Metabolism, bioaccumulation, and toxicity of pesticides in aquatic insect larvae

Abstract

Aquatic insects having a high diversity are good biotic indicators for freshwater quality. Their larvae living in freshwater are sensitive to pesticides, and its impacts has been examined not only through laboratory toxicity studies using water and sediment exposure but also through higher-tier micro-/mesocosm studies and field monitoring. Many sophisticated statistical methods have been applied to assess the impacts of pesticides at levels from species to community, but their body burden has been studied much less, especially in relation to toxicity. We review the uptake, metabolism with relevant detoxifying enzymes, and depuration of pesticides in aquatic insect larvae, which determine their body burden and help to understand the toxicity profiles specific to each chemical class. We also discuss experimental conditions, environmental factors, and species sensitivity in relation to the bioconcentration/-accumulation and toxicity of pesticides.

Introduction

The lotic and lentic freshwater systems consisting of ditches, streams, rivers, and ponds are abundant in many kinds of aquatic insect larvae, as represented by Diptera (flies and mosquitoes),¹⁾ Ephemeroptera (mayflies),²⁾ Plecoptera (stoneflies),³⁾ and Trichoptera (caddisflies),⁴⁾ in each of which more than thousands of species are known. Their larvae are an important food source for fish, amphibians, birds, and semi-aquatic mammals,^5,6) indicating the possible biomagnification of a pesticide if it is bioaccumulated. They also perform ecosystem functions relating to nutrient retention, litter decomposition, rock cleaning, elemental cycling, and stream-bed stabilization, and therefore they are generally used as biotic indicators to protect the freshwater environment. Ephemeroptera, Plecoptera, and Trichoptera (EPT), which represent the highest taxon richness in freshwaters,⁷⁾ are very sensitive to pesticides.^8,9) Catchment urbanization in Georgia, USA, caused poor water quality in streams due to increased erosion of sediment, nutrients, and pesticide residues, resulting in less taxon diversity, with EPT being the most sensitive orders.¹⁰⁾ By examining the macroinvertebrate communities in six German streams, clean or contaminated with ten pesticides, the canonical correspondence and redundancy analyses showed that the variance in their communities, especially EPT and Diptera, was well described by the summation of toxic units, the pesticide concentration in water normalized to the acute LC₅₀ in Daphnia magna.¹¹⁾ As compared with the Brazilian river neighboring forest and meadow, less abundance of macroinvertebrates with poor taxon richness was reported for the river flowing through the agricultural site applied with many insecticides, and some EPT species disappeared altogether.¹²⁾ By comparing species abundance with the concentrations of imidacloprid at various freshwater sites in the Netherlands, an extensive statistical analysis showed a significant negative relationship between them, with high correlations for the most abundant Glyptotendipes pallens (midge), Caenis horaria (mayfly), and Oecetis lacustris (caddisfly).¹³⁾

The toxicity of a pesticide should basically be interpreted from the standpoint of its body burden and mode of action (MOA) in aquatic insects. However, there are a number of difficulties, such as the limit of residue analysis and little information on toxicokinetics and MOA, such that current observations of toxicological signs, such as immobilization, effects on growth and reproduction, and lethality, are based on nominal exposure concentrations. In addition, the taxonomy of small aquatic insects is complex due to their enormous species number, which makes it difficult to monitor their richness and diversity. An alternative classification using functional feeding groups (FFGs) has been conducted from the standpoint of the dominant food resource and feeding mechanism^6,14): shredders, chewing conditional litter or vascular plant tissues; filtering collectors, feeders of suspended detrital particles, algae, and bacteria; gathering collectors, feeders of sediment and deposited loose particles; scrapers, grazing on the surfaces of rocks, wood, and rooted aquatic plants; and predators, preying on living animals. This classification is also useful when considering whether aquatic insect larvae passively take up the dissolved pesticide from water and/or ingest its fraction adsorbed onto sediment, detritus, and biota as part of their diet.

With the above considerations, the metabolic profiles of pesticides in various aquatic insect larvae are reviewed with the relevant enzymes (oxidases, esterases, and glutathione-S-transferases), which determine their body burden. The bioconcentration and bioaccumulation of pesticides are then summarized in relation to pesticide properties such as hydrophobicity and metabolism, interactions with sediment, and species sensitivity. Finally, toxicity at the species to community levels is discussed in relation to assessment methodologies, taking into account bioavailability, sensitivity, and metabolism. Unless specified, the species cited in this review refers to its larvae. Through this review, issues to be further investigated are proposed in order to gain a better understanding of the toxic effects of pesticides on aquatic insect larvae.

1. Metabolism

1.1. In vivo and in vitro metabolism of pesticides

The metabolic transformation is one of the major factors in controlling the body burden and toxicity of pesticide.¹⁵⁾ However, the metabolism of pesticides in aquatic insect larvae has been much less well examined than that in fish and has been mostly limited to Chironomus spp. (Diptera, midge). The corresponding information has been obtained mainly for organochlorine (OC) and organophosphorus (OP) insecticides via water exposure in relation to pesticide resistance and bioaccumulation. In vivo transformation patterns, summarized in Table 1, are common to other aquatic species. ¹⁵⁾ The dehydrochlorination of DDT to DDE is reported in many species, with higher metabolic activity in Hexagenia bilineata (mayfly) compared to the Chironomus sp.^16,17) A trace amount of DDD was detected in Ephemera danica (mayfly) through flow-through water exposure to DDT.¹⁶⁾ DDD was bioaccumulated in Hexagenia spp. from the DDT-contaminated sediment, but the metabolic contribution was unclear.³⁴⁾ The oxidative transformation of aldrin,^17,19,20) chlordane and heptachlor,²¹⁾ and DDT¹⁷⁾ is also reported in some species. Trans and cis isomers of chlordane were oxidized to oxychlordane in Cloeon triangulifer (mayfly) to a similar extent following long-term exposure from eggs to larvae.²¹⁾ Chironomus spp. metabolize OP pesticides^25–29) via either oxidative desulfurization to the respective oxon or hydrolysis to the corresponding phenols. The oxidative dealkylation^23,32,33) at N, O, and Sn atoms and enzymatic hydrolysis^22–24) of amides and carbamates also took place in these midges. An acute toxicity study of acephate in adult Notonectidae (backswimmer) and Corixidae (water boatman) showed the cleavage of its amide linkage to form methamidophos as the main metabolic pathway.³⁰⁾ The reductive metabolism was only known for the lampricide 4-nitro-3-trifluoromethylphenol (TFM) in Chironomus tentans.³¹⁾ Although the metabolism of ¹⁴C-labeled pentachlorophenol (PCP), fenpropidin, and trifluralin was examined in Chironomus riparius, no conclusive identification of metabolites was achieved by chromatographic analysis and mass spectrometry.³⁵⁾ Type II conjugation reactions have been inferred from the detection of polar unknowns, but direct evidence is rarely reported. The conjugation of primary metabolites from pyridalyl in Chironomus yoshimatsui was confirmed after harsh acid hydrolysis of polar metabolites.³²⁾ The release of TFM by the enzymatic hydrolysis of polar metabolites indicated its conjugation with glucuronic acid and sulfate in C. tentans.³¹⁾

