwarf-shrubs 0 to 2 cm tall, true mossesa, and lichens Salix, Dryas, Pedicularis, Asteraceae, Caryophyllaceae, Poaceae, true mossesa
Cushion forb, lichen, and moss tundra Discontinuous cover of rosette plants or cushion forbs with lichens and true mossesa Saxifragaceae, Caryophyllaceae, Papaver, Draba, lichens, true mossesa
Graminoid and forb tundra Predominantly herbaceous vegetation dominated by forbs, graminoids, true mossesa, and lichens Artemisia, Kobresia, Brassicaceae, Asteraceae, Caryophyllaceae, Poaceae, true mossesa
a"true" mosses exclude the genus Sphagnum

 

caption Fig. 7.22. ''Racomitrium''/''Empetrum'' heath in Iceland showing erosion. (Source: Photograph by T.V. Callaghan)

A similar latitudinal decline in abundance occurs with sedges, which are absent from polar deserts, suggesting that this group is also likely to expand northward with warming[21]. Carex stans and C. bigelowii now mark the northernmost boundary of the tundra zone and might be a sensitive indicator of species responses to warming. Sedges have important effects on many ecosystem processes, including methane flux, because of their transport of oxygen to soils, transport of methane to the atmosphere, and inputs of labile carbon to the rhizosphere[22]. Prostrate and dwarf shrubs such as Dryas spp., arctic willow, and polar willow are likely to decline in abundance with warming in the low Arctic, due to competition with taller plants, but are likely to increase in abundance in the current polar deserts. These changes in distribution are very likely to substantially reduce the extent of polar desert ecosystems (section 7.5.3.2), which are characterized by the absence of woody plants.

 

Response to experimental manipulations (7.4.1.2)

Experimental manipulation of environmental factors projected to change at high latitudes (temperature, snow, nutrients, solar radiation, atmospheric CO2 concentrations, and UV-B radiation levels) has substantial effects on the structure of arctic ecosystems, but the effects are regionally variable. The effects of these variables on individual species are discussed in section 7.3; while this section focuses on overall community structure and species interactions.

Plant communities

 

caption Fig. 7.23. Results of long-term (generally 10 years or more) experiments in a range of habitats at Toolik Lake, Alaska, and Abisko, Sweden, showing the responsiveness of aboveground biomass ordered by treatment and degree of responsiveness (X is the mean value of the analyzed characteristic for the experimental (Xe) and control (Xc) groups). Data are given for (a) total vascular plant biomass and (b) lichen biomass. Codes relate to the geographical region (To=Toolik, Ab=Abisko), the site name, and the duration of the experiment[1].

 

Nutrient addition is the environmental manipulation that has the greatest effect on the productivity, canopy height, and community composition of arctic plant communities[23] (Fig. 7.23). Fertilization also increases biomass turnover rate, so eventual biomass may or may not change in response to nutrient addition. In northern Sweden, for example, nutrient addition to a mountain birch (Betula pubescens ssp. czerepanovii) site (cf. a Swedish treeline heath and fellfield) caused an initial biomass increase. This biomass increase was not maintained over the long term, however, because expansion of the grass Calamagrostis lapponica negatively affected the growth of mosses and evergreen shrubs, leading to a negligible change in community biomass[24]. Similarly, addition of nitrogen and phosphorus at a site in northern Alaska increased productivity and turnover within three years[25]. There was, however, little change in biomass because the rapidly growing sedges, forbs, and deciduous shrubs responded most strongly, whereas evergreen shrubs and mosses declined in abundance (Fig. 7.24). After 9 and 15 years, competitive interactions altered the relative abundance of plant functional types, with the tallest species (the deciduous shrub Betula nana; Fig. 7.3) responding most strongly[26]. Litter and/or shade from this species reduced the growth of lichens, mosses, and evergreen shrubs. In vegetation types without any pronounced change in relative proportions of dominant species or life forms following fertilizer addition, as in Swedish treeline and high-altitude heaths and in Alaskan wet-sedge tundra, the biomass of most dominant life forms increased. This resulted in up to a doubling of biomass after five to nine years of treatment[27]. In polar semi-deserts, nutrient addition generally had a negative effect on vascular plants, due to enhanced winterkill, but stimulated the growth of mosses[28], an effect opposite to that in low-arctic tundra. This difference is probably due to the immigration of nitrogen-demanding mosses from nearby bird-cliff communities in the high Arctic compared with loss of existing moss species in the low Arctic.

 

caption Fig. 7.24. Effects of long-term fertilizer addition and experimental warming and shading during the growing season on aboveground net primary production (NPP) of different plant functional types at Toolik Lake, Alaska, showing NPP by functional type and treatment (a) in 1983, after three years of treatment, and (b) in 1989, after nine years of treatment[2].

 

Water additions to simulate increased precipitation have generally had only minor effects on total biomass and production[29].

