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European forests belong to an important economic sector that is potentially affected by climate change and changes in atmospheric CO2 concentrations. Forests also have important interactions with global change processes as a result of their sink potential.
Primarily, temperature and the availability of soil moisture limit the natural range of European tree species. Some forests (particularly in the north) also are nutrient-limited. The structure and composition of many forests is further influenced by the natural disturbance regime (e.g., fire, insects, windthrow). Most European forests are managed for one or several purposes, such as timber production, water resources, or recreation. This management has reduced forest area or strongly modified forest structure in most of Europe, and presently existing forests often consist of species that are different from those that would occur naturally.
In northern Europe, boreal forests are dominated by Picea abies and Pinus sylvestris, and these species grow well across most of their current distribution ranges. Under warmer conditions, these species are likely to invade tundra regions (Sykes and Prentice, 1996). In the southern boreal forests, these species are expected to decline because of a concurrent increase of deciduous tree species (Kellomäki and Kolström, 1993). Most climate change scenarios suggest a possible overall displacement of the climatic zone that is suitable for boreal forests by 150–550 km over the next century (Kirschbaum et al., 1996). This shift in climatic conditions would occur more rapidly than most species have ever migrated in the past (20–200 km per century—see Davis, 1981; Birks, 1989). Also questionable is whether soil structural development would be able to follow. Turnover of current tree populations may be enhanced, however, by changes in management practices and changing disturbance regimes, such as increased fire frequency or increased strong winds in late autumn and early spring (Solantie, 1986; Peltola et al., 1999).
Under most recent climate change scenarios, winters are likely to be still cold enough to fulfill the chilling requirements of the main boreal tree species (Myking and Heide, 1995; Leinonen, 1996; Häkkinen et al., 1998), but earlier budburst can be expected. If summer and winter precipitation increase (as indicated by some scenarios), boreal forests would become less susceptible to fire damages, which currently affect about 0.05% of forests yr-1 (Zackrisson and Östlund, 1991).
At present, cold winters in boreal and some temperate regions protect forests from many insects and fungi that are common further south (Straw, 1995). High summer temperatures and associated drought increase the growth of existing insect populations through enhanced physiological activity and turnover of insect populations. Throughout Europe, forests seem to be quite well buffered against new species coming from outside Europe, but the risk exists. A good example of new organisms with large potential to damage trees is Bursaphelenchus xylophilus, which originates in North America. This pine nematode is easily transported in fresh timber, but its success is related to temperature. Low summer temperatures and short growing seasons have effectively limited the success of this species in northern Europe (Tomminen and Nuorteva, 1987), although it frequently has occurred in imported timber. Reductions in these limitations may result in increased damage to trees.
In western and central Europe, the current forest structure and, in part, tree species composition are determined mainly by past land use and management rather than by natural factors (Ellenberg, 1986). Site-specific assessments of the future composition of near-natural forests suggests that conifers (e.g., Picea abies) may be replaced by deciduous species (e.g., Fagus sylvatica) at some sites (e.g., Kräuchi, 1995). Until recently, our capability to assess long-term forest dynamics at the regional scale was quite limited. Lindner et al. (1997) provided the first assessment of regional-scale patterns of forest composition under current and future climates. Their study suggests that future near-natural forests in the state of Brandenburg (east Germany) would be much more uniform, with little of the differentiation across different site types that shape today’s landscape. A temperature increase of 1–3°C would advance budburst of many tree species by several weeks (Murray et al., 1989). Introduction of phenologically suitable ecotypes or new species has been among the main tools to increase forest growth (Lines, 1987). Minimum winter temperatures seem to be critical for the survival of exotic species with insufficient winter frost hardiness—for example, Nothofagus procera in Britain (Cannell, 1985). Therefore, higher winter temperatures could broaden the potential distribution range of such species in Europe.
In southern Europe, most forests consist of sclerophyllous and some deciduous species that are adapted to summer soil water deficit. Climate scenarios indicate reduced water availability in the summer months and associated responses in forests (e.g., Gavilán and Fernández-González, 1997), although the interactions of this effect with enhanced CO2 concentrations is uncertain. Temperature changes may allow expansion of some thermophilous tree species (e.g., Quercus pyrenaica) when water availability is sufficient. In the Pyrenees, a northward and upward movement of Mediterranean ecotypes is likely to occur with warming accompanied by drier conditions.
Forest growth has increased during the past several decades in northern forests (Lakida et al., 1997; Lelyakin et al., 1997; Myneni et al., 1997) and elsewhere in Europe (Spiecker et al., 1996). Climate warming, increasing CO2, increased nitrogen deposition, and changes in management practices are factors that are assumed to be behind the increase. The impacts of temperature and CO2 have been shown in experiments and are extrapolated by model calculations. For example, under an assumed increase of CO2 by 3.5 µmol mol-1 yr-1 and temperature by 0.04°C yr-1 over 100 years, productivity of Pinus sylvestris increased by 5–15% as a result of the temperature elevation, 10–15% as a result of the increased CO2, and 20–30% as a result of combined temperature and CO2 (Kellomäki and Väisänen, 1997). In northern Europe, the effects of precipitation changes are likely to be much less important than the effects of temperature changes (Kellomäki and Väisänen, 1996; Talkkari and Hypén, 1996). Based on model computations that assume a seasonally uniform temperature increase, Proe et al. (1996) have suggested that growth of Picea sitchensis in Scotland could increase by 2.8 m3 ha–1 yr–1 for each 1°C rise in temperature.
