The Hidden and External Costs of Pesticide Use

2.4.1 A Small Number of Studies

2.4.2 A High Diversity of Costs

Both external and internal costs are associated with the testing and registration, production, distribution – including importation, transport and sales – use and disposal of pesticides. The external costs are the economic burden to the public authorities responsible for organizing controls and checks on the compliance of stakeholders, e.g. public authorities, consumers, sellers and producers, with the regulations. The internal costs are the monetary subsidiaries paid by pesticide handlers, e.g. users, sellers and producers, when they have to comply with mandatory regulations (Ajayi et al. 2002).

Most papers took into account the economic shortfall of crops exceeding the maximum residue limit or the costs of controls and monitoring (Table 2.4). Water decontamination, the regulation of pesticide registration and market monitoring costs were estimated in a small number of papers (Table 2.4) (Fig. 2.2). Other costs, such as those associated with governmental public information campaigns, economic shortfalls for water exceeding the maximum residue limit and public research on pesticides, were considered and estimated even less frequently (Table 2.4). However, these costs may account for a large proportion of the external costs of pesticides. For instance, public information campaigns accounted for about 10 % of the total external costs estimated by Khan et al. (2002) in the Pakistan, and public research costs were estimated at about 10 % of the total external costs by Praneetvatakul et al. (2013) in Thailand.

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Finally, some costs, such as the time and money spent establishing regulations, have never been estimated. This is unfortunate, because it has been acknowledged that such costs may be high, due to the need for research and development, expert advice and a number of official tests (Ajayi et al. 2002; Waibel et al. 1999).

2.4.3 Estimated Costs

Estimates of total annual regulatory costs vary considerably, from US$150,000 (2013) in Niger (Houndekon and De Groote 1998; Houndekon et al. 2006) to US$5 billion (2013) in the United States (Tegtmeier and Duffy 2004) (Table 2.3). We did not carry out a meta-analysis to find the cause of this variation. However, as a first approximation, we can consider this variation to be due to the differences in the categories of costs considered, the detailed composition of each category and the geographic scale of the study. The costs of commonly considered categories were particularly variable and depended strongly on the subcategories included. For instance, monitoring and control costs were frequently considered, but different aspects of these costs were covered. The estimates obtained thus differed considerably between papers, depending, in particular, on whether or not they considered the control of underground water. For instance, Pimentel and coworkers began to consider the costs of monitoring underground water and wells in their papers published in 1991. The consideration of these costs led to an immediate increase in their estimates of the overall cost of pesticide regulations of more than 300 %, with these costs accounting for 90 % of total regulatory costs for pesticide use (Pimentel et al. 1991a, b). Water decontamination and economic shortfalls due to crop contamination have been taken into account by Pimentel et al. since 1992. These costs accounted for about 40 % of the externalities associated with pesticide use.

2.4.4 Actual Versus Theoretical Costs

Most estimates of regulatory costs were based on the actual expenditure of various stakeholders, including public authorities, manufacturers, distributors, sellers and farmers. No attempt was made to estimate non-monetary values. Due to the ‘regulatory’ nature of these costs, estimates were generally based on the official budget reports of public agencies.

However, current costs may be much lower than the theoretical value. For instance, Pimentel (2005, 2009) and Pimentel and Burgess (2014) estimated the current monitoring of wells in the United States at about US$2 billion per year, but indicated that this cost would have reached US$17 billion per year if all the wells in the United States were monitored. Including these theoretical costs made a large difference, increasing the overall regulatory costs estimated by Pimentel (2005, 2009) and Pimentel and Burgess (2014) from US$4.2 billion to almost US$22 billion. Similarly, Jungbluth (1996) noted that costs related to pesticide residues in food in Thailand were difficult to estimate and were based on hypothetical scenarios rather than on real situations. In the absence of pesticide residue control for most food products, Jungbluth (1996) had to extrapolate the proportion of products exceeding the maximum residue limit from scarce data. Assuming that 10 % of all fruits and vegetables were above the maximum residue limit and assuming that these products would be unsaleable according to regulations, Jungbluth (1996) obtained a cost of about five billion Baht in 1996. He considered this value – corresponding to almost 90 % of the regulatory costs – as an upper limit for the costs truly paid by the corresponding stakeholders. Conversely, Jungbluth (1996) noted that if the maximum residue limit was not reached, then only the cost of control and monitoring should be taken into account, corresponding to 48.5 million Baht in 1996. This value should be taken as the lower limit of estimates. Along the same lines, Khan et al. (2002) distinguished between actual and potential costs. The potential costs they considered included the cost of establishing laboratories for pesticide residue analyses, residue monitoring programs, and training programs on the safe use of pesticides. These costs were largely theoretical, because there were no such activities in the region covered by their study in 2002, like in many developing countries (Ecobichon 1999). They reported the existence of regulations, but a lack of enforcement. They pointed out, in particular, that there was no comprehensive national monitoring system, and this may remain the case.

2.4.5 Conclusions

Regulatory costs, in particular, have been underestimated. We will see that this is also true for the other categories of “hidden” and external costs, but this underestimation may be particularly marked for regulatory costs. First, only 24 of the 61 articles assessing the external cost of pesticides included regulatory costs, and these 24 articles were actually based on only nine fully independent datasets. Second, each of these articles considered only a small number of regulatory costs. Finally, current costs are probably much lower than the costs that would have to be paid if the complete control, monitoring and decontamination of pesticide residues were to be undertaken and if all products exceeding the legal maximum residue limit had to be withdrawn from the market.

Although underestimated, regulatory costs could reach very large values such as US$4 billion (2013) yearly in the United States in the 2000s. Our analysis shows that if all regulations were respected, these costs would have jumped to US$22 billion (2013).

2.5.1 Several Studies Based on a Limited Number of Datasets

2.5.2 Estimated Costs

The costs of pesticide poisoning were evaluated at between about US$30 in Thailand and US$600 in Costa Rica (2013) per case, with each farmer/household using pesticides incurring annual costs of US$3 in China to US$187 in Sri-Lanka (2013) per year. In Central America, several authors have reported annual costs of US$32 to US$100 (2013) (see Vaughan (1993) and Villagrán (1976) cited by García (1998) and Castillo and Appel (1990) and Alvarado et al. (1998) cited by Cole et al. (2000)). These costs may be as high as US$850 (2013) per year for a rural establishment. At national level, health costs due to pesticide exposure have been estimated at US$1.1 million in Italy to about US$1.5 billion in the United States (2013) (Table 2.6).

These costs cannot be considered comparable, because they are influenced by several parameters, e.g. the type of pesticide used, the number of treatments applied, the degree to which farm staff spraying pesticides are protected etc., that may differ considerably between countries, with particularly marked differences between developed and developing countries. Moreover, in any given country, these costs have probably decreased over time, for two reasons. First, farmers have certainly become more aware of the effects of pesticide use on health and, therefore, probably protect themselves better against pesticide drifts. Second, some of the most dangerous pesticides have been withdrawn in many countries. Hence, on the one hand, costs actualized to 2013 values in US$ could easily be considered overestimates of current costs. On the other hand, human health costs were probably greatly underestimated at the time at which these reports were published, for three reasons. First, the frequencies of illness and death triggered by chronic exposure to pesticides have rarely been evaluated (see Sect. 2.5.5). Second, acute poisoning events generate various types of costs, and none of the studies performed to date has taken all these costs fully into account (see Sect. 2.5.3). Third, not all pesticide-poisoning events are recorded in databases or reported by farmers, particularly in developing countries (e.g. Lekei et al. 2014; Shetty et al. 2011). Indeed, some of the individuals carrying out pesticide spraying consider the symptoms of poisoning to be ‘normal’ and do not, therefore, pay much attention to them.