Experimental summer warming of tundra vegetation within the range of projected temperature increases (2 to 4 °C over the next 100 years) has generally led to smaller changes than fertilizer addition[30] (Fig. 7.23). For example, temperature enhancement in the high-arctic semi-desert increased plant cover within growing seasons but the effect did not persist from year to year[31]. In the low Arctic, community biomass and nutrient mass changed little in response to warming of two Alaskan tussock sites[32] and two wet-sedge tundra sites[33], coincident with relatively low changes in soil nutrient pools and net mineralization. Tussock tundra showed little response to warming, as some species increased in abundance and others decreased[34], similar to a pattern observed in subarctic Swedish forest floor vegetation[35]. The responses to warming were much greater in Swedish treeline heath and in fellfield[36]. Biomass in the low-altitude heath increased by about 60% after air temperatures were increased by about 2.5 °C, but there was little additional effect when temperatures were further increased by about 2 °C. In contrast, biomass approximately doubled after the first temperature increase (2.5 °C) and tripled after the higher temperature increase (an additional 2 °C) in the colder fellfield. Hence, the growth response increased from the climatically relatively mild forest understory through the treeline heath to the cold, high-altitude fellfield where the response to warming was of the same magnitude as the response to fertilizer addition[37]. A general long-term (10 years or more) response to environmental manipulations at sites in subarctic Sweden and in Alaska was a decrease in total nonvascular plant biomass and particularly the biomass of lichens[38] (Fig. 7.23).

Animal communities

Air-warming experiments at Svalbard (79° N) had greater effects on the fauna above ground than below ground, probably because the soil is more buffered against fluctuations in temperature and moisture than the surface[39]. Species with rapid life cycles (aphids and Collembola) responded demographically more quickly than species (e.g., mites) with slow life cycles[40]. Responses to warming differed among sites.The abundance of Collembola declined at barren sites where higher temperatures also caused drought and mortality due to desiccation, whereas the abundance of Collembola increased at moister sites. In summer, water availability is probably much more important to many invertebrates than is temperature. Mites are more resistant than Collembola to summer desiccation[41] and to anoxic conditions in winter due to ice-crust formation following episodes of mild weather[42]. Ice-crust formation during the winter may increase winter mortality by 50% in Collembola[43]. Freeze–thaw events in spring may also cause differential mortality among species, thus altering community composition[44]. In experiments conducted simultaneously at several sites and over several years, the natural spatial and temporal variability in community structure and population density of soil invertebrates was larger than the effects of the experimental manipulation within years and sites. This demonstrates that there is a large variability in the structure and function of high-arctic invertebrate communities due to current variation in abiotic conditions. It also indicates that arctic invertebrate communities can respond rapidly to change.

Compared to the high Arctic, subarctic invertebrate communities at Abisko responded less to experimental temperature increases[45]. However, nematode population density increased substantially, and the dominance changed in favor of plant- and fungal-feeding species with elevated summer temperatures and nitrogen (N), phosphorus (P), and potassium (K) fertilization, indicating a shift in the decomposition pathway[46].

Microbial communities

Although the biomass of microorganisms is a poor predictor of the productivity and turnover of microorganisms and their carbon, alternative methods focusing on population dynamics of microbial species within communities are extremely difficult to employ in the field. Microbial biomass has therefore been used to quantify microbial processes within ecosystems. The sensitivity of microbial biomass, generally measured as biomass carbon (C), and nutrient content to changed environmental conditions in the Arctic has not been well studied. Long-term addition of easily processed C generally increases the microbial biomass; and addition of inorganic nutrients generally, but not always, increases microbial nutrient content without appreciable effect on the biomass[47]. In some cases, however, a combination of C and nutrient addition has led to a pronounced increase in both microbial biomass and nutrient content[48]. This suggests a general C limitation of microbial biomass production, and increased sink strength for soil nutrients (i.e., increased sequestration of nutrients) if the amounts of both labile C and nutrients increase, but relatively weak effects from increased nutrient availability alone. In the widespread drier ecosystem types in the Arctic, the soil microbial biomass is likely to be further limited by low water supply. Water addition to a high-arctic semi-desert led to a substantial increase in microbial biomass C and microbial activity[49].

Data on the effects on ecosystems of growing-season temperature increases of 2 to 4 °C over five[50] and ten[51] years have not shown appreciable long-term changes in microbial biomass and nutrient stocks. This suggests that an increase in growing-season temperature alone is unlikely to have any strong impact on microbial C and nutrient sequestration, and that changes in soil nutrient availability are likely to lead to greater changes than the direct effects of increased temperature. Temperature effects on ecosystem processes are likely, however, to be different from the observed relatively small effects on microbial biomass and nutrient stocks, because temperature changes are likely to affect rates of decomposition and nutrient mineralization, rather than pool sizes, resulting in altered C balance and rates of nutrient supply to the plants (section 7.4.2.1).

Appreciable seasonality in microbial biomass and nutrient mass have been reported, however, that seemingly are independent of ambient temperature. In general, the masses change little, or fluctuate, during summer[52]. In contrast, pronounced increases in both biomass and nutrient mass have been reported in autumn[53], probably as a function of increased input of labile C and nutrients from plants as they senesce, although these data are from mountain and alpine, rather than from arctic soils. The increase seems to continue through winter, although at a slower rate[54], despite soil temperatures below 0 °C[55] (see section 7.3.2.3). It is followed by a sharp biomass decline in the transition between winter and spring[56], which may[57] or may not[58] coincide with a decrease in microbial N and an increase in mineralized N, indicating a pronounced transformation of microbial N to soil inorganic N[59]. Indeed, this seasonal pattern suggests a temporal partitioning of resource uptake with low competition between plants and microbes for nutrients, as microbes absorb most nutrients in autumn and plants in spring, coincident with the nutrient release from declining microbial populations. However, it may also be an indication that plants compete well for nutrients during the growing season[60], and microbes access nutrients efficiently only when the sink strength for nutrients in plants is low[61].

