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Mesocosm experiments in the approval procedure for pesticides

6 Effect of pesticides in mesocosms

The previous PLS analysis was carried out at a rather high level of taxonomy and organism functionality to satisfy the requirement of data abundance within each group. In effect, detailed information on specific effects of pesticides and differences in sensitivity among different taxonomic groups have not been dealt with. In the following the specific effects (mortality, changes in abundance and sublethal effects) of individual herbicides and insecticides to different taxonomic groups within the major groups (microalgae, zooplankton and macroinvertebrates) are evaluated. In contradiction to the PLS analysis the evaluation has encompassed all mesocosm studies contained in the data base (see Annex B).

6.1 Phytoplankton and microalgae

The insecticide investigations contained in the data base have not demonstrated any directly significant effects such as reduced phytoplankton abundance at the prevalent insecticide concentrations. Therefore, the currently available data do not allow us to determine the maximum permissible insecticide-associated reduction that a phytoplankton population may suffer without becoming extinct. On the contrary it can be concluded that compared with zooplankton and benthic invertebrates, higher concentrations must prevail before a reduction in phytoplankton abundance occurs. This implies that for insecticides the various zooplankton and also invertebrates are affected before the phytoplankton community is directly affected.

A total of 193 records on effects of herbicides on algae (including phytoplankton, epibenthic microalgae and filamentous algae) were distributed between the following end-points LOEC: 83; NOEC: 101 and EC50: 9. We have not attempted to discriminate between phytoplankters and epibenthic algae as their environment (pelagic or benthic) especially in the shallow mesocosms will change rapidly according to mixing conditions. For filamentous algae the number of records was low which excludes a specific analysis. In line with zooplankton and macroinvertebrates structural parameters dominate the effect measures (abundance and biovolume by cell counts, biomass as fresh or dry weight for filamentous algae, chlorophyll a for microalgae). Primary production estimated by oxygen production or 14-C fixation was measured in two experiments (9 records).

In the data base direct effects of herbicides on algae were examined and quantified by relating the dosing of herbicide to changes in abundance relative to corresponding controls (without herbicide dosing). Mesocosm studies with herbicides ranged in duration between 14 and 373 days. Except for one study LOEC’s were recorded during or shortly after termination of herbicide exposure. Hence, most of the data presented below are from this initial period. In the following the relative sensitivity of algae to 3 different herbicides is visualised in diagrams showing LOECs and numeric changes in abundance (Figs. 15-17).

Figure 15.
Summary of effects of Alachlor on abundance of different microalgal species. Tests were carried out in recirculating flumes (175 l) in laboratory dosed at 5 concentrations (1-150 µg l^-1). Samples were taken 5 times during 3 weeks. Position of bars along the concentration axis refer to LOEC for the different groups/species. Numbers shown along bars denote decrease (-) or increase in abundance (in %) of corresponding controls. As a comparison the hazard concentration (HC_5,50%) calculated from distribution based extrapolation of single species toxicity data for Alachlor was 0.73 µg l^-1 (see Table 1).

Figure 16.
Summary of effects of Atrazine on different microalgae. A: recirculating microcosms in laboratory. B: primary production in phytoplankton in large mesocosms (470 m³) dosed once (regular samplings: 39-374 days). C: 1 m³ mesocosms dosed 3 times during 52 days. D: 120 m³ mesocosms dosed twice during 36 days. Biomass was reduced even 280 days after the last application. E: 120 m³ mesocosms dosed twice during 223 days. Periphytes recovered within 7 weeks; phytoplankton still reduced after 7 weeks. F: recirculating microcosms in laboratory. Numbers shown along bars denote decrease in abundance, concentration or primary production (in %) compared to corresponding controls. As a comparison the hazard concentration (HC_5,50%) calculated from distribution based extrapolation of single species toxicity data for Atrazine was 19.9 µg l^-1 (see Table 1).

