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    elephants are not the most abundant species in african grasslands, yet they influence community structure. the grasslands contain scattered woody plants, but they are kept in check by the uprooting activities of the elephants. take away the elephants, and the grasslands are converted to forests or to shrublands. the newly growing forests support fewer species than the previous grasslands. which of the following statements describes why elephants are the keystone species in this scenario?


    Guys, does anyone know the answer?

    get elephants are not the most abundant species in african grasslands, yet they influence community structure. the grasslands contain scattered woody plants, but they are kept in check by the uprooting activities of the elephants. take away the elephants, and the grasslands are converted to forests or to shrublands. the newly growing forests support fewer species than the previous grasslands. which of the following statements describes why elephants are the keystone species in this scenario? from EN Bilgi.

    Many shades of green: the dynamic tropical forest–savannah transition zones

    The forest–savannah transition is the most widespread ecotone in tropical areas, separating two of the most productive terrestrial ecosystems. Here, we review current understanding of the factors that shape this transition, and how it may change under ...

    Immaculada Oliveras

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    Many shades of green: the dynamic tropical forest–savannah transition zones

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    Agroforestry parklands in sub

    Table of Contents

    On the one hand, parklands have been referred to as a vegetation type. They are physiognomically comparable to ‘tree savannas’ and have been referred to as such. The terms ‘savanna parkland’ and ‘park savanna’ or ‘parklike savanna’ are sometimes used by phytogeographers. Specifically, Cole (1986) defined savanna parklands as “tall mesophytic grassland (grasses 40–80 cm high) with scattered deciduous trees (less than 8 m high)” in a savanna classification first proposed in 1963. This vegetation type is intermediate between savanna woodlands, “deciduous and semi-deciduous woodland of tall trees (more than 8 m high) and tall mesophytic grasses (more than 80 cm high)”, and savanna grasslands, defined as “tall tropical grassland without trees or shrubs”. In Cole's analysis, savanna parklands occur in Australia and in Africa, mostly in Central and Southern Africa, but not in South America, with the exception of the region of the High Pantanal. Forests with a very open canopy and a history of frequent fire in Western Canada and northwestern United States are also called parklands (British Columbia Ministry of Forests, 1991).

    However, while tree savannas can occur naturally or as a result of edaphic features, fire and grazing in the absence of cultivation, the term ‘parkland’ as used in this report specifically applies to landscapes derived from human agricultural activities (Pullan, 1974). In parklands the composition and density of the woody vegetation is altered in order to facilitate its use. Most often parklands are not the product of a single agricultural season, but reflect a slow process of species selection, density management, and tree growth over one or several decades. Parklands, in the strict sense of the word, are specific to permanently cultivated fields or fields where fallow duration is shorter than necessary for the regeneration of a second-growth forest (Seignobos, 1982; Raison, 1988). They would not, therefore, include the relic of a natural forest temporarily left standing by frontier farmers (Pélissier, 1980a).

    Parkland trees stand out as an important component of the spatial structure of the landscape (Sautter, 1968, cited in Raison, 1988; Seignobos, 1982). Parkland attributes include a regular distribution of relatively even-aged trees or shrubs and a low tree density so that tree cover is never continuous. Their name derives from their resemblance to urban or rural recreational parks with large scattered trees in expanses of grass. Examples of parklands are more common in the semi-arid or subhumid tropics, and particularly in West Africa which will be the main subject of this report as it is of most of the literature reviewed in it.

    Parklands are not, however, limited to the Sahel and Sudan zones of Africa. While they may not generally be called parklands, systems with scattered trees in fields with similar appearance and purpose are also widespread in Zimbabwe (Campbell et al., 1991) and Malawi (Maghembe and Seyani, 1991) and elsewhere in southern Africa. Several systems in Asia, Oceania, and Latin America would qualify as agroforestry parklands by definition. Information on practices in these areas is more limited, however (Baumer, 1994; Raison, 1988). In India, the well-known Prosopis cineraria is commonly protected in fields planted with millet and legumes, and occurs on fallows and grazing lands in the semi-arid zone of Rajasthan (Mann and Saxena, 1980). Farmers value its high ecological combining ability and suitability for pruning and fodder as well as its socio-cultural significance. Other species found in these systems include Ziziphus nummularia, Acacia nilotica var. cupressiformis and var. radiane, Acacia leucophloea, and Salvadora persica (Shankarnarayan et al., 1987). In the state of Tamil Nadu, several tens of multipurpose tree species are naturally regenerated or planted on farmlands (Jambulingam and Fernandes, 1986). To cite only a few, Borassus flabellifer is grown with cereals and pulses, A. leucophloea with millet and horsegram, and A. nilotica in rice fields.

    Parkland or related systems also exist in temperate regions. Dehesa (Spanish) or montado (Portuguese) systems are centuries old, mainly silvipastoral systems in southwestern Spain and southeastern Portugal where holm oak (Quercus rotundifolia, syn. Quercus ilex) and cork oak (Quercus suber) are scattered in pastures or cereal (oats, barley, wheat) fields (Joffre et al., 1988; Janick et al., 1987). Trees are sometimes seeded and systematically pruned for better fruit and wood production. They provide highly nutritious acorns for fodder and timber, charcoal, tannin, cork (Q. suber), etc. With a density of 20 to 40 trees per hectare, the tree cover may be 5 to 20 percent. This system extends over an estimated 5 million ha in Spain and more than 500 000 ha in Portugal. Similar systems with the same or different species (olive, carob, etc.) also exist in other Mediterranean countries including Morocco, Algeria, Tunisia, southern France (mainly Corsica), Italy (Sardinia) and Greece (Joffre et al., 1988). Moreover, the Acacia caven silvipastoral system is widespread in the semi-arid and subhumid Mediterranean climate zone of Chile, covering 1.5 million ha (ODEPA, 1968, cited in Ovalle and Avendano, 1987). Walnut plantations in the Touraine and Berry regions of France may be included in the parkland family (Raison, 1988). The oak savannas of the eastern United States (New York State) may also be a case of parklands with a silvipastoral focus.

    Among ‘parcs arborés’, the usual French term for parklands, Baumer (1994) distinguished between systems primarily used for cultivation and those used for pastoralism. He suggested ‘forêt-parcs’ as the term for parklands resulting from a high degree of human influence, where trees tend to be monospecific, even-aged, regularly spaced, and form a cover ranging from 1 to 25 percent, and where annual crops, generally cereals, are cultivated, sometimes manured, and grazed after harvest. These trees usually have a positive ecological role (soil fertility or wind reduction) and a strong economic significance. In contrast, ‘parcs arborés’ would describe systems with a parkland appearance but which have not been heavily manipulated by human beings, and are primarily used for pastoralism and gathering of tree products. Examples of such vegetation types include cattle-raising areas in Brazil, such as the open forests of the Sao Paulo and central-east regions, highly forested grasslands of southeastern Queensland in Australia, Miscanthus grasslands under forest relics in Indonesia, and the Daniellia oliveri open dry forests of East Africa. Although the term ‘parcs arborés’ only draws attention to the tree component of the landscape, the term ‘forêt-parcs’ is more problematic. ‘Forêt’ suggests a low degree of human intervention and conceals the major distinction between this and other systems, namely that canopies are spatially scattered rather than forming a closed cover (Depommier, 1996a).

    The major English writings on this topic have used the word ‘parkland’ (Pullan, 1974; Bonkoungou et al., 1994, 1996). Sometimes, however, it is used to describe lands in parks, which bear no relation to the agroforestry parklands considered in this report. The term ‘park’ is sometimes found as a direct English translation of the French ‘parc’, but this is even less explicit than ‘parkland’ and is easily confused with the more common meaning of park as an enclosed or delimited land area managed for preservation or recreation. The term ‘farmed parkland’ is used by Pullan (1974) to encompass parklands being farmed as well as land lying fallow, and does not exclude a pastoralist component. Other labels such as ‘farm parkland’, and ‘cultivation parkland’ have also been used but are less common. The term ‘agroforestry parklands’ incorporates these systems in the fast-growing discipline of agroforestry. It was chosen for this study because it emphasizes the multiple forms and purposes of these systems, and thus includes various schools of thought. As defined by Bonkoungou et al. (1994), agroforestry parklands “are land-use systems in which woody perennials are deliberately preserved in association with crops and/or animals in a spatially dispersed arrangement and where there is both ecological and economic interaction between the trees and other components of the system”. This emphasis on interactions is positive and helpful for a finer understanding of the systems by the research and development (R&D) community, but may be extraneous from a farmer's holistic point of view.

    As mentioned in the introduction, agroforestry is an ancient practice but a relatively new science. In order to assess the significance of the parkland system, it is important to characterize it physically, determine its zone of occurrence, and identify and describe existing parkland types. The next section sets the geographical stage, beginning with a review of the main historical studies of parklands. This is followed by a presentation of the main modes of investigation and existing data, on the basis of which a tentative picture of the wide spatial extension of agroforestry parklands and the distribution and density of prominent parkland species (see Table 1.1) is drawn.

    Being linked to human activities, parklands occur in various latitudes and are not confined to specific agroecological zones. Nonetheless, the most well-known and described parklands are located in semi-arid or subhumid zones where tropical savannas, broadly defined as tropical grasslands with scattered trees (Bourlière and Hadley, 1983), occur.

    We are indebted to the early explorers for the first references to African parklands (Pullan, 1974). Mungo Park described the frequent occurrence of Vitellaria parklands along the Niger River from Segou towards the east in Bambaraland (Miller, 1954), while Caillié (1830) noted the various tree species around houses and Vitellaria and Parkia trees in the surrounding agricultural landscape. On his travels to Hausaland in Nigeria, Clapperton (1829) described well-maintained Vitellaria, Parkia and Tamarindus trees located in agricultural lands. The early reports also focused on how products from these species were prepared for various human and animal uses. In the late 1800s, scientific and commercial interest in savanna and parkland species grew. Plant collections were undertaken and tropical African floras prepared, while tree products were sampled and investigated for commercial use in colonial countries of the North.

    Their origin in people's agricultural activities has given parklands a peculiar scientific status amongst plant communities. Pullan (1974) noted that plant ecologists have generally neglected the study of vegetation in cultivated zones. Thus the degree to which parklands were recognized as distinctive vegetation types, or had their importance in successional stages of vegetation regrowth acknowledged, was variable among early plant scientists. Recognizing the widespread occurrence of 1 in the vegetation of southwestern Senegal, Trochain (1940) characterized it as a climax association (or, more specifically, a stable plant community of anthropogenic origin, which replaces the climax vegetation, including various forms of open forest found in the Sudan zone, on silica and clay-rich soils), developed after the species was introduced by nomadic people and their livestock. Roberty (1956) established distribution maps of parkland types in western Senegal (Fig.1.1), distinguishing major species in part based on a classification he established in the Middle Niger Valley. Pélissier (1953, 1966), Savonnet (1959), and Gallais (1967) also described F. albida parklands in Senegal, southern Burkina Faso and the inland delta of the Niger River, respectively.

    Fig. 1.1

    Lely (1925) recognized a parkland type dominated by Parkia, Vitellaria, Afzelia, Tamarindus and Acacia species in Nigeria. Vegetation maps for northern Nigeria (Clayton, 1962, 1963) took into account Keay's (1959) contribution to the recognition of local parkland types. The Samaru Soil Survey Bulletins (1956-71) also mention parklands in northern Nigeria including Parkia, Vitellaria, Ficus, Balanites, Adansonia, Hyphaene and Borassus species and F. albida. Three types of parklands, dominated by F. albida, Vitellaria-Parkia, and Parkia species, are described by FAO (1969) in northwestern Nigeria. In the northeastern part of the country, Leeuw and Tuley (1972) mapped F. albida, Adansonia and Parkia parklands, while Ceiba parklands were reported by Jackson (1970) around Zaria.

