Comparative toxic and genotoxic effects of chloroacetanilides, formamidines and their degradation products on Vibrio fischeri and Chironomus riparius

Comparative toxic and genotoxic effects of chloroacetanilides, formamidines and their degradation products on Vibrio fischeri and Chironomus riparius

Environmental Pollution 119 (2002) 195–202 www.elsevier.com/locate/envpol Comparative toxic and genotoxic effects of chloroacetanilides, formamidines ...

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Environmental Pollution 119 (2002) 195–202 www.elsevier.com/locate/envpol

Comparative toxic and genotoxic effects of chloroacetanilides, formamidines and their degradation products on Vibrio fischeri and Chironomus riparius O. Osanoa,b,*, W. Admiraala, H.J.C. Klamerc, D. Pastorc, E.A.J. Bleekera a

Department of Aquatic Ecology and Ecotoxicology, IBED, University of Amsterdam, PO Box 94084, 1090 GB, Amsterdam, The Netherlands b School of Environmental Studies, Moi University, PO Box 3900, Eldoret, Kenya c National Institute for Coastal and Marine Management, RIKZ, PO Box 207, 9750 AE Haren The Netherlands Received 21 July 2001; accepted 9 November 2001

‘‘Capsule’’: Pesticides retain acute toxicity in both Vibrio fischeri and Chironomus riparius and genotoxicity in V. fischeri even after degradation.

Abstract Toxic and genotoxic effects of alachlor, metolachlor, amitraz, chlordimeform, their respective environmentally stable degradation products 2,6-diethylaniline, 2-ethyl-4-methylaniline, 2,4-dimethylaniline, and two other related compounds, 3,4-dichloroaniline and aniline were compared. Acute toxicity tests with Chironomus riparius (96 h) and Vibrio fischeri (Microtox1) and genotoxicity tests with a dark mutant of V. fischeri (Mutatox1) were carried out. Our results demonstrate that toxicity and genotoxicity of the pesticides are retained upon degradation to their alkyl-aniline metabolites. In the case of the herbicides alachlor and metolachlor, the toxicity to V. fischeri was enhanced upon degradation. Narcosis alone explains toxicity of the compounds to the midge, but not so for the bacteria suggesting a disparity in the selectivity of the test systems. All compounds showed direct genotoxicity in the Vibrio test, but amitraz and its metabolite were genotoxic at concentrations 103–105 lower than all the other compounds. The observations indicate that stable aniline degradation products of the pesticides may contribute considerably to environmental risks of pesticides application and that genotoxic effects may arise upon degradation of pesticides. # 2002 Elsevier Science Ltd. All rights reserved. Keywords: Chloroacetanilides; Formamidines; Anilines; Metabolites; Vibrio fischeri; Chironomus riparius; Genotoxicity; Toxicity; Lipophilicity

1. Introduction The chloroacetanilides alachlor and metolachlor are some of the most widely used selective herbicides worldwide in corn, soybean and other crop cultures (Krausse et al., 1985; Nesnow et al., 1995). Elevated concentrations of these herbicides and their degradation products have been detected in surface and groundwater (Nielsen and Lee, 1987; Wang et al., 1995). In groundwater, the concentration of the degradation product 2,6diethylaniline can be more than two times that of the parent compound alachlor (Potter and Carpenter, 1995). The formamidines amitraz and chlordimeform * Corresponding author. Tel.: +31-20-525-7718; fax: +31-20-5257716. E-mail address: [email protected] (O. Osano).

are used to control insects and mites. While the use of the latter is banned worldwide there is a notable increased use of amitraz in the cattle industry in Kenya. Both chloroacetanilides and formamidines are metabolised in intraperitoneally treated rats by the hepatic mixed function oxidase systems to 2,4 and 2,6 disubstituted anilines (Kimmel et al., 1986). These are further converted to the corresponding nitrosobenzenes, which are mutagenic in the Ames test (Kimmel et al., 1986). 3,4-dichloroaniline is a common degradation product of many herbicides including diuron and linuron (Crossland, 1990) used alongside the chloroacetanilides in agriculture and aniline bears the basic unsubstituted aniline moiety of the above test compounds. The pesticides and their aniline metabolites are suspect or confirmed oncogens (Weisburger et al., 1978; USEPA, 1990). Alachlor use is banned in Canada