The in vitro metabolism of pesticides has been conveniently examined by using either the whole homogenates of insect larvae or the post-mitochondrial fraction with a protease inhibitor. The epoxidation of aldrin to dieldrin has been found to be common in many aquatic insects as well as in terrestrial ones, but it was not detected in the Ischnura or Enallagma spp. (Odonata, damselflies).³⁶⁾ This reaction proceeded mainly in the gut and fat body of the Limnephilus sp. (caddisfly).¹⁹⁾ The microsomal fraction from the whole homogenate of C. tentans catalyzed either the epoxidation of aldrin³⁷⁾ or the oxidative desulfurizationof chlorpyrifos to its oxon²⁹⁾ in the presence of NADPH. The oxidative degradation of triadimefon in the microsomal fraction of Simulium vittatum (Diptera, black fly) with an NADPH-regenerating system was inhibited by 1-aminobenzotriazole, but the formation of triadimenol was likely to be catalyzed by unknown NADPH-requiring enzymes.³⁸⁾ In the post-mitochondrial fractions from homogenates of several aquatic insects, permethrin was hydrolyzed to form 3-phenoxybenzyl alcohol.³⁶⁾

Table 1. In vivo metabolism of pesticides in the aquatic insect larvae

Pesticide1)	Species2)	Exp. Cond.3)	Metabolic profiles4)	Ref.
DDT†	Ephemera danica (m)	f/w/14/9	DH	16
DDT	Hexagenia bilineata (m)	f/w/21/3	DH	17
	Chironomus sp. (mi)	f/w/21/3	DH	17
	Libellula sp. (dr)	f/w/21/3	DH, O1	17
DDE	Chironomus tentans (4th mi)	s/w/20/8#	stable	18
Aldrin†	Limnephilus sp. (c)	f/w/rt/1	O2	19
	Chironomus riparius (4th mi)	s/w/21/2#	O2	20
Aldrin	Hexagenia bilineata (m)	f/w/21/3	O2	17
Chlordane†	Cloeon triangulifer (egg, m)	s/w/20/43	O2, DC	21
Carbaryl	Chironomus riparius (4th mi)	s/w/10–30/1	H1, U	22
Alachlor	Chironomus spp. (mi)	s/w/rt/<2	OD, ND, H1	23
Niclosamide	Chironomus tentans (4th mi)	s/w/20/4	H1, U	24
Parathion	Chironomus riparius (4th mi)	s/w/10–30/1	O3, U	25
Parathion-methyl	Chironomus tentans (4th mi)	s/w/10–30/4	O3, U	26
Fenitrothion	Chironomus riparius (3rd mi)	s/w/26/5	H2, O3, I, U	27
Chlorpyrifos	Chironomus riparius (2nd–4th mi)	s/w/23/5.5#	H2, O3	28
	Chironomus tentans (4th mi)	s/w/10–30/1–4	O3, U	26, 29
Acephate†	Notonectidae, Corixidae (adult)*	s/w/12/3	H1	30
TFM	Chironomus tentans (4th mi)	s/w/22/3	R (nitro), C	31
Pyridalyl	Chironomus yoshimatsui (3rd mi)	s/w/23/2–4	OD, O2, C	32
Tributyltin†	Chironomus riparius (4th mi)	s/w/20/3	SnD	33

1) Collected from the U.S. EPA EPI-Suite [http://www.epa.gov/opptintr/exposure/pubs/episuite.htm]. ^†, pH 5/8. 2) Stage (X^th-instar) and common name (mi, midge; m, mayfly; c, caddisfly; dr, dragonfly) in the parentheses. *, backswimmer and water boatman. 3) Exposure conditions: s, static; f, flow-through/exposure medium (w, water)/temperature in °C/period of exposure in day or hr (^#). 4) C, conjugation; DC, dechlorination; DH, dehydrochlorination; H1, cleavage of amide or carbamate linkage; H2, cleavage of P-Oaryl linkage; O1, alkyl hydroxylation; O2, epoxidation; O3, oxidative desulfuration; R, reduction; I, isomerization; U, unknown (polar) metabolites; XD, X-dealkylation (X=N, O, Sn).

1.2. Relevant enzymes

Mixed-function oxidases including cytochrome P450 (CYP) and carboxylesterases (CaEs) play a large role in the enzymatic oxidation and hydrolysis (so-called phase-I reaction) of pesticides in aquatic insect larvae, and the primary metabolites can be conjugated with endogenous molecules (phase II).¹⁵⁾ The phase-II enzymes, glutathione-S-transferases (GSTs), are utilized as a typical biomarker for the exposure stress of pesticides, together with CYP and CaE.³⁹⁾ Aquatic insects such as mayfliesand caddisfliesgenerally exhibit a well-developed detoxifying system including CYP, CaE, and GST, and their enzyme activities are equal to or greater than those in some terrestrial insects.³⁶⁾ The conjugation of TFM showed the presence of glucuronyl- and sulfo-transferases in C. tentans,³¹⁾ but the corresponding enzyme profiles in aquatic insects have never been reported. Chromatographic purification followed by electrophoretic analysis is a usual way to characterize an enzyme, but transcriptomic studies have recently begun to be applied to Chironomus spp. for the analysis of differentially expressed genes when exposed to a chemical.^40,41)

1.2.1. Oxidases

The insect CYPs detoxifying xenobiotics are distributed in the microsomal membrane, and many of them in terrestrial insects such as mosquitoesand housefliesare classified into CYP 4, 6, 9, and 12 families.⁴²⁾ Many kinds of CYPs have been confirmed through genome sequencing projects, but in the case of terrestrial insects, this work has been mostly focused on pesticide resistance.^42,43) Aquatic invertebrate CYPs have been investigated from the viewpoint of enzymatic activity, immunology, and gene expression, but this work has been mostly limited to Annelida, Arthropoda, and Mollusca.^15,42,43) Among the aquatic insects, Chironomus spp. are most frequently subjected to enzyme characterization.

The involvement of CYPs in metabolism has been indirectly supported by using the CYP inhibitor piperonyl butoxide (PBO). The co-application of PBO greatly reduced the formation of dieldrin from aldrin in C. riparius²⁰⁾ and that of oxon from chlorpyrifos in C. tentans.²⁹⁾ The ECOD (7-ethoxycoumarin O-deethylase) activity in the post-mitochondrial fraction of C. tentans guts decreased with the addition of PBO,⁴⁴⁾ and the aldrin epoxidase activity in the Limnephilus sp. was reduced by the CYP inhibitors carbon monoxide and SKF525-A.¹⁹⁾ The requirement of co-factor NADPH was also found for the oxidase activity in these species.^19,29,37) The optimal temperature and pH of aldrin epoxidase were reported to be 10–30°C and 8–9 in the Limnephilus sp.¹⁹⁾ and 30–31°C and 7.5–7.6 in Chironomus spp.,^20,29) respectively. The concentration of microsomal CYP has been recorded as 0.05–0.07 nmol/mg protein in Stenacron spp. (mayflies), Hydropsyche and Chemataphyche spp. (caddisflies), and S. vittatum,³⁶⁾ which is slightly higher than that in D. magna.⁴⁵⁾ The MROD and EROD (7-(m)ethoxyresorufin O-de(m)ethylase) activities (5–20 pmol/min/mg protein)^44,46,47) are much higher than for ECOD (0.3–0.6 fmol/min/mg protein)^48–50) in Chironomus spp. A species difference in the aldrin epoxidase activity has been found, with the highest in the black fly and no activity in the damselflies.^36,37)

