1. Pesticide toxicity and occurrence of pesticides in substrates of aquatic biotopes;
2. Acute and chronic effects in aquatic organisms;
3. Acute and chronic effects leading to population and community changes or changes in productivity of the ecosystem;
4. Research needs to improve the scientific bases for measurement of effects on the ecosystem.

13.2 PESTICIDES: TOXICITY TO AND OCCURRENCE IN AQUATIC SYSTEMS

1. The water column, which usually first comes in contact with pesticides; 
2. Organic substrates (algae, mosses, vascular hydrophytes, leaf litter, and branches);
3. Inorganic substrates, which include sedimentary material ranging from microscopic silts to coarse sand particles.

Naturally, pesticide content of interstitial water and sediments is usually much lower than that in the water column; and lithic biotopes are mostly less contaminated than standing waters. For instance, permethrin residues after application to a 640 ha forested block in Ontario attained peak concentrations of 147 µg/l in ponds, but only 2.5 µg/l  in streams; while accumulation and persistence of pesticides in bottom sediment was negligible (Kingsbury and Kreutzweiser, 1980).

The determination of pesticide content in water, organic substrates, sediments (and animal tissues as well) is solely chemical. The solid materials or animal/plant tissues are usually homogenized, and 23 or less are extracted (e.g., with acetone/hexane), evaporated to a small volume, cleaned and dried, and analysed by a gas chromatographic procedure.

The effects, including acute toxicity, of pesticide-contaminated water on non-target organisms (NTOs) should be determined using biological methods. The concept of in situ bioassays is based on exposure of test animals at field sites without disturbing contaminated sediment, and the determination of percentage survival. The process of exposing fish to test the toxicity of water is relatively simple: cages containing fish are hung in the water column or anchored at the bottom, and mortality is measured after exposure for 96 h or longer.

The use of invertebrates to test undisturbed sediments, organic matter, or water presents additional problems such as predation and recovery of test animals. Only planktonic organisms (e.g., Daphnia magna) can be used to test water, since benthic organisms not exposed to the test substrate increase respiratory and metabolic rates by several fold, and results can be easily overestimated (Hynes, 1970). Nebecker et al. (1984) recommends testing sediments with a 9 x 20 cm cylindrical stainless steel screen (1.5 mm mesh) enclosed in petri dishes. Twenty adult Hyalella, juvenile Gammarus, or Hexagenia larvae (only 10 larger specimens) should be used in each case. One quarter of the cage is gently forced lengthwise into the sediment and anchored by stakes. Animals are exposed for 96 h using at least two replicate cages, then gently washed free of sediment and evaluated.

For an acute-lethal in situ mortality test, the crayfish (Orconectes virilis) can also be used at the same moult stage (Leonhard,1974). A suitable cage size for six adult crayfish is 20 x 15 x 10 cm. One cage is removed from each site at each sample period (e.g., 2, 4, 8, 16, 24 h exposure on day 1, and one sample per day for the following 30 days). The number of deaths is recorded; weights, lengths, and other characteristics are noted. Death is defined as complete immobility with no flexion of the abdomen upon forced extension. Animals are frozen for residue analysis to identify bioaccumulation at successive exposure intervals. 

Some laboratory bioassays are unable to test pesticide-contaminated water or sediments. These include the liquid phase elutriate test and the solid phase sediment and water beaker test (Nebecker et al., 1984). The solid phase sediment and water test can be modified according to Prater and Anderson (1977), who use a recirculating system containing both Daphnia (about 1 day old) and Hexagenia (20 and 5 per replicate, respectively). It seems to be more time-consuming to construct, calibrate, and use the beaker test; yet the results are comparable (Malueg et al., 1983). By contrast, a similar device can be used for acute toxicity testing of pesticides on any invertebrates (see the tests for lampricide residues with Hexagenia bilineata in artificial substrates in Fremling (1975).

13.2.1 METHODS TO DETECT ACUTE AND DELAYED TOXICITY AMONG INDIVIDUALS

Of about a million species, only several invertebrate organisms have been selected to test adverse effect of toxicants: predominantly, the snail Helisoma trivolvis, the clams Sphaerium, Corbicula, and Musculium, the amphipods Hyalella azteca, Gammarus lacustris and G. pulex fossarum, the burrowing mayfly larvae Hexagenia rigida and H. bilineata, and larvae of the midge Chironomus tentans. Rarely, marine amphipods or freshwater benthic oligochaetes are used (Swartz et al., 1982; Bailey and Liu, 1980).

The major criterion to select a test organism is the relative ease of mass rearing in the laboratory; others include small size, short life-span, maintenance and testing, and uniformity of organisms (Lawrence, 1981).

