The best-studied macrophyte is the water hyacinth (Eichhornia crassipes) (Gass — mann et al. 2006; Gunnarsson and Petersen 2007). It has been found to produce up to 140 Mgha-1 year-1 of biomass (dry weight) (Gunnarsson and Petersen 2007). Two energetic applications of Eichhornia crassipes which may produce transport bio­fuels have been studied. The first is ethanol production from hemicellulose present in water hyacinths. A yield of 0.14-0.17 (g ethanol) (g dry weight)-1 has been re­ported (Mishima et al. 2008). However, studies of the overall energy efficiency of the production of ethanol from the water hyacinth have so far suggested that the en­ergy balance is negative (Gunnarsson and Petersen 2007). An alternative option is the anaerobic conversion of water hyacinthbiomass into CH4. Though the feasibility thereof has been demonstrated, the process is complicated, among other things by the floating behaviour of water-hyacinth-derived material (Malik 2007). Moreover, the water hyacinth is very effective in adsorbing pollutants (Gunnarsson and Pe­tersen 2007; Malik 2007), and these may interfere with, for example, the sustainable use of residuals (‘digestate’) remaining after anaerobic conversion. More limited re­search has been done regarding another invasive macrophyte: water lettuce (Pistia stratiotes L.), which has growth characteristics similar to water hyacinth (Mishima et al. 2008). The yield of ethanol from hemicellulose conversion is 0.15-0.16 g per gram of dry weight (Mishima et al. 2008); no study has been found regarding the overall energy efficiency of this conversion.

As pointed out in Chap. 2, current transport biofuels often have a cumulative seed — to-wheel fossil energy demand that tends to be smaller than the fossil fuels that they replace. The difference with fossil transport fuels is variable. The difference is probably about zero for CH4 from manure in NW Europe; is as yet unlikely to be positive for current algal biofuels; is smaller than 40% for ethanol from European wheat, biodiesel from European rapeseed or ethanol from US corn and is relatively large for palm oil biodiesel and ethanol from sugar cane, especially when processing is powered by agricultural residues. When the latter applies, the difference may, for example, become greater than 90% for ethanol from sugar cane (Macedo et al. 2008). The overall solar energy conversion efficiency of biofuels suitable for use in internal combustion engines is probably around 0.2% or lower, and expected yield increases per hectare are in the order of roughly 1% per year (cf. Chap. 2). This means that to displace substantial amounts of fossil fuels, land requirements for such transport biofuels are large, as also noted by Dukes (2003).

The contribution that biofuels can make to the national energy security of a coun­try depends on the magnitude of fuel demand and the land area available for sup­plying biofuel feedstocks. For a country such as Brazil, ethanol from sugar cane can make a significant contribution to national energy security. In the USA, the con­tribution of biofuels to national energy security is likely to be much smaller. The reasons for this are that, if compared with Brazil, per capita demand for transport fuels is larger, per capita land availability is lower and feedstocks such as corn and canola are less efficient converters of solar energy into biofuel than sugar cane. Eaves and Eaves (2007) have a point when they argue 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’. In fact, they may even have been too optimistic, because the actual US policy has been using a federal excise tax, making mixtures of conventional gasoline and ethanol cheaper than conventional gasoline, which has an upward effect on overall transport fuel use (Vedenov and Wetzstein 2008). Di­verting all 2007 US soybean cultivation to biodiesel production would cover ap­proximately 2% of US diesel demand, when corrected for fossil fuel inputs and assuming no effect on fuel prices (Bagajewicz et al. 2007; Reijnders and Huijbregts 2008b).

Larger displacements of fossil fuels while using the same area of land can be achieved when biomass is burned in power stations and used for electric traction, as the seed-to-wheel overall solar conversion efficiency thereof is higher than in the case of transport biofuels such as biodiesel and bioethanol, as indicated in Chap. 2. However, as explained in Sect. 1.6, such a strategy is dependent on a major change in social acceptance of plug-in vehicles. The potential for energy security through national transport biofuel supply is low for industrialized countries with high pop­ulation densities, such as Japan and the Low Countries in Europe. For instance, in the Netherlands, a 20% target for the share of transport biofuels in current transport fuel consumption would require an area of arable land that is roughly four to five times the size of current agricultural land in that country when ethanol from starch and sugar and biodiesel from vegetable oil are used and when a correction is made for the cumulative fossil fuel inputs in the biofuel lifecycles.