Aldrin epoxidase in C. tentans (molecular weight (mol.wt.) of 45 kDa by SDS-PAGE (polyacrylamide gel electrophoresis)) cross-reacted with a Drosophila melanogaster (fruit fly) anti-P450 polyclonal antibody using Western blot analysis (protein detection).³⁷⁾ The corresponding genomic DNA of this species was amplified by polymerase chain reaction (PCR) and purified to 444 bp, and its sequence showed a high homology with the CYP4G1 of D. melanogaster.⁴⁹⁾ The authors further indicated the presence of at least two subfamilies of CYP4 using Northern blot analysis (RNA detection). Recently, the CYP4G genes have been characterized in both C. tentans⁵¹⁾ and C. riparius⁵²⁾ with a very high homology (93%), each of which encodes 559 amino acid residues with a predicted mol. wt. of 64 kDa. Additionally, the activities of ECOD, MROD, and aldrin epoxidase were induced in Chironomus spp. to various extents by exposure to triazine herbicides,^{29,37,44,46,49,53)} some OPs,^44,46) pyrethroids,⁵⁴⁾ and naphthalene,⁴⁸⁾ depending on their concentration and exposure period. In relation to these inductions, the transcriptional changes of CYP genes have been recently investigated in C. riparius, but the corresponding mechanism is still unclear. The mRNA expression of the CYP gene was up-regulated by exposure to phenanthrene⁴⁰⁾ and the veterinary antibiotic fenbendazole⁵⁰⁾ in a dose-dependent manner. The up-regulation of the CYP4G gene was caused by tributyltin⁵²⁾ and pentachlorophenol,⁴¹⁾ while both nonylphenol and bisphenol A down-regulated the gene expression.⁵²⁾ By the exposure of C. riparius at successive life stages (eggs to adult) to several chemicals, real-time PCR analysis showed the highest mRNA expression of the CYP9AT2 gene (encoding 528 amino acid residues with a predicted mol. wt. of 61.3 kDa) in the 4^th instar larvae, probably due to higher exposure via the dietary route.⁵⁵⁾ The authors reported down-regulation by nonylphenol and bisphenol A in a manner similar to that of the CYP4G gene, and high up-regulation was observed when exposed to benzo[a]pyrene and chlorpyrifos.

1.2.2. Esterases

CaEs are known to cleave the carboxylic ester linkages of pesticides.¹⁵⁾ The activity of permethrin hydrolases has been detected in several aquatic insect larvae.³⁶⁾ They are B-type esterases typically inhibited by OP chemicals, as reported for Hydropsyche pellucidula (caddisfly), whose enzyme activity against p-nitrophenyl acetate and α-naphthyl acetate (α-NA) was significantly reduced by fenitrothion.⁵⁶⁾ Furthermore, the phosphotriesterase (A-type) activity was detected in this species by using paraoxon as a substrate. The spectrophotometric assay of the CaE activity is generally conducted by using α- and β-NA as substrates. Its specific activity (nmol/min/mg protein) seems to be lower in black flies (3–68),^57,58) the caddisfly (92),⁵⁶⁾ and midges (20–36),^47,59) as compared with D. magna (220)⁶⁰⁾ and Hyalella curvispina (amphipod) (101).⁶¹⁾ Although the number of tested substrates is limited, the CaE activity decreased as follows: black fly>caddisfly≥mayfly>damselfly.³⁶⁾ It is noted that the enzyme polymorphism is known for a few species. The PAGE analysis of the post-mitochondrial fraction of whole-body homogenates showed multiple forms of CaEs in C. tentans, and the banding patterns monitored by the hydrolytic activity against α-/β-NA were independent of the three pesticides used for exposure.⁴⁶⁾ The polymorphisms are similarly reported for Simulium spp., and the different CaE activities were observed for these substrates among the collection sites of the larvae.^57,62)

1.2.3. Glutathione-S-transferases (GSTs)

The cytosolic GSTs catalyze the conjugation of electrophilic substrates including primary pesticide metabolites with glutathione, and their activities are generally assayed using 1-chloro-2,4-dinitrobenzene (CDNB).¹⁵⁾ Dierickx⁶³⁾ showed the distribution of GST activity against CDNB common to aquatic macroinvertebrates, and the activity in Ischnura elegans (damselfly) was an order of magnitude higher than that in E. danica, Hydropsyche angustipennis (caddisfly), and Chironomus sp. In contrast, other authors reported similar enzyme activity against CDNB (nmol/min/mg protein), ranging within an order of magnitude among black flies (190–452),^36,58) caddisflies (90–501),^36,56,63) mayflies (261–337),^36,63) damselflies (445),³⁶⁾ and midges (161–502, C. tentans^46,47,64); 63–250, C. riparius^48,50,59)). These variations may originate not only from species differences but also as a result of different exposure histories to contaminants. GST activity against 1,2-dichloro-4-nitrobenzene (DCNB) was much lower in mayflies,³⁶⁾ damselflies,³⁶⁾ and midges^46,47,64) and was absent in black flies and caddisflies,³⁶⁾ indicating a different distribution of GST isozymes. Similar activity against CDNB has been reported for the D. magna (but no activity against DCNB)⁶⁵⁾ and Hyalella spp.^61,66) The modified activities against CDNB and DCNB by exposure to pesticides have been examined mainly in Chironomus spp. Several OP, OC, and pyrethroid insecticides gave no or slightly increased activities (less than 1.4-fold),^{46,48,56,59,62,67)} while the induction of GST activities by a factor of around two was observed for metolachlor⁴⁷⁾ and 2,4-D.⁶⁸⁾ In the case of alachlor, activities reduced by about half most likely originated from the loss of glutathione by rapid conjugation, as supported by the increased GST gene expression.^47,64) Transcriptional analysis indicated no effect on the mRNA expression of GST genes by the exposure of C. riparius to PCP,⁴¹⁾ and the change in mRNA expression by tributyltin was inconclusive in this midge.^40,52)

Extensive studies on the genome sequences have indicated that insect cytosolic GSTs are assigned to several classes, including δ, σ, and ω.⁶⁴⁾ The cytosolic fraction of the Chironomus sp. was successively chromatographed, and five GST-active bands with a mol. wt. of 23 kDa by SDS-PAGE and an optimal pH of ca. 8 were isolated.⁶⁹⁾ The high homology (56–66%) in the N-terminal sequence for the housefly and mosquito could classify some of these isozymes as σ-GST, which was supported by the activities against not only CDNB and DCNB but also cumene hydroperoxide (σ-GST substrate). Recently, gene analysis using expressed sequence tags has found eleven GST transcripts in C. tentans.⁶⁴⁾ The seven of them were assigned to two δ-, four σ- and one ω-class, all of which were observed in the larvae, but some were lost in the adult. Further, thirteen GST transcripts, including ε-, θ-, and ξ- classes were identified by similar analysis in C. riparius, and one δ-class and one unknown GST in two species are individually homologous with >93% similarity in the amino acid sequence.⁷⁰⁾ These two studies detected the up-regulation of δ- and σ-GST genes by exposure to alachlor and paraquat.