0269-7491/02/$ - see front matter # 2002 Elsevier Science Ltd. All rights reserved. PII: S0269-7491(01)00334-7

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because of its mutagenicity and potential carcinogenicity (Horberg, 1990) while in Kenya its use is still unrestricted (Partow, 1995). Its genotoxic effects have been observed in plants, insects, yeast and mammals, but not in bacteria (Kimmel et al., 1986). Previously we observed that the aniline degradation products of the two chloroacetanilides induced developmental aberrations in the embryos of the frog Xenopus laevis (Osano et al., 2001). The ‘target’ theory (Hansch and Fujita, 1964) states that bioactivities, in our case toxicity, occur as a result of interactions between toxicants and receptors in the organism. The toxicity will depend on the bioavailability and chemical reactivity of the compounds. Based on this, lipophilicity (Log Kow) and different parameters for chemical reactivity of these chloroacetanilides, formamidines, their corresponding degradation products and two related compounds, 3,4-dichloroaniline and aniline are examined here to discuss specific mechanisms of toxicity to midges (Chironomus riparius) and bacteria (Vibrio fischeri). The concentration that is lethal to 50% of the organisms after 96 h [LC50 (96 h)], the concentration that causes 50% reduction in luminescence after 15 min [EC50 (15 min)] and lowest effect concentration (LEC) for induction of luminescence are determined for C. riparius, V. fischeri (Microtox1), and a dark mutant of V. fischeri (Mutatox1), respectively. The selectivity in these different biological test systems is compared and the risk of the tested pesticides and their environmentally stable degradation products is discussed.

2. Materials and methods 2.1. Chemicals The test chemicals, alachlor (ALA), metolachlor (MET), amitraz (AMI), chlordimeform (CHL) (purity: > 97%), 2,6-diethylaniline (DEA), 2-ethyl-6-methylaniline (EMA), 2,4-dimethylaniline (DMA), 3,4-dichloroaniline (DCA) and aniline (ANI) (purity: > 99%), and analytical grade cosolvents, methanol, acetone and dimethylsulfoxide (DMSO) were obtained from Fluka Riedel-de Hae¨n. The Dutch Standard Water (DSW) medium for the C. riparius tests was prepared from analytical grade salts as 0.2, 0.18, 0.1, 0.02 g/l of CaCl2 2H2O, MgSO4 7H2O, NaHCO3, and KHCO3, respectively, all dissolved in deionised water. Microtox1 and Mutatox1 kits were obtained from Azur Environmental (Carlsbad, CA, USA). Table 1 shows the Log Kow and concentration ranges of the tested compounds in our study. Tentatively, we verified the molecular properties of the compounds, i.e. molecular volume, surface area and heat of formation, dipole moment, highest occupied

