In intensive modern farms, grassland areas are ploughed and reseeded (usually with L. perenne in Europe) on a 5-10 year cycle, and their soils in consequence bear more similarity to arable fields than permanent grasslands. Additionally, the past 50 years have seen the widespread use of synthetic fertilizers to improve grassland productivity. Thus, disturbance and eutrophication have led to the demise of most macrofungal fruiting in these habitats, although it has yet to be demonstrated that the mycelia are also absent. Losses of fungal diversity generally mirror declines in plant and invertebrate diversity, and in the case of these better studied groups changes in grassland management can also lead to loss of diversity (Rook and Tallowin, 2003). Shifts from haymaking to silage production or from cattle and sheep to sheep only grazing have also altered patterns of abundance of higher plants and insects. For soil dwelling fungi such changes might be anticipated to have a lesser effect, although changes in patterns of root death and photosynthate translocation will affect the nutrition of soil microbes (Turner et al., 1993). Conversely, microclimatic conditions for basidiocarp formation are altered by sward height variation, and macrofungal fruiting in rank grassland is much reduced compared to adjacent grazed areas (Griffith et al., 2006). However, as appears to be the case in many prairie grasslands where vegetation is much longer than the 3-15 cm sward height typical of European pastures, the health of the underlying mycelium may be little affected by above-ground vegetation height. Mown grasslands, especially historic lawns, represent important refugia for grassland fungi. While these habitats are often spared fertilizer application, the failure to remove clippings can cause eutrophication and loss of diversity, especially in areas of higher nitrogen deposition (i.e. most of Europe).
Fungi are seldom considered in issues of land use but there is a growing body of evidence that sites with diverse fungal communities do not necessarily host diverse plant communities. This is consistent with the idea that soil nutrient conditions are far more important than sward management. While many sites with diverse grassland fungal communities receive some legal protection (SSSI, etc.), fungal diversity is seldom mentioned in the notification statements (Chapter 8). Since site visits by nature conservation staff generally occur in the summer, there is little information about macrofungal diversity. Recent UK legislation (EIA (Agriculture) Regulations, 2001) controls change of use of agricultural land (e.g. ploughing of pasture), but since biodiversity assessments are generally conducted in the summer, low plant diversity can lead to destruction of valuable fungal sites.
With prospective changes in agricultural support, the re-establishment of semi-natural habitats is gaining attention. Dispersal of fungi is not perceived to be a significant factor limiting recolonization but reductions in soil nutrient status, coupled with a latent period between colony establishment and fruiting, can lead to delays in reappearance. Our work at various restoration sites, consistent with other studies (Lange, 1991), suggests that fruiting of the more common member of the more prized grassland taxa (Hygrocybe, Entoloma spp., etc.) may occur within a decade of cessation of nutrient addition. We note, however, that some of the most diverse sites for grassland fungi were subject to significant disturbance in recent centuries (e.g. post-industrial sites such as iron works, canal/reservoir embankments).
Since grasslands contain 12% of the world's SOM (33 kg m-2 in temperate grasslands; Conant et al. , 2001), factors that affect the activity of saprotrophic basidiomycetes in grasslands can impact on atmospheric CO2 levels and consequent climate change (Freibauer et al., 2004). Global warming and changing rainfall patterns combined with changes in agricultural subsidies are likely to lead to changes in climax vegetation types (Raich and Tufekcioglu, 2000), with scrub invasion and afforestation of grasslands generally resulting in increased soil C pools (Smith and Johnson, 2004). However, there are examples where the opposite has occurred. Planting of exotic pines in Andean paramo grasslands has caused loss of SOM, apparently due to the saprotrophic activity (soil C mineralization) of the usually ectomycorrhizal symbiont, Suillus luteus (Chapela et al., 2001). There is already evidence of changes in phenology of basidiocarp production in UK grasslands since the 1970s with grassland species showing contrasting patterns to woodland saprotrophs (Gange et al., 2007; Chapter 5).
Many parts of the world now experience high levels of aerial deposition of 'fixed' N (from intensive agriculture and vehicle emissions), a consequence of anthropogenic fixation of nitrogen (Haber-Bosch process), which has increased 10-fold since pre-industrial times (Fowler et al., 2004), and now exceeds natural fixation by bacteria (Galloway et al., 1995). Even modest nitrogen deposition (5-10 kg N ha-1 year-1) reduces diversity of ectomycorrhizal agarics in boreal forests (Lilleskov et al., 2002), probably due to alteration of soil nitrogen cycles (especially mobilization of organic nitrogen), which are very likely also to affect saprotrophic species. Although critical N loads for grasslands are higher than for woodlands, loss of plant diversity in UK grasslands (receiving 6-50 kg N ha-1 year-1) is correlated with nitrogen deposition (Stevens et al., 2004; Chapter 17). Projected N deposition in 2050 for the world's 34 biodiversity hotspots suggests that half of these, including grassland systems such as the Brazilian cerrado, will be subjected to >15 kg N ha-1 year-1 (Phoenix et al., 2006).
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