Studies regarding annual crop production may provide clues about the factors nec­essary for a type of land-based biomass production that has high productivity and may be maintained indefinitely. Syers et al. (1997),Vance (2000) and Lal (2001a, b)

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

have surveyed such studies and did show that one important factor in maintain­ing high productivity is soil conservation and the maintenance of high levels of organic matter in the upper layer of the soil. Loss of soil due to erosion ultimately leads to a strong decline in crop productivity (Lal 2001). Soil erosion is a major problem in annual crops, including the production of major current biofuel feed­stocks such as corn, sugarcane and soybeans (Mahadevan2008; Smeets etal. 2008), but also may be a problem in plantations and forests (Worrell and Hampson 1997; Perry 1998).

Soil organic matter is an important reserve for plant nutrients such as nitrogen (N) and phosphate (P). It improves soil structure and water-holding capacity (Kahle et al.

2002) and limits erosion (Troeh et al. 1999). Soil organic matter is also involved in weathering that extends the availability of nutrients (McBride 1994). Depletion of organic matter in soils ultimately results in a decrease in yields (Syers et al. 1997; Perry 1998). In many areas of the world, arable land currently shows a net loss of soil organic matter, if compared with its virgin or natural status (Cole et al. 1997; Ogle et al. 2005). Levels of soil organic carbon that are under long-term cultivation with annual crops tend to be lower than under native vegetation. On average, soil carbon levels under arable soils with a long history of cultivation tend to be approxi­mately 18% lower under temperate dry conditions, approximately 30% lower under temperate moist or tropical dry conditions, and approximately 42% lower in tropi­cal moist climates if compared with soils under native vegetation (Ogle et al. 2005). These reductions of carbon levels in soils have contributed to increased levels of greenhouse gases in the atmosphere. In many soils in tropical and subtropical ar­eas, especially in Asia and sub-Saharan Africa, due to systematic excessive residue removal, soil organic carbon pools have decreased to levels that are conducive to soil degradation. Such soils also show substantially reduced production levels (Lal 2008).

Current agricultural practice often leads to further losses of soil carbon from arable soils. Net losses of soil carbon have been documented for the European Union (Vleeshouwers and Verhagen 2002), Eastern Canada (Gregorich et al. 2005), China (Li et al. 2003; Tang et al. 2006; Wright 2006), Nepal (Matthews and Pilbeam 2005; Shrestha et al. 2006), Brazil (Zinn et al. 2005; Jantalia et al. 2007), Sudan (Ardo and Olsson 2003), the southern Ethiopian highlands (Lemenih and Itanna 2004) and in West Africa (Ayanlaja et al. 1991; Ouattara et al. 2006; Bationo et al. 2007; Lufafa et al. 2008). From peaty arable soils losses may be especially high when there is deep drainage and intensive mechanical soil disturbance (Freibauer et al. 2004). Net carbon losses varying from 6 Mg in northern Norway up to 15MgCha-1year-1 in the tropics have been reported (Granlund et al. 2006; Reijnders and Huijbregts 2007, 2008).

It has emerged that low crop yields, the absence of cover crops, low additions of crop residues, low additions of manure, mechanical tillage and high temperatures enhance the loss of carbon from arable soils (Vleeshouwers and Verhagen 2002; Pretty et al. 2002; Freibauer et al. 2004; Lal and Pimentel 2007; Valentin et al. 2008). There is evidence that excessive reliance of synthetic nitrogen fertilizers may be conducive to carbon loss (Kahn et al. 2007; Triberti et al. 2008). It has also been

found that depending on the nature of the soil, climate, crops and crop rotation, crop residues management and tillage system, a partial (20-50%) removal of crop residues from the field reduces the pool of soil organic carbon, can exacerbate soil erosion hazard and negatively impact future yields of crops (Wilhelm et al. 2004; Lal 2005; Blanco-Canqui and Lal 2007; Lal 2008; Varvel et al. 2008). Models with parameters based on empirical data have been developed to estimate the impact of residue removal on soil organic carbon (e. g. Saffih-Hdadi and Mary 2008).

Sustainable use of soil and soil organic matter should be such that levels of soil organic matter do not decrease and that soil loss (erosion) should not exceed addition to topsoil stocks by natural processes. The latter may add between 0.004-0.5 mm of topsoil per year (Cannell and Hawes 1994). A variety of measures has been pro­posed to reduce erosion in annual cropping. These come under the umbrella of the term conservation tillage (Cannell and Hawes 1994; Lal 1997, 2001, 2008). They include the reduction of tillage (preferably to no-till or zero tillage), the use of cover crops and nitrogen-fixing legumes, intercropping, contour farming, increased return of harvest residues or residue mulches to the soil, use of manure, direct seeding, cor­recting effects of soil compaction due to vehicles and catching soil subject to water erosion on sloping soils by terracing or barriers (Gumbs 1993; Cannell and Hawes 1994; Lal 1997, 2001, 2008; Lal et al. 2007; Mills and Fey 2003). High inputs of carbon (residues, manure, compost) in soils that are subject to tillage are conducive to maintaining soil organic carbon levels (Jenkinson et al. 1990; Grace et al. 2006; Reijneveld et al. 2009).

In forests, limitations to harvesting and the use of heavy machinery on erodible soils and judicious planting may be necessary to prevent soil losses from exceeding additions to topsoil stocks (Pimentel et al. 1997b; Worrell and Hampson 1997). To maintain soil organic matter levels, in intensively managed forests and on planta­tions, intensive site preparation involving burning should be avoided, as this leads to volatilisation of soil carbon and prospective soil carbon (Perry 1998; Bauhus et al.

2002) . Also, limitations to removal of harvest residues from forests may be neces­sary to maintain levels of soil organic matter (Worrell and Hampson 1997).

Soils to which crop residues are returned tend to store more soil organic carbon (and nitrogen) than plots where residues are taken away (Vance 2000; Mendham et al. 2002; Dolan et al. 2006; Epron et al. 2006). In this respect, there are two phe­nomena with opposite effects for C4 and C3 plants. Residues from plants that have C4 photosynthesis seem less effective in contributing to soil carbon than the same amounts of residues from plants with C3 photosynthesis (Wynn and Bird 2007). On the other hand, C4 crops rather often generate relatively large amounts of below — and aboveground biomass, if compared with C3 crops (Wright et al. 2001; Wilhelm et al. 2004). Vleeshouwers and Verhagen (2002) estimate that adding to arable soils the cereal straw that is currently taken away may, on average, increase European soil carbon levels by 0.15Mgha~1year~1. Other studies have shown that full return of crop residues to arable soils in temperate climates may increase soil carbon lev­els by up to 0.7Mg C ha-1year-1 (Webb et al. 2003; Smith 2004; Freibauer et al. 2004; Rees et al. 2005). For maintaining soil carbon stocks in tropical soils, return­ing residues, application of other organic matter such as manure, shrub prunings and household composts, cover crops and fallows have been advocated (Lal and Bruce 1999; Bationo and Buerkert 2001; Nandwa 2001; Lufafa et al. 2008).

A changing climate will impact soil carbon stocks. In part, this impact is de­pendent on crop productivity. There is only limited empirical evidence about likely future crop productivity. Kim et al. (2007) have studied the C4 crop corn and con­cluded that under elevated CO2 concentration, productivity may remain unchanged. It has also been suggested that under elevated CO2 concentrations in the atmosphere, productivity of C3 plants may increase and that this may enhance soil carbon se­questration (Marhan et al. 2008), but increased temperature also leads to increased respiration in soils, and soil carbon dynamics may be impacted by changes in precip­itation (Marhan et al. 2008). Overall effects may vary for different climate regions. In the European context, Vleeshouwers and Verhagen (2002) estimate that an in­crease in average temperature of 1 °C caused by an increase of CO2 concentration may, ceteris paribus, lead to an average net loss of soil organic carbon of about 0.04 Mgha~1 year-1.

