aIncluded in the Red/Yellow list of threatened species as "Requiring attention".

The many indirect impacts of pesticide use mean that species that are not directly affected by spraying still have changed conditions after a field has been sprayed. Depending on the products used and the situation in which they are used, the indirect effects will in many cases be the most extensive. Since the indirect effects are not due to toxic effects on the organisms considered, e.g. birds, one cannot protect the organisms by more stringent requirements concerning the toxicity of the products.

There is nothing to indicate that mammals are generally affected in the same way as birds by side effects of pesticide use. In studies of the hare’s population dynamic, pesticides are not mentioned as a possible explanatory factor in the decline in population in the 1960s (Hansen 1991). Among mammals, there are not as many species that are dependent on insects and similar in cultivated fields. For the insectivorous species there are no studies throwing light on the problem.

Unsprayed headlands

Unsprayed headlands are established along fields normally sprayed with herbicides and other pesticides. Unsprayed headlands are in principle a one-year scheme that forms part of the crop rotation. Owing to problems with weeds, unsprayed headlands are not used in crop rotations with beets and potatoes. Placing the zones at the edges of fields, possibly against a hedgerow or other shady vegetation, is normally the least costly solution for the farmer because the yield is often lower along the edges of fields owing to shading and competition for nutrients and water. Unsprayed headlands provide wild plants in the zones with a better possibility of survival and growth. At the same time, they act as a buffer against spray drift to the adjacent small biotopes. After each winter, many of the insect species return to the field from uncultivated areas. Since reproductive capacity is high, spray-free conditions have a big and rapid effect. Insects therefore react with a big and rapid increase in number even though the buffer zones are only one-year zones and are placed differently from year to year. From Danish trials in cereals, we know the effect of unsprayed headlands at pesticide loads slightly below the national average. (Hald et al. 1994). Under these conditions, the total insect densities were just under 50% higher and the species diversity approx. 25% higher than in the sprayed field. The insects considered to be the preferred food of farmland birds were no less than 65% more numerous. These average figures cover a big variation within different insect groups. The effects were greater on larvae than on adult insects and greater on beetles, flies, parasitic wasps, plant bugs and plant hoppers than on other groups. The biggest effect was observed among the herbivores and fungivores. The last-mentioned effects are mainly indirect, since herbicides and fungicides remove the food eaten by these insect groups (Hald et al. 1994; Reddersen et al. 1998).

Importance of headland vegetation to the fauna

The absence of herbicides, in particular, plays a major ecological role in headlands because it results in a much richer flora, with the possibility of both flowering and seeding. The ground cover with different plant species is important as a food resource for herbivores, insects that visit flowers and seed-eaters, although the microclimate and the possibility of concealment also play a role (Reddersen et al. 1998). Buffer zones bordering on uncultivated small biotopes also mean that ground beetles and rove beetles can overwinter there and return to the field in the spring. Certain common insect groups occur in the field only in the buffer zone along uncultivated small biotopes, whether, as in the case of plant bugs and plant hoppers, they live there or, as in the case of harvest spiders, ants and woodlouse, they are simply guests from the boundary biotope.

Only a few species of bird and small mammal live in the cultivated field, while the many species living in uncultivated small biotopes often go on foraging excursions in the field edge: that applies, for example, to different species of mice and species of bird, such as yellowhammer, partridge and pheasant. Such species naturally potentially benefit from the ample food resources available to them in unsprayed headlands.

Effect of pattern of pesticide use on flora and fauna

Wild plants are an important and basic element in the agroecosystem’s species diversity, see earlier sections. Use of reduced herbicide dosages could be expected on average to leave more weed in Danish fields. Surviving plants could form a food resource for herbivorous insects, among others, if the food quality of the host plants were adequate. This has been studied in the case of the knotgrass leaf beetle and the herbicide Glean (Elmegaard, Kjær 1995; Kjær, Elmegaard 1996).

There is generally reason to assume that sublethally treated plants can form a food resource for herbivorous animals. The treated plants do not always have the same quality as untreated plants, but for herbivorous animals and seed-eaters, insects visiting flowers, etc. a little food is better than none.

However, there has not been any clear tendency towards an increase in the amount of weed due to the widespread use of reduced dosages – at any rate not when the registration takes place at harvest time (Kristensen 1994). In other words, farmers have become better at spraying optimally and are achieving the same effect with reduced dosages. Studies have not been carried out of the development of weed throughout the season as a function of changes in the spraying pattern.

Reduced dosages of insecticides are not widely used, but are recommended in certain cases by the Agricultural Advisory Service.

For species with LD₅₀ close to or above the corresponding field dose, repeated spraying will also increase the probability of the individual being hit and affected by an application. The advantage or disadvantage of split dosing (i.e. the full dose is distributed over several applications with reduced dose) for harmless organisms thus seems to depend on the species’ sensitivity to the dosage used. It is therefore interesting to compare these considerations with observations in the field. Split dosing is at present of greatest relevance for fungicides. The side effects are particularly effects on harmless fungi and the derivative effects of reduced fungus densities.

In a study of the effect of split dosing of a leaf fungicide, the effect of 1/3 dose proved to be relatively high, and particularly the effect of the first two (beginning-end May) applications. Depending on the year, 3 x 1/3 dose often has a greater effect on both harmless and harmful fungi than a single full dose. (Reddersen et al. 1998). The results of such comparisons must be expected to depend on the year (the weather) and the actual and relative times when the individual applications are carried out. It is probable that the same applies to fungicides as to herbicides: the recommended dose is normally more than enough to achieve the desired effect. Placed optimally, the effect of 3 x 1/3 dose on sensitive fungi will therefore be more comparable to 3 x full dose than 1 x full dose.

Clear analyses of the importance of split dosing, reduced dosages, spraying time, and weather conditions require simulation models and sensitivity analyses. Such tools are only now being developed.

