Key Concepts

After completing this chapter, you will be able to:

1. Explain the notions of “pest” and “weed,” and provide reasons why it may be necessary to decrease their abundance.
2. Differentiate pesticides by the pests to which they are targeted.
3. Classify pesticides into major chemical groups.
4. Outline the risks and benefits of pesticide use in sanitation, agriculture, forestry, and horticulture.
5. Explain why there is a global contamination of organisms with DDT and related organochlorines, and describe the associated ecological damage.
6. Outline the ecological damage caused by carbofuran, and explain why it took so long for the use of this insecticide to be banned.
7. Describe the economic benefits and ecological risks of pesticide use in forestry.
8. Outline the concept of integrated pest management, and explain whether it is applicable to all pest-management problems.

Introduction

Humans are constantly engaged in struggles against competitors and diseases. One way to gain an advantage in many of those ecological interactions is through the use of pesticides. These substances are used to protect crop plants, livestock, domestic animals, and people from damage and diseases caused by microorganisms, fungi, insects, rodents, and other “pests,” and to defend crops from competition with unwanted but abundant “weeds.”

It is important to understand that the use of words like “pest” and “weed” is highly contextual. In most situations, for example, white-tailed deer are valued for their wild beauty, and they provide economic and subsistence benefits through hunting. However, this animal may also be considered a pest when it feeds in a garden, agricultural field, or forestry plantation. The same is true, to some degree, of other species that are considered to be a pest or weed.

People have been using pesticides for a long time (Hayes, 1991). There are records of unspecified chemicals being used by Egyptians to drive fleas from their home as long as 3,500 years ago. Arsenic has been used as an insecticide in China for at least 2,900 years. In his epic poem the Odyssey, the Greek poet Homer (writing about 2,800 years ago) referred to the burning of sulfur (which generates toxic SO₂ gas) to purge homes of vermin such as fleas.

However, pesticide use has become much more common in modern times, and an enormously wider variety of substances is being used. At least 300 insecticides, 290 herbicides, 165 fungicides, and many other pesticidal chemicals are available in more than 3,000 different formulations. Strictly speaking, a pesticide is a product that consists of a formulation of several chemicals – the “active ingredient” attacks the pest, while various “inert ingredients” enhance its effectiveness (see In Detail 26.1). Even larger numbers of “commercial products” are available, because many involve similar formulations manufactured by different companies.

Almost all pesticides are chemicals. The active ingredients of some of them are based on natural biochemicals that are extracted from plants grown for that purpose, while others are inorganic chemicals based on toxic metals or compounds of arsenic. Most modern pesticides, however, are organic chemicals that have been synthesized by chemists. The costs of developing a new pesticide and testing it for its efficacy (effectiveness against pests), toxicological properties, and environmental effects are quite large, equivalent to tens of millions of dollars per chemical. However, if an effective pesticide against an important pest is discovered, the profits are potentially huge, and therefore industry willingly pays the high development costs.

People have acquired important benefits from many uses of pesticides:

• Increased yields of crops, because of protection from diseases, competition, defoliation, and parasites
• Prevention of much spoilage and destruction of stored food
• Avoidance of certain diseases, thereby conserving health and saving the lives of millions of people and domestic animals

This is not to say, however, that more pesticide use would achieve even better results. In fact, it has been argued that pesticide use in North America could be decreased by half without greatly affecting crop yields (Pimentel et al., 1991). The European Union (EU) has taken some forceful steps to reduce pesticide use within its jurisdiction (Pesticide News, 2003). In 1992, the EU embarked on a full review of the approximately 1,000 different pesticide active ingredients used in the region for agricultural purposes at that time (PAN Europe, 2008). By 2011, there were only 353 approved active substances included in the EU Pesticides Database (Official Journal of the European Union, 2011). In large part, the withdrawals involve obsolete pesticides and others of minor commercial importance, for which the owners do not want to invest the large amounts of money needed to assure EU regulators that their products are safe for people and the environment. Similar actions are also occurring in North America, but they are less advanced than in the EU. Because of the substantial benefits that can be derived from the use of pesticides, their consumption has increased enormously during the past half-century. Overall, the use of pesticides increased ten-fold in North America between 1945 and 1989 (Pimentel et al., 1992), although it has since leveled off. Pesticide use is now a firmly integrated component of technological systems that are widely used in modern agriculture, forestry, horticulture, and public health management in most parts of the world.

Unfortunately, the considerable benefits of pesticides are partially offset by damage their use causes to ecosystems and sometimes to human health. Each year, about one-million people are poisoned by pesticides, with as many as 20-thousand fatalities (Pimentel et al., 1992). Although developing countries account for only about 20% of global pesticide use, they sustain about half of the poisonings. This is because relatively toxic insecticides are used in many developing countries, by a workforce whose widespread illiteracy hinders the understanding of instructions for proper use, and whose safety is further compromised by poor enforcement of regulations and by inadequate use of protective equipment and clothing.

The most tragic case of pesticide-related poisoning occurred in 1984 at Bhopal, India. About 2,800 people were killed and 20 thousand seriously poisoned when a factory accidentally released 40 tonnes of methyl isocyanate vapor into the atmosphere. Methyl isocyanate is a precursor chemical of carbamate insecticides (Rozencranz, 1988).

In addition, many pesticide applications cause ecological damage by killing non-target organisms (organisms that are not considered to be a pest). This damage is particularly important for broad-spectrum pesticides, which are toxic to organisms in addition to the specific pest. Pesticides applied as a broadcast spray are spread over a large area, such as an agricultural field, lawn, or stand of forest.

If a broad-spectrum pesticide is broadcast-sprayed, many non-target organisms are exposed and they may be damaged or killed. For example, in a typical agricultural field or forestry plantation, only a few plant species are abundant enough to compete significantly with crops and reduce their productivity. These are the “weeds” that are the target of a broadcast herbicide application, but many other plants are also affected. The non-target plants may provide habitat or food for animals, and they help to prevent erosion and loss of nutrients. These benefits are degraded by non-target damage – by damage to organisms that are not pests. Similarly, broadcast insecticide spraying causes non-target mortality to many beneficial arthropods in addition to the species that is considered to be a pest. Many birds, mammals, and other creatures may also be poisoned. The non-target mortality may include predators and competitors of the pest, an ecological change that may release it from some of its biological controls. Clearly, the great challenge of pest control is to develop effective, pest-specific pesticides and to invent non-pesticidal methods.

Pesticide use has been expanding rapidly, and this is happening in all countries, although to varying degrees. Although much is known about the environmental damage caused by the use of pesticides, not all of the potential effects are well understood. In this chapter we examine the nature of pesticides and their important uses. We then examine cases of ecological damage caused by their routine use to deal with pest-management problems.

The Nature of Pesticides

Classification by Target

Pesticides are defined by their usefulness in killing or otherwise decreasing the abundance of species that are deemed to be “pests.” Pesticides are, however, an extremely diverse group of substances. To better understand their usefulness and toxicity, and the damage they cause, it is helpful to categorize them in various ways. One classification is based on the intended target of the use:

• A fungicide is used against fungi that cause diseases and other damage to crop plants and animals.
• An herbicide is used to kill weeds, which are unwanted plants that interfere with some human purpose; most use in agriculture and forestry is intended to release crop plants from competition, while horticultural use is mostly for aesthetics (Image 26.1).
• An insecticide is used to kill insects that are pests in agriculture, horticulture, or forestry, or that spread diseases such as mosquito vectors (a path by which a disease is spread) of malaria, yellow fever, and encephalitis.
• An acaricide is used to kill mites that are pests in agriculture, and ticks that are vectors of ailments such as Lyme disease and typhus.
• A molluscicide is used to kill snails and slugs in agriculture and gardens, and aquatic snails that are vectors of diseases such as schistosomiasis.
• A nematicide is used against nematodes, which can damage the roots of agricultural plants.
• A rodenticide is used to control mice, rats, gophers, and other rodents that are pests in agriculture or around the home.
• An avicide is used to kill birds, which are sometimes considered pests in agriculture.
• A piscicide is used to kill fish, which may be pests in aquaculture.
• An algicide is used to kill unwanted growths of algae, for example, in swimming pools.
• Bactericides, disinfectants, and antibiotics are used to control infections and diseases caused by bacteria.

Image 26.1. Glyphosate (commonly known by the brand name “Roundup”) is one of the most widely used herbicides in the U.S. Source: “RoundUp Monsanto” by JeepersMedia is licensed under CC BY 2.0.

Chemical Classification

Because almost all pesticides are chemicals, they can be categorized according to similarities in chemical structure. The most important groups are described below and in Table 26.1. A few “non-chemical” pesticides are based on microbes, and are discussed later under “Biological Pesticides.”

Table 26.1. Some Important Pesticides.

Inorganic pesticides are compounds that contain toxic elements such as arsenic, copper, lead, or mercury. They are highly persistent in terrestrial environments, being only slowly dispersed by leaching and erosion by wind or water. Recently, inorganic pesticides have been widely replaced by synthetic organics. Prominent examples include Bordeaux mixture, a complex of copper-based compounds that is used as a fungicide to protect fruit and vegetable crops, and arsenicals such as arsenic trioxide, sodium arsenite, and calcium arsenate, which are used as herbicides and soil sterilants. Paris green, lead arsenate, and calcium arsenate are used as insecticides.

Organic pesticides are mostly synthesized chemicals, but some are natural toxins produced by certain plants that are extracted and used as pesticides. Important examples include the following:

• Natural organic pesticides are extracted from plants. For example, nicotine and related alkaloids are extracted from tobacco (Nicotiana tabacum) and used as insecticides, usually applied as nicotine sulphate. Neonicotinoids are a synthetic analogue. Pyrethrum is a complex of chemicals extracted from certain chrysanthemums (Chrysanthemum cinerariaefolium and C. coccinium) and used as an insecticide. Rotenone is extracted from several tropical shrubs (Derris elliptica and Lonchocarpus utilis) and used as an insecticide, rodenticide, or piscicide. Red squill, extracted from the sea onion (Scilla maritima), is a rodenticide, as is strychnine, extracted from the tropical shrub Strychnos nux-vomica.
• Synthetic organometallic pesticides are used as fungicides and include organomercurials such as methylmercury and phenylmercuric acetate.
• Phenols include trichlorophenols, tetrachlorophenol, and pentachlorophenol, which are fungicides used mostly to preserve wood.
• Chlorinated hydrocarbons (organochlorines) are a diverse group of synthetic pesticides (see In Detail 26.2). Most are quite persistent, having a half-life of about 10 years in soil because they are not easily degraded by microorganisms or by physical agents such as sunlight or heat. The persistence of organochlorines, coupled with their strongly lipophilic nature (they are highly soluble in fats and lipids, but virtually insoluble in water), means that they strongly bioconcentrate and biomagnify, with the highest concentrations occurring in top predators (see In Detail 17.1 and Figure 26.1). Organochlorines include the following:
• DDT and its insecticidal relatives, such as DDD and methoxychlor, were once widely used insecticides. Because of bans in North America and Europe in the early 1970s, their use is now confined to tropical countries. DDE is a persistent non-insecticidal metabolite of DDT and DDD that accumulates in organisms.
• Lindane is the active constituent of hexachlorocyclohexane, an insecticide.
• Cyclodienes are highly chlorinated cyclic hydrocarbons, such as chlordane, heptachlor, aldrin, and dieldrin, all of which are insecticides.
• Chlorophenoxy acids have growth-regulating influences on plants and are used as herbicides against broad-leafed weeds. The most important compound is 2,4-D, but others are 2,4,5-T, MCPA, and silvex.
• Other organochlorines include polychlorinated biphenyls (PCBs), dioxins, and furans. These are not pesticides but are mentioned here because their ecotoxicological properties are similar to those of the pesticide organochlorines: they are persistent in the environment and are lipophilic, so they bioconcentrate and food-web magnify.
• Organophosphorus pesticides are used mostly as insecticides, acaricides, and nematicides. They are not persistent in the environment, but are extremely toxic to arthropods and also to non-target fish, birds, and mammals. Parathion, fenitrothion, malathion, and phosphamidon are prominent examples of organophosphate insecticides. Glyphosate, a phosphonoalkyl compound, is an important herbicide (it is not toxic to animals).
• Carbamate pesticides have a moderate persistence in the environment but are highly toxic to arthropods, and in some cases, to vertebrates. Aminocarb, carbaryl, and carbofuran are important insecticides.
• Triazine pesticides are used as herbicides and sometimes as soil sterilants. Prominent examples are atrazine, simazine, and hexazinone.
• Synthetic pyrethroids are analogues of natural pyrethrum and are used mostly as insecticides and acaricides. Pyrethroids are highly toxic to invertebrates and fish, but they are of variable toxicity to mammals and of low toxicity to birds. Important examples are cypermethrin, deltamethrin, permethrin, synthetic pyrethrum and pyrethrins, and tetramethrin.
• Biological Pesticides are formulations of microbes that are pathogenic to specific pests and so have a narrow spectrum of toxicity in ecosystems. The best examples are insecticides based on the bacterium Bacillus thuringiensis (B.t.), types of which are used against moths, flies, and beetles. Insecticides based on nuclear polyhedrosis virus (NPV) and insect hormones have also been developed.
• Genetically modified organisms (GMOs; see Environmental Issues 7.1) have been biologically “engineered” by inserting portions of DNA from another species into their genome. This high-tech procedure has been used to develop new varieties of commercial crops that are more resistant to certain pesticides or pests, which can make it easier to cultivate them. For example, GMO varieties of soybean and canola have been developed to be resistant to glyphosate, meaning this herbicide can be used on those crops, providing reduced costs of energy and machinery to control weeds. In addition, there are GMO varieties of maize (corn) that contain DNA of Bacillus thuringiensis, which provides resistance to some insect pests and allows farmers to use less insecticide. These and other GMO crops are widely cultivated in North America, although they are banned in many countries, including most of Europe and Russia. The use of these GMO crops is controversial because there is incomplete knowledge about the biological and ecological risks of their use, including the potential escape of their GMO factors to wild plants.

Figure 26.1. Residues of PCBs in the Food Web of Lake Ontario. Organochlorine insecticides, such as DDT, DDD, and dieldrin, show a similar pattern of bioaccumulation and biomagnification as PCBs, but their residue levels are different. Source: Data from Environmental Protection Agency (2003).

Pesticide Use

The global use of pesticides was about 4.1 million tons in 2018, a total that includes insecticides, herbicides, fungicides, preservatives, and disinfectants (FAO, 2018). Global pesticide trade in 2015 had a value of about US $37 billion (OEC, 2018). About 47.5% (by weight) of pesticides used were herbicides; insecticides accounted for 29.5%, fungicides for 17.5%, and “other chemicals” for 5.5% (mostly used as soil fumigants). In the U.S., for the total expenditures on pesticides 59% were on herbicides, 14% on insecticides and 10% fungicides (Atwood and Paisley-Jones, 2017). Total expenditures for pesticides in the United States were about US $9 billion in 2012 (Hellerstein et al., 2019). The following is a top-20 list of conventional pesticides used in the United States’ agricultural market sector in 2012 (ranked by usage rate range in millions of pounds of active ingredient; Atwood and Paisley-Jones, 2017):

1. Glyphosate; herbicide; 270-290
2. Atrazine; herbicide; 64-74
3. Metolachlor-S; herbicide; 34-44
4. Dichloropropene; fumigant; 32-42
5. 2,4-D; herbicide; 30-40
6. Metam; fumigant; 30-40
7. Acetochlor; herbicide, 28-38
8. Metam Potassium; fumigant; 16-26
9. Chloropicrin; fumigant, 8-18
10. Chlorothalonil; fungicide; 6-16
11. Pendimethalin; herbicide; 6-16
12. Ethephon; plant growth regulator; 7-11
13. Mancozeb; fungicide; 5-9
14. Chlorpyrifos; insecticide; 4-8
15. Metolachlor; herbicide; 4-8
16. Hydrated Lime; fungicide; 3-7
17. Propanil; herbicide; 3-7
18. Dicamba; herbicide; 3-7
19. Trifluralin; herbicide; 3-7
20. Decan-1-ol; plant growth regulator; 3-7

The most important uses of pesticides are in agriculture and forestry, around the home, and in human health and sanitation programs. We examine these uses in the following sections.

Pesticide Use for Human Health

Various insects and ticks are vectors that transmit pathogens among individuals of the same species, or from an alternate host to people, or to domestic and wild animals. Important human diseases that are vectored by invertebrates include the following:

• Malaria, which is caused by the protozoan Plasmodium and spread to people by Anopheles mosquitoes
• Yellow fever, encephalitis, and West Nile virus, caused by viruses and spread by mosquitoes
• Sleeping sickness, caused by the protozoan Trypanosoma and spread by the tsetse fly Glossina
• Plague or black death, caused by the bacterium Yersinia pestis and transmitted by the rat flea Xenopsylla cheops
• Typhoid, caused by the bacterium Rickettsia prowazeki and transmitted by the louse Pediculus humanus
• Schistosomiasis or bilharziasis, caused by the blood fluke Schistosoma, with freshwater snails as the alternate host

To varying degrees, the incidence of these maladies can be controlled by using pesticides against the invertebrate vectors or the alternate hosts. The abundance of mosquitoes, for example, can be reduced by spraying insecticide in their aquatic breeding habitat or by applying a persistent insecticide to the interior walls of buildings, where they rest. Similarly, people infested with body lice may receive a surface dusting with an insecticide – this was an early use of DDT. Plague can be controlled by using rodenticide along with sanitation programs to reduce rat populations. Over the past half-century, pesticides have decreased the abundance of vectors and alternate hosts and have spared hundreds of millions of humans from the debilitating or deadly effects of certain diseases, particularly in tropical countries. (This has been an important factor in reducing death rates and allowed rapid population growth.)

In fact, one of the first important uses of DDT was in Naples, Italy, during the Second World War, to prevent a deadly plague of typhus that could have decimated Allied troops and the civilian population. Because of the success of this use of DDT and its contribution to the victorious war effort, the British prime minister at the time, Winston Churchill, referred to the insecticide as “that miraculous DDT powder.”

Malaria has long been an important disease in tropical countries. During the 1950s, about 5% of the global population was infected with malaria. The use of insecticide to reduce the abundance of mosquitoes achieved huge reductions in the incidence of malaria in some countries. For instance, during 1933-1935, India recorded about 100-million cases of malaria per year and 750-thousand deaths. However, the incidence was reduced to 150-thousand cases and 1,500 deaths in 1966 because of spraying with DDT and the draining of mosquito-breeding wetlands (McEwen and Stephenson, 1979). Similarly, 2.9-million cases of malaria occurred in Sri Lanka in 1934, and 2.8 million in 1946, but DDT use helped to reduce that incidence to only 17 cases in 1963 (Hayes, 1991). However, malaria has recently been resurging in some tropical countries, partly because mosquitoes have developed a genetically based tolerance of previously effective insecticides. Many people are again being exposed to the malarial parasite, although the disease can today be controlled by drugs that prevent Plasmodium from multiplying in the blood. (However, there are also signs that Plasmodium is becoming resistant to those drugs.)

Pesticides and Agriculture

Modern agriculture is a highly technological activity. Machines, energy, fertilizer, pesticides, and high-yield crop varieties are used in intensive management systems to grow crops (see Chapters 12 and 13). The role of pesticides is to help control the abundance of the following problems:

• Weeds that compete with crop plants
• Invertebrates and rodents that feed on crops or stored produce
• Microbial diseases that can kill the crop or diminish its yield

Undeniably, these uses of pesticides are important factors in modern agriculture. Even with pesticide use, the damage caused by pests and diseases around the world are equivalent to about 26-29% of the potential crop of wheat, 37% of rice, 31% of corn (maize), 40% of potatoes, and 26-29% of soybean (Oerke, 2006). Globally, insect pests are known to reduce agricultural yields by up to 32% overall (10–16% before harvest, and another 10-16% following harvest; Bradshaw et al., 2016).

Of course, management practices in agriculture have intensified greatly, particularly during the twentieth century and since. This change has resulted in increases in crop productivity. The gains in agricultural yield have been largely achieved by the combined influences of the following:

• Fossil-fueled mechanization
• Fertilizer use
• Improved crop varieties grown in monocultural systems
• The use of pesticides

The recent intensification of agrotechnology is sometimes referred to as the “green revolution.” Although agricultural yields have increased greatly, it must be recognized that the gains are highly subsidized by the following (see also Chapters 12 and 13):

• Intense use (and depletion) of non-renewable fossil fuels and metals
• Depletion of potentially renewable resources, such as soil fertility and tilth, and groundwater and surface water needed for irrigation
• Loss of soil mass through erosion
• Extensive salinization of soils in semi-arid regions (caused by inappropriate methods of irrigation)
• Ecological damage associated with the conversion of natural ecosystems into agricultural ones
• Ecotoxicological damage caused by the use of pesticides

In the U.S., for example, the yield of corn was typically about 1.4 t/ha-yr in 1933, but it increased to 4.2 t/ha-yr in 1963 and to 11.6 t/ha-yr in 2020 (USDA, 2021). In Mexico, wheat yields increased from 0.75 t/ ha-yr in 1945 to 2.6 t/ ha-yr in 1964. The yield of rice in Japan increased from a pre-war average of 1.8 t/ ha-yr to 4.0 t/ ha-yr in 1963, while in the U.S., rice yields have now reached 8.5 t/ ha-yr (Hayes, 1991; USDA, 2021).

Almost all intensively managed agricultural systems depend to some degree on pest control. High-yield crop varieties are often vulnerable to infestation by insect pests, diseases, and competition from weeds. Pesticides are routinely prescribed to manage those problems. Moreover, monocultural systems (in which only a single crop is grown in a field) result in reduced populations of natural predators and parasites, which can exacerbate existing pests and allow new ones to develop. Some environmentalists have described intensively managed agricultural systems as being a pesticide treadmill, because they rely on pesticides, often in increasing quantities, to deal with unanticipated pests that emerge as “surprises.”