2.5.3 Non-fatal Cases of Acute Poisoning

Acute poisoning, leading to respiratory, gastrointestinal, allergic, and neurologic disorders, is commonly reported by farmers, and particularly by those carrying out pesticide applications (e.g. Hudson et al. 2014; Kishi et al. 1995). For instance, in a broad survey performed in 2010, Lee et al. (2012) found that 25 % of South Korean male farmers had suffered acute occupational pesticide poisoning, suggesting that there may be more than 200,000 cases per year across South Korea. About 12 % of these pesticide-poisoning cases led to the consultation of a medical doctor or hospitalization (Lee et al. 2012). In the United States, the incidence of pesticide poisoning events requiring medical care among the 3,380,000 agricultural workers is thought to be between 10 and 600/100,000 (Calvert et al. 2008 and references therein), corresponding to about 300–20,000 cases annually.

However, absence from work to recover from illness is only one of the various indirect costs associated with pesticide poisoning. Indeed, Wilson (1999a, 2000a, b, 2002a, 2003), who generated what is probably the most comprehensive and complete list of indirect costs to date, also identified (i) a decrease in productivity for farmers not taking time off from work to recover and just after their return to work, (ii) impaired decision-making and (iii) a loss of leisure time (Table 2.7). However, he recognized that it would be difficult to estimate the number of leisure hours lost and the decrease in working efficiency. Leisure hours were defined as ‘any time spent at home after work, such as time spent reading a newspaper, watching television, listening to the radio, playing a game or practicing a hobby, or time spent with the family’. As suggested by Becker (1965), Wilson evaluated leisure time costs on the basis of the hourly wage, given that any loss of leisure time would be likely to affect productivity at work.

Decreases in productivity at work and in decision-making abilities were estimated in a few other cost-of-illness studies (Table 2.7). However, none of these other studies evaluated the loss of leisure time as in the study by Wilson. However, Wilson did not estimate all the indirect costs due to pesticide poisoning and recognized that ‘the costs to the family were not taken into account’. These costs, including the time taken by family members to nurse the victim of illness, were investigated in cost-of-illness studies performed in Nepal (Atreya 2005, 2007, 2008; Atreya et al. 2012, 2013) and Ecuador (Cole et al. 2000; Cole and Mera-Orcés 2003; Crissman et al. 1994; Yanggen et al. 2003). The cost of childcare, which was estimated by Fleischer and coworkers (Table 2.7), is another indirect cost that was not considered by Wilson. Finally, an additional indirect cost, identified but not estimated by Devi (2007), is the time spent traveling to seek medical help. Thus, none of the cost-of-illness studies performed to date fully took into account all the various costs associated with acute pesticide poisoning.

2.5.4 Fatal Cases of Acute Poisoning

Suicide accounts for most of the fatal cases of acute poisoning. Gunnell et al. (2007) estimated that 250,000 people die from voluntary pesticide ingestion each year, accounting for 30 % of all suicides. The costs associated with such deaths cannot be considered an externality of pesticide use. Nevertheless, accidental pesticide poisoning, mostly in the occupational setting, may be fatal in some cases and the costs associated with such deaths can be treated as external costs. Fatal accidents due to occupational pesticide poisoning are very rare in some countries, such as the United States (1 case recorded from 1998 to 2005, Calvert et al. 2008), but may concern several tens or hundreds of workers per year in other countries with higher levels of pesticide use or in which workers are less well equipped with personal protection equipment (Fig. 2.1). For instance, Santana et al. (2013) reported that 2052 deaths, excluding homicides and suicides, were recorded as due to pesticide poisoning in Brazil, between 2000 and 2009. Half of these deaths concerned agricultural workers and most of them were caused by poisoning with organophosphate and carbamate pesticides.

The cost of fatal cases of accidental poisoning was estimated in only six sets of cost-of-illness studies: Ajayi et al. (2002), Choi et al. (2012), Khan et al. (2002), Tegtmeier and Duffy (2004), Pimentel and coworkers and Fleischer and coworkers (Table 2.7). Fatal cases have generally been ignored, mostly due to the type of cost-of-illness studies performed. Indeed, several of these studies involved interviews with a sample of farmers about the costs they incurred during pesticide poisoning incidents (Table 2.5). By definition, studies of this type cannot take deaths into account and, therefore, did not assess the cost of fatal poisoning events.

Two studies estimated the cost of these deaths, by evaluating the corresponding loss of work time. Ajayi et al. (2002) economically quantified the loss of life as the decrease in agricultural gross domestic product per habitant during the mean duration of an economically active life in agriculture set, in their study, at 50 % of 30 years. Similarly, Choi et al. (2012) estimated the loss of productivity loss due to premature death. Age- and sex-specific mean wages and employment rates were used as surrogates for per capita productivity for each sex and age group. Like Ajayi et al. (2002), Khan et al. (2002) included fatal injuries in their overall estimate of health costs. They attributed an overall cost of 224 million Rupees (US$15.1 million (2013)) to such injuries, but provided no details about how this cost was estimated.

David Pimentel and coworkers also considered the cost of fatal cases of pesticide poisoning. They used different sources for their estimates, based on the reasoning that no-one can place a precise monetary value on a human life. In their first estimate, Pimentel et al. (1980a, b) estimated the value of an individual human life at about US$1 million (about US$3.2 million (2013)). This value was considered to be the amount of money that industry and government might reasonably spend to prevent a death, but Pimentel et al. (1980a, b) wrote that ‘obviously it is much less than the true value of a human life’. In their article published in 1992, Pimentel et al. used the monetary ranges computed by the insurance industry and used an estimate of US$2 million (about US$3.4 million (2013)), which they considered to be conservative. Pimentel and Greiner (1997) and Pimentel and Hart (2001) used an estimate of US$2.2 million (about US$3.2 million (2013)) per human life, corresponding to the mean value of the damages paid to the surviving spouses of slain policemen in New York City, which they again considered to be a conservative estimate. Finally, Pimentel (2005, 2009) and Pimentel and Burgess (2014), in their most recent re-evaluation of pesticide externalities, used the United States Environmental Protection Agency standard of US$3.7 million (about US$4.7 million (2013)) per human life. Finally, Fleischer and coworkers estimated the cost of acute fatal poisoning events in Germany, using the estimate of US$2 million per life taken by Pimentel et al. (1993a) (see Waibel and Fleischer 1998).

2.5.5 The (Almost) Uncounted Costs of Chronic Exposure

The most striking feature of cost-of-illness studies on pesticide use is the lack of data concerning the long-term effects of chronic exposure. Several studies have highlighted the possible occurrence of severe health impairment, e.g. cancers, diabetes, depression, neurological deficits, respiratory diseases, fertility problems, cutaneous effects, effects on the unborn embryo, blindness, polyneuropathy, associated with chronic exposure to these chemicals. However, only six estimated the monetary costs of such impairment (Table 2.5). The other studies mostly stated that it was not possible to estimate costs due to chronic exposure because the corresponding illnesses, such as cancers, are multifactorial, making it difficult to estimate the number of cases directly due to pesticide exposure.