Since the beginning of the 1980s, farmers have traditionally sprayed winter cereals twice with fungicides. Ten years ago, they used two full doses. Today, optimisation of the application time and falling grain prices have made two applications of 1/3 of the dose the most economically optimum and, on average, that has replaced two applications of the full dose. Depending on the product, the reduced dosages produce a varying dosage response. Split dosing has not increased the risk of routine application of insecticides. However, when traditional spraying with fungicides is carried out around earing, many farmers are tempted to add an insecticide if they have seen a few aphids in order to avoid perhaps having to spray again one week later, when the damage threshold has been exceeded. All else being equal, the direct effects of pesticides will be less with use of reduced dosages.

Product mixtures in preparations and sprayer tanks

The effect of product mixtures is an unknown factor in the assessment of the effects on the pests and non-target organisms. That is because the substances can affect each other’s toxicity. In many crops the usual thing is to apply several active ingredients in one go, either as finished preparations or by mixing them in the tank. In Denmark, the extent to which tank mixtures are used in practice has not been investigated. The substances can have an additive, intensifying or impeding effect. It is not possible to get a picture of which substances intensify each other’s effect, partly because there are many possible combinations and partly because it depends on the species on which the mixture is tested, the dosages used, and the timing of the test. It has been proven that the effect of many insecticides is intensified by the presence of other substances with an entirely different toxic effect. The intensifying effect may be due to the fact that the substances act on the enzymatic degradation of the toxin, on the penetration of the substance or on the adsorption of the substance. In some cases, an up to 100 times increase in toxicity has been observed, which can radically change a substance’s hazard classification. In 32% of the cases found, where an insecticide is involved, the effect on mammals increases by a factor of 10 or more.

An intensifying effect has also been demonstrated in the interaction with ergosterol-inhibiting (EBI) fungicides and pyrethroids. In the UK it has been found that the fungicide prochloraz increases the toxicity of the insecticide lambdacyhalothrin to honey bees 18 times at field dosages. EBI fungicides also affect the toxicity of the organophosphorous insecticides dimethoate and malathion to partridges when the birds have been exposed to an EBI product before being exposed to the insecticides. The effect is assumed to be due to activation of the enzymes that are responsible for the metabolism of the organophosphorous products into the active metabolite. Since the EBI product, together with dimethoate and the pyrethroids are, respectively, the most widely used fungicides and insecticides in Denmark, it is highly probable that intensifying effects occur and cause greater effects than expected on non-target organisms. This is also substantiated by the fact that, in the UK, greater effects than expected have been observed in bees from applications of product mixtures under normal conditions.

Effects of pesticides in forestry

Very few studies have been carried out specifically of the environmental impact from the use of pesticides in forestry. Such analyses of the environmental consequences of silvicultural use of pesticides as have been carried out are therefore to a very great extent based on experience from studies outside forests and on knowledge of how the forest ecosystem functions (Elmegaard et al. 1996; Strandberg 1998). In the case of the higher fauna, it is estimated that direct effects of pesticides are very low in forests, but that there is a direct effect on the lower fauna from pesticides used to prevent and control aphids and other harmful insects in connection with the production of Christmas trees and ornamental greenery. It is particularly the nordmann fir cultures that are treated with insecticides (Østergaard et al. 1998).

In the case of indirect effects, it is primarily the use of herbicides that has effects on both higher and lower fauna. It is principally in ornamental greenery and Christmas tree cultures that one sees indirect effects on fauna, but the use of herbicides in connection with clear cutting in wood-producing forestry also has an effect by removing a large part of the food resources for a time and by affecting the microclimate and the possibilities of finding concealment (MacKinnon, Freedman 1993). In forestry, one operates with long rotation times, determined by the generation length of the different species of tree, which can easily exceed 100 years. Since long-term effects on fauna cannot be excluded, the lack of such studies of pesticide use in forests is unfortunate.

5.1.1 Conclusions

The effect of the use of pesticides on fauna can be divided into direct toxic effects and indirect effects. The direct toxic effects are caused particularly by applications targeted on harmful species of fauna. Besides the harmful species, a wide range of related species are seriously affected, including the pests’ natural enemies and pollinating insects, depending on the species’ dispersal in time and space. Pesticides also affect fauna by removing their food resources through the effect of herbicides on flora, see section 5.2. The effects must also be seen in relation to other effects, including those of crop rotation and soil treatment.

The following specific conclusions can be drawn:

The indirect effects of the use of pesticides are usually caused by changes in the species’ food resources and are due to both insecticides, herbicides and fungicides etc., since the fauna’s food resources are to be found among insects, plants, fungi, etc. This affects both harmful and beneficial species and species of higher fauna. The indirect effects in farmland and forests are generally thought to be considerably more extensive than the direct effects.

Farmland bird life has been the subject of extensive studies, which have shown a significant decline in most of the species associated with farmland in the period 1976-1996. The decline is due to a number of factors, including the use of pesticides, crop rotation and the management of hedgerows and other small biotopes.

Dressed seed and granulates present a risk to birds and mammals.

There are very few amplifying studies of the effect of pesticides on organisms that live in the soil. Since the diversity and number of processes in the soil are great and exhibit a similarly big variation in sensitivity to pesticides, the long-term effect has not been adequately clarified.

The way in which pesticides are used can affect the environmental load. For example, the use of reduced dosages can reduce the direct effects on fauna. In the case of herbicides, it is uncertain whether reduced dosages have been of much benefit to fauna because very effective control is achieved with low dosages.

Split dosing of fungicides in cereals constitutes a great risk to fungal flora and the associated fauna. Insufficient information on spraying practice is available for analyses at national level.

When pesticides are used in forests, it is the indirect effects that cause the main impact on fauna. The long-term effects on fauna cannot be judged owing to a lack of tools/knowledge.

It has been documented that the use of product mixtures involves a risk of intensified effects, but the significance of this in the field is unclear and there is a lack of information on treatment practice.