molecular orbital (HOMO), lowest unoccupied molecular orbital (LUMO). Chemical reactivity factors were calculated by the Hyperchem (Hypercube Inc., Version 6.1) computer program. Prior to the calculations, the structures of the compounds were drawn and AM 1 optimised. For the Log Kow, surface area and volume calculations these structures were further optimised by MM+. 2.2. Chironomus riparius acute toxicity test Larvae used in each of the tests were taken from a C. riparius culture at the Department of Aquatic Ecology and Ecotoxicology (IBED, University of Amsterdam). Triplicate 96 h static acute toxicity tests were carried out in 180 ml (6.5 cm diameters) glass vessels filled with 100 ml test medium and 50 first instar larvae of less than 24 h old. The larvae were fed 1 ml fish food suspension (containing 52.5 g/l Tetraphil1/Trouvit1 1:20 w/w) and the medium was constantly aerated to maintain the O2 concentration at > 40%. The vessels were covered with a perforated polythene film to minimise evaporation of the medium. An incubation temperature of 20  1  C and a 16:8 h light:dark regime was maintained during the experiment. Test compounds were dissolved in DMSO and added to DSW to prepare a stock solution such that the final test concentration of DMSO did not exceed 80 ml/l. The dissolution of the compounds was facilitated by sonication for 30 min. Test dilutions were made by addition of DSW containing 80 ml/l DMSO to the various quantities of the stock solution. The concentrations of the chemicals tested were verified by HPLC (high-performance liquid chromatography) at the start and at the end of the experiments. The actual exposure concentration was deduced as the geometric mean of start and end concentrations in the media. The LC50 and its 95% confidence intervals for the compounds were deduced from Kaleidograph1 for windows (Synergy Software, version 3.08 for Windows, Reading, UK) fittings using the logistic response model (Haanstra et al., 1985):  Y ¼ c= 1 þ ebðXaÞ where ‘Y’=response (percentage survival), ‘c’=control response (set to 100%), ‘b’=slope, ‘X’=log of the exposure concentration and ‘a’=log LC50. 2.3. Vibrio fischeri acute toxicity test (Microtox1) The Microtox1 tests were done according to the supplier’s protocol (Azur Environmental, 2000a). In brief, the test compounds were dissolved in the cosolvent methanol or acetone to make a stock solution with a 0.5% (v/v) cosolvent concentration and a 1:1 (v:v) dilution series of the stock solutions were made in Microtox1

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O. Osano et al. / Environmental Pollution 119 (2002) 195–202 Table 1 Properties and test concentrations of the pesticides and degradation products Chemical

CAS No.

Molecular weight (g/mol)

Log Kow

Chironomus riparius (96-h acute toxicity; 8–10 dilutions) mM

Vibrio fischeri (Microtox1; 5 dilutions) mM

Vibrio fischeri (Mutatox1; 10 dilutions) mM

Alachlor 2,6-Diethylaniline

Parent Metabolite

15972-60-8 579-66-8

269.77 149.24

3.66 2.99

9.64–92.67 20.10–335.03

50.86–3793.04 156.39–2587.71

1.04–743.73 0.09–1014.78

Metolachlor 2-Ethyl-6-methylaniline

Parent Metabolite

5121-45-2 24549-06-2

283.80 135.21

3.31 2.59

37.35–196.27 34.76–398.64

89.57–7075.07 146.96–2530.45

1.80–1387.27 0.15–4961.66

Amitraz 2,4-Dimethylaniline

Parent Metabolite

33089-61-1 95-68-1

293.41 121.18

5.34 2.20

0.0034–13.63 165.04–825.22

0.31–2518.86 161.08–3540.50

0.0000341–9.65 0.00083–13.56

Chlordimeform

Parent

6164-98-3

196.70

2.79

Nda

3,4-Dichloroanilineb

Metabolite

95-86-1

162.02

2.30

Nda

Anilineb

Metabolite

62-53-3

93.13

1.26

a b

53.69–8590.14

24.40–941.81

0.25–1846.87

3.47–5746.22

0.10–108.95

794.59–12771.22

2.44–1252.08

Nd, not determined. Related aniline compounds.