The use of pesticides in U.S. agriculture has increased greatly in recent decades. An estimated 125 million kg of glyphosate, the most commonly used herbicide in the U.S., was applied in 2014 (Benbrook, 2016). Overall, pesticide use in North American agriculture increased by about ten-fold between 1945 and 1989 (Pimentel et al., 1992). Interestingly, during that same period, crop losses (to insects only) actually increased somewhat, from about 7% during 1941-1951 to 13% during 1951-1974 (Hayes, 1991). These trends, which might seem to contradict each other, may be due to such factors as the development of tolerance by some pests to pesticides, the emergence of new pests because of accidental introductions, changes in predator-prey relationships caused by pesticide use, and the introduction of new crop varieties that are vulnerable to pests.

Agricultural damage caused by arthropod pests varies greatly. Sometimes there is direct competition with humans for a food resource, as when insects defoliate crops in fields or attack stored food. Such depredations can sometimes obliterate agricultural yield, as can happen during a severe infestation of locusts. More commonly, however, insects reduce the yield only somewhat.

In some cases, however, even minor damage by pests can make the produce unsaleable. This can be the case of damage caused to apples by the codling moth (Carpocapsa pomonella) (Image 26.2). “Wormy” apples with larvae of this insect are not saleable to consumers (Image 26.3), and up to 90% of the fruit in unsprayed orchards may be infested (McEwen and Stephenson, 1979). Even a minor discoloration of fruit, such as apple scab and russeting of oranges (neither affect the productivity or nutritional quality of the crop) are considered unacceptable by many consumers. Therefore, seemingly unimportant crop damage can be a critical economic consideration for farmers and the food industry. As with some potentially life-saving drugs, pesticides are over-prescribed for some uses.

Image 26.2. Codling Moth. The caterpillars will bore into fruit, leading to premature ripening. Source: “1261-DSCN2109 Codling Moth (Cydia pomonella)” byajmatthehiddenhouse is licensed under CC BY-NC 2.0.

Image 26.3. Damage from Codling Moth. Source: Photo by Doug Wilson, USDA Agricultural Research Service.

Much pesticide use in agriculture is targeted against weeds, which interfere with crop plants by competing for limited resources of light, water, and nutrients. (Of course, “weediness” is partly a matter of context – in other situations, some weed species have positive attributes.) It is well known that weeds, if abundant, can cause large decreases in the productivity of agricultural crops, even by 50-90%. This is the reason why farmers have always taken measures to reduce the abundance of weeds, initially by hand-pulling or hoeing them and later by mechanical cultivation (ploughing) to disrupt their growth. More recently, chemical herbicides have been widely used to control agricultural weeds. In the U.S., for example, herbicide is used on 90% of the planted area for most crops (Gianessi and Reigner, 2007). Crops receiving the most herbicide are canola (on 99% its cultivated area), dry beans (99%), carrot (98%), maize (98%), rice (98%), sugarbeet (98%), peanut (97%), green bean (96%), soybean (96%), tomato (96%), blueberry (95%), citrus (95%), cotton (95%), and potato (93%). Herbicide use in wheat is relatively low (55%) because mechanical cultivation during sowing is effective in reducing weeds for this crop.

According to Gianessi and Sankula (2003), if herbicide use were discontinued for the 40 crops they studied, then weed management would have to rely on increased mechanical cultivation and manual weeding. They calculated that would have an annual cost of US$14 billion, about double that of herbicide use ($6.6 billion per year). If herbicide use were discontinued and replaced by alternative practices, they estimated that 35 of the 40 crops studied would suffer a decline in productivity, by an average of 21% (range of 5-67%). The productivity loss would have a value of about $21 billion per year (including $13 billion in lost productivity and $8 billion in increased costs of management). Note, however, that the “direct costs” of herbicide use cited above ($6.6 billion/yr) do not include the value of environmental damage that might be caused, for instance, by toxicity to non-target plants, wildlife, or agricultural workers. It must be remembered that non-herbicidal methods of weed control also cause environmental damage, particularly mechanical cultivation, which increases soil erosion and compaction. No-till management systems, which greatly reduce the rate of erosion, must rely on the use of herbicide to control weeds.

Of course, the weeds must be vulnerable to the toxicity of the herbicide being used against them, while the crop plant must be tolerant. Some herbicides are toxic to broad-leaved weeds (dicotyledonous plants) but not to corn, wheat, barley, or other crops in the grass family (monocotyledonous plants). Consequently, herbicides are widely used in grain agriculture in North America. For instance, about 98% of the maize and rice acreage is treated. Some important diseases of agricultural plants can be managed with pesticide. Sometimes, insecticide is sprayed to control arthropod vectors of microbial diseases. More commonly, fungicide is used to control pathogenic fungi such as late blight (Phytophthora infestans) of potato, apple scab (Venturia inequalis), powdery mildew (Sphaerotheca pannosa) of peach, and seedrot and damping-off of many crop species (Pythium spp.). Fungicide also helps to prevent the spoilage of stored crops by fungi such as Aspergillus flavus, which can grow in stored legumes, grains, and nuts, producing deadly aflatoxins that make foods poisonous to humans and livestock.

Pesticides in Forestry

The use of pesticides in forestry raises a great deal of controversy, often more so than use of the same chemicals in agriculture. The controversy partly concerns damage that may be caused to the many native species that are exposed to forestry sprays, compared to mostly non-natives in agriculture. In addition, spraying in forestry is mostly done by government agencies and large companies, while in agriculture it often involves individual farmers working a family farm. Most people have greater empathy for individuals than for big government or big business, and this can influence their opinions about pesticide use.

Pesticides are used in forestry mainly to control epidemics of defoliating insects and to manage weeds in re-forested areas and plantations. The largest insecticide spraying campaigns have been undertaken against spruce budworm (Choristoneura fumiferana) in New Brunswick, Canada, where a cumulative area of about 49-million hectares was sprayed between 1952 and 1992 (this is examined later as a case study). Other large spray programs have included those against gypsy moth (Lymantria dispar, an introduced pest that defoliates many tree species), hemlock looper (Lambdina fiscellaria), and bark beetles (especially species of Ambrosia and Ips).

Pesticides in the Home and Horticulture

Pesticides are commonly used in and around homes. For instance, insecticide may be used to kill bedbugs and cockroaches, and rodenticide to poison rats and mice. As well, large amounts of pesticide are used in horticulture. Herbicides are especially widely applied, mostly to achieve the grass-lawn aesthetic that many homeowners seek. To this end, herbicide is used to kill broad-leaved plants such as dandelion and plantain. The “weed” in common “weed-and-feed” lawn preparations is the herbicide 2,4-D, dicamba, or mecoprop.

Some of the most intensive pesticide usage occurs in the management of golf courses, particularly on putting greens where the lawn quality must be very consistent (Image 26.4). Fungicide is used in especially large amounts to prevent turf-grass diseases. On a per-unit area basis, the use of pesticides on putting greens can be more intensive than almost any in agriculture.

Image 26.4. Las Vegas Paiute Golf Resort, Las Vegas, Nevada. Golf courses require an immense amount of resources to maintain their overall aesthetic including a large amount of chemicals and water. Notice the dichotomy between the green golf course and the surrounding desert environment. Source: “Las Vegas Paiute Golf Resort, Wolf Golf Course, Las Vegas, Nevada” by danperry.com is licensed under CC BY 2.0.

Environmental Effects

Pesticide applications are intended to manage the impacts of pests by reducing their abundance and damage to below an economically or aesthetically acceptable threshold. This objective can sometimes be achieved selectively, thereby avoiding non-target damage. For example, rodenticide can be used judiciously to kill rats and mice around the home, while minimizing toxic exposures to non-target cats, dogs, and children (although risk is never eliminated).

More typically, however, pesticide use involves a less-selective broadcast application, usually by spraying. A crop-dusting aircraft or tractor-drawn sprayer is often used, which results in many non-target species being exposed to the spray (Image 26.5). The non-target organisms may live on the sprayed site, or they may be off-site and suffer exposure from aerial or aquatic drift of a pesticide. Non-target exposures include both direct contact with a sprayed pesticide as well as indirect exposure through the food web.

Image 26.5. Crop Duster. A crop duster plane flies over a field in Washington State. “Crop Duster” by Jenni Reynolds-Kebler is licensed under CC BY-SA 2.0.

The ecotoxicological risk that is inherent in an exposure to pesticide (and to other chemicals) is influenced by a complex of variables, as we previously examined in Chapter 19. Several points should be considered when interpreting exposures of non-target organisms (including people) to pesticides and other chemicals:

• All chemicals are potentially toxic.
• Not all exposures to potentially toxic chemicals result in poisoning (because organisms are to some degree tolerant to pesticides and other chemicals).
• Some pesticides and some naturally occurring chemicals are extremely toxic to many organisms, including humans.
• Humans are subject to both involuntary and voluntary exposures to certain toxic chemicals (the latter includes prescription and recreational drugs).

Of course, pesticides vary enormously in their toxicity. Herbicides, for example, are extremely toxic to at least some plants, but not necessarily to animals, which differ in important physiological respects from plants. In contrast, most insecticides and rodenticides are toxic to a wide range of animals and can cause non-target poisoning of diverse species, including humans.

The acute toxicity of a chemical to animals is defined by its LD₅₀, or the dose needed to kill one-half of a test population that is exposed through food, water, or air. The oral LD₅₀ for rats is an indicator of acute toxicity to mammals. Rats are widely used in toxicological research and are similar to humans in many aspects of their physiology. Table 26.2 compares the acute toxicity of a wide range of pesticides and some other chemicals, using rat oral LD₅₀ (see also Table 14.3). Note that some of the most poisonous chemicals listed are natural biochemicals, such as saxitoxin, a potent neurotoxin produced by certain marine algae. Others are chemicals to which many people expose themselves in their pursuit of pleasure, such as nicotine, the addicting alkaloid in tobacco.

Table 26.2. Acute Toxicity of Various Chemicals to Rats. The oral LD₅₀, measured in milligrams of chemical per kilogram of body weight, is the amount required to kill 50% of a trial population of rats, exposed through their food in a controlled laboratory test. Source: Modified from Freedman (1995).

By poisoning organisms, pesticides may also cause habitat changes to occur, which can indirectly affect many species. For example, herbicide kills plants and thereby changes the habitat of animals, perhaps depriving herbivores of their preferred foods. Similarly, broad-spectrum insecticides kill large numbers of arthropods, which reduces the amount of food available for birds and other animals. These and other indirect effects of pesticide use can result in ecological damage, in addition to the directly toxic effects.

In the remainder of this chapter, we will examine several case studies of particular uses of pesticides. These are useful in illustrating the broader principles and patterns of the ecological damage caused by these chemicals.

DDT and Related Organochlorines

The first case study involves DDT and related organochlorine insecticides, such as DDD, dieldrin, and aldrin. These chemicals were once widely used in the U.S. and most other developed countries. Although these organochlorines were banned here in the early 1970s, they continue to be used in some less-developed nations.

DDT and its relatives are persistent in the environment. Consequently, even though these chemicals have not been used in the U.S. for several decades, there are still substantial residues in the ecosystems of our country. In part, this is also a result of the continued use of these organochlorines in some tropical countries, because small amounts of residue from those ongoing uses are transported to high-latitude countries by global cycling processes. In addition, organochlorines are more persistent in cooler environments than in warmer ones. As a result, these and some non-insecticidal organochlorines (such as PCBs and dioxins) are still important contaminants in North America.