In view of carbon losses, increased use of agricultural residues has been advo­cated in order to maintain (or restore) proper levels of soil organic carbon and ensure agroecosystem sustainability (Lal 1997; Duiker and Lal 1999; Lal 2001,2008; Oue — draogo et al. 2006; Chivenge et al. 2007; Lal and Pimentel 2007). Adding lignocellu — lose to arable soil is more useful in this respect than more easily degradable carbon compounds such as sugars or starches (Sartori et al. 2006). Available evidence is limited but suggests that approximately 4-24% of carbon contained in residues of crops may be converted to refractory soil organic carbon in agricultural soils (Lal 1997; Follett et al. 2005; Razafimbelo et al. 2006; Triberti et al. 2008).

Crop residues contain cellulose in a matrix of lignin and hemicellulose. Lignin is, together with compounds such as cutins, suberins and tannins, largely respon­sible for humus formation in arable soils (Kirk 1971; Rasse et al. 2005) and in such soils is a major contributor to refractory soil organic carbon (Loveland and Webb 2003; Chapman and McCartney 2005). There is evidence that among the compo­nents of lignocellulose (lignin, hemicellulose and cellulose) in arable soils, lignin is by far the most refractory component (Melillo et al. 1989; Spaccini et al. 2000; Quenea et al. 2006). Thus, lignin is more suitable for carbon sequestration in arable soils than hemicellulose. For this reason, removal of residues with high concentra­tion of lignin (such as nut shells) may be expected to be more negative to arable soil carbon stocks than residues with a lower lignin level, such as wheat or rice straw.

Still, the presence of carbon compounds, which are more easily degraded than lignin (with a half life less than 1 year), in arable soil is also important for soil fertil­ity and stability (Spaccini et al. 2000; Loveland and Webb 2003). The carbohydrates hemicellulose and cellulose in harvest residue belong to this category (Spaccini et al. 2000).

Against this background, systematic removal of all crop residues for biofuel pro­duction is not a good idea (Lal 2008; Reijnders 2008; Saffih-Hdadi and Mary 2008). Limitations on crop residue removal will have an upward effect on energy input into, and costs of residue collection for, biofuel production (Higgins et al. 2007). To the extent that crop residues are removed, there is furthermore a case for selecting crop residues for the production of the transport biofuel ethanol that have relatively high concentrations of hemicellulose and cellulose susceptible to conversion into ethanol. In the case of corn stover, this fraction consists of cobs, leaves and husks (Crofcheck and Montross 2004). The crop residue fraction that is relatively rich in lignin may be expected to be a relatively efficient contributor to refractory carbon in arable soils, but also contains a substantial amount of carbohydrates that are more easily degradable and contribute to soil fertility and stability. Thus, it may be that, for example, the scope of residue removal for ethanol production may be widened by selecting residues on the basis of their relative suitability for ethanol production and for the formation of refractory soil carbon, respectively.

Another option is to consider a return of processing ‘wastes’ of crop residues that are relatively rich in lignin. In generating ethanol from crop residues by enzymatic conversion (see Chap. 1), a residue emerges that is rich in lignin and also contains unreacted cellulose and hemicellulose (Mosier et al. 2005). It would seem worth­while to consider applying this residue to arable soils. Such an application would serve the presence of refractory carbon in arable soil, while it may also contribute to the presence of more rapidly degradable carbohydrates. In doing so, one should limit or prevent undesirable side effects of adding this processing residue. A matter to consider in this respect is the accumulation potential of the residue for phenolic carbon compounds. Such accumulation may occur under anaerobic conditions, and this may have a negative effect on soil fertility (Olk et al. 2006). Ionic composition and pH of the processing residue are subject to limitation when use of the residue is to be sustainable (Mahmoudkhani et al. 2007). Also, one should be aware that lignin binds heavy metals such as cadmium much better than cellulose does (Basso et al. 2005). Thus, provisions should be in place to limit the flow of heavy metals to soils when the fraction that is rich in lignin is applied. If the processing residue has acceptable quality, it may well be that the amount of crop residue that can be removed from the field without a negative impact on soil characteristics can be in­creased. The quantitative and qualitative aspects of applying processing residues to arable soils would seem to merit further research. Finally, it should be noted that the refractory character of lignin in the arable soils studies cannot be generalized to all soils. There is, for instance, evidence that in lowland tree plantations in Costa Rica, litter decay increases with increasing lignin content (Raich et al. 2007).

By the beginning of 2008, the share of transport biofuels in the worldwide consump­tion of transport fuels was below 1%, and land use for transport biofuels was esti­mated at 13.8 Mha in the USA, Brazil, China and the EU (Renewable Fuels Agency 2008), but significant upward impacts on the conversion of nature into cropland could be noted (OECD-FAO 2007; Nepstad et al. 2008; Chap. 5). Current policy targets for the expansion of transport biofuel production have been estimated to re­quire between 55 and 166 Mha land (Renewable Fuels Agency 2008). Such a major further expansion of transport biofuel production will (ceteris paribus) stimulate the conversion of nature into cropland (Searchinger et al. 2008; Sukhdev 2008).

Moreover, as indicated in Chaps. 2 and 5, it is to be expected that part of a further expansion of transport biofuel production will come from intensification of agricul­ture. Intensification of agriculture is expected to be associated with higher inputs of nutrients and pesticides and increased irrigation and drainage (Tilman et al. 2001; Datta et al. 2004; Tscharntke et al. 2005; Liira et al. 2008). This, in turn, is expected to lead to a decline in biodiversity among many taxa and a loss of ecosystem services (Tscharntke et al. 2005; Liira et al. 2008).

When, unlike current practice, low-quality land is used for the expansion of bio­fuel production (as suggested in Sects. 6.3 and 6.5) and there is a fixed amount of transport biofuel to be produced, such as mandated under several current regula­tions, then — in view of the probably relatively low productivity of the land — larger areas will be needed than in the case of the use of good-quality land. Moreover, as pointed out in Chap. 5, abandoned and fallow lands by themselves rather often har­bour substantial biodiversity. When such relatively biodiverse abandoned and fallow agricultural lands are exploited, the negative impact on biodiversity may be large (Huston and Marland 2003; Marland and Obersteiner 2008). On the other hand, as pointed out in Chap. 5, there are also fallow and abandoned agricultural lands with relatively low biodiversity, such as parts of the Imperata cylindrica and Saccharum spontaneum grasslands, which may be exploited for transport biofuel production with a relatively low impact on biodiversity.

The size of the impact of expanding transport biofuel production is also depen­dent on other factors. Effects on biodiversity would, for example, be relatively large when current hotspots of biodiversity, such as tropical rainforests, the Cerrado sa­vannah or nature in the Cape region of South Africa (Darkoh 2003; Koh 2007; Danielsen et al. 2008), are converted into land for the production of transport biofuel feedstocks. Also, if precision agriculture and water-efficient irrigation techniques are used for the expansion of feedstock production (cf. Sects. 3.4 and 4.6), the im­pact thereof on biodiversity may well be lower than in the case of conventional practices, because water consumption and the emissions of nutrients and pesticides may be lower.

Though the impact of a maj or expansion of transport biofuel production on living nature may be variable, there would seem to be no scope for a major expansion of biofuel production in such a way that biodiversity loss will be zero. It is likely that a major expansion of transport biofuel production will have a major negative effect on biodiversity and ecosystem services.

As pointed out before, about 1% of primary energy use for transport worldwide con­cerns electricity (de la Rue du Can and Price 2008). Moreover, especially because the electromotor is more efficient than internal combustion engines, electric traction with electricity derived from power stations is relatively fuel efficient (e. g. Reijnders and Huijbregts 2007).