Not only the land area available, but also other factors may limit the extent to which countries may rely on domestically produced biofuel feedstocks for energy security. Climatic change may well have a negative impact on agricultural yields in the developing world (Jepma 2008). As pointed out in Chap. 3, water requirements for producing substantial amounts of biofuel feedstocks are large, and currently, structural water shortages affect about 300-400 million people mainly in Africa and Asia in a band from China to North Africa. Large additional water requirements fol­low from expected population growth and changes in dietary habits, especially the increased consumption of animal produce (Falkenmark and Lannerstad 2005; Liu and Savenije 2008). Assuming business as usual, which does not include a substan­tial production of modern biomass-for-energy, it has been suggested that shortages of freshwater may well become a fact of life for up to 2.5-6.5 billion people by 2050 (World Water Council 2000; Wallace 2002). For instance, water requirements for food consumption are expected to increase greatly in rapidly industrializing coun­tries such as China and India (Falkenmark and Lannerstad 2005; Liu and Savenije 2008). The latter countries are expected to rely increasingly on food export because of limited water availability (Falkenmark and Lannerstad 2005; Liu and Savenije 2008), and this makes it unlikely that they are suitable for large-scale biofuel pro­duction.

Still, in case of major net importers of mineral oil, it may be argued that the availability of biofuels on the world market decreases their reliance on the limited number of countries that are suppliers of mineral oil, and that this diversification may contribute to increased overall energy security. Moreover, one would expect that a substantial production of transport biofuel may have a downward effect on mineral oil prices (Eickhout et al. 2008).

On the other hand, there is the matter of the long-term strategy regarding energy security in transport. Transport biofuels are interesting because they may be used as ‘drop ins’ without a major change in transport technology. But from a long-term perspective, one may argue that major advances should come out of the twin de­velopment of higher energy efficiency in transport (Eaves and Eaves 2007; Royal Society 2008) and supply technologies that are much more efficient than photosyn­thetic organisms in converting solar radiation into usable energy. Using solar cells or concentrated solar power (CSP) to produce H2 for fuel cells or electricity for bat­teries is an interesting example of the latter (Armor 2005; Ros et al. 2009). Though physical conversion technologies such as solar cells presently have higher costs (ex­cluding external costs) than biofuels, it has been argued that in the long run, it may be better to focus on such physical conversion technologies than taking the ‘detour’ of biofuels (cf. Lee and Lee 2008).

The alternative of the twin development of higher energy efficiency and more efficient solar energy conversion technologies will also come up in the context of the next three sections (6.4-6.6) which deal with problems linked to the large areas needed for the production of large amounts of biofuels.

Near-shore perspectives for exploiting macroalgae may be different. Firstly, there are cases where macroalgae have developed into a pest because of eutrophication (Morand and Merceron 2005). In some of these cases, significant amounts of these macroalgae are currently collected and landfilled. For instance, in Europe, this hap­pens in parts of the Venice Lagoon, the Ortbetello Lagoon, the Bay of Brittany and the Peel Inlet, with collected amounts in the order of 103-104 Mg per year (Morand and Merceron 2005; Bastianoni et al. 2008). Such macroalgae may be used for bio­fuel production. However, when nutrient emissions are reduced, macroalgal primary production will also be diminished. Secondly, there is already major near-shore cul­tivation of macroalgae mainly for food and feed (Wikfors and Ohno 2001; Critch — ley et al. 2006). Eutrophication of coastal waters is conducive to yields, and there is also intentional addition of nutrients to further production (Neushul and Wang 2000). Also, it has been suggested to combine cultivation of macroalgae with nu­trient emissions from marine animal aquaculture (Wikfors and Ohno 2001; Chopin et al. 2001; Troell et al. 2006).

4.3.1 Introduction

Transport biofuels made from plants are often called ‘climate neutral’ or ‘carbon neutral’. These terms can be traced back to the participation of plants in the bio­geochemical C cycle. Plants take up CO2 from the atmosphere and convert this into biomass, and when biomass is burned, the CO2 is ‘given back’ to the atmo­sphere. There is said to be C neutrality: over a short time span, sequestration equals emission of CO2, which is a greenhouse gas. Greenhouse gases are transparent for relatively high energy solar radiation, such as visible light, but absorb infrared radi­ation and thereby influence atmospheric temperature. And thus, carbon neutrality is in this case said to equal climate neutrality. However, there is more to the relation between biofuels and climate. In part, this is linked to the direct effect of plants on local climate and in part to the emission of non-CO2 greenhouse gases, such as N2O and CH4. Also, biofuel production can be accompanied by changes in C sequestra­tion by ecosystems.