DDT was first synthesized in 1874, but its insecticidal properties were not discovered until 1939. Its first important use was during World War II in programs to control body lice, mosquitoes, and other disease vectors. DDT was quickly recognized as being an extremely effective insecticide, and it became widely used in agriculture, forestry, and against malaria. The use of DDT peaked in 1970, when 175-million kilograms were manufactured globally. Soon afterward, developed countries began to ban most uses of DDT because it was found to be causing ecological damage, including the contamination of people and their food web. Some researchers thought this contamination could be causing illnesses, such as increases in cancer and liver disease. However, DDT use has continued in some tropical countries, mostly against mosquito vectors of disease.

However, even in those countries the use of DDT and other organochlorine insecticides has been decreasing. This is partly because many pests have developed a genetically based tolerance of these chemicals (sometimes known as resistance), which decreases their effectiveness as pesticides. The development of tolerance is an evolutionary process in which exposure to a toxic substance selects for resistant individuals within a genetically variable population (see Chapters 7). Although tolerant individuals are normally rare in unsprayed populations, they may become rapidly dominant in sprayed habitat. If the insecticide does not kill them, they survive to reproduce, and pass on the genes for tolerance to their offspring. More than 500 species of insects and mites have populations tolerant to at least one insecticide, and there are more than 100 fungicide-resistant plant pathogens, 263 herbicide-tolerant weeds, and five rodents resistant to anticoagulants (Winston, 1997; Landis et al., 2002; Heap, 2021).

Several physical and chemical properties of organochlorines have an important influence on their ability to cause ecological damage. First, they have an extended persistence, or a tendency to remain chemically unchanged in the environment because they are not easily degraded by microorganisms or by physical agents such as sunlight or heat. For example, DDT has a half-life in soil of 3-10 years. The primary breakdown product of DDT is the closely related organochlorine DDE, which has a similar persistence.

In addition, DDT and related organochlorines are essentially insoluble in water and so cannot be “diluted” into that abundant solvent. Instead, these chemicals are highly soluble in fats (or lipids), which occur mostly in organisms. Therefore, DDT and related organochlorines have a strong affinity for organisms, and they accumulate in living things in strong preference to the non-living environment – this process is called bioconcentration.

Moreover, organisms are highly efficient at assimilating organochlorines that are present in their food. As a result, predators at the top of the food web develop the highest residues of organochlorines, particularly in their fatty tissues (this is known as biomagnification or food-web magnification). Both bioconcentration and food-web magnification are progressive with age, so the oldest individuals in a population are the most contaminated (see In Detail 17.1).

These properties of organochlorines are illustrated in Figure 26.1 and Table 26.3. Note that the concentrations are miniscule in air, water, and non-agricultural soil, compared with the much higher residues that occur in organisms. Note also that concentrations in plants are lower than in herbivores, and that residues are highest at the top of the food web, such as in predatory birds and humans.

Table 26.3. Typical Residues of DDT in the 1960s and 1970s. Sources: Data from Edwards (1975) and Freedman (1995).

Another characteristic of organochlorines is their ubiquity – their residues occur in all organisms throughout the biosphere. This widespread contamination occurs because organochlorines enter a global cycle and become widely dispersed in the bodies of migrating organisms and in the atmosphere by evaporation and in wind-eroded dust. Residues of DDT are found even in organisms in Antarctica, far remote from areas where it was ever used. In one study in that far-southern region, the concentration of “total DDT” (almost all of which occur as the metabolic residue, DDE) in the fat of skuas (a marine bird, Catharacta maccormicki) was 5 ppm (or 5 µg/g). Smaller residues (< 0.44 ppm) occurred in birds feeding lower in the food web, such as fulmar (Fulmarus glacialoides) and macaroni penguin (Eudyptes chrysolophus) (Norheim et al., 1982).

Although organochlorine residues are ubiquitous in the biosphere, much higher concentrations occur in animals that live close to areas where these chemicals have been used, such as North America. Because marine mammals feed at or near the top of their food web and are long-lived, they can have extremely high residues of organochlorines. For example, harbour porpoises (Phocoena phocoena) in Atlantic North America have had DDT residues as high as 520 ppm in their fat (Edwards, 1975). High residues of organochlorines also occur in top-predator birds, especially raptors (such as eagles, falcons, hawks, and owls). Prior to the banning of DDT, residues averaged 12 ppm (with a maximum of 356 ppm) in a sample of 69 bald eagles (Haliaeetus leucocephalus), up to 460 ppm among 11 western grebes (Aechmophorus occidentalis), and up to 131 ppm among 13 herring gulls (Larus argentatus) (Edwards, 1975).

Intense exposures to organochlorines caused important ecological damage, including bird poisonings. During the 1950s and 1960s, bird kills resulted when DDT was sprayed in urban areas to kill the beetle vectors of Dutch elm disease, caused by a fungal pathogen (Ceratocyctis ulmi) that was accidentally introduced to North America from Europe. The fungus is transported between trees by bark beetles, which can be controlled to some degree by the use of insecticide. Spraying for this purpose was intensive, and typically involved an application of 0.7-1.4 kg of DDT per tree. Birds feeding on invertebrates in treated areas were exposed to lethal doses. One study in New Hampshire found 117 dead birds in a 6 ha spray area, and estimated that 70% of the breeding robins (Turdus migratorius) had been killed (Wurster et al., 1965). So much avian mortality occurred in sprayed neighborhoods that bird song was markedly reduced – hence the title of Rachel Carson’s 1962 book, Silent Spring (In Detail 26.3).

In addition to acute poisoning caused by organochlorines in sprayed areas, more insidious damage was occurring over large regions. Many species experienced long-term chronic toxicity, even well away from sprayed areas. It took years of population monitoring and ecotoxicological research before organochlorines were identified as the cause of this widespread damage. In fact, we can view the chronic poisoning of birds and other wildlife as an ecological “surprise” that occurred because scientists (and society) had no experience with the long-term effects of persistent, biomagnifying organochlorines.

Raptorial birds were among the prominent victims of organochlorine insecticides. These birds are vulnerable because they are top predators and accumulate high residues of organochlorines. Breeding populations of various raptors suffered large declines. Severely affected species included the peregrine falcon (Falco peregrinus), osprey (Pandion haliaetus), bald eagle (Haliaeetus leucocephalus), and golden eagle (Aquila chrysaetos) (see North American Focus 26.1). In all cases, these species were exposed to a “cocktail” of organochlorines, including the insecticides DDT, DDD (both are metabolized to DDE), aldrin, dieldrin, and heptachlor, as well as PCBs, a group of non-insecticidal compounds with many industrial uses. Researchers have investigated the relative importance of these various organochlorines in causing the population declines of raptors. It appears that DDT may have been the more important toxin to birds in North America, while cyclodienes (particularly dieldrin) were more influential in Britain (Cooper, 1991; Moriarty, 1999).

Damage to raptors was mainly associated with chronic effects on their reproduction, rather than toxicity caused to adults. Reproductive damage included the production of thin eggshells that would break under the weight of an incubating parent, high mortality of embryos and nestlings, and abnormal adult behavior. The numbers of fledged chicks declined, and that resulted in rapid population declines.

Since the bans of DDT and other organochlorines in North America, their residues in wildlife have been progressively decreasing. Data showing this decline have been obtained by analyzing eggs of herring gulls (Larus argentatus) breeding on the Great Lakes (Figure 26.2). Although eggs from various places differ in their residues (partly depending on local sources), they all exhibit large decreases in DDE and PCB. Decreases in residues have also occurred in double-crested cormorants (Phalacrocorax auritus).

Figure 26.2. Changes in Organochlorine Residues in Bird Eggs. Eggs of herring gull have shown decreasing residues since the use of DDT, PCBs, and other persistent organochlorines was banned in North America in the early 1970s. Muggs/Leslie are breeding sites in the Toronto waterfront, while Big Sister Island is in Green Bay in Lake Michigan. Residues are measured in ppm. Sources: Data from Bishop and Weseloh (1990), Environment Canada (1993), and Ryckman et al. (2005).

The Case of Carbofuran

Important replacements for DDT and related organochlorines have been organophosphate and carbamate insecticides. These poison insects and other arthropods by inhibiting a specific enzyme, acetylcholinesterase (AChE), which is critical in the transmission of nerve impulses. Vertebrates such as amphibians, fish, birds, and mammals are also sensitive to poisoning of their AChE system. In all of these animals, acute poisoning by organophosphate and carbamate insecticides causes tremors, convulsions, and ultimately death.

Carbofuran is a carbamate insecticide that can be used for many purposes in agriculture. One formulation is a liquid suspension that can be diluted in water and then broadcast-sprayed against pests such as grasshoppers and leaf beetles. It is also available in a granular formulation, in which the insecticide coats particles of grit and is sown along with seeds to protect tender seedlings from insect damage. The granular formulation has been commonly used when sowing canola and maize.

Unfortunately, wildlife is exposed to toxic doses of carbofuran when either of these formulations is used. For example, if not all of the carbofuran granules are buried in the planting furrows, they remain exposed on the surface (Mineau, 1993). In one method of seeding, previously used for corn in Ontario, 15-31% of the granules remained exposed on the surface, or 515-1065 exposed granules per meter of furrow. Methods used to plant canola in western Canada often left about 5% of the granules on the surface. The exposed granules may be ingested by seed-eating birds, which require hard particles of that size as “grit” for macerating hard-coated seeds in their muscular gizzard. The carbofuran is extremely toxic: consumption of only 1-5 granules can kill a small bird. Raptors and mammals are secondarily poisoned if they scavenge the dead bodies.

In addition, fields treated with carbofuran may become flooded during the spring and autumn, and the surface water may then contain large residues of the insecticide. This is particularly the case if the soil and water are acidic, which greatly reduces the rate of breakdown of carbofuran into less toxic chemicals.

Of all the pesticides used recently in agriculture, carbofuran has probably caused the most non-target mortality of birds and other wild animals. Even though there is no systematic program for reporting bird kills caused by pesticide use in Canada or the U.S., a large number of toxic incidents were documented for carbofuran (Mineau, 1993), a few of which were:

• More than 2,000 Lapland longspurs (Calcarius lapponicus), a seed-eating finch, were killed after eating carbofuran granules in a freshly planted canola field in Saskatchewan in May 1984.
• About 1,200 birds, mostly savannah sparrows (Passerculus sandwichensis), were killed in by granular carbofuran in turnip and radish fields in British Columbia in September 1986.
• More than 1,000 green-winged teal (Anas carolinensis) were killed within hours of landing in a flooded turnip field in British Columbia in the autumn of 1975.
• At least 50 mallards (Anas platyrhynchos) and pintails (A. acuta) were poisoned in a flooded field in British Columbia in December 1973
• 2,450 dead widgeon (Mareca americana) found one day after spraying of an alfalfa field in California in March 1974.

These examples are only a small fraction of the known bird kills caused by the routine use of carbofuran in agriculture. There are also, of course, larger numbers of unreported incidents. Because carbofuran use in agriculture carries such a well-known risk of poisoning birds and other wildlife, ecologists and environmentalists lobbied vigorously to have its registration withdrawn for those uses, or at least more tightly controlled. In 1993, the U.S. Environmental Protection Agency banned the sand-based granular formulation of carbofuran, except for some relatively minor uses and a major one (with rice crops) for which there was no suitable alternative. This was a positive action in terms of pesticide regulation, although in the view of many wildlife toxicologists it took an excessively long time for those necessary steps to be taken. Carbofuran is now banned in the U.S., Canada, and the European Union.