2. Bioconcentration and bioaccumulation

The accumulation of pesticides in aquatic insect larvae proceeds via two routes, through the body surface from water and as a result of dietary uptake from sediment and food items. The former is designated as bioconcentration and both are involved in bioaccumulation.¹⁵⁾ Much less relevant information is available for aquatic insect larvae than for fish. This is due in part to the fact that standardized experimental procedures are only available for Chironomus spp.^71,72) and as a result of difficulties in their evaluation due to the small size and body weight (<0.1 g in many cases).^73,74) The exposure to pesticides is conveniently conducted by using aqueous solutions under static or flow-through conditions or, alternatively, in water-sediment systems applied with pesticides. Since the larvae ingest food or sediment over a relatively long exposure period, the contribution of a pesticide and its metabolites present in their guts needs to be eliminated as far as possible to correctly estimate bioconcentration/-accumulation. Brooke et al.⁷⁵⁾ measured the clearance rates of gut content weight after exposure to several sediments for 10–30 days, and the majority could be depurated after 1 day from Hexagenia limbata (mayfly), C. tentans, and Lumbriculus variegatus (worm).

In order to quantify the accumulation, the following indices are generally utilized.   

C_B, C_W, and C_S are the steady-state concentrations of pesticides in biota, water, and sediment, respectively. f_OC and f_L are the fractions of organic carbon in sediment and lipids in biota, respectively. Lipids are the relevant component for the accumulation of pesticides and mostly amount to 1–8% and 10–20% of a body weight on a wet and dry basis, respectively, with the C₁₆–C₂₀ fatty acid components predominantly detected.^15,76,77)

2.1. Relationship with pesticide hydrophobicity

The linear relationship between log BCF and log K_ow (n-octanol–water partition coefficient) is well known for fish and other aquatic species. However, the physiological profiles of each species affecting the uptake, metabolism, and elimination of a pesticide will deviate a measured BCF value from the predicted one, leading to the result that about one-third of the variance in BCF can be explained by log K_ow of a chemical in aquatic organisms.¹⁵⁾ A good relationship with log K_ow was reported for the log BCF of chloroaromatics and polycyclic aromatic hydrocarbons (PAHs) in Chironomus decorus⁷⁸⁾ and H. limbata⁷⁹⁾ and for the log BAF of polychlorinated biphenyls (PCBs) in H. limbata.⁸⁰⁾ No dependence of the log BSAF on log K_ow was reported for three OC pesticides in the same mayfly species.⁸¹⁾ However, a similar analysis of various pesticides has not been conducted for aquatic insects. The BCF, BAF, and BSAF values of pesticides in the larvae of Chironomus spp. and other aquatic insects are summarized in Tables 2 and 3, respectively. The plot of the observed log BCF of pesticides by water exposure against its log K_ow (Fig. 1) shows no clear relationship between them. This results not only from the different physiology of each aquatic insect larvae but also from the different experimental conditions concerning pH, temperature, exposure period, analytical method, and larval instar. When the ¹⁴C-based log BCF values of fourteen pesticides in Chironomus spp. (Table 2) were analyzed against log K_ow (2.4–8.1), a better linear correlation was obtained (r²=0.48), indicating the importance of insect physiology. As previously reported for OCs in crustaceans and OPs in mollusks,¹⁵⁾ a better relationship was obtained by applying a quadratic equation to the above data: log BCF=−0.17 (log K_ow −6.0)² + 3.3 (r²=0.64). This analysis clearly shows the importance of hydrophobicity in bioconcentration and the probable contribution of metabolism. Accordingly, a standardized condition for specific species should be required to correctly assess the accumulation process.

2.2. Uptake, metabolism and depuration

Each process is kinetically analyzed in a two-compartment model by assuming first-order rate constants (k_U, uptake; k_M, metabolism; k_D, depuration).¹⁵⁾ However, the kinetic analysis of pesticides in aquatic insect larvae is very limited in its availability. Recently, Bayen et al.⁹⁹⁾ proposed the diffusion model for aquatic organisms exposed to PAHs and PCBs dissolved in water. The surface area of the biota/water interface divided by the wet weight of biota was suggested to determine the uptake process. It was considered that diffusion in aquatic insects controlled the uptake at log K_ow <3 and that log k_D was negatively correlated with chemical hydrophobicity at log K_ow >3. A good correlation between log (k_U×mol. wt.) and log K_ow (r²=0.93) was observed in the static water exposure of C. riparius to OCs including DDE and dieldrin, indicating the importance of molecular size in the uptake process.¹⁰⁰⁾ Lower BCFs but with larger k_U values at a higher temperature were reported in Chironomus spp. for several insecticides, which was accounted for by the enhanced transformation to polar metabolites, these being more easily depurated.^22,25,26) The contribution of metabolism to a reduced BCF was also proposed for the water exposure of Pteronarcys californica (stonefly) to DDT.¹⁰¹⁾ A temperature effect on BCF was quantitatively confirmed by the detailed kinetic analysis of the exposure of Chironomus dilutus to chlorpyrifos and two pyrethroids at 13 and 23°C, showing a higher contribution of k_M and k_D than k_U at 23°C.⁸²⁾ The respiratory strategy of an aquatic insect may at least in part account for the temperature effect on the pesticide uptake and BCF. An increase in the k_U of ³H₂O at a higher temperature was observed in P. californica and the Cinygma sp. (mayfly), which have tracheal gills, as a result of lower dissolved oxygen in water, while at the same time the uptake of chlorpyrifos increased.¹⁰²⁾

Not only lower bioavailability due to the association with dissolved organic matter (DOM) and/or colloidal matters but also the dietary uptake changes the bioaccumulation of pesticides in the water-sediment system. Looser et al.³³⁾ reported a lower bioconcentration of the more hydrophobic tributyltin in C. riparius compared to that of triphenyltin, which could be explained by a larger k_M (by two orders of magnitude) for the former pesticide compared to the latter. They observed a lower BCF in the presence of Aldrich humic acids in the exposure water, which was caused by lower bioavailability as a result of the association of pesticides with DOM. Lower uptake of chlorpyrifos to this midge was also reported in exposure water having a higher organic carbon content.¹⁰³⁾ The BSAF value of cypermethrin decreased in a water-sediment system having higher organic carbon. This could be explained by lower bioavailability of the insecticide to C. tentans due to a greater association with DOM and colloidal matter.⁸⁸⁾ As a result of the similar BCF values comparing water and water-sediment exposures of C. riparius to λ-cyhalothrin and with its BAF being inversely proportional to the sediment-water partition coefficient, only the dissolved fraction is considered to be bioavailable.⁸⁹⁾ The direct analysis of pore water in sediment demonstrated that the dissolved polychlorinated benzenes were bioavailable to C. decorus.⁷⁸⁾