The six studies including the costs of health impairment due to chronic exposure provided very rough and incomplete estimates. Steiner et al. (1995) merely considered the cost of chronic illnesses to be as high as that associated with acute poisoning. Pimentel and coworkers based their estimates of the costs of chronic pesticide exposure on a rough estimate of the number of cancers per year. This number varied from 0.5 % of all cancers (Pimentel et al. 1980a, b, 1991a) to 6000 (Pimentel et al. 1991b), <10,000 (Pimentel et al. 1992, 1993a, b), <12,000 (Pimentel and Greiner 1997), 10,000 (Pimentel and Hart 2001) and between 10,000 and 15,000 cases (Pimentel 2005, 2009; Pimentel and Burgess 2014). All but one of these estimates were based on a personal communication from David Schottenfeld indicating that ‘US cases of cancer associated with pesticides in human are less than 1 % of the nation’s total cancer cases’ (see Pimentel et al. 1980a, 1992). Tegtmeier and Duffy (2004) did not provide another estimate for the United States: they incorporated the estimate of Pimentel et al. (1992) into their overall externalities of pesticide use. Houndekon and De Groot (1998) and Houndekon et al. (2006) took chronic exposure into account to some extent in their estimates, but it is impossible to determine to what extent. Indeed, they asked farmers how much money they spent on medication and medical consultations and how many working days per year they lost to illness, without specifying the type of health effect (acute or chronic and, for chronic effects, the illnesses concerned). Similarly, Pingali et al. (1994, 1995) and Rola and Pingali (1993) performed medical tests, providing an assessment of the ailments of each farmer or respondent and their seriousness. Such ailments may or may not be related to chronic exposure to pesticides. Finally, Wilson (1999a, 2000a, b, 2003) considered long-term illness diagnosed by a physician as arising from pesticide exposure. Given the small number of farmers examined (n = 203), long-term illnesses were probably underdetected.

This lack of counts is certainly the major flaw of all cost-of-illness studies performed to date. Indeed, there are good reasons to think that the costs of chronic exposure may be not only as high as those of acute poisoning, as stated by Steiner et al. (1995), but probably higher. One reason for this is that sufferers of irreversible illnesses, e.g. blindness, not only undergo short-term treatments, but may also incur long-term costs over a number of years, sometimes until they die. In their most recent re-evaluation of externalities, Pimentel (2005, 2009) and Pimentel and Burgess (2014) estimated the costs of chronic exposure to pesticides, restricted to cancers, reached US$1 billion, a value four times that estimated for the cost of acute poisoning events. However, this estimate did not include the loss of working days and the cost of death. By taking a death rate of 20 % for people suffering from cancers (Siegel et al. 2014) and a rather conservative estimated 3 months of absence from work for cancer treatment and recovery, and using the same costs of death as for acute poisoning, the costs of chronic exposure estimated by Pimentel and coworkers would have reached US$10.2 billion per year in 2005, 45 times the cost of acute poisoning.

2.5.6 Conclusions

The cost-of-illness studies reviewed here clearly show that the external costs relating to human health associated with pesticide use have always been strongly underestimated. First, most studies considered only the costs associated with short-term effects following acute poisoning events. This resulted in a considerably lower estimate of the overall costs, because severe illnesses, e.g. cancers, diabetes, depression, blindness, potentially triggered by chronic pesticide exposure are probably associated with much higher costs than acute poisoning incidents. The few studies to have taken serious illnesses into account yielded only partial and very crude estimates, for only one of the multiple possible illnesses, cancers, and only some of the costs concerned. Moreover, the cost-of-illness studies generally ignored several direct and indirect costs due to acute poisoning.

Another major flaw in cost estimates to date is the lack of consideration of fatal cases of pesticide exposure. Pesticide exposure-related deaths have sometimes been counted for assessments of accidental acute poisoning incidents, but deaths due to chronic pesticide exposure have been completely ignored. Indeed, even though some authors, such as Pimentel et al. estimated the number of cancers, they did not estimate the corresponding number of deaths. In addition, the value of life has probably been underestimated in the past. Pimentel and coworkers increased the estimate of this cost from US$1 to 3.7 million between 1980 and 2005, but, surprisingly, they retained this value (the value provided by the United States Environmental Protection Agency in the early 2000s) in their reassessments published in 2009 (Pimentel 2009) and 2014 (Pimentel and Burgess 2014). There is no standard concept or tool for placing a precise monetary value on a human life, but the reviews and meta-analyses of Kniesner et al. (2012), Lindhjem et al. (2011), Viscusi and Aldy (2003), and Viscusi et al. (2014) converged on a mean of US$9 to 10 million in 2013, which would correspond to a value of US$7.4 million in 2005. The human health costs estimated by Pimentel (2005, 2009) and Pimentel and Burgess (2014) should therefore be re-evaluated. If we use the re-evaluation of the estimated cost of chronic pesticide exposure of Pimentel (2005) proposed above, then overall human health costs in the article published by Pimentel in 2005 would have reached US$15.65 billion (2005), rather than US$1.23 billion (2005) as originally estimated.

Our review shows that health costs studies generally did not take into account fatal cases due to chronic exposure such as fatal outcomes of cancers. Doing so would increase those health costs by up to tenfold, e.g. US$15 billion instead of US$1.5 billion (2013) in the United States in 2005.

2.6.1 Various Types of Environmental Impact

2.6.1.1 Damage to Animals, Plants, Algae and Microorganisms

The main environmental impact of pesticides is probably the direct or indirect damage they cause to animals, plants and microorganisms, varying from minor injuries to death. This impact is not restricted to the area in and around fields. Indeed, during applications, pesticides drift away in the air and seep into the soil (Gil and Sinfort 2005; Pimentel 1995). Once in the soil, some soluble pesticides may be washed out in runoff water and during soil erosion, resulting in leaching into rivers and lakes (Chopra et al. 2011).

Damage to Vertebrates

Pesticide use has two main unintentional effects on vertebrate (mammals, birds, fish, reptiles and amphibians) wildlife: (i) deaths due to direct or indirect, e.g. feeding on contaminated plants and/or prey, exposure to high doses and (ii) poorer survival, growth and reproduction due to exposure to sublethal doses and a decline in or the elimination of habitats and food sources due to pesticides (Gibbons et al. 2014; Guitart et al. 2010; Sánchez-Bayo 2011).

Pesticides have a particularly strong impact on birds (Mitra et al. 2011), through direct deaths and the reduction or elimination of habitats and food sources. The indirect effects of insecticides, herbicides and fungicides have been identified as one of the main factors contributing to the decline of farmland birds in several European countries (Geiger et al. 2010). For example, herbicides and insecticides, together with certain agricultural practices, decrease levels of cereal grains, weed seeds and arthropods, thereby potentially contributing to the decline of bird species dependent on these resources for survival, e.g. Wilson et al. (1999) for granivorous birds and Hallmann et al. (2014) for insectivorous birds. In North America, the decline of several grassland birds, including songbirds in particular, is thought to be mostly due to a direct impact of insecticides (Mineau (2002) and Mineau et al. (2005) for Canada; Mineau and Whiteside (2006, 2013) for the United States). Birds are particularly susceptible to cholinesterase-inhibiting pesticides, e.g. organophosphates and carbamates, mostly because, unlike mammals, they have low levels of anticholinesterase detoxifying enzymes (Walker 1983). The extensive use of carbofuran, a carbamate, through a granular form resembling plant grains in North America has been reported to lead to the death of millions of birds annually (Mineau et al. 2012) (Fig. 2.4). Other birds, such as those predating on rodents, e.g. owls and other birds of prey, are also directly or indirectly poisoned by rodenticides in many developed countries (Christensen et al. 2012; Elliott et al. 2014; Langford et al. 2013; Thomas et al. 2011).

