When vegetation is cleared to make way for biofuel cropping, there may be sig­nificant emissions of non-CO2 carbonaceous greenhouse gases, including CO and hydrocarbons. In practice, these gases may add, in CO2 equivalents, 10-20% to the emission of CO2 only (Reijnders and Huijbregts 2008a). Also, if compared with forested land, arable land with annual biofuel crops becomes a reduced sink for CH4 (Powlson et al. 1997). Long-term cultivation has been shown to reduce CH4 uptake and oxidation by soils by 85% in a temperate setting. This may correspond to a reduction of the soil sink for CH4 in the order of 100-200 kg CO2 equivalent ha-1year-1. And the nature of cultivation also matters, with synthetic NH4 fertil­izer completely inhibiting CH4 oxidation, whereas manure has no inhibitory effect (Powlson et al. 1997). When organic wastes of biofuel production are anaerobically converted in open ponds or in dumps, there may be very large emissions of methane (Reijnders and Huijbregts 2008a).

Terrestrial plants vary widely in their yearly yields per hectare. Yields are depen­dent on insolation, temperature, the presence of nutrients and water and the nature of plants (Coombs et al. 1987). In natural ecosystems on average, the efficiency of photosynthesis in converting solar energy into plant material is usually in the or­der of 0.1-0.3% (Mezhunts and Givens 2004; Rosing et al. 2006). In the case of cultivated plants, higher conversion efficiencies are achievable. The highest yields are usually achieved in experiments under ‘excellent’ conditions that are highly con­ducive to plant growth. In large-scale commercial cultivation, yields are much lower. In the following, we will use data from large-scale cultivation, as this should be the basis for substantial feedstock production. As there is a tendency of gradual yield increases over time, such data may be biased in favour of crops that have a long tradition of large-scale cultivation. After a similar history of cultivation, the yields of relatively new crops that may serve as biofuel feedstocks such as Miscanthus and switchgrass may well be substantially higher than those that will be presented here.

In practice, the C4 plant sugar cane is relatively efficient in converting solar en­ergy into biomass (Sinclair and Muchow 1999). In subtropical areas, sugar cane may annually yield about 80 Mg per hectare of harvestable biomass (dry weight) when the conditions are excellent (Bastiaanssen and Ali 2003; Braunack et al. 2006). Av­erage sugar cane yields during the mid 1990s in Brazil were about 36.8 tons of biomass (dry weight) ha-1year-1 (Kheshgi et al. 2000). Under excellent condi­tions, another C4 plant, Miscanthus, may yield annually up to about 30-60 Mg of dry weight harvestable biomass per hectare (Long et al. 2006; Heaton et al. 2008a), but more commonly, yields are in the range of 10-13 Mg aboveground dry weight biomass ha^year-1 (Lemus and Lal 2005; Christian et al. 2008).

Oil palms in Southeast Asia yield about 20 Mgyear-1 ha~1 as fresh fruit bunches (dry weight) (Reijnders and Huijbregts 2008a). For sugar beets, good yearly dry weight yields of biomass from large-scale commercial cultivation are also in the or­der of 20 Mgha-1 (Sahin et al. 2004; Tzilivakis et al. 2005). For eucalyptus, yearly biomass yields per hectare tend to be in the order of 10-20 tons (Sims et al; 1999; van den Broek et al. 2001). Yearly dry biomass yields of large-scale cultivation un­der good conditions for switchgrass are in the order of 10-15 Mgha-1, for willow 9Mgha-1, and for poplar 11Mgha-1 (Lemus and Lal 2005; Heaton et al. 2008a). Total yearly (dry weight) aboveground biomass accumulation per hectare in the USA is in the order of 17-18 Mg for corn (Heaton et al. 2008a), and under good conditions, 10-11 Mg for wheat (world average is 5.5 Mg; Wright et al. 2001), in the order of 9 Mg for peas and 4-5 Mg for canola (Lemus and Lal 2005; Malhi et al.

2006) . High yields of photosynthesis in practice usually depend on substantial in­puts of synthetic nutrients derived from non-renewable natural resources (Samson et al. 2005). Sustainable yields that can be achieved when only recycling nutrients that are present in biomass tend to be much lower as will be discussed in Chap. 3 (also Pimentel et al. 2002; Reijnders 2006). Table 2.1 shows the overall energy con­version efficiency (taking account of inputs of fossil fuels) for a variety of crops with relatively good yields.