    In northern Ghana, Taylor (1960) identified Vitellaria-Parkia-Tamarindus and Adansonia-Tamarindus parkland associations, with small parkland patches of F. albida. Irvine (1961) also referred to mature F. albida, Parkia and Borassus parklands, and Vitellaria-Parkia communities were described by Ramsay and Innes (1963). Wills (1962) also referred to farmed parklands in this area. Early citations for Niger include Dundas (1938) and Fairburn (1945), yet the relationship of the described vegetation to parklands was not clearly made (Pullan, 1974). Parkland communities dominated by Vitellaria, Parkia, F. albida, Borassus and Hyphaene were reported by Pias (1955) in the Chari-Logone lowlands of Chad and Cameroon.

    Parklands did not become a major object of study until fairly recently, perhaps because their definition did not fit squarely into any single discipline. Thus, as noted by Pullan (1974), they fell outside of uncultivated plant communities, the ‘classical’ field of study of plant ecologists. Similarly, in the early days there were only isolated scientific investigations of parklands by agronomists or foresters, the latter having focused on natural forests and industrial or village plantations. The first in-depth system descriptions were in fact undertaken by geographers (Sautter, 1968, cited in Raison, 1988; Pélissier, 1964, 1966). It was only the emergence of agroforestry as a scientific discipline in the 1980s that finally established the study of parklands in their own right.

    The Semi-Arid Lowlands of West Africa (SALWA) Programme of the International Centre for Research in Agroforestry (ICRAF) has contributed to gathering and generating information on parklands in its member states: Burkina Faso, Mali, Niger and Senegal. The Macro- and Micro-Diagnostic and Design exercises conducted in the first half of the 1990s advanced the identification, description and general analysis of the various agroforestry land-use systems on a national basis (see for instance ICRAF, 1990). National parkland reviews carried out within the African Research Network on Agroforestry (AFRENA) in SALWA countries (Sall, 1996; Cissé, 1995; S.J. Ouédraogo, 1995; Ounteni, 1998) represent an important step in synthesizing knowledge on these systems. Maps and parkland typologies have also been produced by ICRAF (1996) for the Dori area, Burkina Faso, and the middle Bani-Niger river basin and Gondo-Mondoro region of Mali.

    Parklands occupy a vast land area, representing a large part of the agricultural landscape under subsistence farming in the tropics and constituting the predominant agroforestry system in semi-arid West Africa (Nair, 1993; Bonkoungou et al., 1994). In Mali, the agroforestry parkland system occupies about 90 percent of the agricultural land area (PIRL, 1988, cited by Cissé, 1995) and is practised by an estimated 2.5 million people on the Mandingue and Koutiala plateaux and the Moyen-Bani-Niger, High Dogon Plateau, Seno, Gondo, Bélédougou, Wenia, Falo and Central Delta zones (Djimdé, 1990). In Burkina Faso, parklands are found throughout settled zones where agriculture is practised, i.e. most of the country with the exception of the extreme North, East, and parts of the South and Southwest where human population density is low (S.J. Ouédraogo, 1995). The parkland system is also recognized as the most common production system in Katsina State in northern Nigeria (Otegbeye and Olukosi, 1993). A discontinuous cover of scattered trees in crop fields is traditional in northern Ghana (Rudat et al., 1996). These references highlight the local or national significance of parklands but also serve to illustrate the lack of a coordinated quantitative assessment of this land-use system at the regional level.

    Nowadays, agroforestry parklands are most often characterized by the dominance of one or a few species. Species composition is generally more diverse and variable, however, in areas located farther away from villages and only occasionally cultivated. In parklands, one or a few dominant species may prevail on a local scale or across large land areas with substantial variations in relative abundance, frequency and overall species composition. Thus, parklands are often described by their dominant species (Pullan, 1974; Weber and Hoskins, 1983). Table 1.2 provides a tentative list of the main species in each climatic zone. Seignobos (1982), however, notes that some parklands, such as those in the northern Mandara mountains in Cameroon or around Kimré in southern Chad, include a large diversity of species without apparent dominance. Otegbeye and Olukosi (1993) also report that parklands with a species mix without specific dominance are the most common types in Katsina State, northern Nigeria.

    Whether or not there is a dominant species, parklands usually host a wide variety of tree and shrub species. For instance, 22 and 39 species were recorded in cultivated fields in two sites around Kano, northern Nigeria (Cline-Cole et al., 1990), 43 and 46 in north-central and southern Burkina Faso (Gijsbers et al., 1994; Boffa, 1995), and 46 in northern Côte d'lvoire (Bernard et al., 1996). Species diversity increases when fallows are included.

    There is very little quantitative information regarding the relative representation of major species throughout the parkland range which could help to prioritize conservation and development efforts in a rigorous way. Faiderbia albida parklands have received considerable attention because of the generally observed positive effects of the tree on soil fertility and crop production. It is likely, however, that V. paradoxa and P. biglobosa parklands occupy the largest land area among parkland types. Breman and Kessler (1995) indicate that V. paradoxa may be the most common parkland species in semi-arid zones of West Africa. Prioritization exercises conducted by ICRAF (Franzel et al., 1996) in selected sites have been useful in assessing the importance of parkland species in the eyes of farmers and thus guiding domestication programmes. There is a definite need for more quantitative data on the main parkland species at the regional level, including their distribution, stocking rates and dynamics, to improve the formulation of parkland research and development activities. Some methods for assessing parkland resources are reviewed in Box 1.1

    (Source: adapted from Pullan, 1974)

    The rest of this section reviews recent information on individual parkland species throughout their zone of occurrence. Species are reviewed in order of their importance in the literature. Where parkland studies included information on tree densities, these are listed in Tables 1.3, 1.4 and 1.5 for F. albida, V. paradoxa, and P. biglobosa respectively. Other species mentioned may seem less widespread but are equally important to farmers. Reports of particular species in agroforestry parklands are generally localized, but their wider distribution is by no means limited to these few geographical references.

    Faidherbia albida parklands occur throughout the Sahel and Sudan zones of West Africa, as well as in eastern and southern Africa where it is strongly associated with alluvial soils along perennial or seasonal watercourses. In West Africa, F. albida still shows a preference for alluvial soils but occurs widely as a result of human activities on deep light sands or sandy clays. The African distribution of the species and the two geographical races identified by Brenan are shown in Figure 1.2 below. Because of its unusual characteristic of reverse foliation (i.e. bearing leaves during the hot dry season), it is water-demanding and has a deep fast-growing taproot to reach the aquifer. It occurs in areas with between 500 and 800 mm rainfall and is also found on lateritic soils where it can reach the water table through an opening in the shallow indurated pan. It also makes incursions into the northern Sahel and Sahara in moist sites or areas with a good water table (CTFT, 1988; Wickens, 1969).

    Fig. 1.2

    Faidherbia albida, as a parkland species, is present in virtually all of Senegal (Fig. 1.3) from the Atlantic coast to the Falémé river and from the Senegal river to the Guinea Bissau border (Giffard, 1964). A highly integrated form of the F. albida parkland system among the Sérer in Senegal is described in depth by Pélissier (1953, 1966). This system is also present among the Wolof (Seyler, 1993) and the Mandingue of Casamance (CTFT, 1988). F. albida parklands are most common in the western part of the country on sandy soils within the Thiès-Louga-Kaolack triangle (Portères, 1952; Seyler, 1993; Sall 1996). In Guinea Bissau, they occur among the Brame and Mandjak ethnic groups (Pélissier, 1980a). In Mali, they extend over an estimated area of 8 780 km2, or 17 percent of the country's total estimated parkland area. They are located in the rainfall range between 500 and 1 400 mm in the areas of Gondo-Mondoro, the Bandiagara-Hombori, Koutiala, and Mandingue Plateaux, the Central Niger Delta and Hod (Diallo, 1988, cited in Cissé, 1995). In terms of density and size, particularly outstanding F. albida parklands are found in the Dembéré-Douentza valley and extending into the Seno plain (Gallais, 1965), while Pageard (1971) reported the occurrence of well developed F. albida parklands in the Niger Valley between Bamako and Mopti, and particularly around Segou. A description of those in the inland Niger Delta is also available (Gallais, 1967).

    The most dense and well developed F. albida parklands in Burkina Faso are seen among ethnic groups located near to the Bambara, Dogon and sedentarized Peul cultivators who have a strong F. albida tradition (Pageard, 1971). The F. albida parkland system is dominant in the Bwa and Samo regions extending to the Senoufo, Lobi, Dagara, Birifor and northern Côte d'lvoire regions. It is also well represented in Bissa country and in Yatenga as practised by the Mossi, though most of the Mossi Plateau is Vitellaria-dominated. It is found only in small, scattered areas in the Southeast and is rare in eastern Burkina Faso. In S.J. Ouédraogo's (1995) parkland classification, F. albida is represented in parklands throughout the country except for the extreme Southwest. For instance, he reported their existence around Markoye, Oursi, and Dori in the North, in Kokologo on the Central Plateau, and in Dossi and Boni in the West. Pure stands are reported in the Bulkiemdé Province west of the capital (Yélémou et al., 1993).

    Fig. 1.3

    In northern Côte d'lvoire, a parkland system including F. albida has been described in Dolekaha. This is one of the most southern areas of F. albida parklands described so far (Bernard et al., 1995).

    In Niger, F. albida parklands are limited to the South and Southwest. Well-known examples are the highly stocked and permanently cultivated parklands in the ‘3M’ (Matameye, Myrriah and Magaria) region, resulting from the enforced protection of the species during the rule of Tanimoun, Sultan of Zinder (Montagne, 1986; CTFT, 1988). References to F. albida parklands in southern Niger also include the zone between Damergou (Tanout) and the Nigerian border (Bertrand, 1991), and Madaroumfa next to Maradi (Montagne, 1996). In the Southwest, F. albida parklands occur in Tilly near Sadoré and in Guilleny (Maï Moussa et al., 1993), as well as in the Dosso Department between the Dallol Bosso and the Dallol Maouri (Montagne, 1984). Faidherbia albida densities in parklands in the Dosso Department are lower than in the Zinder area (Montagne, 1996).

    In Nigeria, F. albida is an important farm tree throughout the savanna zone where rainfall is above 1 000 mm (Sanusi, 1993). An early reference was made to F. albida along with Balanites, Borassus, Khaya, Anogeissus and Ziziphus species in fields on the hillsides of the northern Mandara mountains in the Northeast (White, 1941), and the terraced F. albida system was described later by Hallaire (1976). F. albida is also present in acha (Digitaria exilis Stapf) and millet fields of the Biron group on the Jos Plateau in association with the cattle of Fulani herders (Gosden, 1978, cited in Miehe, 1986), as well as in the North among Hausa farmers (CTFT, 1988). Further east in northern Cameroon, a strong tradition of F. albida parklands has survived among the Toupouri and Massa (Seignobos, 1982). Stands are densest in the highly populated zones around Golompui and

    Datcheka and some stands can also be observed around settlements among the Dowayo in the vicinity of Poli. Faidherbia parklands were widespread in the Diamaré area before their destruction during the Fulbe conquest in the nineteenth century. They are gaining ground again due to an increase in human population density (Seignobos, 1982). F. albida has the highest representation among the tree species present in 30 parkland samples within 30 km of Maroua (Triboulet, 1996). In Chad, relics of F. albida parklands are found among the Sara around the town of Koumra and well developed Acacia communities exist among the Kera around Fianga Lake (Seignobos, 1982).