diluent (2% sterile NaCl solution). This mixture was incubated at 15  0.1  C for 15 min. Luminescence was captured in the photoluminescence analyser at the beginning and at the end of 15 min. The EC50 luminescence inhibition and its 95% confidence intervals after 15 min were deduced using an in-built program in the Microbic model 500 program. 2.4. Vibrio fischeri genotoxicity test (Mutatox1) Mutatox1 tests with and without S9 enzyme activation were done according to the suppliers’ protocol (Azur Environmental, 2000a). Guided by the results of the Microtox1 tests described, we chose test dilutions that included concentrations that were sublethal to V. fischeri. Ten 1:1 (v:v) serial dilutions of the stock solution, as in Microtox1 above, were made in Mutatox1 Assay Medium (MAM) and in MAM plus rat S9 plus cofactors for S9-mediated assays in cuvettes for each compound. The S9-mediated assays were incubated at 35 0.1  C for 45 min to allow for S9 enzyme activation. The Mutatox1 tests were then incubated at 27 0.1  C for 24 h. The first luminescence reading was scored at 12 h and hourly until 24 h in the Microbic 500 analyser specially modified for the Mutatox1 test. In this study 16-h readings were chosen for analysis. A positive genotoxic response was recorded when at least two dilution cuvettes showed a light output reading of 4 times or more over the average control value (Ulitzur, 1983). Hence, we set our minimal response at 100 arbitrary light units. The lowest effect concentration (LEC) was deduced by linear interpolation between the dilutions with responses below and above 100 light units.

3. Results and discussion 3.1. Chironomus riparius acute toxicity test The parent pesticide compounds; ALA, MET, and AMI were 1.6, 2.1 and 42.9 times more acutely toxic to 1st instar larvae of C. riparius than their corresponding alkyl-aniline degradation products DEA, EMA, and DMA, respectively (Table 2). For AMI the LC50 (11.2 mM) for C. riparius was obtained by extrapolation as its LC50 lies above its maximum water solubility of 3.4 mM. However, we observed that when a higher concentration of DMSO (1% instead of 0.008% v/v) was used to facilitate dissolving of AMI, a 100% mortality of the larvae in 1.7 mM occurred, while no lethal effect was observed in a corresponding control with a similar concentration of DMSO. Thus, the well-known specific toxicity of amitraz as a monoamine oxidase inhibitor in invertebrates (Aziz and Knowles, 1973) was only apparent after a facilitation of its transport by a cosolvent. It is concluded that the importance of this compound as an acaricide or an insecticide is greatly dependent on its solvent carriers and that its aquatic ecotoxicity is limited by its low solubility in water and high Log Kow. Degradation of amitraz to DMA greatly enhances its water solubility, while decreasing its lethal toxicity. There was a significant correlation between acute toxicity to the midges and Log Kow for the compounds tested (Fig. 1A, solid line, r2=0.90**, slope=0.73, n=6). The data point of AMI was excluded in this analysis. The present correlation was consistent with the fitting for the quantitative structural activity relationship (QSAR) for ‘baseline toxicity’ using relatively

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Table 2 Toxicity values of alachlor, metolachlor, amitraz, their corresponding environmentally stable aniline degradation products 2,6-diethylaniline, 2ethyl-6-methylaniline and 2,4-dimethylaniline, and two other related aniline compounds, 3,4-dichloroaniline and aniline Test organism

Chironomus riparius

Vibrio fischeri

Vibrio fischeri (dark variant)

Cosolvent Chemical tested

Dimethylsulfoxide (mM) LC50

Methanol EC50 (mM)

Acetone EC50 (mM)

Alachlor 2,6-Diethylaniline

Parent 63.53 Metabolite 104.69

Methanol LEC (mM)

Acetone LEC (mM)

(57.51–70.18) (92.21–115.86)

589.39 (422.58–778.44) 28.14 (20.10–40.20)

912.26 75.05

(285.43–2909.89) (53.60–100.51)

21.06 32.27

137.89 49.48

(44.44–49.27 (95.90–103.56)

757.58 (704.72–775.19) 119.07 (79.88–178.98)

775.19 347.61

(549.68–1131.08) (266.25–488.13)

120.35 61.64

178.74 316.23

Amitraz 2,4-Dimethylaniline

Parent 11.18 (8.24–15.18) Metabolite 480.29 (468.22–492.66)

29.31 (17.04–44.31) 306.98 (247.57–412.61)

875.91 800.46

(402.17–19.08.59) (717.12–949.00)

Chlordimeform

Parent

Nda

121.50 (96.59–152.52)