If compared with the original ‘natural’ vegetation, cropping biofuels may differ in determinants of local climate, such as surface roughness (Notaro et al. 2006), evapo(tanspi)ration (Gustafsson et al. 2004; McPherson 2007), precipitation (Liu et al. 2006) and albedo (Gustafsson et al. 2004; Schneider et al. 2004; McPher­son 2007). 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. Net effects of vegetation change may be different dependent on region. In cold regions, replacement of forest by annual biofuel crops tends to have a cooling effect, due to the importance of change in albedo, and in tropical regions, this replacement may cause warming, mainly due to a decrease in evaporation and cloud cover (Betts et al. 2007). When changes in vegetation are widespread, there may be knock-on effects on climate on a wider scale (Delire et al. 2001; Liu et al. 2006; Betts 2007; Betts et al. 2007; McPherson 2007). These will be further discussed in Chap. 5.

Here, of the factors that may impact climate, we will only further consider net greenhouse gas emissions. These may be positive or negative. The latter case corre­sponds with net C sequestration. First, we will consider the major determinants of these net emissions. Thereafter, the actual net greenhouse gas emissions of a number of transport biofuels will be considered.

Potentially important determinants of the net greenhouse gas emissions linked to the transport biofuel life cycle are:

• Carbonaceous greenhouse gas emissions linked to the cumulative demand for fossil fuels.

• N2O emissions linked to N inputs in, and non-product outputs (e. g. NOx emis­sions) of, biofuel production.

• Changes in atmospheric CO2 concentrations following from changes in carbon sequestration. The latter may relate to changes in soil carbon level and/or changes in aboveground biomass.

• Emissions of biogenic, non-CO2 carbonaceous greenhouse gases linked to the biofuel life cycle. These include CH4 emissions linked to anaerobic conversion of biomass and non-CO2 carbonaceous greenhouse gas emissions due to biomass burning.

These determinants will be considered in turn.

2.3.1 Transport Biofuels from Terrestrial Plants

Most studies regarding cumulative fossil energy demand have been done for trans­port biofuels from terrestrial plants, and most agree that the seed-to-wheel cumula­tive demand for fossil fuels associated with transport biofuels from terrestrial plants is lower than the well-to-wheel demand of fossil transport fuels. However, Patzek and Pimentel (Pimentel 2003; Patzek 2004; Patzek and Pimentel 2005; Patzek 2006) have presented calculations for cornstarch-derived ethanol, soybean — and sunflower — derived biodiesel and lignocellulosic ethanol that suggest a higher cumulative de­mand for fossil fuels. The difference between these studies of Patzek and Pimentel and other studies is partly caused by difference in allocation, partly by higher es­timates of fossil fuel input in agriculture and industrial processing, and partly by factoring in the energy demand of the infrastructure needed for transport biofuel pro­duction (factories, vehicles, etc.) into the calculations. However, along with assump­tions that are more favourable to transport biofuels, there seems no denying that in western industrialized countries, the cumulative fossil energy demand for transport biofuels made from starch, sugar and edible oils may be quite high when alloca­tion is on the basis of price. For ethanol from US corn or European wheat or rye, it would seem unlikely that, when allocated on this basis, the ‘seed-to-wheel’ cumu­lative fossil energy demand would be much lower than 80% of the corresponding demand for petrol (Hammerschlag 2006; Hill et al. 2006; von Blottnitz and Curran 2007; Reijnders and Huijbregts 2007; Zah et al. 2007). In the case of biodiesel from rapeseed and soybean, qualitatively good estimates usually suggest that, when allo­cated on the basis of price, the cumulative energy demand may well be in the order of 60-80% of the corresponding demand for diesel (Hill et al. 2006; Zah et al. 2007).

Cumulative fossil energy demand for transport biofuels may be considerably lower when biofuels based on high-yielding crops from developing counties, such as oil palm and sugar cane, are considered, especially when lignocellulosic biomass is used for powering processing facilities (von Blottnitz and Curran 2007; Reijnders and Huijbregts 2008a). When the latter applies, for instance, cumulative fossil fuel inputs in ethanol from sugar cane may become energetically less than 10% of the ethanol output (Macedo et al. 2008). Also, much lower cumulative fossil fuel de­mands have been estimated for transport biofuels from lignocellulosic biomass such as wood or switchgrass when processing is also powered by lignocellulosic biomass (von Blottnitz and Curran 2007). When allocation is based on the energy content or weight of outputs, cumulative fossil energy demand allocated to transport biofuels will tend to be lower than in the case of allocation based on price. Note that cumula­

tive mineral oil demand is often lower than cumulative fossil fuel demand, because natural gas and coal can be significant contributors of energy to the transport biofuel life cycle (Hammerschlag 2006; Kim and Dale 2008). For instance, coal is often an important contributor to electricity supply, which is sometimes used by mills pro­ducing ethanol (Kim and Dale 2008). Natural gas is important in production of fixed nitrogen to be used in agriculture (Hammerschlag 2006).