The sub-committee recommends that more consistent and systematic use than hitherto be made of permanent unsprayed headlands and protection borders. In this connection, it must be ensured that linked dispersal corridors are established. It is also recommended that the natural environment be reinstated by introducing animal species with a small dispersal potential.

5.2 Effect on the flora in cultivated and uncultivated terrestrial ecosystems

Repeated spraying changes the flora

Trials have shown that repeated and effective spraying of herbicides year after year reduces the number of wild plants in cultivated areas. In semi-cultivated areas, herbicides have been applied occasionally – usually in the form of products against dicotyledonous species with a view to promoting grass species and increasing the fodder value of grazing/hay production. Artificial fertiliser has also often been applied to increase fodder production on the area. This has also been done on land with section 3 status under the Nature Protection Act, which does not fully protect against adverse effects of the use of herbicides and fertilisers since the law simply "freezes" existing practice.

Spray drift

During spraying, there is spray drift to the surrounding areas, although this can be reduced with modern spraying equipment. 10-20 years ago, direct herbicide-spraying with an end nozzle on the spray boom was recommended and practised in many places. Spray drift in connection with normal use of herbicides is a particular problem for edge biotopes because of their relatively large edge area in relation to the total area. For the same reason, most of the small biotopes bordering on rotation fields are seriously affected by eutrophication, caused particularly by artificial fertiliser spread centrifugally on adjacent fields.

Herbicides and nutrients change the flora

As a result of the widespread load of herbicides and nutrients, the composition of the vegetation in these types of biotope has developed towards high-growing grass and herbaceous vegetation with low species content and completely dominated by such species as great nettle, wild chervil, common couch grass, cocksfoot grass, tall meadow oat, corn thistle, grey magwort and goosegrass. These species tolerate most herbicides relatively well and can use the ample supply of nutrients for rapid and high growth, thereby shading out most competitors. In such disturbed biotopes, one typically finds increased plant production and plant biomass per unit of area. In all, there are thus fewer species and less locally and regionally characteristic vegetation, while species with high nutritional requirements predominate.

Reestablishment of the flora is difficult

In work on reinstatement of the natural environment, it has been considered whether the original but lost natural values can be recreated by blocking the negative, direct and indirect input of herbicides and fertilisers. Experience with this is disappointing – in practice, there is substantial ‘ecological inertia’ acting against reversion to former types of vegetation, mainly due to two factors:

A tendency towards long-term maintenance of the high-level nutrient status of the biotopes in almost every type of soil except the most sandy and dry. This maintains the competitive high-growing type of vegetation, which effectively prevents remigration of original species.

The species’ poor and often short-lived seed pool and poor dispersal abilities make reestablishment difficult. There are usually big distances, few and poor dispersal paths or actual barriers to dispersal, all working against remigration.

Effects on the seed pool and wild plants in the field

The occurrence of the most common plants in Danish fields generally became less frequent in the 20-year period from 1967-70 to 1987-89 (Andreasen et al. 1996). Weed control is practised with the precise intention of minimising some species (weeds) and favouring others (crops), and the fewer wild plants there are, the less need there is to spray the field in question. There has been a heavy decline in some species because of the intensive weed control with herbicides. In a 25-year period, from 1964 to 1989, increased use of herbicides reduced the average number of species in the seed pool from 12 to 5 per field (Jensen & Kjellsson 1995). In the same period, the total number of viable seeds in farmland halved.

With the object of investigating the relationship between the effects of herbicides on species of weed and the reduction in the number of viable seeds in farmland, a study was carried out to correlate effects at full dose of herbicide mixtures, which were frequently used in cereals in the 1970s and 1980s (effect figures from PC Plant Protection, Per Rydahl, personal communication), with the percentage decrease in the seed pool of individual species (Kjellsson & Madsen, 1998a). However, no clear relationship was found, possibly because other factors also influence the seed pool (e.g. crop rotation, soil treatment and level of fertilisation). A British study shows that the crop rotation plays a major role, affecting the size of the seed pool (Jones et al. 1997).

Spray-free buffer zones

Spray-free buffer zones are established along fields with annual crops, which are normally sprayed with herbicides and other pesticides as described in section 5.1. The wild flora generally reacts strongly and immediately to spray-free conditions because there is plenty of seed in the seed pool to ensure a high plant density. Many wild plants are characterised by a flexible and largely individual growth potential. One could in particular expect considerable quantitative changes and, only secondarily, a number of qualitative changes in the flora. For example, in buffer zones in winter wheat, the total biomass of wild plants was seen to increase 10 to 40 times within a year (Reddersen et al. 1998). The increased growth and seed production would also significantly increase the seed pool after 2-3 years (Hald et al. 1994; Kjellsson & Rasmussen 1995). Buffer zones in cereal crops would have a beneficial effect on common dicotyledonous plant species such as shepherd’s purse, corn pansy, forget-me-not, speedwell, chickweed, white goosefoot and dead nettle, together with a large number of less common species. At the same time, some grasses – particularly annual meadow grass – would decline somewhat. In the short term, one could thus primarily expect an increased total biomass of wild plants caused by the increase of a rather limited number of common species, while an improvement for the less common species would generally require more permanent buffer zones. Hald et al. (1994) thus did not observe any clear increase in the number of species in permanent buffer zones over 5-6 years.

Permanent unsprayed, fertiliser-free headlands

Like buffer zones, unsprayed headlands could help to protect well-preserved vegetation in small biotopes, where such vegetation still exists. In most places, however, the vegetation in small biotopes has been seriously affected by both herbicides and fertilisers in the last few decades. To ensure recolonisation, which normally takes place very slowly, it would be necessary to establish permanent spray-free and fertiliser-free edge zones.