Laboratory experiments have shown that the spring decline of microbial mass is likely to be an effect of repeated freeze–thaw cycles[62]. Indeed, Larsen et al.[63] reported a microbial decline in soils subjected to repeated freezing and thawing but not in the same soils kept constantly frozen before thawing. The seasonal dynamics of microbial biomass and microbial and inorganic soil nutrients therefore suggests that “off growing-season” changes in climate during the transition between winter and spring (e.g., changed frequency of freeze–thaw events and warmer winters) are likely to have greater impacts on nutrient transformations between microbes, soils, and plants than changes during the growing season.

Manipulations simulating enhanced UV-B radiation levels (equivalent to a 15% reduction in stratospheric ozone levels) and a doubling of atmospheric CO2 concentrations for seven years altered the use of labile C substrates by gram-negative bacteria[64]. Although these rhizosphere bacteria are a relatively small component of the belowground microbial biomass, they are likely to be particularly responsive to environmentally induced changes in belowground plant C flow.

Ultraviolet-B radiation also affects the structure of fungal communities. Microcosms of subarctic birch forest floor litter exposed to enhanced UV-B radiation levels showed a reduction in fungal colonization of leaf veins and lamina[65]. Fungal composition was also altered in the UV-B radiation treatments, with a reduction in Mucor hiemalis and a loss of Truncatella truncata. Similar findings of fungal community change were obtained in subarctic Abisko, in an ecosystem that was the source of the litter used by Gehrke et al.[66]. In this field study of the decomposition rates of a standard litter type, there was also a change in the composition of the fungal community associated with litter resulting from elevated UV-B radiation levels[67]. So far, no change in plant community structure has been found in the Arctic in response to artificially enhanced or reduced UV-B radiation levels and CO2 concentrations.

Recent decadal changes within permanent plots (7.4.1.3)

caption Fig. 7.25. Changes in greenness of northern vegetation (depicted by NDVI) between 1981 and 1991 as measured by satellite instruments[3].

Satellite measurements suggest a widespread increase in indices of vegetation greenness (e.g., the normalized difference vegetation index – NDVI) and biomass at high latitudes[68] (Fig. 7.25), although changes in satellites and sensor degradation may have contributed to this trend[69]. Aerial photographs show a general increase in shrubbiness in arctic Alaska[70] and indigenous knowledge also reports an increase in shrubbiness in some areas. These observations are consistent with the satellite observations. However, it has been difficult to corroborate these with studies of permanent plots, because of the paucity of long-term vegetation studies in the Arctic. In arctic Alaska, for example, a trend toward reduced abundance of graminoids and deciduous shrubs during the 1980s was reversed in the 1990s[71]. In Scandinavia, decadal changes in vegetation were affected more strongly by the cyclic abundance of lemmings than by climatic trends[72].

Trophic interactions (7.4.1.4)

Trophic-level structure is simpler in the Arctic than further south. In all taxonomic groups, the Arctic has an unusually high proportion of carnivorous species and a low proportion of herbivores[73]. As herbivores are strongly dependent on responses of vegetation to climate variability, warming is very likely to substantially alter the trophic structure and dynamics of arctic ecosystems. The herbivore-based trophic system in most tundra habitats is dominated by one or two lemming species[74] while the abundance of phytophagous (plant-eating) insects relative to plant biomass is low in arctic tundra[75]. Large predators such as wolves, wolverines, and bears are less numerous in the tundra than in the boreal forest[76] and predation impacts on tundra ungulates are usually low. Thus, the dynamics and assemblages of vertebrate predators in arctic tundra are almost entirely based on lemmings and other small rodent species (Microtus spp. and Clethrionomys spp.[77]), while lemmings and small rodents consume more plant biomass than other herbivores. Climate has direct and indirect impacts on the interactions among trophic levels, but there is greater uncertainty about the responses to climate change of animals at the higher trophic levels.

Plant-herbivore interactions

Plant tissue chemistry and herbivory

Arctic and boreal plant species often contain significant concentrations of secondary metabolites that are important to the regulation of herbivory and herbivore abundance[78]. Secondary compounds also retard decomposition of leaves after litter fall[79]. These secondary metabolites are highly variable in their chemical composition and in their anti-herbivore effects, both within and among species. One hypothesis about the regulation of these compounds that has received widespread discussion is the carbon–nutrient balance hypothesis of Bryant et al.[80], which attempts to explain this variation in part on the basis of C versus nutrient limitation of plant growth. Although many other factors in addition to carbon–nutrient balance are probably important to the regulation of plant–herbivore interactions in the Arctic (e.g., [81]), the abundance of secondary chemicals is often strongly responsive to changes in the environment including temperature, light, and nutrient availability (e.g., [82]). In a widespread arctic shrub species, Betula nana (Fig. 7.3), Graglia et al.[83] found that fertilization and shading generally led to decreased condensed and hydrolyzable tannin concentrations in leaves, whereas warming in small field greenhouses increased condensed tannins and decreased hydrolyzable tannins. There was also a large difference in both the average concentrations and the responsiveness of the concentrations of phenolics in plants from northern Alaska versus northern Sweden, with the plants from Sweden having generally higher concentrations but being less responsive to environmental changes. Such data suggest that the effects of climate change on plant–herbivore interactions are likely to be highly variable, species-specific, and also dependent on the nature of the change and on ecotypic or subspecific differences, perhaps related to local evolution in the presence or absence of herbivores.

Plant exposure to UV-B radiation has the ability to change the chemistry of leaf tissues, which has the potential to affect the odor that herbivores such as reindeer/caribou use to detect food, and the quality of food in terms of palatability and digestibility[84]. In general, enhanced UV-B radiation levels can reduce soluble carbohydrates and increase phenolic compounds and flavonoids. Such changes are expected to reduce forage quality.