About 25 fish species are currently used to test waterborne pesticides. Among them, the common carp (Cyprinus carpio), the rainbow trout (Salmo gairdnerui), the brown trout (Salmo trutta), the goldfish (Carassius auratus), the zebrafish (Brachydanio rerio), and the bluegill (Lepomis macrochirus) represent the most often used test vertebrate organisms. Nevertheless, any invertebrate animal or any fish can be used for chronic tests including those not necessarily advantageous for acute tests. For instance, the freshwater clam Corbicula fluminea has been found to be very useful for long-term field monitoring studies, but is of little or no value in acute studies, since it often closes its shell when exposed to high pulses of toxicants (Nebecker et al., 1984).

Doseresponse relationships are routinely examined in laboratory procedures for acute toxicity of aquatic invertebrates (Muirhead-Thomson,1973). However, the substrate and other abiotic conditions should be similar to natural ones. Pelagic or planktonic organisms are simply tested in natural water. To test benthic organisms, the apparatus suggested by Prater and Anderson (1977) can be used; whereas, that of Rodrigues and Kaushik (1984) is helpful to test organisms living on substrates in the stream (e.g., black flies), and requires that the larvae of test invertebrates be fed and allowed to acclimate for a minimum of 16 h prior to treatment.

Acute toxicity results are usually expressed as LC₅₀/48 h and LC₅/48 h (Mayer and Ellersieck, 1986). In fish, the mechanism of toxic effect is the inhibition of some hydrolytic enzymes, mostly of acetylcholine hydrolase. Pesticides produce inhibition that varies from 40 per cent with fenitrothion to 78 per cent with dichlorvos. The major bioassays used to measure acute or delayed toxic effects of pesticides on aquatic organisms are as follows. 

1. Egg test of the burrowing mayfly Hexagenia. Eggs are obtained from live females, placed in petri dishes (300 to 500 eggs for the standard size of 9 cm diameter), and tested at 20 °C for hatch rate (cumulative percentage hatch per treatment) and initiation of hatching. Eggs should be taken from numerous females to prevent use of unfertilized eggs. Eggs can be stored at 8 °C for up to 10 months and returned to 20 °C without considerably affecting hatch parameters (Friesen, 1979). With this test, the effects of methoxychlor ranged from partial suppression of hatching to a delay in hatching to total suppression of embryogenesis with the incipient LC₅₀ estimated to be less than 0.06 mg/1.

2. Chironomus adults emergence test used to test for pesticides in sediments. A 2 or 3 cm deep layer of sediment is placed in 201 containers overlaid with 15 cm of aerated water. The test starts with 10-day-old (20 °C) second instar larvae; the endpoint is the number of emerged adults collected in an erlenmayer flask trap. One hundred larvae fed on a mixture of 600 mg caerophyl and 100 mg tetralin are always used in replicate (Nebecker et al., 1984).

3. Daphnia magna life-cycle test. This test exposes 5-day old Daphnia for 10 days through maturation and release of young (three broods), at which time adults and young specimens are counted. Experimental animals are fed at the rate of 2 mg/1 of algae such as Selenastrum every other day. For water or sediment toxicity, total number of survivors are reported.

4. Measurement of mortality, growth, and fecundity of successive generations of the ramshorn snail, Helisoma trivolvis. Snail are cultured according to standard conditions and exposed to various concentrations (Flannagan, 1974). If mortality occurs during the egg-laying period, fecundity can be adjusted by multiplying the number of offspring produced by the standard number of days of egg laying divided by the actual number of days of egg laying (Flannagan and Cobb, 1979).

5. Sublethal crayfish Orconectes virilis test on intermolt phase and haemolymph calcium. This test is based on cyclical changes throughout the moult cycle, with normally higher levels of exoskeletal calcification present during premoult than during intermoult. Absence of elevation of calcium concentration in specimens exposed to pesticides in the laboratory defines failure to moult. Hemolymph samples are taken directly from the heart and analysed for calcium spectrophotometrically (Leonhard, 1974).

6. Testing embroyonic and larval life stages of fish. Acute-lethal tests are based on direct mortality; however, the most common effects of pesticides are changes in rates of respiration, fin and body movements, heartbeat, and response to light or touch. Chronic sublethal responses can be monitored in long-term toxicity tests. Rates of development, growth, yolk utilization, chorion strength, the occurrence of abnormalities, the timing and duration of hatch, and the initiation of larval feeding behaviour are all indicators of toxicity. For instance, triazine pesticides cause accumulation of an exudate in the body cavity and sometimes even ruptures of the body wall. The most common test is performed in petri dishes using eggs of rainbow trout; eggs and larvae of Pimephales promelas, Salvelinus fontinalis, Schizostedion vitreium, and Brachydanio rerio can be used as well (McKim, 1977; McDonald, 1979).

7. Residual oxygen test to estimate acute-lethal toxicity in fish. Using Salmo gairdneri, a 25g trout contained in a 300 ml vessel is tested with five replicated pesticide concentrations at 1015 ^°C. The time to death and the residual oxygen concentration at death are recorded. Residual oxygen concentrations, as mg O₂/1, are plotted against the logarithm of pesticide concentration, and the relationships are examined for a well-defined inflection point (Giles and Klaprat, 1979).