In Sect. 1.3, the types of organic materials that are currently used in electricity production have been outlined. It has also been suggested to use herbaceous crops that generate dried downbiomass, such as horseweed and sunflower (Kamm 2004) for electricity production. Substantial expansion of biofuel-based electricity production for transport is dependent on the social acceptability of electric traction in cars. Inter­estingly, there have been periods in which the social acceptability of electric traction has been high for types of car transport now dominated by internal combustion en­gines. In 1899/1900, electric motorcars outsold other types of cars in the USA (H0yer 2008), electric taxis were then highly popular and between 1900 and 1920, electric vans were important in intra-urban and suburban transport of a variety of goods in the USA (Mom 1997; Mom and Kirsch 2001). All in all, the 1880-1925 period was a golden age for electric cars in the USA and parts of Western Europe (Mom 1997; H0yer 2008). In the 1940s, vans used for the German postal services and for milk and bread delivery in Britain were usually electric (№yer 2008). Still, electric cars meet a substantial demand of large fleet owners in urban settings (e. g. as post office and street cleaning vans). Cars powered by electricity from power stations have, how­ever, only had very limited success among individual users (Gjoen and Hard 2002). Opinions diverge about their future potential. Some view the re-emerging interest in electric cars as an episode in a series of unsuccessful attempts to substantially increase the use of electric cars. Others predict that there will be a rapid and fast increase in the use of electric cars. Lache et al. (2008) suggest a rapidly increasing market share for plug-in electric vehicles in Europe, with lithium batteries as key enabling technol­ogy. Others suggest that large socio-cultural changes, major technical changes and substantial financial incentives are necessary to make plug-in, battery-powered, all­electric traction for cars much more popular in the future (Delucchi and Lipman 2001; Gjoen and Hard 2002; Chalk and Miller 2006; H0yer 2008).

When electric traction gets a much larger share in road transport, it is likely that two technologies will contribute significantly to its success. The first is better batteries. Prime candidates are currently lithium ion batteries, which for a specified electrical performance are, over their life cycle, less of an environmental burden than competing batteries, such as the lead-acid and nickel-based batteries (Matheys et al. 2007). The second is plug-in hybrid cars, which in their life cycle energy use may have an advantage over current hybrid cars (Samaras and Meisterling 2008).

1.7 Recent Development of Transport Biofuel Production: Volume, Costs and Prices

1.7.1 Volume of Biofuel Production

Some companies operating means of public transport which use electric traction have opted for ‘green electricity’, which may include biomass-based electricity pro­duction. No data have been found that allow a worldwide estimate of biomass-based electricity in electric traction. Still, the production of bioethanol apparently ac­counts for most of the current volume of transport biofuel production. The focus in the USA is largely on ethanol made from cornstarch, and in Brazil, it is mostly on ethanol made from sugar cane. China has also emerged as a major producer of bioethanol, preferentially from sugar cane, cassava and yams (Cascone 2007), and so has the European Union, producing bioethanol from wheat and sugar beet (Bern — des and Hansson 2007). India, Russia, Southern Africa, Thailand and the Caribbean are emerging as important producers of ethanol as a transport fuel (Cascone 2007; Szklo et al. 2007; Barrett 2007; Amigun et al. 2008; Nguyen et al. 2008). Estimated bioethanol production volumes for 2006 in the main production areas are given in Fig. 1.2 and sum up to a world estimate of 51 x 106 m3. The estimated world pro­duction of bioethanol in 2007 was 54 x 106 Mg (Monfort 2008).

The worldwide production of biodiesel in 2006 was probably in the order of

6.4 x106 Mg, with the share of the EuropeanUnionbeing approximately 77% and of

the USA approximately 12% (Canakci and Sanli 2008). In 2007, estimated biodiesel production was approximately 7.6-8 x 106Mg (Von Braun 2008; Monfort 2008). Recently, there has been rapid growth especially in Argentine biodiesel production capacity. Argentine biodiesel production is expected to grow to about 3.5 x 106 Mg per year by 2008/2009 (Lamers et al. 2008). Biodiesel production on the basis of castor oil is expanding in Brazil, though it has been argued that in view of properties and price, castor oil is unlikely to be competitive with palm oil and rapeseed oil (Mathews 2008b; Scholz and da Silva 2008). India, Malaysia, Nicaragua, and several Pacific island states are also involved in substantial biodiesel production (Grimm 1999; China Chemical Reporter 2007; Cascone 2007; Cloin 2007; Fairless 2007; da Costa et al. 2007; Runge and Senauer 2007). Much of the biofuels produced are for domestic use, but increasingly, biofuels are traded internationally. Brazil and Argentina are, for instance, emerging as major transport biofuel exporters.

Most of the biofuels which were produced for transport applications in 2006 were based on substances that are also applied as foodstuffs. Such biofuels have been called ‘first generation biofuels’. Biofuels can also be produced on the basis of other substances, such as lignocellulosic feedstocks and oils that are not foodstuffs. This category of biofuels has been called ‘second generation biofuels’. However, ethanol production as a by-product of the sulphite pulping process, Russian ligno — cellulose-based ethanol production and the application of Jatropha oil for biodiesel production have evolved contemporaneousto or even before first generation biofuels such as cornstarch-based bioethanol and rapeseed biodiesel (Grimm 1999; Zverlov et al. 2006; McElroy 2007). Also, algal biofuels are often called second generation biofuels, but several of the algae considered for this purpose have current applica­tions as food.

Moreover, as it is apparently felt that second generation is somehow better than first generation, the former designation is used in strange ways, for instance for de — oxygenated and hydrogenated edible vegetable oils (Rantanen et al. 2005; Mikko — nen 2008). For these reasons, the designations first and second generation will not be further used. It seems likely that biofuels made from substances that may also serve as food or feed will dominate the supply in the near future. Plans for other types of biofuel, when implemented, will by 2010 probably not be able to supply more than 1% of overall biofuel production, and such biofuels are unlikely to allow for large-scale replacement of biofuels from substances such as sugar, starch and edible vegetable oil before 2020 (Gibbs et al. 2008; OECD 2008).

Growth of biofuel production and/or consumption is foreseen in a number of countries. In Brazil, sugar cane production, as feedstock for ethanol, is expected to grow from 425 x 103 Mt in 2006 to 728 x 103 Mt in 2012 (Macedo et al. 2008), while the mandate for Brazilian biodiesel is set at 5% for 2013. In 2007, the USA mandated a growth of bioethanol production from 4.7 billion gallons in 2007 to 36 billion gallons by 2022. In 2008, there were, however, calls for revision of this target in the US Congress (Doering 2008). In 2008, Canada mandated a 5% ethanol blend in gasoline by 2010 and a 2% biodiesel blend in on-road diesel by 2012. The European Union in 2007 suggested a 10% share of biofuels in transport fuels by 2020, which in 2008 was hotly debated.

The production of transport biofuels can be accompanied by changes in carbon se­questration. Firstly, there can be changes in the carbon content of ecosystems. There may be both losses from, and increases of, C in the ecosystem, which in turn will change atmospheric CO2 concentrations. In early life cycle assessments, quantifica­tion of such changes was largely neglected (with some exceptions, such as Reijnders and Huijbregts 2003, Delucchi 2005; Kim and Dale 2005 and Cowie et al. 2006). However, by now, changes in C sequestration by ecosystems are increasingly rec­ognized as a major determinant of net seed-to-wheel greenhouse gas emissions (e. g. Fritsche 2007; Danielsen et al. 2008; Fargione et al. 2008; Gibbs et al. 2008; Ne — chodom et al. 2008; Searchinger et al. 2008; Wicke et al. 2009). In a number of cases, the link between change in C sequestration and expansion of biofuel produc­tion is direct. For instance, because forest is cleared or degraded land is planted for oil palm plantations serving the biodiesel market. In these cases, there is a large ef­fect on C sequestration (Danielsen et al. 2008; Fargione et al. 2008). There are also cases in which land use change has more limited direct effects on C stocks — for instance, cases where oil palms have replaced other plantation crops, such as rubber or coconut, that gave lower revenues (Tan et al. 2009).

