Food, agriculture and fisheries, and biotechnology teampest project
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Source: Eurostat (2008) The main users of pesticides at a European level are Italy, France, Spain, Germany and the United Kingdom (Table 1). On the other hand the lower positions are shared between the Scandinavian countries that appear to have the lowest amount of pesticide purchases. In most of the countries, pesticide sales during the period 1996-2007 appear to experience slight fluctuations but in general it can be concluded that either they increased slightly or remained at the same level. Among the exceptions are Italy, Poland, Portugal, and Greece where pesticide sales increased, and Denmark, France and Germany that have achieved considerable reductions. The increased pesticide sales in the above mentioned countries can be attributed to the fact that the economic growth that they are experiencing is translated into agricultural intensification in the rural areas with increased use of production inputs like pesticides. Different climatic conditions around Europe are responsible for differences in the types of pesticides used. Northern European countries that have humid climatic conditions use more herbicides and fungicides (Appendix, Table 1, 2) while Mediterranean countries use mainly insecticides (Appendix, Table 3) as the warm climate is responsible for a plethora of insects. The sales of other pesticides (Appendix, Table 4) like growth regulators and wood preservatives have a small share in comparison to insecticides and herbicides.
Upon reviewing the trajectory of pesticide use in European Union (EU), Eurostat (2008) trends are presented separately for EU-15 countries and EU-25 countries as data for the latter were obtained after 2000. Total pesticide use increased steadily during 1990’s while after a period of stabilization at the end of 1990’s, started to decrease (Figure 16). 
Figure 16. Total pesticide consumption: EU15, EU25, 1992-2003. Source: Eurostat (2008) Figure 17 illustrates the use of different plant protection products (PPPs) at the EU-15 level. The most widely used type of pesticides is fungicides followed by herbicides, other PPPs (growth regulators, wood preservatives, rodenticides), and insecticides. The use of fungicides increased in the mid 1990’s but after this period there is a continuous decrease. This decrease can be attributed in a shift to substances active at low dosages, dryer climatic conditions at the Northern EU countries during the last years, and increasing information and knowledge at the farm level (extension services, IPM) that leads to the application of more environmental friendly agricultural practices. 
Figure 17. Use of PPPs at an EU-15 level (1992-2003) Source: Eurostat (2008) Herbicides use has followed an increasing trajectory with a decrease after 2002 while the use of other PPPs increased in mid-1990’s but after this period remained at the same level. Finally, insecticides use seems to follow a steady path throughout 1992-2003. The EU enlargement and the data availability after 2000 have contributed to an increase in the use of all PPPs. However, after 2000 there is a stabilizing or a decreasing trend in the use of all types of pesticides.
The calculation of pesticide demand elasticities is important in order to design an EU wide regulatory framework for levies on pesticides. If pesticide demand is inelastic the tax or levy introduced will not affect pesticide use significantly but it will create revenues that can be reimbursed to the agricultural sector. Table 2 presents a review of the pesticide demand elasticity estimates of European Countries and United States. A general conclusion based on this table is that the price elasticity of demand for pesticides is quite low (in most of the cases), indicating that pesticide use is indifferent to pesticide price increases. Inelastic demand can mean that there is a lack of knowledge among farmers on alternative production practices or there is a strong intention towards risk-aversion or even due to behavioral factors like professional pride derived from weed-free fields. Another important point is that the more specific the pesticide (fungicides, insecticides) is, the higher the elasticity of demand is. The reason behind this is that there are not so many substitutes to these specific products, with the result the producers to face difficulties in adjusting their agricultural practices. Table 2. Pesticide Demand Elasticity Estimates 
Study Country/Region Elasticity
Aaltink (1992) Netherlands -0.13 to -0.39 Antle (1984) USA -0.19 Bauer et al. (1995) German regions, wheat -0.02 Brown & Christensen (1981) USA -0.18 Carpentier (1994) France, arable farms -0.3 DHV & LUW (1991) Netherlands -0.2 to -0.3 Dubgaard (1987) Denmark -0.3 (threshold approach) Dubgaard (1991) Denmark -0.7 (herbicides) Dubgaard (1991) Denmark -0.8 (fungicides + insecticides) Ecotec (1997) UK -0.5 to -0.7 (herbicides) Elhorst (1990) Netherlands -0.3 Falconer (1997) UK (East Anglia -0.1 to -0.3 arable production) Gren (1994) Sweden -0.4 (fungicides) -0.5 (insecticides) -0.9 (fungicides) Johnsson (1991) Sweden -0.3 t0 -0.4 (pesticides) Komen et al. (1995) Netherlands -0.14 to -0.25 Lichtenberg et al. (1988) USA -0.33 to -0.66 McIntosh & Georgia (USA) -0.11 Williams (1992) Oskam et al. (1992) Netherlands -0.1 to -0.5 (pesticides) Oskam et al. (1997) EU -0.2 to -0.5 (pesticides) Oude Lansink (1994) Netherlands, arable farms -0.12 Oude Lansink & Netherlands -0.48 (pesticides) Peerlings (1995) Papanagioutou (1995) Greece -0.28 Petterson et al. (1989) Sweden -0.2 Rude (1992) Sweden -0.22 to -0.32 Russell et al. (1995) UK (Northwest) -1.1 (pesticides in cereals) SEPA (1997) Sweden -0.2 to -0.4 Schulte (1983) Three German regions -0.23 to -0.65 Villezca-Becerra & Texas & Florida (USA) -0.16 to -0.21 Shumway (1992) S  ource: Hoevenagel & van Noort (1999); Falconer & Hodge (2000); Fernandez-Cornejo et al. (1998)