    Elsewhere in Africa F. albida parklands also occur in Sudan in the Fur farmlands in the Jebel Marra Highlands of Western Darfur (Miehe, 1986). The species dominates montane habitats between 1 800 and 2 300 metres above sea level, but F. albida parklands are also conspicuous in the Wadi Azoum alluvial system in the pediplains west of the Jebel Marra massif (Radwanski and Wickens, 1967). Faidherbia albida occurs along with other tree species in fields in the Nuba mountains in southern Kordofan province (Miehe, 1986). It is widespread in the Hararghe highlands of eastern Ethiopia (Poschen, 1986) and is reported further west around the town of Debre Zeit (Kamara and Haque, 1992), as well as associated with Cordia abyssinica in the Rift Valley among the Galla and Arussi groups (Miehe, 1986). Faidherbia albida also used to be maintained in arable areas of the Upper Jordan valley (Karschon, 1961).

    Some F. albida parklands are found in Southern Africa. They have been studied in Tanzania (Fernandes et al., 1984; Okorio and Maghembe, 1994; Wickens, 1969), in Malawi's Lakeshore plain and upland central region (Saka et al., 1994; Rhoades, 1995), and in Zimbabwe (Campbell et al., 1991).

    Fig. 1.4

    Vitellaria paradoxa is probably the most common parkland species in semi-arid West Africa (Breman and Kessler, 1995). There are two subspecies in the genus, the subspecies paradoxa occurring from Senegal to the Central African Republic and the subspecies nilotica in southern Sudan and Ethiopia, Uganda and northeast Zaire (see Fig. 1.4). V. paradoxa is generally found in wooded grassland and farmland within the rainfall zone of 600 to 1 400 mm, annual potential evapotranspiration from 1 400 to 2 300 mm, and a dry period of 5 to 7 months. It preferentially grows on colluvial soils of reasonable depth (> 30 cm above parent rock or indurated layer), free drainage, and a predominantly sandy topsoil texture (Hall et al., 1996). Like other parkland species, its distribution has also been shaped by human influence.

    A few estimates of the surface area occupied by Vitellaria species are available for Mali. Ruyssen (1957) has Vitellaria parklands occurring south of the line going through Bafoulabe, Segou and Bandiagara with an estimated distribution area of 19.4 million ha, revised to 22.9 million ha by Maïga (1990, cited in Cissé, 1995). Ruyssen (1957) estimated that 943 000 ha had Vitellaria as a significant tree component on land cultivated annually with rainfed crops. Based on a five-year rotation (one cropping and four fallow seasons), the area of productive Vitellaria parkland was estimated at 4.7 million ha. Clamagirand and Bruffaerts (1983, cited in Raison, 1988) state that at least 3.6 million ha in Mali have Vitellaria with a density of over 40 trees/ha. The species was represented in parklands covering 68 percent of the 415 700 ha middle Bani-Niger river basin mapped by ICRAF (1996).

    Vitellaria parklands occur throughout Burkina Faso south of 14° latitude (S.J. Ouédraogo, 1995). In Bulkiemdé Province, Vitellaria has the highest density among parkland species (Yélémou et al., 1993). In Niger, the distribution is irregular and linked to human activities (Ounteni, 1998), with Vitellaria occurring in the south central and southwestern regions of the country (Hall et al., 1996). Vitellaria paradoxa and Parkia biglobosa are the most common indigenous parkland species in northern Ghana (Rudat et al., 1996) and spread through northern Togo and Benin where they have been studied in the Bassila and Parakou areas (Schreckenberg, 1996; Agbahungba and Depommier, 1989). They are also referred to in the north of Côte d'lvoire (Bernard et al., 1995; Louppe and Ouattara, 1996).

    Fig. 1.5

    In Nigeria, Vitellaria parklands cover large areas. They were observed around Birnin Kebbi in the Northwest, south of Bida (southwest of Kaduna) in an area of old Nupe settlements (Pullan, 1974) and west of the Mandara mountains in the Northeast (Seignobos, 1982). Vitellaria paradoxa was cited among the most important farm trees during farmer interviews around the towns of Saki, Ilorin, Minna, Jos, Kaduna and Kano (Teklehaimanot et al., 1995). In northern Cameroon it is represented in many bushfallow communities outside F. albida parkland zones, in Toupouri country for example, but there are few well-developed Vitellaria parklands except on terraces of the Benoue valley west of Garoua (Seignobos, 1982). In southern Chad, there are large expanses of Vitellaria parklands (Seignobos, 1982) with tree densities higher than in F. albida parklands. References are also made to Vitellaria parklands among the Ngambay (area of Moundou in the Southwest) and around Koumra (west of Sarh, northern part of Sara country) mixed with F. albida (Seignobos, 1982). Vitellaria also dominates parklands south of Domo Dambali in Moussey country.

    In Ouham province in the northwest of the Central African Republic, agriculture is traditionally practised under the parkland cover of V. paradoxa associated with P. biglobosa in high densities (Depommier and Fernandes, 1985). Further east, V. paradoxa subsp. nilotica is reported in fields in northern Uganda, where its density is highest in Otuke county in Lira district (Masters and Puga, 1994). Reported V. paradoxa densities in parklands throughout their distribution zones are listed in Table 1.4. Additional information on the species can be found in several monographs (Hall et al., 1996; Bonkoungou, 1987).

    Parkia biglobosa is often found in parklands in association with V. paradoxa. This species is mostly present in areas with between 800 and 1 500 mm of rainfall, and 1 400 and 2 100 mm of potential evapotranspiration, and is generally associated with a dry season of 5–7 months. Its range is similar to that of V. paradoxa but extends further south. Parkia biglobosa typically occurs on mid-toposequence positions on deep soils and sometimes, through farmer protection, on well-drained soils in floodplains and riparian sites. It is absent from depressions where soil drainage is impeded (Hall et al., 1997). The species naturally occurs in the dry forests of the Sudano-Guinean zones where it is associated with Pterocarpus erinaceus (Sall, 1996). In Senegal, it is found on various soils in the Sine, Cayor, Laghem, Saloum, Tambacounda and Niololo Koba areas, but not in the North (Sall, 1996). Parkia biglobosa in parklands is found in the southern tip of Gaya in Niger (Ounteni, 1998), in south Mali (Kater et al., 1992; Bagnoud et al., 1995b), throughout Burkina Faso except north of the south Sahel zone (Teklehaimanot et al., 1997), in northern Côte d'lvoire (Bernard et al., 1995), northern Ghana (Rudat et al., 1996), northern Togo among the Kabre (Enjalbert, 1956), and the northern half of Benin (Schreckenberg, 1996; Agbahungba and Depommier, 1989). In Nigeria, Parkia genus was reported in fields in Bomo, north of Zaria (Pullan, 1974), and was studied in sixteen sites throughout the lowland forest zone up to the Sudan savanna zone (Teklehaimanot et al., 1996b). Pure stands were also observed in northern Cameroon (Seignobos, 1982) and the species pervades parklands in the Ouham province of northwestern Central African Republic (Depommier and Fernandes, 1985). Information regarding the taxonomy, biology, ecology, management, and uses has recently been compiled by Hall et al. (1997). Reported P. biglobosa densities in parklands throughout their distribution zones are listed in Table 1.5.

    Fig. 1.6

    Overall, F. albida, V. paradoxa and P. biglobosa are probably the most widespread parkland species in the Sahel and Sudan zones of Africa. This explains their predominance in the existing literature. However, this should not conceal the importance of a large number of other parkland species, whose representation may be more restricted geographically but which may, on a local scale, be more abundant and more economically valuable. It is worth emphasizing that one of the major assets of agroforestry parklands is their biological (and genetic) diversity. This implies a wide diversity of uses and applications with significant economic value and a variety of management techniques and objectives (both discussed in later chapters).

    The baobab (Adansonia digitata) is one of the best known and often reported tree species in semi-arid Africa, due to its large size and associated mythical and spiritual powers. Adansonia parklands throughout the zone are associated with both old and recent settlements. The species is characteristic of plant communities of the Sudano-Zambezian lowlands with 200–800 mm annual rainfall, but it has extended into higher rainfall areas, possibly with human assistance (Wickens, 1982). It is very common in the intensively managed and permanently cultivated fields around residential compounds. High densities are present in western Senegal around Dakar, Bargny (densest stands) and Thies, north of the Gambia river between Kaolack and Tambacounda, as well as in the Kedougou area in the Southeast (Sall, 1996), and were used as a defensive barrier against attacks by men on horseback (Baumer, 1994). Adansonia parklands are found in Burkina Faso's Yatenga and Bam provinces (S.J. Ouédraogo, 1995), and the species is represented in parklands on 12 percent of the middle Bani-Niger river basin in Mali (ICRAF, 1996). It occurs on old settlement sites in parklands of the southern Mandara, in northwestern Nigeria (Seignobos, 1982) and in northern Togo among the Kabre (Enjalbert, 1956). Additional review information on the species, including its distribution and ecology by regions, is provided in Wickens (1982).

    1 Distribution based on Herbarium and flora records;

    2 Specimens known to be cultivated or introduced;

    3 Distribution based on published and unpublished photographs;

    4 Distribution based on the Kew ‘Baobab Survey’ information;

    5 Records obtained from travel literature, maps etc.

    Fig. 1.7

    Fig. 1.8

    Borassus aethiopum, or fan palm (rônier in French), also has a wide distribution. It is found on a variety of soil types but demands a high water holding capacity and shallow groundwater tables. Its development is favoured by fertile and weeded soils as found in parklands (Niang, 1975; Weber and Hoskins, 1983; Cassou et al., 1997). Among other places, it occurs along all the major rivers and other water bodies of the Sahel. According to Pélissier (1980a), the most beautiful stands are in villages around the mouth of the Soungrougou river (Ziguinchor and Kolda), Casamance in Senegal. Other stands exist in the North in the Goumel forest next to Dagana, in the Northeast in the Matam département, in the Southwest near Tabacounda and on the Gambia river, as well as west around the towns of Thiès, Dakar and Fatick. Very high densities of 50–100 and 115 trees/ha in almost pure stands were reported in the Leolgeou area in Mali (Gallais, 1965) and Wolokonto, Burkina Faso. The species forms one of the ten parkland types classified in Mali (Cissé, 1995). In Burkina Faso, they are most common in the southwestern provinces of BoboDioulasso and Banfora in village or compound fields (S.J. Ouédraogo, 1995). In Niger, B. aethiopum occurs in six major populations estimated at 27 000 ha in the Dallol Maouri, Gaya Department, and 3 000 ha along the Niger river and on Leté island (Ounteni, 1998).

    The species occurs in northern Ghana associated with F. albida. This combination is also found in Nigeria near Foggo, southwest of Azare, an area of Hausa and Fulani settlement (Pullan, 1974), and in northern Cameroon around Mboum settlements in the Guidar region (Seignobos, 1982). Borassus aethiopum is also observed in the North Cameroon inlet of Chad between and along the Logone and Chari river valleys, for example among the southern Kotoko and Mousgoum people around Goffa, Holom, Marmay and Pouss, and among the Massa in Ziguey, Goufka, Domo, Bosgoye, Geme, Nahaide and Dana (Seignobos, 1982). In Chad, the Borassus area extends east and southeast from the Logone river along the southern fringe of the Baguirmi region going through the towns of Logone Gana, Morno, Ngam (Kwang group) to the Sarwa region and even into the Day area (Seignobos, 1982).