Plant exposure to increased CO2 concentrations can also affect plant tissue quality and consequently herbivory[85]. Enriched CO2 concentrations may lead to the accumulation of carbohydrates and phenolic compounds while reducing N concentrations in leaves. However, these phytochemical responses can be significantly modified by the availability of other resources such as nutrients, water, and light. Unfortunately, little information about the impacts of increased CO2 concentrations on herbivory is available for the Arctic.

Herbivore abundance and vegetation production

Invertebrates

Insect population outbreaks seldom extend into the tundra. However, in the forest near the treeline, insect defoliators can have devastating impacts on the ecosystem. Climate change is very likely to modify the population dynamics of such insects in several ways[86]. In the autumnal moth, eggs laid on birch twigs in autumn cannot tolerate winter temperatures lower than -36 °C. For this reason, the moth is destroyed in parts of the terrain (e.g., depressions) where winter temperatures drop below this critical minimum[87]. Warmer winters are very likely to reduce winter mortality and possibly increase outbreak intensity. Moreover, lower minimum temperatures are likely to allow the autumnal moth and the related, less cold-tolerant winter moth (Operophtera brumata) to extend their geographic distributions into continental areas with cold winters[88]. However, predicting the effect of a changing climate is not straightforward because moth responses are season-specific. For instance, increasing spring temperatures are likely to cause a mismatch between the phenology of birch leaves and hatching of larvae that are currently synchronized[89]. Moreover, natural enemies such as parasitoid wasps and ants are likely to increase their abundances and activity rates if summer temperature rises. Currently, there is cyclicity in the populations of the autumnal moth and outbreak proportions occur approximately every 10 to 11 years[90]. The defoliated forests require about 70 years to attain their former leaf area, although insect outbreaks in subarctic Finland followed by heavy reindeer browsing of regenerating birch shoots have led to more or less permanent tundra[91]. There are no population outbreaks in the autumnal moth further south in Fennoscandia, most likely due to the high abundance of generalist parasitoids that keep moth populations below outbreak levels[92]. However, the border between outbreaking and non-outbreaking populations of geometrid moths is likely to move northward if climate changes.

Enhanced UV-B radiation levels applied to birch leaves alters the chemistry or structure of the leaves such that caterpillars eat three times as much leaf biomass to maintain body development[93]. There is also a tendency for enhanced UV-B radiation levels to increase the immunocompetence of the caterpillars, which could possibly make them more tolerant to the wasp parasitoid[94]. Although the effects of winter warming on eggs, increased UV-B radiation levels on leaves, and immunocompetence on caterpillars are likely to increase future damage to subarctic birch forests, it is not known to what extent other processes susceptible to spring and summer climate variability may alleviate these effects.

Vertebrates

The herbivore-based trophic system in most tundra habitats is dominated by one or two lemming species[95]. Lemming abundance is the highest in coastal tundra, especially in moist sedge meadows that are the optimum habitat for Lemmus. Collared lemmings (Dicrostonyx) usually do not reach as high densities in their preferred habitats on drier ridges where herbs and dwarf shrubs dominate. Voles (Microtus and Clethrionomys spp.) are likely to become more abundant than lemmings in some low-arctic tundra habitats and forest tundra[96]. At the landscape scale, lemmings and voles are very patchily distributed according to the abundance of their preferred food plants, as well as the distribution of snow[97] (section 7.3.2.2). Lemming peak densities exceed 200 individuals per hectare in the most productive Lemmus habitats in both Siberia and North America[98] and the standing crop of lemmings may approach 2.6 kilograms (kg) dry weight per hectare. The population builds up during the winter (due to winter breeding) and peak densities may be reached in late winter/early spring when the standing crop of food plants is minimal. The diet of Lemmus consists mainly of mosses and graminoids, while Dicrostonyx prefers herbs and dwarf shrubs[99]. Lemmings have a high metabolic rate, and Lemmus in particular has a low digestive efficiency (about 30%, compared to 50% in other small rodents). Consequently, their consumption rate and impact on the vegetation exceeds that of all other herbivores combined (with the exception of local effects of geese near breeding colonies). Moreover, lemmings destroy much more vegetation than they ingest and after population peaks typically 50% of the aboveground biomass has been removed by the time the snow melts[100]. In unproductive snow beds, which are favored winter habitats of the Norway lemming[101], up to 90–100% of the mosses and graminoids present during the winter may be removed[102]. If winters become so unfavorable for lemmings that they are unable to build up cyclic peak densities, the species-rich predator community relying on lemmings is likely to collapse (see next subsection). Moreover, their important, pulsed impact on vegetation as a result of grazing and nutrient recycling is likely to cease. Changes in snow conditions, relative abundances of preferred food plants, and climate impacts on primary production are all very likely to affect lemming populations, and are likely to result in a northward displacement of the climatically determined geographic borders between cyclic and non-cyclic populations of small herbivores (small rodents and moths), as well as the species distributions per se.

Wild populations of other herbivorous mammal species in the tundra, such as hares, squirrels, muskox, and reindeer/caribou, never reach population densities or biomass levels that can compare with peak lemming populations[103]. Moving herds of reindeer/caribou represent only patchy and temporary excursions in numbers, biomass, and impacts on vegetation; averaged over space and time some of the largest herds approach only 0.01 individuals and 0.5 kg of dry weight per hectare[104] on their summer pastures and usually take less than 10% of the vegetation[105]. The only cases where reindeer/caribou have been shown to have large impacts on vegetation seem to be in unusual circumstances (stranding on islands[106]) or under human intervention (e.g., removing top predators or introductions to islands) where overshooting reindeer/caribou populations have led to vegetation destruction, habitat degradation, and subsequent population crashes.