Using transport biofuels may change the emissions of non-greenhouse gases, if com­pared with the original (fossil) fuel. For instance, the substitution of fossil diesel by biodiesel (fatty acid ester) reduces sulphur dioxide emissions but tends to increase the emissions of nitrogen oxides (NOx) from diesel, whereas the acute effects on res­piratory organs do not change significantly (Ban-Weiss et al. 2007; Lin et al. 2007; Swanson et al. 2007; Szybist et al. 2007). The increase of NOx emissions caused by switching to biodiesel can be reduced by adjusting timing of the injection pump (Kegl 2008). The impact of substituting fossil diesel by biodiesel on particulate matter emissions by motorcars is apparently complex, with evidence that biodiesel substitution impacts the nanostructure of diesel soot, enhances oxidative reactivity and cytotoxicity but reduces mutagenicity (Bunger et al. 2000; Szybist et al. 2007). It appears that the overall amount of particulate matter and the number of particles that is emitted is reduced when fossil diesel is progressively replaced by biodiesel, which seems indicative of reduced risk. But the average particle size is also reduced (Kegl 2008; Keskin et al. 2008; Lapuerta et al. 2008), and smaller particle size is correlated with increased risk of a specified mass of particulate matter (Lapuerta et al. 2008). The overall effects of all these changes on human health impacts await further research (Swanson et al. 2007).

It would seem likely that, if compared with fossil gasoline, the admixture of ethanol to gasoline may be able to reduce emissions of CO and reduce ambient O3 concentrations (Ahmed 2001). On the other hand, the emission of acetaldehyde is in­creased by such a substitution, and there may also be an increase in the atmospheric concentration of peroxylacetate nitrate (PAN) (Ahmed 2001). What the overall im­pact thereof on health will be awaits further research. Moreover, in practice, ethanol (or ETBE) may not substitute fossil hydrocarbons but other oxygenates of MTBE. It would seem doubtful that, as far as its impact on inhaled air is concerned, such a substitution would benefit health (Ahmed 2001).

Changes in non-greenhouse gas emissions are not confined to cars; they concern the complete life cycles. And indeed, a substantial part of the seed-to-wheel non­greenhouse gas emissions is, for instance, associated with the cropping stage. This stage is associated with the input of fertilizers (‘nutrients’) and pesticides. Nutrients (including conversion products thereof) may be emitted into the wider environment. Well known is the leaching of P and N nutrients into water. Leaching of these nu­trients from arable soils in the US Midwest, where corn is grown to supply ethanol, is a primary contributor to the hypoxic zone in the Gulf of Mexico (Powers 2007). Hypoxic zones due to elevated levels of nutrients also occur in the East China Sea and several continental European seas, whereas continental shelves of Africa, South America and India are relatively vulnerable to increases in nutrient emissions (Diaz and Rosenberg 2008). More in general elevated concentrations of nutrients may lead to eutrophication. Eutrophication is linked with harmful algal blooms and reduced biodiversity (Graneli and Turner 2006; Ptacnik et al. 2008).

Even the cropping of Jatropha, which produces nuts with well-known insectici­dal properties, may require substantial pesticide inputs to reduce the impact of pests (Grimm 1999; Grimm and Somarriba 1999; Carvalho et al. 2008). More generally, cropping is also associated with the use of pesticides, which may lead to ecotoxicity and toxic effects on humans. In some cases, handling of harvested materials may have a large impact on non-greenhouse gas emissions. A case in point is the burning of harvest residues of sugar cane, which serves as feedstock for the production of bioethanol. This has an adverse impact on populations living in areas where sugar cane is harvested, especially on the respiratory systems of children and the elderly (Cangado et al. 2006).

The most comprehensive study regarding transport biofuel life cycles is the work of Zah et al. (2007), who compared traditional fossil fuels with a variety of plant-based biofuels, such as rapeseed methylester, palm oil methylester, soybean methylester, methanol and ethanol from various biomass sources and countries of origin, regarding seed-to-wheel non-greenhouse gas emissions. Allocation was on the basis of prices. Zah et al. (2007) considered the life cycle emissions that may lead to oxidizing smog, eutrophication and ecotoxicity. In many cases, the emis­sion of ecotoxic substances was found by Zah et al. (2007) to be lower for crop — based transport biofuels than for fossil fuels. However, there were also exceptions. Biodiesel from Malaysian palm oil and Brazilian soybean oil gave rise to seed-to — wheel emissions that were at least five times more ecotoxic than the fossil petrol or diesel life cycle emissions. As to eutrophication, plant-based biofuels tended to do worse than fossil transport fuels over their respective life cycles, with the exception of some wood — and grass-based products that scored rather similar to fossil trans­port fuels. Regarding the emission of hydrocarbons which may lead to oxidizing or photochemical smog, biofuels did often somewhat better than fossil fuels. How­ever, soybean-based biodiesel, Malaysian oil-palm-based biodiesel and bioethanol from sugar cane in Brazil did much worse regarding their seed-to-wheel emissions of compounds that may cause oxidizing smog.

