Biological Conservation 143 (2010) 2395–2404
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Beyond the forest edge: Ecology, diversity and conservation of the grassy biomes William J. Bond a,*, Catherine L. Parr b,1 a b
Botany Department, University of Cape Town, Private Bag, Rondebosch 7701, South Africa Environmental Change Institute, School of Geography and the Environment, University of Oxford, South Parks Road, Oxford OX1 3QY, UK
a r t i c l e
i n f o
Article history: Received 31 July 2009 Received in revised form 6 December 2009 Accepted 13 December 2009 Available online 20 January 2010 Keywords: Savanna Tropical grassland C4 grass Cerrado Campos Fire regime Forest–grassland mosaic
a b s t r a c t Forests and grassy vegetation (savannas and grasslands) are alternative ecosystem states in many tropical landscapes. Relative to forests the grassy ecosystems are poorly known and poorly conserved, partly because they were thought to be products of forest clearance. However many grasslands have proved to be ancient. Commensurate with their antiquity, grassy biomes have distinct suites of plant and animal species that contribute a large fraction of the diversity of forest–grassland mosaics. Grasslands differ strikingly from forests in their ecology and in the nature of threats to their future. Here we highlight the high biodiversity value of grassy biomes and, in contrast to tropical forests, we illustrate the importance of ﬁre in maintaining these systems. We discuss the major threats to, and consequences for, biodiversity in these regions including land clearance and elevated CO2-driven forest expansion. Finally we focus on the difﬁculties of grassland restoration. A new approach to understanding and conserving grassy ecosystems, free from cultural prejudices of the past, is long overdue. Ó 2009 Elsevier Ltd. All rights reserved.
1. Introduction Tropical forests rarely dwindle away to woodlands, scattered trees, and then grasslands along a gradient of declining rainfall. Instead forests patches change, usually abruptly, to grassy biomes. Grasslands and savannas form landscape mosaics with more or less forest cover over vast areas of the tropics (Fig. 1); these grassy biomes can be considered the mirror image of forests. Where forest plants grow in deep shade, grassland plants are shade avoiders, where fallen leaves decompose rapidly in the forest ﬂoor, grassland plants decompose slowly with litter often consumed by ﬁres, where ﬁre destroys the structure of the forests, ﬁre is essential for maintaining large areas of tropical savannas and grasslands. Both systems are threatened by changes in the ﬁre regime but for opposite reasons. Where forests are considered remnants of pristine landscapes, the grassy biomes have long been viewed as anthropogenic artefacts, ‘degraded’ lands, ‘secondary successional’ stages, neglected by scientists and of no interest to conservation (see for example Perrier de la Bâthie, 1936; Dalfelt et al., 1996; Banerjee, 1995). The lack of interest in tropical grassy biomes is highlighted by Web of Science searches (years: 2004–2009; databases: SCI-EXPANDED, CPCI-S) using the search terms ‘biodiversity’ and * Corresponding author. Tel.: +27 21 650 2439; fax: +27 21 650 4041. E-mail addresses: [email protected]
(W.J. Bond), [email protected]
(C.L. Parr). 1 Tel.: +44 1865 285537; fax: +44 1865 275885. 0006-3207/$ - see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.biocon.2009.12.012
‘tropical’ with either ‘forest’, ‘grassland’ or ‘savanna’. The search revealed 1343 papers on tropical forests, and only 61 and 103 papers on grassland and savanna systems respectively. Using ‘conservation’ instead of ‘biodiversity’ as a search term, conﬁrmed this ﬁnding with more than eight times as many papers on tropical forests than grassy biomes. Consequently, where loss of tropical forests and their diversity is of wide public concern, loss of grassy vegetation to croplands and plantations has proceeded with little opposition. Indeed afforestation of grasslands is increasingly promoted as a public good – a means of combating climate change through carbon sequestration. The curious cultural hostility to tropical grasslands was prevalent for much of the last century and remains widespread today. However there is also growing recognition of the remarkable diversity of tropical grassy systems and belated attempts are being made to conserve threatened remnants. Where tropical forests face an uncertain future from global warming and drying, tropical grasslands face an uncertain future from increasing carbon dioxide in the atmosphere. The dominant grasses evolved under low atmospheric CO2. In the next 20 or 30 years, CO2 concentrations will rise above the levels at which tropical grasses ﬁrst appeared on earth and below which they have a photosynthetic advantage over their temperate competitors. The consequences for the ecology and conservation of these vast grassy ecosystems have hardly been addressed. This contribution illustrates the high biodiversity value of grassy biomes, highlights the importance of ﬁre in shaping these systems, discusses major threats, and ﬁnally, emphasises the
W.J. Bond, C.L. Parr / Biological Conservation 143 (2010) 2395–2404
Fig. 1. (A) The savanna–forest mosaic landscape and, (B) an abrupt savanna–forest boundary in Lopé National Park, Gabon. Photo credit: D. Maniatis.