Figure 17.
Summary of effects of Linuron on microalgae (abundance) and macrophytes (Elodea nuttalii, growth, EC50). Tests were carried out in laboratory mesocosms (600 l) dosed almost continuously through 28 days at 5 concentrations (0.5 – 150 µg l^-1). Numbers shown along bars denote decrease (-) or increase (%) in abundance or growth of corresponding controls. As a comparison the hazard concentration (HC_5,50%) calculated from distribution based extrapolation of single species toxicity data for Linuron was 19.7 µg l^-1 (see Table 1).

On the basis of these comparisons it is evident that:

1. The effect of herbicides on the different algal groups (and macrophytes) is very variable, with both reductions and increases occurring within one systematic group (Fig. 14). For Alachlor increases were observed only at the lowest test concentration (1 µg l^-1).
2. Atrazine as the mostly studied herbicide consistently led to reductions in algal biomass (Fig. 16). Generally, the effects were rather persistent in accordance with the slow dissipation of Atrazine, e.g. in one study primary production was impaired more than one year after application (Fig. 16 B).

In a study with a mixture of Atrazine, Diuron and Alachlor, some of the most sensitive species were Cyanophytae filaments and Monoraphidium sp. demonstrating inhibited growth. Cryptomonas sp., Chlorophyceae coccales, Diatoma sp. (single cell) and Scenedesmus sp. were less adversely affected, while the growth of Chlamydomonas sp. and Stephanodiscus sp. was stimulated. Several species were virtually unaffected by herbicides, e.g. pennate diatoms, Cyanophytae coccales and Anabaena sp.

Whether the phytoplankton can recover after a herbicide-related reduction is difficult to conclude from the mesocosm studies contained in the data base. In one study with Atrazine dosed at 100 µg l^-1, primary production was not fully recovered even one year after the application, while in a comparable study (80 µg l^-1) periphyton biomass and species composition recovered within 49 days. For Alachlor, almost full recovery was attained within 3 weeks for most algal groups except at the highest concentration tested (1000 µg l^-1). However, based on the dynamics of the phytoplankton communities observed in lakes, phytoplankton seem capable of recovering even after a pronounced reduction. It is a well-known phenomenon that some phytoplankton species may disappear from lakes for several years, only to reappear when growing conditions improve.

6.2 Zooplankton

Eighteen studies in the data base have examined the effects of herbicides. None of these studies have demonstrated direct effects on zooplankton.

The data base contains a total of 43 individual mesocosm experiments with zooplankton effect concentrations distributed among 18 different insecticides. A total of 1564 records on effects of insecticides on zooplankton were distributed between the following end-points LOEC: 706; NOEC: 817 and EC50: 14. The majority of records concern Copepods (674 records), Cladocerans (543 records) and Rotifers (279 records), while unspecified zooplankton records amounts to 32. Abundance of individuals is by far the most used effect parameter (1549 records), while species diversity and biomass are scarce at 13 and 2 records, respectively.

In the data base direct effects of insecticides on zooplankton were examined and quantified by relating the dosing of insecticides to changes in abundance relative to corresponding controls (without insecticide dosing). To include results from both single and multiple pesticide application experiments the average decrease in abundance within the period 3-14 days after the first application of insecticide was used. By this approach studies with single and multiple applications could be compared and bias due to different recovery was eliminated. In cases where sufficient data were available EC50 was calculated after probit transformation.

Figure 18.
Temporal variation in abundance (% of control) of zooplankton following a single dose of esfenvalerate in 5 different concentrations ( µg l^-1)to mesocosms (after Lozano et al. 1992 ).

In Fig. 18 is shown an example of the temporal variation in abundance of Cladocera, Copepoda and Rotifera after a single dose of esfenvalerate. Both the initial impact and the subsequent recovery were dependent on the dose. By combining 2-3 different mesocosm studies distinct dose-response curves can be established (see Fig. 19). Noticeable features are dose dependent decreases in Cladocera and Copepoda and increases in phytoplankton and Rotifera.

Figure 19.
Dose-response relations of plankters in mesocosms exposed to esfenvalerate. Mesocosms were shallow (0.5-1.1 m depth), had sediment and macrophytes and ranged between 25 – 1100 m³ in volume (from Fairchild et al. 1992;Lozano et al. 1992, Webber et al. 1992).