Zah et al. (2007) did not consider acidifying substances (NO*, SO2, NH3, HCl), but other studies suggest that in this respect, biofuels often do worse than fossil fu­els (Kaltschmitt et al. 1997; Sheehan et al. 2003; Reinhardt et al. 2006; Kim and Dale 2008a), when allocation is on the basis of prices. Reinhardt et al. (2006) con­sidered a variety of processes that convert lignocellulosic biomass into transport fuels via synthesis gas. Apart from the life cycle emissions of acidifying substances, they looked at plant nutrients and compounds that may be toxic to humans, while allocating on the basis of prices, and concluded that such transport lignocellulosic biofuels did in these respects mostly worse than fossil fuels. When allocation would have been on the basis of energy content or weight of output, the emissions allo­cated to transport biofuels would have been lower than in the case of allocation on the basis of prices. Kim and Dale (2008a) looked at ethanol derived from US corn grain by dry milling and found that this did worse than conventional gaso­line as to eutrophication and photochemical smog. In this case, allocation was done by substitution. Graebig (2006) has considered the relative environmental impacts of electricity from photovoltaics and from biogas generated by the conversion of maize. It was concluded that photovoltaics were better in all life cycle assessment categories, except eutrophication.

Zah et al. (2007) also studied waste-to-wheel emissions associated with methane production from a variety of wastes and compared these with natural gas. They found that the emission of hydrocarbons, which contribute to oxidizing smog asso­ciated with methane from wastes, was somewhat larger, and the emission of eutro —

phying substances much larger than in the case of natural gas. Emissions of ecotoxic substances were roughly similar or somewhat larger. The outcomes of the study of Zah et al. (2007) seem more favourable to transport biofuels made from wastes than to transport biofuels made from food crops. However, one should keep in mind that this verdict is based on the assumption that life cycle impacts up to the waste can be neglected. When wastes change into secondary resources, fetching a price, or when the allocation in life cycle assessment is based on mass or energy, differences between transport biofuels made from, for example, starch and from residues would become smaller.

Besides biological processes, there are also physical conversion processes for solar energy. Efficiencies for a number of physical methods of converting solar radiation into heat, H2 or electricity are in Table 2.3. It can be seen that solar conversion efficiencies of photovoltaic cells are much higher than the conversion efficiencies for the transport biofuels in Table 2.2.

As to the effect of biofuels on food security, it has already been noted in Chap. 1 that substantial production of transport biofuels will, under market conditions, have an upward effect on food prices. Prices of food crops which also serve as major feed­stocks for biofuels are likely to show linkage with fuel prices. The price of sugar in Brazil is now linked to the price of ethanol, and large-scale use of carbohydrates and vegetable oils as transport biofuel feedstocks may be expected to link the prices thereof to fossil fuel prices, corrected for differences in ‘energy content’ (Naylor et al. 2007; von Braun 2007; Eickhout et al. 2008; Westhoff 2008). The effect of an expanding transport biofuel production on food prices may lead to an increased insufficiency of food for the world’s poorest people that are not net food producers and currently spend 50-80% of their total household income on food (Naylor et al. 2007; Runge and Senauer 2007, Daschle et al. 2007; von Braun 2007). Fast expan­sion of transport biofuel feedstock production might be expected to have a relatively strong upward effect on food prices (von Braun 2007).

The upward effect of transport biofuel production on food prices partially follows from competition between food crops and biofuel crops for good-quality land. This competition occurs both when transport biofuels are based on feedstocks that can be used for food or feed and in the case that feedstocks for lignocellulosic biofuels are grown (Christersson 2008). Thus, the competition extends to part of the lig — nocellulosic transport biofuels, including biofuels made from lignocellulosic crops, such as Miscanthus (e. g. S0rensen et al. 2008), and biofuels from lignocellulosic by­products which are currently used as animal feed (e. g. Linde et al. 2008; Murphy and Power 2008).