    Fig. 1.9

    Dominant species in the Sahelian (northern) parklands include Acacia raddiana (syn. Acacia tortilis), Acacia senegal, Balanites aegyptiaca and Hyphaene thebaica. In Senegal, A. raddiana parklands are located throughout the North (Sall, 1996). Trochain (1940) also mentioned mature A. raddiana trees in groundnut fields in the Sahelian part of the country. Acacia raddiana is found in the Sahel zones of Mali (Cissé, 1995) and Burkina Faso, sometimes in pure stands (S.J. Ouédraogo, 1995). The species occurs almost exclusively in cultivated or agro-pastoral areas, in contrast to A. senegal which is characteristic of rangelands in West Africa. In Sudan, virtually pure stands of A. senegal are systematically rotated for crop and fallow/gum production and are sometimes referred to as ‘gum gardens’ (Seif El Din, 1981; Jamal and Huntsinger, 1993).

    Balanites aegyptiaca is a spiny, fairly short tree, occurring most frequently on the 400–800 mm rainfall zone. The species avoids shallow and gravelly soils and prefers deep sands, sandy clay loams or clays in lowlands. As a shrub, it is an important species on aeolian sands in the Sahel (Hall and Walker, 1991). It is associated with A. raddiana in northern Senegal (Nizinski and Grouzis, 1991; Sall, 1996), Mali (Cissé, 1995) and Burkina Faso (S.J. Ouédraogo, 1995). It is also well adapted to the climatic conditions prevailing in Niger where it is widespread (Ounteni, 1998). It is the most frequent species in parklands in Léré, Koro arrondissement (district) in the Fifth Region of Mali (PRSPR, 1993), and is reported in Vitellaria-dominated parklands in Birnin Kebbi, northwestern Nigeria (Pullan, 1974). Additional data on the species are available in Hall and Walker (1991).

    Hyphaene thebaica, or dum palm, is another economically significant Sahelian species which is easily recognized by its dichotomous branching habit. It occurs in the Sahelian and Sudano-Sahelian zones with 200 to 600 mm rainfall, even extending to the southern edge of the Sahara, generally on light soils with shallow aquifers, in dune depressions or gallery forests but not on rocky soils (Raison, 1988; von Maydell, 1983). It is mentioned in parklands in the Ferlo river valley, Senegal (Freudenberger, 1993b), in the Leolgeou area, Mali (Gallais, 1967), in Tilly, southwestern Niger (Maï Moussa et al., 1993) and in many dallol and goulbi valleys and along the Niger river (Ounteni, 1998). It also occurs in northern Nigeria (Sanusi, 1993) and is characteristic of parklands in the southern Sahelian region of Burkina Faso (S.J. Ouédraogo, 1995).

    Other important parkland species include the oil palm, Elaeis guineensis, found in Benin (Schreckenberg, 1996; Raison, 1988) and northern Togo (Enjalbert, 1956). The exotic species, Azadirachta indica (neem), has expanded widely in Sahelian countries, to the point of sometimes being considered an invasive pest (Ganaba, 1996). It thrives in the Bulkiemdé province of Burkina Faso, especially in compound fields (Yélémou et al., 1993) where it is the second most common parkland species after V. paradoxa (Yélémou, 1993). Prosopis africana is a major, yet declining parkland species along the Chad-Cameroon border north of Fianga Lake and further east in Ngam in Chad (Seignobos, 1982; Bernard, 1996). It also occurs in the southern part of the Dosso and Tillabéry Departments as isolated individuals and as monospecific stands in the Matameye area, Zinder Department in southern Niger (Ounteni, 1998).

    Fig. 1.10

    Ceiba pentandra parklands have been noted around Zaria, Nigeria (Jackson, 1970) and in northern Togo (Enjalbert, 1956). Cordyla pinnata (dimb) was ranked second in relative abundance after F. albida in Sob, Senegal (Lericollais, 1989), and after V. paradoxa in Mali (Ohler, 1985). It thrives in farming systems with high population density in the Sine Saloum area of Senegal because of its nutritional contribution, the kernels being used as a substitute for meat (Niang, M., 1998). Cordia africana dominance was reported in fields devoid of F. albida in the Alemaya area, Ethiopia (Poschen, 1986).

    The list of species occurring in parklands reviewed so far is by no means exhaustive. The following, among many others, can be added: Sclerocarya birrea, Parinari macrophylla, Cola cordifolia, Anogeissus leiocarpus, Bombax costatum, Tamarindus indica, Sterculia setigera, Ziziphus mauritiana, Lannea microcarpum, Lannea acida, Pterocarpus erinaceus, Diospyros mespiliformis, Detarium microcarpum, Combretum glutinosum, Celtis integrifolia, Piliostigma reticulatum, Cordia myxa, Khaya senegalensis, Blighia sapida, Tectona grandis.

    As noted earlier, savanna landscapes have evolved under the influence of natural processes such as fire and grazing. In contrast, parklands reflect deliberate human manipulation of trees in agricultural production systems. Various classification systems attempt to capture regional and local variations in parkland structure and composition according to factors such as the degree of human intervention, their main functional uses, their physical structure, and their reflection of the different natural resource management systems of diverse ethnic groups.

    Geographers first classified parklands in relation to the degree of human intervention that contributed to their formation, an approach which was particularly useful in revealing gradations of management intensity and giving them a broad, inclusive representation among human-manipulated vegetation types. Categories first established by Pélissier (1964), and later adapted by Seignobos (1982) and Raison (1988), are presented below and an attempt made to examine how effective they are in reflecting actual and significant field realities.

    The second type develops with conditions created by ‘clearing’. Such parklands are generally composed of shade-intolerant species, released after clearing, and species which become dominant due to their ability to regenerate quickly from root-suckers (Seignobos, 1982). Typical species include some acacias and Prosopis africana, which benefit crop production and provide excellent fodder. Prosopis africana, however, also persists in old parklands (Bernard, 1996). Elaeis is a similar case (Pélissier, 1964). The two kinds of parkland presented above are often short-lived, but their recognition may be useful to situate their current stage and potential for evolution into a more elaborate and stable type.

    ‘Selected’ parklands are composed of trees which were part of the initial vegetation and have been deliberately protected in fields owing to their various uses including food, fodder or soil fertility. Their composition and density correspond to family and community needs. They reflect old, stable and relatively dense settlements as well as landscapes more deeply influenced by human beings than the previous two types. Classic examples are the karité (Vitellaria paradoxa) and néré (Parkia biglobosa) parklands. Given the slow growth rate of Vitellaria, such parklands indicate a certain degree of permanence. Systems based on the fan palm, Borassus aethiopum (source of palm wine and seedlings eaten during famines), and to some degree the oil palm, Elaeis guineensis, are other examples of selected parklands.

    Quantitative evidence of differences in tree density and species composition in fields as compared to the original savanna vegetation illustrates the impact of human selection on parkland formation. Farmers tend to reduce the total number of stems and species and favour preferred species. In Vitellaria and Parkia parklands of southern Burkina Faso, a third of the species found in uncultivated woodland savanna were not found in adjacent agricultural fields, while nine out of ten woody individuals were thinned during clearing and cultivation. At the same time, the relative abundance of Vitellaria had increased five times from 16 to 83 percent, and that of Parkia nine times from 0.4 to 3.5 percent (Boffa, 1995). Similarly in Benin, V. paradoxa and P. biglobosa represented 10.6 and 0.7 percent of trees in their original savanna vegetation and 39.2 and 3.6 percent in fields, a 3.7 and 5.4 fold increase respectively (Schreckenberg, 1996).

    Fig. 1.11

    The fourth type, ‘constructed’ parklands, is made up of species which may not always be present in the initial vegetation, at least not in the densities observed after farmer selection. They are more elaborate as trees are not only protected but also pruned and tended in order to reach large height and crown dimensions. The best example is that of F. albida which is naturally bushy but which will develop into tall trees when pruned early on. This type has also been called substitution parkland in the case of F. albida (Seignobos, 1982), because the species, which can be absent from climax communities, may partially or completely replace spontaneous vegetation (Pélissier, 1980b).

    The fifth type of parkland displays species disseminated by people in a type of ‘proto-arboriculture’ (Seignobos, 1982). Such parklands are strongly linked to settlements. For example, the baobab, Adansonia digitata, the fruits, leaves and bark of which are systematically gathered, is a familiar sight in Sahelian villages. Years after settlements relocate, its presence may still indicate areas used for cultivation. Borassus aethiopum is another such species. Among others, it was associated with the Baïnouk people and is found along their migration paths from Senegal to north Cameroon (Pélissier, 1980b).

    Finally, ‘planted’ parklands have the highest degree of human intervention and may more appropriately be referred to as intercropped orchards. Such is the case, for instance, of orchards of mango trees interplanted with food crops. However, whether such stands of specially planted trees qualify as parklands is debatable (Raison, 1988).

    The lack of clear definitional boundaries tends to make the different parkland types less operational than desired. Selected and constructed parklands can overlap as in the case of Vitellaria and Parkia parklands. Both are selected from the pre-existing vegetation, but they are also constructed through farmer management. As suggested by the above data, tree size and relative abundance are greatly enhanced through the selective process and their density in parklands may also be increased through fallow enrichment, transplanting or planting. The same can be said of the Acacia senegal (and Acacia laeta) parklands of Sudan, often called gum gardens2. Present in the original vegetation, they are selected and constructed through various cycles of cultivation consisting of coppicing, fallow and gum collection. As parkland trees become more valuable and are managed more intensively by manuring, trimming and pruning practices, the distinction between selected and constructed parklands will probably become less pronounced.

    Similarly, distinctions between constructed, proto-arboricultural and planted parklands can appear artificial. For example, in areas where land use becomes sedentary, A. digitata parklands may be considered to be constructed rather than products of proto-arboriculture. The case of Azadirachta indica, which is both planted and disseminated by birds, encompasses all three categories. Similarly, the notion of substitution of the original vegetation by one species, as in F. albida parklands, is at least potentially present in the (trans-) planting of parkland species. There is no clear consensus on the inclusion of intercropped orchards (mango, citrus, palm) in the parkland family, yet the ‘artificialization’ of tree cover in parklands is likely to become more pronounced as research and development efforts give increasing recognition to trees in these systems. One may speculate that the dissemination and planting of domesticated parkland trees will further challenge the available classification and may increasingly divert attention from categories for whole systems towards a parkland mosaic of niches specific to particular species. Clearly, there are opportunities for refining and operationalizing parkland classifications.

    By allowing certain trees to persist in their fields, local populations have shaped the tree component of their farmland to fulfil specific needs. Among the most impressive and probably insufficiently recognized characteristics of agroforestry parklands are the diversity of tree species they contain and the variety of products and uses they generate. This has led researchers to develop functional classifications for parklands as in the one adapted from Seignobos (1982) in Table 1.6. Some of these categories, such as the use of tree fibre for clothing, appear less relevant today. In addition to the productive roles presented in Table 1.6, trees are maintained in fields for less obvious purposes, such as the provision of shade, fodder storage (crop residues like cereal stalks and cowpea stems stored in branches of tall trees), and a vast array of medicinal products. Chapter 6 elaborates on the quantitative and qualitative diversity of parkland products and services.

    (Source: adapted from Seignobos, 1982)

    ‘Compound fields’ are usually located within a 50 m radius of the compound. They are heavily fertilized with household wastes and manure produced by small livestock and fowl kept close by. Soil fertility and moisture conditions are generally higher here than on other farmland types and can support crops such as maize and red sorghum. Vegetables and other crops used daily are grown near the house for convenience.