8. Test on cardiovascular and respiratory functions in fish. A surgical procedure is used to implant a buccal or opercular catheter to measure cough frequency, ventilation rate, and changes in pressure. Generally, an anaesthetized fish weighing from 0.32.0 kg is used. For electrocardiogram data, needle electrodes are inserted under the skin slightly anterior to the pectoral fins (recording electrode) and midventrally to the' pelvic fins (reference electrode). An 
   alternative method has also been suggested (Majewski and Klaverkamp, 1975). Other `physiological or ethological' tests include optomotor response or responsiveness to overhead light stimulation.

9. Fish serum and brain cholinesterase bioassays. Dissected brain or samples of serum are analysed by laboratory procedures using acetylcholine as the substrate. The supernatant of brain extract from 10 µl of serum is examined for absorbance changes at 406 nm against air every 30 s for 3 min. The enzyme activity is related to the slope of the absorbance change. Other changes may occur in serum glucose, serum protein, or total serum lipid (Lockhart et al., 1951). Since most organophosphate pesticides are known to inhibit cholinesterase, this method seems to be most efficient. It was applied to monitor field applications of the organophosphates fenitrothion and malathion and the carbamate propoxur.

Tests with fish for chronic toxicity and bioaccumulation are somewhat difficult, because of the relatively long life cycles (e.g., 23 years in the carp). A possible exception is the zebrafish (Brachydanio rerio) because this small, egg-laying, tropical cyprinid has a life cycle of approximately 75 days. The fish are placed in chambers that allow the eggs to drop to the bottom, and eliminates their predation. The eggs are collected by scraping a slanted divider and siphoning them out (Lillie et al., 1979).

Morphological and pathological effects caused by pesticides are known to be non-specific. They are practically unknown in invertebrates, where generally only the nervous system shows morphological changes. For instance, treatment of mayfly larvae with nematocides at various concentrations led to a brownish precipitate in numerous cells of the sub-oesophageal ganglion. The deposit was intensely concentrated in a well-localized connective tissue, and sporadically it was found in the adjoining commissures and ganglia (Gysels, 1975).

Histopathological effects of pesticides in fishes have been studied intensively. Pathological changes occur mainly in the liver, blood vessels, kidneys, and gills. Liver cells exhibit cytoplasmic granularity, partial loss of liver plate radial orientation, and shrinkage of some liver cell mass. Glomeruli in the posterior kidney show pycnotic changes of cell nuclei, vocalization of cytoplasm, and atrophy of some cells. Gill filaments and lamellae show the precipitated masses that have plugged the central capillaries. The pathological changes of large blood vessels caused by methoxychlor are described in bluegill by Kennedy et al. (1970). Also changes in haematocrit levels and in morphology and quantity of blood cells have been found in various fish species. Boyd (1964) found that sublethal amounts of several chlorine insecticides induced abortions. The highest levels of DDT (2.0 mg/kg per week for 156 days) produced more mature ova than the untreated fish; also mortality among sac-fry was higher when one of the gametes came from a treated parent than when both gametes came from untreated fish. The reproduction of some other species (e.g, cutthroat goldfish) seems to be unaffected by DDT treatment (Allison et al., 1964).

13.3 BIOCONCENTRATION AND BIOINDICATION OF PESTICIDES

The bioconcentration tests are of principal importance, second only to those for determining chronic toxicity in the environment. A large number of bioconcentration assays have been described.

The Chironomus larval survival and growth test follows the same procedure used for the adult emergence test, except that exposure is for 15 days. The larvae are then placed in water overnight to clear the gut; then they are killed with warm water, dried, weighed, measured, and frozen for later tissue analysis. Other tests include the Hyalella partial life cycle test (for 28 days) with Gammarus or Hexagenia (Arthur, 1980; Fremling and Mauck, 1980). In larger animals, limited numbers of tissues are used. For example, at the conclusion of the Orconectes virilis sublethal test, only the following tissues are extracted; gills, liver, stomach, intestine, gonad, heart, tail muscle, haemolymph, and carapace (Leonhard, 1974).

Kinetics (i.e., rates of accumulation and elimination) of pesticides are determined using ¹⁴C-labelled insecticides, followed by deputation of radioactivity after transfer to clean water for 96h. A two-component curve consisting of a rapid initial uptake rate during the first 2 h followed by a reduced linear rate for the remainder of the 24 h exposure is characteristic for 20 plant and animal species in the case of the lampricide uptake (Figure 13.1). Rates of uptake ranged from 0.36 µg/g/h for the crayfish Orconectes propinquus to 17.9 µg/g/h for the caddis fly larvae Brachycentrus americanus (Maki and Johnson, 1977). Rates of loss (Figure 13.2) and half-life figures varied from 7.2 h for the crayfish to 5295 for annelid worms. Continued accumulation of labelled residues from the organic matter is suggested for some pool species in the case of fenitrothion. Although the level of an initial 64 µg/l had declined to traces of 0.1 µg/l in water a few hours after treatment, it peaked in the whole crayfish Orconectes virilis at 1.37 µg/l at 19 days after application, and persisted at lower levels for 30 days (Leonhard, 1974).