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Many studies have documented direct and indirect effects of both high and sublethal doses of pesticides on several wild vertebrates other than birds. Herbicide treatments can be lethal for amphibians. For instance, one of the surfactants added to glyphosate, the most widely used herbicide worldwide, has been shown to be highly toxic to several species of amphibians in North America (Relyea 2005). Recent reviews and meta-analyses have confirmed that several pesticides decrease amphibian survival (Baker et al. 2013; Egea‐Serrano et al. 2012). It has also been shown that pesticides have indirect and sublethal effects on this class of vertebrates, reducing their growth (Baker et al. 2013; Egea‐Serrano et al. 2012) and increasing the frequency of abnormalities (Egea‐Serrano et al. 2012). For instance, the herbicide atrazine, one of the most commonly used pesticides worldwide, adversely affects amphibians by disrupting metamorphosis, reducing antipredator behavior, decreasing immune function and increasing the frequency of infection (Rohr and McCoy 2010). The endocrine disruptor activities of atrazine, which decreases both time to metamorphosis and size at metamorphosis, can be enhanced by the presence of insecticides and fungicides. The effects of such mixtures of pesticides have probably played a major role in the global decline of amphibians (Hayes et al. 2006). Atrazine also disrupts several life history traits in fish (Rohr and McCoy 2010). Several pesticides, including atrazine, have been shown to have immunotoxic effects (Dunier and Siwicki 1993) and to cause oxidative stress (Slaninova et al. 2009) in fish, and these compounds can also interfere with olfaction in these organisms (Tierney et al. 2010).

Finally, pesticides also injure wild and domestic mammals. Rodenticides, particularly second–generation compounds, kill not only target pests, but many non-target rodent species (Elliott et al. 2014; Fournier-Chambrillon et al. 2004). Species abundance and diversity in rodent communities can also be altered by herbicides, particularly in situations in which these chemicals are used to convert bushwood to grassland (Freemark and Boutin 1995). Pesticides can also poison several domestic mammals (Wang et al. 2007; Berny et al. 2010). In the United States, and probably also in many European countries, the incidence of poisoning is highest in cats and dogs (Berny et al. 2010). These animals often wander freely around homes and farms. They are therefore much more likely to come into contact with pesticides than other domesticated animals. The presence of sprayed chemicals on fodder or of pesticide residues in feed for livestock may lead to fatal poisoning events in domestic farm animals, particularly in developing countries (Ajayi et al. 2002).

Damage to Invertebrates

Insecticide treatments controlling pests also have damaging effects on many non-target terrestrial arthropods in agroecosystems, including the natural enemies (predators, parasites and parasitoids) of agricultural pests (Croft and Brown 1975). Damage to these species may be greater than initially thought, because such damage can occur even at low non-lethal doses of insecticides (Desneux et al. 2007). For instance, sublethal doses of neonicotinoids (a new generation of insecticides) have clearly been shown to affect the foraging success, survival, colony growth, and queen production of honey and bumble bees (Henry et al. 2012; Schneider et al. 2012; Whitehorn et al. 2012) (Fig. 2.5). Beneficial arthropods are also affected by herbicides. This impact may be direct (Norris and Kogan 2000), but it is generally indirect. By killing weeds and non-target plants, herbicides reduce the fitness of many of the arthropods developing or resting on weeds, thereby decreasing the growth of their populations (Freemark and Boutin 1995; Norris and Kogan 2005). Even if herbicides do not actually kill non-target plants, they may still suppress flower formation in some species (Schmitz et al. 2014a), or markedly delay flowering time and decrease flower production in many other species (Boutin et al. 2014). As a consequence, herbicide treatments may indirectly decrease the fitness of pollinating insects in non-crop habitats during periods in which crop plants are unavailable for pollination. Egan et al. (2014) showed that changes in the structure and function of arthropod communities depend on species composition, crop rotation patterns and the timing of herbicide exposure.

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Pesticides can also have an impact on aquatic invertebrates (Rasmussen et al. 2013), particularly during pulses of contamination triggered by surface runoff and through tile drains during heavy rain. Invertebrates may also be injured during short pulses of contamination due to pesticide desorption from suspended solids or sediment particles. Finally, they can be poisoned via the ingestion of “polluted” particles. Several studies have found associations between pesticide concentrations and decreases in the numbers and abundances of taxa and changes to invertebrate community structure (e.g. Friberg et al. 2003; Liess and von der Ohe 2005; Schäfer et al. 2007, 2011, 2012). These studies were performed at many sites in Europe, Siberia and Australia, and the authors concluded that there was little doubt that pesticides were responsible for the observed changes in aquatic invertebrate communities. Liess and von der Ohe (2005) and Schäfer et al. (2007) showed that the number and abundance of aquatic invertebrate taxa could be compensated, probably through recolonization from undisturbed sections of the stream. Nevertheless, Beketov et al. (2013) found that pesticides had significant effects on regional species and family richness in Germany, France and Australia, with up to 42 % of the taxa from the recorded taxonomic pools lost. Furthermore, in Europe, effects were detected at concentrations considered environmentally benign in current legislation (Beketov et al. 2013).

Damage to Plants, Algae and Corals

Pesticides can accidentally injure crops. First, the crops protected by the pesticide may be damaged by it. In particular, some pesticides may disrupt photosynthesis, thereby decreasing both growth and yield. Such an effect has been shown for several fungicides, on many crops (Petit et al. 2012), and for some herbicides, on cotton (Reddy et al. 1990) and soybean (Hagood et al. 1980). Similarly, insecticide treatments may also lower yields when applied to lettuce (Toscano et al. 1982) and cotton (Youngman et al. 1990). Second, pesticides may disperse by drifting during spray applications. They may reach non-target crops in neighboring fields, weakening these plants and reducing yields. Such crop injuries have been reported, in particular, for aerial applications of glyphosate (e.g. Ding et al. 2011; Reddy et al. 2010). Third, as some herbicides persist in the soil, other crops (notably vegetables) in the rotation may be affected and display lower yields (e.g. Felix et al. 2007; Mahmoudi et al. 2011). These carryover injuries may be accentuated in fields previously treated with several herbicides. For instance, the addition of atrazine to mesotrione treatments in the year before planting has been shown to increase injury rates by 3–55 % in broccoli, carrot, cucumber, onion, and potato (Robinson 2008).

In some agroecosystems, field margins and boundaries (e.g. hedgerows, woodlots, etc.) are the only remaining habitats for many wild plant species, some of which are beneficial, considered of heritage value or protected (Türe and Böcük 2008). The long-term maintenance of their populations, particularly close to edges of crop fields, may be jeopardized by the drift of herbicide treatments. Several studies have shown that non-target plants are affected by herbicides (e.g. Freemark and Boutin 1995; Gove et al. 2007; Schmitz et al. 2014a), leading to short- and long-term changes in the richness and/or structure of plant communities (e.g. Egan et al. 2014; Gove et al. 2007; Schmitz et al. 2014b). Changes also occur among weed communities within crop fields (e.g. Andreasen and Streibig 2011). These changes in the composition of weed plant communities may reflect lower rates of reproduction in the species most affected by herbicides, as demonstrated by Boutin et al. (2014).