    ‘Bush fields’ further out are where the bulk of staple crops such as sorghum, millet, cowpeas, and groundnuts (on relatively sandy soils) are cultivated. Historically, this is a zone of shifting cultivation where the area of fallow grass is much greater than that under crops (Morgan, 1969), the ratio being highly dependent on population density. In recent decades, there has been a strong tendency to shorten or eliminate fallow periods. Fertile and moist lowland areas, where farmers grow vegetable crops, rice, sorghum, etc., may occur regardless of distance from settlements.

    Tree selection and conservation are closely linked to agricultural production activities over a span of years or decades. The spatial structure of parklands is influenced, therefore, by the specific management of each of the above-mentioned cultivation rings. In fact, a terroir villageois (village land) is often said to harbour three distinct zones of tree cover, namely a ‘core’ and a ‘margin’ with an intervening zone of ‘plain parkland’ (Seignobos, 1982). The core or residential zone has species with particular uses which may not be found elsewhere such as Mangifera indica and other fruit and shade trees, Ficus thonningii, Azadirachta indica, Eucalyptus spp., Cassia siamea, Adansonia digitata, Ceiba pentandra, Tamarindus indica, etc. (Savonnet, 1959; Pullan, 1974). The plain or pure parkland zone is found on permanently or repeatedly cultivated fields and visually represents the agricultural system practised. It is usually dominated by one or a few species, such as F. albida, or V. paradoxa. In Birnin Kebbi, northwestern Nigeria, this zone is characterized by the presence of mature trees of species including Vitellaria, Tamarindus, Vitex, Ficus and Balanites (Pullan, 1974). In the Jebel Marra massif of western Sudan, the most extensive F. albida parklands are primarily found in limited areas of very dense and continuous habitation (Miehe, 1986). In parkland margins, a greater variety of species may be selected or selection may be less precise.

    This general threefold zonation is reflected in several studies which record the different names given to each field type in local languages. For instance, among the Bwaba-Bobo-Oulé of Burkina Faso, ka fields around compounds, and extending 25 to 50 metres out, are heavily manured and enclosed with fences, and contain C. pentandra, baobab and tamarind trees (Savonnet, 1959). Outside the ka fields, there are wa fields corresponding to the area covered with F. albida trees in a radius of up to one kilometre. Still further out are the more intensive ma or bush fields, which undergo fallowing and are dotted with V. paradoxa, P. biglobosa and Ficus spp. Likewise, Gallais (1967) notes that the ring of permanent fields is called so-foro in Bambara, or fio by the Bwa people in the inland Niger delta region in Mali, while bush fields are called kongodian-foro and mwese respectively by these two ethnic groups. In Wolokonto, Burkina Faso, compound gardens include mango, citrus and cashew trees, while Borassus aethiopum parklands develop on permanently cultivated village fields which are themselves subject to different intensities of land use. A decreasing gradient of Borassus tree density of 155, 127 and 87 trees/ha was found in three adjacent rings of village fields located at increasing distance from the village. Beyond these, bush fields are characterized by Vitellaria and Parkia parklands and the absence of Borassus (Cassou et al., 1997).

    Fig. 1.12

    Not only do parkland composition and structure result from the type and management intensity of agricultural activities, but they are also influenced by demographic changes. Reflecting such trends, Gallais (1967) outlined stages in the spatial expansion of F. albida parklands in the eastern part of the Kounari region in Mali. The continuum included villages with ‘non-existent’, ‘limited’ and ‘extensive’ parklands, as well as the separate case of ‘distended’ parklands. Villages in this region were devastated by wars and abandoned during the nineteenth century, after which access was open to exploitative wood cutters and herds. More recently, they have been recolonized. In places, trees were cut down and these lands may still remain treeless. In others, degraded parklands were regenerated, but where human density was low, the zone of so-called ‘limited’ parkland is now surrounded by a growing belt of permanently and extensively cultivated fields. Elsewhere in areas of higher population density and tight settlements, ‘extensive’ parklands ranging from the village houses through the cultivated zone were rapidly regenerated to support an intensively-managed farming system. Also, in agropastoral communities Gallais observed ‘distended’ parklands where an inner treeless zone used for cattle pens in the settlement area was surrounded by a parkland ring. The intensive use of F. albida branches for fencing pens and cattle pathways through the cultivated fields to outer grazing zones explained the low tree density in the village centre. Such a spatial pattern was also reported by Pullan (1974) in Birnin Kebbi in northwestern Nigeria.

    Adjacent parkland areas with stands of trees of different sizes have often been reported (Pullan, 1974). One may find a central zone of cultivation with large scattered trees surrounded by an outer zone of smaller farm trees in far higher density, the latter being a zone of expansion of cultivation where conditions for natural regeneration are favourable. The condition of the woody cover on these newly cultivated areas depends on the intensity of wood collection, burning, and grazing. Gallais (1967) also noted variations in tree size ranging from 10–15 cm to 70–80 cm in diameter in the F. albida parkland stands of Mali which revealed their dynamic condition. These observations indicate a relationship between the size of parkland trees, which is often reported as relatively uniform within stands, and time of village occupancy or at least duration of cultivation. In southern Burkina Faso, average diameter at breast height of Vitellaria in farmers' fields was highly correlated with the number of years fields had been cultivated (Boffa, 1995).

    Ethnic groups differ in their settlement and land-management patterns, even in comparable arable landscape units. Variations in social organization may be reflected in general land-use patterns including parkland management strategies at village and regional scales, as illustrated in examples from Burkina Faso (Box 1.3) and Senegal (Box 1.4).

    The structure and composition of parklands may reflect the way an ethnic group manages the vegetation in fields and fallows. The composition of dominant species indicates the general type of agricultural system while secondary parkland variables may be determined more by cultural and ethnic factors (Seignobos, 1982). This may explain why early explorers observed variations in parkland composition as they crossed West Africa (Kirk-Greene, 1962, cited in Pullan, 1974). For instance, Seignobos (1982) illustrated how F. albida and V. paradoxa parklands reflected two fundamentally different agrosystems and rural societies.

    Faidherbia albida is the dominant species of the Sérer parklands in Senegal (Pélissier, 1966). The cultivated area is divided into at least two or three zones cropped and fallowed in rotation on a biennial or triennial basis. Elsewhere F. albida parklands are usually associated with permanent cultivation. During the cropping season, cattle are restricted to fallow areas and circulation to and from the outer zones of uncultivated forest is facilitated by live and dead fences around fields and along paths. At night, they are tethered and progressively moved in order to ensure homogeneous manuring of fields. After harvest, they are free to roam throughout the village, but remain secured at night.

    Trees, cattle and crops are strongly interconnected in an intensively managed system. Trees contribute to the maintenance of soil fertility through nitrogen fixation and cycling of mineral elements from deep soil layers, but their pods and leaves also provide substantial quantities of fodder for cattle which, in turn, are a source of manure for soil productivity.

    In contrast, V. paradoxa parklands may indicate an absence of cattle in the local farming system and little or no contact with pastoralists; Vitellaria kernels, from which a vegetable butter is extracted, compensate for the absence of milk and animal butter. This type of parkland is usually associated with a relatively stable settlement and a more extensive cropping regime than in the case of F. albida. Farmers typically clear and cultivate fields at the village periphery, beyond a limited zone of permanent cultivation around settlements, and these are shifted further out after three or four years of cultivation, while a fallow period of 12 to 15 years is applied. Land availability permitting, residence sites may be moved after two or three cycles to allow the land to rest longer. Over several decades, a stratum of selected trees, often dominated by V. paradoxa and P. biglobosa, develops. This land-management system requires a large area of land and population density associated with V. paradoxa parklands therefore tends to be lower than in F. albida parklands. Unlike F. albida, the Vitellaria system supports only one aspect of the economy rather than the whole agricultural system.

    Some authors argue that ethnic specificities could result in the mutual exclusion of the two species. For instance, Pélissier (1980a) claimed that the Vitellaria genus is absent from communities which are either purely agropastoral or have maintained close relations with cattle raisers. This would explain why it is not found west of the Falémé river in Senegal. However, natural factors limiting the extension of the species could also be responsible. The presence of both species in the Dagomba area in Ghana as well as in other places in West Africa illustrates their compatibility (Sène, 1998). The two species are also interspersed where distinct ethnic groups used the same area either subsequently or simultaneously as a result of adoption (Marchal, 1980; Gallais, 1967; Savonnet, 1959; Dubourg, 1957). For example, in Dakola, Burkina Faso, stands of F. albida are found only in very old settlement sites established by Dogon and Samo populations before the Mossi conquest, which subsequently introduced V. paradoxa, P. biglobosa and Tamarindus indica (Kohler, 1971). This is also what Izard-Héritier and Izard (1959) and Marchal (1978) observed in Yatenga. Historical research in areas where the species occur together or exclude each other would provide highly interesting insights into the promotion of parkland species by various ethnic groups.

    The influence of specific ethnic groups on the extension of parkland systems is also referred to by Pageard (1971). In Mali, the most elaborate and extensive F. albida parklands are located between Bamako and Mopti where sedentary cultivators of the Niger valley, namely the Bambara, the sedentary Peulh descendants of the Macina empire, and the Dogon people, live. In Burkina Faso, parklands and the emphasis on F. albida in myths and religious practices are mostly found among ethnic groups of the western provinces who were geographically closest to Malian groups associated with F. albida parklands. These parklands are rare in the Gourmanché villages of eastern Burkina Faso.

    In northern Cameroon, from the Mandara mountains in the east to the Logone river in the west, F. albida parklands are most frequently encountered. However, Prosopis africana stands out as the most common species in some villages in the Bec de Canard region (Bernard, 1996). Interestingly, these stands are circumscribed in a land pocket occupied by the Musey ethnic group, whose largest area of settlement is in Chad and surrounded by the more spatially dominating Massa populations. This group had a strong warrior and hunter tradition with a well established reputation. They are known for having fought the neighbouring Massa and Fulani and their strong cultural identity persists today. Prosopis africana was actively maintained in parklands because it provided stakes for decorating tombs and recalling the war or hunting feats of old Musey warriors. Whereas the surrounding Massa villages emphasize livestock production, the Musey raise Logone horses used in combat and hunting. However, cultural and technological changes have made the practice of regenerating P. africana less significant among young segments of the Musey.

    Whereas the above-mentioned agricultural systems stand out because of the prominence of one or a few species, this is not always the case. In the Mafa region of Cameroon, where terraced agriculture is practised on the steep slopes of the northern Mandara mountains, parklands include several species, none of which is predominant. In this remote area farmers are relatively poor and self-sufficient. They maximize the use of trees, while minimizing their negative effect on crop production (Seignobos, 1982).

    Parklands (or ‘parcs arborés’ in French) are landscapes in which mature trees occur scattered in cultivated or recently fallowed fields. They are extremely variable, leading to continued discussion about appropriate terminology. They have been referred to as a vegetation type similar to ‘tree savannas’ but differ from these in that they are of specifically human origin, with the composition and density of their woody component manipulated in order to facilitate its use. Although occurring in other areas of the world, such as southern Africa and the Mediterranean, it is in the Sahel and Sudan zones of Africa that parklands are most widespread. No accurate figures exist at a regional level, but parklands constitute the predominant agroforestry system in semi-arid West Africa. They are most often characterized by the dominance of one or a few species. Faidherbia albida has received most attention in the literature because of its positive effect on soil fertility but it is likely that parklands dominated by Vitellaria paradoxa and Parkia biglobosa occupy the largest area.