The overall solar energy conversion efficiencies in Table 2.1 are below 1% and range roughly between 0.15% (for rapeseed/canola) and 0.9% (for sugar cane). For comparison, a percentage is added for sustainably grown wood in Western Russia (Nabuurs and Lioubimov 2000). In this case, the conversion efficiency is about 0.05%.

There have been efforts to improve the solar-energy-to-biomass conversion by transgenic approaches. These have focused on increasing the net carboxylation ef­ficiency of 1,5-biphosphate carboxylase and the introduction of enzymes charac­teristic for C4 plants in C3 plants (Heaton et al. 2008; Raines 2006). So far, such efforts have not led to a substantial improvement in the conversion of solar energy to biomass (Raines 2006).

Biodiversity of areas under annual crops, perennial grass crops or plantations may be different from that of the nature that they replace(d). To the extent that transport biofuel production leads to losses of natural habitats, there is apparently a consistent negative effect on species diversity (Fahrig 2003). The effect appears to be stronger at higher trophic levels (Dobson et al. 2006).

In practice, expansion of biofuel production is often associated with cutting forests, and this will probably also hold for future expansion (Gurgel et al. 2007). In such cases, land with annual biofuel crops functions in a number of other ways dif­ferently from a forest. Vegetation influences surface roughness, which in turn may alter wind, turbulence and moisture convergence, and forest and cropland may be significantly different in this respect (Notaro et al. 2006). Change of forest into crop­land in semi-arid climates may increase dust emissions, which in turn may change radiative forcing (Betts 2007). The albedo of arable land, on which biofuel crops are grown, is often different from the albedo of land under natural vegetation, and actual differences may also be crop dependent (Gustafsson et al. 2004; Schneider et al. 2004; McPherson 2007). The albedo is a measure of the reflection of solar ra­diation by the earth’s surface (including vegetation), which in turn is a determinant of net radiation. Biofuel crops tend to be shorter and have less foliage than forests, and so the surface albedo of arable land with biofuel crops tends to be higher than the albedo of a forest. The release of water to the atmosphere by evapotranspira — tion linked to biofuel crops may also be different from natural vegetation or other crops (Gustafsson et al. 2004; McPherson 2007). The difference in evapotranspira — tion may have impacts on soils and the atmosphere. An example of the former is the large-scale salinization of soils in Australia, following the replacement of native woody vegetation by crops. The change in evapotranspiration caused groundwater tables to rise, and the rising water mobilized salt (Folke et al. 2004).

As to the atmospheric effect of changed albedo, it may be noted that, besides net radiation, evapotranspiration influences local climate, including temperature (Schneider et al. 2004; Notaro et al. 2006; McPherson 2007), and there may also be an impact on precipitation (Liu et al. 2006). Moderate and local deforestation may lead locally to enhanced rainfall (Malhi et al. 2008). Also, in tropical regions, replacement of forest by annual biofuel crops may cause warming, mainly due to a decrease in evaporation and cloud cover (Betts et al. 2007). In temperate regions, the overall effect of turning forest into cropland may lead to regional cooling that may be partly offset by increased aerosol concentrations (Betts 2007).

When changes in vegetation are widespread, there may be an impact on mesoscale climate (McPherson 2007; Liu et al. 2006). For instance, a simulation study assum­ing a major reduction in forested area and a major increase in annual cropping in Amazonia found a substantial reduction in precipitation over Amazonia and a posi­tive radiation forcing at the top of the atmosphere linked to increased airborne dust (Betts 2007). Another model study suggests that removal of 30-40% of the Ama­zonian rainforest may push much of Amazonia into a permanently drier climate (Malhi et al. 2008). Such a climate change may lead to additional damage to the current Amazonian rainforest (Nepstad et al. 2008).

Mesoscale changes in climate may have knock-on effects. For instance, mesoscale changes in Amazonia may in turn have effects on precipitation in the Northern Hemisphere and across the rest of South America (Malhi et al. 2008). A simula­tion regarding deforestation in Indonesia has suggested that the surrounding ocean surfaces may be warmed and that this may have a widespread impact on atmospheric circulation (Delire et al. 2001).

It has furthermore been found that cultivation of land for arable crops reduces the uptake rate of CH4. In an experiment in Rothamsted, it appeared that extended (150-year) cultivation of arable land decreased CH4 uptake and oxidation by 85%, if compared to that in the soil under woodland (Powlson et al. 1997). Also, the ‘leakiness’ of nutrients and, more in general, soluble minerals is much increased in agricultural land, if compared with soils under native forest cover (Williams and Melack 1997). Moreover, forests are better in the capture of nutrients (such as P and N) from air than annual crops (Lawrence et al. 2007). So the function of annual crops on arable land is much different from that of forests.