Although the cooling since the mid-1970s in the Hudson Bay region has affected the reproduction of snow geese (Anser caerulescens), the mid-continental population is currently growing by 5% per year[107]. This, in combination with the staging of snow geese in La Pérouse Bay, Manitoba, because of bad weather further north, leads to increasing foraging for roots and rhizomes of the graminoids Puccinellia phryganodes and Carex subspathacea[108]. The rate of removal of belowground organs in the salt marshes combined with intense grazing of sward during summer exceeds the rate of recovery of the vegetation. It is estimated that geese have destroyed 50% of the salt marsh graminoid sward of La Pérouse Bay since 1985. This loss of vegetation cover exposes the sediments of the salt marshes, which have become hypersaline (salinities exceeding 3.2) as a result of increased evapotranspiration. This further reduces plant growth and forage availability to the geese. In turn, this is reducing goose size, survivorship, and fecundity. Other factors that are affected by the trophic cascades initiated by the geese include reduced N mineralization rates and declines in the populations of soil invertebrates, waders, and some species of duck such as the widgeon (Anas americana).

Cyclic populations

Herbivore–plant interactions have been proposed to produce population cycles in arctic herbivores through several mechanisms including nutrient recycling[109], production cycles inherent in food plants[110], induced chemical defense in plants[111], and recurrent overgrazing[112]. The empirical evidence is mixed. There is at least partly supporting evidence for induced chemical defense in the Epirrita–birch system[113] and for overgrazing in the Lemmus–plant system in unproductive tundra habitats[114]. There is little evidence, however, for mechanisms involving nutrient cycling and chemical defense in the case of lemmings and voles[115]. Climate is somehow involved in all the hypotheses of population cycles related to plant–herbivore interactions. For example, allocation strategies in plants and the amount of secondary compounds (induced chemical defense hypothesis) depend on temperature and growing-season length (see plant tissue chemistry subsection). Plant production and biomass are also controlled by temperature (overgrazing hypothesis). Climate change is thus likely to modify the population dynamics patterns and roles of key herbivores such as lemmings and moths because the dynamics of herbivore-plant interactions are likely to change. As early as 1924, Charles Elton pointed out the potentially decisive role of climate in determining the generation of cycles in northern animal populations[116].

Mathematical modeling shows that specialist resident predators such as small mustelids and the Arctic fox can also impose prey population cycles due to sufficiently strong numerical and adequate functional responses[117]. Moreover, nomadic specialists such as birds of prey can dampen lemming cycles and decrease the degree of regional asynchrony if their predation rates are sufficiently high[118]. The impacts of bird predators have a strong seasonal component since most migrate south for the winter[119]. Reliable estimates of predation rates on cyclic lemming populations are rare. Indirect estimates based on the energy requirements of predators at Point Barrow, Alaska, indicated that avian predators could account for 88% of the early summer mortality, but it was concluded that neither this nor winter predation by weasels could stop lemming population growth under otherwise favorable winter conditions[120]. In the Karup Valley, Greenland, the combined impact of different predators both limited population growth and caused population crashes in collared lemmings[121]. In a declining lemming population in an alpine area in Norway, almost 50% predation was demonstrated by following the fates of radio-tagged individuals[122]. Using the same methodology, Reid et al.[123], Wilson D. et al.[124], and Gilg[125] showed that predation was the predominant mortality factor in populations of collared lemmings at various locations in northern Canada and eastern Greenland.

Predator–prey interactions

The dynamics and assemblages of vertebrate predators in arctic tundra are almost entirely based on lemmings and other small rodent species (Microtus spp. and Clethrionomys spp.[126]). Birds of prey such as snowy owls, short-eared owls (Asio flammeus), jaegers (skuas – Stercorarius spp.), and rough-legged buzzards (Buteo lagopus) are lemming and vole specialists that are only able to breed at peak lemming densities and which aggregate in areas with high lemming densities. Since lemming cycles are not synchronized over large distances[127], the highly mobile avian predators can track lemming population peaks in space. Mammal lemming and vole specialists in the Arctic, such as the least weasel (Mustela nivalis) and the ermine, are less mobile than birds but both have high pregnancy rates and produce large litters in lemming peak years[128]. In lemming low years, weasel and ermine reproduction frequently fails and mortality rates increase[129]. In coastal and inland tundra habitats where bird colonies are lacking, the Arctic fox also exhibits the population dynamics typical of a lemming specialist[130]. The lemming cycles also impose cyclic dynamics in other animals such as geese and waders because they serve as alternative prey for predators in lemming crash years[131]. Recently observed increased predation pressure on water birds in various arctic regions might reflect a change of the lemming cycle in response to climate change, with secondary effects on predators and water birds as an alternative prey[132]. Thus, a large part of the tundra vertebrate community cycle is in a rhythm dictated by the lemming populations[133].

This rhythm is likely to be disrupted by projected future variations in snow properties (e.g., snow-season length, snow density, and snow-cover thickness[134]). For small mammals living in the subnivean space, snow provides insulation from low temperatures as well as protection from most predators such as foxes and raptors[135] and increases in snow are likely to be beneficial. The effect on large mammal prey species (ungulates) is likely to be the opposite, as deeper snow makes reindeer/caribou and moose (Alces alces) more vulnerable to predators such as wolves[136], but more extensive snow patches provide relief from insect pests (section 7.3.3.2). If climate change results in more frequent freeze-thaw events leading to a more shallow and icy snowpack (section 6.4.4), this is likely to expose small mammals to predators, disrupt population increases, and thereby prevent cyclic peak abundances of lemmings and voles. For nomadic predators whose life-history tactic is based on asynchronous lemming populations at the continental scale, an increased frequency of large-scale climatic anomalies that induces continent-wide synchrony (the “Moran effect”[137]) is very likely to have devastating effects.