Muir et al.^83,87) examined the effect of sediment in bioconcentration by utilizing a screen to isolate C. tentans from the sediment applied with a number of ¹⁴C-pesticides, including several pyrethroids. The adsorption of the pesticides onto sediment in the presence of the screen reduced their BCF by a smaller k_U but an almost constant k_D. By comparing the ¹⁴C-residues in the midges without the screen, it was found that the concentration of each pesticide in pore water provided a good indicator of its BCF, but the dietary route via ingestion of sediment contributed more to the pesticide uptake. The different BCFs due to geometrical isomerism (cis>trans) in permethrin and cypermethrin⁸⁷⁾ most likely originate from the preferential ester cleavage in trans isomers by CaE.¹⁵⁾ Larvae, non-feeding pupae, and dead larvae of Hydropsyche bidens (caddisfly) were individually exposed to dibenzofuran and dioxins via a dietary route in a stream microcosm.¹⁰⁴⁾ The comparison of bioaccumulation showed the dietary route to be predominant for these hydrophobic contaminants (86–87%), with less contribution from the uptake via water (7–10%) and the body surface (3–7%). However, the effect of sediment on pesticide uptake seems more complicated. More uptake but less depuration of ¹⁴C-chlorpyrifos from a non-sterile sediment was observed in C. riparius compared to that from a sterile sediment. These changes were emphasized by the addition of sediment microbial extracts with a pyrophosphate buffer.¹⁰⁵⁾ The passive uptake was reduced by the association of the pesticide with DOM, while more feeding on the fortified sediment by the larvae increased the dietary uptake. The depuration of DDE and PCP taken up by C. riparius was enhanced in a water-sediment system having higher organic carbon, but the mechanism was unclear.¹⁰⁶⁾

Table 2. Bioconcentration and bioaccumulation of pesticides in Chironomus spp.

14C-Pesticide	log Kow1)	Species (instar)	Exp. cond.2)	BCF×10−3 3)	BAF3)	BSAF3)	Ref.
DDT	6.91	Chironomus sp.	f/w/21/3	48			17
		C. dilutus (4th)	s/w/13, 23/4	49, 30*,+			82
Methoxychlor	5.08	C. tentans (4th)	s/s/—/2	0.13–1.3			83
Aldrin	5.40	Chironomus sp.	f/w/21/3	22.8			17
Hexachlorobenzene	5.73	C. decorus (4th)	f/s/—/2	0.85–0.88	0.58		78
Atrazine	2.61	C. riparius (1st)	s/s/18–23/42	1.9			84
Terbutryn	3.74	C. tentans (4th)	s/s/—/4	0.01–0.07			83
Benthiocarb	3.40	C. plumosas (4th)	f/w/20/5	0.03–0.08			85
Carbaryl	2.36	C. riparius (4th)	s/w/10–30/1	0.05–0.1*			22
Fenitrothion	3.30	C. riparius (3rd)	s/w/26/5	0.07–0.23*			27
Chlorpyrifos	4.96	C. riparius (2nd–4th)	s/w/23/7#	0.43–1.0			28
		C. dilutus (4th)	s/w/13, 23/4	0.83, 0.34*,+			82
Phosmet	2.78	C. plumosas (4th)	f/w/20/1	0.009			86
Fonofos	3.94	C. riparius (1st)	s/s/18–23/42	0.75–1.2			84
Permethrin	6.50	C. dilutus (4th)	s/w/13, 23/4	0.32, 0.09*,+			82
trans-isomer		C. tentans (4th)	s/s/—/1–4	0.01–0.3			83, 87
cis-isomer		C. tentans (4th)	s/s/—/1–2	0.008–0.42			87
Cypermethrin	6.40	C. tentans (3rd–4th)	s/s/23/4	0.3–2.3	0.06–0.62	0.08–0.63	88
trans-isomer		C. tentans (4th)	s/s/—/1–2	0.04–0.25			87
cis-isomer		C. tentans (4th)	s/s/—/1–2	0.03–0.39			87
Deltamethrin	6.20	C. tentans (4th)	s/s/—/1–2	0.06–0.31			87
Fenvalerate	6.22	C. tentans (4th)	s/s/—/1–2	0.02–0.17			87
λ-cyhalothrin	6.90	C. tentans (4th)	s/w, s/23/4	1.5–2.0	0.39		89
		C. dilutus (4th)	s/w/13, 23/4	0.17, 0.09*,+			82
Pentachlorophenol	5.12	C. riparius (4th)	s/w/20/16#	0.46			90
Fluridone	3.16	C. tentans (4th)	s/s/—/4	0.005–0.07			83
Trifluralin	5.34	C. riparius (1st)	s/s/18–23/42	0.21			84
Ioxynil	3.43	C. riparius (1st)	s/s/20/10		0.37–1.6		91
Fipronil	4.00	C. annularius	s/w/—/2	0.06			92
Bentazone	2.34	C. riparius (1st)	s/s/20/10		0.11–1.3		91
Pyridalyl	8.10	C. yoshimatsui (3rd)	s/w/23/2–4	0.06*			32
Tributyltin	2.8/4.1†	C. riparius (4th)	s/w/20/3	0.14/0.9+,†			33
Triphenyltin	3.0/3.5†	C. riparius (4th)	s/w/20/3	2.2/2.7+,†			33

1) Same as key 1) in Table 1. 2) Same as the key 3) in Table 1. s (exposure medium), sediment; —, not available. 3) On the ¹⁴C-basis; BCF, bioconcentration factor in L/kg; BAF, bioaccumulation factor in kg soil/kg larvae; BSAF, biota-soil accumulation factor in kg organic carbon/kg lipid. *, parent-basis; ⁺, kinetically estimated.

Table 3. Bioconcentration and bioaccumulation of pesticides in the aquatic insects other than Chironomus spp.

Pesticide1)	log Kow2)	Species3)	Exp. cond.4)	B-factor5)	Ref.
DDT†	6.91	Ephemera danica (m)	f/w/14/9	BCF: 2.9–8.3×103	16
DDT		Hexagenia bilineata (m)	f/w/21/3	BCF: 3.3×104	17
		Siphlonurus sp. (m)	f/w/21/3	BCF: 2.3×104	17
		Ischnura verticalis (d)	f/w/21/3	BCF: 3.5×103	17
DDD†	6.02	Hexagenia sp. (m)	s/s/23/28	BSAF: 0.2–0.4*	34
DDE†	6.51	Hexagenia limbata (m)	s/s/20–22/32–33	BSAF: 6.1–6.9*	81
Methoxychlor†	5.08	Pteronarcys dorsata (s)	f/w/15/28	BCF: 0.35–1.1×103*	93
		Hydropsyche sp. (c)	s/w/rt/0.5#	BCF: 8–16*	94
		Simulium vittatum (b)	s/w/rt/0.5#	BCF: 0.8–31*	94
Aldrin	5.40	Hexagenia bilineata (m)	f/w/21/3	BCF: 3.1×104	17
Endrin†	5.40	Pteronarcys dorsata (s)	f/w/15/28	BCF: 0.61–1.0×103*	93
trans-nonachlor†	6.20	Hexagenia limbata (m)	s/s/20–22/32–33	BSAF: 6.7–9.4*	81
Hexachlorobenzene†	5.73	Hexagenia limbata (m)	s/s/20–22/32–33	BSAF: 1.1–1.2*	81
Atrazine	2.61	Baetis sp. (m)	e/w/30/60	BCF: 52–124*	95
		Nemoura sp. (s)	e/w/30/60	BCF: 29–47*	95
Benthiocarb	3.40	Hexagenia bilineata (m)	f/w/20/5	BCF: 36–85	85
Chlorpyrifos	4.96	Hydropsyche spp. (c)	s/w/20/6#	BCF: 40	96
		Molanna angustata (c)	s/w/17/2	BCF: 5.3×103	77
		Stenacron spp. (m)	s/w/20/6#	BCF: 13	96
		Cloeon dipterum (m)	s/w/17/2	BCF:1.8×103+	77
		Sialis lutaria (a)	s/w/17/2	BCF:9.6×103+	77
		Ischnura spp. (d)	s/w/20/6#	BCF: 5.4	96
		Anax imperator (dr)	s/w/17/2	BCF:100+	77
Phosmet	2.78	Ischnura verticalis (d)	f/w/20/1	BCF: 1.0	86
Permethrin	6.50	Hydropsyche spp. (c)	s/w/20/6#	BCF: 30	96
		Stenacron spp. (m)	s/w/20/6#	BCF: 24	96
		Ischnura spp. (d)	s/w/20/6#	BCF: 6.9	96
		Simulium vittatum (b)	s/w/20/6#	BCF: 18	96
Permethrin†	6.50	Pteronarcys dorsata (s)	f/w/15/4	BCF: 43–570*	97
TFM	2.87	Hexagenia sp. (m)	e/w/rt/1	BCF: 8.7	98
		Glossosoma sp. (c)	e/w/rt/1	BCF: 34	98
		Limnephilus sp. (c)	e/w/rt/1	BCF: 19	98
		Bachycentrus americanus (c)	e/w/rt/1	BCF: 62	98