However, it may be expected that when lignocellulosic biofuels contribute sub­stantially to transport biofuel production, the upward effect on food prices will be reduced. This is even more so when lignocellulosic crops are converted into electric­ity for electric traction, because the seed-to-wheel solar energy conversion efficiency is relatively high (see Chap. 2). There will also be an effect when lignocellulosic biomass is converted into biofuel for internal combustion engines, because in this case, more cropped biomass can be turned into transport biofuel. The magnitude of this effect is uncertain, however, as it is not clear how much lignocellulosic biomass can be diverted to transport biofuel production without having a negative impact on soil organic matter levels and animal feed supplies.

The competition between food crops and biofuel crops also appears to apply to biofuel crops which are well adapted to growth on poor-quality land. One example thereof is Eucalyptus. Around 1900, Eucalyptus was promoted for growth on ‘waste lands, where few other trees would grow’ (Doughty 2000). However, now it is often grown on good-quality land in competition with food crops, which has led to restric­tions on Eucalyptus cultivation in some countries (see Chap. 3). More recently, the oil crop Jatropha has been promoted because of its ability to grow on marginal land (Kaushik et al. 2007; Achten et al. 2009). However, as evidenced by the eviction of small-scale farmers in Tanzania for large-scale Jatropha cropping (Gross 2008) and the replacement of rice production by Jatropha in Burma (Ethnic Community De­velopment Forum 2008), in practice, Jatropha cultivation may well compete with food production. It has also been found on the basis of experience with Jatropha cultivation in Belize, Nicaragua and India that to be competitive, cultivation has to be intensified beyond that of a rain-fed, low-input and drought-resistant crop (Euler and Gorriz 2004). This should be no surprise as on the biodiesel market, Jatropha oil also has to compete with vegetable oils, which have been grown under good conditions which are conducive to high yields.

To the extent that the competition between food and transport biofuel crops for good-quality land has been studied for the United States, a rather general upward effect on food prices has been found (Walsh et al. 2003; Johansson and Azar 2007; Schneider et al. 2007). However, there may also be differential effects of biofuel crops on the prices of specific foods. These depend on actual crops that are used for the production of transport biofuels. Elobeid and Hart (2007) have modelled the effect of expanding bioethanol from corn production in the USA and found the biggest impact on food-basket costs in sub-Saharan Africa and Latin Amer­ica, where corn is a major food grain. A lower impact was expected in Southeast Asia where rice is a major food grain, with countries where wheat and/or sorghum are major staples falling in between. To lessen the effects of biofuel feedstocks on Chinese food prices, in 2008, China began to import cassava as feedstock from Malaysia, the Philippines, Indonesia and Nigeria (Tenenbaum 2008). When China is to heavily rely on cassava as a feedstock for bioethanol production, it would seem likely that prices of this ‘poor man’s food’ may be much increased (Naylor et al. 2007).

Is there a strategy for developing transport biofuels that will not have an upward impact on food prices? The answer to this question should take account of a down­ward pressure on crop production associated with climate change and increasing land claims associated with agricultural production for an increasing world popu­lation with consumption patterns that increasingly favour animal produce (Tilman et al. 2001; Reijnders and Soret 2003; Swedish Environmental Advisory Council 2007; Koneswaran and Nierenberg 2008; von Braun 2007). The latter development will in all probability intensify competition for good-quality land.

To the extent that one relies on crops for the supply of transport biofuel feed­stock, while relying on market forces, direct competition with food and feed pro­duction therefore seems inevitable, as does an upward effect on food prices. In this respect, there are likely to be quantitative differences linked to the relative yield of transport biofuels per hectare. These differences can be substantial, as shown in Chap. 2. From the data presented in Chap. 2, it would seem that crops proposed for the generation of lignocellulosic feedstock are not necessarily superior to current food crops such as sugar beet and sugar cane as to their net efficiency in converting solar radiation into biomass. How they will perform in net yield of biofuels is rather

uncertain because technologies for the conversion of lignocellulosic biomass into transport biofuels are under development, and there is uncertainty about yields that may be possible in the future and the extent to which aboveground biomass should be returned to soils to maintain soil organic matter levels.

Still, it is to be expected that some biofuels would not lead to an upward move­ment of food prices. Firstly, biofuels produced from what are currently ‘wastes’, such as organic urban wastes, biomass from forest remediation and residues from forestry and agriculture, which are not used as animal feed, may partly qualify as such. The worldwide amount of these wastes is currently estimated at between 50 and 100 EJ (Swedish Environmental Advisory Council 2007; Lysen and van Egmond 2008). Unfortunately, it is not clear how much thereof is necessary for maintaining soil organic carbon stocks in a steady state to safeguard the future productivity of arable lands and forests (see Chap. 3). However, even when only 10-20% thereof could be diverted to transport biofuel production, this would still represent a substantial contribution to the transport fuel supply.