Long-term monitoring (>50 years) of small rodents near the treeline at Kilpisjärvi in subarctic Finland has shown a pronounced shift in small rodent community structure and dynamics since the early 1990s[138] (see Fig. 7.13). In particular, the previously numerically dominant and cyclically fluctuating grey-sided vole has become both less abundant and less variable in abundance. The Norway lemming and Microtus voles also have lower peak abundances, and the small rodent community is currently dominated by the relatively more stable red-backed vole (Clethrionomys rutilus). Similar changes took place in the mid-1980s in the northern taiga[139] and are still prevailing. For predators that specialize in feeding on small rodents, the lack of cyclic peak abundances of small rodents, especially in the spring[140], is likely to have detrimental consequences, as they need to breed successfully at least every three to four years to sustain viable populations. At Kilpisjärvi, the least weasel has become rare. Moreover, the severe decline of the Arctic fox and the snowy owl in Fennoscandia, both of which prey on Microtus voles and lemmings in mountain and tundra habitats, may be due to lower peak abundances of small mammal prey species in their habitats[141]. In Alaska, a similar decrease in lemming cyclicity occurred in the 1970s[142].

Large predators such as wolves, wolverines, and bears are less numerous in the tundra than in the boreal forest[143]. Consequently, predation impacts on tundra ungulates are usually low. While 79% of the production in small herbivores (voles, lemmings, ptarmigan, and Arctic hares) was consumed by predators averaged across a number of sites in arctic Canada, the corresponding number was only 9% for large herbivores (caribou and muskox[144]).

Insect pests, parasites, and pathogens

Plants

Disease in plants is likely to increase in those parts of species distribution ranges where a mismatch between the rate of relocation of the species and the northward/upward shift of climatic zones results in populations remaining in supra-optimal temperature conditions. Under these conditions, species can experience thermal injury (particularly plants of wet and shady habitats[145]), drought, and other stresses that make plants more susceptible to disease.

Very little is known about the incidence and impacts of plant diseases in arctic ecosystems. However, recent work has shown that a fungal pathogen (Exobasidium) of Cassiope tetragona and Andromeda polifolia reduces host plant growth, reproductive investment, and survival[146]. As the incidence of disease increases with an increase in temperature downward along an altitudinal gradient, climate warming is likely to increase the incidence of at least this naturally occurring disease in the Arctic. The incidence of new diseases from increasing mobility of pathogens with a southern distribution is a possibility.

Animals

Ultraviolet-B radiation can reduce the impact of viral and fungal pathogens on insects. The nuclear polyhedrosis virus is a major cause of death of the defoliating autumnal moth. However, this virus is killed by UV-B radiation[147]. Species and strains of the fungus Metarhizium are important agents of insect disease, but some, particularly high-latitude strains, are sensitive to UV-B radiation[148].

Parasitism is perhaps the most successful form of life, but until recently has been underestimated, especially in the Arctic[149]. Parasitism in the Arctic has been poorly studied with respect to both taxonomy and biodiversity as well as the ecological impact parasites may have on animal species and communities.

Recent research on the evolution and phylogeography of typical arctic animals like lemmings has revealed how greatly the alternating glacial and interglacial periods have influenced their distribution and genetic diversity[150]. The impact seems to be at least as profound on the helminth parasites of arctic rodents[151]. Such impacts of past climatic fluctuations can be used to project some possible consequences of the present warming. If the arctic host populations become fragmented due to the northward expansion of southern biogeographic elements, extinction of parasites in small host populations and/or cryptic speciation (isolation events seen in parasites, often only by using molecular methods, that are not evident in host populations) in refugia are likely to follow. Phylogeographic structure (often cryptic speciation) can be seen in rodent cestodes in the Arctic even if there is no such structure in the host. This is true also for ruminant parasites.

Phylogenetic studies have shown that host switches have occurred in many clades of rodent cestodes[152]. It seems plausible that host switches have been promoted by climatic events that force host assemblages, earlier separated by geography or habitat, to overlap in their distribution.

Macroparasites, such as intestinal worms, often have complicated life cycles. In the main host, in which the parasite reproduces, parasites are controlled by host immunity. On the other hand, the free-living intermediate stages (eggs and larvae), and those in intermediate hosts, are subject to extrinsic environmental conditions like temperature and humidity. Temperature strongly affects the development rate of parasite larvae. For example, a small increase in temperature has a clear effect on the development of the muskox lungworm Umingmakstrongylus pallikuukensis in its gastropod intermediate hosts[153]. Therefore, a slight increase in temperature and in growing-season length is very likely to profoundly affect the abundance and geographic distribution of potentially harmful parasites such as lungworms. Lungworm infections have become conspicuous in recent years as summer temperatures in the Arctic have increased.

The free-living stages of parasites are prone to desiccation. In addition to temperature effects on their development, the survival and abundance of free-living intermediate stages depend greatly on humidity. In addition, the same factors affect drastically the abundance, survival, and distribution of the intermediate hosts of parasites, like insects, gastropods, and soil mites. Haukisalmi and Henttonen[154] found that precipitation in early summer was the most important factor affecting the prevalence of common nematodes and cestodes in Clethrionomys voles in Finnish Lapland. Temperature and humidity also affect the primary production and development of the free-living stages of abomasal nematodes of reindeer/caribou[155]. Recently, Albon et al.[156] showed that abomasal nematodes affect the dynamics of Svalbard reindeer through fecundity. Consequently, even slight climatic changes are likely to have surprising effects on the large ungulates, and possibly on humans exploiting them, through enhanced parasite development (Chapter 15).