difﬁculties of restoring of these habitats (Table 1). In doing so, we invite forest ecologists and conservationists to cross the great divide, step into the sunshine, and consider the grassy world. 2. Grasslands: successional stage or alternative ecosystem state? Grassy vegetation has long been viewed as a successional stage seral to forest (e.g. see Thomas and Palmer, 2007). The assumption is that open habitats suitable for grasslands are created by natural or human disturbance. ‘Natural’ grasslands would only occur where the climate or soils exclude forest trees. Successional grasslands would revert to forests unless actively maintained by
ﬁre or herbivory. The classic example is the Central Grasslands of North America. According to Axelrod (1985), these are of post-glacial origin, do not pre-date human settlement, and were maintained by anthropogenic burning. The recent origin of these grasslands is indicated by ‘the occurrence of most of its species in forest and woodlands, presence of few endemic plants . . . insects, . . . or birds . . ., the relict occurrence of a variety of tree species throughout the region, and the current invasion of woody plants into the grassland’ (Anderson, 2006). Many of the vertebrates of the Central Grasslands are thought not to have evolved in the grasslands but in openings in the eastern forests, the south-western deserts or in the ’natural’ mountain meadows (Anderson, 2006). The grassy biomes of the tropics and sub-tropics share many of the same grass taxa as the North American prairies but have a startlingly different Pleistocene history. Far from being a post-glacial vegetation cobbled together after the ice retreated, C4 grasslands and savannas were even more extensive in Africa at the last glacial maximum (Dupont et al., 2000) and covered extensive areas of South America (reviewed by Furley and Metcalfe, 2007). Forests are frequently the recent invaders of grasslands rather than the reverse (e.g. Delegue et al., 2001; Burbridge et al., 2004). The grassy biomes are a very old land cover and have persisted in some landscapes for tens of thousands of years (Mayle et al., 2007). Their biota has been unevenly studied but includes, in the Cerrado biome of Brazil, one of the world’s biodiversity hotspots with a ﬂora and fauna unique to this open vegetation (Furley, 1999; Ratter et al., 1997). Successional concepts would seem singularly inappropriate for ‘early successional’ vegetation that has persisted for millennia. The concept of alternative ecosystem states seems a far more appropriate framework for analysing grass/forest relationships in the tropics and sub-tropics. Here more than one stable state is possible for a given set of environmental conditions (see Scheffer and Carpenter, 2003; Warman and Moles, 2009). Each state is persistent and controlled by a different set of processes from the alternate state. Rapid shifts (‘regime shifts’) can occur between states (Scheffer and Carpenter, 2003). The idea is that positive feedbacks promote the necessary environmental conditions to maintain one ecosystem state but that these conditions are hostile to the alternative ecosystem state. The result is that alternative vegetation states are characterised by sharp boundaries often producing land-
Table 1 Key questions for conservation of tropical (C4 dominated) grassy biomes, especially in forest/grassland mosaics. The comments column provides summary answers from studies discussed in this paper. The relevant section in this manuscript is given for further reference. USOs are underground storage organs. Key questions
How old are the grasslands?
Often considered to be secondary vegetation produced by forest clearing but carbon isotope and other studies show many are ancient and were even more extensive in the late Pleistocene Diversity is still poorly documented in many regions Grassy biomes include global biodiversity hotspots rich in endemic species Grassland communities are often very rich in non-grass species Grasslands typically have lower alpha diversity of vertebrates while invertebrates may be richer than forests Many vertebrates and invertebrates are endemic to open (non-forest) habitats Plant studies report one third to one half of the diversity is in non-forested habitats
(a) Woody plants typically shade-intolerant, often with USOs, ﬁre tolerant (b) Grasses are shade-intolerant, highly ﬂammable, ﬁre tolerant (c) Forbs are shade-intolerant, often with USOs, ﬁre tolerant (a) Land clearing is the major threat and usually more extensive than in forests (b) Afforestation is common (a) Natural areas: biome switches to closed forest, often from ﬁre suppression (b) Natural areas: alien invasives, especially where ﬁre has been suppressed (a) increasing CO2 favours woody plants promoting forest invasion (b) Increasing CO2 may favour C3 grasses over C4 grasses with poorly known consequences (c) Increased ﬁre activity from extreme weather may favour grass expansion at expense of forests While grasses can colonise abandoned land quickly, restoration of forb and other plant diversity typical of primary grasslands may be very difﬁcult and very slow.
How diverse are grasslands? The plants The animals
What do grasslands contribute to the biodiversity of forest–grassland mosaics? How do grassland plants differ in functional traits from forests? What are the major threats to grassland conservation?
How may global change inﬂuence grasslands?
Can grassland loss be reversed? How effective is grassland restoration?
4 4.1 4.2
6.1 6.2 6.3
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scape-scale mosaics since each state is relatively uninvasible by the other. When boundaries do shift, the shifts are rapid relative to the periods of stability. For a lucid account of the theory in the context of (Australian) tropical forest vs. ﬂammable savannas, including its relevance for conservation and management, see Warman and Moles (2009). From a conservation perspective, it is important to distinguish between successional grasslands, such as those that form after felling forests, from ancient grasslands that have persisted as alternative ecosystem states in forest/grassland mosaics for millennia. The latter would be expected to be rich in species endemic to open grassy habitats (Bond et al., 2008) though, in most parts of the world, the biodiversity of grassy biomes is still poorly known. 