But how about the functioning of tree plantations, which would seem rather sim­ilar to forests? Differences in biodiversity between oil palm plantations and the forests which these plantations replaced have been analyzed by Danielsen et al. (2008). They found that total species richness of vertebrates, invertebrates and flora on oil palm plantations was impoverished. Similarly, Lindenmayer and Hobbs (2004) found reduced faunal diversity in Australian eucalypt plantations if com­pared with native forests. Barlow et al. (2007) studied differences between native forest and eucalypt plantations in Brazil for 15 taxonomic groups and found overall diversity reduced in plantations, with major differences in biodiversity change be­tween the taxa. They pointed out, however, that faunal diversity can be improved by integrating elements of original biota into plantations, by modifications to plantation management (e. g. regarding harvesting and thinning) and by having extensive areas of remnant native vegetation adjacent to plantations.

The functioning of tree plantations has been found to be different from the func­tioning of primary forests. This has consequences for ecosystem services. Com­parison of tropical tree plantations with secondary tropical forests showed, for in­stance, that in secondary forests, root densities and nutrient concentration in roots were higher and root penetration was deeper in forests than in plantations (Lugo 1992). This allows secondary forests to better recapture nutrients which become available by mineralization and could otherwise be lost to water and the atmosphere. In the highly productive plantations, which are important for achieving a large fu­ture biofuel supply (e. g. Hoogwijk et al. 2003), there is furthermore an intensive use of herbicides and other pesticides (Tuskan 1998; Robison and Raffa 1998) which will cause increased leakiness regarding nutrients and will limit nitrogen fix­ation.

Primary forests are more efficient than plantations in protecting watersheds, in reducing peak flows, soil erosion and nutrient emissions, in maintaining good water quality, in stabilizing local climate and in generating OH radicals that are impor­tant cleansing agents in the atmosphere (Hartshorn 1995; Perry 1998; Daily 2000; Monson and Holland 2001; Brauman et al. 2007; Wallace 2007). The risk of pests tends to be increased in plantations if compared with native biota, but such risk may be reduced by integrating patches of native vegetation into plantations (Lin — denmayer and Hobbs 2004). Li et al. (2005) found that secondary forests may stock more long-term soil organic carbon than plantations in the wet tropics. This is rel­evant to the sustainability of plantations (cf. Chap. 3) and to the net emission of greenhouse gases from plantation-derived biofuels (cf. Chap. 4). Still, differences in non-monetary ecosystem services between primary forests and arable land tend to be larger than those between forests and plantations (Brauman et al. 2007).

The effect of habitat loss on non-monetary ecosystem services also has world­wide aspects. These are closely linked to the biogeochemical cycles of relatively mobile elements, of which the carbon cycle is a good example. There are sev­eral links between natural biomass and worldwide atmospheric concentrations of CO2. Reduced C sequestration in ecosystems, which is often associated with ex­panded biofuel production (see Chap. 4), increases the atmospheric CO2 concentra­tion (Houghton 2007). It is furthermore likely that natural terrestrial biomass may give rise to a better negative feedback by fixing increasing amounts of CO2 when the atmospheric concentration thereof increases than most agro-ecosystems (Cao and Woodward 1998; Luo 2007). And, overall, natural systems are better at seques­trating carbon than agricultural systems (Vitousek et al. 1997). Moreover, when land use is changed back from agriculture to nature, the functioning of secondary forests in biogeochemical cycles such as the carbon cycle may for a long time be substan­tially different from primary forests (Grau et al. 2003).

Losses of natural terrestrial biomass are contributing significantly to the increase in the atmospheric CO2 concentration (IPCC 2001; Houghton 2007). This may also have monetary effects. An increase in atmospheric CO2 concentration will (ceteris paribus) on average increase costs of production and consumption in the world econ­omy (IPCC 2001; Stern 2007).

All in all, to the extent that transport biofuel production leads to habitat loss for living nature, this affects functions of nature that can be considered beneficial to mankind such as its contribution to a benign biological, chemical and physical environment and socio-cultural fulfilment (Vitousek et al. 1997; Wallace 2007).

