The Impact of Expanded Biofuel Production on Living Nature

5.1 Introduction

In 2007, there was a lively discussion about a plan to replace the Mabira Forest Reserve in Uganda with a sugar cane plantation for the production of the trans­port biofuel ethanol. This forest reserve is home to almost 300 bird species (among which is the very rare Nahan’s Francolin) and supports 75 endemic species. In Oc­tober 2007, the Ugandan government announced that the plan had been scrapped, because the income from conserving the Mabira Forest would dwarf the profits from bioethanol production (Williams 2007). In the case of the Mabira Forest, much of this income is derived from ecotourism (Williams 2007), with additional revenues coming from harvesting timber, making charcoal and the collection of fuelwood (Naidoo and Adamowicz 2005). There may also be other sources of income from such forests, such as the collection of food, ornamental plants and organisms that have medicinal value (Brown and Rosendo 2000; Shanley and Luz 2003; Brennan etal. 2005;Mutimukuruetal. 2006). In being a target for ecotourism and providing natural resources, living nature may be said to provide ecosystem services that have monetary value.

Decisions about clearing nature to make way for biofuel production may also turn out differently. For instance, in 2007, the replacement of the high ecological value Tanoe Swamp Forest (Ivory Coast) by an oil palm plantation for biofuel production was also started (Sielhorst et al. 2008).

Direct replacement of nature by growing biofuels is only a part of the conse­quences of the expansion of biofuel production. There are also indirect effects of this expansion. These follow from the relative inelasticity of the demand for food (von Braun 2007; Searchinger et al. 2008). When cropping for biofuels replaces cropping for food or feed, food or feed crops largely have to be grown somewhere else. When the expansion of cropping for biofuel production is small, it may be pos­sible that the extra production of food and feed can be accommodated on existing agricultural soils, due to increasing productivity of agriculture. But when there is a fast and major expansion, this is not possible, and food and feed production may

L. Reijnders, M. A.J. Huibregts, Biofuls for Road Transport © Springer 2009

have to expand in areas that were so far left to nature (Searchinger et al. 2008). When the expansion of biomass for biofuel production is on highly productive soils, the effect can be relatively large, as it may well be that part of the expansion of food or feed production has to take place on soils with lower productivity, which will lead to relatively large land claims.

Replacement of living nature by agriculture directly or indirectly related to the expansion of biofuel production is a significant matter in the current debate about transport biofuels. And financial considerations are important to the outcome of the conflicts between nature and (agri)culture. However, financial interests are not the only matters that are at stake. It has been argued that natural species have an in­trinsic value, which requires protection against extinction. This type of argument has led to laws such as the (US) Endangered Species Act, which aims at protec­tion of endangered species. Secondly, ecosystems provide non-monetary ecosystem services to humankind conducive to a benign biological, chemical and physical en­vironment and to socio-cultural fulfilment (Daily 1997; Moberg and Folke 1999; Daily 2000; Batabyal et al. 2003; D^az et al. 2006; Brauman et al. 2007; Marrs et al. 2007; Wallace 2007). The non-monetary ecosystem services include beneficial im­pacts on water quality and quantity, climate, soil retention, pest and disease control and pollination. A more extended list is in Table 5.1.

Table 5.1 Non-monetary ecosystem services to mankind (Daily 1997; Diaz et al. 2006; Lelieveld et al. 2008)

Non-monetary ecosystem service to mankind

— Cleansing of air and water

— Contribution to preservation of soil fertility and stability

— Regulation of water quantity and quality available to humans, crops, animal husbandry and domestic animals

— Pollination of plants important to humans

— Pest and disease control in agriculture

— Resistance to invasive organisms that have negative impacts

— Climate regulation

— Protection against natural hazards (floods, fires, storms)

— Contribution to productivity and stability of plant production important to humans

— Prevention of leakage of nutrients and metals from soil to surface and ground water

Ecosystem services have been linked to biodiversity (Daily 1997, 2000), under­stood here as the diversity of species present in an ecosystem. There are a large num­ber of empirical studies that underpin this link (Salonius 1981; Naeem et al. 1995; Walker 1995; vanderHeijdenetal. 1999; Schlapfer and Schmid 1999; Duarte 2000; Emmerson et al. 2001; Engelhardt and Ritchie 2001; Hector et al. 2001; Lyons and Schwarz 2001; Loreau et al. 2001; Tilman et al. 2001b; Duffy 2002; Emmerling et al. 2002; Symstadet al. 2003; Tilman etal. 1996; Armsworth et al. 2004; Heems — bergen et al. 2004; Reusch et al. 2005; Balvanera et al. 2006; Cardinale et al. 2006; Worm et al. 2006; Brussaard et al. 2007; D^az et al. 2007; Fargione et al. 2007;

Turner et al. 2007; Flombaum and Sala 2008; Fornara and Tilman 2008; Ptacnik et al. 2008; Weigelt et al. 2008). From the studies done so far, it appears that there may be considerable differences between species in terms of what they contribute to ecosystem services (Cardinale et al. 2006; Jordan et al. 2006). On one hand, there are ‘keystone’ species that appear to have a large impact on such services. Keystone species have key functions in ecosystems. For instance, in sea grass communities, there are engineering species which, by changing the environment, facilitate the presence of species that would otherwise be absent (Duarte 2000). In arid envi­ronments, nurse plants such as Cercidium microphyllum and Carnegiea gigantea (saguaro cactus) have been identified that promote the establishment and survival of other species, including a variety of trees and shrubs (Withgott 2000; Drezner 2006,

2007) . And in the US Great Basin, sagebrush serves as a nurse plant for pinyon pine (Withgott 2000).