1) ¹⁴C-label unless marked with ^† (non-labeled). TFM, 4-nitro-3-trifluoromethylphenol. 2) Same as the key 1) in Table 1. 3) Common name in the parentheses. s, stonefly; d, damselfly; a, alderfly; b, black fly; others, same in the key 2) in Table 1. 4) Same as the key 3) in Table 1. e, eco-system; rt, room temperature. 5) Same as the key 3) in Table 2.

2.3. Species differences

The field monitoring of PCBs and PAHs in the western part of Lake Erie in the USA showed that their BSAFs in Hexagenia spp. were higher than in mussels, crayfish, and amphipods, possibly due to their ingestion of the contaminated sediment and detritus.¹⁰⁷⁾ However, there seems to be no clear ranking in the bioconcentration/-accumulation of pesticides among orders, judging from the large variation in BCF, BAF, and BSAF values reported for the various aquatic insects.^17,34,85,95) Even in the same order, four caddisfly species showed different BCF values of TFM, varying by a factor of 5.⁹⁸⁾ A wide variation of lipid content⁷⁶⁾ and different metabolic activities¹⁷⁾ among aquatic insect larvae may, at least in part, cause these results, but the relevant information is not available to examine the mechanism in detail. The species differences can be also related to either the ecology or physiology of each aquatic insect. After the outdoor application of fenitrothion to two Canadian headwater streams, much higher body residues in the mayfly and black fly were detected as compared to the caddisflies.¹⁰⁸⁾ Fenitrothion in water was predominantly associated with suspended particles, and the highest residue was detected in the liverwort. Since the mayfly is a herbivore scraping surfaces for periphyton and the black fly is a filter feeder, while the caddisflies are detritivores, their feeding habit may affect the uptake of fenitrothion. By measuring the uptake of ³H₂O and ¹⁴C-chlorpyrifos to the larvae of various species, Buchwalter et al.⁷³⁾ have shown that either body size or respiratory strategy determines the bioconcentration. The dissolved oxygen breathers with a smaller size through epithelial surfaces (e.g., Chironomidae) or tracheal gills (e.g., Ephemeroptera and Trichoptera) required a shorter period to attain a steady-state uptake of ³H₂O, together with a more rapid uptake of the pesticide, than the air breathers (e.g., Hemiptera) with a larger size. The kinetic analysis of the static water exposure of fifteen freshwater arthropods (larvae and adults) to chlorpyrifos showed lipid-normalized BCF values decreasing as follows: midge>caddisfly>mayfly>dragonfly.⁷⁷⁾ The estimated k_U and k_D values mostly accounted for this result, and they were likely to depend on differences in body size, lipid content, respiratory strategy, and metabolic activity.

Fig. 1. Bioconcentration of pesticides in the aquatic insect larvae. ◆, midge; ■, mayfly; ▲, caddisfly; ●, stonefly; ○, damselfly; □, black fly.

2.4. Effects of life stage

In the static water exposure of caddisfly egg masses (Triaenodes tardus) to ten pesticides, the negative correlation of log BCF with log (pesticide water solubility) showed pesticide hydrophobicity as the determining factor in its uptake, and the presence of maxima in the uptake curves implied their release from the gelatinous egg matrix.¹⁰⁹⁾ The progress from the 2^nd to 4^th instar in C. riparius decreased the body residues of chlorpyrifos by water exposure.²⁸⁾ Since analysis revealed almost a constant ratio of metabolites to chlorpyrifos, less uptake due to a smaller surface-to-volume ratio might account for the lower residues in the later instars. In contrast, when Chironomus spp. were continuously exposed to DDE¹¹⁰⁾ or trans-chlordane¹¹¹⁾ in water from egg masses to adults, the BCF values increased with their growth, probably due to their insignificant metabolic degradation. When the pesticide concentration in adults is significantly higher than that in larvae, more biomagnification in predators is concerned.¹¹²⁾ Information on the dependence of bioconcentration/-accumulation on the life stage of other aquatic insects is only available for polychlorinated contaminants. The residue analysis of PCBs in sediments and Hexagenia spp. collected from the contaminated sites showed that the total PCB concentration in the mayfly remained almost constant irrespective of the growth stage.¹¹³⁾ However, the BAF value for the male imago significantly increased as compared with the larval and sub-imago stages due to the reduced body weight (lipid) caused by the reproductive costs during mating swarms. Therefore, male adults with a low nutritional value and a higher content of PCBs may have a greater potential for biomagnification in predators.

3. Toxicity

The use of standard species is indispensable to understand the relative toxicity among pesticides, but the relevant guidelines for the acute and chronic toxicity in aquatic insects are only available for Chironomus spp.^71,72,114) Since these midges do not always represent all aquatic insects in relation to their different ecology and sensitivity, some researchers have proposed new testing methods for the other species. The 1^st instar larvae of the net-spinning caddisfly Cheumatopsyche brevilineata were introduced to examine the 2-day acute toxicity of thirty pesticides by water exposure without a substrate.¹¹⁵⁾ A 14-day sediment toxicity study in the mayfly was proposed for H. limbata with the toxicological signs including immobility as endpoints.¹¹⁶⁾ It should be noted that the ecotoxicological safety assessment of pesticides is generally conducted on the margin-of-safety concept by comparing the toxicity and the relevant environmental concentration of a pesticide. Brock and Wijngaarden¹¹⁷⁾ have demonstrated that the acute toxicity data with a safety factor of 100 can conservatively evaluate the impact of the insecticides observed in many micro-/mesocosm studies when the standard species, such as D. magna, Americamysis bahia (mysid shrimp) and/or Chironomus spp., are taken into consideration. The EFSA⁹⁾ has recently proposed a tiered approach starting with laboratory acute and chronic studies using the standard species and progressing to geometric mean toxicity or a species sensitivity distribution (SSD) with additional species, and then an outdoor microcosm study, as appropriate. At each stage, an appropriate assessment factor is used.