3. Geographic scope of tropical grassy biomes Tropical forests most often abut onto grassy ecosystems dominated by C4 grasses. The ﬁrst C4 grasses appeared some 30 million years ago coinciding with declining atmospheric CO2. The C4 grassy biomes began to spread rapidly from their equatorial origin only 8–9 Million years ago spreading south and north during the Pliocene and Pleistocene (Cerling et al., 1997). Thus, they are very recent biomes in earth history, but ancient ones in human history. Contemporary C4 grassy biomes occupy about 20% of the vegetated land surface in a latitudinal band 30°N and S of the equator, being particularly prominent in Africa, northern Australia and South America (Sage, 2004; Ehleringer, 2005). The C4 grassy biomes include savannas and grasslands. Savannas have a more or less continuous grass cover but a discontinuous tree cover, while grasslands lack trees but may have scattered shrubs. Grasslands are clearly distinct from forests but there has been considerable confusion over the distinction between forests and savannas. We take the view that savannas, regardless of tree densities, are characterised by a shade-intolerant understorey of which C4 grasses are the most prominent component. Forests are closed communities casting too much shade to support shade-intolerant understorey plants. One important implication is that ﬁres are frequent and a natural consequence of the grass understorey in savannas (Hennenberg et al., 2006). In contrast, ﬁres are typically very destructive killing mature and juvenile trees in forests. By these criteria, ‘forests’ with a C4 grassy understorey, that burn frequently (such as the dry deciduous ‘forests’ of the Western Ghats; Kodandapani et al., 2004) should be classiﬁed as savannas. Grasslands often occur in montane regions or upland plateaux at higher elevations which may have cooler winters than savannas (we do not consider alpine regions dominated by C3 grasses). Grasslands also occur in seasonally waterlogged lowland areas. Though increased grassy cover is often assumed to be associated with aridity, especially by paleobiologists, the C4 grassy biomes occur across a very broad rainfall range from 200 mm MAP to 3000 mm MAP. Forest patches are common in higher rainfall savanna and grassland landscapes and savanna patches occur in predominantly forested landscapes (Bond, 2008; Sarmiento, 1992; Scholes and Archer, 1997). Here we focus on the higher rainfall grasslands and savannas (P900 mm MAP) which most often border tropical forests. 4. Diversity of grassy ecosystems 4.1. Flora The diversity of grassy ecosystems is still poorly known for many regions. While the woody ﬂora is generally well known, grasses are difﬁcult to identify and the non-grass herbaceous elements (forbs) are both difﬁcult to collect and to identify in rapid
biodiversity surveys. Many forbs have ﬁre-stimulated ﬂowering and are best censused and identiﬁed in the ﬁrst post-burn ﬂowering season. Unfortunately grasses cannot be easily identiﬁed at this stage. Compromises have to be made over the timing of grass layer surveys so that snapshot studies will typically underestimate the diversity of the herbaceous layer. Covering about 2 million km2, the Cerrado biome of Brazil is the best known grassy biota. Ranked twelfth on a list of global ‘hot spot’ areas, ecoregions that contain high levels of plant endemism and are under threat (Mittermeier et al., 1998), it has a notably diverse ﬂora (about 6000 plant species, Ratter et al., 1997), with a particularly high endemism of native angiosperms (Ratter et al., 1997; Myers et al., 2000; Silva and Bates, 2002). The ‘Campo’ (grassland) biome of southern Brazil is also remarkably species rich with an estimated 3000–4000 plant species in an area of 137,000 km2 (Overbeck et al., 2007). This is comparable to the highveld grassland biome of South Africa with about 3800 species in 112,000 km2 containing centres of diversity for many speciose genera (Cowling et al., 1989; Cowling and Hilton-Taylor, 1994). At more local scales, the richness of high rainfall grasslands is especially notable. In the grasslands of southern Brazil, Overbeck et al. (2007) reported 450 plant species in an area of 220 ha ‘placing these grasslands as among the most species-rich grassland communities in the world’. Sankaran (2009), working in the savanna grasslands of the Western Ghats in India, another of the world’s biodiversity hotspots (Myers et al., 2000), recorded 278 plant species in 40 10 10 m plots, distributed in 14 distinct assemblages. More than 90% (256 species) of the plant species contributed less than 5% to grassland cover with most species being highly restricted in their distribution. His study included higher altitude ‘shola’ grasslands which make up a grassland-forest mosaic, and include many endemic plant species (e.g. many species of Strobilanthes). Stott (1990), reporting on the high rainfall savanna woodlands of Thailand, also noted the high diversity of plant species in the grass layer which included many geophytes. The overstory in these savannas is made up of dipterocarps but with a distinct species composition different from adjacent tropical dipterocarp forests. In southern Africa, high levels of plant diversity, especially of the forb layer, are characteristic of the high rainfall savannas and grasslands (Hoare, 2003; Uys et al., 2004; Uys, 2006; Parr, Gray, Bond, unpublished), including the coastal grasslands on infertile dune sands of the Indian Ocean coastal belt (Zaloumis and Bond, in press). Uys (2006) found 158 forb species distributed among 43 families in just ﬁve 0.1 ha plots in a mesic savanna (Hluhluwe). Family diversity of forbs in South African grasslands and savannas is particularly striking with Uys (2006) reporting a mean of 26 families per 0.1 ha plot. This is two to three times richer than for similar sized plots from the Cape fynbos, famous for its high diversity. In the grasslands bordering coastal dune forests of the Indian Ocean coastal belt, Zaloumis and Bond (in press) found 305 plant species, 201 of which were forbs, crammed into 64 sampling plots of radius 5 m. 4.1.1. Grassy biomes vs. forests How distinct are the ﬂoras of grassy vegetation and adjacent forests? Are savannas merely a collection of scrub forest and ruderal species as implied by the notion that grasslands are seral to forest? Or do they have their own distinct species restricted to the grassy ecosystems? Studies of the tree ﬂora in South America, Africa and Australia typically ﬁnd distinct savanna and forest species with little overlap (e.g. Felﬁli and Silva, 1992; Ratter, 1992; Adejuwon and Adesina, 1992; Nangendo et al., 2002). Perez-Garcia and Meave (2006) analysed the complete vascular ﬂora of a tropical dry forest–savanna mosaic in a 90 km2 area of southern Mexico. Only 5% of the 600 species, 10.6% of the genera and less than half of
