Population extinction can result from many factors. Even though most cases are hidden from observers, its processes can be viewed in different ways. The other chapters of this book focus on single features of species biology that contribute to species extinction. Superficially, this chapter treats yet another factor: the extension of the spatial stage on which the extinction drama is played. Nonetheless, space is something different. It affects and is, in turn, affected by the other aspects of species biology that contribute to extinction.

The spatial extension of ecological systems, which considers the actual locations of organisms in the landscape, is not routinely incorporated into theoretical formulations of population genetics, demography, population dynamics, and community ecology. Our theories typically present variables such as N's and p's that summarize, with a single number, the ecological and genetical states of a system over some conceptually delimited region of physical space. That is, N is a count of all the animals in this space; it does not tell where they are, or how they are clumped or otherwise associated. Similarly, p represents a gene frequency in a ‘population’ of organisms, but the region over which this estimate is valid is not normally specified.

This reluctance to address questions of spatial extension results from at least two important considerations.

INTRODUCTION

In comparing the reactions of different aquatic organisms to toxic chemicals (whether they are inorganic in the form of heavy metals, ammonia, cyanides etc. or more complex organic pesticides), it has long been the aim of the toxicologists involved to attain some standardisation or uniformity in the experimental techniques they devise. This approach to unification of methods has been particularly marked in the case of freshwater fish, for which group there are now standard methods acceptable in most countries.

These standard methods lay down guidelines on the exact conditions of the tests vis-à-vis water quality, temperature, number and size of test organisms, duration of exposure to the chemical dilution and, finally, the criteria to be adopted when measuring the effect in terms of immobility or mortality.

The need for the same degree of uniformity in testing other forms of freshwater life, invertebrates in particular, though slow to develop has gained increasing momentum in the last 10 years or so. Much of this incentive has come from the US Environmental Protection Agency (EPA) which now plays an international role in the screening and clearance of all pesticides and other toxic chemicals likely to have a harmful effect on the environment. One of the first fruits of this increased interest in aquatic invertebrates appeared in 1972 in the form of an exhaustive summary – compiled by the EPA – of all published information on their reactions in the laboratory to a wide range of pesticides, mainly insecticides, and herbicides (NTIS, 1972).

INTRODUCTION

The use of herbicides to control undesirable plant growth first developed on a large scale shortly after World War II and has been extending rapidly ever since that time. With their increasing use in agriculture, forestry and water-way clearance particularly in developing countries these chemicals now rank alongside insecticides as major environmental contaminants (Balk & Koeman, 1984). The continuous monitoring programme of streams flowing into the Great Lakes over the last 10 years for example, has shown that herbicide use in agricultural land has now increased to such an extent that they now constitute more than half the total volume of pesticides used in agriculture (Frank et al., 1982). Even in the UK where there are unusually stringent regulations controlling pesticides in the environment – particularly with regard to natural water bodies – many of the long-established herbicides such as 2, 4-D, dalapon, dichlobenil and diquat have been cleared under the Pesticides Safety Precautions Scheme, 1973, for use as aquatic herbicides for control of submerged and emergent aquatic weeds, and for the control of vegetation along the banks of rivers and drainage channels (Ministry of Agriculture, Fisheries & Food, 1985).

Early recognition of possible effects on fish life of herbicides applied directly to water or contaminating water by run-off from agricultural land, led to very thorough laboratory investigations in the UK on fish toxicity, and established the relative lethal levels of about 20 common herbicides based on 24-h LC50 values (Alabaster, 1969).