315.71

(233.86–421.96)

68.56

1051.60

3,4-Dichloroanilineb

Metabolite

Nda

47.97

(18.52–92.58)

13.70

33.85

Anilineb

Metabolite 2549.77 (2528.72–2570.82) 1467.84 (966.39–2147.54) 2759.58 (1267.05–6013.10)

Metolachlor Parent 2-Ethyl-6-methylaniline Metabolite

a b

46.80 99.66

6.80

(5.55–8.02)

0.000281 0.0027

>1000

<0.01 0.38

Ndb

Nd, not determined. In brackets: 95% confidence limits. See text for comments.

non-reactive chemicals (2-propanol, 2-methyl-2propanol, 3-pentanol, 1,1,2-trichloroethane, toluene, 1,4-dichlorobenzene, 1,2,3-trichorobenzene, 1,2,3,4-tetrachlorobenzene, pentachlorobenzene) and 3rd instar C. riparius larvae in a 48-h test (Fig. 1A, broken line, Roghair et al., 1994). The slopes of Roghair et al. (1994) and our fitting were not different (P=0.051). The relatively lower (P=0.026) elevation in the present study may be attributed to a higher test sensitivity arising from the use of earlier stages of C. riparius larvae (1st instar) and the longer test duration (96 h). None of the molecular reactivity parameters tentatively explored (Section 2) yielded a significant correlation with toxicity. Clearly, in the Chironomus acute toxicity test, the pesticides and their environmentally stable aniline degradation products tested exerted their effects through narcosis (Fig. 1A) and not by a specific mode of action, except that the toxicity of amitraz may have been underestimated because of its low solubility. 3.2. Cosolvent effects in Vibrio fischeri assays The cosolvent strength is the ratio between pollutant concentrations at saturation in cosolvent–water mix to pollutant concentration at saturation in water. Therefore the stronger cosolvents would facilitate a higher availability of the test compounds to the bacteria. In both the Microtox1 and Mutatox1 assays all the compounds elicited a stronger effect when methanol was used as a cosolvent instead of acetone (Table 2), in spite of acetone’s higher cosolvent strength (Pastor et al., 1998). This was probably due to the higher toxicity of acetone (Lowest observed effect concentration (LOEC)

=600 mM) compared to methanol (LOEC=1000 mM) to V. fischeri (Azur Environmental Corporation, 2000b), thereby reducing the sensitivity of the test. Reported toxicity data in literature should therefore take account of the cosolvents used. The toxicity values in methanol cosolvent are considered hereinafter. 3.3. Vibrio fischeri (Microtox1 test) Bioluminescence of V. fischeri bacteria is controlled by a population density-responsive regulatory mechanism called quorum sensing (Dunlap, 1999). Cytotoxic compounds (bacteriocides) or inhibitors of any of the enzymes in the bioluminescence mechanism will result in luminescence reduction (Dunlap, 1999). The degradation products DEA and EMA were 20.9 and 6.3 times more acutely toxic than their corresponding parent compounds ALA and MET (Table 2). These results contrast with those for C. riparius where the parent compounds showed higher toxicity than their respective degradation products (Table 2). Amitraz affected the Vibrio test only at concentrations exceeding its solubility in water, while the degradation product DMA was slightly more toxic in the Vibrio test than in the midge test. The role of lipophilicity (Fig. 1B) and other chemical reactivity parameters (Section 2) in toxicity in the Microtox1 test could not be established, suggesting a more complex mechanism without any clear limiting step. The degradation products of the herbicides alachlor and metolachlor were more toxic than the parent compounds despite their predicted low affinity for uptake in the bacterial membranes by their low Log

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Fig 1. Correlation of toxicity data with Log Kow (A) Chironomus riparius LC50 (96 h). (B) Vibrio fischeri EC50 (15 min Microtox1). ALA, alachlor; MET, metolachlor, AMI, amitraz, DEA, 2,6-diethylaniline; EMA, 2-ethyl-6-methylaniline; DMA, 2,4-dimethylaniline and ANI, aniline. Bold line, linear regression fit and thick dotted lines: 95% CLs for our test. Broken line: Roghair et al. (1994) fitting for QSAR of ‘Baseline toxicity’ for C. riparius and thin dotted lines: its 95% CLs. AMI data was not considered in our fitting in A graph. For toxicity of amitraz see text.