    Their origin in human agricultural activities meant that their study was generally neglected by early plant ecologists and the first in-depth system descriptions were undertaken by geographers in the 1960s. Several approaches have been used to distinguish types of parklands. First, they have been categorized according to the intensity of human management as residual, selected, constructed, proto-arboricultural or planted parklands. This perspective is useful in revealing the various degrees of management, but parklands characterized by a dominant species often do not fit exclusively into one or other of these categories. With increasing management intensity and dissemination of species and technology, the development of a mosaic of parkland types adapted to various niches may be expected.

    Parklands have also been classified according to broad use categories. These systems have been established for the production of essential food complements, oils, wine, famine foods, wood, browse, crafts and for soil fertility restoration or a combination of these purposes.

    The age structure, composition, and density of parklands have also been analysed in terms of spatial variation. This corresponds in part to soil management practices, which decrease in intensity from compound fields through village fields to bush fields. Thus Faidherbia albida may be favoured on permanently cultivated village fields, while alternating fallow and cultivation cycles in bush fields may promote the development of floristically more diverse parklands with a Vitellaria paradoxa and Parkia biglobosa dominance. Tree size is correlated with the time during which fields have been cultivated. Thus, contrasts in size between parkland stands reveal land-use changes sometimes associated with demographic events. The growth rate of trees in parklands is generally higher than in uncultivated locations due to more favourable conditions.

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    Carbon dioxide and the uneasy interactions of trees and savannah grasses

    Savannahs are a mixture of trees and grasses often occurring as alternate states to closed forests. Savannah fires are frequent where grass productivity is high in the wet season. Fires help maintain grassy vegetation where the climate is suitable for ...

    Carbon dioxide and the uneasy interactions of trees and savannah grasses

    William J. Bond

    1Botany Department, University of Cape Town, Rondebosch 7701, South Africa

    Guy F. Midgley

    2Climate Change and Bioadaptation Programme, South African National Biodiversity Institute, Private Bag X7, Claremont, Cape Town 7735, South Africa


    Savannahs are a mixture of trees and grasses often occurring as alternate states to closed forests. Savannah fires are frequent where grass productivity is high in the wet season. Fires help maintain grassy vegetation where the climate is suitable for woodlands or forests. Saplings in savannahs are particularly vulnerable to topkill of above-ground biomass. Larger trees are more fire-resistant and suffer little damage when burnt. Recruitment to large mature tree size classes depends on sapling growth rates to fire-resistant sizes and the time between fires. Carbon dioxide (CO2) can influence the growth rate of juvenile plants, thereby affecting tree recruitment and the conversion of open savannahs to woodlands. Trees have increased in many savannahs throughout the world, whereas some humid savannahs are being invaded by forests. CO2 has been implicated in this woody increase but attribution to global drivers has been controversial where changes in grazing and fire have also occurred. We report on diverse tests of the magnitude of CO2 effects on both ancient and modern ecosystems with a particular focus on African savannahs. Large increases in trees of mesic savannahs in the region cannot easily be explained by land use change but are consistent with experimental and simulation studies of CO2 effects. Changes in arid savannahs seem less obviously linked to CO2 effects and may be driven more by overgrazing. Large-scale shifts in the tree–grass balance in the past and the future need to be better understood. They not only have major impacts on the ecology of grassy ecosystems but also on Earth–atmosphere linkages and the global carbon cycle in ways that are still being discovered.

    1. Introduction

    Concentrations of carbon dioxide (CO2) in the atmosphere have varied greatly over the history of terrestrial plants [1]. Major changes in CO2 in the geological record have been associated with large changes in vegetation. Over the last 65 million years (the Cenozoic) CO2 rose to greater than 1000 ppm in the Eocene (50 Ma) and forests spread to their greatest extent [2]. When CO2 fell in the Oligocene, forests retreated and grasslands began to spread. C4 grassland biomes became prominent from the late Miocene (approx. 8 Ma) [3] and their appearance has also been controversially associated with low CO2 [46]. During the Pleistocene, tropical grasslands expanded under low CO2 conditions of the glacials and contracted as forest expanded in the interglacials [4,7,8]. In historical times, anthropogenic increases in CO2 following industrialization have been proposed as a factor contributing to woody plant increases in grasslands (figure 1). Shifts between grasslands and forests are likely to have feedbacks on the atmosphere via diverse mechanisms [9,10]. Over geological time, changes in forested landcover may also have feedbacks to CO2 in the atmosphere by influencing rates of weathering, and therefore the terrestrial sink for CO2 [11]. The role of rising atmospheric CO2 in driving such shifts, and their global implications, are poorly studied in comparison with CO2 effects in temperate forest ecosystems. Here, we review studies on CO2 impacts on tree–grass interactions. CO2 effects cannot be studied in isolation from frequent disturbances by fire and grazing which are characteristic of savannahs. Consequently, we have selected studies where interactions between CO2 effects and the disturbance regime have been considered. Our main focus is on those grassy biomes where frequent fires are important in reducing tree cover.

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    Woody thickening over the last century near Queenstown, eastern Cape, South Africa. Acacia karroo is the most common tree species in this mesic savannah (approx. 750 mm mean annual precipitation). (a) 1925, (b) 1993, and (c) 2011. Note the large woody increase since the early 1990s. The original photograph was taken by the late IB Pole Evans (South African National Botanical Institute) and repeat photos courtesy of Timm Hoffman and James Puttick (Plant Conservation Unit, University of Cape Town).

    2. Fire and the distribution of grassy biomes

    In large regions of the world, direct climate control over vegetation is overridden by fire producing low biomass ecosystems where conditions are warm enough and wet enough to support forests [12]. C4 grasslands and savannahs cover greater than 15 per cent of the vegetated land surface, mostly at lower latitudes. These pyrophytic grassy biomes alternate with closed forests in seasonally wet tropical landscapes. They have been interpreted as alternative stable states which can switch on the order of decades from grassland to forest (or vice versa), depending on the interaction between plant growth rates and disturbance frequency and severity [1315]. Because CO2 influences plant growth rates, changes in atmospheric CO2 could be a significant factor influencing the proportions of forests and grasslands in a landscape. Changing CO2 can also influence tree cover within savannahs. Savannahs are a mixed community of trees with C4 grasses usually dominant in the herbaceous layer. The trees are compositionally and functionally distinct from forest trees being more tolerant of competition with grasses and of the frequent fires characteristic of seasonally wet savannahs. Trees recovering from disturbance may be particularly sensitive to changes in CO2. After a burn, light, water and nutrients are least likely to limit growth thus facilitating maximum CO2 responsiveness. Top-killed plants have to re-build their stems so there is a strong carbon sink [16]. Furthermore, frequent surface fires in savannahs select for woody plants with underground storage organs, or clonally spreading root systems, which promote rapid post-burn recovery [1719]. Post-burn resprouting of smaller size classes of savannah trees, whether using stored reserves or not, will produce large carbon sinks so that these plants should be particularly responsive to CO2 fertilization with less sink limitation of photosynthesis.

    CO2 can influence plants by altering water use efficiency, photosynthetic rates, light and nutrient use efficiency, and the relative performance of C3 and C4 photosynthesis [20]. The latter effect has been invoked as an explanation for the timing of the evolution of the C4 photosynthetic pathway with a drop in CO2 [4,21,22]. However, C4 photosynthesis requires high-light environments and high temperatures for maximum photosynthetic performance [23]. The implication is that C4 grasses were restricted to open, well-lit ecosystems and that the Late Miocene–Pliocene expansion of the grassy biomes depended on the retreat of forests. Palaeoecologists often attribute the contraction of forests to increasing aridity at low latitudes (or cold at high elevation or latitudes). However, bunch grasses (whether C3 or C4) are highly flammable where decomposition is slow and, given sufficient water to produce continuous fuels, support very frequent fires. Fires are a major factor influencing the distribution of forests in the modern world, especially where C4 grasses predominate in low to mid latitudes, but also in flammable C3 grasslands in the steppes of Asia [14,24] (figure 2).

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    Global fire activity from 2001 to 2006 from MODIS active fire counts. Tropical and sub-tropical grasslands and savannahs account for the largest burnt area over the period but grasslands in the steppes of Central Asia also burnt frequently. Reproduced with permission from Chuvieco et al. [24]. Copyright © Wiley.

    Changing atmospheric CO2 can influence woody plant expansion into both arid grasslands and more humid flammable grasslands but by different hypothetical mechanisms. An increase in water use efficiency under elevated CO2 owing to reduced transpiration has been proposed as a key mechanism favouring woody plant thickening in semi-arid savannahs [25]. This has the indirect effect of increasing soil moisture supply which may favour tree seedling recruitment and subsequent growth in competition with grasses. CO2 enrichment experiments in grasslands have commonly shown an increase in soil moisture consistent with this mechanism [26,27].

    In more humid grasslands, where fires are frequent, Bond & Midgley [28] suggested that direct CO2 effects on plant growth would favour an increase in tree cover. They noted that the rates of recovery of plants after a disturbance should vary inversely with the proportion of carbon allocated to support structures. This is because plant relative growth rates (RGRs) increase in proportion to leaf area ratio, the ratio of photosynthetic leaf area to total plant mass [29]. Woody plants which have to re-build large non-photosynthetic stems after a disturbance will have lower RGR than herbaceous plants such as grasses. Tree seedlings and saplings surviving fire and resprouting after a burn in a flammable grassland will soon face a repeated fire, as grasses accumulate fuel rapidly. Woody plants can increase the chance of escaping topkill by building below-ground reserves of carbon and nutrients that promote rapid regrowth. However, sprouting from underground storage organs comes at a cost of initially slower growth rates in seedlings that have to build below-ground reserves resulting in poorer establishment and reduced ability to colonize new areas [30]. Within savannahs, growth rates of trees to fire-resistant sizes may be particularly sensitive to CO2 as the replacement of stem tissue and investment in storage organs are easier to make under elevated CO2 [20]. Thus, regardless of whether grasses are C3 or C4, increasing CO2 should favour trees more than grasses in flammable (and other disturbed) ecosystems.

    3. Tests of the fire–carbon dioxide hypothesis

    Savannah trees are a very useful model system for testing the effect of CO2 on plants recovering from disturbance. Seedlings and smaller plants not only compete with grasses but are also subjected to frequent grass-fuelled fires. Fires topkill the stems of smaller plants. Given sufficiently rapid stem growth rates, or an unusually long interval between fires, stems grow to fire-resistant sizes. Mortality from fires is very low in established plants and small plants can persist for decades suffering repeated topkill by fires [17,31,32]. Release from the flame zone (the ‘fire trap’) depends both on sapling growth rates and the rare periods between fires that are long enough for a small plant to grow to fire-resistant size. Unlike CO2 effects on closed forests, the impact of a change in growth rate on savannah trees has clearly quantifiable demographic effects on the transition to large trees. Demographic changes in large tree densities, in turn, have large ecosystem impacts on fluxes of water, nutrients, carbon and even, possibly, rates of weathering [11]. This chain of CO2 effects from increased growth rates to changes in tree cover has been simulated for South African savannahs [33]. A detailed demographic model of fire–savannah tree demography [31] showed great sensitivity of tree cover to changes in sapling growth rates. CO2 effects on sapling growth rates, simulated with a Dynamic Global Vegetation Model (DGVM) [34], and coupled to the demographic model, indicated a striking effect of changes in CO2 since the last glacial on tree cover in the region (figure 3). Simulations predicted that sapling growth rates would be so slow at Last Glacial Maximum (LGM) CO2 levels that trees would have been eliminated from flammable savannahs. The largest subsequent CO2-induced change to tree cover was predicted from pre-industrial to modern CO2 levels, especially over the twentieth century [33].