Diazinon and Monocrotophos

Other modern insecticides are also poisoning North American birds. Diazinon, an organophosphate insecticide, has caused numerous cases of mass mortality. For instance, in the late 1980s, there were at least five events in which entire flocks of Canada goose (Branta canadensis) were killed when they fed on grass on golf courses in southern Ontario that had been treated to reduce infestations of turf insects (Mineau, 1999). The geese died within minutes, as is typical of acute poisoning by AChE inhibitors. Similar toxic events have occurred in the U.S., including one in which 700 Brant geese (Branta bernicla) were killed on a golf course in New York. Diazinon has now been banned for use on golf courses in the U.S., but it can still be used in Canada for that purpose and also for horticultural and agricultural purposes, although its use is declining.

In 1996, it was discovered that agricultural use of monocrotophos and other organophosphate insecticides against grasshoppers in Argentina was killing large numbers of Swainson’s hawks (Buteo swainsoni). This raptor breeds in western Canada and the U.S. and migrates to winter on the pampas of South America. Populations of this hawk had been declining for several years. However, it was not until some birds were fitted with satellite transmitters and followed to Argentina that wildlife toxicologists discovered a probable cause of the decline – poorly regulated use of monocrotophos on the wintering grounds. Field studies in Argentina discovered that more than 20 thousand Swainson’s hawks had been killed in just one agricultural area (other regions were not surveyed), out of a total breeding population of only 400 thousand (of which up to one-quarter breed in Canada).

Monocrotophos is extremely toxic to birds, although it is not persistent in the environment. Because of the risk of ecological damage, monocrotophos has been banned in the U.S. In Argentina, however, the insecticide could be legally used. Swainson’s hawks were exposed to lethal doses of monocrotophos when they fluttered behind spray tractors to feed on grasshoppers flushed by the machinery, and also when they later fed on insecticide-contaminated prey. Argentina has since banned monocrotophos, replacing its applications with pyrethroids.

Neonicotinoids

Perhaps the most contentious pesticide-related issue of recent years is the use of neonicotinoid insecticides agriculture (Borsuah et al., 2020; Craddock et al., 2019). These are a class of synthetic AChE-inhibiting insecticides that are chemically similar to nicotine, the key alkaloid in tobacco. Neonicotinoids are much less toxic to vertebrate animals that carbamate and organophosphate insecticides, but they are persistent and broad-spectrum poisons to non-target arthropods. The “neonic” group includes a variety of compounds, with imidacloprid recently being one of the most widely used insecticides in the world.

Neonicotinoids are a systemic insecticide. The chemical can be applied with irrigation water, as a water-based spray, or as a coating on seed intended for planting. In all of those cases, the neonic is absorbed and then distributed throughout the plant to confer protection against herbivorous insects. These have become popular uses – since the introduction of neonics in the early 1990s, they have grown to account for several billions of dollars of annual sales. Neonicotinoids are used on a wide range of crops. In the U.S., they are used on 95% of the canola and maize acreage, and at least half of cotton, sorghum, soybean, and sugar beet. They are used on many fruit crops, including almonds, apples, berries, cherries, grapes, oranges, and peaches, as well as vegetables such as leafy greens, potatoes, and tomatoes, and even cereal grains. They are also used as a wood preservative.

Recently, the use of neonicotinoids has been linked to some important environmental problems. One is the general decline in pollinating insects, which are vital both to agricultural production and to natural ecosystems. The agricultural linkage involves that fact that most commercial fruits, ranging from apples to zucchinis, rely on insects to pollinate their flowers so that fruit development can occur. Honeybees are especially important in this regard, yet these vitally important pollinators have been badly declining because of a syndrome referred to as colony collapse disorder. The cause of that damage is not exactly known, but neonics are suspected as having a contributing influence. Another indirect effect of the widespread use of neonics could be declines of some bird species due to a reduction of their arthropod food base (Hallmann et al., 2014).

As a result of the growing concerns about the environmental effects of neonicotinoid insecticides, many countries have been restricting or banning their use. In 2013, a study by the European Food Safety Authority reported an unacceptably high risk to honeybees from many uses of neonics, and in 2014 a critical integrated study was published (van der Sluijs et al, 2014). In 2014, in response to this and other research, 15 of the 27 European Union (EU) member states voted to restrict the use of three neonicotinoids (clothianidin, imidacloprid, and thiamethoxam) for two years while additional studies were undertaken. As of 2018, all outdoor uses of these three neonicotinoids in the EU are banned. While the U.S. EPA has not banned the use of neonicotinoids, in 2020 the agency proposed guidelines to better manage the application of these chemicals, including improving targeted application, restrictions on when the pesticides can be applied, and canceling use on residential turf (EPA 2021).

Pest Problems in Forestry

Pesticides are used much more extensively in agriculture than in forestry (about 89% of pesticide sales in the U.S. are for use in agriculture, 5% for domestic and industrial use, and 6% for home and garden use (Atwood & Paisley-Jones, 2017). However, forestry case studies better illustrate many of the ecological effects of pesticide use because the treated habitats support mainly native species and natural or semi-natural ecosystems. In contrast, agricultural habitats are dominated by non-native species and are intensively managed, making them less amenable to the examination of some of the ecological effects of pesticides.

Spruce Budworm

The largest insecticide spray programs in forestry have been mounted against spruce budworm (Choristoneura fumiferana). Because the spraying occurs over extensive natural forests, whose biodiversity consists of native species, it represents an excellent case study to examine the ecological effects of insecticide use. Spruce budworm is a native moth whose larvae are pests of fir- and spruce-dominated forest (Image 26.8). An infestation can affect tens of millions of hectares, and trees are killed after several years of defoliation. Mature stands dominated by balsam fir (Abies balsamea) are particularly vulnerable. White spruce (Picea glauca) is also a preferred food, but red spruce (P. rubens) and black spruce (P. mariana) are less apt to suffer lethal damage. Spruce budworm is always present in small populations in its fir-spruce habitat, but it occasionally irrupts to an enormous abundance and becomes a pest. Under normal conditions, only about five larvae may occur on each conifer tree, but this increases to 2 thousand at the beginning of an irruption, and to more than 20 thousand during the peak. A local outbreak is typically sustained for 6-10 years, and then collapses. Studies in North America have shown that outbreaks have occurred at an average interval of about 35 years (Blais, 1985). An outbreak is typically synchronous (occurring at the same time) over an extensive area of vulnerable forest, although there are large variations in the abundance of budworm among stands. The exact reasons for the irruptions are not known, but they may involve several years of warm, dry weather in the springtime, which favors the survival of larvae.

Image 26.8. Intensive Forest Damage Caused by Western Spruce Budworm in Glacier National Park, 2012. Source: Chris Peterson via National Park Service.

Forest Damage

It appears that the extent of damage caused by spruce budworm may have increased during the three outbreaks of the twentieth century. The outbreak that began in 1910 affected about 10-million hectares, one starting in 1940 involved 25-million ha, and another in 1970 affected more than 55-million ha (Figure 26.3). The enlarging areas of infestation may be related to an increase in the extent of vulnerable fir-spruce forest, possibly due to the following influences:

• Regeneration of conifer stands on abandoned farmland, particularly since the 1920s
• Protection of forests from wildfire
• Forestry practices such as clear-cutting
• Spraying of infested stands with insecticide, which may help maintain the habitat in a condition suitable for budworm

Figure 26.3. Forest Area Defoliated by Spruce Budworm in the Twentieth Century. The area corresponds to stands suffering severe or moderate defoliation. Source: Modified from Kettela (1983).

Enormous damage has been caused by budworm to economically important forest resources. During the most recent outbreak (1971-1984), tree mortality was equivalent to more than 38-million cubic meters of saleable timber. During the peak of the infestation, substantial tree mortality occurred over about 26.5-million hectares in eastern Canada (Ostaff, 1985). For example, in one studied location (Ostaff and MacLean, 1989), no defoliation by budworm was observed in 1973, but in 1974 there was moderate-to-severe defoliation over 165-thousand hectares. This increased to 486-thousand ha in 1975, and to 1.22-million hectares in 1976, when essentially all of the vulnerable fir-spruce forest was infested.

An irruption of budworm can last for 10 or more years, with the damage increasing over time. During the first two years of severe defoliation in the previously referenced study, 4% of the balsam fir trees died. The cumulative mortality increased to 9% after four years of heavy defoliation, 37% after six years, 48% after eight years, 75% after 10 years, and 95% after 12 years (Ostaff and MacLean, 1989). Across eastern Canada, tree mortality averaged 85% in mature fir-dominated stands, 42% in immature fir stands, and 36% in mature spruce stands (MacLean, 1990).

Mature trees are much more vulnerable to budworm than are smaller immature ones, which commonly survive an outbreak. Consequently, the understory of a damaged stand typically contains a dense population of small fir and spruce. Known as advanced regeneration, this is important in re-establishing the next fir-spruce forest after an infestation collapses. Severely damaged stands typically have an advanced regeneration of 45-thousand small fir plus 3-thousand spruce per hectare, most of which survive the infestation (MacLean, 1988).

After the mature trees die and the canopy opens up, the small conifers grow rapidly and establish another fir-spruce forest, which becomes vulnerable decades later to another irruption of budworm. These observations suggest that, over the long term, the interaction between the forest and budworm can be viewed as an ecologically stable, cyclic succession. The natural cycle of disturbance and recovery has probably been occurring for thousands of years, although human influences may have increased its scale since the nineteenth century.

Of course, spruce budworm causes great economic instability in the forest-products industry, which is in competition with this moth for the fir-spruce resource (Image 26.9). The periodic irruptions of budworm severely damage the forest, making it difficult for humans to plan their own orderly harvesting and management of the trees. Spraying is one way of dealing with this problem, because it can limit the defoliation and prevent some tree mortality. The objective of spraying is not to eradicate the budworm, but to decrease the damage it causes, and thereby maintain the forest resource and its dependent economy.

Image 26.9. Balsam Fir Trees Afflicted by Spruce Budworm. This is a ground-level view of a stand damaged by spruce budworm. Although the mature balsam fir trees have been killed, dense regeneration is occurring in the understory. After 40-50 years, another mature forest will have developed, ready to be harvested by budworm or by humans. Source: B. Freedman.

Insecticide Spraying

After the Second World War, an early use of DDT was against spruce budworm. The early popularity of DDT, a member of the chlorinated hydrocarbon group, was due to its reasonable cost, effectiveness, persistence, and versatility. During the 30 years prior to its cancellation, a total of approximately 1.35 million pounds of DDT was used domestically (EPA, 1975). DDT was sprayed at 1 pound per acre from airplanes for the control of spruce budworm on 71,678 acres in the Yellowstone River drainage in 1957 (Cope, 2011). During the late 1980s, a biological insecticide based on the bacterium Bacillus thuringiensis var. kurstaki (abbreviated B.t.) began to displace the synthetic pesticides in budworm spraying. B.t. is toxic to most lepidopterans (moths and butterflies) and to some other insects, including blackflies and mosquitoes. Otherwise, this insecticide causes little non-target damage. Initially, budworm control using B.t. was variable in effectiveness, and the cost was high compared with fenitrothion. Since then, however, the effectiveness and cost of spraying B.t. have improved. This, along with concern about ecological damage caused by fenitrothion, has resulted in B.t. becoming the insecticide of choice in spray programs against budworm.