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Kow. The Log Kow values for all parent compounds tested were above 3, apart from that of chlordimeform (Log Kow= 2.79) and the Log Kow values for all the degradation products were below 3. Hermens et al. (1985) reported that Microtox1 was not sensitive to very lipophilic compounds and Shusterman et al. (1992) observed higher toxicities to Salmonella typhimurium of substances with intermediate Log Kow values. Hence, lipophilicity alone does not explain toxicity of the aniline based pesticides and their degradation products in the Microtox1 test. Once transported into the cytoplasm, the aniline based compounds (parent compounds and degradation products) may have acted by inhibiting essential bioluminescence enzymes. Hermens et al. (1985) suggested that molecular refractivity (a steric factor) of aniline and dichlorinated anilines could negatively influence the activity of the luciferase enzyme or reduce the flavin mononucleotide (FMNH2) donating system of V. fischeri. The dependence of the activity of the degradation products on lipophilicity as well as the evidence of the hydrophobic site(s) on NADPH oxidase, an essential enzyme in the bioluminescence production (Nakata et al., 1997), further suggest a possible enzyme inhibition mechanism of toxicity. This enzyme inhibition of bioluminescence may explain why toxicity of aniline compounds to V. fischeri bacteria was found to be unrelated to their Log Kow values as opposed to E. coli (Jaworska and Schultz, 1994). Toxicity of substituted nitrosobenzenes including NO2, OH, NH2, OCH, methyl, and halogen substituents to a river bacteria inoculate and V. fischeri was found to be controlled mainly by the electronic factors (Yuan et al., 1997). We suggest that toxicity or bioluminescence inhibition depends on transport of the toxicants to the cell membranes bilayer as well as transport from the membrane to luciferase enzyme receptors and agree with Huang et al. (1996) that toxicity of the anilines to V. fischeri could be controlled by electronic factors and Log Kow combined. The limited set of our compounds hampered a search for correlations between toxicity and electronic factors. 3.4. Vibrio fischeri genotoxicity (Mutatox1 test) The Mutatox1 genotoxicity test is affected by sublethal concentrations of substances that damage or intercalate DNA, inhibit synthesis of new DNA, are direct mutagens causing base substitution or frameshifts, or are SOS inducing agents (Weiser et al., 1981; Ulitzur and Weiser, 1981; Ulitzur, 1983; Johnson, 1992). All the compounds were genotoxic in the direct test (Table 2) and appeared nongenotoxic when S9 activated (results not shown). Lipophilicity, physico-chemical and reactivity parameters of the compounds (Section 2) did not correlate with toxicity in the Mutatox1.