The concept of damage abatement input was first introduced by Hall and Norgaard (1973) and Talpaz and Borosh (1974). Lichtenberg and Zilberman (1986) were the first to specify production functions that are consistent with the concept that pesticides are damage abatement input that have an indirect effect on output rather than a direct yield-increasing effect. The use of damage control inputs can have both positive and negative effects on output like the development of pest resistance that can lead to decreased output even if there is increasing use of pesticides. Damage control inputs reduce damage from natural causes and, except of pesticides, this class of production inputs include windbreaks, buffer zones and antibiotics. The damage control framework proposed by Lichtenberg and Zilberman (LZ) (1986) has important economic value. This framework enabled economists and policy makers to observe that the (then standard) Cobb-Douglas formulations were resulting in an upward bias in the optimal pesticide use estimations (underuse of pesticides) while recent evidence suggests an overuse. Additionally, the damage control specification accounts for changes in pesticide productivity and enables the prediction of producers’ behavior. Pest resistance initially triggers farmers to apply more pesticides until alternative damage control measures become more cost effective. The LZ damage control specification was applied by Babcock et al. (1992), Carrasco-Tauber and Moffit (1992), Chambers and Lichtenberg (1994) and Oude Lansink and Carpentier (2001). The results are mixed with some studies indicating overuse of pesticides and other underuse. Although in general the LZ specification has been successfully applied and constitutes a considerable innovation, some authors have expressed various critiques. Oude Lansink and Carpentier (2001) have shown that in a quadratic production function the lack of differentiation between damage abatement inputs and productive inputs does not lead to overestimation of the marginal product as Lichtenberg and Zilberman (1986) argued. Additionally, they separate inputs into those that increase productivity and those that reduce damage and assume that there is an interaction between damage abatement and other production inputs, where the LZ specification precludes these interactions. Oude Lansink and Silva (2004) challenge the assumption of a nondecreasing damage control function and assumptions imposed on parameters in the damage control model.
IPM aims at farming with a relatively low input of plant protection products and a very high efficiency of their use. Based on ecological, sociological and economic factors, it emphasizes the development of alternative pest control practices (genetic, biological, mechanical, and cultural). Allen and Bath (1980) define IPM is “extremely pluralistic” as different disciplines see pesticides as the dominating element of IPM, while other focus on natural enemies, and mechanical and cultural practices. The United States Environmental Protection Agency (EPA) defines IPM as “an effective and environmentally sensitive approach to pest management that relies on a combination of common-sense practices”. Information on the life cycle of different pests and their interaction with the environment has a central role in an IPM program. This information in conjunction with existing pest control practices are used in order to address the problem, of crop losses due to different pests, in an environmentally friendly and economically viable way. IPM constitutes a mixture of pest control methods and decisions at the farm level. Based on EPA’s four steps, the first is the setting of action thresholds (pest levels) at which pests can pose an economic threat and therefore action needs to be taken. Then a farmer has to monitor and identify the pests of his/her field. This step is essential as it enables farmers to recognize innocuous and beneficial species, to judge if there is a real need to use pesticides, and if it seems necessary to use some type of pesticide, to use the correct one. In the third step, methods/practices should be undertaken in order to prevent pests from becoming a threat. Among these methods are cultural methods like crop rotation, and plantation of pest-resistant varieties. The final step prescribes rigorous action should be taken. Initially, less risky control methods (e.g. mechanical control) are chosen, but when farm operators identify that they are not effective enough, then additional control methods can be employed (e.g. targeted spraying of pesticides).