At the time when I last carried out a review of the subject of pesticides and freshwater fauna (Muirhead-Thomson, 1971) it was still possible for a single author to do justice, within one book, to the information then available regarding all forms of freshwater animal life and all types of freshwater body. In the 15 years since that book was published, there has not only been an enormous proliferation of knowledge about this subject but also noteworthy changes in emphasis and priorities. There has been increasing specialisation within this general subject as well, making it increasingly difficult for a single author to encompass all aspects of this problem. For all these reasons, the scope of the present review is restricted to running waters, rivers and streams, and to the macroinvertebrate fauna of such water bodies. The restriction to macroinvertebrate fauna is dictated in part by the fact that a great deal of the voluminous literature in the last 15 years deals with studies on the reactions of freshwater fish, to such an extent that a competent review of that aspect, including all the physiological work on uptake and retention of pesticides by different organs, would require a separate volume. However, one aspect of those fish studies cannot be omitted from any review devoted to aquatic macroinvertebrates, that is the effect of pesticides and allied toxic chemicals on feeding habits of fish in so far as these are influenced by drastic changes in the availability of different invertebrate fish food organisms, as measured by changes in the composition of the stomach contents.

INTRODUCTION TO BLACKFLY CONTROL

Blackflies (Simuliidae) are biting flies which are widely distributed in both temperate and tropical regions. In some northern countries such as Canada their main economic importance is as biting and bloodsucking pests of humans and domestic stock. In other regions such as tropical Africa and Central America their main importance is in their role of vectors of human diseases such as onchocerciasis caused by a parasitic filarial worm. A feature common to all species of Simulium is their association with running waters which form the larval habitat. According to species and country, these habitats or breeding places may range from quite small trickling streams to very large rivers of Africa such as the Niger, the Zaire, the Volta and the Nile. In many of these rivers and larger streams the highest larval populations tend to be concentrated in the fast-flowing sections of turbulent water such as those associated with rapids and dam spillways.

Shortly after the discovery and rapid developments of DDT in the early 1940s, it was found that Simulium larvae are extremely sensitive to this insecticide and that, in some cases, effective larval control was still achieved many miles downstream from the point of application. This opened up entirely new possibilities for Simulium control on a large scale by effectively reducing or eradicating larval populations with insecticide. In one of these early and successful control operations DDT was applied at high dosages of 5–10 ppm, or even greater on occasions, which produced massive fish kills and drastic effects on other stream fauna.

THE SPRUCE BUDWORM IN CANADA AND THE US

Since the early 1950s the spruce budworm (Choristoneura fumiferana) has posed a serious threat in parts of eastern Canada, particularly New Brunswick, (Eidt, 1975, 1977; Symons, 1977a) and adjacent states of the US (Nash, Peterson & Chansler, 1971). In order to protect the valuable timber trees against defoliation, the method of control originally adopted was aerial spraying with DDT, which was practised from 1952 onwards. Since that time the intensity and extent of the infestation has increased. In New Brunswick for example, between 1952 and 1957 the sprayed area increased from 75 × 103 ha to 2.3 × 106 ha (8876 sq. miles). Many of the areas treated twice a year recorded a total application of 560 g/ha DDT per annum, and it was after such heavy treatment that Atlantic salmon, living in streams and rivers in the sprayed forest area, were found to be severely affected.

DDT began to be phased out in 1968, and by 19 70 was replaced completely by organophosphorus compounds, mainly fenitrothion. By 1976 the sprayed area in New Brunswick had increased to 4.0 × 106 ha (15000 sq. miles). By that year infestation had extended to other provinces, Quebec, Ontario and parts of Newfoundland and Nova Scotia up to a total area of 30 × 106 ha.

Fenitrothion continues to be the insecticide of choice, in Canada, and at the time of writing appears unlikely to be superseded by other insecticides (D. C. Eidt, personal communication).

The term ‘pesticide’ embraces a wide range of toxic chemicals used for controlling or eradicating undesirable forms of life. Compounds specifically designed for the control of insects and other arthropods, i.e. insecticides, make up the bulk of these; another range of pesticides is designed to deal with undesirable fish (both predatory and competitive) while still others were developed for use against the aquatic snails which harbour intermediate stages of human parasites. The term pesticide now conveniently includes herbicides, chemicals specifically designed for control of undesirable plant growth, and this inclusion recognises the fact that the greatly increased use of herbicides in recent years has, in many cases, placed these chemicals equal to or ahead of insecticides as major environmental contaminants (Balk & Koeman, 1984)

CONTAMINATION AS A DIRECT CONSEQUENCE OF PEST CONTROL OPERATIONS

DIRECT APPLICATION OF PESTICIDE TO WATERBODY

Pesticide contamination of running waters can occur in many different ways and from many different sources, and may be only of short duration or it may be prolonged. In view of the emphasis in this review on the problems of evaluation, it would be convenient to consider pesticide contamination under two main categories. In the first category would be listed all those cases where the presence of pesticide in running water is the direct consequence of control operations carried out against undesirable fauna or flora.