Overall, during the first 1-2 weeks after insectide application Cladocerans were more sensitive to insecticides than Copepoda, even though large variation were evident in the data (see Figure 20). Within the copepod population nauplii were more sensitive (» 10%) than adult copepods (see Figure 21). Besides, the variation in the plot was limited reflecting that the two groups probably represent the same species within each experiment.

Figure 20.
Scatter plot of decreases in abundance of Cladocera and Copepoda 3-14 days after insecticide application (Diflubenzuron; Methoxychlor, Hexazinon, Chlorpyrifos, Esfenvalerat, Deltamethrin, Permethrin, Bifenthrin). Regression line and X=Y shown (stippled). Decrease in Cladocera was significantly larger (i.e. Cladocera being more sensitive) than corresponding decrease in Copepoda (Kolgomorov-Smirnov test).

Figure 21.
Scatter plot of decreases in populations of Nauplii and adult Copepoda after insecticide application (Fenvalerat, Methoxychlor, Chlorpyrifos, Esfenvalerat, Deltamethrin, Permethrin, Bifenthrin, Trahalomethrin). Regression line and X=Y shown. Only reductions lower than 100% were included. Decrease in Nauplii was significantly larger (i.e. Nauplii being more sensitive) than decrease in adult Copepoda (Kolgomorov-Smirnov test).

The larval stage of the dipteran Chaoborus is an important pelagic predator in lakes and ponds. In 6 mesocosm studies the abundance of Chaborus was sufficient to calculate the impact of insecticides and compare its sensitivity to Cladocera and Copepoda (Figure 22). In these studies representing different classes of insecticides Chaoborus consistently was more sensitive than Copepoda, but had a sensitivity similar to Cladocera’s.

Figure 22.
Scatter plot of decreases in abundance of Chaoborus and crustacean zooplankton after insecticide application (Chlorpyrifos, Permethrin, Lindan, Methoxychlor). X=Y shown. Only reductions lower than 100% were included. Decrease in Chaoborus was significantly larger (i.e. Chaoborus was more sensitive) than decrease in adult Copepoda () but identical to impact on Cladocera () (Kolgomorov-Smirnov test).

Hitherto, effects have been evaluated at the Order level (e.g. Cladocera) as dictated by the level of taxonomy reported in the majority of mesocosm studies. This invariably will lead to variation in the aggregated data in case of interspecies differences in sensitivity (see Figure 20). Two studies allow extracting quantitative information on variability in sensitivity within Cladocera.

In mesocosms exposed to Azinphos-methyl LOEC ranged between 4 and 20 µg l^-1 while for esfenvalerate the observed range in LOEC was markedly wider at 0.01-5 µg l^-1 (Figure 23). In both studies Sida was the most sensitive genus and Pleuroxus the least sensitive. The difference may be related to size and habitat of species. For comparison the calculated Hazard Concentration of esfenvalerate to Cladocera is 0.18 µg l^-1 and 0.02 µg l^-1 for HC_5,50 and OECD₁₀, respectively (see Table 1). The range in LOEC for different species within Copepoda varied between 0.08 – 5 µg esfenvalerate l^-1.

Figure 23.
Lowest observed effect concentration for different Cladoceran species in mesocosms exposed to Azinphos-methyl (A) and Esfenvalerate (B); (*) all observations were identical, range of LOEC recorded during exposure.

In summary, mesocosm studies have demonstrated that zooplankters are very sensitive to insecticide exposure. At the group level:

1. Cladocerans and Chaoborus are the most sensitive followed by copepod nauplii and adult Copepoda.
2. At a given concentration the Cladoceran population on average will show larger reductions (20 %) than the copepod population (based on regression analysis).
3. Copepod nauplii on average will show 10 % larger reductions than the adult population and observed reductions in one group are very good predictors of the reductions of the other group.