Effect on seed production

It has been shown that sub-lethal doses of herbicides lead to a reduction in plants’ seed production. The reduction is related to the dose, and this is probably directly related to a smaller biomass production (Rasmussen 1993; Andersson 1994). For example, a sub-lethal dose (1/2 the normal dose) of isoproturon resulted in a 50 per cent reduction in seed production in common pennycress (Hald 1993). In a Danish study it was shown that seed production in unsprayed field plots was 6-14 times higher than the seed production in sprayed plots (Kjellsson, Rasmussen 1995). At the same time, spraying with herbicides (dichlorprop + 2,4-D/MCPA in normal dosage) resulted in a smaller proportion of the surviving plants being able to propagate. It has been shown that some herbicides (tribenuron-methyl and, to a lesser extent, MCPA) can result in a smaller seed size in some species, such as black bindweed, goosegrass and common pennycress (Andersson 1994). With this knowledge and different scenarios for herbicide use (0 and 100%), it would be possible to model and estimate the effects on the seed pool in farmland and the consequences in the form of changes in the frequency and composition of the vegetation (Kjellsson, Rasmussen 1995; Madsen et al. 1996, 1997, 1999).

The wild flora in hedgerows and small biotopes

Herbicides are normally not used in hedgerows and small biotopes, so any impact on these areas is due to unintended effects from agricultural use of herbicides, e.g. through spray drift (see section 4.1.6.2). During a 5-year period, the flora in Danish field hedges showed a slight tendency to contain more species along unsprayed edges of fields than along sprayed ones (Hald et al. 1994). A 50% standard trial treatment with the herbicide fluroxypyr of a fallow field on sandy soil in the Netherlands reduced the content of species (Kleijn, Snoeijing 1997). Effects on survival at lower dosages (5 and 10 %) were only found for a few species in some years. In another Dutch study, de Snoo (1997) found, in a three-year trial, increased species diversity in unsprayed field edges in sugar beet, potatoes and winter wheat, primarily as a consequence of an increase in the number of dicotyledonous plants.

Effect on the flora through spray drift

There exists a set of standardised values from Germany for spray drift. The values are based on 16 field trials in the period from 1989 to 1992, where, on the basis of a 95% percentile, a so-called "realistic worst case" was established (Ganzelmeier et al. 1995). On the basis of these values and known effect thresholds (PC Plant Protection), a qualitative analysis can be carried out of the effects on plant growth. The method used by Ganzelmeier is not necessarily ecologically relevant. The deposition was measured during spraying with a single spray plume, and the spray product was collected on flat targets laid on the ground. Such data are relevant for estimating the effect on plants that have not yet germinated and for "flat" areas like ponds and lakes. However, plants that have germinated and are established "catch" more of the spray product. Both Davis et al. (1993) and Bui et al. (1998) found that different "targets" had different "catch efficiencies". "Targets" that rise above the ground and have a complex structure catch more spray product than objects lying flat on the ground. Finally, Nordby and Skuterud (1975) found that the dose of spray product needed to trigger a given effect is smaller for plants in the drift zone than for plants directly under the sprayer. An American study showed that sweet cherry is damaged by herbicide drift (chlorsulfuron) at doses down to 1/100 the normal field dose (Al-Khatib et al. 1992). At doses 1/3 to 1/10 the normal dose, several herbicides (2,4-D, glyphosate) caused significant damage. A number of studies (Marrs et al. 1989, 1993; Davis et al. 1993, 1994) of the effect of spray-product drift on plants showed that plants were affected up to 50 m from the sprayed area. However, the majority of the plants were only affected over a distance of 0 to 5 m from the field.

Effects on forest-floor flora

Actual studies of the effect of pesticides on forest-floor flora are rare, so the assessment has been based on knowledge concerning the mode of action of herbicides and knowledge concerning the ecology of forest-floor flora. It is only the effect of herbicides that has been considered because, according to Elmegaard et al. (1996), there are not stated to be any known effects from other groups of pesticides on forest-floor flora.

Østergaard et al. (1998) state that glyphosate is applied once before clear cutting of deciduous trees and conifers and once or twice in the first few years thereafter. This application practice is thought to have a radical effect on the forest-floor plants, so that all individuals of the flora associated with the type of forest in question may be eradicated (F. Rune, Danish Forest and Landscape Research Institute, personal communication). There will still be a seed pool, but it will also be seriously reduced, partly because of the repeated spraying and partly because of changes in the forest climate for a time when renewal takes place by means of clear cutting. Plants from seeds that germinate in the first period after clear cutting, possibly provoked into germinating by the increased amount of light, will often die, either as a consequence of spraying or because of drought, frost-nipping or scorching by the sun. In addition, forest-floor plants have a short-lived seed pool, so these species have no possibility of surviving an unfavourable period during the seed stage (Graae 1999). Even if the treatment is carried out over a limited number of years, it will have a serious impact on the forest-floor flora. Seen over a rotation period, the relatively small treatment frequency index thus has a big impact. In Christmas tree and ornamental greenery cultures, where herbicides are used during the entire lifetime of the culture, there is virtually no forest-floor flora. So massive are the effects that there is no flora at all – which is, of course, the whole idea of using herbicides in these cultures. When Christmas trees and ornamental greenery are produced without herbicides, a flora develops that is dependent on the way the soil is treated, and the method of renewal can be more or less authentic (M. Strandberg, DMU, personal communication).

The rate of recolonisation by forest-floor flora species is very slow (0 - 1m/year) (Brunet & von Oheimb 1998), so it is only in areas in the immediate vicinity of the forest with elements of the original flora that one can expect partial recolonisation by the species that have disappeared. Indeed, Brunet and von Oheimb (1998) also recommend that the slow recolonisation rate of forest-floor flora be taken into account in forestry planning. Inghe and Tamm (1985) have shown in Swedish forests that individuals of blue anemone are more than 40 years old, and it is not unlikely that many species of forest-floor flora can reach a higher age than trees. Danish studies have shown that forests with long continuity (>200 years) have a better developed forest-floor flora than younger forests (Graae 1999), so for Danish forests, too, it is reasonable to expect impacts on the forest-floor flora if herbicides are used in connection with clear cutting and reestablishment of cultures. According to Graae (1999), species propagated mainly by clonal growth will recolonise very slowly, while those dispersed by animals and the wind recolonise more quickly. Mechanical soil treatment instead of the use of herbicides presumably has similar effects on the forest-floor flora, but this has not yet been studied.