Another option that has been suggested in this context is growing microalgae (Chisti 2007, 2008; Dismukes et al. 2008; Groom et al. 2008). However, as pointed out in Chap. 2, an overall positive energy conversion efficiency of microalgal bio­fuels currently seems uncertain. There is also the water demand associated with growing algae, which may, per kilogram of dry weight biomass, be larger than, for example, sugar cane (see Chap. 3). This may lead to claims which may easily com­pete with agricultural land use.

Still another option is the use of abandoned cropland and lands that sequester little carbon today (Searchinger et al. 2008). The use of terrestrial plants to reclaim deserts may make it possible to harvest lignocellulosic biomass or oil (from, e. g. Ja — tropha). Of course, sustainable productivity of reclaimed drylands is relatively low. Apart from human intervention, actual productivity depends on rainfall (Webb et al. 1978). In tropical and subtropical areas with precipitation below 500mmyear-1, abovegroundC sequestration may be roughly between 0.15 and 1.5Mgha-1year-1 (Hadley and Szarek 1981). Increased sustainable yields may be possible by effi­cient water management and conservation practices (Thomas 2008). In semi-arid (500-750 mm rainfall year-1) and sub-humid (750-1,000mm rainfall year-1) envi­ronments with relatively high insolation, aboveground C sequestration may amount to 2-3 Mg C ha-1year-1 (Lal 2001). In humid Icelandic deserts, restoration activ­ities have led to the sequestration of 0.6-1.1Mg C ha-1year-1 (Aghstdottir 2004). After reclaiming lands with little C sequestration, there are often competing uses. For example, biomass may be exploited for grazing (Brown 2003; Darkoh 2003; McNeely 2003; Lal 2008; Ludwig et al. 2008), and this may lead to limitations on use for biofuel production.

If compared with reclaimed deserts, biomass production may be higher on cur­rently abandoned and fallow agricultural lands, with proper use of organic amend­ments and fertilizers and appropriately adopted plant species (Lal and Bruce 1999). Field et al. (2008) and Campbell et al. (2008) have estimated that such lands com­prise about 385-472 x 106 ha. In Chap. 3, it has been estimated that, after restoration of soil organic matter and nutrient levels, the worldwide sustainable feedstock pro­duction on such lands may be in the order of 23-28 EJ. When one assumes that the conversion efficiency thereof to transport biofuels is 40-50%, this would allow for the production of 8.6-14 EJ of transport fuels, which would be a substantial contri­bution to the 85-90 EJ of transport fuels that is currently used in means of transport.

Higher yields may be possible by intensifying cultivation of lands which cur­rently sequester little C and agricultural lands that have been abandoned or fallow, but it is highly doubtful that the biofuels generated in this way could be considered sustainable. As pointed out in Chap. 5, to be successful, this way to exploit fallow and abandoned agricultural land should be viewed by the local population as being in line with their pressing needs. Moreover, for the large-scale cultivation on aban­doned and fallow croplands and lands that currently sequester little C, one has to go beyond the market mechanism. For instance, in the case of palm oil, in Malaysia, planting oil palms on abandoned land is currently rare, because degraded land does not provide revenue from initial timber extraction and entails relatively high estab­lishment costs and possibly reduced yields (Wicke et al. 2009). More in general, the profitability of abandoned cropland and land that currently sequesters little C tends to be less than for good-quality agricultural land (Huston and Marland 2003; Johansson and Azar 2007). Therefore, large-scale cultivation of crops for transport biofuels on degraded cropland and in the context of desert reclamation will probably depend on government interventions (see Sect. 6.7).

While aiming at transport biofuels, the growth of microalgae with high levels of oil (triacylglycerol) followed by lipid extraction has drawn most attention (Scragg et al. 2002; Wijffels 2008; www. oilgae. com; Dismukes et al. 2008; Liu et al. 2008). Such lipids can then subsequently be converted into replacements for fossil fuels, in ways similar to vegetable oils from terrestrial plants. There is currently some use of biodiesel based on algal oils, as pointed out above. There have also been proposals to convert algal biomass into methanol via synthesis gas or into bio-oil via pyrolysis (Hirano et al. 1998; Sawayama et al. 1999). Strains of the photosyn­thetic microalga Botryococcus braunii may contain and secrete substantial amounts of isoprenoid hydrocarbons: n-alkadienes and trienes, methylated squalenes and terpenoids (Guschina and Harwood 2006). When subjected to catalytic cracking, these hydrocarbons can be converted into transport biofuels (Banerjee et al. 2002). It has also been suggested that an intermediate in the synthesis of isoprenoids by Botryococcus braunii (isoprenylpyrophosphate) may be converted into isopentanol, which may be used as a gasoline additive (Fortman et al. 2008). The slow growth of Botryococcus braunii has not been conducive to its application.