Figure 13.1. Rate of uptake of ¹⁴C-labelled lampricide by the amphipod Gammarus pseudolimnaeus during 24 hour exposure to 9 mg/l (Maki and Johnson, 1977)

The bioconcentration factor (BCF) is defined by solving the uptake equation for total body residue after 24 h exposure, and dividing this value by the mean initial pesticide water concentration (Maki and Johnson, 1977; Muir et al., 1975). For instance, the BCF of some synthetic pyrethroids ranged in invertebrates from 135 for trans-permethrin to 316 for -methrin, but they were two- or fourfold less than those observed in oysters (Schimmel et al., 1983) or carp (Ohkawa et al., 1980).

In contrast to that for organophosphates and pyrethroids, the bioaccumulation of DDT and PCBs is a very aggressive and long-term process. Detritus-feeder organisms especially are exposed to a pronounced pesticide stress. For instance, mayfly larvae (Ephemera danica) accumulate DDT and PCB (Figure 13.3) to a steady state of 310 µg/l and 130 µg/g, respectively, according to a first-order kinetic equation (Södergren and Svensson, 1973). The same species showed DDE/DDT accumulation greater than that of detritus, Plecoptera, Ancylus, Gammarus, Trichoptera, Tipula, Baetis, or Salmo (Södergren et al., 1972). The bioaccumulation and magnification of pesticides are very efficient, especially in the food chain of detritivore invertebrate fish. The bioconcentration factor in fish may even reach levels of 10⁴ (Macek and Korn, 1970). Tolerances for DDT and PCBs in food range in most countries from 0.2 p.p.m. in baby food to 5.0 p.p.m. in fish. In the past, at the time of large-scale use of DDT, these tolerances were exceeded (Mauck and Olson, 1977).

Figure 13.2. Rate of elimination of ¹⁴C-labelled lampricide by the amphipod Gammarus pseudolimnaeus following 24 hour exposure to 0.9 mg/l (Maki and Johnson, 1977)

Figure 13.3. Accumulation of DDT and PCBs by sediment-dwelling larvae of the mayfly Ephemera danica (Sodergren and Swenson, 1973)

The study of distribution of pesticides in different aquatic biotopes is urgently needed. Invertebrates as primary consumers and relatively short-lived organisms seem to be more useful than top consumers (fish), which accumulate pesticide for a long time. For instance, Södergren et al. (1972) successfully used an amphipod, Gammarus pulex, to indicate the levels of DDT and PCB in southern Swedish streams. Residue levels (less than 5 p.p.m. of lipid weight) were generally low in February, when streams were fed by underground water and the ground was frozen and snow-covered. They were very high (more than 35 p.p.m. of lipid weight) in April, associated with peak flow volumes caused by melting snow sweeping organic and inorganic material into streams.

When evaluating bioindicators of pesticide levels in the environment, one might keep in mind the reduced resistance and different accumulation rates of young and smaller specimens (Freeden, 1972; Eidt, 1975) and those with higher metabolic rates (Jensen and Gaufin, 1964) compared to others. Food selection and microhabitat may influence the susceptibility to soluble and insoluble pesticides (e.g., methoxychlor and its derivatives); filter-feeding organisms (unwanted blackflies and caddis flies) showed that a large number of non-target, scrapper, and collector feeding category species (Cummins, 1973) were heavily influenced (Flannagan et al., 1980). The populations of animals that live on the upper surface of stones and pelagic organism (most fish) as well as those entering aquatic habitats occasionally are more likely to be exposed to a pesticide during short to moderate exposure than those that dwell deeper in a substrate (Wiederholm, 1984). Moreover, a short pulse of pesticide at high concentration usually exerts a more pronounced effect than a long exposure to that pesticide at low concentration, even though the product of the concentration times the period of expsoure may be identical (Muirhead-Thomson, 1973).