Impact on the flora of nature areas through atmospheric transport of pesticides

Herbicides are not normally used on nature areas, but many herbicides are transported over long distances and small amounts of them are thereby deposited on nature areas. No information has yet been found on measurements of direct impacts from pesticide drift on the flora in nature areas apart from boundary areas such as hedgerows. It is therefore necessary to use model calculations. In the Netherlands, the average combined herbicide consumption is 1.35 dose equivalents per year, 5.5% of which evaporates. On this basis and import/export considerations concerning atmospheric transport, it was calculated in a Dutch model study (Klepper et al. 1998) that the Dutch nature areas received an average of 0.02 dose equivalents per year. By means of a dose-response model for the potentially affected natural vegetation, this deposition was used to predict how large a percentage of the species were affected beyond NOEC (No Observed Effect Concentration). The result was that 2% (median value) of the species were affected beyond their NOEC. The percentage of affected species was highest in agricultural areas and in areas with fruit and berry growing. There are no similar calculations for Danish conditions, but the use of herbicides with a treatment frequency index of 1.65 in 1997 (DEPA 1998a) is comparable to the Dutch consumption, since the frequency index corresponds to the Dutch dose equivalent. It cannot be concluded directly from the Dutch calculations that around 2% of the species in Danish nature areas are unacceptably affected by herbicides. The level of impact depends, in part, on how the nature areas are situated in relation to areas in which herbicides are used and on the magnitude of the emission, the wind conditions and the sensitivity of the local plant community. In the Dutch calculations, the biggest uncertainties were reportedly due to the emission calculation and the effect models (Klepper et al. 1998). A diffuse dispersal of pesticides, e.g. via the atmosphere with precipitation, must be regarded as having less effect on the composition of the flora in nature areas than increased supply of nutrition and changed nature management.

The effect of pesticides in rainwater

The occurrence of pesticides in rainwater in Denmark is discussed in section 4.5. Herbicides have also been detected in rainwater in countries in Scandinavia and Northern Europe (Kirknel, Felding 1995ab), but no direct effects on flora as a consequence of this have yet been found (Felding 1998b). It is known from American studies that deposition of atmospheric herbicide residues (sulphonylurea) can produce symptoms of damage in some crops, such as peas and beans (Felsot et al. 1996). It has similarly been shown that even small doses of chlorsulfuron (1/100th to 1/1000th part of normal dose) greatly reduce the plant biomass and seed production in persecaria (Fletcher et al. 1996). In a Danish project on the occurrence of pesticides in precipitation and effects on plants and plant communities, the effects of mechlorprop have been investigated in concentrations corresponding to those found in rainwater. The study covers sensitive species – particularly the crucifers – but no effects have been found (Solveig Mathiassen, DIAS, personal communication). One can assess possible effects by combining data for the deposition of pesticides with effect data.

5.2.1 Conclusions

Together with crop rotation and other cultivation measures, the use of herbicides has considerably reduced the flora in farmland under cultivation. Accidental spray drift can have adverse effects on plants in hedgerows and small biotopes. In particular circumstances, herbicides in rainwater can presumably cause damage to plants outside the cultivated area.

The following specific conclusions can be drawn:

The number and frequency of plant species in an average field have halved in the last 20-25 years. From an agricultural point of view, this has been a desirable development, but it has adverse consequences for the natural species. The main reason for the decline is the use of herbicides, but changes in cultivation practice have also had a major effect.

Herbicide drift can affect the flora in hedgerows, small biotopes and nature areas.

In Denmark, pesticides have been detected in rainwater, but effects in plants have not yet been detected. However, American studies have shown that even small doses of some herbicides can harm particular plant species. The studies in question are limited in numbers and cover only a few pesticides.

There is a lack of systematic studies of how pesticides in large, connected areas affect wild plants and associated animals in hedgerows, ditches and other small biotopes, and neighbouring nature areas.

The impact on the flora caused by long-distance transport and deposition of herbicides in Denmark is not known. We know from a Dutch study that effects are probable, but studies of both the effects and the atmospheric transport are needed in order to determine them more exactly.

There is also a lack of specific studies of the effect of herbicide use on ground flora in forests, but there is no doubt that even the limited use of herbicides in forestry has a great and adverse effect on the original ground flora. Many woodland plant species have a very slow recolonisation rate (0-1 metre per year), which makes them very sensitive to herbicides, even if herbicides are used only in connection with clear cutting and establishment.

There have been no systematic studies of the combined effect of fertilisers, pesticides and crop rotation on the flora in areas near fields.

The sub-committee recommends more consistent and systematic use than hitherto of permanent unsprayed headlands and protection borders as buffer zones to protect original vegetation in small biotopes and remaining nature areas. In this connection, it must be ensured that linked dispersal corridors are established. Where the vegetation of small terrestrial biotopes has been seriously affected by the last few decades’ intensive use of both herbicides and fertilisers, recolonisation will normally take a very long time. Here, it will be necessary to establish both spray-free and fertiliser-free boundary zones where restoration of the vegetation and the associated fauna is desired. Sowing of wild plants is also recommended as a way of re-establishing original vegetation.