There may also be indirect effects of expanding biofuel production on land use. These follow from the relative inelasticity of demand for food. When land used for food or feed (fodder) production is diverted to use for biofuel production, one may expect that food or feed production to a very large extent moves elsewhere (Searchinger et al. 2008). Some examples may illustrate this. Currently, in Brazil, there is substantial conversion of pasture to arable land to grow soybeans. Conver­sion of pasture to arable land tends to lower carbon in the ecosystem. But this is not the ‘whole story’. Farmers that used the pasture may move to new pastures or may import fodder from elsewhere, for which deforestation may be necessary (Nepstad et al. 2008). Also, the expansion of soybean cropping in Brazil is partly linked to the expansion of corn-for-biofuel production in the United States, which has dis­placed US soybean production (Nepstad et al. 2008; Scharlemann and Laurance 2008). Tilman et al. (2006) have proposed the use of native prairie species as a ba­sis for transport biofuel production from the US prairies, but to the extent that such production replaces current grazing by livestock, the latter has to be accommodated somewhere else. A last example refers to European wheat. In 2007, the price thereof was much elevated, in part due to the use of wheat starch for ethanol biofuel pro­duction. This led to a massive rise of corn imports with Brazil as a major supplier, which in turn was conducive to change of Brazilian land with natural vegetation into cropland.

There may also be other changes in carbon sequestration linked to biofuel life cycles. For instance, both in the case of electricity production of biomass and the production of synthesis gas from biomass, CO2 may be captured and stored in soils (e. g. aquifers or abandoned gas fields). In the case of electricity production, de­pending on the technology used, life cycle emissions of greenhouse gases linked to electricity production may be reduced by 75-84%, though other emissions such as those of eutrophying and acidifying substances may be increased (Odeh and Cock — erill 2008). There is currently only limited application of such capture and storage (abbreviated CCS), but this may change in the future (Odeh and Cockerill 2008; Gibbins and Chalmers 2008). Also, in generating transport biofuels from biomass by pyrolysis, significant amounts of charcoal, also called ‘biochar’, may be gener­ated (Demirba§ 2001), which in turn may be added to soils. This approach is cur­rently not practiced, but again, this may change in the future (Marris 2006; Renner

2007) .

Changes in carbon sequestration in ecosystems are currently common in trans­port biofuel production. Clearing forests to allow directly or indirectly for culti­vation of transport biofuel crops corresponds with substantial emissions of green­house gases (Fearnside and Laurance 2004; Righelato and Spracklen 2007; Far — gione et al. 2008; Reijnders and Huijbregts 2008a; Searchinger et al. 2008). This follows to a large extent from the large changes in aboveground and soil carbon stocks. For instance, estimates of aboveground C stocks in rainforests range from approximately 130-270Mg C ha-1 (Danielsen et al. 2008; Fargione et al. 2008; Reijnders and Huijbregts 2008a), and similar estimates for aboveground C stocks in temperate forests range from approximately 100-160 Mgha-1 (Searchinger et al.

2008) . Clearing other types of natural vegetation for transport biofuel production may lead to a lower but still substantial desequestration of C. For instance, clearing Cerrado savannah for soybean biodiesel production leads to an average deseques­tration of about 23 Mg C ha-1 (Reijnders and Huijbregts 2008b), whereas losses of C stocks in soil due to conversion of US grasslands into cropland for biofuels have been estimated at approximately 17-40 Mg C ha-1 (Fargione et al. 2008). On the other hand, Germer and Sauerborn (2007) have suggested that planting oil palms on degraded grassland may lead to a substantial accumulation of aboveground and soil organic carbon, estimated at 135 Mg C ha-1 over a period of 25 years. Ne — chodom et al. (2008) have suggested very favourable life cycle assessments regard­ing biomass from forest remediation in California, as it is assumed that without using such biomass, forest fire risk would be much higher.

If one takes account of changes in carbon sequestration due to ecosystem changes in life cycle environmental impact assessments, the distribution of changes in the carbon content linked to land use change over the subsequent period should be established, as pointed out in Sect. 4.2. The Intergovernmental Panel on Climate Change (IPCC 2006) has suggested the use of a 20-year period for this purpose, but calculations also have been made for periods of up to 100 years (Reijnders and Huijbregts 2008a; Wicke et al. 2009).

Another way to approach this matter is to balance the net reduction of C emis­sions from the use of biofuels against the ‘biofuel carbon debt’ due to C losses from ecosystems. This gives rise to a number of years to pay off the carbon debt. An ex­ample thereof for palm oil, reflecting a large effect on C stocks, is given in Fig. 4.1.

Direct changes in land use linked to expanded transport biofuel production are relatively easy to include in life cycle assessments. Including indirect changes is more difficult. A first possibility to take account of changes in C sequestration due to indirect changes in land use has been suggested by Fritsche (2007). Fritsche pro­posed that for such indirect effects, there should be a ‘risk adder’ or ‘iLUC’, re­flecting the risk that there may be clearance of natural vegetation or other forms of land use change that affect C sequestration and the net emission of CO2 (usually distributed over a 20-year period in line with recommendations of the Intergovern­mental Panel on Climate Change). The impact of emissions associated with differ­ent assumptions as to indirect effects on land use change (different risk adders or iLUCs) may be considerable. Illustrations are given by Table 4.1.

A more sophisticated approach uses modelling to estimate land use and land use change induced by expanding transport biofuel production. A proposal for such modelling has been published by Kkrverpris et al. (2008). Searchinger et al. (2008) have applied this type of modelling using a representation of outcomes in terms of carbon debt. Searchinger et al. (2008) evaluated expanding corn cropping to achieve the US government goals set for the transport biofuel supply in 2016. They used a model to estimate worldwide land use changes following from this development, which, they estimate, will divert 12.8 million ha of US cropland to transport bio­fuel production. Searchinger et al. (2008) estimate that this will bring 10.8 million

ha of additional land into cultivation. Searchinger et al. (2008) have calculated that over a 30-year period, including land use change linked to expanding US corn — based ethanol production will add about 93% to greenhouse gas emissions, if com­pared with fossil gasoline. The payback time for the carbon debt caused by land use change by corn ethanol is estimated by Searchinger et al. (2008) at 163 years. Mod­elling such as performed by Searchinger et al. (2008) is dependent on assumptions. Important among these are assumptions regarding the future productivity of land. Searchinger et al. (2008) assumed per hectare an average worldwide corn yield in­crease of 11.5% between 2007 and 2016. A lower increase would lead to a larger indirect effect on land use, a higher increase to a lower indirect effect. For instance, an additional 20% increase in yield per hectare would decrease the carbon debt from 163 to 133 years, whereas a yield increase of 1.5% between 2007 and 2016 would increase the carbon debt to 183 years (Searchinger et al. 2008).

Thus, land use change may have a very large impact on seed-to-wheel green­house gas emissions. Biofuel crops may, furthermore, differ in their consequences for carbon stocks in the soil on which they are grown after land use change. An­nual cropping of European arable land has been associated with losses of, on aver­age, 0.84Mgha~1year~1 soil carbon from the upper soil layer (Vleeshouwers and Verhagen 2002), and this has implications for the net greenhouse gas emissions as­sociated with biofuel crops such as rapeseed and wheat (Reijnders and Huijbregts 2007, 2008b). Net losses of soil carbon from arable soils under annual crops are not restricted to Europe. In Eastern Canada, arable land on average loses 0.07 Mg C ha-1year-1 to the atmosphere (Gregorich et al. 2005). In China, the estimated loss on average is 0.81 Mg C ha-1year-1 (Tang et al. 2006). The loss of soil or­ganic carbon in China parallels the removal of approximately 300 million tons of straw (Wright 2006). From Nepal, losses of soil carbon have been noted in a va­riety of cropping systems, with, for instance, losses in maize/millet cropping sys­

tems ranging between 0.11 and 0.23 Mg C ha-1year-1 (Matthews and Pilbeam 2005; Shrestha et al. 2006). In a set of 14 cropped Brazilian soils sampled annu­ally, losses of soil organic carbon were on average 0.15 Mgha-1year-1 (Zinn et al.

2005) , whereas for soybean-based crop rotation in the Brazilian Cerrado region, losses were reported to be between 0.5 and 1.5 Mg C year-1 ha-1 (Jantalia et al. 2007). In Africa, losses of carbon from soil are expected as little or no agricul­tural residues are returned to soils in many cropping systems (Syers et al. 1997). In the case of semi-arid Sudan, an annual loss from cropland during the twentieth century of, on average, 0.29 Mg C ha-1 year-1 has been found (Ardo and Olsson 2003). Long-term experiments in East Africa have suggested that losses of 0.69 Mg C ha-1 year-1 are common (Nandwa 2001). Over relatively short and recent time spans, measured C losses in the southern Ethiopian highlands amounted to approx­imately 0.85-1.75Mg C ha-1year-1 (Lemenih and Itanna 2004), and in Western Burkina Faso to approximately 0.31 Mg C ha-1 year-1 (Ouattara et al. 2006).

From peaty arable soils, losses may be much higher. They are especially high when there is deep drainage and intensive mechanical soil disturbance (Freibauer et al. 2004). Net carbon losses varying from 6 Mg in northern Norway up to 15 Mg C ha-1 year-1 in the tropics have been reported (Granlund et al. 2006; Reijnders and Huijbregts 2007, 2008a).

Mechanical tillage and deep ploughing tend to favour net losses of soil carbon stocks (Fontaine et al. 2007). No-till practices, ceteris paribus, lead to higher lev­els of soil carbon than tillage (Fontaine et al. 2007), so does the use of cover crops (Pretty et al. 2002). Also, returning harvest residues to soil is conducive to carbon sequestration in soils (Reijnders and Huijbregts 2007; Lal 2008). Intemperate cli­mates, soil tillage can be combined with a stable level of soil carbon, when fresh inputs of C (e. g. manure, crop residues) are large enough (Reijneveld et al. 2009).

Whereas losses of soil carbon are not uncommon under annual crops, net carbon sequestration can occur when perennials or multiannual crops are grown.

Growing switchgrass that may serve as a cellulosic feedstock for transport bio­fuel production is associated with the accumulation of soil carbon in the upper soil layer (Ma et al. 2000; Lal 2008). Carbon accumulation rates depend on cli­mate and soil and time period chosen. McLaughlin and Adams Kszos (2005) found in the mid-Atlantic region of the USA over a 6-year period an accumulation of 1.2-1.6Mg C ha-1 year-1. A simulation study considering a 30-year period of switchgrass cropping in the Eastern USA suggests a carbon accumulation rate of 0.53 Mgha-1 year-1 (McLaughlin et al. 2002). Long-term trials with ley grass in Sweden suggested an annual accumulation of 1.0-1.3Mg C ha-1 year-1 over a 30-year period (Boijesson 1999). Bjoresson (1999) has suggested that a switch from annual crops to perennial biofuel crops may be associated with a gain of about 0.5Mg C ha-1 year-1 in Swedish mineral soils. However, net sequestration does not always occur under perennial crops. In a grassland with Miscanthus sinensis in Japan, harvested yearly, reductions of carbon stocks were observed in the order of 0.56-1 Mgha-1year-1 (Yazaki et al. 2004). Finally, soil organic carbon levels may change when the climate changes. The effects of climatic change are complex and may be different dependent on geography and the (agro)ecosystem (Luo 2007)

For Europe, Vleeshouwers and Verhagen (2002) estimated that an increase in aver­age temperature of 1 °C may be associated with an average net loss of soil organic carbon in arable soils of about 0.04 Mgha^year-1.

The intercept by the Earth of solar energy exceeds the present input of fossil and uranium fuels into the world economy by a factor of about 10,000 (Lewis and No- cera 2006). The average daily solar irradiation varies, dependent on latitude, climate and season. When on the equator, maximum irradiation is on a horizontal plane, but away from the equator, for the maximum intercept of solar radiation by a fixed plane, the plane should have an angle corresponding to latitude (e. g. Qelik 2006). Average daily solar irradiation (measured on a horizontal surface) that may support feedstock for biofuel production varies roughly between 7 and 25 MJm~2. The daily worldwide average irradiation is about 15.5MJm~2, or 180 Wm~2. Differences be­tween days can be large. For instance, in Amsterdam (52°21′ N), the average daily irradiation is approximately 3 MJm~2 in January and 17 MJ m~2 in July (Akkerman et al. 2002). The greatest annual input of solar radiation tends to occur in subtrop­ical regions at latitudes between 20 and 30° and little cloud cover. Humid tropical regions have somewhat lower irradiation (Sinclair and Muchow 1999). When go­ing poleward from a latitude of about 30°, solar irradiation tends to decrease. As for major areas for current biofuel production, in Brazil, where sugar cane ethanol is produced, daily solar irradiation is on average about 220 W m~2 (approximately 19 MJday-1 m~2 or 694 x 102 GJyear-1 ha-1). In the US, average daily irradiation varies between 12 and 22 MJm~2, whereas in the US Midwest, where there is large — scale corn ethanol production, solar irradiation is about 170 Wm~2 (approximately 14.7MJday-1m~2 or 536 x 102GJyear-1ha-1) (Kheshgi et al. 2000; Vasudevan and Briggs 2008).

In establishing the overall conversion efficiency of technologies for the conver­sion of solar energy, there should be a correction for the cumulative energy demand associated with the biofuel life cycle and the life cycle of physical conversion tech­nologies (Reijnders and Huijbregts 2007). For instance, if the lower heating value of fossil fuel inputs amounts to 20% of the lower heating value of a biofuel, the solar energy conversion efficiency will be corrected by this percentage. The result thereof is the overall energy efficiency of the biofuel. This is summarized in the following equation:

SCEX = Yx ‘Ex’FEx • 100 x F

-r^solar

where SCEX is the solar energy conversion efficiency of biomass or biofuel type x (%), Yx is the yield of biomass type x (kg/ha/year), Ex the energy content of biomass or biofuel type x (MJ/kg), FEx the correction factor for fossil fuel input in the life

cycle of biomass or biofuel type x (MJ/MJ), and Esolar is the yearly solar irradiation (MJ/ha/year). SCEx is a measure that can help in estimating the ability of biofuels to displace fossil fuels.

As pointed out in Chap. 1, conversion of solar energy into biomass occurs by photosynthesis. Harvestable biomass that can be used for energy generation (yield) depends on a number of factors. At the present atmospheric CO2 concentration for C4 terrestrial plants, the maximum conversion efficiency is estimated at 5.5-6.7% and for C3 plants at 3.3-4.6% (Hall 1982; El Bassam 1998; Heaton et al. 2008b). A 6.7% solar energy conversion efficiency would correspond with a dry biomass yield of approximately 250Mgha-1year_1 at 40° latitude (El Bassam 1998). Ac­tual yields are much lower than theoretical yields, because there are factors — such as in the case of terrestrial plants, the absence of a full canopy, shading, photosatu­ration and limited availability of nutrients and water — which in practice reduce the efficiency. Due to such factors, the theoretical differences in conversion efficiency between, for instance, C3 and C4 plants may not materialize in real life differences in conversion efficiency. For instance, sorghum is a C4 plant that tends to be roughly as efficient as the C3 cereals. And C3 plants such as sugar beet and oil palm are in practice often more efficient in converting solar radiation into biomass than the C4 plant Miscanthus.