13.4 EFFECTS OF PESTICIDES ON NON-TARGET COMMUNITIES

The successive physiological disturbances induced by pesticides (including hyperactivity, loss of equilibrium, tremors and convulsions) are capable of producing an avoidance mechanism, causing population movements. This mechanism in vertebrates (fish) has been reported frequently (Alabaster and Lloyd, 1980; Baier et al., 1985; Murty, 1986), but is much less known than the drift of aquatic invertebrates. Some species, such as rainbow trout (Salmo gairdneri), tend to remain motionless for several minutes; others, such as whitefish (Coregonus clupeaformis), show consistent swimming patterns. The apparatus to test for avoidance behaviour has been described in detail by Scherer and Nowak (1973). The avoidance of fenitrothion by goldfish provides an immediate read-out of time spent in places with concentration variations as great as 50 per cent (Scherer, 1975). Avoidance chamber responses of invertebrates and fry of rainbow trout to some herbicides (copper sulphate, acrolein, xylene emulsion, Dalapon, and Roundup) have been described by Folmar (1978).

Pesticide-induced downstream drift of invertebrates can be easily sampled with drifting nets (usually 1 x 1 m or smaller, mesh size about 0.140.30 mm). Pesticides may often induce invertebrate drift of catastrophic proportions. Numbers of drifting specimens increase at least several times during the 24 h after application, as in the case of large-scale treatment of forest with Actellic and Ambush in Czechoslovakia (Tonner et al., 1983; Figure 13.4). For instance, Flannagan et al. (1980) found that 50 000 specimens entered a 15 cm diameter sampler during a 4 h period after treatment of the Athabasca River with methoxychlor. An estimated 2.5 billion animals drifted past a particular site during that period of time. Dead specimens reached more than 50 per cent (Figure 13.5; Tonner et al., 1983). Gibbs et al. (1984) described the effects of carbaryl application in Maine. Invertebrate drift increased up to 170 times two days after treatment, and virtually all organisms in the samples were dead. Surface benthic communities (mainly Ephemeroptera, Trichoptera, and Plecoptera) can be eliminated (Wallace and Hynes, 1975).

Figure 13.4. Avoidance of fenitrothion by goldfish Carassius auratus; values are median response, and bars are standard errors (Scherer, 1975)

Benthos recovery is usually accomplished within several months. For fenitrothion applied at 73 µg/l benthos recovered after its total elimination at about 350 m downstream (23 µg/l 50 days after treatment (Eidt, 1981). Organophosphates and carbamates rarely penetrate the sediments deeply, and substrate-dwelling animals remain mostly unaffected (e.g., oligochaetes and Chironomidae). On the other hand, penetration by DDT and PCB is more pronounced, and recovery of benthic invertebrates required at least 4 years in New Brunswick (Ide,1960).

Toxic effects of pesticides also influence more tolerant species of aquatic communities, in that their qualitative presentation and types of competitors, predators, and prey-organisms may expand or contract as niche space is vacated or eliminated. Changing densities are detected by quantitative samplingusually electrofishing or Surber benthometry of invertebrates. The effects are summarized as follows:

1. Top predator elimination usually results in a several-fold increase in density of benthic insects. For example, elimination of fish from a Wisconsin lake by toxaphene was followed by a 200-fold increase in the Chironomus population, which declined again when fish were restocked (Hilsenhoff, 1965).

2. Reduction of invertebrate predators showed similar effects; for example, the number of blackflies (filtrators) increased greatly following the treatment of streams with DDT and the subsequent reduction of predatory Plecoptera and Trichoptera (Hurlbert, 1975).

3. Similarly, massive increases in attached algae in streams have been attributed to the elimination of grazers by pesticides (Yasuno et al., 1982). However, increased densities of benthic invertebrates following some herbicide treatment presumably resulted from the changes in oxygen levels associated with decaying plants; this may also give rise to higher phytoplankton production before the macrophytes begin to reappear. Species closely associated with macrophytes (e.g., some Ephemeroptera, Plecoptera, or Chironomidae) can be reduced greatly when substrate availability and complexity are reduced (Brooker and Edwards, 1975).

4. The removal of competitors causes expansion of species with similar niche requirements. For example, in a malaria eradication programme, pesticide-resistant populations of Anopheles mosquitoes have benefited from the pesticide-caused mortality of species that compete for their preferred prey (Hurlbert, 1975).

5. Non-target species populations can be eliminated as a consequence of effects on some abiotic factors, as in the case of herbicide treatment. The decay of large amounts of plant material can reduce the amount of dissolved oxygen and alter pH conditions, thus affecting many organisms (especially fish).

Figure 13.5. Drifting specimens of the stonefly Protonemura auberti during application of pirifosmethyl (Actelic) and decamethrin (Ambush) (Tonner et al., 1983) 

Data on pesticide effects at the aquatic ecosystem level are very scarce. Thus far, measurements have been made of only the quantity of nutrients and levels of energy. Both primary and secondary production have been measured by hydrobiological methods, mostly in conjunction with methoxychlor or lampricide treatment of running waters. While methoxychlor caused reduction of densities and biomass to less than 10 per cent, the following parameters were unaffected: counts of fungal hyphae, respiration rates of conditioned leaf discs, and adenosine triphosphatase (ATP) content of coarse organic matter. However, ATP levels of fine particulate organic matter were significantly higher in a treated stream representing potentially increased food quality and generation of subsequent transport to downstream reaches (Cuffney et al., 1984). Generally, both secondary production of benthic organisms and the turnover ratio (production +biomass) are decreased by any toxic quantity of a pesticide, mostly as a consequence of density reduction. The production quotient between primary and secondary consumers seem to be very suitable for assessing the effects of pesticides on running water biocenoses. The production quotient is approximately 15 in undisturbed biocenosis and increases with greater disturbances (Frutiger,1985). 