5.3 Impacts on flora and fauna in watercourses, lakes and coastal waters

Pesticides can be transported to watercourses via the atmosphere, through wind drift from field spraying or long-distance transport, via groundwater, drain water and surface run-off, and through unlawful spraying in or near watercourses within the 2-metre buffer zones. From the watercourses, the pesticides can be led to lakes or coastal waters. These can also receive pesticides by the same transport paths as watercourses. All transport of pesticides to fresh water in Denmark is undesirable. Data are available on the occurrence of pesticides in watercourses and ponds, as described in sections 4.2 and 4.3, but there have not yet been any systematic studies of the occurrence in Danish lakes and coastal waters. Studies in lakes have been initiated in connection with Aquatic Environment Plan II.

In this section we describe the impacts of pesticides in fresh water. Since there are only a few Danish studies, the description is based primarily on foreign studies, with the main emphasis on field studies (Friberg 1998). In order to assess the effects of current practice for use of pesticides, concentration levels given in Mogensen and Spliid (1997) are combined with results from the literature. Denmark has distance requirements for many pesticides, which must not be applied closer than 10 or 20 metres to lakes and watercourses. Table 5.3 shows the active ingredients to which such distance requirements currently apply. Some of these ingredients affect organisms in the aquatic environment at lower concentrations than 1 microgramme per litre and a few at lower concentrations than 0.1 microgramme per litre.

Table 5.3 Look here!
Pesticides subject to special distance requirements. The distance requirement is generally 2 m to watercourses and lakes. The table shows the pesticide, its active ingredients, the type (F = fungicide; H =herbicide; I = insecticide; P = growth regulator). Also shown is the lowest concentration for individual substances that cause 50% mortality or inhibition of activity in laboratory tests with aquatic organisms (fish, crustaceans and algae) (m g/L = microgramme per litre).

The effect of pesticides depends on whether the recipient is a watercourse or stagnant water. The effect of pesticides on aquatic organisms generally depends on the retention time and thus on the exposure time. In watercourses, the retention time depends on the average rate of flow and the presence of areas with a low rate of flow. In a watercourse to which a herbicide was added, a stretch with a high rate of flow directly downstream from the treated area was analysed and compared with a more slow-flowing stretch 225 m downstream (Thomson et al. 1995). Although the concentration was briefly higher in the upstream stretch in connection with the spraying, the retention time of the substance in a concentration of more than 1 microgramme per litre was twice as long in the stretch of water 225 m downstream. In watercourses, the concentration of pesticides also decreases at a given distance downstream of the source, depending on dilution and deposition/metabolism. In stagnant water, the effect of the pesticides depends particularly on the volume of water and the rate of water replacement.

Small and slow-flowing watercourses and small ponds are therefore most exposed to any pesticides. Moreover, these systems are closely linked to the surroundings and are therefore more likely to be affected by any use of pesticides in the surrounding areas.

Biologically, watercourses differ from stagnant waters by being open systems. There is permanent drift of organisms with the current, which means that stretches of watercourses are generally recolonised quickly. That means that watercourses are generally very resilient and therefore quickly return to normal once a chemical impact has ended.

Effects on primary producers

There are many documented instances of effects of herbicides on primary producers in freshwater ecosystems, whereas insecticides seldom seem to have any direct effect on plants. The effect of herbicides is usually that they reduce productivity, whereas, in the short term, the biomass is not affected. There is a clear tendency for the adverse effects to end very quickly when the exposure is over. Besides the direct effects of herbicides on primary producers in watercourses, indirect effects have been found in a number of studies. For instance, photosynthesis in an algae community increased at a concentration of more than 4 microgrammes per litre of the insecticide lindane owing to a reduction in the number of grazing invertebrates (Pearson, Crossland 1996).

With the concentrations found for the herbicide atrazine, it is unlikely that algae and macrophytes will be affected. Generally speaking, the concentration has to be much higher than 50 microgrammes per litre before the ecosystems are affected (Solomon et al. 1996). In the case of the herbicide hexazinone, the measured concentrations are also so low that no serious effect on the primary producers is expected. However, up to 43 microgrammes per litre have been found in a drain water sample, which could affect aquatic plants locally if there were little dilution in the recipient watercourse. EC₅₀ (4 hours) for hexazinone has been found to be 3.6 microgrammes per litre for aquatic plants in watercourses, and a concentration of 145-432 microgrammes per litre reduced the productivity of aquatic plants by 80% (Schneider et al. 1995).

Effects of insecticides on fauna

A large number of studies show that pesticides can affect the fauna in freshwater ecosystems, both directly and indirectly. Generally speaking, insecticides, in particular, have an adverse effect on the fauna, whereas direct effects from herbicides are often seen only at relatively high concentrations. Indirect effects are seen particularly in the highest trophic levels (fish, amphibians and birds) owing to reduced quantities of prey or in the form of changes in the function of the entire ecosystem. For example, Wallace et al. (1991) found that the decomposition of leaves into fine organic matter fell considerably in a forest watercourse after treatment with the insecticide methoxychlor owing to elimination of the insects that comminute the leaves. Methoxychlor may no longer be used in Denmark. If the possibility of recolonisation of a watercourse is poor, e.g. owing to barriers or the geographical location, it can take a long time (years) for the watercourse’s biological conditions to be reestablished, even if the supply of pesticide ceases. In the case of stagnant freshwater systems, the effects of pesticides seem to disappear within a similar period of time. These ecosystems are generally affected by both fertilisers and pesticides.

Effects of herbicides on fauna

In the case of the herbicide atrazine, the concentrations found in watercourses are generally so low that no effect on the fauna is likely. For example, in a study by Grande et al. (1994), the mortality among trout fry, which are more sensitive than adult fish, only increased at atrazine concentrations of more than 50 microgrammes per litre. However, Lampert et al. (1989) showed that concentrations right down to 0.1 microgramme per litre could affect daphnia in stagnant water, even though, in a single-species test, EC₅₀ was 2 microgrammes per litre. The study shows that the sensitivity at community level is far greater because both direct and indirect effects play a part. Atrazine concentrations of more than 0.1 microgramme per litre have been found in a pond near Køge, but there are very few other measurements. As in the case of atrazine, the concentrations of hexazine that have been found are so small that no serious effect on the fauna is likely.