3.1. Pesticide and species sensitivity

Only water exposure has been conveniently used to exclude the effect of pesticide adsorption onto a substrate. Even so, the different experimental conditions make it difficult to simply compare the existing toxicity data among both insect orders and also pesticide classes. For example, the acute toxicities (LC₅₀ in µmol/L) of eight insecticides after water exposure for 1–4 days were compared in the larvae of mayflies,^118,119) stoneflies,^120,121) caddisflies,¹¹⁸⁾ and midges.^{20,119,122–128)} The LC₅₀ values varied by up to two orders of magnitude, demonstrating the large variation in species-dependent sensitivity even within each order (Fig. 2). This variation may be caused by different profiles in uptake, metabolism, and/or depuration. The collection site of larvae should be also taken into account. Thus,

Fig. 2. Acute toxicity of pesticides in aquatic insect larvae. 1, fipronil; 2, imidacloprid; 3, DDT; 4, dieldrin; 5, parathion; 6, malathion; 7, deltamethrin; 8, fenvalerate. Numbers of species and toxicity data / days of exposure are in the parentheses. ○, mayfly; ◇, stonefly; △, caddisfly; □, midge.

C. riparius collected near a sewage treatment plant exhibited 13–228-fold lower acute toxicity against four insecticides compared to a laboratory culture, which was likely due to induced GST activity.¹²²⁾ The black fly larvae collected from a pesticide-free upstream area exhibited approximately two orders of magnitude higher acute toxicity against fenvalerate and deltamethrin than did those from an irrigation channel exposed to insecticides, which was explained by a 3–6-fold lower CaE activity.⁵⁸⁾

Variation has also been shown between pesticide classes. Endrin was most acutely toxic in Pteronarcys spp. (stoneflies),¹²⁰⁾ while some pyrethroid and neo-nicotinoid insecticides showed very high acute toxicities in midges and mayflies. A wide variation in the acute toxicity of different chemical classes (log LC₅₀ in µmol L⁻¹) is reported for midges and mayflies^129–133) with the related references in Fig. 2: −3.7 – −0.1 and −3.0 – −1.7 (neo-nicotinoids), −4.7 – −1.0 and −2.5 – −2.0 (pyrethroids), −3.2 – −1.2 and −3.3 – −0.3 (OPs), and −2.9 – −1.9 and −1.9 – 1.2 (OCs), respectively. In general, fungicides and herbicides are much less acutely toxic than insecticides.

The aquatic insect orders sensitive to pesticides have been examined, but not systematically. One-day water exposure to chlorpyrifos and three pyrethroids showed that the mayfly and damselflies were more sensitive than the black fly and caddisflies.¹³⁰⁾ Pteronarcys dorsata (stonefly) was less sensitive to fenvalerate under flow-through water exposure as compared to the Ephemerella sp. (mayfly).⁹⁷⁾ A 4-day acute toxicity study on fipronil was conducted for C. dilutus, seven mayfly species, two stonefly species, and three caddisfly species, giving useful information on the species sensitivity.¹¹⁸⁾ The EC₅₀ values in ppt were estimated to be 30–35 (midge), 52–>1229 (mayflies), 101–>184 (stoneflies), and 267–634 (caddisflies). The midge larvae were the most sensitive, and the caddisflies were less sensitive than the stoneflies, but the variation in EC₅₀ values made it difficult to clearly rank the orders. Three mayfly species were more sensitive to λ-cyhalothrin and bifenthrin than were Daphnia galeata (2-day water exposure)¹³¹⁾ and C. tentans (10-day sediment exposure).¹¹⁶⁾ C. brevilineata was more sensitive to many pesticides following 2-day water exposure as compared to D. magna.¹¹⁵⁾ Based on the toxicity studies of various aquatic insect larvae, the SSD approach gives a useful insight.⁹⁾ A recent review of the acute and chronic toxicities of neo-nicotinoids in twelve invertebrate orders shows that the aquatic insects are more sensitive than crustaceans, with the acute toxicity decreasing as follows: Ephemeroptera>Trichoptera>Chironomus spp.>Diptera>Odonata.¹¹⁹⁾ However, the potential sensitivity relating to the pesticide class and insect order should also be assessed by considering the results from higher-tier studies under more realistic exposure conditions. This is considered in Section 3.5.

3.2. Effects of experimental conditions

Water chemistry and temperature vary in the environment, and their effects on toxicity should be taken into account in a risk assessment. The acute toxicities of some pesticides change with the pH and temperature of the exposure water in a complicated manner. The 4-day toxicity in Pteronarcella badia (stonefly) decreased for acephate and carbaryl with an increasing pH, while the opposite change was observed for trichlorfon, and there were insignificant effects on aminocarb and fenitrothion.¹³⁴⁾ The lower toxicity at a higher pH was reported in C. riparius for carbaryl²²⁾ and parathion.²⁵⁾ Much higher toxicity at a higher temperature was common to three OP insecticides in Chironomus spp.,^25,26) but the differences were smaller for carbaryl.²²⁾ Detailed analysis using ¹⁴C-pesticides indicated that these dependencies should be considered in combination with the pesticide stability in water, uptake rate, activities of relevant enzymes such as CYPs and CaEs, depuration rates of a pesticide and its metabolites, and re-absorption of metabolites.^22,25,26)

Sediment is necessary as a habitat and/or food for many aquatic insect larvae. Pesticides are partitioned between water and sediment mainly depending on their hydrophobicity and the adsorptive capacity of sediment, and at the same time, they undergo microbial degradation.¹³⁵⁾ The adsorption of pesticides onto sediment significantly reduced their short-term toxicity in C. riparius, taking into account the minimal effects of degradation and dietary uptake.^136,137) The 1-day EC₅₀ values showed a moderately positive correlation with log K_ow of pesticides, supporting the effect of adsorption.¹³⁶⁾ In a 10-day exposure of C. tentans to ¹⁴C-cypermethrin in three sediments with different organic carbon contents, the higher the organic content, the more the adsorption onto the sediment was associated with a lower BSAF value, resulting in larger LC₅₀ and NOEC values.⁸⁸⁾ A similar change dependent on the sediment organic carbon was reported for ioxynil and bentazone with this midge.⁹¹⁾ In contrast, the dietary uptake route predominates for highly hydrophobic chemicals,¹⁰⁴⁾ indicating the possible importance of sediment toxicology for hydrophobic and persistent pesticides.

When a metabolite formed via microbial degradation is more toxic than the parent, its effect on toxicity should be assessed by considering the fraction of its formation.¹¹⁸⁾ The sulfide and sulfone metabolites of fipronil, formed in a water-sediment system,¹³⁵⁾ were more toxic than the parent in C. brevilineata. However, the risk quotient calculated from the monitored concentration and EC₅₀ of each metabolite with a safety factor of 10 did not exceed 1 for these mixtures in a Japanese river.¹³⁸⁾ Two metabolites showed higher acute toxicity in many species of EPT compared to the parent fipronil, and a toxic concern for the most sensitive species (C. dilutes) was reported for the California creek receiving urban storm-water runoff.¹¹⁸⁾ Furthermore, periodical feeding should be conducted in a longer-term study to assess the effects of metabolites on growth where the dietary exposure to pesticides needs to be considered.