Living nature is an important supplier of natural resources which have monetary value, such as fuel, materials (e. g. rubber, aloe gel, wax, tannins, wood, thatch and broom grass), food, medicines and ornamental plants (Carr et al. 1993; Brown and Rosendo 2000; Brennan et al. 2005; Mutimukuru et al. 2006; Shackleton et al. 2007). Poor people, especially, often depend directly on the natural resources pro­vided by living nature (Shackleton et al. 2007; Vedeld et al. 2007). An estimated more than 1.6 billion people depend for their livelihood to varying degrees on forests, and about 60 million people are fully dependent on forests (World Bank

2004) . Vedeld et al. (2007) analyzed 51 case studies regarding rural dwellers from 17 developing countries in Africa, East Asia and Latin America and found that in­come from forests represented, on average, 22% of total income. Main contribu­tors to income were the collection of food, fodder, fuelwood, thatch and medicine. When food prices are high, collection of wild foods has added importance for the poor (e. g. Delang 2006). In the studies reviewed by Vedeld et al. (2007), medicines from forests contributed about 7% to the income of rural dwellers. Also, natural ecosystems other than forests, such as savannahs, are important providers of natural medicines (Shackleton et al. 2007).

Currently, three-quarters of the world population depend at least partly on natu­ral remedies (Sukhdev 2008). In China alone, 5,000 of the 30,000 recorded higher plant species are used for therapeutic purposes (Sukhdev 2008). Natural medicines have been found especially important to urban poor, for instance in countries such as South Africa and Brazil (Shanley and Luz 2003; Shackleton et al. 2007). An exam­ple of people currently affected by biodiversity loss are the city dwellers in Eastern Amazonia (Shanley and Luz 2003). Medicinal plants in this region are negatively affected by repeated cycles of forest burning and cutting and even more by the re­placement of forests by biofuel crops. This, in turn, affects the availability and price of such medicinal plants, which for plants with pharmacologically demonstrated effectiveness tended to be cheaper than their counterparts from the pharmaceuti­cal industry (Shanley and Luz 2003). There may also be a long-term effect on the availability of medicines produced by the worldwide pharmaceutical industry (Grifo et al. 1997). More than half of the medicines prescribed in the USA in 1993 con­tained at least one active compound ‘derived from or patterned after compounds derived from biodiversity’ (Grifo et al. 1997). With many species not investigated as yet as to their potential medicinal value, it may well be that the decrease of bio­diversity negatively affects the future availability of new medicines.

Finally, the case of the Mabira Forest, mentioned at the beginning of this chap­ter, illustrated the monetary importance of tourism. Nature-oriented tourism now accounts worldwide for about 10% of international tourism expenditures and ap­proximately 1% of total employment. It is increasingly important as a source of revenue for a wide variety of countries, accounting in some for 40-60% of all inter­national tourists (Carr et al. 1993; Watkins 2002; Nyaupane et al. 2004; Mowforth and Munt 2005; Cochrane 2006; Shackleton et al. 2007).

There are a variety of ways to use biofuels for transport. The first category focuses on electric traction, which currently accounts for about 1% of energy use in the transportation sector worldwide (de la Rue du Can and Price 2008). Electric trac­tion is common in train transport, but there are also ships powered by electricity, and a battery-powered small airplane has been demonstrated (Sanderson 2008). All­electric cars currently have limited application, but more recently there has been a rapid increase in the use of hybrid cars that use both internal combustion engines and electromotors (Mom and Kirsch 2001; Wurster and Zittel 2007; H0yer 2008).

Electricity can, for instance, be generated in power plants fired by biomass and stored in batteries. Also, electricity can be generated by onboard fuel cells fed with, for example, H2 derived from biomass or H2-producing organisms. Hydrogen used in fuel cells is, from a life cycle perspective, more energy efficient than the applica­tion of H2 in Otto or diesel motors (EUCAR et al. 2007; Hussain et al. 2007; Kleiner

2007) . Fuel cells may also be used for the propulsion of ships and airplanes (Little­field and Nickens 2005; Lapena-Rey et al. 2008; Sanderson 2008). Introduction of hydrogen as a major transport fuel requires concerted action of many stakehold­ers (Wurster and Zittel 2007) and includes large changes in fuelling infrastructure and a major effort to reduce fire and explosion risks (MacLean and Lave 2003; Ag — nolucci 2007; Astbury and Hawksworth 2007; Markert et al. 2007; Melaina 2007; Ng and Lee 2008). Also, major advances in several key components of motorcars are necessary for a successful large-scale introduction of all-electric or H2-powered cars (Chalk and Miller 2006; Matheys et al. 2007; H0yer 2008; Lache et al. 2008; Samaras and Meisterling 2008).

In practice, wood, animal wastes, harvest residues, municipal and industrial or­ganic wastes, landfill gas, ‘energy’ grasses (such as reed canary grass) and veg­etable oils have been used in power generation (e. g. Reijnders and Huijbregts 2005; Berggren et al. 2008; Heinimo 2008; Junginger et al. 2008; Reijnders and Huijbregts

2008) . Sewage sludges and wastewater treatment sludges are also applied, though these tend to be net users instead of net producers of energy due to their high water content (Wang et al. 2008c).

There is, furthermore, scope for the co-production of electricity and ethanol from sugar cane (Macedo et al. 2008). In producing electricity, both direct burning of biomass and burning after gasification or fermentation are practiced (Wheals et al. 1999). Problems in generating electricity from biomass have arisen due to slagging, corrosion and fouling mainly linked to the presence of inorganic elements such as Cl and K; in the case of gasification, fouling has also been linked to tar formation (Monti et al. 2008). Ways to decrease such problems, such as lowering Cl and K concentrations by judicious choice of feedstocks, have been researched (Monti et al. 2008), though there are types of biomass, such as macroalgae, that still appear un­suitable for direct combustion or gasification (Ros et al. 2009).

The second possibility is to produce liquid or gaseous biofuels that can be burnt in transport engines that currently burn fossil fuels. In 2006, such biofuels accounted for about 1% of energy use in the transportation sector worldwide (de la Rue du Can and Price 2008). Various engines operate under a variety of conditions, and not all liquid and gaseous biofuels are suitable to all applications. Quantitatively speaking, two engine types dominate road transport, and also transport in general: the diesel engine and the Otto motor. A variety of gaseous and liquid biofuels produced have been proposed for these engine types. As to the way these biofuels are produced, most of them can be allocated to three categories (Ahman and Nilsson 2008). The first category relies on the biochemical conversion of biomass into transport bio­fuels. Biochemical conversion is now used for the production of ethanol, butanol and methane. The second category is based on lipids (oils and fats) derived from organisms. Such oils may be applied directly or after processing (e. g. transesterifi­cation or catalytic cracking). The third category uses thermochemical conversion of biomass via pyrolysis or gasification into a variety of fuels.

A part of the transport biofuels which have been proposed are currently produced on an industrial scale and widely applied in means of transport. Ethanol obtained from starch or sugar by fermentation and biodiesel based on lipids from terrestrial plants are currently the main transport biofuels. Other substances that have potential as transport biofuels are produced on an industrial scale but hardly or not applied in Otto and diesel motors. A third category of transport biofuels include those in the laboratory and pilot plant stage. All these are shown in Table 1.1.