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    Simulated change in tree growth and sapling response for African savannah trees. (a) Sapling growth rates to fire-resistant size (2.5–3.5 m) at varying CO2 levels simulated with a DGVM [34]. (b) Simulated median tree densities as a result of these growth rates simulated with a detailed demographic model of savannah trees burnt by fires [31]. Adapted from Bond et al. [33].

    (a) Glasshouse experiments

    The simulated response to CO2 reported in [33] was tested for common savannah plants grown experimentally in open-top chambers in a greenhouse [35] (B. S. Kgope, G. F. Midgley & W. J. Bond 2011, unpublished data). Savannah tree seedlings were grown under well-watered conditions with nutrients supplied for a full growing season, after which they were cut to simulate fire in the subsequent dry season. After a second full growing season, the plants were harvested. Results were consistent with the DGVM simulations (figure 4) with very slow growth at 180 ppm (LGM), slightly greater at pre-industrial, and then very large increases at 360 ppm. Plants grown at elevated CO2 did not grow well probably because of pot size constraints [35]. In a second set of experiments, the soil volume was increased to 200 l and a broad-leaved savannah tree, Terminalia sericea, was included (figure 5) (B. S. Kgope, G. F. Midgley & W. J. Bond 2011, unpublished data). Kgope et al. [35] reported large changes in root size and starch concentration within the roots, especially from 260 ppm (pre-industrial) to 360 (‘ambient’) CO2 treatments (figure 6). The impact on sapling growth rates, scaled from total biomass differences across CO2 treatments, exceeds those used in the simulation study (figure 3). These very large effects imply that, during the last century, increasing CO2 has fuelled changes in savannah tree growth helping to create ‘super-seedlings’. Relative to a century ago, young trees have faster growth, massive root systems with larger starch reserves and a greatly enhanced capacity for resprouting after injury. In the Acacia seedlings, there is the added bonus of larger spines and higher tannin content in the leaves providing better protection against seedling herbivory.

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    Acacia karroo growth response to CO2 treatments. Plants were grown for a year, cut to simulate fire (shoot) and harvested after a second growing season (resprout, root) [35].

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    Terminalia sericea growth response to CO2 treatments. Plants were grown for a year, cut to simulate fire (shoot) and harvested after a second growing season (resprout, root). Adapted from B. S. Kgope, G. F. Midgley & W. J. Bond 2011, unpublished data.

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    Acacia karroo root responses to a gradient of CO2 treatments. The scale on the left of each plant is a metre long. Plants were grown for a year, cut to simulate fire and harvested after a second growing season. Adapted from B. S. Kgope, G. F. Midgley & W. J. Bond 2011, unpublished data.

    Responses of savannah trees to CO2 have been studied from below to above ambient levels in North American Prosopis species [36,37], and to elevated CO2 for Australian [38,39] and South American woody species [40]. As has been found for plants from other ecosystems, responses vary among species with faster growing species tending to respond more strongly [37,39]. Hoffman et al. [40] also showed that CO2 responses were sensitive to nutrient supply which suggests that CO2 effects may be muted on nutrient poor soils.

    (b) Field experiments

    There are no field studies manipulating CO2 in tree–grass ecosystems in Africa. However, there are several long-term burning experiments in which the same fire treatments have been applied for several decades. If the CO2 effects on tree growth observed for African species were carried over to the field setting, then there should be CO2-enhanced seedling recruitment and more frequent transitions from smaller to larger tree size classes. During the twentieth century, CO2 levels rose from 296 ppm in 1900 to 310 ppm in 1950, 330 ppm in 1975 and 376 ppm by 2003 [41]. The physiological responses to CO2 are steepest from sub-ambient to ambient CO2 concentrations relative to above ambient [42]. If physiological responses fed into changes in the ecology of woody plants, then long-term burning experiments would be expected to show changes in tree responses in the early versus later years of the experiments. Buitenwerf et al. [43] have analysed the results of two long-term burning studies in South Africa testing for temporal trends. The longest running experiments are at Kruger National Park [4446]. Figure 7 shows trends in densities for a semi-arid and mesic savannah for censuses conducted in the 1950s, 1970s, 1990s and 2000s. The 1970s and 2000s censuses used permanently marked grids providing spatially explicit records of changes in tree densities and sizes. Data for the most common tree and shrub species in the mesic savannah are shown in figure 8. There were negligible temporal trends in woody plant densities in the semi-arid savannah (figure 7). In the mesic savannah, there were negligible changes in the first two decades (1950s–1970s) but very large increases from the 1970s to the 2000s. Both the most common tree and shrub species increased in density and both showed an increase in recruitment into the larger size classes (figure 8; see also Higgins et al. [45]). Terminalia sericea, the tree species, responded strongly to CO2 in glasshouse experiments (figure 5) so that its increase in the field is consistent with CO2 effects.

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    Changes in woody plant densities in a long-term burning experiment in the Kruger National Park, South Africa. Results from (a) a semi-arid savannah (mean annual precipitation (MAP) approx. 500 mm) and (b) a mesic savannah (MAP approx. 750 mm). Adapted from Buitenwerf et al. [43].

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    Changes in density across size classes of the (a) most common tree and (b) shrub species in a mesic savannah from the 1970s to the 2000s. Data is from identical areas sampled 30 years apart and subjected to the same fire treatments since the 1950s in the experimental burn plots, Kruger National Park, South Africa. Adapted from Buitenwerf et al. [43].

    A second fire experiment has been maintained for three decades near the southern limit of savannahs in South Africa (latitude 32° S). Figure 9 shows changes in Acacia karroo, the most common savannah tree species in this system. Changes in tree densities were similar to the mesic savannah of Kruger National Park with a large increase in both densities and size over the second half of the experiment from the mid-1990s to 2010. Interpretation of these results is complicated by the removal of trees before the experiment started [43]. However, large increases in this tree species have been reported elsewhere in this region, especially in higher rainfall areas, over the last few decades (figure 1). Acacia karroo was very responsive to CO2 increases from pre-industrial to current values in glasshouse experiments (figures (figures44 and and6;6; [35]) so that the field observations are consistent with experimentally observed CO2 fertilization effects.

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    Changes in Acacia karroo populations subjected to the same fire treatments for 30 years. Eastern Cape, South Africa. Adapted from Buitenwerf et al. [43].

    There are a number of long-term burning experiments in Africa and elsewhere. Where treatments have been maintained and sufficient censuses are available, these would be worth exploring for temporal trends as predicted from tree responses to increasing CO2. For example, Briggs et al. [47] reported longitudinal trends of woody plant responses to long-maintained fire regimes in tall grass prairies in Kansas. In this system, different catchments have been burnt annually, at moderate frequency or left unburnt since 1980. Bison were introduced into some of the treatment areas in the 1990s. In the ungrazed treatments, two woody species showed large increases in the intermediate fire frequencies but no change under annual burning. The greatest increases were in catchments that were both grazed and burnt. Briggs et al. [47] noted that the increase in these woody species was paradoxical. Historical accounts reported the area to be tree-less. Fire frequencies and bison densities several centuries ago were thought to be similar to those applied over the experimental period. Thus, the increase in woody plants in all but the most extreme annual burning treatment is a departure from the historical condition. They attributed the woody increase to habitat fragmentation and changes in seed dispersal and did not consider CO2 fertilization as a possible factor.

    4. Carbon dioxide and woody plant cover increase

    Savannah ecosystems are likely to be particularly sensitive to CO2 effects because small increases in stem growth rates can have very large effects on the number and biomass of trees. The potential for change from an open to a wooded savannah is greatest in higher rainfall savannahs [48]. Once trees have escaped the firetrap and a woodland patch has developed, CO2 effects on further changes in woodland structure are likely to greatly diminish. Many studies from around the world have reported woody thickening in savannahs (see list in Archer et al. [49]). Attribution to global drivers has been controversial since changes in grazing and fire are a common and well-studied alternative explanation [12,45,46,50]. CO2 is unlikely to be implicated in all cases of woody thickening. We suggest that the most probable environmental settings for CO2 effects on woody increase are higher rainfall savannahs which burn frequently. These contain fast growing woody plants that depend on stored carbon for resprouting so that the potential for sink-driven increases in carbon gain are particularly strong. In more arid savannahs, CO2 effects may be small relative to the effects of other drivers [50] (but see Morgan et al. [51] for large increase in a shrub in short grass steppes under elevated CO2). CO2 effects are least likely where edaphic constraints, such as shallow rooting depth, limit tree growth [52]. Low soil nutrients may also reduce CO2 responses in nutrient poor savannahs. Despite these uncertainties, there is clear evidence for CO2 fertilization contributing to woody thickening over the past few decades in some savannahs. The implication is that land users will have to work much harder than in the past to maintain open grassy ecosystems.

    5. Forest expansion into grasslands

    Our discussion thus far has been on CO2 effects on tree cover within savannahs. There is growing evidence that elevated CO2 may also be promoting conversion of savannahs to closed forests. Forests are distinct from savannah woodlands in their species composition, structure and function [19]. They are too shady to support a C4 grass understorey so that grass-fuelled fires seldom spread beyond the forest edge. Forest trees are far more sensitive to fire than savannah trees so that the two ecosystems have been characterized as pyrophytic and pyrophobic alternative ecosystem states [1315]. Tropical forests have changed in recent decades with increased net primary productivity, and increased tree growth, recruitment, mortality rates and forest biomass [53]. There has been much debate over the possible cause of these changes, but a growing consensus that increasing atmospheric CO2 is the most probable cause [53,54].

    Tropical forests are also expanding into grassy biomes in many regions where there is no human deforestation. Expansion is the most conspicuous in landscapes with mosaics of grasslands and forests. Gallery forests (along riparian areas) and other forest patches have expanded into flammable grassy ecosystems in South America [55], North America [56,57], Australia [58,59] and different parts of Africa [6063]. Forest expansion seems to be the most common in wetter climates [59]. In semi-arid central Australia, for example, mosaics of closed ‘mulga’ (Acacia aneura) woodland and open spinifex (Triodia spp.) grasslands have remained fairly stable for decades [59]. Thus, forest invasion of grassland may be similar to tree increases within savannahs, with the greatest changes in wetter climates and the least in more arid regions.

    The causes of forest expansion are also a subject of vigorous debate over the relative importance of changes in fire regimes versus global drivers, especially CO2 increase. There are far fewer studies on forest edges than tropical forest interiors. Changes in the forest boundaries have mostly been studied by historical aerial photography and not from plot data. Despite the difficulties, studies of forest expansion into flammable grassy ecosystems provide key insights into the potential for major landcover changes. If savannahs were to be replaced by forests on large continental scales, the repercussions would be far greater than a shift from a less to a more wooded savannah.