The fungicide propiconazole has been found in concentrations up to 0.8 microgramme per litre in a watercourse and 0.1 microgramme per litre in a pond. These concentrations are below the concentration found to have an effect on zooplankton, invertebrates and fish.

The insecticide dimethoate has been found to affect (increase) activity in watercourse invertebrates at concentrations of more than 1 microgramme per litre and to reduce the density of some species (Bækken, Aanes 1994). In the case of zooplankton, an LC₅₀-value of around 20 microgramme per litre has been found in mesocosm experiments (Hessen et al. 1994). The concentrations in Danish watercourses and lakes lie below these values but not significantly below the concentrations that produce effects.

The pyrethroid insecticides are generally very toxic to largely all freshwater fauna, i.e. to zooplankton, invertebrates, crustaceans, fish and amphibians, in very small concentrations – normally less than 1 microgramme per litre. Besides that, the substances accumulate in the organisms (e.g. Andersen 1982), but are eliminated when the exposure ends. The pyrethroid fenvalerate has been found in ponds in a concentration of 0.12 microgramme per litre. Anderson (1982) observed behavioural changes and mortality in invertebrates – particularly amphipods – at fenvalerate concentrations from 0.022 microgramme per litre.

Effects on amphibians

There are only a few studies of the effect of pesticides on amphibians. In a Danish study, the insecticide esfenvalerate caused paralysis in embryos of the fire-bellied toad and the xenopus at a concentration of 1 microgramme per litre. At 5 microgrammes per litre, 80% of the embryos of the xenopus had deformed bodies, oedema and deformities of the spine, brain and intestine (Larsen, Sørensen in prep). Esfenvalerate was detected in a stream, Lillebæk, on Funen in the period 1994-1996 in a concentration of 0.2 microgramme per litre. (Funen County 1997). In that period, the recommended field dosage of esfenvalerate was 25 grammes per hectare. In a pond test in northern USA, esfenvalerate was found to have an adverse effect on daphnia and copepods at a concentration of 0.01 microgramme per litre (Lozano et al. 1992).

Unlawful use of pesticides

With our present level of knowledge, it is not possible to determine how large a part of the measured concentrations of pesticides in fresh water are due to unlawful use.

Watercourses must generally be deemed to be very sensitive to accidents with pesticides at washing sites and similar because, in many cases, the water is led directly to the nearest watercourse. Danish watercourses are also generally small, i.e. more than 80% of them are less than 2 metres wide, and they have low rates of flow. Therefore, even a short spillage of pesticides could have a major, acute effect on the ecological conditions in them.

Effects of pesticides in rainwater

On the basis of rainwater data given in Mogensen and Spliid (1997), the pesticide content must be assumed to be generally so low that it will have no effect on the biological conditions after further dilution in the fresh water system (irrespective of size). However, that does not exclude the possibility of drift from neighbouring fields resulting in local increases in the concentration of pesticides to a level that can be harmful to the freshwater environment. All the same, this problem is probably limited provided the distance requirements for spraying are observed.

Other environmental factors and pesticides

The effects of a given pesticide depend on the other ecological factors in the freshwater environment. For example, Caux and Kent (1995) found that a green alga Selenastrum capricornutum was affected differently by atrazine, depending on the chemical composition of the water. The role played by this has not been taken into account in this report. Furthermore, pesticides in watercourses often occur in pulses, especially with unlawful use. It is during such an episode that the concentration of pesticides must be measured for the effects to be assessed. This is seldom possible because the water containing the pesticide passes the sampling site within a very short space of time, so the chance of extracting a sample of the polluted water is extremely small. The measured concentrations are therefore probably often lower than the maximum that has occurred in the system. Another factor that probably has a significant influence on the effect of pesticides in watercourses is the physical conditions. Most Danish watercourses have been physically altered to ensure drainage of surrounding fields, whereby they have been made unnaturally wide and physically uniform. Both the degradation time and the effect on the ecological conditions presumably depend on the physical heterogeneity, so – all else being equal – physically uniform watercourses will be affected most by any supply of pesticides.

5.3.1 Conclusions

There are very few Danish studies of the occurrence and effects of pesticides in Danish watercourses and ponds and, as yet, no systematic studies in lakes and coastal waters. The assessment of the effects of pesticides has therefore been based primarily on foreign studies, and importance has been attached to field studies. It has not been possible, either, to find any studies of the effects of all the pesticides detected in Danish freshwater ecosystems. Since there are often differences in the species composition and the structure and function of the ecosystems in Denmark, compared with the foreign studies, the sensitivity to a given pesticide should be treated with caution. Owing to the relatively low water temperatures in Denmark, compared with more southerly countries, the degradation time for pesticides is shorter in Denmark, so the exposure time and thus the risk of effects are greater. With our present level of knowledge it is therefore difficult to judge how the present use of pesticides affects Denmark’s freshwater systems. However, several measurements indicate that pyrethroids and thiophosphate insecticides have been detected in concentrations at the level that has effects according to the existing literature. In particular, the available concentration levels indicate that it is insecticides – and especially the pyrethroids – that can have an adverse effect. In addition, by reason of their persistence, the pyrethroids can occur in freshwater ecosystems for a long period, during which they can be absorbed by, for example, invertebrates that live from dead organic material.

In connection with authorisation of pesticides, tests are carried out to determine their toxic effect on individual species of algae, daphnia, fish and other organisms. Mesocosm analyses that simulate entire ecosystems are also performed. However, these analyses do not provide insight into the many natural factors that interact in nature and they do not cover the combination of the many different pesticides detected in the aquatic environment.