Non-Target Damage

The synthetic insecticides used against budworm caused a great deal of non-target mortality. Typical spray rates and toxicities for the insecticides are summarized in Table 26.4. DDT is the least toxic to budworm, but the most toxic to salmonid fish. Perhaps the most important reason for the 1968 ban on using DDT against budworm was the mortality caused to sportfish, particularly Atlantic salmon (Salmo salar) and brook trout (Salvelinus fontinalis).

Table 26.4. Acute Toxicity of Insecticides Used against Spruce Budworm. Typical application rates for forestry purposes are in kilograms per hectare. Acute toxicity to budworm larvae and selected vertebrate species was determined under laboratory conditions. The units of toxicity to budworm are in micrograms per square centimeter of body surface; for trout, data are ppm in water; and for pheasant and rat, data are ppm in food. Note: the 96-hr LC₅₀ is the concentration in water that killed 50% of a population of trout after a 96-hour exposure. Source: Freedman (1995).

The insecticides that initially replaced DDT are more toxic to budworm and so could be sprayed at lower rates while achieving a similar degree of pest control. Although these insecticides are less toxic to fish than DDT was, they can be very poisonous to other animals. Fenitrothion and aminocarb, for example, are highly toxic to all arthropods, so their use results in an enormous kill of non-target insects and spiders, including many predators of budworm. One study estimated that a typical fenitrothion spray killed as many as 7.5-million individuals of hundreds of species of arthropods per hectare, although more than 90% of the dead biomass was budworm (Varty, 1975). Overall, a typical spray with fenitrothion caused a short-term decrease in arthropod biomass of 35%, and a decrease in total individuals of 50% (Agriculture Canada, 1993). However, studies found that the non-target damage to arthropods was temporary and long-term decreases in abundance were not detectable (Varty, 1975; Millikin, 1990). The post-spray recovery was due to recolonization from non-sprayed forest, along with increases by arthropods that survived the spraying.

Bird populations are unusually abundant in infested forest because budworm is a plentiful and nutritious food for insect-eating animals. In fact, some birds are uncommon except in budworm-infested forest. In one study, the breeding population of bay-breasted warbler increased from only 0.25 pairs/ha in non-infested forest to 30 pairs/ha during a budworm outbreak, while Tennessee warbler increased from 0 to 12.5 pairs/ha (Morris et al., 1958).

During an outbreak of budworm, most birds rely heavily on the larvae as food for raising nestlings. One study estimated that birds eat 89-thousand budworm larvae and pupae per hectare of infested forest, compared with 6 thousand/ha in stands without an irruption (Crawford et al., 1983). However, despite the enthusiastic effort, bird predation has no substantial effect on the abundance of budworm during an outbreak – birds consume only about 2% of the larvae. In essence, insectivorous birds are satiated by the extremely abundant food resource and are incapable of controlling the huge population of larvae. Predation by birds may, however, be important in reducing less-abundant budworm populations, and may help to lengthen the interval between outbreaks.

Because birds are abundant in a budworm-infested forest, they are exposed to any insecticide that might be sprayed. Even though some of the insecticides used against budworm are extremely toxic to birds, it has proved difficult to document actual damage to avian populations by spraying. First, it is extremely difficult to find dead or dying birds in forest habitat because they occur in a small density and are quickly scavenged by other animals. Even with all the insecticide spraying in New Brunswick between 1965 and 1987, the Canadian Wildlife Service has records of only 125 dead birds (Busby et al., 1989), which is a gross underestimate of the actual mortality.

In addition, it is difficult to detect population changes resulting from mortality of birds in forest. A census of forest birds is taken by mapping the locations from which male birds sing; this information is then used to determine the boundaries of their territories. The song censuses are conducted in the spring over a period of four to six weeks. During that time, bird populations are dynamic because migratory species are returning from their wintering grounds, and they arrive at different times. Moreover, if a territory-holding bird is killed by an insecticide spray, it may be quickly replaced from a “surplus” of non-breeding individuals that wander extensively, searching for suitable habitat that is not occupied by another of their species. Because of the temporal variations in bird abundance and the rapid replacement of killed individuals, it is difficult to document population changes caused by insecticide spraying.

Phosphamidon is very poisonous to birds (Table 26.4), and its use likely caused severe mortality of some species. One study suggested that as many as 376-thousand ruby-crowned kinglets were killed in New Brunswick during the 1975 spray season, mostly by phosphamidon (Pearce and Peakall, 1977). Because they forage high in the canopy, kinglets are particularly vulnerable to insecticide exposure during an aerial spray. The presumed damage to birds was the key reason why phosphamidon was banned for use against budworm after 1975.

Fenitrothion is less poisonous to birds, but it nevertheless has a small margin of toxicological safety during operational sprays. While a normal application appears to cause little avian mortality, exposure to a double spray, as commonly occurs with overlapping spray swaths, can be lethal. Studies of white-throated sparrows found much greater mortality and behavioral impairment after a double application of fenitrothion, compared with the normal spray rate (Busby et al., 1989).

Table 26.5 shows the effects of a fenitrothion spray on birds. When interpreting these census data, the trends in the sprayed habitat should be compared with those of non-sprayed forest. The comparison is necessary because the “pre-spray” census was conducted in mid-May, when many migratory birds had not yet returned to their breeding habitat. Consequently, the total avian abundance in the pre-spray census was much smaller than occurs later on. If the data are considered in this relative sense, they suggest that the fenitrothion spraying had no obvious effects on the abundance or species composition of the avian community. It is likely, however, that some birds were poisoned by fenitrothion during the spray and that the damage was not reflected in the census for the reasons we noted earlier.

Table 26.5. Effects of Fenitrothion on Bird Populations. The data show the effects of forest spraying with fenitrothion on breeding birds in the Gaspé region of Quebec, Canada. The operational spraying procedure is to treat stands twice, about one week apart. In this study, birds were censused for five days before spraying, then for seven days after the initial spray on May 21, 1976, and again for five days after the second spray on May 30. The sprayed stand (“Spr”) was treated with fenitrothion at 0.56 kg/ha, while the unsprayed reference stand (“Ref”) monitors changes that are unrelated to spraying. Bird density is expressed as numbers per 10 hectares. Only prominent species are listed. Source: Data from Kingsbury and McLeod (1981).

It should also be recognized that a budworm outbreak causes severe damage to the forest, and this affects the habitat of wildlife. Table 26.6 illustrates the effects of this habitat change on birds. Defoliation by budworm caused relatively little damage to Stand A, so the decreased population of some birds (such as solitary vireo and many warblers) was caused by a decrease in the availability of food because the outbreak had collapsed and budworm is a critical resource. In contrast, Stand B was intensely damaged by budworm, with large habitat changes that included many dead trees and a lush growth of understory plants. In this stand, the decreased abundance of some birds (such as solitary vireo, Tennessee warbler, black-throated green warbler, blackburnian warbler, and bay-breasted warbler) resulted from changes in vegetation as well as a reduced availability of larvae as food. Note, in addition, that some birds (such as least flycatcher, magnolia warbler, and white-throated sparrow) do well in recently disturbed stands: they benefit from the habitat associated with budworm damage.

Table 26.6. Abundance of Birds During and After a Spruce Budworm Outbreak. The data are from a study in New Brunswick. Stand A was monitored for eight years during an infestation (1952-1959), and then for six post-outbreak years (to 1965). However, Stand A suffered little damage to its trees – the average age of balsam fir trees was 120 years in both sampling periods. Stand B was censused for five infestation years (1955-1959) and for five post-outbreak years (to 1964). Stand B suffered intense damage to its trees – the fir averaged older than 80 years when the infestation began, but < 10 years old afterward because so many mature trees had died. The bird data are in numbers per 40 hectares. Only prominent species are listed. Source: Data from Gage and Miller (1978).

Overall, it appears that the demonstrated ecological effects of post-DDT spray programs against spruce budworm were relatively short in duration and moderate in intensity (with the exception of the effects of phosphamidon on birds). Residues of such chemicals as fenitrothion and aminocarb are not long lived, and no magnification occurs in the food web. Although substantial mortality was caused to many non-target species, long-term decreases in their populations have not been documented (bearing in mind that such effects are hard to demonstrate, particularly at larger spatial scales). For example, although many individual birds have undoubtedly been poisoned by insecticide toxicity, measurable damage to their populations has not been demonstrated.

Spray Policies

Spraying insecticide on forest infested by spruce budworm has economic benefits that are associated with the protection of conifer trees, an important natural resource. As a result, decision makers and regulators in some areas have considered spray programs to be necessary. Many environmentalists, however, come to different conclusions about the benefits and costs of spraying, because they valuate ecological damage more highly than do resource managers and regulators. Ecologists and environmental activists are not, however, the people who make the decisions about undertaking insecticide spray programs to manage populations of budworm or other pests.

After 1986, fenitrothion was the only synthetic insecticide used against budworm. In 1993, a risk assessment of this use (Pauli et al., 1993) concluded that many of its ecotoxicological damages are significant: “The weight of evidence accumulated with respect to the identified and potential negative impacts caused by the forestry use of fenitrothion on non-target fauna and their potential ecological implications, supports the conclusion that the large-scale spraying of fenitrothion for forest pest control, as currently practiced operationally, is environmentally unacceptable.”

Partly because of the strong conclusions of that risk assessment, the registration of fenitrothion for use in budworm spray programs in Canada was withdrawn in 1995. This action left B.t., a bacterial insecticide that causes little non-target damage, as the major insecticide available for spraying against this forest pest.

Alternatives to Insecticide

In addition to ecotoxicity caused by any insecticide, spray programs against budworm have other drawbacks from a forest-management perspective. Within limits, treating infested stands with insecticide maintains fir-spruce forest “alive and green” and therefore available as a resource for the economically important forest industry. However, spraying also maintains good habitat for budworm, and so may prolong its outbreaks. Agencies that spray insecticide can become “locked” into this pest-management tactic and must continue to spray if the forest resource is to be maintained in an economically viable condition. In the absence of an alternative control practice, spraying may be perceived to be the best available short-term tactic. Clearly, however, spray programs over extensive areas are not desirable.

An alternative to spraying insecticides may be the use of silvicultural practices that are aimed at reducing the vulnerability of stands to budworm infestation. For example, relatively tolerant species such as black spruce might be planted extensively. Alternatively, the landscape could be structured so that the total area of mature but vulnerable forest is kept small. If the landscape mosaic was dominated by less vulnerable stands that could be harvested by industry at about the rate at which the trees mature, a smaller area would be vulnerable to budworm infestation. It must be recognized, however, that such actions would represent a huge intensification of forest management and would cause enormous changes in the ecological character of the landscape. Little is known about the long-term economic viability or ecological consequences of such changes in forest-management strategy.