When keystone species disappear, the loss of ecosystem services may be dispro — portional. On the other hand, there are species which are in one or more respects rather similar to others in what they do in ecosystems. The loss of such a species may give rise to a less-than-proportionate loss of ecosystem services. The latter is reflected in studies that suggest that halving the number of plant species, on average, leads to a reduction in primary production of about 10-20% (Tilman et al. 1996). Still, there is also the possibility that cumulative biotic changes which at first ap­pear to have little effect may give rise to a sudden collapse of ecosystem services (Scheffer et al. 2001; Folke et al. 2004; Balmford and Bond 2005).

Such a collapse and the disproportionate effect of the loss of keystone species ex­emplify the possibility that the relationship between biodiversity and ecosystem ser­vices may show non-linearities (Lovelock 1988; Scheffer et al. 2001; Strange 2007). There may also be other causes for non-linearity. For instance, it has been found that there may be synergistic relations between invasive species (Grosholz 2005). And there may be interactions between losses of biodiversity and other human interven­tions that may give rise to non-linear effects. Malhi et al. (2008) discussed such a possibility in the context of deforestation and fire use in Amazonia. Here, there is a synergism between forest fragmentation and fire. Once burnt, a forest becomes more vulnerable to further burns and loses many primary forest species. Malhi et al. (2008) suggest that a tipping point may be reached when gasses establish in the forest understory, providing a source of fuel for repeated burns.

Monetary valuation of non-monetary ecosystem services is inherently problem­atic. One cannot buy them on markets. Provided that nature is there, they are freely provided. Without the ecosystem services of living nature, we would not even exist. The latter may be argued to suggest an infinite value, the former a zero value. And there have also been estimates in between (Costanza et al. 1997, 2007; Sukhdev

2008) . Given the problematical character of monetary valuation of non-monetary services, the following overview considers non-monetary ecosystem services with­out attributing monetary values.

In Chap. 2, it was pointed out that increased yields per hectare, partly linked to intensification of agriculture, are expected to contribute to the displacement of fossil fuels by biofuels, but that large-scale displacement of fossil fuels by biofuels requires large areas of land. Current policy targets for 2020 would require the use of between 55-166 million hectares (Mha) for biofuel cropping (Renewable Fuels Agency 2008). One option is to use surplus and degraded or abandoned and fallow agricultural land for this purpose (e. g. Hoogwijk et al. 2003). An estimate suggest­ing that by 2050, up to 300 EJyear~1 of liquid biofuels can be produced worldwide indeed assumes that 80% of the land area needed for that purpose will be aban­doned land (de Vries et al. 2007). In this case, there will probably be a large effect on biodiversity (Huston and Marland 2003; Marland and Obersteiner 2008), and biodiversity on abandoned and fallow land is linked to ecosystem services (Borner et al. 2007; Williams et al. 2008). It is likely that abandoned and fallow land will have a lower productivity than good-quality land. As a result, when fixed amounts of transport biofuels have to be produced, as mandated under current regulations in the USA, Canada, Brazil, India and the European Union, larger areas of land will be needed for biofuel production.

It is likely that abandoned and fallow land rather often harbours substantial biodi­versity (vanNoordwijk 2002; Zechmeister et al. 2003; Karlowski 2006; Bowen et al. 2007; Royal Society 2008). Especially after long periods of abandonment, biodiver­sity may be much increased (Fournier and Planchon 1998; van Noordwijk 2002; Williams et al. 2008) and may approach biodiversity in undisturbed ecosystems. However, there are also abandoned and fallow lands with relatively low biodiversity. Examples are the Imperata cylindrica and Saccharum spontaneum dominated grass­lands in formerly forested areas (Hooper et al. 2005; Germer and Sauerborn 2007). Though there are parts of such grasslands of great ecological importance (Peet et al. 1999), often they are not (MacDonald 2004). There are hundreds of mega-hectares of Imperata cylindrica grasslands, mainly in Africa and Asia (MacDonald 2004). In Southeastern Asia, these grasslands cover an estimated 25-35 Mha (Garrity et al. 1996; Otsamo 2000). Such grasslands are currently used for feeding livestock and thatching material (MacDonald 2004) but can also be used for biofuel production by harvesting the grasses as lignocellulosic feedstock for biofuel production or by the cultivation of, for example, short rotation woody crops that may serve as feedstock. In practice so far, the use of abandoned agricultural land for biofuel production has been very limited. For instance, expansion of Brazilian sugar cane production for the biofuel ethanol has largely been in the Cerrado region, a hotspot for biodiversity (Klink and Machado 2005; Koh 2007), and it has been suggested that further expan­sion may mainly take place on current pastureland (Goldemberg 2008; Goldemberg et al. 2008). Similarly, in Malaysia and Indonesia, there has been large-scale con­version of tropical forest into plantations that produce palm oil, notwithstanding the presence of large areas of degraded land in these countries (Germer and Sauerborn 2007).