There is also an option which may be called virtual biofuels. The pyrolytic produc­tion of ‘black carbon’ (also charcoal, biochar) from biomass has been advocated as an alternative to biofuels (Fowles 2007). Such black carbon would be added to soils, where it said to be ‘very stable’ and able to fulfil useful functions. This is argued to offset CO2 emissions (Lehmann et al. 2006; Fowles 2007; Mathews 2008a). Saun­ders (2008) has proposed to landfill purpose-grown biomass as a ‘virtual biofuel’, which he considers ‘more practical, economic and immediate’ than the use of actual biofuels from lignocellulosics. There are also ‘climate compensation schemes’ of­fered to users of transport (especially car and air transport) to offset their emissions of CO2. Planting trees tends to be major contributor to such schemes. The idea be­hind this is that the C emitted as CO2 into the atmosphere due to the burning of fossil fuels will be sequestered in biomass. For instance, it has been estimated that refor­estation of abandoned tropical land may lead to an aboveground C sequestration of approximately 1.4Mgha-1year-1 and a sequestration in soils of approximately

0. 4Mg C ha^year-1 over an 80-100-year period (Silver et al. 2000). Offsetting by forest conservation or reforestation leads to much lower costs for the alleged reduction of CO2 emissions from transport than the production of biofuels.

The obvious question about these proposals is: are virtual biofuels indeed a so­lution to the impact of fossil transport fuels on climate? This depends on the dura­tion of carbon sequestration in virtual biofuels. Before being used as transport fuel, mineral oil was destined to remain for many millions of years outside the carbon cycle in which biomass participates, and one may argue that to really offset the use thereof, C in black carbon, purpose-grown landfilled biomass and forests should also remain outside the biogeochemical carbon cycle for many millions of years or ‘in perpetuity’. Also, one may focus on CO2 emitted due to the consumption of fossil fuels. Full elimination of the effect of the CO2 emission from fossil transport fuels on the climate is expected to take a very long time. One quarter of fossil-fuel — derived CO2 remains airborne for several centuries, and complete removal may take 30,000-35,000years (Archer 2005; Hansen etal. 2008). So it may be argued that the sequestration in biochar, landfills or forests should at least be for many thousands of years.

Whether C sequestration for at least many thousands of years applies to biochar has been studied in a limited way (Eckmeier et al. 2007; Wardle et al. 2008). There have been reports about the decomposition of biochar and oxidation of the aromatic backbone of biochar, partly depending on the production procedure (Lehmann et al. 2006; Steiner et al. 2007). However, there is also evidence that carbonblack particles may persist in soils over thousands of years (e. g. Carcaillet and Talon 2001; Long et al. 2007), allowing for the possibility that part of buried biochar sequesters C for a very long time indeed (Lehmann et al. 2006). Secondly, there is evidence that biochar may have an effect on soil biological processes: experimental data suggest that this effect may result in loss of native soil carbon (Wardle et al. 2008). As it stands, it would seem likely that the net carbon sequestration by biochar is partial and may show a decrease overtime.

Landfilled purpose-grown biomass will not for thousands of years or in perpetu­ity remain outside the biogeochemical carbon cycle. Landfilled biomass will largely be converted into CH4 by anaerobic processes. This has the added disadvantage that CH4 is, over a 100-year period, 21 times a more potent greenhouse gas than CO2, the main carbonaceous product of biofuel combustion (Barlaz 2006). The rate of con­version of biomass into methane is dependent on a variety of factors, including tem­perature, lignin content, moisture and pH, and such conversion is often a matter of decades (Barlaz 2006, Themelis and Ulloa 2007). Capture of CH4 emitted by land­fills is an option but has in practice limited efficiency (Themelis and Ulloa 2007).

There is also a major snag with planting forests as virtual biofuels. Storage of carbon in trees should be guaranteed for many thousands of years. At the level of individual trees, this is impossible, as storage in perpetuity or for many thousands of years is well beyond the maximum lifespan of tree species. And guarantees at the level of forests are also a problem. Current ‘climate compensation schemes’ have guarantees for forests that do not exceed a hundred years. Even this guaranteed timeframe is questionable in view of (increasing) risks that forests may be destroyed by wildfires and extreme weather events such as storms and droughts (Kirilenko and Sedjo 2007; Gough et al. 2008; Nepstad et al. 2008). The social arrangements safeguarding forests are also unlikely to persist for many thousands of years or in perpetuity. So it would seem that forests as virtual biofuels rather delay than fully offset the emission of CO2 from fossil transport fuels. There is also another problem with planting forests to limit the increase in atmospheric CO2. This has been called ‘leakage’ (Sathaye and Andrasko 2007; Ewers and Rodrigues 2008). When forestry projects are established, people dependent on that area may move elsewhere, where they may reduce C stocks. There has been a series of case studies regarding this phenomenon. In some cases, high levels of leakage have been demonstrated. For instance, Boer et al. (2007) studied forestation projects in the Jambi province of Indonesia and found that reductions in C stock due to leakage exceeded gains in C stock linked to forestation over a 10-year period. Other forestation projects showed lower leakage, and worldwide, an average percentage of about 50% leakage seems to be associated with forestation projects (Sathaye and Andrasko 2007). So forestry as a virtual biofuel is subject to major problems when it comes to full compensation of fossil fuel consumption.

Apart from problems with C sequestration over at least thousands of years, a main problem is that virtual biofuels clearly would not solve the problem of dependence on mineral oil, as one cannot drive or fly on virtual fuels. For this reason, the ‘real’ biofuels that are used or have been proposed for use as transport biofuels will be the main topic of this book.

Obtaining biofuels derived from plantations and forests heavily relies on including parts of trees, such as crowns, that have relatively high concentrations of nutrients (Manley and Richardson 1995; Perry 1998; Sims and Riddell-Black 1998; Pare et al. 2002; Rytter 2002). In the relatively young trees that characterize plantations, nutri­ent concentrations are, moreover, higher than in older trees (Rytter 2002). The use of feedstocks with high nutrient levels adds to losses of nutrients associated with common consequences of tree-harvesting practices such as erosion, increased leach­ing of nutrients and lowered rates of N-fixation by leguminous understory plants (Hamilton 1997; Heilman and Norby 1998; Richardson et al. 1999; Bernhardt et al.

2003) .

Overall losses of nutrients may well have an impact on future productivity. For instance, studies of whole tree harvesting, with branches, tops and needles used as biofuels, as it is currently practiced in Sweden, show deficits in base cations (K, Mg and Ca) (Akselsson et al. 2007). Akselsson et al. (2007) suggest that compensatory fertilization with K, Mg and Ca is necessary to keep forestry sustainable. Studies in French forests show that the budget for Ca and probably Mg is negative, as­suming a 60-year rotation time and a conservative scenario of biomass harvesting (Ranger and Turpault 1999). In the southern USA, P deficits have been noted (Pit­man 2006). In Scandinavian forests, thinning involving whole tree removal has been found to cause significant reduction in stand volume increment linked with nutrient loss (Nord-Larsen 2002). And in tropical dry forests, repeated harvesting may well lead to reduced primary production due to a reduced capture of P from air (Lawrence et al. 2007). Keeping forest soil concentrations of nutrients in a steady state while removing feedstocks for biofuel production with relatively high concentrations of nutrients may, in the absence of nutrient amendments, force the application of long rotations or even an end to harvesting. In Sweden, harvesting trees from nutrient — poor soil is in fact discouraged (Manley and Richardson 1995).

There is also the option of the recycling of nutrients to soils. The extent to which this can be done depends on the use of biomass. For instance, fixed N is almost totally lost during combustion (Sander and Andren 1997). On the other hand, it may be expected that fixed N can, to a large extent, be conserved in the production of methane and ethanol from biomass. In the context of ethanol production from switchgrass, Anex et al. (2007) have proposed a process that may recover about 78% of the fixed N input, which then may be recycled.

Nutrient elements other than N tend to be largely conserved in (fly and bottom) ash during burning and can be retrieved when proper controls are in place. In power plants, biomass is often co-fired with coal, and this will lead to ashes that are often considered unacceptable for nutrient application in forests or on arable soils (Reijn — ders 2005). When only biomass is burned, this may be different. It has been pro­posed to return ashes, especially for their base cations, to forest soils after burning forest-derivedbiomass (Hanell and Magnusson 2005; Pitman 2006; Ozolineius et al.

2007) . Pettersson et al. (2008) have suggested extracting phosphate from ashes for reuse as a nutrient. Also, digestate remaining after anaerobic conversion of biomass to methane, and fermentation residue remaining after converting lignocellulosic biomass into ethanol, may be returned to soils (e. g. Zwart et al. 2007; Reijnders

2006).