Microalgae may be produced in open ponds converting solar irradiation into biomass which may be harvested and converted into biofuels. Open ponds used for growing microalgae are man-made structures (made from, for example, plas­tic or concrete) with 10-20 cm of water that are subjected to circulation and mix­ing (Chisti 2007). Closed systems (‘bioreactors’) have also been proposed for the purpose of growing photosynthetic micro-organisms to produce transport biofuels (Chisti 2007, Wijffels 2008). In closed systems, heterotrophs, organisms that graze on algae (zooplankton) and viruses can be excluded, and monocultures of desirable species can be maintained. In open ponds, sustained generation of a specific pho­tosynthetic micro-organism with relatively little contamination of other species and subject to low heterotrophic conversion would seem only possible under extreme circumstances, such as very high salinity and/or a high pH (Joint et al. 2002; Ugwu et al. 2008). Sustained open pond production has been successful for a limited num­ber of algae such as Spirulina, Chlorella and Dunaliella (grown at high pH and/or NaCl concentrations). For other organisms, most growth can take place in a closed bioreactor, which then may be eventually followed by a short period in an open pond (Huntley and Redalje 2007).

Freshwater Macrophytes

In fresh waters, there has been an emergence of invasive macrophytes with high pri­mary production per hectare. Increased levels of nutrients (‘eutrophication’) and the import of macrophytes from other continents have been conducive to this emergence (Gassmann et al. 2006; Gunnarsson and Petersen 2007). Among these macrophytes, the water hyacinth (Eichhornia crassipes) has been studied in the context of bio­fuel production (Gunnarsson and Petersen 2007; Malik 2007). The water hyacinth has emerged as a major invasive organism (‘pest’) in tropical freshwater systems especially outside its natural range (South America). Water hyacinth biomass forms floating mats which interfere with shipping, power generation, drinking water pro­duction and irrigation, are detrimental to fish stocks and may be conducive to a num­ber of infectious diseases (Odada and Olago 2006; Gunnarsson and Petersen 2007; Malik 2007). Due to these negative impacts, there are efforts to reduce the pres­ence of Eichhornia crassipes in tropical surface waters, which have met with at least some success (Odada and Olago 2006). The need to control the water hyacinth is evidently at variance with high yields, but when the water hyacinth generates substantial amounts of floating biomass, energetic use thereof may be considered (Gunnarsson and Petersen 2007; Malik 2007).

Transport biofuels replace fossil fuels. But because, as pointed out in Chap. 2, much production in current societies is dependent on fossil fuels, it will come as no sur­prise that burning fossil fuels is often an important contributor to the greenhouse gas emissions associated with biofuels. N fertilizers are often made on the basis of natu­ral gas; tractors and transport are often powered by fossil fuels based on mineral oil. Factories, involved in converting biomass into fuels fit for powering vehicles, are more mixed in their fuel use. There are production facilities doing without burning fossil fuels. In Brazil, factories converting sugar into ethanol are often powered by harvest residues of sugar cane (Macedo et al. 2008). In Sweden, biofuel production in factories tends to use wood chips from forest logging residues (Borjesson and Mattiasson 2008). But, for example, in France and Germany, factories producing bioethanol or biodiesel are usually powered with fossil fuels (Reijnders and Huij — bregts 2007,2008b). Life cycle assessments of biofuels are characterized by a long­standing interest in the CO2 emissions linked to cumulative fossil energy demand (e. g. Pimentel et al. 1973; Weisz and Marshall 1979), and by now the standard is that they include the CO2 emission linked to burning fossil fuels to power the bio­fuel life cycle, though there are still exceptions to this rule (e. g. Chisti 2007, 2008; Dismukes et al. 2008).

There may also be non-CO2 carbonaceous emissions linked to fossil fuel use. For instance, there may be leakage of CH4 (methane) during transport and use of natural gas, and, if burning is not optimized, there may be emission of hydrocarbons and carbon monoxide (CO). The non-CO2 carbonaceous gases on a molecule-for — molecule basis tend to have a greater greenhouse effect than CO2. The non-CO2 carbonaceous greenhouse gases are often neglected in life cycle assessments, which will lead to an underestimate of the greenhouse effect of actual emissions. However, in advanced industrial economies, the error linked to this underestimation will be small, in the order of a few percent maximally.