The dietary uptake of diflubenzuron applied to pieces of yellow poplar leaves was confirmed for the larvae of two stonefly species, and the observed minimal toxicity was accounted for by its rapid metabolism.¹³⁹⁾ It should be noted that aquatic insect larvae have different feeding strategies,^6,14) which should be considered in the dietary exposure of pesticides. When the larvae of two stonefly species were exposed to a particulate formulation of methoxychlor, more uptake by P. dorsata (detritivorous) than by Acroneuria lycorias (carnivorous) was observed, most likely due to the higher consumption of the particles as a result of the scraping habit in the former species.¹⁴⁰⁾ The effect of the dietary uptake of esfenvalerate on long-term toxicity was investigated for a few aquatic insect larvae.¹⁴¹⁾ Brachycentrus americanus (caddisfly) and Cinygmula reticulata (mayfly) fed on esfenvalerate-treated algal cultures showed increasing mortality with case abandonment and growth inhibition with lower egg production, respectively, clearly indicating the dietary uptake of the insecticide. In contrast, when esfenvalerate-treated C. tentans larvae were fed to Hesperoperla pacifica (stonefly), no significant toxic effects were observed without a prey rejection, most likely due to the rapid metabolism of the insecticide.

3.3. Prediction of acute toxicity

The body burden of pesticide as a result of its accumulation is the primary factor determining toxicity, and hence, some correlation of acute toxicity with log K_ow is expected. Such an approach is very limited for aquatic insect larvae, and only a few analyses are available for the midges. The 2-day LC₅₀ values in the water exposure of C. riparius to nine simple organic compounds were found to correlate well with log K_ow (r²=0.99).¹⁴²⁾ These chemicals are narcotics without any particular MOA, but a more complicated relationship was reported for biologically active pesticides. No simple correlation of 1-day EC₅₀ was observed in C. riparius with either mol. wt. or log K_ow for twelve carbamate and OP insecticides following water exposure, but the linear solvation energy and molecular connectivity indices greatly improved the correlation.^136,137) In the case of water-sediment exposure, a similar relationship was confirmed but with a slightly lower correlation. These studies show the importance of size, shape, dipolarity, and basicity of a pesticide molecule in comparison with its acute toxicity, which might relate to its interactions with a target site influencing toxicity. In water exposure (1–4 days at 21–25°C) of various midge species to fourteen insecticides, the correlation of log LC₅₀ values in Section 3.1 and other studies^{92,118,119,129,131)} was low against log K_ow (r²=0.42, n=136), showing that about half of the variance in the toxicity was accounted for by pesticide hydrophobicity. An even poorer correlation (r²=0.33) was obtained for six OPs (n=77) having the same MOA. In the case of mayflies, there was no correlation between log LC₅₀ (Section 3.1) and log K_ow (r²=0.04, n=80). Therefore, the insignificant or poor correlation indicates the contribution of factors other than metabolism and MOA, for example the different interactions with a target site. The simple prediction of acute toxicity therefore seems very difficult in aquatic insects.

3.4. Combined toxicity

In addition to several active ingredients occurring in one commercial formulation, there may also be a co-existence of different pesticides and anthropogenic contaminants in a single habitat that results in unexpected toxicity. In order to evaluate combined toxicity, concentration addition and independent action models have been proposed for pesticides having the same and different MOA, respectively.¹⁴³⁾ However, many researchers have reported deviations from these models. By the addition of very weakly toxic atrazine, the 4-day acute toxicity in C. tentans in static water exposure increased by a factor of 2–4 for chlorpyrifos, parathion-methyl, and diazinon, but no effect was observed for malathion.²⁹⁾ Much greater inhibition of acetylcholine esterases in the same species occurred with exposure of both chlorpyrifos and atrazine, as compared to a single application.⁵³⁾ CYP induction by atrazine leading to greater formation of toxic oxon derivatives most likely accounts for this enhanced toxicity, which was confirmed by the time- and concentration-dependent induction of the CYP activity.⁴⁴⁾ The antagonistic action by PBO has been frequently reported in Chironomus spp. for OPs and other insecticides with their toxicity activated by oxidation, while enhanced toxicity has been observed when CYP is involved in the detoxification.^{20,29,44,128)} Synergism in the mixture of chlorpyrifos and imidacloprid was reported for the acute toxicity in C. dilutus, which similarly might be due to CYP induction by imidacloprid.¹⁴⁴⁾ The authors observed either synergism in the mixture of chlorpyrifos and dimethoate (common MOA) or antagonism in that of dimethoate and imidacloprid (different MOA), indicating the importance of other mechanisms in considering the combined toxicity. A different mechanism for alachlor and metolachlor was proposed for the enhanced toxicity of chlorpyrifos.⁴⁷⁾ The constant activities of ECOD and CaEs but a lower GST activity with the former herbicide might reduce detoxification, while the latter additionally decreased ECOD activity due to lower protein production.

Conclusion

The aquatic insect larvae, especially those classified as EPT and Diptera, are among the most sensitive organisms to pesticides in freshwater, as evidenced by many monitoring and micro-/mesocosm studies. The toxicity of pesticides to each species, especially acute and short term, is useful as basic information to evaluate its impact on the macroinvertebrate community. However, short-term toxicity shows considerable variation even in one species, most likely due to the usage of individuals collected from different sources, which may have different exposure histories to contaminants under the various experimental conditions. Among EPT and Diptera, the non-biting midge Chironomus sp. is the only known standardized species with specific preparation procedures and experimental conditions. Although evaluation methods for the mayfly and caddisfly species have been proposed, improved standardization is required for each order.

Although the primary metabolic profiles of pesticides, such as ester hydrolysis and oxidation, seem to be common to other aquatic species, the relevant information for aquatic insect larvae is fragmentary and mostly limited to OC and OP pesticides in Chironomus spp. The metabolism, at least in EPT, should be examined further to consider the role of species sensitivity in toxicity and accumulation for each chemical class of pesticide. The chemical identification of an aglycon or the metabolite itself via secondary conjugation reactions should be conducted by chromatographic and instrumental analyses. In addition to phase-I enzymes, the investigation of phase-II enzymes such as glucuronyl- and sulfo-transferases by applying both biochemical and transcriptional methods is required.

More species and pesticides have been examined for bioconcentration/-accumulation than for metabolism, but the apparent BCF/BAF/BSAF values obtained under the different conditions (larval stage, temperature, period of exposure, and presence/absence of sediment) are difficult to interpret without metabolic information. Further, the kinetic analysis of uptake, metabolism, and depuration of pesticides should be investigated at least for the EPT and Diptera. Since the existing studies on toxicity, bioconcentration/-accumulation, and metabolism have not been always conducted with the same combination of test species and pesticides, this information cannot be utilized to completely explain pesticide toxicity in an individual species. Therefore, in addition to the standardization of species and experimental methods, studies using the same species and comparing the body burden of a pesticide with its toxicity are required.

Finally, the combined toxicity of pesticides should be further investigated at least on pesticide-sensitive orders, not only in laboratory studies but also using higher-tier micro-/mesocosms. The former studies will help to find a potential change in toxicity for a mixture, and the latter conducted in accordance with good agricultural practice are useful to help understand realistic impacts of the most probable mixtures, taking into account the environmental behavior of each pesticide. Mechanistic studies, explaining synergism or antagonism for combined toxicity, are required, using both biochemical and transcriptional approaches. Concerning the micro-/mesocosm studies, the biological impacts should be further examined in relation to the body burden of a pesticide and its metabolites by using a radio-labeled pesticide if possible and/or applying the high-sensitivity LC(GC)-MS/MS analysis.

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