As yet, however, nutrient recycling is very limited. Ashes from burning biofuels are not usually returned to soils used for biofuel production, but largely diverted to other destinations, such as landfills (Reijnders 2005; Saikku et al. 2007). There may also be complications in returning nutrients. Most studies have focussed on the recycling of ash, and especially the recycling of base cations such as Ca2+, Mg2+ and K+. Additions of such ashes to forests on mineral soils have shown disappoint­ing effects, which have been linked to N deficiency (Augusto et al. 2008). It has been found that the chemistry of base cations in ashes tends to be different from the original chemistry in soils, and so are local concentrations of base cations in the soil after applications of ash. It has been argued that such differences may be limited by keeping temperatures between 600 and 9000 C during burning and using granulated ashes (Pitman 2006). Furthermore, it has been advised to apply ash to adult stands and not to seedlings (Augusto et al. 2008).

Experience with return of wood ash for its base cations to forest soil shows sub­stantial side effects, for instance, on the levels of aluminium in soil solution and an increased soil emission of CO2 (Maljanen et al. 2006; Ring et al. 2006), sup­pression of denitrification (Odlare and Pell 2009) and especially in the case of high fixed N presence in soils, increased leaching of nitrate (Pitman 2006). Reductions of Mn levels in biomass have been associated with wood ash recycling (Augusto et al.

2008) . Moreover, hazardous elements, such as lead and cadmium, and hazardous organics, such as polycyclic aromatic hydrocarbons, may be present in ashes in sub­stantial amounts. It has been found that even in the apparent absence of substantial anthropogenic contamination, levels of heavy metals in combustion ashes may be remarkably high (Reimann et al. 2008). For example, combustion ashes from South Norwegian birch and spruce wood ashes contained up to 1.3% lead and 203 mgkg~1 cadmium (Reimann et al. 2008). Johansson and van Bavel (2003) and Enell et al.

(2008) looked at the presence of polycyclic aromatic hydrocarbons (PAHs) in wood ash and found that the concentration thereof in a substantial number of cases ex­ceeded the standard applicable in Swedish forests of 2 mgkg-1 of 16 PAHs.

Thus, high concentrations of inorganic and/or organic contaminants may repre­sent a barrier to the sustainable recycling of nutrients. This may require input con­trols for burners (e. g. excluding wood with unacceptable levels of heavy metals), facilities for burning that minimize the formation of hazardous compounds such as polycyclic aromatic hydrocarbons and chlorinated dioxins and/or treatment of ashes to eliminate hazardous compounds. As pointed out in Sect. 3.2, in the case of wastes from lignocellulosic ethanol production, phenolic compounds, ionic compo­sition and pH should be controlled, and the flow of heavy metals should be limited. All in all, it is unlikely that recycling of nutrients after biofuel processing and use can or will be as efficient as nutrient recycling in natural systems.

Regarding plantations, processes such as natural weathering and symbiotic N-fix — ation that are important in providing undisturbed forests with nutrients may well be less productive (Perry 1998). Intensive site preparation, common in plantations, in­volving burning biomass may negatively affect productivity as it leads to the volatil — isationof nutrient N (Perry 1998). Because root systems in short rotation plantations may be less extensive than in undisturbed forests, the leakage of nutrients may well increase (Ong and Leakey 1999). Harvesting practices on plantations tend to in­crease denitrification and leaching, which may lead to increased nutrient deficits (Hamilton 1997; Heilman and Norby 1998). It is likely that erosion on plantations will exceed the value range 0.004-0.05 Mgha-1year_1 that is found in undisturbed forests. Still, by judiciously planting and harvesting trees, it may be possible to keep erosion rates below the level of 1 Mgha-1year_1 (Pimentel et al. 1997b).

Overall, as rates of biomass harvesting also tend to be much higher on plantations than in forests, large deficits in nutrients are to be expected. For instance, regarding aspen-for-fuel plantations, Rytter (2002) estimated the yearly deficit per hectare of N at 30 kg, for P at 4 kg, for Ca at 30 kg, for Mg at 4 kg and for S at 2.5 kg. Indeed, the productivity of growing short rotation trees strongly depends on external nutrient inputs (Adegbidi et al. 2001). Remarkably, current human activity, especially in in­dustrialized countries, has led to increased environmental fluxes of wasted nutrients such as S, N and P, which reach soils via air and/or water (Smil 1991; Kvarnstrom and Nilsson 1999). For instance, in North America, unintended N depositions on soils may be up to 53 hgha^year-1, and in Europe up to 115 kg N ha-1year-1 (Heilman and Norby 1998). It is probable that the unintentional addition of nutri­ents to soils has been important to maintaining productivity in the absence of inten­tional nutrient amendments. For aspen plantations in Southern Sweden, it has, for instance, been calculated that deposition exceeds the yearly S deficit (2.5 kgha-1) and may cover a substantial part of the N deficit. In forests in Southern Sweden, there may still be accumulation of N with current biofuel harvesting practices (Ak — selsson et al. 2007). However, such unintended additions of nutrients to soils are not designed to fit all actual deficits in nutrients. For instance, though S deficits are com­pensated for in Nordic spruce forests, nutrient deficiencies for P, K and B still occur (Rytter 2002). Unintended additions of nutrients to soils may also cause new prob­lems, especially when unintended nutrient additions are relatively high. They may well contribute to long-lasting eutrophication and acidification and deterioration of ground water quality (Galloway et al. 2008).

