Environmental applications of earthworm esterases in the ...

J. Pestic. Sci., 35(3), 290–301 (2010) DOI: 10.1584/jpestics.R10-08

Review (Special Topic)

Environmental applications of earthworm esterases in the agroecosystem Juan C. SANCHEZ-HERNANDEZ* Laboratory of Ecotoxicology, Faculty of Environmental Science, University of Castilla-La Mancha, Avda. Carlos III, 54071, Toledo, Spain. (Received April 24, 2010; Accepted June 12, 2010)

Chemical control of pests is still necessary in agriculture, despite the growing efforts to introduce biocontrolbased strategies. Many studies have evidenced the harm of pesticide side-effects on natural populations of pest enemies and other non-target organisms. Moreover, pesticide-contaminated soils can act as a secondary pollution source causing contamination in environmental compartments of critical concern to public health (water resource). These environmental risks need to be assessed and monitored for decision-making related to the post-authorization management of pesticides. Under these considerations, earthworm esterases can be a suitable tool for these regulatory and environmental purposes. Herein, it is suggested the use of earthworm esterases as biomarkers to be included in a field toxicity test. Furthermore, the potential role of gut carboxylesterases (CEs) in the modulation of pesticide toxicity is discussed in view of their contribution to the natural tolerance of earthworms to pesticides, and consequently the appropriate selection of earthworm species for regulatory toxicity testing. Finally, it is postulated that CE secretion into the earthworm gut could be an environmental friendly methodology in the enzymatic bioremediation of pesticide-contaminated soils. © Pesticide Science Society of Japan

Keywords: earthworms, carboxylesterases, biomarkers, enzymatic bioremediation.

Introduction The concept of soil quality has been a matter of continuous scientific debate. Parr et al.1) defined it as the “capacity or capability of a soil to produce safe and nutritious crops in a sustained manner over a long term, and to enhance human and animal health, without impairing the natural resource base or adversely affecting the environment”. The Soil Science Society of America defined soil quality as the “capacity of a specific kind of soil to function, within natural or managed ecosystem boundaries, to sustain plant and animal productivity, maintain or enhance water and air quality, and support human health and habitation”.2) These definitions agree to include both agronomic and environmental qualities of soil function. Therefore, soil degradation describes the situation in which the capacity for crop production and maintenance of healthy ecosystem is lost. Natural phenomena and human ac* To whom correspondence should be addressed. E-mail: [email protected] Published online August 7, 2010 © Pesticide Science Society of Japan

tivities cause soil degradation.3) The European Union (EU) has identified erosion, decline in organic matter, contamination, salinisation, compaction, loss of biodiversity, soil sealing, landslides and flooding as major threats to soil quality.4) Among them, contamination is probably the most concerned and dangerous phenomena to human health, because human are a significant vector for soil contamination and, in turn, humans are a biological receptor of soil toxicants. Agricultural pesticide applications can lead to a set of environmental risks that justify the monitoring of pesticide exposure and toxicity in soil system.5) Pesticide-contaminated soils can behave as a secondary pollution source causing contamination in environmental compartments of critical concern to public health such as groundwater or streams.6–9) Likewise, chemical control of pests causes side-effects on natural populations of pest enemies and other non-target organisms.5,10–12) Integrated pest management (IPM) is a strategy of growing use in current agriculture practices. However, this approach does not totally exclude the use of agrochemicals.5) In fact, the International Code of Conduct on the Distribution and Use of Pesticides defines IPM as “the careful consideration of all available pest control techniques and subsequent integra-

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tion of appropriate measures that discourage the development of pest populations and keep pesticides and other interventions to levels that are economically justified and reduce or minimize risks to human health and the environment”.13) According to this definition, a rational and justified use of pesticides is compatible with the implementation of biological measurements to combat pests. However, pesticide side-effects on natural populations of beneficial predators should be taken into account for pesticide regulatory purposes.10) Moreover, indirect effects of pesticide contamination such as reduction of prey populations can have negative consequences for top predators.14,15) Despite the introduction of measurements for reducing pesticide and fertilizer inputs in the agroecosystems (e.g., introduction of agri-environment schemes in the EU—Council Regulation No. EEC 2078/92), production and use of these chemicals have not decreased significantly. As an example, Fig. 1 shows the worldwide pesticide use per unit of agricultural land surface. Asiatic countries and some EU member states have pesticide uses higher than 0.2 tons/km2 of agricultural land.16) Among the vast variety of agrochemicals, organophosphorus (OPs) and carbamates (CBs), so-called anticholinesterase (anti-ChE) insecticides, as well as the synthetic pyrethroids (SPTs) continue to be an important group of crop protection products in agriculture, especially in some Asiatic countries such as India.17) In the last decades, trends in pesticide consumption reveal that chemical control of pests is still widely used in the agriculture. For example, a third of the Organisation for Economic Co-operation and Development (OECD) member countries have increased their pesticide consumption in the period between 1990–1993 and 2001–2003.18) Likewise, the extended use of certain types of crops (e.g., corn and soybean) for biofuel production could increase the

demand of agrochemicals with a subsequent risk for the environment.19–21) In view of this scenario, it is rather justified that field monitoring of pesticide side-effects and the development of methodologies for reducing pesticide residues in soil be among the priorities of pesticide management in the agroecosystem. The purpose of this review is to offer a viewpoint on the environmental applications that earthworm esterases can provide in that context. Accordingly, two major issues are discussed: the use of cholinesterases (ChEs) and carboxylesterases (CEs, EC 3.1.1.1) as biomarkers for field monitoring of pesticide exposure. Particular attention will be put in the CE activity because of its potential role in the modulation of acute toxicity of pesticides. Family, it is suggested the potential environmental application of earthworm gastrointestinal CEs as an enzymatic bioremediation system of pesticide-contaminated soils. The reason for such an analysis with earthworms solely is supported by the ecological, toxicological and agronomic importance of these organisms in the agroecosystem. Earthworms are key components of the soil system contributing to improve its function and structure,22) with beneficial properties for plant growth and health.23) Furthermore, earthworms are the favorite prey of many invertebrate (centipedes or beetles) and vertebrate (gulls, starlings, magpies, moles, foxes, shrews or badgers) predators.24) In environmental toxicology, these annelids are the recommended organisms in soil toxicity testing and in the field monitoring of metal pollution.25,26) Thus, many reviews have examined different aspects of earthworm ecotoxicology such as biomarkers,27–29) or practical issues when these organisms are used in toxicity testing.30,31)

Historical Overview of Earthworm Esterases as Ecotoxicological Biomarkers 1.

Fig. 1. Pesticide use in OECD countries. Data were taken from OECD (http://dx.doi.org/10.1787/703206110145).16)

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The Past

Early investigations on earthworm esterases were primarily aimed to describe the enzymological properties of these hydrolases. The use of selective substrates and specific inhibitors for mammalian ChE activities, i.e., acetylcholinesterase (AChE, EC 3.1.1.7) and butyrylcholinesterase (BChE, EC 3.1.1.8), enabled two ChE activities in Eisenia fetida to be distinguished, although overlapping of mammalian ChEs-type features was observed in both ChEs.32) Incubation of the earthworm homogenate in the presence of carbaryl allowed to distinguish both ChEs.33) The existence of a carbaryl-resistant ChE accounted for the survival of E. fetida and Eisenia andrei when they were exposed to 700 m g carbaryl using the paper contact test,34) compared to Eisenia veneta that does not have this carbaryl-resistant ChE.35) These two ChEs in E. fetida was later corroborated by Scaps et al.,36) although Andersen et al.37) reported the presence of a ChE activity solely in this earthworm species. Esterase characterization has also been performed with other earthworm species. For example, three ChEs were identified in the soil-dwelling

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species Allolobophora (⫽Aporrectodea) caliginosa, corresponding probably to multiple forms of an AChE because of its preference for acetylthiocholine (AcSCh, Km⫽32 m M) or propionylthiocholine (PrSCh, Km⫽48 m M) compared to butyrylthiocholine (BuSCh, Km⫽152 m M).38,39) Although less investigated, earthworm tissues display also CE activity. Haites et al.40) found multiple CE isozymes with molecular weights raging between 60,000 and 72,000 in muscular body wall, pharynx, intestine and reproductive tissues of Lumbricus terrestris using in-gel staining native polyacrylamide gel electrophoresis.

2.

Ongoing

In the current decade, earthworm esterases have been investigated as potential biomarkers of pesticides exposure, although description of their enzymological properties is still a matter of intense work. Caselli et al.41) suggested that ChE activity in E. andrei is a mammalian-type AChE. This esterase hydrolyzed the substrates AcSCh (Km⫽140 m M; Vmax⫽ 41 mU mg⫺1 protein) and PrSCh (Km⫽180 m M, Vmax ⫽46 mU mg⫺1 protein) equally well, and no enzyme inhibition by substrate excess was observed. In addition, BW285c51 (1,5bis(4-allyldimethylammoniumphenyl)pentan-3-one dibromide) was a potent inhibitor of ChE activity, whereas isoOMPA (tetraisopropyl pyrophosphoramide) had no effect even when BuSCh was used as the substrate.41) Contrary to the observations by Stenersen et al.,35) who identified a carbaryl-resistant ChE activity in E. andrei, Caselli et al.41) found that 10⫺6 M of carbaryl fully inhibited (90% inhibition) the ChE activity of this earthworm species. Ribera et al.42) reported a strong inhibition of ChE activity in E. andrei exposed for 2, 7 and 14 d to soils spiked with carbaryl (12–50 mg kg⫺1). However, they found a residual ChE activity (11–24% compared to controls), that was attributed to the carbaryl-resistant ChE activity identified by Stenersen.32,33) In a related study, Gambi et al.43) examined the response of E. andrei ChE activity to different scenarios to carbaryl exposure and type of carbaryl formulation. In the paper contact OECD test, carbaryl caused a ChE inhibition higher than 50% in all treatment groups (6.3⫻10⫺4 to 6.3 mg cm⫺2) and the carbamylated enzyme recovered its normal activity within 5 days. Most likely, spontaneous reactivation of the carbaryl-inhibited enzyme explains this rapid recovery of ChE activity because the CB-ChE complex is unstable.44) In contrast, when carbaryl was added to natural soils, no recovery of the carbamylated ChE activity was found during the period of observation (5 d). In that investigation, ChE activity was not fully inhibited by carbaryl concentrations as high as 48 mg kg⫺1 (soils spiked with carbaryl solution) or 429 mg kg⫺1 (soils spiked with Zoril 5). Similar to the study by Ribera et al.,42) the carbaryl-resistant ChE isozyme described by Stenersen32,33) could account for this observation. Most of the toxicological studies on the “insecticide-ChE” interactions have been performed with the epigeic earth-

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worms, E. fetida and E. andrei. However, similar investigations are still needed with soil-dwelling earthworms (endogeics and anecics) which have a permanent contact with soil, and further, they are common species in the agroecosystem.45) A comparative study by Rault et al.46) demonstrates that generalizations about earthworm ChEs should be taken with cautions. Besides species-specific differences in ChE activity (from 72 mU mg⫺1 protein for Allolobophora chlorotica to 260 mU mg⫺1 protein for Lumbricus castaneus), the response of this esterase to selective substrates and inhibitors was markedly different among the species, suggesting the existence of multiple ChEs with atypical behaviors compared to those displayed by the mammalian AChE and BChE activities.46) Recovery rate of phosphorylated ChE activity has been also a matter of continue research because of its implications in biomonitoring. Transient response of biomarkers following exposure to toxicants often limits their application in the field. Earthworm ChE activity generally shows a slow recovery following acute exposure to OP compounds. For example, the tropical earthworm Drawida willsi recovered its normal ChE activity within 45 and 75 d of continuous exposure to carbofuran and malathion, respectively.47) This slow recovery of D. willsi ChE activity could be the result of a reduced pesticide degradation rate in the test soils and a low rate of new enzyme synthesis. In a laboratory study aimed to examine the food web transfer of OP poisoning from earthworms to the common shrew (Sorex araenus), individuals of E. fetida were exposed to artificial soils spiked with dimethoate (150 and 200 mg kg⫺1 fresh weight) for 48 h and then transferred to clean soil. Cholinesterase activity of earthworms was still depressed (35% of residual ChE activity compared to controls) 40 d following acute exposure.48) The same species was used by Aamodt et al.49) to investigate the recovery rate of the two ChEs described by Stenersen.32,33) Whereas the carbaryl-sensitive ChE form recovered its normal activity within one month following 48-h exposure to chlorpyrifos (240 mg kg⫺1 dry weight)-spiked soils, the carbaryl-resistant isozyme remained strongly inhibited for 84 d. This study demonstrates that dynamic of ChE inhibition and recovery in the earthworm E. fetida should be analyzed considering both ChE forms individually to gain a better understanding of “OPChE” interaction. Species-specific differences in the recovery rate of phosphorylated ChE activity must be taken into account for field monitoring purposes. For example, Rault et al.50) found that A. chlorotica exposed to soils spiked with 1 mg kg⫺1 parathion recovered its normal esterase activity after 64 d of transferring individuals to clean soil, however, ChE activity of A. caliginosa remained still depressed (24–37% inhibition compared to controls). Comparatively, a few investigations have examined the sensitivity of CE activity to OP insecticides, as well as its recovery following OP inhibition. For example, CE activity of L. terrestris presented a marked tissue-specific sensitivity to

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chlorpyrifos-oxon, paraoxon or dichlorvos, which was also dependent on the substrate used in the esterase assay.51–52) The occurrence of multiple isozymes may explain the OP- and substrate-dependent variations of the enzyme activity. These in vitro results were later reproduced in a microcosm laboratory trial, where L. terrestris was exposed during 48 h to chlorpyrifos (3–48 mg kg⫺1 dry weight)-spiked soils. Inhibition of CE activity was tissue-specific, and the substrate 4-nitrophenyl valerate (4-NPV) revealed a highly sensitive CE activity in the body wall muscle, seminal vesicles (male reproductive organ), pharynx and foregut.53,54) In addition, recovery of 4-NPV-CE activity was little or absent in these tissues during the 35-d period following the acute exposure to chlorpyrifos.

3.

Future potential approaches

Some reviews have already indicated what areas of biomarker research with earthworms demand further investigation.27–29,55) Earthworm esterases have had little attention in ecotoxicology compared to other biomarkers such as metallothionein induction or the lysosomal membrane fragility. Moreover, earthworm biomarkers have been extensively investigated in response to metal contamination. The limited number of field and laboratory studies involving pesticides have used the responses of ChE and glutathione S-transferase activities as primary sublethal biochemical endpoints.56) Yet, to the best of my knowledge, no field studies with earthworms have used the measurement of CE inhibition as a complementary biomarker of anti-ChE pesticides.57) In this way, the developments achieved with aquatic and vertebrate esterases could be explored with earthworms to increase the understanding and use of esterases in the assessment of pesticide exposure. Thus, it is highlighted two areas of future research with earthworm esterases: 1) the validation of earthworm ChE and CE activities in the field and 2) the development of complementary techniques for supporting the esterase responses in the field. 3.1. Need for field validation of earthworm esterases Field survey of pesticide impact on non-target biota is, or should be, an essential step of the ecological risk assessment (ERA) scheme for plant protection products. Indeed, Newman et al.58) suggested that post-authorization monitoring of pesticides in the field should be performed to provide more accurate predictions of adverse effects on terrestrial ecosystems. However, field studies to monitor pesticide exposure and toxic effects have not always provided the correct answers; often because of the lack of right questions59) or due to the complexity of field toxicity testing.60) Earthworms are recommended organisms in soil toxicity testing. Although detailed guidelines for conducting acute toxicity testing with earthworms are currently available,61) no similar guidance exists for carrying out and interpreting semi-field or field surveys. However, Edwards60) proposed a well-designed and standardized—although flexible for some of its variables—field toxic-

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ity test to confirm adverse effects from pesticides previously observed in the laboratory. As with other many organisms, earthworm ChE activity generally responds in a dose-dependent manner to OPs and CBs under laboratory conditions. However, very few field studies have addressed the use of ChE and CE activities as biomarkers. For example, the earthworm A. caliginosa was used to test for ChE inhibition following spraying with diazinon and chlorpyrifos at the recommended field application rates.62) No inhibition of ChE activity was observed during the monitoring period (28 d). However, Reinecke and Reinecke63) recorded significant ChE inhibition in caged A. caliginosa in a plum orchard one and two weeks following chlorpyrifos spraying. Inhibition of ChE activity in A. chlorotica was related to the pest control strategies used in apple orchards in France.64) A level of ChE inhibition higher than 30% was recorded in earthworms collected from orchards with conventional and IPM agricultural practices compared to individuals sampled in organic agriculture or abandoned fields. Comparisons of ChE responses between these field studies are rather difficult because of the marked differences in the field design. A complementary approach to the field survey of pesticide impact on natural populations of earthworms could be the use of mesocosms (outdoor manmade systems simulating field conditions as closely as possible). The flexible field toxicity testing proposed by Edwards60) could be standardized (e.g., earthworm species, use of indigenous or sentinel earthworms, plot design features, esterase assay conditions) to this purpose. Thus, sublethal effects of pesticides using earthworm esterases could be tested under an environmentally realistic scenario. 3.2. Need for complementary methods of pesticide exposure Many researchers agree that the use of a set of biomarkers covering multiple levels of biological organization and screening multiple biological targets is an appropriate strategy for a better assessment of contaminant exposure and effect.65–67) Yet, multivariate statistical analysis68) and biological health indexes69) have improved the interpretations of multiple responses of biomarkers. In addition, chemical analysis of residues in both abiotic and biotic samples should be a solid line-of-evidence complementing the biomarkers’ outcomes. However, biomarkers often replace analytical chemistry mainly because the former approach enables simple, low-cost and rapid assays, whereas the latter require sophisticated equipments at high cost. In the case of ChE inhibition, complementary biomarkers are recommended because of the short half-life of OP within the organism, the high interindividual variation of normal ChE activity and the transient ChE response in some organisms following pesticide exposure.62,64,70) One of the most popular methodologies for determining the cause for ChE inhibition is the chemical reactivation of the esterase activity in the presence of nucleophylic compounds such as oximes. This approach has been used with birds71–77) and reptiles.78,79) In earthworms, the oxime pyridine-2-aldoxime methochloride (2-PAM) reactivated in

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vitro the phosphorylated ChE activity of E. fetida and L. terrestris, but this oxime lost its reactivating power after one week of ChE inhibition.80) Indeed, a half-aging time (AT50)— defined as the time elapsed since in vitro ChE inhibition at which 2-PAM reactivated the 50% of normal activity—of 22 and 14 h was estimated for ChE activity of L. terrestris and E. fetida, respectively.80) This observation was previously documented by Andersen et al.,37) who found that the AT50 of E. fetida ChE activity previously inhibited with paraoxon (5⫻10⫺6 M) was 12 h using 5⫻10⫺5 M 2-PAM. These in vitro results were reproduced in an in vivo trial with the species L. terrestris exposed for 48 h to chlorpyrifos-spiked soils. Chemical reactivation of phosphorylated ChE activity using 5⫻10⫺4 M 2-PAM or obidoxime occurred during the first week following exposure to the chlorpyrifos-spiked soils.53) Therefore, this methodology becomes a complementary approach to the comparison of ChE inhibition levels. However, further studies are needed to examine the real potency of this methodology under field situation where multiple and frequent pesticide applications are a common practice. A few studies have also explored this methodology with CEs. For example, Maxwell et al.81) reported a full reactivation of rat plasma phosphorylated CE activity in the presence of diacetylmonoxime. Chemical-induced reactivation was efficient with alkoxy-containing OPs, whereas the CE aged when inhibited by aryloxy-containing OPs. In a toxicological context, the use of oximes might help to understand the dynamic of “OP-esterase” interaction (spontaneous reactivation, aging and synthesis of new enzyme). Despite this growing concern in field or semi-field studies using earthworms to examine the impact of pesticides on the biological component of soil, factors other than the environmental variables should be taken into account as potential sources of variation in the biomarker responses. Among the biological factors, BChE and CE activities are accepted as modulators of pesticide acute toxicity in mammals. These esterases play a pivotal role as bioscavengers of OP compounds thus reducing the acute toxicity of these chemicals.82–85) The high affinity of these esterases to bind OPs, the tissue where they are preferentially expressed, and their high activity levels could be determinant factors in the protective mechanism exerted by these enzymes. For example, brain AChE and serum ChE activities showed a significant non-linear relationship in reptiles86) and small wild mammals70) experimentally exposed to parathion and dimethoate, respectively. Serum esterase activity was more severely inhibited by the OP compared to brain AChE activity. This marked difference in esterase sensitivity suggested a ‘protective’ role attributable to the serum esterases when these vertebrates are exposed to sublethal doses of OPs.70,86) Similar investigations might be performed with earthworms to determine to what extent esterases expressed in tissues of toxicological concern such as the gastrointestinal epithelium modulate the impact of OPs and CB pesticides on nervous AChE activity. Moreover, a comparative

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study examining species-specific differences in CE activity would allow to obtain important information on the ability of earthworms to naturally tolerate pesticides or in the selection of the suitable earthworm species for regulatory toxicity testing. Recently, some laboratory studies have evidenced a high CE activity in the gastrointestinal tissue and content of the earthworm L. terrestris compared to other tissues and organs.51,52) In the following two sections, some potential toxicological and environmental implications derived from this gastrointestinal CE activity are detailed.

Gastrointestinal CEs: An Enzymatic Barrier Against Pesticide Uptake Carboxylesterase activity is particularly abundant in the liver of mammals, although they are also present in other tissues such as gonads, kidney, plasma or intestine.87) Yet, expression of this esterase in the intestinal epithelium is believed to be a defensive barrier against esterified xenobiotic uptake such as SPTs.88) Earthworm gastrointestinal tract displays a high CE activity towards naphthyl or nitrophenyl esters.52) In addition, CE activity was surprisingly abundant in the gut contents (ingested soil) collected from the “crop-gizzard-foregut” segment of the L. terrestris alimentary canal (a -NA-CE activity⫽5.5⫾1.6 to 8.5⫾2.8 U mg⫺1 protein; 4-NPA-CE activity⫽5.3⫾1.9 to 6.0⫾2.1 U mg⫺1 protein, mean⫾standard deviation) compared to the esterase activity in the corresponding gut epithelium (a -NA-CE activity⫽0.6⫾0.11 to 0.9⫾0.10 U mg⫺1 protein; 4-NPA-CE activity⫽0.7⫾0.10 to 0.9⫾0.14 U mg⫺1 protein) (Fig. 2). Moreover, this luminal CE activity was sensitive to inhibition by dichlorvos (IC50 s⫽6.6 to 8.6 nM) and paraoxon (IC50 s⫽7.4 to 8.4 m M). Intestinal secretion of CEs seems a regular phenomenon in invertebrates. For example, Turunen and Chippendale89) found seven esterases and two lipase-like enzymes in the midgut epithelium of the southwestern corn borer (Diatraea grandiosella); some of them were secreted into the gut lumen during feeding. The intestinal esterases in this insect species play a major role in the digestion of lipids. In-gel staining electrophoresis was used by Geering and Freyvogel 90) to investigate changes in the esterase profile of female mosquitoes (Aedes aegypti) under different physiological stages (feeding and emerging processes). Blood uptake by mosquitoes caused an increase of esterase activity in the midgut epithelium and a luminal esterase activity was observed two days after blood ingestion. In that study, a potential role in lipid metabolism was also attributed to the increase of esterase secretion. Multiple CEs were also found in the digestive juice of the spider Tegenaria strica,91) some of which were able to hydrolyze long-chain acylglycerols such as b naphthyl laureate, p-nitrophenyl laureate or p-nitrophenyl palmitate, although the hydrolysis rate was low in comparison to that for short-chain naphthyl esters. In invertebrates, gut CE activity seems therefore to play a significant function in the digestion of lipids. This hypothetical physiological role

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Fig. 3. Conceptual model illustrating how gastrointestinal (tissue and luminal) CE of Lumbricus terrestris may play a pivotal role in the reduction of OP and SPT uptake. Secretion of CEs from gut epithelium can contribute to extracellular hydrolysis of SPTs or stoichiometric binding of “oxon” metabolites formed from phosphorothioate-type OPs (“tion” form). Bioactivation (oxidative desulfuration) of OPs is catalyzed by either CYP450-dependent monooxygenases (or related enzymes) or microorganisms. Phosphate- and phosphonate-type OPs do not require such a bioactivation to inhibit esterase activity.

Fig. 2. Carboxylesterase activity (mean and standard deviation) in the gut tissue (bars) and ingested soil (circles) of the earthworm Lumbricus terrestris. Esterase activity was measured using the substrates a -NA, (graph A) and 4-NPA, (graph B). Native polyacrylamide gel electrophoresis of tissue homogenates (lanes 1–4) and soil extracts (lanes 6–8) revealed multiple esterase (ES1-ES11) isozymes (graph C). Figure taken from Sanchez-Hernandez et al.52) with kind permission from Elsevier.

could explain their gastrointestinal secretion. It could be postulated that the gut CE activity (luminal and tissue fractions) of L. terrestris play an important mechanism to reduce the uptake of pesticides (Fig. 3). In the case of OP insecticides, it is well-known that the highly toxic oxon

metabolites of phosphorothioate-type OPs show a higher affinity for the active site of CEs than their parent compounds.92) Such an interaction leads to a stable inhibitorenzyme complex, which results an effective stoichiometric mechanism of OP detoxification. The oxidative desulfuration needed for OP bioactivation can take place in the enterocyte by CYP450-dependent monooxygenases (or related enzymes),93,94) or by microorganisms present in the soil or in the earthworm gut.95,96) Indeed, González Vejares et al.54) reported a high inhibition of both ChE and CE activities in the first segment (pharynx) of L. terrestris gut after two days of exposure to soils spiked with 12 and 48 mg kg⫺1 chlorpyrifos (31–56% ChE inhibition, 41–50% a -NA-CE inhibition and 78–84% 4-NPV-CE inhibition). They suggested that such a depression of esterase activity could be due to absorption of chlorpyrifos-oxon from the ingested soil. Earthworms are exposed to soil contaminants by two ways, i.e., the skin and the gastrointestinal epithelium. Exposure to non-polar organic chemicals such as OP pesticides occurs mainly by ingestion of contaminated soil.97) Thus, high levels of CE activity in the gut may contribute to decrease OP uptake and toxicity. For example, Henson-Ramsey et al.98) examined the toxic effects and body concentration of malathion in the earthworm L. terrestris using two different exposure setups (malathion-spiked soils and the paper contact OECD test). Symptoms of OP intoxication (coiling) were more severe in earthworms exposed to the insecticide using the contact test (dermal exposure). In

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addition, a higher concentration of malathion was found in earthworms exposed to 50 and 100 m g cm2 using the contact test compared to the same concentrations applied on soil surface. The authors suggested that hydrolysis of the malathion by CEs, among other toxicokinetic factors, could explain the marked differences in malathion body burden and toxicity.98) Similarly, Gambi et al.43) examined the inhibition and recovery of AChE activity in E. andrei exposed to carbaryl using three exposure scenarios: the OECD contact test, natural soils spiked with carbaryl and natural soils spiked with Zoryl 5 (5% carbaryl). Similar to the results by Henson-Ramsey et al.,98) the impact of carbaryl in terms of AChE inhibition was more prominent in earthworms exposed to the carbamate through the OECD contact test than those exposed to carbaryl-spiked soils. Although multiples factors evidently contribute to pesticide toxicity in these examples, the expression and secretion of CEs in the gastrointestinal tract (Fig. 2), could reduce acute toxicity when the alimentary canal is the major uptake route. In the case of SPT insecticides, no available data exist about pyrethroid hydrolysis by earthworm CE. However, the hydrolytic activity of CEs using pyrethroids as substrates has been investigated in the human and rat small intestine. The CE activity of intestinal microsomes efficiently hydrolyzed trans-permethrin,99) whereas low or absent hydrolytic activity was observed with deltamethrin or bioresmethrin.88) Moreover, CE activity is regionally located along the length of the intestinal tract. For example, this esterase displays a higher activity in the jejunal region as compared to duodenum or ileum,100–102) which indicates that this intestinal region has an active role in the detoxification of esterified compounds. A spatial distribution of CE activity along the gastrointestinal tract has also been described for L. terrestris.51,54) The intrinsic susceptibility of an individual to pesticide toxicity can be altered by mutations in genes involved in detoxification processes. This is a common and well-described phenomenon in some pest species. An overexpression of CE activity, among other pesticide-metabolizing enzymes, account for insecticide resistance in many pest species.103) Moreover, biological factors such as type of diet can be a determining factor in the modulation of CE activity with subsequent impact on OP toxicity. This relationship has been investigated in birds and mammals. In general, low liver CE activities are found in avian and mammalian carnivorous species compared to herbivorous or omnivorous species.104) The role of CE activity in the detoxification of lipophilic esters present in the diet (e.g., plant secondary metabolites) could account for this marked difference in avian blood CE activity.105) Wheelock et al.57) suggested that CE activity could be used as a biomarker of susceptibility because of its significant role in pesticide detoxification. In their opinion, the establishment of normal levels of CE activity and its capability to hydrolyze SPTs or bind anti-ChE pesticides would allow the development of a susceptibility index. To develop such an index, fac-

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tors contributing to interindividual (and interspecific) variations of CE activity must be identified. Again, past studies performed with invertebrates and birds could provide some clues. Those works strongly suggest that gut esterases of arthropods are actively involved in the metabolism of lipid materials. Distinction between esterases and lipases has been traditionally attributed to the substrates that these enzymes preferentially cleavage. Esterases hydrolyze short-chain acylglycerols, whereas long-chain acylglycerols are the preferred substrates of lipases.106,107) However, invertebrate esterases are able to hydrolyze typical substrates of lipase activity.91) If we assume that earthworm esterases participate in lipid digestion, then lipids could modulate the expression and secretion of CEs in the gut. In support of this hypothesis, some investigations suggest that enhanced CE activity in blood or gastrointestinal tract is highly dependent on a lipid-rich diet. For example, van Lith et al.108) demonstrated that plasma and intestinal esterase activity (using 4-NPA as the substrate) in male rats changed as a function of the type of fat or the concentration of lipids in the diet. Plasma esterase activity increased with the fat content in the diet. Similarly, an increase of esterase activity was found in both the duodenum and jejunum as the concentration of ingested fat enhanced. The type of fat also caused a different response in plasma esterase activity. Coconut-fat diets yielded the lowest response in plasma esterase activity, whereas olive oil led to a positive dose–response relationship between olive oil amount and esterase activity.108) Comparable results were previously observed with mice.109) A large increase of esterase activity (ES-2) was found in the lymph and serum of mice administrated with a lipid-rich infusion by duodenal fistula. It was suggested that this esterase isozyme is involved in lipid absorption from the duodenum. Taken together, these investigations encourage future research on the impact of lipid-rich diets on expression and secretion of CEs in the earthworm gut in order to examine their role as biomarkers of susceptibility.

Intestinal Secretion of CEs: A Promising Tool for Enzymatic Bioremediation Many soil remediation procedures are currently available, involving both engineering (chemical and physical treatments, solidification, stabilization or thermal treatment) and biological (aerobic and anaerobic bioremediation, phytoremediation or rhyzoremediation) approaches.110) However, methodologies that involve microorganisms and plants—i.e., bioremediation—are desirable because of their low cost, in situ treatment and apparently minimal environmental impact.111,112) Bioremediation of soils contaminated by low-persistent pesticides such as OPs frequently use indigenous or non-indigenous microorganisms, or even the inoculation of genetically engineering microorganisms.95,113–115) However, microorganismassisted bioremediation presents several limitations related to biochemical, physiological and ecological aspects of the microorganisms (Table 1). Bioremediation

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using microorganisms often requires long times (months) to be effective because it is highly dependent on the microorganism growth rate.116) Addition of organic matter-rich substrates and soil aeration help to solve this limitation. Sensitivity to pesticides, or their metabolites, is another critical factor in the bioremediation with microorganisms. For example, degradation of the OP insecticide chlorpyrifos yields the main metabolite 3,5,6-trichloro-2-pyridinol. This metabolite inhibits the proliferation of soil microorganisms, thus metabolism of the parent compound is reduced.95,117) An attractive alternative is the direct inoculation of pesticide-metabolizing enzymes such as CEs or phosphotriesterases (EC 3.1.8).118) Such an approach, called enzymatic bioremediation, is generally supported by genetic engineering.118,119) However, the use of free-enzyme preparations can become a tedious and expensive method that involves three main steps116,118): the identification of a suitable enzyme from the organism, the isolation Table 1. Bioremediation strategy Microorganismsa)

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of the gene encoding the enzyme of interest, and the improvement of enzymatic properties in the laboratory when necessary. Exploration of other natural biological sources of pesticide-metabolizing enzymes is therefore a matter of future concern. Another practical approach to bioremediation would be to use soil-dwelling earthworms. These organisms offer several qualities as a potential ‘biological reactor’ for remediation of pesticide-contaminated soils. They have already been used as biostimulation agents in the in situ bioremediation of soils contaminated by non-polar chemicals such as polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and herbicides.120) As mentioned previously in this review, L. terrestris display a high CE activity in its gut lumen (Fig. 2). This extracellular CE activity could be an efficient enzymatic bioremediation system for removing pesticides from the ingested soil (Fig. 3). The use of soil-dwelling earth-

Main strengths and limitations of the bioremediation methodologies for recovering pesticide-contaminated soils

Advantages – High capacity to degrade a wide variety of pesticides. – Degradation of contaminants can be significantly increased by addition of nutrient-rich amendments and aeration (biostimulation), or by inoculation of non-native microorganisms with enhanced catalytic capability (genetically engineered) for pesticide degradation.

Limitations – High pesticide concentrations, and their metabolites, can reduce microbial biomass because high toxicity. – Environmental impact of microorganism-assisted bioremediation need to be explored, particularly when genetically modified microorganisms are used. – Microorganisms require optimal conditions for growth, and often bioremediation results a slow (weeks to months) or incomplete process. – Occasionally, microbial metabolism of pesticides may cause highly toxic metabolites to non-target organisms. – Sorption of pesticides to organic matter can reduce their bioaccessibility by microorganisms.

a)

Free enzymes

– Release of genetically engineered microorganisms into the environment is avoided. – Modification of catalytic properties of enzymes for enhancing pesticide degradation is viable. – Manipulation of the amount of the enzyme in the soil.

Earthworms

– They exert a biostimulation effect increasing therefore the biodegradation capacity of microorganisms. – Gut luminal esterases can operate jointly to soil microorganisms in the degradation of pesticide residues. – Biological factors (e.g., diet type) can lead to an overexpression of earthworm gut esterases. Their secretion in the gut could result in an esterase-enrich vermicompost or cast (?). – Earthworm activity can increase pesticide bioaccessibility to degradation.

– High costs of production. – Need of continuous monitoring because of protein degradation in soil. – Genetically engineered enzymes often provide a limited substrate range. – Sensitivity of earthworms to toxic effects from pesticides (earthworm avoid contaminated soil). – Soil features (e.g., moisture, temperature, pH) can reduce earthworm survival. – Lack of information about the real impact of gut luminal CEs in the degradation of pesticides present in the ingested soil.

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worms to remove pesticide residues in soil provides some advantages from an environmental and agronomic viewpoint compared to the bioremediation using microorganisms or free enzyme preparations (Table 1). Earthworms increase soil nutrient content, soil aeration and stimulate microorganism biomass; which are favorable conditions for microorganism-assisted bioremediation. Furthermore, the secretion of CEs in the earthworm gut offers a complementary system for removing soil pesticides. Interactions of non-polar pesticides with soil organic matter determine their environmental fate and microbial degradation. It is tempting to speculate that earthworm burrowing and feeding activities might increase the accessibility of luminal CEs to pesticides that are entrapped in the organic matter fraction. However, the sensitivity of earthworms to pesticides is the main limitation in their use as enzymatic bioreactors. A threshold of viability based on acute and chronic toxicity tests25) and avoidance behaviour bioassays56) should be firstly defined for the pesticides of interest. Alternatively, vermicompost (a nutrient-rich material that is the result of the biooxidative, fragmentation and stabilization of the organic matter by earthworm activity)121) could be an attractive source of extracellular CEs if compost’s earthworms (e.g., Eisenia species) are able to secrete these esterases as L. terrestris do. The use of this biological material is not novel as some studies have shown its capacity for pesticide breakdown. For example, Fernández-Bayo et al.122) reduced the persistence of the herbicide diuron in sandy loam soils by the addition of vermicompost. Stimulation of microorganism biomass and increase of organic matter in soil following vermicompost addition account for the impact of this type of amendment in pesticide degradation. Whether CEs are present in the vermicompost and they contribute notably to pesticide degradation is a stimulating matter of future research particularly when soil conditions or pesticide exposure levels are inhospitable to earthworm survival.

Conclusions Pesticide use continues being a key strategy to combat pests in the worldwide agriculture, despite it has led to serious risks to humans and the environment. Toxicity testing with earthworms is an important laboratory-based tier to support authorization of pesticides. However, what are the long-term effects of agrochemicals on non-target organisms? As Newman et al.58) suggested, monitoring of biological effects in the field where authorized pesticides are being used is a logical and recommended approach for determining the harmful side-effects of pesticides on biota. In this sense, earthworm esterases (ChEs and CEs) are a suitable tool for assessing pesticide impact in the below-ground system. However, a set of practical considerations should be taken into account when using earthworm esterases for monitoring anti-ChE pesticide exposure. First, selection of earthworm species is critical. Most of the laboratory studies have used Eisenia species to investigate the interaction between ChE activity and OP or CB pesti-

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cides,32,33,35–39,41–43,46,47) and similar studies are comparatively limited with soil-dwelling species.45) Indeed, sensitivity of ChE activity to pesticides and its recovery rate vary markedly with the species.50) Second, selection of the target tissue for esterase measurements is also an important element in a biomonitoring program. Although most of the studies use the whole organism for esterase assay, others have demonstrated that ChEs and CEs show a tissue-dependent distribution.51) Furthermore, response of esterase activities to anti-ChE pesticides is highly dependent on the tissue where they are expressed.53,54) Whenever possible, dissection of several tissues for determination of ChE and CE inhibition is recommended. Third, monitoring of biological effects of pesticides in the field generally use the inhibition of ChE activity as a sublethal endpoint.62–64) However, CE activity has demonstrated to be a complementary and highly sensitive biomarker of OP exposure.53,54) Moreover, this esterase activity seems to modulate the acute toxicity effects from pesticides. Parameters such as normal levels of activity, isozyme abundance and affinity to OPs/CBs could define the susceptibility of earthworms to these agrochemicals.57) Field studies are needed to validate the use of multiple esterases as biomarkers of exposure in combination with other methodologies such as chemical reactivation using oximes.80) Carboxylesterases are abundant in the gastrointestinal tract of L. terrestris.51) They show a high sensitivity to inhibition by OPs. In vitro studies have shown that these esterases are secreted by the gut epithelium, although gut microorganisms inhabiting the earthworm gut can also contribute to the total hydrolytic activity by CEs.52) This luminal CE activity might play an important role in reducing the acute toxicity of antiChE and SPT pesticides. Likewise, this luminal esterase activity provides an exciting field of research in the bioremediation of pesticide-contaminated soils. However, these suppositions need to be tested to know their real scope in an enzymatic bioremediation process, especially when secretion of esterases in the earthworm gut could be enhanced (e.g., lipidrich diet). Acknowledgements I thank Dr. Craig E. Wheelock for his invitation to review this topic, and Dr. Matthew Ross for a critical reading of a draft of this paper. This review was elaborated with financial support from the Ministerio de Ciencia e Innovación (CTM 2006-11828/TECNO, CTM 201019375/TECNO) and the Consejería de Educación y Ciencia (PCI080049-0228).

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