Further research into alternative methods of budworm control may yield novel methods that are effective. Some promise has been demonstrated by the release of large numbers of tiny Trichogramma wasps that are parasites of budworm, by the use of synthetic hormones that disrupt molting or mating of the budworm, and by several other innovative biotechnologies. So far, however, none of these methods have proven sufficiently effective to be used in operational programs to reduce an irruption of budworm.

If regulators decide that the resource damage caused by spruce budworm must be controlled, it is desirable to utilize the least damaging, but still effective, options. At present, effective control appears to require insecticide, with the only available one being B.t. The use of synthetic organic insecticides in budworm spray programs, with their attendant ecotoxicological damage, appears to be history.

Herbicides in Forestry

The most common use of herbicide in forestry is to keep weeds from competing with young conifers, allowing them to grow more rapidly so that harvests can be more frequent (Newton and Knight, 1981; Freedman, 1995) (Image 26.10). In 2011, herbicides were applied to 4.4% of the total U.S. forest area under management (294,000 ha; Weatherford et al., 2015). However, forestry use affects the habitat of many native species, whereas this is not the case in agriculture and horticulture.

As in agriculture and horticulture, there are alternatives to the use of herbicide in forestry. These options include the crushing of competing vegetation with large machines, manual cutting of weeds with brush saws, prescribed burning, and even using sheep to selectively browse weedy plants. These alternatives can provide a degree of weed control, but foresters generally regard them as more expensive and less effective than herbicide use.

Image 26.10. This is a four-year-old clear-cut of a coniferous forest. Almost all of the vegetation is considered (by foresters) to be “weeds” that compete with desired conifer seedlings for space, water, and nutrients. The objective of an herbicide treatment is to reduce the abundance of weeds, so as to allow the conifers to grow more quickly. Source: B. Freedman.

As with any pesticide, the successful use of herbicide for a management purpose requires the selection of an appropriate chemical and its proper application. If the right choices are not made, the weeds will not be adequately controlled, and the conifer crop may be injured. In addition, the suppression of “weeds” affects the useful ecological services that they provide, such as helping to control erosion and reducing losses of nutrients by leaching. Using herbicide also reduces employment opportunities available in manual weed-control programs. For these and other reasons, including the fears that many people have about the potential toxicological risks of pesticides in the environment, the use of herbicide (and insecticide) in forestry has been very controversial.

Weeds in Forestry

Any disturbance of a forest is followed by vigorous regeneration involving many species of plants that compete for space, nutrients, and moisture. This is true whether the disturbance is caused by a natural fire, insect attack, or clear-cutting. During the first 10 to 15 years of succession, the plant community is dominated by many plants other than the conifers that are desired by the forest industry. This can be illustrated by examining vegetation data for young clear-cuts, where conifers contributed only 4-9% of the plant cover (Figure 26.4). Other species that are not economically desirable (from the forestry perspective) are much more abundant – these “weeds” include ferns, monocotyledonous plants such as sedges and grasses, dicotyledonous herbs such as asters and goldenrods, low shrubs such as raspberries and blackberries, and taller shrubs such as birches, maples, and cherries. The dominance of the site by “undesirable” plants inhibits the growth of commercially desired conifers, and provides an economic justification for a weed-management treatment.

Figure 26.4. Dominant Plants in Disturbed Forest. The vegetation was surveyed in four clear-cuts of conifer forest that were 4-6 years old. The data represent the range of average plant cover among the four sites, expressed as the percentage of ground that is obscured by foliage. Because of overlap, cover values can exceed 100%. Source: Data from Freedman (1995).

The effects of competing plants on conifer productivity are illustrated by a study of a site in Canada that had been treated with herbicides 28 years previously (MacLean and Morgan, 1983). Prior to the herbicide treatment, the vigorously growing vegetation had been dominated by angiosperm shrubs that formed a dense, 2-m-tall canopy that overtopped the shorter conifers. The herbicide spray had released the conifers from some of the stresses of competition. The study consequently found, 28 years after the herbicide treatment, that the biomass of balsam fir on the sprayed plots was about three-times larger than on an adjacent unsprayed plot. From the forestry perspective, this means that the herbicide treatment allowed a conifer-dominated stand to develop more quickly, shortening the time to the next harvest.

Toxicological and Ecological Effects

The silvicultural objective of herbicide spraying is to manage vegetation by changing its character. An herbicide treatment in forestry reduces the abundance of competing vegetation, but it rapidly recovers. In essence, an herbicide treatment returns the post-harvest regeneration (usually post-clear-cutting) to an earlier successional stage, while for several years releasing small conifer plants from some effects of competition. The changes in vegetation are illustrated by a study in Nova Scotia, in which a substantial recovery occurred within only one growing season after an herbicide treatment (Figure 26.5). The recovery involved species whose seeds colonize sprayed sites, as well as plants that were not killed by the herbicide. This study found that no species were eliminated from sprayed clear-cuts, although large differences occurred in their relative abundance between sprayed and reference (unsprayed) plots. This is because species vary in susceptibility to herbicides and in their ability to recover from a disturbance.

Figure 26.5. Recovery of Vegetation after Herbicide Treatment. Regenerating clear-cuts were treated with the herbicide glyphosate at two sites. The reference plots were not sprayed and illustrate the normal recovery of vegetation after clear-cutting. The sprayed plots were sampled for one year prior to the herbicide treatment, and the post-spray recovery was then monitored for several years. The data are percentage plant cover at the end of the summer. Source: Data from Freedman et al. (1993).

Compared with many insecticides, herbicides used in forestry (such as 2,4,5-T, 2,4-D, and glyphosate) are not very toxic to animals (see Table 26.2). At exposures encountered by animals during typical forestry uses, the direct toxicological risks are probably unimportant. This is particularly true of glyphosate, the most commonly used herbicide (Image 26.11).

Glyphosate is extremely toxic to most plants, acting by blocking the synthesis of several essential amino acids. All plants and some microorganisms use this metabolic pathway, but animals obtain these amino acids in their food. Consequently, glyphosate is non-toxic to animals (see Tables 14.2 and 26.2).

Nevertheless, glyphosate causes large changes to occur in habitat because it affects the productivity and biomass of plants. Birds and other animals can be affected by a decreased availability of berries and other plant foods. In addition, the reduced foliage biomass on sprayed areas sustains a lower abundance of insects and spiders, which are important foods for most birds. These are indirect ecotoxicological effects of herbicide spraying, and they affect birds and other wildlife even if they are not directly poisoned.

Image 26.11. Application of Glyphosate. A helicopter releases a silvicultural application of the herbicide glyphosate to a clear-cut. Source: B. Freedman.

A study in Canada found only small changes in the abundance of birds that were breeding on clear-cuts treated with glyphosate (Table 26.7). Although the avian population decreased between the pre-spray and first post-spray years, this also occurred on the reference plot, suggesting it was caused by a factor unrelated to the herbicide treatment, such as bad weather. In the second year after spraying, the abundance of birds on the sprayed plots was similar to the first post-spray year, while on the unsprayed plot it increased to about the pre-spray value.

Table 26.7. Populations of Breeding Birds on Herbicide-Treated Clear-Cuts. Average data are presented for four sprayed plots and one unsprayed reference plot. The sprayed plots were treated with glyphosate. The data are pairs of breeding birds per square kilometer, surveyed for one year before the herbicide treatment (year 0), and then for four post-spray years. Only abundant species are listed here. Source: Data from MacKinnon and Freedman. (1993).

The most common species were white-throated sparrow and common yellowthroat, which had a decreased abundance on both the sprayed and reference plots up to the second year after spraying, and then recovered by the fourth post-spray year. On the reference plot, song sparrow and Lincoln’s sparrow declined during the course of the study, whereas on the sprayed plots they were most abundant in the second and fourth years after spraying. The reference plot became colonized by some new species, including black-and-white warbler, red-eyed vireo, ruby-throated hummingbird, and palm warbler. These species did not invade the sprayed plots because the herbicide treatment caused the habitat to revert to a younger stage that was less favorable to these birds.

Most studies of the effects of herbicides on deer and moose have examined the availability of their food. Broad-leaved shrubs are a preferred food (known as browse) for species of deer, but they are also important weeds in forestry and are a target of herbicide treatments. The quantity of browse, although initially reduced by herbicide spraying, is often increased in the medium-term through the regeneration of shrubs. For example, studies in Maine found that several years after spraying, the availability of browse was greater on treated clear-cuts, partly because the height of the shrub canopy was lower, which gave white-tailed deer easier access to this food (Newton et al., 1989). This is not always the case, however, and some studies have shown that herbicide spraying may decrease the habitat quality for deer.

Overall, field research suggests that herbicide use in forestry has relatively small effects on birds and other wildlife that utilize clear-cuts as habitat. Other disturbances associated with forestry cause much larger effects on wildlife, particularly clear-cutting and the conversion of the natural forest into a plantation (see Chapter 27).

The hazards to people from herbicide use in forestry have also been scrutinized. The concerns include occupational risks for people who are engaged in spraying or are working in recently sprayed areas, as well as risks for the general population. These issues are highly controversial. Much of the concern has focused on the phenoxy herbicides 2,4,5-T and 2,4-D, partly because 2,4,5-T may contain a trace contamination of TCDD, a toxic dioxin. However, the use of other herbicides, such as glyphosate, is also controversial.

It is somewhat reassuring to know that many scientists believe that herbicides can be used safely in forestry (and in agriculture and horticulture), provided the instructions for their use are followed carefully. Many scientists also believe that herbicides do not cause undue risks to sprayers or people living near the treated areas. These are among the reasons why governments have registered these pesticides for uses that are economically beneficial by allowing greater productivity of both agricultural and forest crops. It must be remembered, however, that scientists have not achieved a full consensus on these issues. Partly for this reason, the use of pesticides in forestry and for other purposes continues to be highly controversial.

Integrated Pest Management

Pesticides are commonly used in agriculture, horticulture, and forestry. It is clear from this fact that most politicians, bureaucrats, and resource managers – and many scientists – have decided that the environmental “costs” associated with the use of pesticides are “acceptable” in view of economic benefits that are achieved. It is debatable, however, whether reliance on pesticide use is desirable over the long term. This is particularly true for those pesticides that are toxic to a broad spectrum of organisms. Most people would prefer that less reliance be placed on such non-specific methods of pest control.

A much preferable approach is known as integrated pest management (IPM; Figure 26.6), which employs an array of complementary tactics toward achieving pest control, with the aim of there being fewer environmental and health risks. Elements of an IPM system can include the following:

• Use of natural predators, parasites, and other biological agents that can help to control a pest, while causing few non-target damages
• Use of crop varieties that are resistant to pests
• Management of habitat to make it less suitable for pests
• Careful monitoring of pest abundance, so control measures are undertaken only when necessary
• Use of pesticides, but only if required as a component of an IPM strategy

Figure 26.6. Pyramid of Integrated Pest Management Tactics for Crops, Lawn, and Garden. Source: Penn State University, 2011 via EPA.

A successful IPM program can greatly reduce, but not necessarily eliminate, reliance on pesticides. For example, for many years the cultivation of cotton in the southern U.S. relied on the intensive application of insecticides against pests such as the boll weevil. Widespread use of an IPM system to control this insect in Texas cottonfields reduced insecticide use from 8.8 million kilogram in 1964 to 1.05 million kilogram in 1976 (Bottrell and Smith, 1982). Nevertheless, insecticide use against this pest remained necessary.