The variation in sensitivity within each zooplankton group as demonstrated in mesocosm studies is considerable. For esfenvalerate LOEC varied 2.5 orders of magnitude for the different species among cladocerans. This variation is probably related to the size of the different species, their habitat and/or feeding mode.

6.2.1 Statistical power of impact of insecticides on zooplankton in mesocosm experiments

Overall, the statistical power in the mesocosm studies was rather low. The average reduction in abundance of zooplankters (i.e. excluding indirect effects) exposed to insecticides at recorded LOEC’s was 75.4 % (± 21.3 %; SD). The low power is due to low number of replicates, low number of and/or large range in test concentrations. The use of few test concentrations spanning 2-3 orders of magnitude invariably will lead to crude estimates of LOEC.

In Table 17 is shown the distribution of reductions in abundance of zooplankton at the various combinations of replicate number and number of test concentrations applied in the different studies. The different combinations are based on observations ranging from 6 to 124 in number and from 1 to 4 different studies carried out at different locations and using different mesocosms (volume, ± macrophytes etc.). Hence, conclusions drawn should not be too firm. Still, the data suggest that in order to obtain a sufficient resolution and sensitivity the experimental design should be a hybrid approach encompassing more than 4 test concentrations and at least two replicates at each concentration. As the size of experimental design usually is constrained by economy with a maximum number of units of 15-16 (see Table 17) based on the results shown in Table 17 they should be distributed between 5 (8) test concentrations each with 2 (3) replicates. Still, to achieve a sufficient sensitivity the range in concentrations applied should not be unduly large, i.e. less than 2.5-3 orders of magnitude.

Table 17.
Average reduction (%)± SD in zooplankton abundance at LOEC in mesocosm studies of different experimental design. Number of observations in brackets. – # observations below 5.

* 2 replicates in control and one replicate per test concentration.

6.2.2 Recovery of zooplankton populations

Recovery of zooplankton populations following insecticide exposure relies on reproduction from surviving individuals, hatching of resting stages (eggs) or immigrations from outside of the system. For the dipteran Chaoborus recovery may also take place by egg laying from imagos. Whether the zooplankton community may endure a 100% reduction depends on whether the resting stages of the various zooplankton groups are tolerant to pesticides, which remains to be elucidated.

To be able to examine recovery of zooplankters, mesocosms need to include sediment and in addition to be in operation for several weeks-months after pesticide dosing has stopped. However, in more than 50 % of the mesocosm studies where zooplankton were followed the post exposure period was too short and/or the doses of insecticides too high to observe complete recovery of zooplankton.

Based on the recovery studies, an attempt can be made at defining the lowest level to which zooplankton populations may be reduced as a consequence of pesticides without being at risk of extinction. For Cladocera the time elapsed for full recovery after the insecticide dosage varied between 10 and 120 days. In Figure 23 is shown a plot of the initial (and maximum) reduction in population size (relative to corresponding control) and the time elapsed after dosage had stopped until full recovery of the population. In mesocosm experiments where cladocerans had been reduced severely (i.e. > 95 %) it took more than 12-15 weeks for full recovery. Such lengthy recovery is probably the result of slow dissipation of the insecticide in the mesocosm and thus continued toxic effects after dosage was stopped.

Figure 24.
Scatterplot between initial reduction in abundance of Cladocera and time elapsed for full recovery of the population. The relation can be described by: R = 8.5 e^0.019x, r²=0.6, where 8.5 (y-axis intercept) indicate the generation time for non-affected populations. Only reductions below 100 % were included.

The relation between the initial reduction in population size and time until full recovery was met could be described by an exponential function (see Figure 24). At reductions below 80 % of the initial population size recovery was fast, i.e. less than 20 days. However, recovery time increased markedly if the initial population was reduced by more then 85 %. Still, even at population reductions close to 100 % full recovery of Cladocerans was observed in the mesocosms where the length of observation period was sufficient long.