Aquatic plants, algae and coral species may also be affected by pesticide use. The large distances between sprayed fields and bodies of fresh and inshore waters should theoretically provide some protection, through the adsorption of some of the drift by bank vegetation and, probably, also through the dilution of the herbicides in water. In some ecosystems, aquatic and algal species are, indeed, considered to be not necessarily at risk (e.g. Cedergreen and Streibig 2005). However, there may be a major impact on aquatic species in bodies of water subject to intense agricultural runoff (Fabricius 2005). A textbook example is provided by the inshore waters of the Australian Great Barrier Reef. This lagoon has World Heritage status, but is widely contaminated with insecticides and herbicides (Haynes et al. 2000; Lewis et al. 2009; Packett et al. 2009). Kroon et al. (2012) estimated that >30,000 kg of herbicides enter the Great Barrier Reef lagoon each year. Despite their dilution in the water, concentrations exceeding 1 μg L⁻1 have been reported for some herbicides within the lagoon (Lewis et al. 2009). These concentrations may be high enough (Lewis et al. 2012) to have deleterious effects on corals (Cantin et al. 2007; Jones et al. 2003; Negri et al. 2011), seagrasses (Flores et al. 2013), foraminifera (van Dam et al. 2012), benthic microalgae (Magnusson et al. 2008, 2010, 2012) and coralline algae (Negri et al. 2011). The Great Barrier Reef is probably the most widely studied ecosystem threatened by pesticides, but other species in several other coastal water systems are also threatened by the effects of pesticide runoff. The ecosystems concerned include Chesapeake Bay in the United States (Hartwell 2011), the Seto Inland Sea (Balakrishnan et al. 2012) and two lagoons (Yamamuro 2012) in Japan.

Damage to the Soil Community

The effects of pesticides on earthworms (Yasmin and D’Souza 2010), microarthropods (Adamski et al. 2009), nematodes (Zhao et al. 2013), fungi (Morjan et al. 2002) and microorganisms (viruses, protozoa and bacteria) (Imfeld and Vuilleumier 2012; Lo 2010) within the soil may have major environmental consequences. The soil community plays a critical role in crop production and crop protection (Barrios 2007). These small organisms are essential to the functioning of all ecosystems, because they break down waste, thereby recycling the chemical elements required for life. Bacteria and fungi make nitrogen and other elements available to plants (Bonfante and Anca 2009) and, like nematodes, some soil-borne fungi are natural enemies of pest insects (Kaya and Gaugler 1993; Klingen and Haukeland 2006). Earthworms, which are widely recognized as ‘ecosystem engineers’, contribute to several ecosystem services through pedogenesis, the development of soil structure, water regulation, nutrient cycling, primary production, climate regulation, the remediation of pollution and cultural services (Blouin et al. 2013).

Damage Due to Interactions Between Species and Between Stressors

Species are not isolated from their environment or from other interconnected species. Pesticide exposure may, therefore, have indirect effects on biotic interactions, such as host-parasite relationships (Köhler and Triebskorn 2013). For instance, Rohr et al. (2008) showed that atrazine use was the best predictor of the abundance of larval trematodes (parasitic flatworms) in the declining northern leopard frog Rana pipiens. Pesticides can also increase the frequency of deformities associated with trematode infection in amphibians (Kiesecker 2002). More generally, interactions between pesticides and other environmental stressors may play a key role in the decline of amphibian populations (Mann et al. 2009). Synergistic effects of pesticides and natural stressors, such as heat, desiccation, oxygen depletion and pathogens, have already been documented in many other classes of animals (Holmstrup et al. 2010). Pesticides can also affect food webs and competition between species (Köhler and Triebskorn 2013). For instance, benomyl, a widely used fungicide, suppresses populations of arbuscular mycorrhizal fungi in grasslands, altering floral display at the patch level. Such changes have been shown to induce a shift in the community of floral visitors, from large-bodied bees to small-bodied bees and flies, and to decrease the total number of visits to flowers (Cahill et al. 2008).

2.6.2 Economic Consequences Considered to Date

The environmental impacts described above are obviously costly, in many ways. The various economic consequences considered in the 15 sets of studies are shown in Table 2.9.

Pimentel et al. (1980a, b, 1991a, b, 1992, 1993a, b), Pimentel and Greiner (1997), Pimentel and Hart (2001), Pimentel (2005), Pimentel and Burgess (2014), followed by Steiner et al. (1995), Khan et al. (2002) and Tegtmeier and Duffy (2004), tried to carry out a complete evaluation of the economic consequences of pesticide exposure in bees (Table 2.9). They evaluated colony losses, but also considered (i) losses of honey and wax due to bee colonies being either seriously weakened by pesticides or suffering losses when moved by beekeepers to minimize the risk of pesticide damage, (ii) losses of potential honey production because heavy pesticide applications on some crops may result in beekeepers being excluded from sites otherwise suitable for beekeeping, (iii) the lack of pollination due to losses of bee colonies and (iv) bee rental to compensate for this lack of pollination. Pollination losses were the greatest loss by far, accounting for more than 60 % of the total economic impact of pesticide exposure in bees.

A thorough analysis, such as that performed for bees, has never been undertaken for plants, microorganisms or animals other than bees. Considerations of the economic consequence of arthropod and microorganism depletion have focused on the loss of natural enemies of agricultural pests (Table 2.9). This loss of beneficial arthropods, fungi, bacteria and viruses increases pest pressure on crops. First, such losses allow the primary pests themselves to occur at higher densities. Several outbreaks of primary pests have been accounted for by the depletion of their natural enemies by pesticides (Bommarco et al. 2011; Hardin et al. 1995; Wilson et al. 1998). Second, many secondary pests, i.e. species that were once minor or unimportant crop pests, may become major pests if no longer controlled by their natural enemies (Hardin et al. 1995; Eveleens et al. 1973). Primary and secondary pest outbreaks due to the depletion of natural enemies have two main economic consequences: they increase pesticide use and decrease yields.

Pesticide resistance increases the amount of pesticide used, because higher doses are required to kill resistant pests. The use of alternative pesticides to which the resistant pests are still susceptible, or of a mixture of pesticides, which may be more expensive, may prove necessary. Resistance also decreases yields, because some pests become so resistant that they can no longer be fully controlled by pesticides or because the larger amounts of pesticides required to control resistant pests damage the crops treated.

The annual cost of mortality in birds and fish has been evaluated by multiplying the number of individuals actually killed due to direct or indirect exposure to pesticides by the estimated mean price of the individuals concerned. For birds, two additional types of environmental costs have been considered: the monitoring of species threatened by pesticide exposure and the re-establishment of endangered species, e.g. the bald eagle, Haliaeetus leucocephalus, affected by pesticides (Table 2.9).

Three economic consequences have been associated with damage to domesticated animals: the cost of illness, e.g. veterinary fees, the cost of dead livestock and the loss of productivity of animals weakened by poisoning, with affected individuals producing less milk, meat or eggs, for example (Table 2.9).

Yield loss is the principal economic consequence of accidental injury to crops from pesticide use. Contractors applying pesticides can be sued for damage to the crop during or after treatment. In many states of the United States, contractors applying pesticides must provide evidence of financial responsibility before spraying. Most are insured, to protect themselves against expensive lawsuits, and this increases the environmental cost of pesticide use.