As pointed out in Chap. 1, much of the impetus for the development of transport biofuel production has come from governments. Looking back on the results of government intervention, as discussed in the previous pages of this book, this has been a very mixed blessing. So, are there suggestions for government policy which may be conducive to more beneficial results?

The growth of biofuel production and consumption for automotive transport is now the subject of a lively debate. This debate has led to revisions of transport biofuel policy in a number of countries, and it is likely that this debate will have further impacts on the development of biofuel production. Major items in this debate are the following.

Energy Security

To the extent that transport biofuels are advocated to provide for energy security, it has been stated that their potential may be very limited. Eaves and Eaves (2007) have argued that devoting 100% of US corn to ethanol, while correcting for fossil fuel inputs, would displace 3.5% of gasoline consumption, ‘only slightly more than the displacement that would follow from properly inflated tires’. Moreover, they have pointed out that historical US corn yields have shown considerable volatility, with corn yields about once every 20 years more than 30% lower than on average, which is not conducive to national energy security (Eaves and Eaves 2007). On the other hand, it has been argued that worldwide biofuels may end the dependence of transport on mineral oil, or even all fossil fuels. As pointed out above, de Vries et al. (2007) suggest that by 2050, up to about 300 EJ (= 300 x 1018 J) of liquid biofuels may be produced worldwide, mainly on abandoned agricultural soils, which would, as pointed out in Sect. 1.1, in all probability be sufficient to power transport.

The data presented in Sect. 4.3 for the four maj or determinants of net greenhouse gas emissions will be used in this section to calculate life cycle emissions for specific biofuels and categories of transport biofuels.

4.4.1 Net Greenhouse Gas Emissions for Specific Biofuels

Example 1: Electricity from Woody Biomass Instead of from Fossil Fuels

Substituting fossil fuels with woody biomass from sustainably managed forests in electricity production will strongly reduce the life cycle greenhouse gas emissions per kWhe, as indicated by Fig. 4.2.

However, when there are changes in C sequestration due to forestry, emissions may deviate in a major way from the values in the last column of Fig. 4.2. Nechodom et al. (2008) have given a far more favourable assessment of electricity generated by burning woody biomass from forest remediation in California, as such remediation is supposed to reduce C losses due to fires, whereas use of woody biomass associ­ated with clear cutting forests will lead to greenhouse gas emissions that are much

larger than those associated with the fossil fuels given in Table 4.1 (Reijnders and Huijbregts 2003). There may also be temporal deviations from the woody biomass value given in Fig. 4.2 when the requirement is that cutting trees should be bal­anced by planting new trees. In this case, there may be a period of up to 20-40 years after harvesting in which forests are net sources of atmospheric carbon, and even longer periods before initial C losses are fully compensated (Reijnders and Huijbregts 2003).

For some applications, biomass as it is produced in solar energy conversion may be used as such. This applies, for instance, to the generation of electricity, which in turn may be used for electrical traction. However, diesel or Otto motors or fuel cells need the use of specific biochemicals (transport biofuels) such as specific alcohols and acylesters, as discussed in Chap. 1. This has an impact on the efficiency of solar conversion. Only a part of the biomass originating in solar energy conversion can be turned into such chemicals. It may be that part of the biomass that cannot be con-

verted into the biochemicals needed is used to power the production from biomass of specific biochemicals. It may also be that a part of the original biomass emerges from processing as waste. Furthermore, in many processes generating biochemicals from biomass, there is an input of fossil fuels that is to be taken into account when determining overall conversion efficiencies. Figure 2.1 gives estimated efficiencies for some conversions of biomass into transport biofuels.

Table 2.2 shows solar conversion efficiencies for a number of biofuels from ter­restrial plants. In this case, the allocation has been done on the basis of energy content of marketable products.