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the 94 families occurred in both forest and savanna, (Table 2). The savanna ﬂora differed in growth form mix from the forest with most of the diversity contained in forbs and graminoids whereas forest diversity is highest in trees and climbers. In this landscape, both the forest and savannas contribute to the rich diversity of the area. Loss of the savannas would result in the loss of nearly a quarter of the species (23.8%). A similar pattern has been found in a South African dry forest/ savanna mosaic in the hilly landscapes of northern Hluhluwe (Gray, Bond, Parr, unpublished). The forests are expanding into the savannas in this landscape having increased from 10% of the area in the 1930s to 60% of the area in 2004 (Wigley et al., in press). Table 3 shows species richness of woody plants, grasses and forbs in the savanna vs. the forest. The forest patches included areas where forest has invaded savannas only in the last few decades. More species occur in the savanna than the forest in this area. Loss of the savanna habitat would result in the loss of nearly half the ﬂora (41% of the species), with the biggest losses in the abundant forb species that are restricted to open grassy habitats. 4.2. Faunal diversity and speciﬁcity Although little attention has been paid to animals in grassy biomes abutting tropical forests, as with plants, available information indicates that faunal assemblages in grass-dominated systems are diverse and distinct. The fauna of grassy biomes is, most often, not simply a depauperate sub-set of the forest community: rather on leaving the forest one is plunged into an entirely new community. The high diversity of grassland vegetation of the Cerrado of Brazil is mirrored in the high diversity and endemism of a range of faunal groups including birds (Da Silva, 1997; Piratelli and Blake, 2006; Cavalcanti et al., 1999), small mammals (Mares et al., 1986; Lacher and Alho, 2001), herpetofauna (Vitt and Caldwell, 1993) and insects (ants, Vasconcelos et al., 2008; Drosophila, da Mata et al., 2008). The importance of conserving open grassy vegetation is particularly highlighted for birds; for example more than a quarter of species known to breed in the Cerrado region do so only in open vegetation (Da Silva, 1997), and several endemic species at risk are associated primarily with open grassland (Cavalcanti et al., 1999). High habitat speciﬁcity of dung beetle assemblages has also been noted across a savanna/forest ecotone in Bolivia (Spector and Ayzama, 2003). Elsewhere too distinct faunal assemblages can be found between forest and grass-dominated systems. In Hluhluwe in South Africa, there is a large turnover in insect community composition
Table 2 Plant taxon richness and overall similarity index (SI, Sørenson’s Index) between tropical dry forest (TDF) and savanna (SAV) of Nizanda, Mexico. Area is 90 km2. Data from Perez-Garcia and Meave (2006). Taxon
Species Genera Families
457 306 85
174 109 43
31 40 34
600 375 72
10 21 72
Table 3 Plant species richness and shared species between sub-tropical dry forest and savanna in Hluhluwe, South Africa. Study area is 80 km2. Data ex Gray 2008 (unpublished). Growth form
Woody Graminoids Forbs Total
65 6 58 129
41 14 98 153
30 4 27 61
76 16 129 221
between dry sub-tropical forests and adjacent savanna habitats with several species characteristic of savanna and others of forest habitats (Parr, Gray and Bond, unpublished). In northern Australia, distinct ant assemblages occur in the tropical forests with little overlap in species composition with savanna ant assemblages; for example, van Ingen et al. (2008) found that 75% of the 58 ant species in the savanna woodlands were restricted to this habitat and did not occur in forests. Fisher and Robertson (2002) found similar patterns in Madagascar with higher species richness of ants in grasslands than adjacent montane forests and marked species turnover. As in Australia most of the ant species collected in the grasslands were restricted to this habitat. Although grassy biomes may have lower richness than adjacent tropical rainforests, these areas clearly warrant conservation attention due to their diverse and distinctive ﬂora and fauna. Importantly, although in many areas grasslands are regarded as a sub-climax vegetation state, faunal diversity including levels of endemism can be high. For example, the grassland biome of South Africa contains 10 endemic grassland bird species (six of which are threatened), and 10 of the 14 globally threatened bird species present in South Africa have major strongholds here (Neke and du Plessis, 2004). In India, the Shola grasslands of the Western Ghats harbour high faunal diversity and a number of endemic species. These include restricted endemic birds (e.g. Nilgiri Pipit, Broad Tailed Grassbird, Birdlife International), mammals (e.g. the Nilgiri tahr, an Asian goat-antelope, Thomas and Palmer, 2007), insects (e.g. butterﬂies, Kunte et al., 1999) and amphibians such as caecilians which have been found in a range of habitats including open, disturbed areas (Bhatta et al., 2007). Furthermore, Sankaran (2009) notes that the grassland is important for sustaining the nine large mammal herbivores present which in turn provide the prey for tigers and leopards in this world heritage site. In other areas where grasslands are considered degraded ecosystems (e.g. Madagascar and Indonesia) there is evidence of both richness and specialisation. The Imperata grasslands of East Kalimantan contain distinctive assemblages of parasitic wasps dominated by idiobionts of herbivores with relatively high habitat speciﬁcity (nearly 15% of all species collected only occurred in the grassland habitat) (Maeto et al., 2009). A similar pattern was found in Madagascar with bird assemblages in ‘degraded’ savanna habitats differing in composition to forested areas: included within the savanna assemblages were a number of granivores such as the Madagascan mannikin, Madagascan lark, and the range-restricted endemic, the Madagascan sandgrouse (Pons and Wendenburg, 2005). The presence of many grassland specialists and endemics across a range of other taxon groups is highly suggestive of the antiquity of Madagascan grasslands (Fisher and Robertson, 2002; Bond et al., 2008).