The energy contents of the liquid and gaseous transport biofuels mentioned in Table 1.1 may be different from the fossil petrol and diesel that they replace. Table 1.2 gives a survey of the energy contents (lower and higher heating values) in megajoules (MJ) of the liquid fossil and biofuels per kilogram (kg) and per litre (l). The lower heating value (LHV) represents net energy content, and the higher heating value (HHV) represents gross energy content (including the heat of condensation of water vapour produced by combustion (Piringer and Steinberg 2006)).

The differences in heating values indicate that when the amount of transport kilo­metres for a full tank is to be maintained, a substantial adaptation of tank size may be necessary when transport fuels contain high percentages of biofuels with relatively low heating values, such as dimethylether and ethanol (Semelsberger et al. 2006). This is not the only adaptation that may be necessary when switching to biofuels. Table 1.3 gives a brief summary of other adaptations for a number of biofuels.

Table 1.3 Problems and adaptations necessary for the use of biofuels Biofuel Problems and adaptations

Ethanol — Ethanol is relatively corrosive, and ethanol-gasoline blends may sep­

arate in pipelines; this limits the scope for pipeline transport. Also, ethanol is hygroscopic, and high water concentrations may lead to phase separation. So, in storage and distribution, exposure to water should be severely limited (Antoni et al. 2007; Atsumi et al. 2008).

— Limited admixture of ethanol (whether or not as ETBE: the tertiary butylether of ethanol) up to 5% is possible without adaptation of cars. If ethanol-fossil hydrocarbon blends with percentages of ethanol over 5% are used, however, changes in cars are needed (Antoni et al. 2007). Such changes regard the fuel-sending unit, the fuel injector, the fuel filter, fuel management and flame arrestors. When the percentage of bioethanol be­comes 85 or 100%, changes necessary for the engine become substantial (Antoni et al. 2007; Hammond et al. 2008). This has led to the develop­ment of flex vehicles that are able to run on blends with high percentages of ethanol, and also on conventional petrol.

Another factor important in productivity is the availability of sufficient mineral nutrients, such as fixed nitrogen (N), phosphorus (P), sulphur (S), potassium (K), calcium (Ca) and magnesium (Mg). As to the indefinite availability of sufficient nutrients, difficulties may well emerge (Manley and Richardson 1995; Sims and Riddell-Black 1998; Perry 1998; Ranger and Turpault 1999; Pare et al. 2002). These partly follow from the limitations of natural processes involved in making minerals available to the generation of biomass. These are deposition on soil and weather­ing (Hedin et al. 2003). On a time scale valid for forestry and cropping, there is only a very small addition to total reserves of minerals that can be made avail­able to biomass due to geologic processes such as weathering (Ranger and Turpault 1999). Thus, the availability of nutrients from reserves generated by processes such as weathering may well go down over time on a time scale relevant to cropping and forestry due to losses linked to harvesting and erosion. External inputs of N and P in cropping and forestry are, moreover, dependent on the large-scale use of geochem­ically scarce natural resources (phosphate ore and fossil fuels) that are formed in slow geological processes, which does not allow for sustainable us. Fossil fuels are used to produce N-based synthetic fertilizers (Galloway et al. 2008). We will now first consider nutrients in forests and plantations and thereafter nutrients in arable soils.

In this book, ‘sustainable’ is taken to mean that a practice can be continued in­definitely. As explained in Chaps. 2 and 3, this severely limits the extent to which geochemically scarce resources which have been formed in slow geological pro­cesses, such as fossil fuels and phosphate ore, can be converted into wastes. Sus­tainability also requires that renewable resources such as fertile soil, soil organic matter, groundwater and nutrients are maintained and retain their quality. Regarding cropping, this in turn leads to preferences for conservation tillage, much improved nutrient recycling and improved water efficiency. In forestry, this leads to a prefer­ence for long rotations and nutrient recycling. Sustainability also limits achievable biomass production. In Chap. 3, biomass production from currently abandoned, in­cluding fallow, agricultural land was estimated to have a sustainable yearly yield of approximately 23-28 EJ, about one order of magnitude below the yield suggested by de Vries et al. (2007).

The cost of biofuels is a longstanding topic of discussion, especially in relation to the costs of competing fossil fuels. Two types of costs are involved: the costs of producing biofuels and costs that users have in adapting to biofuels. The latter costs are highly variable. Co-firing wood pellets in power plants to power electric trains has low adaptation costs, and the same holds for adding low percentages of biofuel to conventional gasoline and diesel. However, for instance, switching from diesel and gasoline to (biofuelled) electric traction is a major operation. Here we will further focus on the costs of transport biofuel production.

For producers, there again are two types of costs. Firstly, there are costs borne by the producer. Secondly, there are external costs or externalities (Pigou 1920). External costs are (fuel-linked) costs that are not reflected in actual prices. Such costs are associated with negative environmental impacts, including negative im­pacts of air pollution on health (Johansson 1999) and on ecosystems, and the future availability of natural resources. But there are also other external costs associated with fuels, such as the costs of strategic stockpiling and in the case of mineral oil, military costs involved in safeguarding the supply (Zaldivar et al. 2001; Delucchi and Murphy 2008). Such costs are substantial and may vary strongly between fuels (Johansson 1999). However, as long as governments do not succeed in fully ‘inter­nalizing’ such external costs, they will have very little impact on economic decision making. So here, only costs borne by the producer will be considered. Figure 1.3

shows cost estimates for per-litre, fossil-fuel-based transport fuels and the energetic equivalent thereof for varieties of bioethanol in 2006.

What have also emerged are major regional differences in biofuel production costs, probably linked to differences in costs of land and labour and yields of feedstocks. This is shown by Fig. 1.4, which gives biodiesel production costs for 2006 while not taking account of external costs. All prices in Fig. 1.4 refer to bio­fuels from terrestrial plants. Estimates about the costs of large-scale production of biodiesel from algal oil are in the order of US $2.90 per litre (Chisti 2007), whereas transport biofuels from cultivated macroalgae would even be more expensive as the price range of the latter is more in line with their use as a delicacy (Neushul and Badash 1998; Buschmann et al. 2001). The cost of biodiesel made from used cook­ing oil and animal fats has been estimated at about US $0.22-0.74 per litre (Johnston and Holloway 2007; Canakci and Sanli 2008; Royal Society 2008).

In Brazil, as indicated by Fig. 1.3, during 2006, ethanol from sugar cane could compete with fossil-fuel-based transport fuels, but in the USA and the European Union, ethanol prices in 2006 were such that they were not competitive with gaso­line when external costs of fuels and fuel production were not included. By mid­May 2008, the situation was changed. Then, when costs were compared, corn-based ethanol in the United States was competitive with fossil gasoline (Westhoff 2008).

Figure 1.4 shows that in 2006, biodiesel from vegetable oil produced in Europe or the USA was not competitive with fossil-fuel-based transport fuels, but that in Malaysia and Indonesia, it was competitive when external costs were not included. Costs for different types of biofuel partly reflect differences in maturity of the pro­
duction process. The relatively low price for sugar-cane-based ethanol partly reflects a long learning curve (Goldemberg et al. 2004), though there is still scope for ad­ditional cost reduction (Hayes 2008; Macedo et al. 2008). Ethanol from lignocellu — lose is in an early stage of development, and there is much scope for cost reduction (Hayes 2008). It has been claimed that lignocellulosic ethanol can ultimately be­come competitive with ethanol from corn (Frederick et al. 2008; Lynd et al. 2008).