The efficiencies in the last column of Table 2.2 are typically lower than the effi­ciencies shown in Table 2.1. Most of them are below 0.2%. For ethanol from Euro­pean wheat starch, the efficiency is 0.024-0.03%, and for biodiesel from European rapeseed, it is approximately 0.034%. Apart from Jatropha, which has quite an un­certain conversion efficiency, the best efficiency in Table 2.2 is for ethanol from switchgrass, with ethanol from sugar cane coming second. However, it was assumed in this table that in the case of sugar cane, only sugar is to be converted into ethanol. If also a substantial part of the lignocellulosic aboveground biomass of sugar cane is converted into ethanol, sugar cane may as efficient as or better than switchgrass.

5.4.1 Cropping and Crop Harvesting Practices

There are a variety of aspects of cropping and crop harvesting practices regarding terrestrial biofuels that may impact biodiversity. The use of cropping systems which include a relatively wide crop genetic diversity may allow for more services in the fields of pest control and pollination than cropping systems that have a narrow ge­netic base (Tilman et al. 2006; Hajjar et al. 2008). And annual cropping systems that use cover crops may well be better in soil conservation than cropping systems that do not use such crops (Jarecki and Lal 2003). There may also be differences in the non-crop biodiversity of production systems. For instance, extensively managed perennial grass crops (e. g. Miscanthus, switchgrass, mixtures of prairie grasses) may allow for more invertebrate diversity than intensively managed annual crops, and willow coppice plantations may benefit some bird species (Anderson and Fergus- son 2006). Replacement of extensively used grasslands by arable land for biofuel cropping may negatively affect bird species that rely on such grassland habitats (e. g. Schleupner and Link 2008). Sage et al. (2006) compared bird populations in a short rotation willow coppice, used for biofuel production, and other arable crops in England and concluded that it was unlikely that the planting of willow coppice on unimproved farmland would lead to a conservation gain. However, planting willows in small blocks of different age classes and no harvest in summer would, in their view, benefit bird populations in short rotation willow coppice fields.

One likely future development to which increased transport biofuel production will contribute is intensification of agriculture (Goldemberg 2008; Searchinger et al. 2008; Sukhdev 2008). Intensified agriculture has a variety of effects on living nature. Higher production may provide more resources for a number of mammals, birds and insects. For instance, populations of bumblebees may increase in landscapes with intensive rapeseed cropping (Tscharntke et al. 2005). On the other hand, high inten­sities of nutrient and pesticide use tend to reduce biodiversity (Tilman et al. 2001a; Ptacnik et al. 2008). Intensification of agriculture may also decrease edge habitats such as hedges (Tscharntke et al. 2005). Intensified agriculture is furthermore often associated with lowering water tables (Tscharntke et al. 2005), and this may lead to changes in biodiversity. Also, increased irrigation, which is expected to contribute to intensified agriculture, is likely to have an impact on biodiversity by lowering the

availability of water to natural species. This is exemplified by the negative impact of increased irrigation on biodiversity and ecosystem services of wetlands and rivers in the European Union, the United States, China and Australia (Gerakis and Kalburtji 1998; Gleick 2003; Castaneda and Herrero 2008; Postel 2008). On average, the net effect of intensified agriculture is a decline in biodiversity among many different taxa (Liira et al. 2008). Such a decline may have a rebound effect on crop produc­tivity. For instance, it has been found that crop pollination from native bees may be at risk from agricultural intensification (Kremen et al. 2002).

Handling harvest residues may also matter to biodiversity. There are bird species which depend to at least some extent on harvest residues. Well known is the depen­dence of waterfowl on residues from harvesting in flooded rice fields (van Diepen et al. 2004), the consumption of harvest residues in the Mississippi delta by water­fowl (Gallagher et al. 2003) and the dependence of cranes on residues from corn harvesting in Northern Germany. Residue removal for biofuel production would in such cases reduce bird populations. In some cases, this may have knock-on effects. For instance, in rice-growing areas, waterfowl dependent on residues may be im­portant in maintaining productivity as they reduce weed pressure and pests and in­crease N cycling (Bird et al. 2000; van Diepen et al. 2004), and they may also serve as a food source. There is little information about the impact on living nature of the cultivation of algae. Open water seaweed farming near the coast of Zanzibar has, however, been found to be associated with less sea grass and reduced abundance and biomass of macrofauna (Eklof et al. 2005).