    Experimental and modelling studies have shown that large areas of savannahs could support forests in the absence of fire. Forest expansion into savannahs has commonly been attributed to reduced fire activity near forest margins. Global drivers are implicated where forests have expanded despite stable or even more frequent fires. In northern Australia, for example, Bowman et al. [59] noted that forest expansion has occurred from small rainforest patches even though there have been landscape-wide increases in the size and frequency of fires. Annual rainfall and the number of rain days have increased in the region but Bowman et al. [59] attribute the forest expansion primarily to the physiological effects of CO2 on forest trees exposed to fire at the forest edge. In South Africa, Wigley et al. [63] compared forest expansion over a near 70 year period in three adjacent areas with sharply contrasting fire and herbivory regimes. This study included a nature reserve, effectively a natural ‘control’, with an extant African megafauna, including elephants and other large browsers and grazers, and frequent fire. At the other extreme was a densely populated subsistence farming area where trees are used by people and livestock for building and fuels. Forest expansion was greatest in the nature reserve ‘control’ increasing from 14 per cent in 1937 to 58 per cent by 2004. The subsistence farming area showed the smallest increase (6–25%). However, the striking fact is that scrub forests increased in all land use types regardless of the intensity and types of disturbance. This strongly suggests a global driver tipping the balance towards trees. The most probably driver is increasing atmospheric CO2 because no significant trends in temperature and rainfall were identified in the region [63].

    Recent studies suggest that processes operating at forest–savannah edges are similar to these within savannahs and are likely to be similarly sensitive to changing CO2 [15,64]. Working in Brazil, Hoffmann et al. [64] studied the dynamics of forest recovery after fire. On the rare occasions where fire spreads into a forest from an adjacent savannah, smaller forest trees are top-killed by fire while larger trees are much more fire-resistant. As in savannahs, most top-killed smaller trees resprout and mortality rates are low, at least after a single fire. The ability to repair the forest edge is a function of growth rates to fire-resistant size versus the fire return time. If growth rates are slow, then the forest edge will retreat and grasses will invade. If growth rates are fast, then the forest edge may expand into the grassland, despite the fires. Forest trees have denser canopies than savannah trees and can suppress fires by shading out the grasses [65]. Forest trees had thinner bark than savannah trees and were therefore more vulnerable to topkill for a given size. It is interesting to note that forest trees at this forest edge site had much thicker bark and were far more resilient to burning than trees in the interior of Amazon forests [64,66]. Because of their thinner bark, forest trees would need at least twice as long as savannah trees to reach fire-resistant sizes making them far more sensitive to burning.

    This analysis of the mechanistic responses of forest trees to grassland fires is very similar to the fire trap models for savannah tree populations [15,31,64]. It suggests the same sensitivity of the process to atmospheric CO2. After a fire, resprouting forest trees would be growing in well-watered, high-light and high-nutrient environments and carbon gain will be strongly sink-driven as trees re-build their stems. CO2 effects are likely to be the most pronounced under these conditions. We know of neither any glasshouse experiments testing CO2 responsiveness of forest margin trees bordering on savannahs nor any field studies of CO2 responsiveness of resprouting tropical forest margins. The closest analogue seems to be a Florida scrub oak system exposed to elevated CO2 after the ecosystem was burnt. The scrub oak ecosystem showed a 67 per cent increase in above-ground stem biomass under elevated CO2, with the response sustained over the 11 years that the experiment has been maintained [67].

    Forest margins have been avoided as study sites being neither fish nor fowl. However, local processes at these margins may determine large-scale patterns of forest–savannah distribution. A continent-wide study of tree cover in Africa, for example, found that patterns of forest versus savannah distribution over a rainfall gradient were invariant across scales implying that the distribution of the two vegetation types is determined by local processes [15]. Hoffmann et al. [64] make the same point. The implication is that changes in the processes operating at a forest–grassland boundary a few metres wide are the frontline for wholesale biome switches. If this is the case, then changes in atmospheric CO2, interacting with fire, could have major effects on the distribution of forest in the tropics with forest retreats in glacial periods and expansion in interglacial periods. Focused CO2 studies on boundary dynamics will go a long way in showing the sensitivity of forest boundaries to future shifts in CO2.

    6. Palaeoecology of tree–grass interactions

    Studies of CO2 and other drivers of tree–grass interactions in contemporary ecosystems are complicated by their short duration and multiple interacting factors. Current woody cover may reflect past events, such as the rinderpest epidemic in Africa [68] or severe droughts [69,70] rather than current factors [71]. It has also been argued that CO2 responses reflect short-term changes and that other factors, such as nitrogen supply, will become limiting as ecosystems equilibrate [72,73]. One way out of the morass is to explore the palaeorecord. The last glacial period was not only colder but also had much lower atmospheric CO2 (185–200 ppm) than interglacial periods (280–300 ppm). If CO2 effects were indeed trivial over longer periods of time, then the palaeorecord of biome shifts should be predictable from palaeoclimates alone. That this was not the case began to emerge from attempts to model the distribution of last glacial vegetation [7476]. Prentice & Harrison [77,78] have recently summarized diverse lines of evidence showing that palaeoclimate is not a good predictor of biome shifts and that CO2 effects are fundamental to explaining vegetation change from glacial to interglacial periods.

    Firstly, they note that changes in δ13C in the shells of marine sub-fossil foraminifera indicate a reduction in terrestrial C in the last glacial on the order of 300–750 Pg with recent estimates suggesting a loss of approximately 600 Pg of C. Reductions of terrestrial carbon of this magnitude cannot be predicted from palaeoclimate models. Indeed some models simulate a gain in carbon largely because low temperatures in glacial periods would have slowed decomposition of litter thereby causing an increase in C stored in soils. Only models that incorporate both climate change and the physiological effects of low CO2 produce vegetation change of the magnitude necessary to explain the estimated reduction in terrestrial C. Palaeo-reconstructions of biome distribution can also be used to test CO2 effects. Attempts to simulate biome distributions from General Circulation Model (GCM) reconstructions of last glacial climates have failed to produce forest reduction of the magnitude necessary to explain the pollen evidence, especially in the tropics where savannahs expanded and forests contracted. Biome models incorporating the physiological effects of CO2 simulate far greater forest reduction consistent with the palaeodata [77,78]. Finally, recent studies have used inverse modelling to test the kinds of climates needed to produce the biome distributions reconstructed from pollen data. Inverse modelling failed to reconstruct climates matching those from GCM modelling unless physiological CO2 effects were also incorporated [77].

    The evidence from the palaeodata clearly supports a major role for CO2 in large-scale biome shifts from low CO2 glacial periods to higher CO2 in the interglacials (though the role of fire–CO2 interactions over these timescales has been little studied [33]). Palaeoecologists have been aware of the potential importance of CO2 for some time but have been slow to interpret atmospheric CO2 as a contributor to past vegetation change. Wooller et al. [7] suggested that fire, together with low CO2, helped drive the retreat of forests and expansion of flammable grasslands in equatorial Africa. Damste et al. [8] used diverse proxies to test for low CO2 and fire synergies in their analysis of forest–savannah interactions in Africa. Studies of the palaeoecology of fire have also increased greatly over the past decade allowing comparisons of glacial and interglacial fire activity. Though data for tropical fire regimes (savannahs, C4 grasslands) are still rare compared with woody-fuelled ecosystems, long temporal record of fires is accumulating and is beginning to contribute major insights into the determinants of global fire regimes [79]. Coupled with increasing availability of data on past vegetation and past fires are new developments in global fire modelling [80] and DGVMs [78,81]. These new tools should help greatly in testing the relative importance of CO2, fire and climate on forest–savannah distributions over far longer periods than are feasible in contemporary experimental studies.

    7. Biogeochemistry of savannahs versus forests

    Changes in tree cover in savannahs or switches from savannah to forest could have wide-ranging consequences, including feedbacks to the Earth–atmosphere system [9]. Savannahs cover more than 17 million km2 and account for about 21–23% of global productivity [51]. With their large spatial extent, even small increases in woody biomass would represent a substantial carbon sink [49,82,83]. Savannahs differ from forests in their energy budget and hydrology [84]. They typically have a higher albedo than forests, partly because they retain dead, reflective leaves over the dry season. Trees generally transpire more water than grasses because of their more extensive root system, greater leaf area, and greater canopy roughness. Consequently, large-scale change in tree cover, or biome switches to forest, would have feedbacks to the regional climate [84]. Extensive grass-fuelled fires also have feedbacks to climate [10]. Black aerosols alter energy budgets and reduce the size of cloud droplets causing precipitation to be less frequent but more intense [8587]. Savannah fires are also important sources of NOx which promote increased tropospheric ozone formation and transport reactive N to potentially N-limited ecosystems [88].

    A recent estimate of global vegetation fire emissions from 1997 to 2009 (when MODIS satellite imagery became available) estimated average emissions of 2.0 Pg C per year [89]. Grasslands and savannahs accounted for 44 per cent of this, with (mostly tropical) woodlands an additional 16 per cent. Thus, the tropical grassy systems accounted for about 60 per cent of global fire emissions. African savannah fires alone contributed approximately 1 Pg C per year. Increases in tree cover could profoundly alter these global burning patterns. An analysis of fire spread in Africa has shown threshold effects with fire activity declining sharply once tree cover exceeds about 40 per cent [90]. If large-scale tree thickening occurs across the continent and fires cease to burn, then there are likely to be multiple effects on the Earth–atmosphere system [9,91].

    Over longer timescales (million of years), Pagani et al. [11] have suggested that reductions in tree cover at low atmospheric CO2 may have limited further CO2 decline. These authors argue that silicate weathering of rock, the major sink for atmospheric CO2, would have decreased with loss of forest cover and their replacement by C4 grasslands and savannahs. This negative feedback mechanism opposing higher rates of silicate chemical weathering may explain why CO2 did not decline further during warmer periods (e.g. from the Early to Mid-Miocene) when high weathering activity should have promoted further CO2 reduction.

    8. Conclusions

    Savannahs span from climates too arid to support trees to climates so productive that they switch to forest within a decade or two of fire suppression. Within this diverse climate setting, the ecology of trees in savannahs is frustratingly complex with multiple factors interacting to determine changes in tree populations. For these reasons, it is difficult to predict the future of savannahs, or to identify the key drivers of changes in woody cover in the immediate past. There is enough evidence to indicate that elevated CO2 is increasingly tipping the balance in savannahs in favour of trees, especially in the more open, frequently burnt savannahs. But, there are still far too few data to predict which parts of this complex biome are most and which least likely to respond to increasing CO2.

    Contemplation of the more distant past can help provide perspective. C4 grasses evolved in a low CO2 world in the Oligocene more than 25 million years ago and first assembled as a major biome, the world's youngest, in the low CO2 atmospheres from about eight million years ago. Within the next hundred years, savannah plant species are therefore likely to experience CO2 levels outside their entire evolutionary history. Given the decreasing physiological advantages of C4 grasses at high CO2, and the ecological and palaeoecological studies reviewed above, we suggest that it is not unreasonable to postulate that the rapid rise of the savannah biome from the Late Miocene may be matched by an abrupt decline owing to anthropogenic impacts on the carbon cycle. The likelihood of large-scale loss of savannahs (other than by land clearing), and the repercussions for local land users, biodiversity and Earth–atmosphere feedbacks deserve far more attention.


    Many people have contributed to the studies reported here from simulation modellers and physiologists to field staff who maintained field experiments and burning treatments for decades. We particularly thank Ian Woodward, Steve Higgins, Winston Trollope, Barney Kgope, Ben Wigley and Rob Buitenwerf for their diverse contributions. For three decades, Andre Potgieter and his staff maintained the Kruger burn plots and Winston Trollope and Wellington Shabangu maintained the eastern Cape burn plots. We are very grateful to the Royal Society for hosting the Research Fellows International Scientific Seminar on Atmospheric CO2 and Green Evolution at the Kavli Centre and to David Beerling for inviting us to participate and inspiring us to take a broad perspective on CO2 and plant ecology and evolution.


    Carbon dioxide and the uneasy interactions of trees and savannah grasses

    Source : www.ncbi.nlm.nih.gov

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