Both mineral oil and biomass prices are subject to change, and this may strongly affect the relative attractiveness of biofuels. For instance, Brazilian ethanol produc­tion did well when oil prices were relatively high, but demand slumped when such prices were low. It is often argued that mineral oil prices will in the future probably remain relatively high, which seems to bode well for the competitive position of bio­fuels. However, experience shows that predictions as to when biofuels become com­petitive with fossil fuels are subject to a major uncertainty — the prices of feedstocks. This is exemplified by the situation in 2008. Crude oil prices temporarily achieved price levels in the order of greater than US $100 per barrel, but biofuel feedstock prices also rose sharply. So, for instance, in 2008 and dependent on feedstock, the biodiesel unit price was 1.5-3 times higher than that of mineral-oil-derived diesel (Canakci and Sanli 2008).

For feedstocks that may also serve as a basis for food, major changes in prices are well known from the past. For instance, coconut oil prices varied by more than a factor of seven over the last 40 years (Cloin 2007). The nominal (US $) price of vegetable oil changed by about a factor of two in the 1997-2000 period, and the nominal (US $) price of wheat increased by about a factor of two between 1999 and 2006 (OECD-FAO 2007). And over the February 2007 to February 2008 period, the price of palm oil roughly doubled (www. palmoil. com). From early 2006 to early 2008, the price of US corn went from US $87 per metric ton to US $217 per metric ton (Tyner 2008). Price volatility may increase due to climate change (Eaves and Eaves 2007; Lobell et al. 2008).

High feedstock prices have a strong impact on biofuel prices. In the mid-1990s, the cost of biodiesel feedstock was 60-75% of the total cost of biofuel, and by 2008, this was 85%, with a $0.20 per litre biodiesel price increase when the feedstock price increased by US $0.22 per kilogram (Canakci and Sanli 2008). Similar changes occurred for starch — and sugar-based alcohols (Claassen et al. 1999; Qureshi and Blaschek 2001; Huber et al. 2006; Demirbas 2007; Koizumi and Ohga 2007; You et al. 2008). Furthermore, changes in prices of by-products do not necessarily favour the profitability of biofuel projects. Mainly due to expanding biodiesel production, a glycerol glut has emerged, which has negatively affected glycerol prices (Willke and Vorlop 2004; Yazdani and Gonzalez 2007). In 2007, glycerol prices were low­ered to a level well below that previously used in the calculation of biofuel prices (e. g. Francis et al. 2005; Huber et al. 2006). That co-products of biofuel production may be subject to change may have consequences for prospective biofuel prices. For instance, the relatively low price for microalgal biodiesel suggested by Huntley and Redalje (2007) is dependent on the current high value for the co-product astaxan — thin, but the price of astaxanthin may plummet if the production of algal biodiesel were to expand greatly (Vasudevan and Briggs 2008).

Further rapid expansion of biofuel production has also been argued to contribute to relatively high prices for the major commodities from which biofuels are made: crops for vegetable oil, starch and sugar (Runge and Senauer 2007; Daschle et al. 2007; Naylor et al. 2007). Actual predictions about future prices for vegetable oil, starch and sugar crops are extremely variable (Naylor et al. 2007). So firm predic­tions as to the relative future costs of fossil and biofuels, if based on the crops from which they are currently largely made, are hard to make. However, when biofuel production from food crops becomes large scale, prices of crops that serve as major biofuel feedstocks are expected to follow the price of crude oil, when corrected for the energy content of the biofuel (Naylor et al. 2007; Westhoff 2008).

It has been argued that the situation will be different when lignocellulose is used as a basis for transport biofuel production. Here, estimates of feedstock costs are often in the order of 20-33% of total operational costs when feedstocks are currently ‘wastes’, while processing costs usually are usually in the 70-80% range (Dien et al. 2003; Huber et al. 2006; Lin and Tanaka 2006; Solomon et al. 2007; Dale 2008). In the case of specific wastes, the share of feedstock costs in operational costs may even be lower. Joelsson and Gustavsson (2008) have, for instance, argued that a synthesis of transport biofuels based on the gasification of black liquor in the paper industry is competitive with mineral oil when the price of crude oil is at least US $40 per barrel. Black liquor is a co-product of paper that is relatively rich in lignin. The gas can be used for powering the paper plant and the production of transport biofuels such as methanol and dimethylether. In the case of crops grown as lignocellulosic feedstocks, the share of feedstock costs in operational costs may be higher than in the case of feedstocks that are currently wastes. Borgwardt (1999) considered lignocellulosic ethanol production with switchgrass or hybrid poplar as a feedstock and found that the feedstock cost was nearly 60% of operational costs.

Also, whether the current low (zero or even negative) costs of wastes and the relatively low costs of other lignocellulosic feedstocks can be maintained when they turn out to be good feedstocks for transport biofuel production is very doubtful. Indeed, in the long term, it seems likely that in this case, biofuels will follow the cost of competing fossil fuels, when corrected for differences in energy content (Naylor et al. 2007). When lignocellulosic biofuels turn out to be competitive, this may offer scope for substantial prices to be paid for what is currently considered a waste.

Still, it has been argued that as there are many sources of lignocellulose, it may well be that the price of feedstocks will be more stable than in case of starch or oil crops. This, however, is not necessarily relevant for production units turning out lignocellulosic biofuels. These may well restrict themselves to a limited range of feedstocks. Both in the case of enzymatic production and in the case of gasification, one would expect that production units, as they will be built in the near future, would be fit for a limited part of the broad range of lignocellulosic materials (e. g. Nathan 2007; Hayes 2008; Olofsson et al. 2008). On the other hand, it may well be that further technological development may allow for the use of broader ranges of lignocellulosic feedstocks.

Capital costs for converting lignocellulosic biomass into biofuel will be much higher than the capital costs for, for example, starch-based ethanol (Nathan 2007; Rotman 2008). Also, the operational costs of current enzymatic ways to produce lignocellulosic transport fuels are relatively high, even when currently available op­tions for cost cutting and increasing the expected credit for co-products are imple­mented. For instance, in the case of enzymatic conversion, such costs are estimated to be greater than US $0.60 per litre (Sassner et al. 2008), whereas the 2006 costs for ethanol from Brazilian sugar cane were US $0.28-0.31 per litre (see Fig. 1.3). So, much reduced operating costs and increased yields would seem essential to the long­time financial viability of biochemical conversion of lignocellulose into ethanol. Overcoming the recalcitrance of cellulosic biomass, lower pre-treatment costs and lower costs of enzymatic conversions are priorities in this respect (Wyman 2007). Whether further research will indeed lead to much lower costs is an open question.

As to the prospects for future cost reduction of non-hydrolytic/fermentative ways to convert lignocellulosic biomass into transport fuels, the following may be noted. Some of the processes proposed for converting lignocellulose into transport fuels, such as the processes to convert synthesis gas into transport fuels, have been well researched and developed (Huber et al. 2006; Haryanto et al. 2007). However, gas­ification of biomass has only been subject to limited research, and it would seem that much can be done to optimize gasification of the wide range of feedstocks available (Nathan 2007; Wang et al. 2008b). In the field of gasification, there also would seem to be scope for cost reduction linked to technological developments such as mem­brane separation, supercritical water gasification and better control technology for tar, char and ashes (Han and Kim 2008; Haryanto et al. 2007; Wang et al. 2008b). The production of methane by anaerobic conversion of biomass is a well-developed technology, but scope for cost reduction and improvement of efficiency in the case of the conversion of lignocellulosic biomass to CH4 may still be substantial (Bagi et al. 2007; Boijesson and Mattiasson 2008; Rodriguez et al. 2008).

If crop prices remain high, it may well be that, while excluding external costs, prices for many road transport biofuels may remain higher than fossil fuels in the near future. The higher cost of biofuels in the past has led to government policies favouring the application of biofuels. For the long-term viability of transport bio­fuels, however, it would seem unlikely that they can be more expensive than com­petitive fossil fuels. This may have a strong selective effect on production processes and producer countries.