13.5 REFERENCES 

Alabaster, J. S. and Lloyd, R. (1980) Water Quality Criteria for Fresh-Water Fish, Butterworths, London, pp. 297-298.

Allison, D., Kallman, B. J., Cope, O. B. and Van Valin, C. C. (1964) Some Chronic Effects of DDT on Cutthroat Trout, Research Report 64, US Department of Interior, Bureau of Sport Fisheries and Wildlife, Washington, DC, 30pp. 

Arthur, J. W. (1980) Review of freshwater bioassay procedures for selected amphipods. In: Buikema, A. L. and Cairns, J. (Eds) Aquatic Invertebrate Bioassays, ASTM Special Technical Publication 715, American Society for Testing and Materials, Philadelphia, Pennsylvania, pp. 98-108.

Baier, Ch., Hurle, K. and Kirchhoff, J. (1985) Datensammlung zur Abschatzung des Gefahrdungspotentials von PflanzenschutzmiettelWirkstoffen fur Gewasser, Paul Parey, Verlag.

Bailey, H. C. and Liu, D. H. E. (1980) Lumbriculus veriegatus, a benthic oligochaete, as a bioassay organism. In: Eaton, J. C., Parrish, P. R. and Hendrick, A. C. (Eds) Aquatic Toxicology, ASTM Special Technical Publication 707, American Society for Testing and Materials, Philadelphia, Pennsylvania, pp. 205-215.

Boyd, C. E. (1964) Insecticides cause mosquitofish to abort. Prog. Fish-Cult. 26, 138. 

Brooker, M. P. and Edwards, R. W. (1975) Review paper: aquatic herbicides and the control of water weeds. Water Res. 9, 1-15.

Cuffney, T. F., Wallace, J. B. and Webster, J. R. (1984) Pesticide manipulation of a headwater stream: invertebrate response and their significance for ecosystem processes. Freshwater Invertebr. Biol. 3, 153-171.

Cummins, K. W. (1973) Tropic relations of aquatic insects. Ann. Rev. Entomol. 18, 183-206.

Eidt, D. C. (1975) The effect of fenitrothion from large-scale forest spraying on benthos in New Brunswick headwater streams. Can. Entomol. 107, 743-760.

Eidt, D. C. (1981) Recovery of aquatic arthropod populations in a woodland stream after depletion by fenitrothion treatment. Can. Entomol. 113, 303-313.

Flannagan, J. F. (1974) Influence of trisodium nitrilotriacetate on the mortality, growth and fecundity of the freshwater snail (Helisoma trivolvis). Fish. Res. Board Can. 31, 155-161.

Flannagan, J. F. and Cobb, D. G. (1979) The use of the snail Helisoma trivolvis in a multingeneration mortality, growth and fecundity test. In: Scherer, E. (Ed.) Toxicity Tests for Freshwater Organisms, Can. Spec. Publ. Fish. Aquat. Sci. 44, 19-26. 

Flannagan, J. F., Townsend, B. E. and de March, B. G. E. (1980) Acute and long-term effects of methoxychlor larviciding on the aquatic invertebrates of the Athabasca River, Alberta. In: Haufe, W. O. and Croome, G. C. R. (Eds) Control of Black Flies in the Athabasca River, technical report, Alberta Environment, Edmonton, pp. 151-158. 

Folmar, L. C. (1978) Avoidance chamber responses of mayfly nymphs exposed to eight herbicides. Bull. Environ. Contam. Toxicol. 17, 312-318.

Freeden, F. J. H. (1972) Reactions of the larvae of three rheophilic species of Trichoptera to selected insecticides. Can. Entomol. 104, 944-953.

Fremling, C. R. (1975) Acute toxicity of the lampricide 3-trifluoro-methyl-4-nitrophenyl (TFM) to nymphs of mayflies (Hexagenis sp.). Invest. Fish Control 58, 1-6. 

Fremling, C. R. and Mauck, W. L. (1980) Methods for using nymphs of burrowing mayflies (Ephemeroptera, Hexagenia) as toxicity test organisms. In: Buikema, A. L. and Cairns, J. (Eds) Aquatic Invertebrate Bioassays, ASTM Special Technical Publication 715, American Society for Testing and Materials, Philadelphia, pp. 81-97. 