There have been earlier proposals and attempts to exploit Imperata cylindrica grasslands for the production of wood and lignocellulosic feedstocks for the pulp and paper industry (Potter 1996; Lamb 1998; Otsamo 2000). These have met with some success (Marjokorpi and Otsamo 2006). In Southeast Asia, an estimated 2 Mha of former Imperata cylindrica grassland is now converted into Acacia mangium plantations, serving the supply of wood and lignocellulosic feedstock for the pulp and paper industry (Yamashita et al. 2008). However, attempts to convert Imperata cylindrica grasslands has had limited success because of limited support by local people who made use of those grasslands, e. g. for keeping livestock, and/or felt such plantations at variance with their pressing needs (Potter 1996; Otsamo 2000; Marjokorpi and Otsamo 2006), which is illustrated with a quote by one of those affected: ‘The trees are healthy, but the people are sick at heart’ (Potter 1996). And the production of lignocellulosic feedstock is not the only option for the conversion of Imperata cylindrica grasslands. It has, for instance, been suggested that, where possible, replacement of Imperata cylindrica grasslands by agroforests may be more in line with the needs of local populations (De Foresta and Michon 1996).

Assuming ‘business as usual’, a strong future expansion of transport biofuel pro­duction is expected to cause large-scale replacement of nature (Germer and Sauer — born 2007; Gurgel et al. 2007; Johansson and Azar 2007; Sivaram 2007; Christers — son 2008). At a regional scale, this seems to be confirmed by studies about a future expansion of biofuel production. For instance, in studies regarding the perspectives for ‘sustainable’ modern biomass production in Asian countries such as China, In­dia, Sri Lanka, Malaysia and Thailand, production of biofuels means to a consid­erable extent conversion of forests into plantations (Bhattacharya et al. 2003). And use of marginal lands for biofuel production in Southwestern China is doubted as much of this land is on sloping land that is prone to serious erosion (Naylor et al.

2007) . In Africa, wetlands of high ecological value are increasingly considered for biofuel crops such as sugar cane and oil palm (Sielhorst et al. 2008). Moreover, the land claims associated with the expansion of transport biofuel production should be considered against the background of increasing food production.

Currently on land, about one fifth of net primary production, or somewhat more, is appropriated by humankind (Imhoff et al. 2004; Haberl et al. 2007), and about 38% of land is in agricultural use (FAO 2007). Tilman et al. (2001a) have esti­mated that the area needed for expansion of agriculture for food production until 2050 may be about twice as large as the area of surplus and degraded agricul­tural land identified by Hoogwijk et al. (2003). And the total amount of arable land where wheat, maize, oilseeds and sugar is grown for food and feed purposes is pro­jected to grow between 2000 and 2020, while assuming substantial improvements in yield per hectare (Eickhout et al. 2008). Growing additional crops for ethanol or biodiesel production or establishing plantations of trees will, assuming business as usual, only add to the conversion of nature into ‘culture’, and thus to loss of habitats for living nature (Koh 2007; del Carmen Vera-Diaz et al. 2008; Searchinger et al.

2008) .

Replacement is not the only impact that an expanded production of transport biofuels may have on living nature. It is likely (e. g. Goldemberg 2008) that expan­sion of biofuel production will partly result in intensification of agriculture, which is often associated with the increased use of inputs such as nutrients, pesticides and (irrigation) water and increased drainage (Tilman et al. 2001a; Tscharntke et al. 2005; Liira et al. 2008). This will have side effects that affect biodiversity. It is also possible to harvest nature for biomass that may serve as the basis for biofuel pro­duction. For a specified amount of biofuel, this may well affect a larger area than

for a similar amount of biofuels from crops. The removal of biofuels from forests and other natural ecosystems may impact biodiversity. A case in point is the use of forestry residues, which may negatively affect a variety of species (Norden et al. 2004; Rudolphi and Gustafsson 2005). Natural habitats may also be invaded by species planted for biofuel production (Zedler and Kercher 2004; Lavi et al. 2005; Raghu et al. 2006; Nash 2007).

Effects of replacement of nature by agriculture on biodiversity and ecosystem services will be discussed in Sect. 5.3. The effects on biodiversity and ecosystem services of cropping and harvesting practices and of invasive species used in biofuel cropping will be considered in Sects. 5.4 and 5.5, respectively. But first, in Sect. 5.2, we will go briefly into the impact of biodiversity loss on natural resources which have monetary value.