It is therefore likely that the freshwater environment is affected by the present use of pesticides, but it is not yet possible to determine the magnitude of the impact owing to lack of data and knowledge. The county authorities, in their regulatory capacity, have provisionally estimated that around 2% of the unfulfilled targets in approx. 11,000 km of watercourses are due to toxic substances, including pesticides (Windolf 1997). However, this figure is based on a subjective evaluation and will also vary with the region, the sampling method and the frequency.

The following specific conclusions can be drawn:

Some pesticides – particularly pyrethroid insecticides and thiophosphates – are toxic to aquatic organisms in lower concentrations than the limit value of 0.1 microgramme per litre for drinking water.

	Pesticide concentrations in watercourses are transported with the current. The highest concentrations are probably rarely detected because the water containing the pesticides passes the sampling site within such a short space of time that there is very little chance of extracting a sample of the polluted water.
	Small, slow-flowing watercourses and small ponds are most exposed to any pesticide load.

The impact on the flora and fauna in watercourses is brief, but even a short pulse of pesticides can have a major, acute effect on the watercourses’ ecological conditions. If the possibilities of recolonisation are poor, e.g. due to barriers or geographical location, the effects can in some cases be observed for more than one year.

The pesticide concentration in ponds and lakes will be longer-lasting than the concentration in watercourses. It will not be possible to evaluate the occurrence and effect of pesticide concentrations in lakes until data are available from the monitoring programme in connection with Aquatic Environment Plan II.

In watercourses, insecticides have been detected in concentrations that very probably affect the fauna, and there are documented cases of effects of herbicides on algae and other primary producers.

There is no national registration of the effects on the aquatic flora and fauna.

The sub-committee recommends that more consistent and systematic use than hitherto be made of permanent spray-free zones and protection borders, which, as buffer zones, would help to protect watercourses, lakes and ponds.

5.4 The sub-committee’s conclusions and recommendations concerning environmental impacts

The main impacts occur in connection with the application of pesticides, with organisms being hit directly, and with indirect effects occurring as a consequence of the effect on food chains. Here, plants play a key role as the first link in the food chains. A Danish study has shown that the number and frequency of plant species in field studies have halved in the last 20-25 years. From an agricultural point of view, this has been a desirable development, but it has adverse consequences for the natural species. The main reason for the decline is the use of herbicides, but changes in cultivation practice, including the use of fertilisers, have also had a major effect.

During spraying, there is spray drift to the surrounding areas. However, hedgerows, dikes, dry stone walls and other small biotopes are of such small width that they should in practice be included in the area that is affected by spray products. Spray drift can affect both terrestrial and aquatic ecosystems. In the case of the aquatic environment, any effect from pesticides is undesirable, including changes of flora and fauna in coastal waters, lakes, ponds and watercourses. Among the aquatic ecosystems, it is particularly ponds, watercourses and lakes near fields that could potentially be affected.

The freshwater environment is in all probability affected by the present use of pesticides, but it is not possible on the basis of the existing data to quantify the magnitude of the impact nationally. On the basis of information from the county authorities, it is provisionally estimated that around 2% of the unfulfilled targets of about 11,000 km watercourses can be due to toxic substances, including pesticides. In particular, the concentration levels measured indicate that it is insecticides – and especially pyrethroids – that can have an adverse effect. Owing to their persistence, pyrethroids can also occur in the freshwater ecosystems for a long period of time. There are also documented cases of effects of herbicides on algae and other primary producers. Several measurements indicate concentrations of pyrethroids and certain thiophosphate insecticides that are close to the level that has effects according to the existing literature. For some pesticides, this level is lower than the limit value of 0.1 microgramme per litre for drinking water.

In both cultivated areas and the adjacent biotopes, the use of pesticides involves a risk of reduced populations of flora and fauna, changed biodiversity, changes in the cultivation medium and natural pest regulation, and of food-chain effects and other indirect effects. Seen overall, it is not the individual field and its possible loss of wild flora that are the problem, but rather the total, countrywide effects on characteristic farmland flora and the associated fauna.

In forestry, very little use is made of pesticides, whereas, in Christmas tree and ornamental greenery cultures, the same quantities are used as in farming. The treatment frequency index in nurseries and market gardens is high. There is a lack of specific studies of the effect of herbicides on forest-floor plant species, but there is no doubt that even the limited use made of pesticides in forestry has adverse effects. Many woodland plant species have a very slow recolonisation rate of less than 1 metre per year, which makes them particularly sensitive to herbicides, even though these are only applied in connection with clear cutting and afforestation.

The sub-committee’s recommendations concerning the impacts of pesticides in the environment

For the scenarios in which pesticides are used, there are no systematic studies of how pesticides in large, connected areas affect wild flora and the associated fauna in hedgerows, waysides and other small biotopes and in neighbouring nature areas. The effects on the flora caused by precipitation of long-distance transported herbicides in Denmark is not known. Foreign studies show that effects probably exist, but further investigations of both the atmospheric transport and the effects are needed for any closer determination. There is also a need to assess the effect of pesticides on aquatic organisms in relation to the actual finds in watercourses and surface water. The sub-committee recommends that the necessary knowledge be built up and that time series be established to document any effects on the terrestrial and aquatic ecosystems seen in relation to the proven occurrence of pesticides in the environment.

The sub-committee recommends more consistent and systematic use than hitherto of permanent unsprayed headlands and protection borders as buffer zones to help protect watercourses, lakes and ponds, together with well-preserved vegetation in small biotopes and nature areas, where such still exist. In this connection, it must be ensured that linked dispersal corridors are established. Where the vegetation of small terrestrial biotopes has been seriously affected by the last few decades’ intensive use of both herbicides and fertilisers, recolonisation will normally take a very long time. Here, it will be necessary to establish both spray-free and fertiliser-free boundary zones where restoration of the vegetation and the associated fauna is desired. The sub-committee also recommends nature restoration by sowing of wild plants and the introduction of species of fauna with a low dispersal potential.

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