Wherever possible, IPM systems utilize control methods that are as specific as possible to the pest so that non-target damage can be avoided or greatly reduced. Some of the best examples of such specific methods involve biological control (the use of a biological agent). The usefulness of biological control can be illustrated with the following successful controls of introduced agricultural pests (Freedman, 1995):

• The cottony-cushion scale (Icyera purchasi) is a sap-sucking insect that was accidentally introduced to the U.S., where it became a threat to citrus agriculture. Research in its native Australia discovered that the pest was naturally controlled by certain insect predators and parasites. In 1888, two of its predators, a lady beetle and a parasitic fly, were introduced to California. This allowed almost total control over this potentially disastrous pest. Unfortunately, this biological control was disrupted when DDT and other broad-spectrum insecticides were used to deal with other orchard pests beginning in the late 1940s.
• St. John’s wort (Hypericum perforatum), a common weed, is toxic to cattle. It became a serious pest in pastures after it was introduced from Europe. In 1943, two leaf beetles that feed on this plant were released to North America, and this pest is no longer an important problem.
• The prickly pear cactus (Opuntia stricta) was imported to Australia from North America and grown as an ornamental plant and a “living fence.” It escaped and became a serious weed in rangelands (Image 26.12). This pest was controlled by the introduction of one of its herbivores, a moth whose larvae feed on the cactus.
• Ragwort (Senecio jacobea) is a Eurasian plant that has been introduced to the Americas and Australia. It became an important weed in rangeland because it crowds out native plants and is toxic to cattle. Several of its Eurasian herbivores are now being used to control its abundance, including the cinnabar moth, ragwort flea beetle, and ragwort seed fly.
• The screw-worm fly (Callitroga hominivorax) causes damage to cattle when its larvae feed on open wounds. This pest has been controlled in some areas through the release of large numbers of male flies that were reared in laboratories and sterilized by irradiation. Because the female flies mate only once, copulation with a sterile male results in unsuccessful reproduction. If this happens to enough females, the abundance of the pest decreases to an acceptable level.

Image 26.12. Prickly Pear Cactus (Opuntia stricta). This species, native to Australia, is now invasive in North American rangelands. Source: “Opuntia stricta” by Christian Ferrer is licensed under CC BY-SA 4.0.

Unfortunately, biological control may not be suitable for all pest problems, and in fact it has not succeeded in most cases in which it has been attempted. The list of failures includes forest pests such as spruce budworm and gypsy moth. Researchers have investigated the potential for controlling budworm using pest-specific bacteria, viruses, and other agents of diseases, wasps that parasitize and kill larvae, and sex and developmental hormones to disrupt mating and growth. Some of the biological control methods have shown promise, but they do not yet achieve a consistent kill of budworm and are relatively expensive. For these reasons, they are not considered ready for routine use against this important pest.

The only viable alternative to the broadcast spraying of synthetic insecticides to control budworm is an insecticide based on the bacterium B.t. (examined earlier). Research into other biological controls continues, and some may yet prove successful and would allow managers to develop an effective IPM system that does not rely on broadcast spraying of insecticides, even relatively specific ones such as B.t.

At the present time, however, it appears that society will continue to rely heavily on pesticide use in intensively managed systems in agriculture and forestry. This will happen even though the intensive systems cause ecological damage (partly because this damage is not fully accounted for as an economic “cost”).

It is important that additional research is undertaken to develop viable methods of biological control and other elements of IPM systems. This is necessary if the present reliance on the pesticide treadmill is to be replaced with less-damaging methods of pest management. Such a change would deliver substantial benefits to society, because the agricultural and forestry systems that we require for sustenance could be managed on a more ecologically sustainable basis. In part, this will require that more attention be directed to the ecological damage caused by the intensive use of pesticides.

Because of this damage, it is highly desirable that non-pesticidal alternatives to pest control are discovered as quickly as possible. Until this happens, pesticide use should be reduced to the lowest levels that continue to effectively control the pests. Some environmentalists have argued that pesticide use in North America is much greater than necessary and that it could be decreased without causing a significant adverse effect on crop yields. In fact, the European Union has already passed legislation requiring that agricultural pesticide use be reduced by half, and many pesticides have been banned. Some argue that serious consideration should be given for similar legislation in other countries such as the U.S. and Canada.

Conclusions

Pesticides are a wide range of substances that are used to gain an advantage over species that cause diseases or are pests in agriculture, forestry, or horticulture. However, the use of many pesticides carries risks of causing damage to human health or the environment. Pesticides have become an integrated component of most of the intensive systems by which foods and other crops are grown, and there are not yet good replacements for all of their uses. For this reason, the use of pesticides will continue into the foreseeable future. Nevertheless, it is important that more research be done to find effective ways of reducing the dependence on pesticides, especially on integrated pest-management systems and on means of biological control. In the meantime, it is important that pesticide use be reduced to the lowest amounts possible and that the most damaging chemicals are withdrawn from legal use.

Questions for Review

1. Classify pesticides according to their intended targets and also by their major chemical groups.
2. What is a “pest”? Why do people consider it necessary to manage their abundance in agriculture, horticulture, forestry, and public health?
3. What characteristics of organochlorines have caused them to become global contaminants? Why do they pose special toxicological risks to top predators?
4. In view of the fact that spruce budworm is a native insect, why is the damage it causes to conifer forest considered to be a problem?

Questions for Discussion

1. Identify and compare the benefits and environmental risks associated with pesticide use in one of agriculture, horticulture, or forestry.
2. Are there effective alternatives to the continued use of pesticides? Consider the roles of integrated pest management, biological controls, and other options.
3. For many common uses of pesticides, there are alternative ways of managing the targeted pest. For example, weeds in a lawn can be controlled by digging them out, rather than by using herbicide. Pest rodents could be trapped instead of being killed with a rodenticide. What do you think should be the key considerations when deciding whether to use pesticides or alternative means of control?
4. Some people believe that the use of pesticides should be allowed only in extreme cases, for example, to save human lives or to prevent a food catastrophe. Most pesticide use, however, is more routine than this. What do you think about this issue? Should it be made more difficult for people to use pesticides? Or should farmers and other potential users have freedom to make their own choices about pesticide use?

Exploring Issues

1. A large corporation is seeking permission from the government to market a new insecticide in the U.S. You are a wildlife biologist and have been asked to recommend studies to identify whether unacceptable damages would be caused to animals, plants, or ecosystems by the new insecticide. How would you design such a study? What specific questions would you want to answer?
2. You are an environmental scientist and have been asked to provide expert advice about a proposed new municipal bylaw regulating the cosmetic use of pesticides in horticulture. From the environmental perspective, what would you consider to be the benefits and problems with the proposed bylaw?

References Cited and Further Reading

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Benbrook, C.M. 2016. Trends in glyphosate herbicide use in the United States and globally. Environ. Sci. Eur. 28: 3 10.1186/s12302-016-0070-0.

Beyond Pesticides. 2017. Lawn and Garden Pesticides Facts and Figures. https://www.beyondpesticides.org/assets/media/documents/bp-fact-lawnpesticides.081817.pdf. Accessed June 16, 2021.

Borsuah, J.F., T.L. Messer, D.D. Snow, S.D. Comfort, A.R. Mittelstet. 2020. Literature review: global neonicotinoid insecticide occurrence in aquatic environments. Water 12(12): 3388, https://doi.org/10.3390/w12123388.

Bottrell, D.G. and R.F. Smith. 1982. Integrated pest management. Environmental Science and Technology, 16: 282A-288A.

Bradshaw, C.J.A., B. Leroy, C. Bellard, D. Roiz, C. Albert, A. Fournier, M. Barbet-Massin, J.M. Salles, F. Simard, and F. Courchamp. 2016. Massive yet grossly underestimated global costs of invasive insects. Nature Communications 7: 12986, doi: 10.1038/ncomms12986.

Briggs, S.A. 1992. Basic Guide to Pesticides: Their Characteristics and Hazards. Taylor & Francis, Washington, DC.

Busby, D.G., L.M. White, P.A. Pearce, and P. Mineau. 1989. Fenitrothion effects on forest songbirds: A critical new look. Pp. 43-108 in: Environmental Effects of Fenitrothion Use in Forestry. Environment Canada, Dartmouth, NS.

Carson, R. 1962. Silent Spring. Houghton-Mifflin, Boston, MA.

Cooper, K. 1991. Effects of pesticides on wildlife. Pp. 463-496 in: Handbook of Pesticide Toxicology. Vol. 1, General Principles. (W.C. Hayes and E.R. Laws, eds.). Academic Press, San Diego, CA.

Cope, O.B. 2011. Effects of DDT spraying for spruce budworm on fish in the Yellowstone River System. Transactions of the American Fisheries Society 90(3): 239-251, https://doi.org/10.1577/1548-8659(1961)90[239:EODSFS]2.0.CO;2

Craddock, H.A., Huang, D., Turner, P.C. et al. 2019. Trends in neonicotinoid pesticide residues in food and water in the United States, 1999–2015. Environ Health 18, 7. https://doi.org/10.1186/s12940-018-0441-7

Crawford, H.S., R.W. Titterington, and D.T. Jennings. 1983. Bird predation and spruce budworm populations. Journal of Forestry, 81: 433-435.

Donley, N. 2019. The USA lags behind other agricultural nations in banning harmful pesticides. Environmental Health 18(44), https://doi.org/10.1186/s12940-019-0488-0.

Edwards, C.A. 1975. Persistent Pesticides in the Environment. CRC Press, Cleveland, OH.

Ennis, T. and E.T.N. Caldwell. 1991. Spruce budworm, chemical and biological control. Pp. 621-641 in: Tortricid Pests, Their Biology, Natural Enemies, and Control. (L.P.S. van der Geest and H.H. Evenhuis, eds.). Elsevier, Amsterdam.

Environment Canada. 1993. Toxic Contaminants in the Environment: Persistent Organochlorines. Ottawa: State of the Environment Reporting. Environment Canada. 1996. The State of Canada’s Environment, 1996. Government of Canada, Ottawa, ON.

Environment Canada. 2001. Assessment Report; Nonylphenol and Its Ethoxylates. Environment Canada. Ottawa, ON.

Environmental Protection Agency (EPA). 1975. DDT Regulatory History: A Brief Survey (to 1975). https://archive.epa.gov/epa/aboutepa/ddt-regulatory-history-brief-survey-1975.html

Environmental Protection Agency (EPA). 2003. The Great Lakes; An Environmental Atlas and Resource Book. U.S. Environmental Protection Agency, Washington, DC.

Environmental Protection Agency (EPA). 2008. “Inert or “Other” Ingredients in Pesticide Products. U.S. Environmental Protection Agency, Washington, DC.

Environmental Protection Agency (EPA). 2010. Nonylphenol (NP) and Nonylphenol Ethoxylates (NPEs) Action Plan. https://www.epa.gov/sites/production/files/2015-09/documents/rin2070-za09_np-npes_action_plan_final_2010-08-09.pdf. Accessed June 16, 2021.

Environmental Protection Agency (EPA). 2021. EPA Actions to Protect Pollinators. https://www.epa.gov/pollinator-protection/epa-actions-protect-pollinators. Accessed June 16, 2021.

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