Biodiversity is a concept that comprises the totality of species in an area. Its conservation has received great importance in recent years, as its loss can be irreversible and reduce the ecosystem value as well as farm productivity. European agri-environmental schemes constitute an initiative towards biodiversity conservation.
Noss (1990) and Brock and Xepapadeas (2003) state that it is difficult to find a simple, comprehensive and fully operational definition of biodiversity. Diversity measures are influenced by the richness and evenness. Richness is the number of species at an ecosystem while evenness expresses the distribution of species. Many researchers and organizations tried to formulate biodiversity definitions. McNeely et al. (1990) state that biodiversity expresses the degree of nature’s variety, including species, genes and ecosystems. The World Resources Institute (WRI), the World Conservation Union (IUCN), and the United Nations Environment Programme (UNEP) (1992), define biodiversity as “the totality of genes, species and ecosystems in a region”. The United Nations Earth Summit in Rio de Janeiro (1992) (Convention on Biological Diversity), defined biodiversity as “the variability among living organisms for all sources, including, 'inter alia', terrestrial, marine, and other aquatic ecosystems, and the ecological complexes of which they are part: this includes diversity within species, between species and of ecosystems”. Gaston and Spicer (2004) provide a more straightforward definition by stating that biodiversity is “the variation of life at all levels of biological organization”.
The economic valuation of biodiversity is among the most pressing and challenging issues confronting today’s economists. There are many reasons behind valuing biodiversity. Biodiversity values can be compared with economic values of alternative options, a corner stone of any cost-benefit analysis. Additionally, valuing biodiversity can provide useful insights in environmental assessments, accounting and consumer behavior. A challenge in modeling biodiversity is to identify the states of nature that summarize its value. Economists deal with similar issues in the modeling of products in terms of their attributes. Duelli (1997) valued biodiversity by developing a conceptual ‘mosaic’ model in which biodiversity evaluation is based on structural landscape parameters like landscape heterogeneity, habitat diversity and on meta-community dynamics. Nijkamp et al. (2008) summarize the different methods of economic valuation of biodiversity (Figure 18).The valuation of biodiversity can be done in different ways. Monetary indicators of biodiversity can be extracted from market prices, for instance, by valuing the financial revenues from tourism to natural parks. Revealed Preference (RP) Techniques are other methods that can be applied to valuing biodiversity. Among them are the travel cost method, the hedonic pricing method and the averting behavior (Bockstael et al., 1991; Palmquist, 1991; Cropper and Freeman, 1991). The emphases of these techniques lies in valuing biological resources in an indirect way by investigating peoples’ preferences in purchasing goods that are related in some way to environmental goods. In other words, they lack of direct questions like how much a consumer may be willing to pay to preserve a natural resource or a plant or animal species. 
Figure 18. Methodologies for economic valuation of biodiversity. Source: Nijkamp et al. (2008) On the other hand, there are the Stated Preferences (SP) Techniques that are based on price observations of the good that is going to be valued. Data are collected by means of questionnaires while the best known method of this category is the Contingent Valuation Method (CVM). CVM enables researchers to avoid systematic bias and therefore underestimation of the different values as it uses also the non-use values. CVM allows for ex ante environmental valuation, offering greater scope and flexibility in comparison to RP techniques. According to Nijkamp et al. (2008), the most popular techniques for valuing biodiversity are the SP techniques and especially the CVM. Among the reasons for this classification are their easy format, the fact that they are more informative and the ease of isolation of the good in interest from other closely related goods.
Agrochemicals, overexploitation of natural resources, intensification of agricultural landscapes and trade of endangered species can have irreversible effects on biodiversity. Experimental studies have underlined the difficulty in enhancing the botanical diversity of fields especially after a period of intensive use that has depleted the local seed bank (Berendse et al., 1992; Bekker et al., 1997). Results from the evaluation of European Union agri-environmental schemes are underlying the difficulty in enhancing farmland biodiversity (Kleijn & Sutherland, 2003). Dietz and Adger (2003) estimate their parabolic Kuznets curve showing that biodiversity loss is expected to decrease and then rise with increasing income (Figure 19). However, the rising limb cannot be of the same magnitude as the falling limb as the species are not replenished at the same level. Other biodiversity indicators such as the presence of arthropods and birds have shown positive patterns in relation with changes in agro-environmental schemes (Kruess & Tscharntke 2002). Thus, there is a great uncertainty concerning the optimal time of intervention and policy makers have to weigh and monitor carefully all the costs that are created. 
Figure 19. Possible forms of the income – environmental degradation relationship. Source: Dietz and Adger, 2002.