5. Grassy biomes and ﬁre 5.1. Fire regimes Grassy biomes of the humid seasonal tropics have all the necessary ingredients for high ﬁre activity. The C4 grasses are very productive during the wet season, but dry out rapidly in the dry season. C4 clades in the wet tropics have high nutrient-use efﬁciency, producing leaf tissue with low nitrogen concentrations that decompose slowly (Bond, 2008). The accumulation of dead standing litter produces highly ﬂammable fuels at a rate surpassing any ﬂammable woody vegetation. C4 grassy biomes support the highest ﬁre frequencies on earth (Mouillot and Field, 2005). Natural ignition is most common from thunder storms during the transition from the dry to the wet season in monsoonal climates. In an interesting analysis, Cardoso et al. (2008) noted that savannas occur where the climate favours high C4 grass productivity and con-
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vective storms associated with lightning in the wet–dry transition. In geographic regions which have a climate suitable for savannas, but where convective storms are absent at the end of the dry season, forests are the dominant vegetation. Cardoso et al. (2008) suggest that, in regions otherwise suitable for savannas, lightning in the right season is the key ingredient for determining whether forests or savannas were the dominant natural vegetation. Human ignitions pre-empt lightning ignitions in most parts of the world and are generally shifted earlier in the dry season than lightning ignitions (e.g. Laris, 2002; Andersen, 2003). The consequences of shifts from natural to human ignitions are not well understood. Early dry season ﬁres are less intense and less damaging to trees and saplings. Thus early season ﬁre regimes lit by people result in increases in woody cover and, probably, more diverse woody communities than lightning ﬁre regimes (Laris, 2002). Shifts from lightning to anthropogenic ﬁres may also have consequences for the landscape mosaic of forest and grassland. For example, Geldenhuys (1994) noted that forests occurred on landscape facets which were in the lee of prevailing winds during extreme, lightning-ignited, ﬁre weather conditions. Landscape patterns developed under lightning ﬁre regimes may develop ﬁre-resistant forest margins. Anthropogenic ﬁres, burnt at different seasons with different prevailing wind directions, may penetrate more deeply into vulnerable forest boundaries that are less ﬁre-proof. The idea has yet to be tested. 5.2. Plant traits 5.2.1. Grasses Many species in temperate shrublands and forests with crown ﬁre regimes have ﬁre-stimulated reproductive traits such as ﬁrestimulated ﬂowering, seed release and seed germination. In contrast, savannas and grasslands, with surface ﬁre regimes and very frequent ﬁres, have few if any species with ﬁre-stimulated seed germination or seed release. However many forb species show ﬁre-stimulated ﬂowering exploiting the brief post-ﬁre period before grass growth suppresses them (reviewed by Miranda et al., 2002 for Cerrado). Nearly all plants have vegetative traits that allow them to persist by resprouting after ﬁres. Their dependence on ﬁre is indirect where ﬁres are essential for maintaining an open habitat. Most plants tolerant of C4 grasses are intolerant of shade and dependent on an open, well-lit environment. The C4 grasses themselves have a high light requirement which may be intrinsic to the C4 photosynthetic pathway (Sage and McKown, 2006). Some of the most common C4 grasses in humid grasslands are so intolerant of shade that they die from self-shading if standing litter is not removed (usually by ﬁre) (Everson et al., 1988). Themeda triandra in South African mesic grasslands, for example, declined from 75% cover to <10% cover after four years without burning due to shelf shading (Uys et al., 2004), and Sarga intrans from the savannas of Northern Australia also declines with ﬁre exclusion (Scott, 2008). 5.2.2. Forbs A similar dependence on frequent defoliation has been reported for grassland forbs. The problem for most forbs is that they are easily overtopped by grasses and shaded out. Vernal forbs survive by exploiting the brief seasonal window when grasses are not shading them. This occurs either in the early growing season before grasses have started growth or immediately after ﬁre. A remarkable feature of grassland forbs reported in South Africa (Uys, 2006; Bews, 1925), the Cerrado (Filgueiras, 2002) and campos grasslands of southern Brazil (Overbeck and Pfadenhauer, 2007), is the widespread occurrence of underground storage organs (USOs). In contrast to North American prairies, annuals and short-lived pauciennials are very rare in these ancient grasslands. Instead
Fig. 2. Underground storage organ in an African grassland forb, Raphionacme lucens. Underground storage organs are very common in frequently burnt grassy vegetation in the tropics and sub-tropics, and facilitate rapid post-burn sprouting. Photo credit: N. Zaloumis.
the perennial ‘herbs’, both monocots and dicots, have swollen roots from which they resprout rapidly after ﬁre. These USOs can be very large (Fig. 2) suggesting that the plants that bear them may be very old. The forbs sprout very rapidly after ﬁre, and typically ﬂower before grass growth is sufﬁcient to shade them. Many of these forbs, in both African (Bews, 1925) and Cerrado (Filgueiras, 2002) grass layers, are poisonous to livestock and therefore protected from herbivory in their most vulnerable stages. In the absence of ﬁre to create the necessary light gap, these ﬁre-dependent forbs senesce and die out. In an analysis of forb diversity in long term burning experiments in South African grasslands, Uys et al. (2004) found that the ﬁre-dependent forbs disappear after ﬁre-free intervals of 8 years or more. In Brazilian campos grasslands, forbs have a similar dependence on ﬁre and diversity declines when grasslands are left unburnt or ungrazed (Overbeck and Pfadenhauer, 2007; Overbeck et al., 2007). 5.2.3. Woody plants Savanna trees diverge from forest trees in a number of key functional traits (Hoffmann et al., 2003, 2004; Rossatto et al., 2009). The common woody plants of savannas are intolerant of shade as indicated by their absence as juveniles in expanding forests and their disappearance from ﬁre exclusion plots as woody cover increases (e.g. Louppe et al., 1995; Woinarski et al., 2004). In contrast to forest trees, savanna trees are well-equipped to cope with frequent ﬁres. Both forbs and woody plants of savannas have underground storage organs and other belowground traits that promote persistence under very frequent burning (Hoffmann et al., 2003; Wigley et al., 2009; Schutz et al., 2009). Tree seedlings rapidly acquire the ability to sprout and differ in this respect from forest trees species that do not allocate signiﬁcant resources to below ground USOs (Hoffmann et al., 2004). Humid savanna saplings are distinguished from forest and arid savanna trees by the presence of large swollen roots and a pole-like stem that facilitates rapid emergence from the ﬂame zone (Gignoux et al., 1997; Wigley et al., 2009; Schutz et al., 2009). Clonal spread by root suckering is another common strategy for surviving frequent savanna ﬁres found in shrubs and trees (Lacey and Johnston, 1990; Wakeling and Bond, 2007). Among the more remarkable examples of these are the ’geoxylic suffrutices’, described as ‘the underground forest of Africa’ by White (1977) but also present in South America (Filgueiras, 2002) and Australia. These look like populations of forbs with leaves emerging from short woody shoots near ground-level.