In Russian boreal forests, some studies predict large shifts in distribution (up to 19% area reduction) and productivity (e.g., Kondrashova and Kobak, 1996; Krankina et al., 1997; Izrael, 1997; Raptsun, 1997). It could be concluded that climate change and CO2 increase would be favorable for northern forests (e.g., as a result of increased regeneration capacity). Indeed, some studies suggest a significant increase in productivity of forests in higher latitudes. The largest changes are expected for forest tundra and the northern taiga, reaching 12–15% additional growth per 1°C warming (e.g., Karaban et al., 1993; Shvidenko et al., 1996; Lelyakin et al., 1997). The same studies estimate the impact in more southern forest zones (southern taiga, mixed and deciduous forests) to be less: 3–8% per 1°C.
In central and southern Europe, limited moisture resulting from increasing temperature and (possibly) reduced summer rainfall may generate productivity declines regionally, but this cannot be predicted because of uncertain rainfall scenarios. In addition, CO2 enrichment is likely to increase water-use efficiency (WUE), which makes growth less drought-sensitive. Forest growth conditions in the southern parts of eastern Europe (Russia, Ukraine, Moldova) are likely to decline as a result of increased drought, specifically in the steppe. Secondary problems could arise in the protective shelterbelts in the south of the forest-steppe zone that now covers about 3 Mha.
In the Mediterranean region, elevation of summer temperature and reduction of precipitation may further increase fire risk. Colacino and Conte (1993a,b) examined the pattern of forest fires in the Mediterranean region in connection with the number of heat waves. An increase of 70% in the number of heat waves was recorded in the period 1980–1985 with respect to the period 1970–1975, and a similar increase was recorded in the extent of forest burned. In temperate eastern Europe, forest fire increase is less likely, but very dry and warm years could occur more frequently and promote pest and pathogen development. Large areas of pine forests in Ukraine, Belarus, and central regions of Russia might face some increased risk.
Increased forest fire risk is a crucial factor in the survival of boreal forests in Russia. Most dangerous are large forest fires, which occur during extremely dry and warm years. Such climatic conditions occur periodically (every 15–20 years) in parts of the Russian boreal zone. Currently these large fires account for about 1–2% of the total number of forest fires, but burned areas reach 70–80% and losses are as much as 90% of the total values. Most climate scenarios indicate that the probability of large fires will increase.
Estimates of the possible influence of climate change on insect infestation are uncertain because of complex interactions between forests, insects, and climate. The probability of outbreaks of pests such as Dendrolimus sibirica or Limantria dispar is expected to increase, especially in monocultures. Short-period warming also could promote infestation with new pest species that presently do not occur in the boreal zone. Increases of climate aridity would promote occurrence of some diseases (e.g., root and stem fungi decays).
Permanent grassland and heathland occupy a large proportion of the European agricultural area. The type of grassland varies greatly, from grass and shrub steppes in the Mediterranean region to moist heathland in western Europe. The annual cycle of many temperate grasses is limited by low temperature during the winter and spring and by water stress during the summer. Climate change can affect the productivity and composition of grasslands in two ways: directly through the effects of CO2, or indirectly through changes in temperature and rainfall. Different species will differ in their responses to CO2 and climate change, resulting in alterations in community structure (Jones and Jongen, 1996). Legumes, which are frequent in these communities, may benefit more from a CO2 increase than nonfixing species (Schenk et al., 1995).
Intensively managed and nutrient-rich grasslands will respond positively to the increase in CO2 concentration and to rising temperature, as long as water resources are sufficient (Thornley and Cannell, 1997). The direct effect of doubling CO2 concentration by itself may cause a 20–30% increase in productivity in nutrient-rich grasslands (Jones et al., 1996; Cannell and Thornley, 1998). The importance of water management (including drainage) may be even more important, however, under changed climatic conditions in northern Europe (Armstrong and Castle, 1992). This positive effect of increased CO2 on biomass production and WUE can be offset by climate change, depending on local climate and soil conditions (Topp and Doyle, 1996a; Riedo et al., 1999). These effects also will determine the spatial distribution of agricultural grassland. An analysis by Rounsevell et al. (1996a) showed that grassland production in England and Wales is resilient to small perturbations in temperature and precipitation, but larger temperature increases may cause drought stress and reduced suitability for grassland production.
There is a greater controversy regarding the response of nitrogen-poor and species-rich grassland communities. Experimental studies in such grasslands have shown little response or even a reduction in production with CO2 enrichment (Körner, 1996). On the other hand, simulation studies have shown that this could be just a transient response and that the long-term response of nitrogen-poor grassland ecosystems may be relatively larger than that of nitrogen-rich systems (Cannell and Thornley, 1998). This effect is caused by a reduction in nutrient losses and an increase in nitrogen fixation at elevated CO2.
Because of its impacts on primary productivity and community structure, the long-term effect of elevated CO2 on grasslands is an additional carbon sink. By contrast, increasing temperatures alone are likely to turn grasslands into a carbon source because soil respiration would be accelerated more than NPP. The net effect of current scenarios for CO2 and temperature is likely to be a small carbon sink in European grasslands (Thornley and Cannell, 1997).
Arid and semi-arid environments (e.g., certain steppe-like habitats), which are well represented in the Mediterranean area, are crucial for the preservation of rich species diversity in this region. These regions seem to be the only places within Europe that certain insect species, such as Lepidoptera, can inhabit because of the abundant availability of their foodplants. Furthermore, the lack of winter climatic stress makes arid lands quite suitable as wintering grounds for birds. Overgrazing, fire, urbanization, and changes in land use can be considered the main threats to these regions. The potential distribution of these semi-arid environments may increase under drier and warmer climatic conditions, leading to landscape fragmentation at the local scale and consequent local extinctions (del Barrio and Moreno, 2000).
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