2.6.3 Counting Environmental Costs: From Specific to Overall Costs

Some studies have focused on a particular impact. For instance, James (1995) specifically estimated the cost of bird losses in Canada. Some studies have been devoted to a specific crop in a specific area, such as the Punjabi cotton zones in Pakistan (Khan et al. 2002). Others have focused on externalities sensu stricto: Steiner et al. (1995) and Tegtmeier and Duffy (2004) in the United States and Pretty et al. (2000, 2001) for the United Kingdom. Steiner et al. (1995) therefore chose to ignore the costs associated with pesticide resistances and the loss of natural enemies, because these costs are mostly met by users (see also Pearce and Tinch 1998). Finally, Pimentel et al. (1980a, b, 1991a, b, 1992, 1993a, b), Pimentel and Greiner (1997), Pimentel and Hart (2001), Pimentel (2005) and Pimentel and Burgess (2014) in the United States, Jungbluth (1996) in Thailand and Ajayi et al. (2002) in Mali assessed the total environmental costs associated with pesticide use on all crops at the national level.

Estimates of economic costs due to environmental damages are therefore highly variable, from US$270,000 (2013) for the birds killed in Canada (James 1995) to about US$8 billion (2013) for total environmental impact in the United Sates (Pimentel et al. 1992, 1993a, b) (Table 2.8).

Similar variations were observed in estimates of fish losses in the United States. The cost of a fish varied from US$0.40 to US$10 between papers, resulting in estimates of the annual cost of fishery losses of between US$2.53 million (2013) in 1980 and US$170 million (2013) in 2005, 2009 and 2014, in studies by the same authors (Table 2.10).

2.6.4 Underestimated and Uncounted Costs

The costs provided by these studies are probably far from the actual costs. There are, indeed, several reasons for thinking that the counted costs were underestimated. In addition, several types of environmental damage have yet to be assessed.

2.6.5 Conclusions

The cost of the environmental impact of pesticides has been poorly investigated to date. Only 15 sets of studies have evaluated these costs, and these studies were actually based on only 11 independent datasets. Only six studies provided an overall cost assessment at national level. The pioneering work of David Pimentel in the United States remains the key reference, but this work dates from the 1980s and 1990s, with a partial update published in 2005, 2009 and 2014. Although Pimentel and coworkers provided the most complete evaluation of environmental impairment available, we have shown that this assessment was probably highly incomplete, with a strong underestimation of costs.

It should be borne in mind that the current environmental impact of pesticide use is probably very different from that during the 1980s and 1990s (see Sect. 2.8.4). In North American and European countries, the most dangerous and persistent pesticides (e.g. DDT, carbofuran) have been banned and partly replaced by less toxic and less persistent compounds, strongly decreasing the impact on birds and fish. However, other countries, such as India and China, are still producing, exporting and using DDT (van den Berg et al. 2012). Moreover, pesticide resistance has steadily increased over the last 30 years (Rex Consortium 2013). The doses of pesticides applied to many crops are, therefore, almost certainly higher than in the past, resulting in a greater impact on the environment.

To conclude on environmental costs of pesticide use, we show that they suffered large underestimation and most of them were never considered in the literature. They were nevertheless estimated to up to US$8 billion (2013) in the United States in 1992.

2.7.1 Defensive Expenditures for Pesticide Handling and Spraying

Only 13 articles have estimated the cost of defensive expenditures, and these estimates were based on only seven independent datasets (Table 2.1). This small number of studies considering defensive expenditures may be accounted for by defensive expenditures not being an externality sensu stricto. These costs are paid by farmers, which accounts for their lack of inclusion in studies focusing on the external costs of pesticide use such as those performed by Pimentel and coworkers.

Two groups of authors, in particular, have explored the defensive expenditures of farmers: Clevo Wilson (Wilson 1999a, b, 2000b, 2002b, 2003, 2005) and Athukorala et al. (2012) in Sri Lanka and Kishor Atreya (Atreya 2005, 2007, 2008 and Atreya et al. 2012, 2013) in Nepal. We were able to identify only one other studying exploring defensive expenditures, by Ajayi et al. (2002), in Mali.

In Nepal and Sri Lanka, farmers were found to spend a mean of between US$6 and US$32 (2013) per year on defensive expenditures (Table 2.11). Ajayi et al. (2002) estimated that farmers in Mali would need to spend US$30 to US$60 (2013) per year on equipment to ensure that they were protected against pesticide exposure. Wilson (Wilson 1999a, 2000b, 2003, 2005) and Athukorala et al. (2012) used data obtained directly from farmers to estimate the annual cost for the whole of Sri Lanka. They estimated these costs at between US$1 million (2013) if only 5 % of the farmers used pesticides and US$10 million (2013) if 20 % of the farmers used pesticides (Table 2.11).

In Nepal, defensive expenditures accounted for about 15 % of the total cost of pesticide use and 27 % of pesticide expenditure, i.e. the amount spent on purchasing pesticides in a year. Defensive expenditures were slightly higher (Atreya 2008) or slightly lower (Atreya et al. 2012, 2013) than the cost-of-illness, but essentially of a similar magnitude. In Sri Lanka, Athukorala et al. (2012) found these costs to be one quarter those for medical expenditure and one seventh the loss of earnings; cost-of-illness was thus 11 times higher than defensive expenditures (Wilson 1999a). Nevertheless, in this country, annual defensive expenditures corresponded to 12 % of the monthly income of a farmer (Athukorala et al. 2012; Wilson 1999a, 2000b, 2003, 2005). These costs, although low, could be a significant burden to farmers, whose incomes fluctuate greatly, due to adverse biotic, e.g. pest and disease damage, and abiotic, e.g. weather conditions, crop price fluctuations, conditions.

Several types of defensive expenditures have not been considered, probably due to data, time and financial constraints. The elements not analyzed include the purchase of more expensive sprayers less likely to malfunction and place the user at risk of exposure. They also include the time spent purchasing, cleaning and fixing defensive/protective equipment, and reading ‘warnings and instructions’. Precautionary drug treatment to protect against pesticide exposure and leisure time given up in favor of aversive behavior should also be taken into account. The estimates to date therefore almost certainly constitute the lower limit of the range of actual defensive expenditures paid by farmers to reduce their exposure to pesticides.

Moreover, in developing countries, these costs could probably be increased to levels much higher than those currently observed, as pesticide users often adopt few protective measures (Food and Agriculture Organization 2011b). Spraying is sometimes carried out without protection and even those farmers who do try to protect themselves generally limit this protection to the wearing of long-sleeved shirts and long pants. Low levels of income, awareness and education, the hot and humid climate, cultural taboos, fashion and discomfort are significant factors accounting for the lack of personal protection (Atreya et al. 2013) (Fig. 2.1).

Sivayoganathan et al. (1995) reported that some Sri Lankan farmers were keen to use protective measures but did not do so due to cultural taboos, such as wearing shoes in the field. The field is seen as a sort of “temple” because the land within it produces food. Another cultural taboo mentioned concerned the wearing of long pants during pesticide applications, which many farmers, especially the elderly, were reluctant to do, due to their low socioeconomic status.

Finally, not only might farmers be unable to afford adequate precautionary/defensive measures, but the protective gear required may be unavailable as it may not be sold by any shop to which the farmer has access. Hence, defensive expenditures have never been correctly counted, both because the actual expenses were not fully estimated and because they could potentially be much higher than they currently are, particularly in developing countries.

2.7.2 Defensive Expenditures for Safe Drinking Water

The presence of pesticides in tap water may be one of the key reasons for consumers buying bottled water or drinking purified or filtered water. These sources of water are much more expensive for the consumer than tap water. The excess costs of purified or bottled water over tap water could be considered as both a private cost borne by farmers if they drink such water and as an external costs to non-farming consumers buying such water. The production and transportation of bottled water also require the consumption of massive amounts of fossil fuels (Gleick and Cooley 2009). Finally, the bottles degrade slowly, and their incineration can produce toxic byproducts. Bottled water thus has an environmental impact between 90 and 1000 times greater than that of tap water (Jungbluth 2005). The resulting pollution can be considered as a negative externality for society as a whole. However, if the production, transportation and purchase of bottled water and all devices for water purification or filtration are to be considered as defensive expenditures, and hence as external costs, these expenditures should be made specifically to protect against pesticide residues. This relationship is anything but simple.

Consumers choose to drink bottled, purified or filtered water for two main reasons: because they think this water tastes better and/or is safer than tap water (Doria 2006; Doria et al. 2009; Dupont et al. 2010). Several factors are known to influence the public perception of drinking water quality: organoleptic properties, risk perception, attitude towards water chemicals, past problems attributed to water quality, trust in water companies, information from the mass media and family members (Doria 2010). Hence, the presence of pesticides, whether real or imagined, in tap water may be only one of a number of factors pushing people to buy bottled water and/or to drink purified or filtered water. Unfortunately, we were able to identify no study specifically exploring this question. Studies on factors influencing drinking behavior have considered chemical pollutants either as a general entity, i.e. with no specification of the type of chemical substance (e.g. Auslander and Langlois 1993), or have concentrated on lead, chlorine and/or water hardness, e.g. the survey of Statistics Canada (2009), which specifically mentioned chlorine. Pesticides, like other chemical substances including fluoride, nitrates, heavy metals and industrial chemicals, are sometimes specified, but, according to Doria (2010), their relevance to the perception of drinking water safety appears to be very limited or restricted to specific locations.

No specific data are available for pesticides, but several studies have explored the influence of chemicals on the water-drinking behavior of consumers, notably in Canada. In Toronto, 73 % of those questioned felt that tap water contained “some” or “a lot” of chemical pollutants, but half the households overall rated this source of water as “good” or “very good” (Auslander and Langlois 1993). In a more recent national survey of a representative sample of 1633 Canadians, 62 % felt that tap water posed no problem for health (Dupont et al. 2010). Only 12 % and 3 % believed that this source of water posed moderate or serious problems for health, respectively. In their study focused in one Canadian province, McLeod et al. (2014) also found that no more than 12 % of the 2000 respondents believed tap water to be unsafe to drink. Noteworthy, those respondents who believed tap water to be unsafe appeared more likely to choose bottled water (McLeod et al. 2014). In other countries with reliable supplies, surveys generally indicate that most people perceive the risk associated with drinking tap water to be small (Doria 2006). In low- and medium-income countries, in which tap water quality is often poorer, surveys of the motives for choosing bottled water over tap water have not been performed. However, in such countries, the average per capita consumption of bottled water is low.

In conclusion, the extra cost of drinking bottled, purified and filtered waters, rather than tap water, cannot be firmly attributed to the presence of pesticides. Of course, consumers indirectly pay for the monitoring and elimination of pesticides from the tap water they use, as these costs are passed on by water companies, through the billing process. We decided to count these costs as regulatory rather than as defensive expenditures because, as indicated in Sect. 2.4, monitoring and decontamination processes are mandatory in most countries: see the United States Safe Drinking Water Act (http://water.epa.gov/lawsregs/rulesregs/sdwa/index.cfm), for example.

Pesticides trigger defensive expenditures when they are detected in tap water at levels beyond the threshold considered acceptable, thus causing a decrease in quality. The monitoring of private wells, which are generally not regulated by public authorities, and the use of filtering/purifying devices for detecting and eliminating pesticides from these wells can also be considered as defensive expenditures.

Water quality violations may trigger aversive behavior, such as the purchase of bottled water. When such violations are due to pesticide contamination (e.g. Zaki et al. 1982), the increased in the purchase of bottled water in the area concerned may be considered defensive expenditures. Zivin et al. (2011) estimated that, in 2005, United States citizens spent US$47.15 million (2005) in response to element/chemical violations of water quality. They indicated that this estimate probably constituted the lower limit of the cost of defensive expenditures, because they only considered bottled water consumption and did not include other responses to violations, such as purchasing alternative beverages, e.g. juice, other actions people may have taken, e.g. boiling water, and more permanent responses, e.g. installing water filters. Zivin et al. (2011) did not provide details of the elements/chemicals responsible for the quality violations. We know only that they did not include nitrate, which was counted separately. It is therefore difficult to determine what proportion of the costs corresponded to pesticide contamination. Similarly, Dupont and Jahan (2012) estimated that Canadian households spent almost US$600 (2010) per year on tap water substitutes (purchase of bottled water and devices for filtering/purifying tap water), to decrease the perceived health risks associated with tap water consumption. Unfortunately, the influence of pesticides on this perception was not investigated.

The second type of defensive expenditures concerns the monitoring and decontamination of private wells and small-scale public systems. As indicated above, in the United States, state and federal authorities do not generally regulate these sources of drinking water. The householders concerned therefore pay for the detection of pesticides in these wells and their elimination. In the United States, 15 million households regularly obtain drinking water from their own private wells (United States Environmental Protection Agency 2002) and the groundwater in those wells may be contaminated with pesticides, particularly in rural areas (Toccalino et al. 2014). Pesticides, such as atrazine, deethylatrazine, simazine, metolachlor, and prometon are, indeed, regularly detected in groundwater and wells (Goss et al. 1998; Hallberg 1989; Ritter 1990, 2001; Toccalino et al. 2014). However, pesticide concentrations in North American domestic wells were found to be generally low. In Ontario, for instance, only six of the 1292 water-wells surveyed contained pesticide residues at concentrations above the maximum acceptable value (Goss et al. 1998). Similar findings were reported for the United States: for the 1993–2011 period, pesticide concentrations exceeded human-health benchmarks in only 1.8 % of the 2541 samples collected from 1271 wells in well networks distributed nationwide (Toccalino et al. 2014). However, pesticide contamination rates and concentration may reach higher values in some countries. In the Netherlands, several pesticides were detected in 27 % of groundwater samples taken from 771 monitoring wells. In 11 % of these samples, the concentration exceeded the upper regulatory limit (Schipper et al. 2008).

Worldwide, the most important contaminant of groundwater and private wells, in terms of health concerns, is arsenic (Nordstrom 2002). Arsenic contamination may have diverse sources, some of which are entirely natural, as in Bangladesh (Nickson et al. 1998). However, arsenic contamination may also result from local anthropogenic activities, such as mining (Mukherjee et al. 2006). In Canada and the United States, significant amounts of arsenic contamination result from the use of arsenic-based pesticides (Smedley and Kinniburgh 2002; Wang and Mulligan 2006).

According to the Massachusetts Department of Environmental Protection, testing a well for arsenic costs US$15 to US$30. Treatment systems for removing arsenic (reverse osmosis, activated alumina) cost at least US$400 per year (Sargent-Michaud et al. 2006). In addition to the costs of monitoring and testing, the presence of arsenic may also increase the consumption of bottled water (Jakus et al. 2009). As arsenic comes from diverse sources, which may vary over space and time, it is not easy to evaluate defensive expenditures due to arsenic-based pesticides. However, in the United States, where 15 million households regularly obtain drinking water from their own private wells, this cost might reach several hundred million US$ per year.