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White (1977) likened them to underground trees with the branches buried beneath the soil and just the tips with the leaves emerging. They occur in frequently burnt, often nutrient poor soils and many have close relatives that are trees; they do not occur in forests. 5.3. Plant traits, ﬁre and open vegetation specialisation The traits of the grasses, the forbs, and the woody plants of C4 grassy biomes all point to high tolerance of frequent ﬁres and intolerance of prolonged shade. Of course species vary in their tolerance of both ﬁre and shade but the abrupt compositional changes from grassy to forest vegetation suggests distinct ﬂoras with opposing light requirements and disturbance tolerance, and not a continuum. Reduced frequency and intensity of burning should lead, ﬁrst, to the loss of the rich shade-intolerant forb ﬂora, switches in grass species composition, then colonisation of forest precursors and ultimately a complete biome switch to forest with cascading losses of the grassland biota. Managers of protected areas in grassland/ forest mosaics will have to maintain a grassland ﬁre regime to maintain the grassland diversity and even the grasslands themselves. Fire management has become increasingly difﬁcult in conservation areas because of threats to neighbouring properties and infrastructure within reserves. However surface ﬁres are far easier to manage effectively than the crown ﬁres of shrublands and woodlands such as Californian chaparral or Australian eucalypt woodland. 5.4. Fauna and ﬁre Although relative to vegetation there are fewer studies focusing on fauna and ﬁre (see Parr and Chown, 2003), nevertheless it is clear that overall fauna in grassy biomes exhibits striking resistance and resilience to burning. This contrasts greatly with forests where the fauna, like the ﬂora, is much more susceptible to ﬁre, with even a single-ﬁre event having a detrimental impact on the biota (Barlow et al., 2002; Barlow and Peres, 2004, 2006). High overall resilience has been shown across faunal groups for both single-ﬁre events and repeated ﬁres. Savanna insects show high resilience to burning with little change in richness, abundance or assemblage composition following ﬁres (Andersen et al., 2007; Parr et al., 2004; De Souza et al., 2003). Birds, too, are little affected; indeed many species make active use of ﬁres whether during the ﬁre for feeding or post-ﬁre for nesting (e.g. Bronze-winged courser in southern Africa, Dean, 1974). And, large mammals also make use of recently burnt areas: in Africa and South America they have been documented returning to burnt areas post-ﬁre to feed on new grass re-growth (Riginos and Grace, 2008; Prada, 2001). Small mammals are relatively unaffected by burning (Viera, 1999; Layme et al., 2004), although post-ﬁre successional use of the Cerrado by small mammals has been reported in a few studies (Briani et al., 2004), and extremely high frequency ﬁre such as annual burning is considered detrimental to small mammals in the tropical savannas of northern Australia (e.g. Pardon et al., 2003). There are a range of functional traits that grassland and savanna animal species exhibit; these facilitate their survival in grassy biomes with frequent ﬁres. An examination of these traits can provide clues to differences in ﬁre-resilience between forest and grassy biomes (Andersen et al., 2007). Key traits in ﬁre-dominated systems relate to mobility (ability to move out of affected area and ﬁnd refugia both during and post-ﬁre), persistence (ability to survive post-ﬁre) and resilience (ability to recover) (Moretti and Legg, 2009). Thus, savanna species that are often favoured by burning tend to be those that are competitive, have broad diets, high mobility, and preference for open environments (see Parr et al., 2007). For example, frequent burning promotes competitively dominant
ant species that are also omnivores with subterranean nesting habits (Andersen et al., 2007). In contrast, forest ant species with leaflitter nesting habits and lower competitive ability would be more susceptible to ﬁres. The removal of ﬁre from grassy biomes results, as with plants, with the gain of rainforest-associated species (Andersen et al., 2006). However studies show that where some savanna elements still remain in the community, ﬁre-resilience can be maintained. For example, a study in northern Australia where ﬁre had been excluded from a tropical savanna for 13 years, found that the ant assemblage (richness, abundance and composition) was unaffected by the re-introduction of burning (Parr and Andersen, 2008).