Farmland biodiversity is mainly composed from different plant species, insects, breeding birds, rodents and small mammals. The largest number of farmland flora and fauna is mainly found at the field boundaries (Kleijn, 1997; Wossink et al., 1999) as they provide forage, shelter, and reproduction sites. Field boundaries support many flowering plants and insects, such as bees and butterflies, which are not only important for bird species but can also contribute to plant pollination. The intensive agriculture of the last decades has caused considerable environmental problems including the decline of farmland species. Species poisoning, accumulation of chemical substances in their bodies and transition to other organisms though the food chain are some of the common impacts of uncontrolled agrochemical use. Some chemicals can be directly toxic while some others can be responsible for reducing breeding success to levels that that could not maintain populations. Pimental et al. (1992) have shown that wild birds are subject to pesticide contamination and poisoning while Heard et al. (2003) report a 3% annual decrease of arable weeds since 1940. Donald et al. (2000) propose that farmland birds constitute a good indicator of overall farmland biodiversity and their populations (in Europe) have declined during the last decades. European Union data confirm this trend by indicating an overall decline of a selected group of breeding bird species dependent on agricultural land for nesting or feeding (Figure 20). 
Figure 20. Farmland bird index* (EU-25) Source: Eurostat (2008) * Indices are calculated for each species independently and are weighted equally when combined in the aggregate index using a geometric mean. Aggregated EU indices are calculated using population-weighted factors for each country and species. Pest control methods intend to provide an ideal habitat for crop plants by promoting better water and nutrient absorption and access to light, but on the other hand the non-crop plants share on the pre-mentioned resources declines (Firbank, 2005). These non-crop plants constitute a refuge and a source of food for many farmland birds and the winter food supply that weed seed banks can provide is important for their survival (Siriwardena et al. 2000). Moreover, Boatman et al. (2004) and Hawes et al. (2003) have shown that insecticide application during the breeding season of some farmland birds can be responsible for the decline of their populations as the supply of invertebrates for feeding chicks is reduced. Additionally, plant species variety has declined since the increased use of fertilizers favors the growth of nutrient demanding plants that are highly competitive and impede the growth of other species. Finally, the homogeneity of agricultural landscapes has dramatically increased and poses a serious threat to biodiversity (Benton et al., 2003). The reasons behind this is that many farmland species require different food resources that a homogeneous habitat cannot provide and also the decrease of mixed farming systems deprived small mammals and birds from feeding and nesting sites. Among the proposals for enhancing farmland biodiversity are crop rotation and mixed farming systems that can provide important food reserves for birds and mammals and the maintenance of crop edges (physical boundaries, trees, bushes and lack of spraying) that provide forage, shelter, reproduction and over-wintering sites for the farmland fauna.
Agri-environmental schemes were introduced in European agriculture under the 2078/92 regulation. Their introduction was a response to the increasing concerns for the environmental impacts of agricultural intensification. Among their main objectives are biodiversity protections, reduction of nutrient and pesticide emissions, restoring landscapes and preventing rural depopulation. Farmers receive payments in order to apply environmental friendly agricultural practices. Among the measures of agri-environmental schemes that aim at conserving and enhancing biodiversity are conservation of headlands for arable weeds, conservation of wet meadows, grassland and grazing extensification, botanical management agreements, meadow bird agreements, conservation of field margin stripes and agreements concerning wetlands and coastal habitats. Kleijn and Sutherland (2003) reviewed the literature dealing with the effectiveness of the pre-mentioned schemes but they were unable to express how effective these schemes are in protecting biodiversity as some studies indicated positive effects of agri-environmental schemes in terms of increased species diversity while other showed negative or no effects, or both some negative and positive effects.
Modern agricultural practices are moving towards the simplification of ecosystems. Pesticides are widely used in an effort to optimize the growing conditions of target species and/or to reduce those of competing species. Tilman et al. (2001) note that pesticide use may increase by 2- to 3-fold resulting in a further decrease of global biodiversity. The harm to biodiversity arises from the direct toxic effects of pesticides and their potential to reduce the number of competing plant species. The different weeds that are present around or inside the parcels are providing a breeding environment and a seed bank for different species such as birds and insects that can act as beneficial predators (Firbank, 2005). Therefore, increased intensification and further loss of the diversity of natural habitats is considered to be among the drivers of biodiversity loss. Biodiversity is closely related to agricultural productivity. Table 3 summarizes the relationships between pesticide use, biodiversity and agricultural productivity. While pesticide use may increase agricultural productivity in the short run it can be negatively impacted due to the development of resistance in the long run. On the other hand, biodiversity seems to have a positive effect on agricultural productivity. Table 3. Pesticide Use, Biodiversity and Agricultural Productivity
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