The complicated life cycles of parasites cause intrinsic lags in their capacity to track the changes in the population density of their hosts, and these lags are further retarded by unfavorable arctic conditions. Any climatic factor promoting the development of a parasite, so that it can respond in a density-dependent way to host dynamics, is likely to alter the interaction between parasite and host, and the dynamics of both.

There is considerable uncertainty about the possibilities for invasion of pathogens and parasites into the Arctic as a result of climate warming (but see section 15.4.1.2). However, increased tourism combined with a warmer climate could possibly increase the risk of such invasions.

Climate change is likely to affect the important interaction between parasitic insects and reindeer/caribou. Insect harassment is already a significant factor affecting the condition of reindeer/caribou in the summer (section 7.3.3.2). These insects are likely to become more widespread, abundant, and active during warmer summers while many refuges for reindeer/caribou on glaciers and late snow patches are likely to disappear.

Microbe–plant and microbe–microbivore interactions

Although data on the dynamics and processes in arctic microbial communities and on processes in the soil–microbial–plant interface are accumulating rapidly, it is not yet possible to reach firm conclusions about how the dynamics and processes will change in a changing climate. However, the following can be stated. First, short-term (seasonal) changes in microbial processes may not necessarily have major influences on longer-term (annual to multi-annual) processes. Second, microbes and plants share common nutrient resources, although they may not be limited by the same resource. For example, while nutrient supply rates generally control plant productivity, microbial productivity may be constantly or periodically controlled by the abundance of labile C. Third, the nutrient supply rate to the pool available to plants may not be controlled principally by continuous nutrient mineralization, but rather by pulses of supply and sequestration of nutrients linked to microbial population dynamics and abiotic change, such as freeze–thaw cycles.

Jonasson et al.[157] showed that despite no appreciable effect on the microbial biomass and nutrient mass, warming increased plant productivity. Because plant productivity was limited principally by a low rate of N supply, it appears that the mineralization of litter or soil organic matter, or microbial solubilization of organic N, increased, and that the plants rather than the microbes sequestered the "extra" N in inorganic or organic form. However, microbes increased their nutrient content in cases when the sink strength for nutrients in the plants decreased (e.g., after shading) at the same time as soil inorganic N also increased. This suggests either that plants compete successfully with microbes for nutrients, or that the microbial requirement for nutrients was satisfied, and the microbes absorbed a "surplus" of nutrients, which is likely if they were limited by C rather than nutrients. This does not fully preclude nutrient competition, however, because it is possible that the plants accessed the nutrients from pulse releases from microbes during periods of population dieback. If so, seasonal changes in the frequency of such pulses are of importance for projecting changes in ecosystem function and need further investigation. This is particularly obvious, considering that the microbial N and P content typically exceeds the amounts annually sequestered by plants several-fold, and should constitute an important plant nutrient source[158].

The plant–microbe interaction may also be mutualistic through the mycorrhiza by which the fungal partner supplies nutrients to the plant in exchange for C supplied by the plant. A large proportion of the plant species in shrubby vegetation, common in the Arctic, associate with ecto- or ericaceous mycorrhizal fungi. These mycorrhiza types have enzyme systems able to break down complex organic molecules and thereby supply the plant partner with N[159], the most common production-limiting element for plants. Changes in plant species composition as a consequence of climatic changes are very likely to substantially affect microbial community composition, including that of mycorrhizal fungi. Unfortunately, studies of the effects of projected climate change on mycorrhizal associations in the Arctic are virtually nonexistent. However, a decade of warming of a fellfield led to a substantial increase in willow (Salix) biomass, but few changes in the community of associated ecto-mycorrhizal fungi[160].

The effects of microbivores on the microbial community are yet poorly explored and can only be listed as potentially important for projecting effects of global change. It appears, however, that populations of nematodes increase strongly with warming. Because nematodes are the main predators of fungi and bacteria, it may be that increased biomass production of microbes is masked in a warmer environment because of predation by strongly responding microbivores[161]. If so, the release rate of plant-available nutrients is likely to increase (e.g., [162]), which may explain the enhanced nutrient sequestration by plants in warmer soils (rather than pulse sequestration after microbial dieback).

Summary (7.4.1.5)

Changes in climate and UV radiation levels are very likely to affect three important attributes of ecosystem structure: spatial structure (e.g., canopy structure and habitat), trophic interactions, and community composition in terms of biodiversity. Ecosystem structure varies along a latitudinal gradient from the treeline to the polar deserts of the high Arctic. Along this gradient there is a decreasing complexity of vertical canopy structure and ground cover ranging from the continuous and high canopies (>2 m) of the forest tundra in the south to the low canopies (~5 cm) that occupy less than 5% of the ground surface in the polar deserts. Within each arctic vegetation zone, there are often outliers of more southerly zones. Changes in vegetation distribution in relation to climate warming are likely to occur by local expansion of these intrazonal communities and northward movement of zones. Satellite measurements, aerial photographs, and indigenous knowledge show a recent increase in shrubbiness in parts of the Arctic.