2.7.3 Defensive Expenditures to Avoid Pesticide Residues in Food: The Purchase of Organic Food

Consumers choose to purchase organic food for several reasons, some of which are linked to the externalities of pesticides and to a demand for pesticide-free food (Fotopoulos and Krystallis 2002; Misra et al. 1991; Squires et al. 2001; Tsakiridou et al. 2008; Williams and Hammitt 2001) (Fig. 2.6). Most consumers of organic food declare that the main reasons for this choice are connected to personal health and the avoidance of environmental damage (e.g. Huang 1996; Hughner et al. 2007; Magnusson et al. 2003; Saba and Messina 2003; Schifferstein and Oude Ophuis 1998; Schlegelmilch et al. 1996; Squires et al. 2001; Tregear et al. 1994; Wier et al. 2008). In Greece, about 90 % of general consumers consider organic food to be healthier than conventionally farmed food, and 75 % think that it is better for the environment; even higher percentages were recorded among the consumers of organic food (Tsakiridou et al. 2008). Animal well-being, taste or simply fashion are other factors less frequently proposed by consumers to explain their choices (Pearson et al. 2011). Parents of young children and babies are among those most likely to consume organic food, as a proactive measure, to prevent health problems (Pearson et al. 2011). Another reason cited for buying organic food is also linked to health, with some ill individuals choosing to buy organic food because they hope that it will help them to recover more rapidly (Pearson et al. 2011). Health is thus a key motive behind organic food consumption. Another reason often given for purchasing organic food is that it decreases damage to the environment, and this idea is generally supported by scientific evidence (e.g. Mäder et al. 2002; Gomiero et al. 2011). Buying organic food is thus partly a consequence of the perceived negative risk of pesticides to the environment and to the consumer.

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Organic food consumption can thus be considered, at least in part, as an externality of pesticide use if organic food is more expensive than non-organic, conventional foods. Comparisons of the organic and conventional food markets show that organic food is generally more expensive than conventionally produced food (e.g. Bonti-Ankomah and Yiridoe 2006). The excess cost of organic food varies considerably between countries and products (Bonti-Ankomah and Yiridoe 2006) and is dependent on several factors. However, according to several studies, the lower limit for this price premium would lie somewhere between 10 % and 20 % (e.g. Bonti-Ankomah and Yiridoe 2006; Rodríguez et al. 2008), although price premiums of between 50 % and more than 100 % were reported in the United States in 2013 for fruits and vegetables, respectively (see the web page on Organic prices of the United States Department of Agriculture Economic Research Service: http://www.ers.usda.gov/data-products/organic-prices.aspx#.VAmF0mTV_sk). This price premium, paid by the consumers of organic food thus corresponds, at least in part, to the consumers’ willingness to pay for avoiding pesticide risks (Onozaka et al. 2006) and, more precisely, to the hedonic estimation of willingness to pay for a reduction of the presence of pesticides in food. The range of values for the mean price premium of organic food has been confirmed by studies of the willingness to pay for organic food carried out with the contingent valuation technique. Consumers were asked to set a value on the premium they would be prepared to pay for organic food rather than conventionally produced food. These studies also highlighted considerably variability in the responses obtained (e.g. Zehnder et al. 2003; reviewed by Bonti-Ankomah and Yiridoe 2006), but they frequently suggested that the minimum value was about 10–20 % (e.g. Bonti-Ankomah and Yiridoe 2006; Gil et al. 2000; Onozaka et al. 2006; Rodríguez et al. 2008).

The worldwide organic food market was of the order of US$64 billion in 2012 (Sahota 2014), equally split between Europe (US$29 billion) (Schaack et al. 2014) and the United States (US$29 billion) (Fitch Haumann 2014). In Europe, the organic food market in 2012 represented about US$9 billion in Germany, US$5 billion in France and US$2.5 billion in the United Kingdom (Schaack et al. 2014). The world market for organic food has grown considerably over the last 15 years: it almost tripled between 2000 and 2008 and continued to grow thereafter, from US$50 billion in 2008 to US$64 billion in 2012 (Sahota 2014).

If we assume that prices in this market are 20 % higher than those of conventional food and that about 50 % of the reasons for consumers choosing organic food are directly linked to the avoidance of pesticide risk (e.g. Schifferstein and Oude Ophuis 1998), then the added cost of pesticide use is about 10 % of the total market value of organic food. This amounts to US$2.9 billion for the United States and Europe, and about US$0.9 billion for Germany, US$0.5 billion for France, and US$0.25 billion for the United Kingdom. Griffith and Nesheim (2008) used hedonic prices and purchase quantities for 2003 and 2004 in the United Kingdom to estimate the aggregate lower limit of willingness to pay for organic products. They obtained a value of about 22 % of the annual expenditure on organic products, corresponding to about US$0.55 billion, based on the figures obtained for the organic market in the United Kingdom in 2012. Griffith and Nesheim (2008) estimated that about 20 % of the lower limit of the willingness to pay was directly linked to health and environmental concerns – about US$110 million, corresponding to 44 % of our estimate of US$0.25 billion.

2.7.4 Conclusion

Defensive expenditures have rarely been considered among the external and “hidden” costs of pesticide use. For instance, we found no study considering the defensive expenditures of both farmers and consumers. In particular, the consumption of organic food as a defensive action against pesticide residues has never been fully considered as a negative externality of pesticide use. Indeed, all studies to date on the economics and rationale of organic food consumption have been completely disconnected from studies analyzing the benefit-cost ratio of pesticide use.

In general, aversive actions have been little studied and, when considered, they have generally been restricted to the protection of the body and respiratory system by farmers handling or applying pesticides. However, these costs are only part of the costs directly borne by farmers.

Furthermore, aversive actions could be carried out on a much wider scale than is currently the case. This is certainly true for protective clothing, which is rarely worn by farmers in most developing countries, and for the monitoring and decontamination of drinking water. If all owners of private wells carried out monitoring and were equipped with a filter/purifier, or if the consumption of bottled water continues to grow, then defensive expenditures to avoid residues in drinking water could rise exponentially. However, it should be borne in mind that these costs are somewhat linked to cost-of-illness. If tap water contains pesticide residues at levels that may injure human health, then an increase in defensive expenditures should lead to a decrease in cost-of-illness. Put another way, some of the current cost-of-illness could be due to a lack of aversive action. Alternatively, an increase in defensive expenditures might decrease the overall cost of pesticide use if these additional defensive expenditures are overcompensated by the decrease in cost-of-illness they trigger. Similarly, an increase in the consumption of organic food might decrease the cost-of-illness by reducing chronic illness although the relationship between exposure to low pesticide doses and chronic illnesses remains very difficult to quantify.

Here, we show that defensive expenditures have rarely been considered in the literature of pesticide use cost. These costs include at least the extra cost of organic food consumption due to aversive behavior linked to pesticide use. This cost reached more than US$6.4 billion worldwide in 2012.

2.8.1 A Small Numbers of Estimates

We found only ten independent groups of papers combining estimates of regulatory, environmental and human health costs at the national level. These groups of studies are those of Ajayi et al. (2002) for Mali, Houndekon and De Groot (1998) and Houndekon et al. (2006) for Niger, Jungbluth (1996) and Praneetvatakul et al. (2013) for Thailand, Khan et al. (2002) for Pakistan, Pimentel and coworkers (Pimentel et al. 1980a, b, 1991a, b, 1992, 1993a, b; Pimentel and Greiner 1997; Pimentel and Hart 2001; Pimentel 2005; Pimentel and Burgess 2014), Steiner et al. (1995) and Tegtmeier and Duffy (2004) for the United States, Pretty et al. (2000, 2001) for the United Kingdom, and Fleischer and coworkers (Fleischer 1999; Waibel and Fleischer 1998; Waibel et al. 1999) for Germany.

2.8.2 Overall Costs Are Underestimated