Friesen, M. K. (1979) Toxicity testing using eggs of the burrowing mayfly Hexagenia rigida (Ephemeroptera, Ephemeridae), with methoxychlor as toxicant. Proceedings of the Fifth Annual Aquatics Workshop, Report 862, pp. 266-277.

Frutiger, A. (1985) The production quotient PQ: a new approach for quality determination of slightly to moderately polluted running waters. Arch. Hydrobiol. 104, 513-526.

Gibbs, K. E., Mingo, T. M. and Courtemanch, D. L. (1984) Persistence of carbaryl (Sevin-4-oil) in woodland ponds and its effects on pond macro invertebrates following forest spraying. Can. Entomol. 116, 203-204.

Giles, M. A. and Klaprat, D. (1979) The residual oxygen test: a rapid method for estimating the acute-lethal toxicity of aquatic contaminants. In: Scherer, E. (Ed.) Toxicity Tests for Freshwater Organisms, Can. Spec. Publ. Fish. Aquat. Sci. 44, pp. 37-45.

Gysels, H. (1975) Electrophoretical and histochemical investigations on some noxious Lepidoptera and on the impact of pesticide upon Ephemera danica, a water-dwelling, innocuous mayfly. Acta Zool. Pathol. Antverp. 62, 127-141.

Hilsenhoff, W. L. (1965) The effect of toxaphene on the benthos in a thermally-stratified lake. Trans. Am. Fish. Soc. 94, 210-213.

Hurlbert, S. H. (1975) Secondary effects of pesticides on aquatic ecosystems. Res. Rev. 57, 81-148.

Hynes, H. B. N. (1970) The Ecology of Running Waters, University Press, Toronto, 555pp.

Ide, F. P. (1960) Effects of forest spraying with DDT on aquatic insects, food of salmon and trout, in New Brunswick. Proc. Entomol. Soc. Ontario 91, 39-40.

Jensen, L. A. and Gaufin, A. R. (1964) Long-term effects of organic insecticides on two species of stonefly naiads. Trans. Am. Fish. Soc. 93, 357-363.

Kennedy, H. D., Eller, L. L. and Walsh, D. F. (1970) Chronic Effects of Methoxychlor on Bluegills and Aquatic Invertebrates, Tech. Pap. Bureau Fish. Wildl. 53, pp. 3-18. 

Kingsbury, P. D. and Kreutzweiser, D. P. (1980) Environmental Impact Assessment of a Semi-Operational Permethrin Application, Forestry Service Report FPM-X-30, Environment Canada, Ottawa, pp. 1-47.

Lawrence, S. G. (Ed.) (1981) Manual for the Culture of Selected Freshwater Invertebrates, Can. Spec. Publ. Fish. Aquat. Sci. 54, 169.

Leonhard, S. L. (1974) Uptake of fenitrothion by caged crayfish Orconectes virilis in Pine Creek, Manitoba. Manitoba Entomol. 8, 16-18.

Lillie, W. R., Harrison, S. E., Macdonald, W. A. and Klaverkamp, J. F. (1979) The use of the zebrafish (Brachydanio rerio) in whole-life-cycle tests. In: Scherer, E. (Ed.) Toxicity Tests for Freshwater Organisms, Can. Spec. Publ. Fish. Aquat. Sci. 44, pp.104-111.

Lockhart, W. L., Rosenbrough, N. J., Farr, A. L. and Randall, R. J. (1951) Protein measurement with folin phenol reagent. J. Biol. Chem. 193, 267-275.

Macdonald, W. A. (1979) Testing embryonic and larval life stages of fish. In: Scherer, E. (Ed.) Toxicity Tests for Freshwater Organisms, Can. Spec. Publ. Fish. Aquat. Sci. 44, pp. 131-138.

Macek, K. J. and Korn, S. (1970) Significance of the food chain in DDT accumulation by fish. J. Fish. Res. Board Can. 27, 1496-1498.

Majewski, H. S. and Klaverkamp, J. F. (1975) An acute bioassay for cardiovascular and respiratory functions in rainbow trout. In: Craig, G. R. (Ed.) Proceedings of the Second Annual Aquatic Toxicity Workshop, 1975, Toronto, Ontario, Ontario Ministry of the Environment, Ottawa, pp. 165-178.

Maki, A. W. and Johnson, H. E. (1977) Kinetics of lampricide (TFM, 3-trifluoromethy-l4-nitrophenol) residues in model stream communities. J. Fish. Res. Board Can. 34, 276-281.

Malueg, K. W., Schuytema, G. S., Gakstatter, J. H. and Krawczyk, D. F. (1983) Effects of Hexagenia on Daphnia response in sediment toxicity tests. Environ. Toxicol. Chem. 2, 73-82.

Mayer, F. L. and Ellersieck, M. R. (1986) Manual of Acute Toxicity: Interpretation and Data Base for 410 Chemicals and 66 Species of Freshwater Animals, US Department of Interior, Fish and Wildlife Service, Washington, DC, 582pp.

Mauck, W. L. and Olson, K. E. (1977) Polychlorinated biphenyls in adult mayflies (Hexagenia bilineata) from the upper Mississippi River. Bull. Environ. Contam. Toxicol. 17, 387-390.

McKim, J. M. (1977) Evaluation of test with early life stages of fish for predicting long-term toxicity. J. Fish. Res. Board Can. 34, 1148-1154.

Muir, D. C. G., Rawn, G. P., Townsend, B. E., Lockhart, W. L. and Greenhalgh, R. (1975) Bioconcentration of cypermethrin, deltamethrin, fenvalerate and permethrin by Chironomous tentans larvae in sediment and water. Environ. Toxicol. Chem. 4, 51-61.

Muirhead-Thomson, R. C. (1971) Pesticides and Freshwater Fauna, Academic Press, London, New York, 248pp.

Muirhead-Thomson, R. C. (1973) Laboratory evaluation of pesticide impact on stream invertebrates. Freshwater Biol. 3, 479-498.

Murty, A. S. (1986) Toxicity of Pesticides to Fish, CRC Press, Boca Raton, Florida, 143pp.

Nebecker, A. V., Cairns, M. A., Gakstatter, J. H., Malueg, K. W., Schuytema, G. S. and Krawczyk, D. F. (1984) Biological methods for determining toxicity of contaminated freshwater sediments to invertebrates. Environ. Toxicol. Chem. 3, 617-630.

Ohkawa, H., Kikuchi, R. and Miyamoto, J. (1980) Bioaccumulation and biodegradation of the (S)-acid isomer of fenvalerate (Sumicidin) in an aquatic model ecosystem. J. Pest. Sci. 5, 11-22.

Prater, B. L. and Anderson, M. A. (1977) A 96-hour bioassay of Duluth and Superior harbor basins (Minnesota) using Hexagenia limbata, Asellus communis, Daphnia magna, and Pimephales promelas as test organisms. Bull. Environ. Contam. Toxicol. Chem. 2, 73-82.

Rodrigues, C. S. and Kaushik, N. K. (1984) A bioassay apparatus for the evaluation of black fly (Diptera simuliidae) larvicides. Can. Entomol. 116, 75-78.

Rosenberg, D. M. (1975). Food chain concentration of chlorinated hydrocarbon pesticides in invertebrate communities: a re-evaluation. Quest. Entomol. 11, 97-110. 

Scherer, E. (1975) Avoidance of fenitrothion by goldfish. Bull. Environ. Contam. Toxicol. 13, 492-496.

Scherer, E. and Nowak, S. (1973) Apparatus for recording avoidance movements of fish. J. Fish Res. Board Can. 30, 1594-1596.

Schimmel, S. C., Garnas, R. L., Patrik, J. M. and Moore, J. C. (1983) Acute toxicity, bioconcentration and persistence of AC222, 705, benthiocarb, chloropyrifos, fenvalerate, methyl parathion and permethrin in the estuarine environment. J. Agric. Food Chem. 31, 104-113.

Södergren, A. and Svensson, B. (1973) Uptake and accumulation of DDT and PCB by Ephemera danica (Ephemeroptera) in continuous-flow systems. Bull. Environ. Contam. Toxicol. 9, 345-350.

Södergren, A., Svensson, B. and Ulfstrand, S. (1972) DDT and PCB in south Swedish streams. Environ. Pollut. 3, 25-36.

Swartz, R. C., DeBeb, W. A., Sercu, K. A. and Lamberson, J. O. (1982) Sediment toxicity and the distribution of amphipods in Commencement Bay, Washington, USA. Pollut. Bull. 13, 359-364.

Tatton, J. O'. G. and Ruzicka, J. H. A. (1967) Organochlorine pesticides in Antarctica. Nature (Lond.) 215, 346-348.

Tonner, M., Vavra, N., Syrovatka, O. and Soldan, T. (1983) Einfluss der Luftbespritzung gegen grauen Larchenwickler (Zeiraphera diniana) auf die Entomofauna des Rhithrons in Krkonose (Riesengebirge). Vest. Cesk. Spol. Zool. 47, 293-303.

Wallace, R. R. and Hynes, H. B. N. (1975) The catastrophic drift of stream insects after treatments with methoxychlor (1,1,1-trichloro-2,2-bis-[o-methoxyphenyl] ethane). Environ. Pollut. 8, 255-268.

Wiederhohn, T. (1984) Responses of aquatic insects to environmental pollution. In: Resh, V. S. and Rosenberg, D. M. (Eds) The Ecology of Aquatic Insects, Praeger, New York, pp. 508-557. 

Yasuno, M., Ohkita, J. and Hatakeyma, S. (1982) Effects of temephos on macrobenthos in a stream of Mt Tsukuba. Jpn. J. Ecol. 32, 29-38.