Experimental manipulations of environmental factors projected to change at high latitudes show that some of these factors have strong effects on the structure of arctic ecosystems, but the effects are regionally variable. Nutrient addition has the greatest effect on the productivity, canopy height, and community composition of arctic plant communities. Nutrients also increase biomass turnover, so biomass may or may not respond to nutrient addition. Summer warming of tundra vegetation within the range of projected temperature increases (2 to 4 °C over the next 100 years) has generally led to smaller changes compared with fertilization and always to greater responses compared with irrigation. Plant growth response increased from a climatically relatively mild forest understory through a treeline heath to a cold, high-altitude fellfield. Total nonvascular plant biomass and particularly the biomass of lichens decreased in response to 10 years or more of environmental manipulations at sites in subarctic Sweden and in Alaska. Warming experiments in the high Arctic had a greater effect on the fauna above ground compared with fauna below ground and in the subarctic. Spring freeze–thaw events are important, and will probably cause differential mortality among species, thus altering community composition. In general, arctic invertebrate communities are very likely to respond rapidly to change. In contrast, long-term data on the effects of summer warming (2–4 °C) of ecosystems have not shown appreciable changes in microbial biomass and nutrient stocks. This suggests that a temperature increase alone is unlikely to have any strong impact on microbial carbon and nutrient sequestration. Manipulations simulating enhanced UV-B radiation levels and a doubling of atmospheric CO2 concentration for seven years altered the use of labile carbon substrates by gram-negative bacteria, suggesting a change in community composition. UV-B radiation also affects the structure of fungal communities. So far, no change in plant community structure has been found in the Arctic in response to manipulations of UV-B radiation levels and CO2 concentration.

Trophic interactions of tundra and subarctic forest plant-based food webs are centered on a few dominant animal species, which often have cyclic population fluctuations that lead to extremely high peak abundances in some years. Small herbivorous rodents of the tundra (mainly lemmings) are the main trophic link between plants and carnivores. Small-rodent population cycles with peak densities every three to five years induce strong pulses of disturbance, energy, and nutrient flows, and a host of indirect interactions throughout the food web. Lemming population cycles are crucial for nutrient cycling, structure and diversity of vegetation, and for the viability of a number of predators and parasites that are specialists on rodent prey/hosts. Trophic interactions are likely to be affected by climate change. Ice crusting in winter is likely to render vegetation inaccessible for lemmings, deep snow is likely to render rodent prey less accessible to predators, and increased plant productivity due to warmer summers is likely to dominate food-web dynamics. Long-term monitoring of small rodents at the border of arctic Fennoscandia provides evidence of pronounced shifts in small rodent community structure and dynamics that have resulted in a decline in predators (including Arctic fox, snowy owls, buzzards, and skuas) that specialize in feeding on small rodents.

In subarctic forests, a few insect defoliators such as the autumnal moth that exhibit cyclic peak densities at approximately 10-year intervals are dominant actors in the forest food web. At outbreak densities, insects can devastate large tracts of birch forest and play a crucial role in forest structure and dynamics. Trophic interactions with either the mountain birch host or its insect parasitoids are the most plausible mechanisms generating cyclic outbreaks in Epirrita. Climate is likely to alter the role of Epirrita and other insect pests in the birch forest system in several ways. Warmer winters are likely to increase egg survival and expand the range of the insects into areas outside their present outbreak ranges. However, the distribution range and activity of natural enemies are likely to keep the insect herbivore populations below outbreak densities in some areas.

Climate change is likely to also affect the important interaction between parasitic insects and reindeer/caribou. Insect harassment is already a significant factor affecting the condition of reindeer in the summer. These insects are likely to become more widespread, abundant, and active during warmer summers while refuges for reindeer/caribou on glaciers and late snow patches are likely to disappear. There are large uncertainties about the outcome of the potential spread of new trophic interactants, especially pests and pathogens, into the Arctic.

Disease in plants is likely to increase in those parts of species distribution ranges where a mismatch between the rate of relocation of the species and the northward/upward shift of climatic zones results in populations remaining in supra-optimal temperature conditions. The incidence of new diseases from increasing mobility of pathogens with a southern distribution is a possibility, but increases in UV-B radiation levels could possibly reduce the impact of viral and fungal pathogens.

Microbe–plant interactions can be competitive for nutrients and also mutualistic through mycorrhizal associations. Warming will probably affect both types of relationship, but information is scarce.

Chapter 7: Arctic Tundra and Polar Desert Ecosystems

7.1 Introduction
7.2 Late-Quaternary changes in arctic terrestrial ecosystems, climate, and ultraviolet radiation levels
7.3 Species responses to changes in climate and ultraviolet-B radiation in the Arctic
                7.3.1 Implications of current species distributions for future biotic change
                7.3.2 General characteristics of arctic species and their adaptations in the context of changes in climate and ultraviolet-B radiation levels
                7.3.3 Phenotypic responses of arctic species to changes in climate and ultraviolet-B radiation
                7.3.4 Genetic responses of arctic species to changes in climate and ultraviolet-B radiation levels
                7.3.5 Recent and projected changes in arctic species distributions and potential ranges
7.4 Effects of changes in climate and UV radiation levels on structure and function of arctic ecosystems in the short and long term
                7.4.1 Ecosystem structure
                7.4.2 Ecosystem function
7.5 Effects of climate change on landscape and regional processes and feedbacks to the climate system
7.6 Synthesis: Scenarios of projected changes in the four ACIA regions for 2020, 2050, and 2080
7.7 Uncertainties and recommendations

 

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Committee, I. (2012). Effects of changes in climate and UV radiation levels on structure of arctic ecosystems in the short and long term. Retrieved from http://www.eoearth.org/view/article/152364

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