INTRODUCTION TO FISH TOXICANTS: DEVELOPMENT OF SELECTIVE PISCICIDES

Fish toxicants are widely used to eradicate some or all of the fish in a body of water in order that desirable fish may be stocked, free from predation, from competition or from other interference from undesirable fish. Fish poisons have a long history of use in many countries but it is only in the last 40 years that the subject has really been scientifically investigated, and only within the last 20 that the full environmental or ecological effect of such toxic chemicals has been examined critically, particularly in the US and in Canada. Fish toxicants have been used in all types of water body, both static and running. Earlier progress in their study has been exhaustively reviewed (Lennon et al., 1971), and information available at that time regarding the reactions of freshwater fauna in general, including fish, to those fish toxicants was also the subject of a separate review at that time (Muirhead-Thomson, 1971). In the present review, space limitations would now make it extremely difficult to do justice to the mass of new information covering all freshwater fauna and all types of water body. Accordingly, in keeping with the scope of the coverage, progress since that time will deal only with the use of fish toxicants in running water, and only with the macroinvertebrate fauna at risk.

INTRODUCTION

For many years, control of tsetse fly in Africa was carried out by a variety of methods based on environmental manipulation, such as bush clearing, game exclusion, habitat destruction by burning etc. The choice of methods was mainly determined by the nature of the habitats characteristic of different species of tsetse, and also by whether the objective of these operations was tsetse control or tsetse eradication.

With the advent of the synthetic insecticide DDT and its allies, increasing emphasis has been on the application of insecticide to the tsetse environment either by means of heavy residual dosages to tsetse-resting sites, or by repeated non-residual applications at lower dosage rates (Jordan, 1974). Initially the insecticides of choice were DDT and dieldrin, the latter being favoured because of its higher toxicity to tsetse. However, it was recognised early that such tsetse control measures had a serious immediate effect on wildlife, mammals, birds, reptiles and fish (Graham, 1964). Over the last 20 years therefore, the preferred insecticide for tsetse control has been the allied organochlorine chemical, endosulphan (Thiodan) (Goebel et al., 1982) selected because of its high lethal effect on tsetse combined with less inimical effect on wildlife (Hocking et al., 1966; Park et al., 1972). That period has also been marked by operational changes; insecticides originally applied by means of ground spraying or fogging equipment, are now applied almost entirely from the air, both by fixed-wing planes and by helicopter.

From the very wide range of activities dealt with in this review, dealing with a great variety of problems in many countries, it is clear that the period under review has been one marked by a great upsurge of interest, and by remarkable progress in a hitherto much-neglected aspect of running water contamination. In fact this period of intense interest in the special problems of the macroinvertebrate fauna (as distinct from those of freshwater fish) of streams and rivers, may indeed encompass a peak period of study which is perhaps already on the decline. This is perhaps inevitable. Problems of particular interest or importance, or ones for which unusually ample funds are available for research, attract and encourage the most competent researchers, whose interest and scientific dedication in turn engenders further enthusiasm (Brown, 1973). This leads to the build up of multidisciplinary teams capable of making massive contributions to knowledge during their tenure.

This has certainly been the case with some of the major projects examined in detail in this review. For example, the OCP (Onchocerciasis Control Programme) inaugurated in 1974 on a 20-year basis has produced team work of the highest productivity during its first 10 years. Key research workers tend to move to other fields, or to retire; the teams become disbanded and funds for research tend to dwindle. It is difficult to visualise that the same intense research effort on the impact of Simulium control measures on non-target macroinvertebrates, and on stream ecosystems, will continue at the same high pitch for the next 10 years of the programme.

TOLERANCE LEVELS UNDER STANDARD CONDITIONS

REACTIONS IN STATIC TESTS

With the widespread use of DDT and allied chlorinated hydrocarbon insecticides for pest control in the 1950s and 1960s it became increasingly clear that stream invertebrates were highly vulnerable to these chemicals. This destructive effect was particularly evident when stream and river habitats were unavoidably contaminated by repeated aerial applications of insecticides against forest pests such as the American spruce budworm. Experiences with mass destruction of river fauna in one of these campaigns in the Yellowstone National Park in the US, stimulated the need for more precise information about the susceptibility of stream invertebrates to the different insecticides then in use. This resulted in perhaps the first serious development of a laboratory evaluation programme for stream invertebrates in general as distinct from such target fauna as Simulium larvae (Gaufin, Jensen & Nelson, 1961; Jensen & Gaufin, 1964, 1966; Gaufin et al., 1965). That and the other contemporary work has already been reviewed in depth (Muirhead-Thomson, 1971), but there are still aspects of those studies which are of particular significance in the light of developments in the 20–25 years since that period. First of all, two aquatic invertebrates which played a prominent part in those laboratory tests, namely the stonefly (Plecoptera) nymphs of Pteronarcys californica and Acroneuria pacifica, were typical of clean unpolluted running water and were also important food organisms of trout. Second, both species were robust creatures, easy to obtain and to maintain in a healthy condition in the laboratory.

INTRODUCTION

In Chapter 1 it was pointed out that many of the striking advances in knowledge over the last 10–15 years regarding pesticide impact have been the outcome of field studies in practical pest control programmes. In all these projects, the environmental studies relate to the known chemical and formulation which is either applied directly to the stream at a predetermined application rate calculated to produce the desired concentration of the chemical in the water, or which contaminates the stream indirectly as a result of aerial application of pesticide to control terrestrial pests in the environs of the stream or river, and where the dosage rate in terms of kilograms per hectare has again been predetermined. In both instances the actual time of application is also known precisely.

The object of these environmental studies is to find out the extent to which the pesticide treatment produces significant changes in the composition of stream fauna produced by mortality or downstream movement, with particular reference to the macroinvertebrates which are the subject of this review. In order to measure these effects and to ascertain significant population changes among the different organisms of running water community – both in the short term and the long term – pesticide ecologists have to rely on a variety of sampling methods. It is these various capture or trapping techniques which provide the essential data for measuring changes in population density or in population composition attributable to pesticide impact.

ARTIFICIAL STREAM COMMUNITIES

The laboratory streams described so far have been designed mainly for single species at a time, or for limited select groups of macroinvertebrates. The logical progress from this point is to design simulated streams in which impact of toxic chemical on a whole community can be studied under conditions akin to those in the natural habitat, but allowing certain factors to be controlled and studied separately in a way that is not feasible in the stream complex. Noteworthy developments along these lines have indeed been made, though not necessarily with the same objective. For example, in studies on the community effect of the lamprey larvicide TFM (see page 214) in Michigan, six fish hatchery channels 8 m long and 0.6 m wide were used, allowing three complete systems, each comprising one control and one adjacent experimental channel (Maki & Johnson, 1977). Each channel was divided into a 4 m upper pool section and a 4 m lower riffle section. The upper pool section was allowed to become colonised by introduction of organic matter and by drift of fauna in the gravity-fed water supply from an adjacent creek. The riffle section was also colonised from natural stream substrates, with associated fauna and flora introduced from a natural source. These communities were allowed to grow and become stabilised for a period of 2 months before experiments started, by which time a very good representation of stream organisms was established, including five species of stonefly (Plecoptera), three of mayflies (Ephemeroptera) and no fewer than nine species of caddis (Trichoptera), as well as the crustaceans Gammarus and Asellus.