6. Threats to grassy biomes 6.1. Land clearance for crops, urbanisation and forestry Grassy vegetation has been extensively cleared for crops and plantation forestry or used as rangelands for livestock. Afforestation projects aim to improve ‘degraded grasslands’ and make them more productive; they involve the establishment of exotic and indigenous tree species as well as through the planting of biofuels such as oil palm (e.g. Jussi et al., 1995, Otsamo, 2000). Afforestation is increasingly motivated by carbon sequestration to mitigate greenhouse gas emissions. The Cerrados of Brazil, a global biodiversity hotspot, are estimated to have lost 40 (Ratter et al., 1997) to 80% of their area (Mittermeier et al., 2000) to agro-pastoral use, outpacing the rate of rainforest loss (Ratter et al., 1997). The large range of these estimates has been attributed to difﬁculties over scale of analysis, methodology used and, particularly, failure to include reversals of land cover (Jepson, 2005). Thus Jepson (2005) reported that in her study area, Cerrado cleared for crops often reverted to savanna so that, over a 13 year period, gross reduction of Cerrado was 1338 km2, but net reduction was only 385 km2. This indicates the potential for rapid structural regeneration of Cerrado when land use is changed. Jepson’s paper raises the important issue of the reversibility of land cover changes. Unfortunately the quality of secondary grasslands as habitat for plants and animals seems seldom to have been evaluated. There are good reasons to expect that restoration of the species diversity of the grassy biomes is a far more difﬁcult prospect than restoring some sort of grassland structure (see section below) but studies are urgently needed. The rich Campos grasslands of Brazil have also undergone extensive land transformation. Between 1970 and 1996 alone, some 25% of the 14 million ha of Brazilian campos were converted to agriculture or plantation forestry (Bilenca and Miñarro, 2004). It is estimated that about 48% of Brazilian campos is still natural grassland but only 0.36% is formally protected as conservation units (Bilenca and Miñarro, 2004). Much of the area is being considered for afforestation with monocultures of eucalypts and conifers. In South Africa also, afforestation is a major conservation threat to the C4 grassland biome. The grassland biome covers an area of 349,174 km2, or 16.5% of South Africa’s land surface. In 1995, some 53% remained as semi-pristine grassland (largely used for livestock farming), of which only 1.6% was formally protected (Neke and Du Plessis, 2004). The remainder had been converted to cropland, plantation forestry, urban and industrial land, or areas invaded by non-native woody species (Neke and Du Plessis, 2004). Afforestation, although occupying a relatively small total area of the biome (3.3%), is perceived to be one of the most severe agents of transformation since plantations were established in areas with the highest species richness of grassland birds, including globally threatened species (Neke and Du Plessis, 2004; Allan et al., 1997).
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6.2. Biome switches and forest invasion of grassy ecosystems Whereas ﬁres and savannization are major threats to the future of forests (Cochrane et al., 1999: Barlow and Peres, 2004), one of the key threats to high rainfall grassy biomes, and the biodiversity they contain, is a switch of system state through forest invasion. The grassy portion of forest–grassland mosaics is shrinking. Indigenous forest expansion into savannas and grasslands is especially widespread and has been reported in Australia, Africa, India, and North and South America (e.g. Banfai and Bowman, 2006; Puyravaud et al., 2003; Silva et al., 2008; Wigley et al., 2009). Though often attributed to land use changes, such as reduced ﬁre (Nangendo et al., 2002; Woinarski, 2010) or heavy livestock grazing, forest expansion may also be due to global drivers favouring trees at the expense of grass. Increased growth rates of forest margin trees and/or increased sprouting ability of forest margin species could promote expansion of forest margins. Increased atmospheric CO2 is a plausible candidate for a global driver promoting woody plants in frequently burnt grasslands (Hoffmann et al., 2000; Bond et al., 2003; Kgope et al., in press). Wigley et al. (in press) compared forest expansion in three adjacent but strongly contrasting land use systems in a mesic South African savanna. Forest increase from 1937 to 2004 was least in a densely populated communal farming area, greater in a commercial ranching area, and greatest in a conservation area. This is surprising since the conservation area is perhaps the closest approximation to a natural ’control’ for a forest/ savanna mosaic. It had the full complement of African megafauna native to the region, including elephants, giraffe and black rhino, and has been subjected to frequent burns by park managers. Yet despite the presence of agents of natural and managed disturbance, forests expanded from 10% of the study area in the 1930s to 60% in 2004 swallowing up the species-rich grasslands. Though trajectories of tree cover varied among the three study areas conﬁrming the importance of local land use on forest expansion, all showed massive increases in tree cover over the 20th century consistent with a global driver (Wigley et al., in press). Forest expansion in forest/grassland mosaics is likely to be a major threat in smaller protected areas or in areas with frequent natural barriers to ﬁre spread such as valleys and rivers. This is because changes in the forest mosaic, such as the expansion of forests along riparian areas, may create natural ﬁrebreaks preventing the free movement of ﬁre across the landscape. Fire exclusion experiments in Africa and the USA show that high rainfall savannas can be replaced by forest in as little as 20–30 years (Louppe et al., 1995; Peterson and Reich, 2001). Where fragmentation of grassy vegetation stops spread of ﬁres, there is therefore the potential for a landscape scale hysteresis (regime shift) and major changes in vegetation cover from open grassy ecosystems to closed scrub forest. Loss of grassland specialist birds may be a useful early warning sign of landscape scale shifts to forest. The few available studies of bird community responses to changes in woody cover in grassy systems indicate the presence of specialists for different levels of woody plant cover with signiﬁcant species turnover from open grasslands to woodlands to forests (van Rensburg et al., 2000; Skowno and Bond, 2003). In a study of a South African grassland region undergoing afforestation with eucalypts and conifers, Allan et al. (1997) found 90 grassland bird species, 26 endemic to southern Africa, compared to 65 woodland and forest species of which nine were endemic to southern Africa. Of the grassland species threatened by afforestation, 25 were of conservation concern and ten listed as globally threatened. Of the forest and woodland species which would be promoted by afforestation, one was listed as threatened and none as globally threatened. Allan et al. (1997) noted that grassland bird diversity generally, and globally threatened grassland birds in particular, was signiﬁcantly and nega-