Activation of the compounds with the S9 enzyme system apparently eliminated their genotoxicity or gave rise to highly bactericidal products or bioluminescence inhibitors that prevented expression of any effects in the present test. It is also possible that the products of S9 bioactivation were less able to cross the membrane barriers into the site of photoluminescence activity within the cytoplasm (De Maagd and Tonkes, 2000). However, genotoxicity of ALA after S9 activation at 30 mg/l (Canna-Michaelidou and Nicolaou, 1996), a value 10 times lower than our lowest test concentration, suggest that the products of S9 activation which includes nitrosobenzenes and hydroxylamines (Kimmel et al., 1986) may have been too toxic to allow fluorescence induction or expression. In accordance, the nitrobenzenes, which are hypothesised to be one of the anilines’ breakdown products, were found to have Microtox1 EC50 values lower than our test concentrations (Deneer et al., 1989; Yuan et al., 1997). Furthermore, nitrosobenzenes are potential inhibitors of NADPH-oxidase (Nakata et al., 1997), an essential enzyme in the bioluminescence production. So, the evidence available suggests that the lack of response after S9 activation in the present study may have been a false negative. The direct genotoxicity of the test compounds in the present study contrasts with the findings in the Ames test (Kimmel et al., 1986) where the same compounds tested had to be S9 activated before they could become mutagenic. Direct reactions with or modification of the DNA may have been responsible for the direct genotoxicity of the compounds in the present study. For example ALA and its metabolites are capable of forming DNA adducts (Brown et al., 1988; Nesnow et al., 1995) and Ribas et al. (1995) found ALA to be genotoxic when evaluated in the Single-cell gel electrophoresis assay both with and without S9 activation. The range of the concentrations of ALA tested by CannaMichaelidou and Nicolaou (1996) may have been too low for substantial interaction with DNA. It is possible that the compounds provided modification(s) to the DNA that never lead to mutagenic events in the Ames test or that the Mutatox1 test depends on regulatory events in the Vibrio that differ from that in Salmonella. In conclusion all the three test systems indicate that toxicity is retained after degradation of the pesticides to their environmentally stable aniline compounds. For the midges this effect is likely to be controlled by narcosis (baseline toxicity), while the Vibrio test showed very different selectivity. The Microtox1 test revealed relatively higher toxicities for the degradation products of the herbicides than their parent compounds. All the pesticides and degradation products in the present study gave positive genotoxic response in the Mutatox1 test and should be considered potential genotoxicants capable of altering gene functions. Special attention should be given to amitraz and its degradation product, which

O. Osano et al. / Environmental Pollution 119 (2002) 195–202

were genotoxic at very low concentrations (< 0.005 mM). The widespread application of aniline based pesticides, both herbicides and acaricides, in Kenya and elsewhere might lead to accumulation of relatively stable degradation products. Ecotoxicological effects of these compounds in the environment can be diverse due to the different selectivities in different test systems as observed in the present study.

Acknowledgements This research was sponsored by the Nuffic, Nertherlands Organization for International Cooperation in Higher Euduation. We are grateful to Dr. Michiel Kraak (University of Amsterdam, IBED) for his supportive comments on earlier versions of this manuscript.

References Aziz, S.A., Knowles, C.O., 1973. Inhibition of monoamine oxidase by the pesticide chlordimeform and related compounds. Nature 242, 417–418. Azur Environmental Corporation, 2000a. Azur Mutatox1. Azur Corporation 2232 Rutherford Road Carlsbad, California 92008, USA. Azur Environmental Corporation, 2000b. The Microtox1 Chronic Toxicity Testing System. Available: http://www.azurenv.com/ chron.htm. Brown, M.A., Kimmel, E.C., Casida, J.E., 1988. DNA adducts formation by alachlor metabolites. Life Sciences 43, 2087–2094. Canna-Michaelidou, S., Nicolaou, A., 1996. Evaluation of the genotoxicity potential by Mutatox1 test of ten pesticides found as water pollutants in Cyprus. The Science of the Total Environment 93, 27– 35. Crossland, N.O., 1990. A review of fate and toxicity of 3,4-dichloroaniline in aquatic environments. Chemosphere 21, 1489–1497. De Maagd, P.G.J., Tonkes, M., 2000. Selection of genotoxicity tests for risk assessment of effluents. Environmental Toxicology 15, 81– 90. Deneer, J.W., Van Leeuwen, C.J., Seinen, W., Maas Diepeveen, J.L., Hermens, J.L.M., 1989. QSAR study of the toxicity of nitrobenzene derivatives towards Daphnia magna, Chlorella pyrenoidosa and Photobacterium phosphoreum. Aquatic Toxicology 15, 83–98. Dunlap, P.V., 1999. Quorum regulation of luminescence in V. fischeri. Journal of Molecular Microbiology and Biotechnology 1 (1), 5–12. Haanstra, L., Doelman, P., