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a Dep. of Biological Sciences, Macquarie Univ., North Ryde NSW 2113 Australia
b Environment Protection Authority, NSW, located at EPA and Univ. of Technology, Sydney, Centre for Ecotoxicology, Westbourne Street, Gore Hill NSW 2065 Australia
c School of Pharmaceutical Sciences, Univ. of South Australia, Adelaide, SA, 5000, Australia
Corresponding author (chapmanj{at}epa.nsw.gov.au)
Received for publication October 8, 1999.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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Abbreviations: AChE, acetylcholinesterase • ANOVA, analysis of variance • OP, organophosphorus • PCCM, Pearson Correlation Coefficient Matrix
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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It is often assumed that the low bioconcentration potential of OPs by aquatic organisms rapidly reduces the risk of profenofos exposure in ecosystems (Eto, 1974; Yu and Sandborn, 1975) after application. Hence, evaluation of the rate of recovery of the ecosystem would permit adequate assessment of any long-term consequences of profenofos spraying.
Measurement of the concentrations of pesticide in spot samples of water may underestimate the exposure levels encountered by inhabiting aquatic organisms. However, an analysis of accumulated profenofos residues in the sediment or fish tissue may give a better indication of exposure (Nowak and Julli, 1991) and toxicological risk. In general, the most significant route of exposure for fish appears to be the direct uptake of insecticide from water (Tsuda et al., 1994). This study attempts to establish whether or not wild fish accumulate sufficient residues of profenofos for it to represent a risk to the aquatic system.
Acetylcholinesterase (AChE) activity in fish has been used as an indicator to monitor poisoning by OP pesticides (Coppage and Braidech, 1976; Zinkl et al., 1991). The use of AChE activity measurements in fish could be developed as a useful indicator, provided some prior knowledge of the profenofos use patterns is known. Correlation between depressed AChE levels and manifestations of toxic effects have been reported for other OPs (Post and Leasure, 1974; Lockhart et al., 1985). In the case of profenofos, a quantitative relationship needs to be developed to permit the use of AChE activity as an important indicator of the exposure–effect relationships. If residues of chemicals in tissues of organisms could be related to observable tissue damage (Connolly, 1985), it may be possible to increase the predictability of laboratory and field ecotoxicology data. Connolly (1985) related residues of diazinon to AChE levels as a measure of OP damage. This study also attempts to relate AChE effects to residues as a first step to evaluating whether the accumulation of profenofos may actually be causing harm to fish populations inhabiting the water bodies in the cotton-growing areas.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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River systems in this district are more difficult to study, as they are heavily regulated to supply irrigation water for cotton farming and there is a greater movement of some animal species. During the 1994 survey, two creeks were also included to assess the effect of profenofos spraying on river systems. The location of the lagoons, creeks, and some irrigation canals are detailed in Fig. 1 and descriptions of sites are in Kumar (1995). All lagoons and creeks except Lowana Lagoon were either adjacent to cotton fields, or in the case of Weetawaa Lagoon and Gunidgera Creek, had cotton fields nearby. Kerribee Lagoon also received tailwater from adjacent cotton fields, and irrigation canals were in the middle of fields. Lowana Lagoon was considered to be the reference site (Napier, 1992), as it was located in the same geographical area but about 1.5 km from the areas of spraying.
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Samples were collected in March 1993 and, at more sites, in March 1994, during the profenofos spraying seasons. Repeat samples were also taken between 16 and 20 May 1994, 6 wk after completion of profenofos spraying, to determine how long profenofos was persisting in water, sediment, and fish tissues; if AChE levels were recovering after cessation of spraying; and whether profenofos spraying was having a long-term effect.
Fish sampling was designed for one-factor analysis of variance (ANOVA) of each species to monitor the differences in the species sampled during each survey. Data from 1994 were also analyzed by two-factor ANOVA to determine the effect of location and season on the percentage lipid content, profenofos residue levels in fish, and their AChE activities. The fish collected from creeks were analyzed separately. Before ANOVA was performed, data were tested for normality and homogeneity of variance (Underwood, 1981). If a significant difference (P < 0.05) was found in ANOVA, then the test was followed by Dunnett's multiple comparisons. If the variances were heterogeneous, the data were transformed and tested again (Underwood, 1981). If the variances were still heterogenous, the Kruskal–Wallis rank test was used (Zar, 1984, p. 620). Statistical significance was determined at 
= 0.05. Pearson Correlation Coefficient Matrices (PCCMs) were performed to find the correlations between different variables. Regression analysis was performed to investigate the relationship between profenofos residue levels and AChE activity in various fish species.
Field Sampling
Duplicate water samples (500 mL) were collected in acetone–hexane rinsed amber glass bottles with aluminum foil–lined caps from the surface waters of each sampling site before commencement of any other procedures. Physico–chemical parameters (pH, dissolved oxygen [DO], conductivity, and temperature) were also measured at each site. Logistical difficulties in sampling in a remote area, with locations at different distances from each other, make direct measurements at the same time impossible. However, generally water quality parameters were measured in late evening or early morning when fish nets were set.
Replicate sediment samples to about 6 cm deep were collected using an ice-corer with a sleeve filled with dry ice and ethanol, and were immediately wrapped in aluminum foil, labeled, and dropped into liquid nitrogen for further analysis.
Fish were collected using gill nets, seine, nets and pond nets. Large fish were dissected on site to isolate gill, liver, and brain tissues while small fish were wrapped whole in aluminum foil and labeled. Each sample vial was immediately stored in liquid nitrogen then kept at -70°C for subsequent analysis of profenofos residue levels and AChE activity.
Laboratory Analysis
Profenofos in water samples was analyzed according to the method described by Abdullah et al. (1994). Profenofos residues in fish tissue was analyzed following the procedure outlined by Kumar and Chapman (1998). Recoveries for the samples of profenofos in water varied from 98 to 105% and for tissue samples were greater than 89%. The method has a detection limit of 0.001 mg/kg. The percentage lipid content in the gill and liver tissues of the fish sampled for residue analysis was determined as described by Nowak and Julli (1991). The AChE activity was measured in fish brain for bony bream and carp and in fish heads (up to operculum) for mosquitofish by the method of Ellman et al. (1961) and expressed as µmol acetylthiocholine/min/g protein. Percentage inhibition was compared with appropriate reference fish as 100% activity.
Spraying History
In March 1993, profenofos had been applied to the fields near Weetawaa Lagoon 2 d before sampling, and to the fields near Kerribee Lagoon 1 wk before sampling. Lowana Lagoon had no history of nearby recent profenofos application.
Profenofos spraying began in the Wee Waa district from 21 Feb. 1994, and fields near to Weetawaa, Kerribee, and Cudgewa Lagoons and Gunidgera Creek were sprayed with profenofos 7 to 9 d prior to our sampling. The property close to Galathera Creek and adjacent to Myall Vale irrigation canal was sprayed the night before sampling. Profenofos spraying was completed by early April and there was no history of profenofos spraying 6 wk before the May survey.
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RESULTS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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The pattern of profenofos concentrations in sediment samples was similar to that observed in water samples (Table 2), with an increase in the 1994 spraying season from low levels in 1993 (not detected to 160 µg/kg), followed by a decrease in May 1994. The highest concentrations of profenofos in sediment were found in March 1994 in irrigation ditches and canals at Jabiru (up to 2200 µg/kg; Kumar, 1995) but levels between 390 and 670 µg/kg were common at many sites. There was a positive correlation between the level of profenofos contamination in water and corresponding sediment samples.
Lipid Content in Fish
The lipid contents in fish tissues were remarkably similar within species from one survey to another (Table 3). Lipid content in gills and liver of bony bream was two to four times higher than in carp. The lipid content in gravid female mosquitofish was significantly higher than for nongravid mosquitofish, carp, or bony bream (Table 3).
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Residue Levels in Fish from Reference Sites
No profenofos residues were detected in fish from reference sites in March 1993 (Tables 4 and 5). However, in March 1994, low levels of profenofos (0.1–0.2 mg/kg) were found in gills and liver of both bony bream and carp (Table 4). In May 1994, as no bony bream were caught in Lowana Lagoon, 10 bream (average length: 32 cm) caught from the Darling River at Wentworth, NSW were used as controls. Some of these had profenofos residues at very low levels (Table 4), significantly lower than in March 1994. Although profenofos was not detected in water and sediment from Lowana Lagoon in May, two of the six carp were contaminated with very low levels of profenofos in liver (Table 4).
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Location and season had a significant effect on the amount of residues present in both gills and liver of carp. Residues of profenofos in liver in 1994 were significantly higher in carp from Weetawaa Lagoon and Gunidgera Creek compared with those from the reference site, Lowana Lagoon (Table 4). In May, 6 wk after completion of spraying, there was significant loss of residues from both gills and liver at all three comparative sites, except that residues in gills of fish from Weetawaa Lagoon remained low (Table 4). The high residues in whole small carp from Galathera Creek in March 1994 (1.5 mg/kg) had significantly reduced to 0.13 mg/kg by May 1994 (Table 5). In contrast, the residues in liver of large carp from Gunidgera Creek were still relatively high at 0.9 mg/kg.
Using a Pearson Correlation Coefficient Matrix (PCCM), both residue levels in carp and fish size in March 1993 were significantly correlated to their lipid content (Table 6). In March 1994, the only significant correlations for profenofos residues in large carp were between length and residue levels in liver. Profenofos concentrations in water and sediment were correlated to residue levels in the liver of large carp (r2 = 0.9) in both March and May 1994 (Table 6). Lipid content in small carp from Galathera Creek was significantly correlated with length and weight but not with residue levels, although length was correlated with residue levels in May.
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The greater lipid content of bony bream (Table 3) may have contributed to their accumulating more profenofos, as there was a significant correlation (PCCM) between lipid content and residue levels in individuals from Kerribee Lagoon in all surveys and (for liver tissues) from Weetawaa Lagoon in March 1993. Profenofos concentrations in water and sediment were correlated to residue levels in the liver of large bony bream (r2 = 0.9). Lipid content in small bony bream from Gunidgera Creek was significantly correlated with size of fish but not with profenofos residue levels.
Residue Levels in Exposed Mosquitofish
Mosquitofish were mostly caught from irrigation ditches and canals. The highest residue levels were found in gravid mosquitofish in March 1994 (9.5–10.7 mg/kg; Table 5), around twice the levels in nongravid mosquitofish (4.3–5.7 mg/kg). The residue levels in all mosquitofish had significantly reduced by May (Table 5).
There was a significant correlation (PCCM) between residue levels and lipid content in gravid and nongravid mosquitofish in March but not in May (Table 7). However, there was no significant correlation between residue levels in mosquitofish and profenofos in water and sediment.
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Acetylcholinesterase Activity in Wild Fish
Acetylcholinesterase Activity at Reference Sites
The AChE activities were almost the same in carp from Narrandera in March 1993 and Lowana Lagoon in May 1994, but there was significant inhibition (29%) in carp from Lowana Lagoon in March 1994 (Table 7). A similar pattern was detected in bony bream. The mean AChE activity in May 1994 in bream from the Darling River was not significantly different from that in bream from Lowana Lagoon in March 1993. However, the AChE activity in bony bream from Lowana Lagoon in March 1994 was significantly lower than in 1993 (only 63% of March 1993 activity levels; Table 7). Hence, the respective March 1993 AChE activities were adopted as the reference levels for both species (i.e., 0% inhibition). This slight depression of AChE activity in March 1994 was in accord with slightly elevated levels of profenofos in water and sediment in Lowana Lagoon (Table 2). As insufficient numbers of mosquitofish were caught from Lowana Lagoon in 1994, AChE activities of mosquitofish caught from an artificial lake at Macquarie University, Sydney were used as reference levels for this species (i.e., 0% inhibition). Profenofos was not detected in these fish.
Acetylcholinesterase Activity in Carp
Location and season had significant effects on the AChE activity in carp. Carp from Lowana Lagoon had significantly greater AChE activity than those from all other sites, except Galathera Creek in May. There was a universal trend toward recovery in AChE activity by May (Table 6), but it was not statistically significant for individuals from Weetawaa Lagoon.
There was a significant negative correlation (PCCM) between AChE activity and profenofos residue levels in gills of carp from Weetawaa Lagoon in March 1993 and residues in whole carp from Galathera Creek in March 1994 but not in May (Table 6). There were no significant correlations between AChE activity and other variables in May 1994.
Acetylcholinesterase Activity in Bony Bream
The AChE activities in bony bream from both the reference sites, Lowana Lagoon and the Darling River, were significantly greater than in those from Kerribee Lagoon and Gunidgera Creek during 1994 and Weetawaa Lagoon in 1993. There was recovery of AChE activity in bony bream from all sites (Table 7) by May 1994 but AChE activity in those from Kerribee Lagoon was still reduced by around 50% of the 1993 reference levels (Lowana Lagoon), and recovery of AChE was not significant.
There were strong correlations (PCCM) between AChE activity and residue levels in liver, as well as lipid content in tissues and fish length in bony bream from Kerribee Lagoon in March 1993. Residue levels in gills and liver of large bony bream and AChE activity were correlated in both March and May 1994 (Table 6) There was significant correlation between profenofos residue levels and AChE activity in small bony bream from Gunidgera Creek in March 1994. There was a significant effect of season on both the amount of residues and percentage AChE inhibition in the small bony bream from Gunidgera Creek.
Acetylcholinesterase Activity in Mosquitofish
Data from all the mosquitofish were pooled to analyze for the effect of size and season on residue levels and on AChE activity. Gravid females had significantly greater residue levels than nongravid individuals and significantly lower AChE levels in both 1994 surveys (Table 6). Greater AChE inhibition (64–72%) in gravid females, in comparison with 42 to 56% inhibition in the nongravid mosquitofish, was associated with higher profenofos residue levels. By May, 6 wk after the last spraying, all mosquitofish showed significant recoveries in their AChE activities, corresponding with significant elimination of residues from their bodies. However, gravid females still had 50 to 56% inhibition in AChE activity in May as compared with the reference samples.
There was significant correlation (PCCM) between residue levels and AChE activity in nongravid mosquitofish in May only, but the correlations were not significant between residue levels in gravid mosquitofish and their AChE activities.
Comparisons of Acetylcholinesterase Inhibition between Species
There was significantly greater AChE inhibition in bony bream and mosquitofish at exposed sites than in carp (Table 7). Overall correlations were calculated between the magnitude of residues detected and the level of AChE inhibition (% compared with control). Data were pooled from different sites and seasons, separately for each species, except that fish from lagoons were analyzed separately from those in creeks. There were significant correlations between percentage AChE activity and the amount of residues present in the small carp collected from Galathera Creek (r2 = 0.78) and bony bream from Gunidgera Creek (r2 = 0.89). However, the correlations were not high between residues in gills and liver tissues and percentage AChE activity for large carp (r2 = 0.2) and bony bream (r2 = 0.3) from lagoons, and in whole mosquitofish (r2 = 0.4).
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DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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The concentrations of profenofos in water in 1994 were much higher than in 1993, reflecting the greater usage of profenofos in 1994. By May 1994, 6 wk after spraying, profenofos levels in water had declined significantly and profenofos could not be detected in most lagoons and in Galathera Creek. However, the retention of high levels in Kerribee Lagoon (30% of the March level) and Gunidgera Creek (27–42% of the March level) was surprising. Its persistence in Gunidgera Creek may be explained in part by the drought conditions that resulted in the creek becoming a series of isolated waterholes. Some other sites, such as Weetawaa Lagoon, were also greatly affected by the long dry season, but no profenofos was detected in the water. Tomlin (1994)(p. 836–837) reported that profenofos is unstable in alkaline solutions with a 50% hydrolytic degradation time of 5.7 h at pH 9 at 20°C, compared with 14.6 d at pH 7. At the typical pH of these waters of 8, profenofos would hydrolyze by 50% in around 3 d, and the high concentrations in Kerribee Lagoon (assuming these to represent the levels when spraying ceased in early April) should have declined to around 0.2 ng/L in 6 wk. Even at the lowest measured pH of 6.9 (Galathera Creek in March), the 50% hydrolysis time would be 17 d. Clearly, other factors are resulting in profenofos retention at some sites.
The log Kow for profenofos is 4.44 (Tomlin, 1994, p. 836–837), indicating that profenofos would be expected to bind strongly to sediment. Profenofos levels in sediment generally correlated with the levels in water. They were high in March (0.8–0.7 mg/kg in exposed sites) and declined at all sites by May. Sediments may, in turn, be contributing to the retention of profenofos in the water column.
Profenofos concentrations up to 2.6 µg/L in stream water samples were measured after an alleged overspray incident from cotton-growing areas in 1991 (Environment Protection Authority versus Barlow; unreported, Land and Environment Court, Stein J, Case no. 50003/92). Although OPs are known to degrade rapidly in water, several authors have confirmed the presence of OPs in water samples collected in agricultural areas (Eto, 1974; Anees, 1975). Brunetto et al. (1992) detected dimethoate, diazinon, methyl parathion, and methamidophos in watercourses from Venezuela. Tsuda et al. (1994) surveyed seven rivers flowing into Lake Biwa in 1991 and detected high levels of diazinon, fenthion, and fenitrothion in water.
This study found that profenofos persisted in tissues of fish from lagoons and creeks close to cotton fields, reflecting the general levels of pesticide usage. The high profenofos residue levels in liver of bony bream from Kerribee Lagoon could be due to a combination of the high lipid content in their livers and the high Kow of profenofos. Liver seems to be the target organ for profenofos. Similar observations have been reported for other OPs in fish (Bender, 1969; Barron et al., 1991; Ferrando et al., 1992; Barron et al., 1993), and also for endosulfan (Nowak and Julli, 1991). Profenofos levels in gills of some fish (1.5 mg/kg) in the alleged overspray incident in 1991 (Case no. 50003/92) were similar to those in the present study but the levels in liver (0.5 mg/kg) were lower than in most sites in March 1994. The lower residue levels in liver are consistent with fish recently exposed to acute lethal levels of pesticide (Nowak et al., 1995).
Collection of a variety of fish species is necessary to characterize residue levels in fishes from a river system (Schmitt, 1981) and this provided an insight into the variability that can occur in the accumulation of profenofos. Residue levels in fish vary with species due to differences in lipid content, biology (trophic level habitat and reproductive season), exposure, detoxification capability, and ecology (Nowak and Julli, 1991). Nowak and Julli (1991) found higher residue levels of endosulfan in bony bream than in carp and this was partly explained by the much higher lipid content in the bony bream.
Despite the lack of statistical correlations between lipid and residue levels for mosquitofish and some carp, there was a similarity between the general trends for fish lipid content and profenofos levels. Lipid content decreased in the order of bony bream liver > gravid female mosquitofish > nongravid mosquitofish > carp. Residue levels decreased in the order of gravid female mosquitofish > nongravid mosquitofish > bony bream liver > carp. Napier (1992) also found much higher residues of endosulfan in gravid female mosquitofish, compared with nongravid, from these same lagoons. The differences in residue levels may also be a reflection of different exposure regimes and innate individual differences in metabolism.
The inhibition of AChE activity has been used as a tool in the monitoring of exposure of organisms during spraying programs involving OPs (Coppage et al., 1975; Busby et al., 1989). Acetylcholinesterase depression has been used successfully to detect exposure to cholinesterase-inhibiting pesticides near manufacturing plants. Moribund Atlantic menhaden taken from the Ashley River, South Carolina, USA had 47% inhibition of brain AChE, while fish from the same site but not showing signs of toxicity effect had only 17% inhibition (Williams and Sova, 1966). Exposed croakers from the same study also had decreased brain AChE activity (36%).
The interpretation of AChE activity depends on the knowledge of "normal" levels. To minimize errors and the confounding variations, it is preferable when determining the effect of chemicals on AChE levels to use appropriate unexposed "controls" from the same geographical area, collected at the same time, to account for natural and nonpesticide stress-induced factors. The AChE activities of bony bream and carp from Lowana Lagoon were measured where possible but it was not always possible to catch all these species in each survey. Hence, on occasion, appropriate species from other locations were used; carp from Narrandera NSW in March 1993, bony bream from the Darling River at Wentworth NSW in May 1994, and mosquitofish from a pond at Macquarie University NSW. During March 1994, water, sediments, and fish tissue samples from Lowana Lagoon, the reference site, were found to be contaminated with low levels of profenofos, possibly from runoff from rain a day before our survey and spray drift from fields more than 1 km away. Whyte and Conlon (1990) suggested that drift from aerial spray could be detected up to 2 km away from its source. This low level of contamination coincided with significant AChE inhibition in bony bream and carp, compared with March 1993. Hence, using values from the 1993 survey of Lowana Lagoon as reference levels of AChE may provide a better estimation of AChE inhibition, despite any disadvantages of making comparisons over different seasons. There is confidence in this comparison because of the similarities of AChE activities in reference bony bream and carp from March 1993 and May 1994, despite their different geographical sources. It is of even greater importance to compare AChE activities with those from fish in which little or no profenofos had been detected.
Difficulties arise when one tries to assign biological significance to a level of AChE inhibition. Inhibition of brain AChE in excess of 20% of normal or two standard deviations below the normal has been used as an indicator of significant exposure, following known application of AChE-inhibiting pesticides (Zinkl et al., 1980; Busby et al., 1989). Findlay et al. (1982) assessed poisoning in honey bees, based on the following three levels of AChE inhibition:
Level 1: treated animals, whose brain AChE levels were 67 to 100% of control, were considered normal;
Level 2: treated animals, whose brain AChE levels were 33 to 67% of control, were diagnosed as suffering from possible OP poisoning; and
Level 3: treated animals, whose brain AChE levels were 0 to 33% of control, indicated that OP poisoning was probable.
The degree of inhibition for bony bream and gravid mosquitofish during the spraying period would fall into Level 3 of the scale of Findlay et al. (1982), and, for most carp and nongravid mosquitofish, Level 2. Zinkl et al. (1991) suggested that brain AChE activity is usually depressed about 70% or greater in fish that have died or are in danger of dying from the AChE-inhibiting poisoning.
Although AChE inhibition has been successfully used for monitoring OPs, it is more powerful if it can be linked with measures of exposure, such as concentrations in water or tissues. This linkage is not often reported. In the present study, high residues were always associated with decreased AChE levels, and correlations were significant for bony bream, even when concentrations in water or sediment were low. There have been no previous reports of correlations between body residues of OPs in fish and AChE inhibition.
By May, 6 wk after cessation of spraying, the fish appeared to have metabolized the profenofos in most cases, and this corresponded with the recovery of inhibited AChE activity. The AChE activity after cessation of spraying recovered in different species in the following order: carp, nongravid mosquitofish, bony bream, gravid female mosquitofish.
Prolonged AChE inhibition has been found in fish exposed to several organophosphorus insecticides. Macek et al. (1972) found that after treatment of ponds with chlorpyrifos at 0.056 kg/ha, brain AChE activity of bluegills and largemouth bass was inhibited by 80%, and did not return to the levels of control fish until 28 d after pulse addition of the pesticide. Enzyme activity in brains of spot remained significantly below pre-spray levels for more than 40 d after spraying of malathion (Coppage et al., 1975). The persistence of elevated levels of profenofos residues and reduced AChE activity in the fish from Kerribee Lagoon could be due to this lagoon receiving tailwater from the nearby cotton fields as well as contributions from the contaminated sediment, although sediment contamination was similar or greater at other sites. In contrast, the AChE activity in brook trout, rainbow trout, and coho salmon had completely recovered 25 d after exposure to malathion (Sanders et al., 1981). Anderson et al. (1977) and Wang and Murphy (1982) found that differences in species sensitivity to AChE-inhibiting compounds can often be explained on the basis of inhibitor binding affinity and phosphorylation rate constants for the AChE enzymes.
Acetylcholinesterase inhibition can only be used as an important indicator of profenofos exposure if there is evidence of exposure in terms of profenofos being present in water, sediment, or the inhabiting fish species. It is noteworthy that AChE inhibition (Table 6) appears to be a more stable and robust measure (i.e., less variable) than residue levels in the fish tissues (Tables 4 and 5). The AChE levels in May 1994 in carp from Weetawaa Lagoon remained low (36% inhibited) despite a significant reduction of liver residue levels. Hence, this prolonged inhibition may not only be due to profenofos contamination, but could also be related to the decreasing water levels in the lagoon and deteriorating water quality, possible lack of food for the fish, or stress from disease. Zinkl et al. (1991) reported that environmental factors such as temperature might affect AChE activity in fish. As fish are poikilotherms, warmer temperatures will denature and inactivate enzymes like AChE more rapidly. At temperatures causing severe stress to rainbow trout, brain AChE activity was very low or non-existent (Zinkl et al., 1991). Underfeeding and exposure to elevated ambient temperatures have been shown to reduce AChE activity in Japanese quail (Rattner, 1982). However, comparisons of birds with fish may be questionable.
Connolly (1985) suggested that the toxic effect of a chemical could be related to the accumulation of the chemical at the site of action or to the accumulation of the injury. Once these relationships are established for all OPs, then it would conceivably only be necessary to measure AChE activity as an indicator of OP toxicity. A recent hazard assessment for fenitrothion use in Canadian forestry was based largely on the evidence that a large number of songbirds were found with a high degree of inhibition of brain AChE. This was related to studies observing effects, such as mortality, sublethal manifestations of toxicity, and reduced breeding success at similar levels of AChE inhibition (Busby et al., 1990). More work is required to establish such relationships for these fish.
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CONCLUSIONS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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ACKNOWLEDGMENTS |
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0.05) (this coding is not applicable for comparisons between species).