tively correlated with the extent of afforestation. Further afforestation poses a signiﬁcant extinction threat to the grassland specialist species. More studies are needed exploring the structure of landscape mosaics that will best promote the full diversity of species. Given the large habitat requirements of birds, they may make particularly useful indicators of the appropriate scale of landscape mosaics necessary to support both open and closed habitat specialists within forest/savanna mosaics. 6.3. Global change and the future of C4 grassy biomes Besides the projected loss of grasslands and savannas to crops, plantation forestry and urban spread, C4 grassy biomes face a very uncertain future in a high CO2 and a warmer world. C4 grass evolution is closely tied to a low CO2 world and the photosynthetic advantages of the C4 photosynthetic pathway depend on low CO2 (Ehleringer et al., 1997; Sage, 2004; Ehleringer, 2005). In the next few decades atmospheric CO2 is expected to increase to levels not seen for more than 20 million years, exceeding levels when C4 grasses ﬁrst appeared. In a high CO2 world, tree seedlings and saplings can recharge root starch reserves more rapidly than ever before so that, after ﬁre damage, saplings may grow to ﬁre-proof sizes more frequently than ever before (Hoffmann et al., 2000; Bond et al., 2003; Kgope et al., in press). The net effect should be that forests expand at the expense of savannas. The reverse pattern, where grassy vegetation expands into forest, can occur naturally when rare ﬁres under extreme ﬁre weather conditions penetrate deep into the forest interior (Cochrane and Laurance, 2002). Repeated ﬁres in the forest interior, triggered by extreme weather conditions compounded by logging activities, can also lead to ‘savannization’ of tropical forests (Cochrane et al., 1999; Barlow and Peres, 2004). If extreme weather conditions increase under global warming, then even well conserved forests will shrink at the expense of the grasses. Both trends, forest expansion and contraction, have been observed at local scales in diverse geographic regions. It is hard to predict what the future holds for forest vs. grassy biomes given these contrasting threats. The one certainty is that both will suffer extensive land clearance for crop farming (Fearnside, 2001). A separate CO2-related effect is likely to cause additional major disruption of grasses in tropical and sub-topical savannas. The climatic conditions under which C4 grasses outperform C3 grasses are predicted to virtually disappear at the end of the century under a 2 CO2 scenario according to Collatz et al. (1998). Whether the grasses will disappear too, presumably replaced by C3 grasses, and at the same rate as changes in CO2, is unlikely (Sage and Kubien, 2003) but has been poorly explored. However it is clear that the future of C4 grassy biomes under current and future high atmospheric CO2 is very uncertain and the vast grassy ecosystems in the warmer parts of the world will be severely disrupted. 7. Grassland restoration If grasslands are successional to forests then grassy vegetation should be easy to restore when crop land is abandoned. Some grasses are indeed capable of rapidly colonising disturbed land. However secondary grasslands may show little resemblance to primary grasslands in community composition just as secondary forests differ from old growth primary forests. Indeed we would expect slow rates of recovery of species composition for the many species with strong persistence traits such as root suckering, large swollen starch-storing roots, and thick bark. Persistence is traded off against traits that promote seedling recruitment so that strong persisters tend to be poor recruiters and consequently poor colonisers (Bond and Midgley, 2001). There are, as yet, few studies of
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restoration of tropical grassy ecosystems. However in a South African coastal grassland restored after afforestation, forb diversity was less than half that of pristine grasslands (73 vs. 163 species), while sprouting species had shrunk to nearly a quarter of the pristine grasslands (36 vs. 130 species). Major differences in species composition between secondary and pristine grassland were apparent, especially in the forbs. No increase in forb species richness was apparent with increasing time since deforestation (Zaloumis and Bond, in press). For these grasslands, restoration is not really feasible and every effort should be made to conserve the tiny intact remnants remaining. Similar slow recovery of species richness might be expected in Cerrado and campos grasslands of South America where the non-grassy ﬂora is dominated by vigorous sprouting species (Overbeck and Pfadenhauer, 2007; Coutinho, 1990; Miranda et al., 2002). 8. Conclusions Tropical grassy biomes have suffered a bad press for nearly a century. In part, we suggest, this is because of an inappropriate conceptual model for understanding non-forested vegetation. Grasslands were seen as successional stages to forest and ﬁres as intrinsically ‘unnatural’ and anthropogenic. One legacy is that we still know very little about patterns of diversity in grassy biomes. The future of C4 grasses is very uncertain in a high CO2 world. Their photosynthetic advantages over C3 grasses will diminish and may result in replacement by C3 species while CO2 fertilisation of woody plants may results in loss of grasslands altogether and their replacement by forests. Savannas are also easy targets for clearing for cropping in tropical regions. They are easier to convert to crops and do not attract the same public censure as clearing of forests. Where mosaics of forest and grassy vegetation occur in protected areas, it is particularly important to study and understand both ecosystem states so as to secure a landscape pattern that can sustain diversity of both ecosystem states into the future. Spatial scale of landscape units is critical for providing sustainable habitat for specialist species. In the context of alternative stable states, we need to have a far better understanding of the stability of biome boundaries and what causes them to change. It is far from clear whether a high CO2 world will favour forest advance and the destruction of grassy vegetation, or whether global warming will result in more extreme ﬁres that burn deep into the forests causing grassy vegetation to advance. Conservation biologists tend to work either within grassy ecosystems or within the forests. We need more folk who work in both systems to help develop a richer understanding and appreciation of the full diversity of wet–dry tropical ecosystems. Acknowledgments Thanks to Emma Gray and Nick Zaloumis for unpublished data and Mahesh Sankaran and Jayshiree Ratnam for exposure to shola grasslands and other ecosystems of the Western Ghats. WB would also like to thank Caroline Lehmann and participants in the Darwin savanna working group for promoting a global perspective on tropical grassy ecosystems. WB thanks the Mellon Foundation and the National Research Foundation of South Africa for sustained funding. CLP is grateful to the Trapnell Fund for support. We thank Jos Barlow for the invitation to contribute a perspective from ‘outside the forest’ to this special issue. References Adejuwon, J.O., Adesina, F.A., 1992. The nature and dynamics of the forest–savanna boundary in south-western Nigeria. In: Furley, P.A., Proctor, J., Ratter, J.A. (Eds.),
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