For copepods an almost identical relation between initial decrease and recovery was obtained (see Figure 25). Fast recovery within Cladocera (usually analysed at Order level) as observed in numerous studies is likely to be governed by parthenogenetic reproduction. However, to maintain populations of cladoceran species sexual reproduction is essential at intervals. Therefore, recovery studies terminated successfully within 3-4 months may not be sufficient to describe the recovery on the long term.

Figure 25.
Scatterplot between initial reduction in abundance of Copepoda and time elapsed before full recovery of the population. Curve fitted by eye.

Without being at risk of extinction, a significant reduction of cladoceran and copepod numbers may, however, result in reduced species diversity and thus a decline of environmental quality. It may also be that a less diverse community/ecosystem is more sensitive to sudden outside influences such as increased nutrient input or additional pesticide inputs during the recovery phase. Unfortunately, the data contained in the data base do not allow examination of such relations.

6.3 Indirect effects of insecticides on plankton.

The most prominent indirect effect of insecticides on the plankton community includes increases in phytoplankton and rotifers. Following a decrease in population size of crustacean zooplankton, phytoplankton biomass generally will increase due to relaxation of grazing control. In addition, planktonic rotifers that are less sensitive to insecticides will increase in abundance due to increased food availability and reduced competition from crustacean zooplankton (see Figs. 18&19).

In Figure 26 is shown that the phytoplankton biomass increases, when the crustacean zooplankton becomes affected by insecticides. As expected being an indirect effect the scatter is substantial, however, the relation is highly significant. It seems that low impacts on the crustacean zooplankton will not result in increased growth of phytoplankton. If however, zooplankton becomes reduced by more than 50 % dramatic increases in phytoplankton (>100 %) must be expected (Fig. 26).

Figure 26.
Decrease in crustacean zooplankton (Copepoda & Cladocera) and corresponding change (increase) in phytoplankton biomass (Chla) in mesocosm experiments with insecticides (diflubenzuron, endosulfan, deltamethrin, esfenvalerate).

Planktonic rotifers constitute direct competitors to cladocerans and copepods. Reductions caused by insecticides in these groups generally will lead to increases within Rotifera. Due to high reproductive potential increases in abundance up to 3000 % have been observed. In Figure 27 is shown a scatter-plot of changes in crustacean zooplankton and corresponding observations in rotifer abundance in mesocosm experiments with insecticides. Note that the increase in rotifer abundance has been scaled to 100 % within each experiment. On average the decrease in crustacean zooplankton only explains about 20 % of the observed variation in rotifer abundance. Still, the inverse relation is highly significant. Despite increases in rotifer abundance the pelagic grazing control in insecticide affected systems become impaired and phytoplankton biomass will increase (Fig. 26).

Figure 27.
Decrease in crustacean zooplankton (Copepoda & Cladocera) and corresponding change (increase) in Rotifer abundance in mesocosm experiments with insecticides (methoxychlor, esfenvalerate, fenvalerate, cyfluthrin). Within each experiment the increase in Rotifera has been normalised to 100 %.

Table 18 shows an overview of recorded effects on plankton communities with 12 different insecticides in 19 different mesocosm studies. While direct effects on Cladocera and Copepoda are very consistent, indirect effects on Rotifera are more variable. In addition, it is striking that changes in phytoplankton biomass were observed in only 5 out of 19 mesocosm studies. Presence of non-eatable phytoplankters may be responsible

Table 18
Overview of direct og indirect effects of insecticides in mesocosm experiments. ß = significant and consistent decrease in abundance (direct effect), Þ = no effects, Ý = significant and consistent increase in abundance (indirect effect), Ý ß = both decrease and increase observed , - no records.

In conclusion, indirect effects of insecticides on the plankton community have been recorded in more than 50 % of mesocosm studies. In those studies the indirect effects were at least as sensitive as direct effects, e.g. a 75 % reduction in crustacean zooplankton on average will be followed by a 500 % increase in rotifer abundance and a 200 % increase in phytoplankton biomass (see Figure 26). However, indirect effects are very variable in both magnitude and direction and thus less robust compared to direct effects.

6.4 Effect of Insecticides on Macroinvertebrates

Macroinvertebrates generally are insensitive to herbicides. Hence, in only 3 out of 7 mesocosm experiments involving macroinvertebrates in the data base were effects on the macroinvertebrate community detected. They included reduced emergence of Chironomids due to food limitation (reduction of epibenthic algae due to Atrazine), increased drift in streams (Triclopyr-ester & Hexazinone). However, effect concentrations were above calculated hazard concentrations for these herbicides (see Table 1).

The data base contains a total of 41 individual mesocosm experiments with macroinvertebrates distributed among 19 different insecticides. A total of 935 records on effects of insecticides on macroinvertebrates were distributed between the following end-points LOEC: 424, NOEC: 491, EC50: 17 and NEC: 3. Dipterans were used in 29% of the experiments followed by mayflies, Ephemeroptera, which was used in 21% of the experiments. All other macroinvertebrate orders were followed in less than 10% of the experiments. In total, insects constituted 73% of all macroinvertebrates sampled and non-insects 27%.

The majority of recorded effect concentrations concern Dipteran (316 records) with the majority belonging to the family Chironomidae, Ephemeroptera (131 records), Amphipoda (93 records), Isopoda (75 records), Tricoptera (67 records), Hemioptera (58 records), Gastropoda (53 records), Coleptera (50 records), Oligochaeta (32 records), Odonata (29 records), Plecoptera (10 records) and Lepidoptera (4 records).
Abundance of individuals was by far the most used effect parameter (872 records), followed by drift (26 records), mortality (17 records), emergence (13 records) and survival (4 records).

In the data base direct effects of insecticides on macroinvertebrates were examined and quantified by relating the dosing of insecticides to changes in abundance relative to corresponding controls (without insecticide dosing). To be able to compare studies with different application schemes the average decrease in abundance within the period 28-56 days after the first application of insecticide was used. By this approach studies with single and multiple applications could be compared.

In Fig. 28 is shown an example of the temporal variation in abundance of Amphipoda, Chironomidae and Oligochaeta after a single dose of esfenvalerate. Both the initial impact and the subsequent recovery (Chironomidae and Oligochaeta only) were dependent on the dose.

Figure 28.
Temporal variation in abundance (% of control) of macroinvertebrates following a single dose of esfenvalerate in 5 different concentrations to mesocosms (after Lozano et al. 1992 ).

The sensitivity of different macroinvertebrate groups and effect parameters to insecticide exposure in mesocosms were evaluated by comparing corresponding LOECs and numeric reductions. In Figure 29 is shown an example of reductions in abundance of various macroinvertebrate groups exposed to esfenvalerate along with effect on the emergence of insects. These studies demonstrate the general pattern among macroinvertebrates: amphipods and mayflies being rather sensitive to insecticides, while gastropods, Odonata and oligochaetes are rather insensitive.

Figure 29.
Dose-response relations of macroinvertebrates in mesocosms exposed to esfenvalerate. Mesocosms were shallow (0.5-1.1 m depth), had sediment and macrophytes and ranged between 25 – 1100 m³ in volume (from Fairchild et al. 1992; Lozano et al. 1992, Webber et al. 1992).

In the following the relative sensitivity of macroinvertebrate groups to individual insecticides is visualised in diagrams showing LOECs and numeric reductions of macroinvertebrate abundance or alternative endpoints such as increase in drift in artificial streams. Only experiments with more than one group or two or more endpoints followed within one group are presented and discussed. Because of differences in mesocosm volume, season and latitude that all influence the measured toxicity of insecticides (see chapter 5) comparisons can only be evaluated within single mesocosm experiments.

On the basis of these comparisons it is evident that:

1. The sublethal effect drift in stream macroinvertebrates generally appears to be a more sensitive endpoint than changes in abundance (Figure 30AB). In stream ecosystems drift is a natural behaviour of crustaceans and insect larvae for dispersal and colonisation of substrate. When exposed to insecticides (and several other toxic substances) arthropod macroinvertebrates may leave the substrate and drift to avoid the toxicant. Hence, in the short term, drift and population size are reciprocal measures: increased drift invariably will lead to reduced abundance. The seemingly higher sensitivity of drift compared to abundance presumably is related to differences in sample size and stronger statistics in drift data.
2. The endpoint emergence of adult insects seems to be as sensitive as changes in abundance of larvae (e.g. Figure 30). Insecticides may increase the mortality of larvae and reduce growth rate. In effect, emergence will decrease or be delayed. Rate of emergence usually is assessed using float traps that integrate samples from a fairly large bottom area and therefore show less spatial variability than benthos samples. On the other hand, the timing of emergence in affected populations of insect larvae often will differ from the emergence in non-affected populations (i.e. controls) which may complicate sampling and interpretation.
3. The insect order Tricoptera consistently was the most sensitive macroinvertebrate group to insecticides (Figs. 32-34), followed by Plecoptera/Hemiptera/Ephemeroptera/Cole-optera/Amfipoda/Isopoda (no particular order). Chironomidae as a very diverse group (individual size, mode of feeding etc.) showed a rather large variation in sensitivity (e.g. Fig. 31).
4. Odonata, Gastropoda and Oligochaeta consistently were the groups with the lowest sensitivity to insecticides.

Figure 30.
Summary of effects of Lindane on drift, insect emergence and abundance of different macroinvertebrate groups. Experiment A&B (artificial streams) received Lindane continuously for 4 weeks, while in experiment C (1000 l stagnant mesocosm) Lindane was dosed only once. Numbers shown along bars denote the increase in drift (in percentage) or decrease in emergence or abundance of corresponding controls. As a comparison the hazard concentration (HC_5,50%) calculated from distribution based extrapolation of single species toxicity data for Lindane was 2.9 µg l^-1 (see Table 1).

Figure 31.
Summary of effects of Chlorpyrifos on macroinvertebrate groups in laboratory mesocosms (A & B), experimental ditches (C) and in artificial streams (D). In all 4 experiments reduction in abundance was used as end-point. Numbers shown along bars denote the reduction in percentage of corresponding controls. Experiment A-C received Chlorpyrifos as a single dose, while in experiment D Chlorpyrifos was dosed continuously for 21 days. The different colours denote different species within one group. As a comparison the hazard concentration (HC_5,50%) calculated from distribution based extrapolation of single species toxicity data for Chlorpyrifos was 0.04 µg l^-1 (see Table 1).

Figure 32.
Summary of effects of Lambda-cyhalothrin on abundance of different macroinvertebrates in mesocosms. In experiment A and C (25 m³) Lambda-cyhalothrin was dosed 4 times during 42 days, while B (450 m³) was dosed every 14 days through 147 days. Numbers shown along bars denote the reduction in percentage of corresponding controls. The different colours denote different species within one group. As a comparison the hazard concentration (HC_5,50%) calculated from distribution based extrapolation of single species toxicity data for Lambda-cyhalotrin was 80 ng l^-1 (see Table 1).

Figure 33.
Summary of effects of Diazinone on abundance of different macroinvertebrates in mesocosms. Numbers shown along bars denote the reduction in percentage of corresponding controls. As a comparison the hazard concentration (HC_5,50%) calculated from distribution based extrapolation of single species toxicity data for Diazinon was 0.03 µg l^-1 (see Table 1).

Figure 34.
Summary of effects of Fenvalerate on abundance of different macroinvertebrates in mesocosms. In experiment A (small recirculating flume) Fenvalerate was dosed once times and abundance was monitored after 30 days, while B was followed through 84 days. Numbers shown along bars denote the reduction in percentage of corresponding controls. The different colours denote different species within one group. As a comparison the hazard concentration (HC_5,50%) calculated from distribution based extrapolation of single species toxicity data for Fenvalerate was 50 ng l^-1 (see Table 1).

Figure 35.
Summary of effects of Esfenvalerate on abundance of different macroinvertebrates in 1100 m³ mesocosms. Esfenvalerate was dosed every week through 10 weeks. Only LOECs, but no numeric reductions were given in the report. The different colours denote different species within one group. As a comparison the hazard concentration (HC_5,50%) calculated from distribution based extrapolation of single species toxicity data for Esfenvalerate was 180 ng l^-1 (see Table 1).

6.4.1 Statistical power of impact of insecticides on macroinvertebrates in mesocosm experiments

The statistical power for effects on macroinvertebrates was comparable to the impact on zooplankton with an average reduction in abundance (i.e. excluding indirect effects) at recorded LOEC’s of 76.5 % (± 20.3 %; SD). In Table 19 is shown the distribution of reductions in abundance of macroinvertebrates at the various combinations of replicate number and number of test concentrations applied in the different studies. The different combinations are based on observations ranging from 7 to 103 in number and from 1 to 4 different studies carried out at different locations and using different mesocosms (volume, ± macrophytes etc.), which may set limits to conclusions drawn.

Overall, the general pattern resembles the data for zooplankton suggesting that a sufficient sensitivity may be obtained by a hybrid approach with more than 4 test concentrations and at least two but preferably 3 replicates at each concentration.

Table 19.
Average reduction (%)± SD in macroinvertebrate abundance at LOEC in mesocosm studies of different experimental design. Number of observations in brackets. – # observations below 5.

6.4.2 Recovery of macroinvertebrates after insecticide exposure.

Recovery is essential when evaluating effects of pesticides. In macroinvertebrates recovery may take place by invasion from non-affected populations outside the affected area (e.g. by drift in streams, reproduction) and reproduction by surviving individuals within the affected area. In order to evaluate recovery, mesocosm studies need to be carried out in the field (to allow flying insects to lay eggs) and should at the minimum extend a full life cycle of the organisms studied after insecticide dosage. And obviously, repeated sampling of macroinvertebrates will be necessary to follow changes in populations. In the data base not all of the mesocosm experiments included a time series. Furthermore, the majority of experiments in the data base were terminated within 150 days although a few experiments ran for a whole year.

Taking the general life-cycle length for macroinvertebrates into consideration (ranging from less than a month to several years), the experimental time frames in most mesocosm studies appear to be too short. This might partly explain why there are only very few examples of recovery in mesocosms contained in the data base, none of them being a full recovery (Fig. 36). There was no sign of recovery in 81 % of the observations. Signs of recovery were found in 13% of the observations and a moderate recovery in 6% only.

Chironomids (belonging to the order Diptera) and Isopoda were the most important taxonomic groups in the "slight recovery group" (Fig. 37 left) whereas Chironomids and Ephemeropterans dominate the "moderate recovery group" (Fig. 37 right). Both Chironomids and Ephemeropterans in general are considered as good colonisers with short life cycles and this probably explains why they show the most rapid recovery.

Figure 36.
Recovery of macroinvertebrate populations in mesocosm studies contained in data base. An observation includes changes found over time in a macroinvertebrate taxon. No recovery is defined as a less than 5% change (increase) after the initial decrease; slight recovery less than 25% change and moderate between 25 and 75% change.

Figure 37.
Percentage composition of macro-invertebrate orders that showed a slight recovery (left) or moderate recovery (right).

Overall, it is surprising that there is so little evidence of recovery within macroinvertebrates. This finding might reflect that most studies have been too short. One reason for this might be that studies in general involve other taxonomic groups such as zooplankters with shorter life spans and that the duration of the experiments are set to reflect their life span and not the macroinvertebrates. More specific studies targeted towards macroinvertebrates might be needed or the duration of the experiments should be increased when mimicking whole ecosystems. Generally, one should expect that organisms with limited ability for colonising, i.e. non-insect groups such as Isopods, Amphipods and Gastropods or insects with long generation times such as Odonata would be the slowest to recover following insecticide exposure. However, both Gastropoda and Odonata are among the least sensitive to insecticides and the recovery will only be an issue after excessive insecticide exposure. For Amphipods, however, the limited ability to recover can be very critical as these organisms also are among the most sensitive to insecticdes.

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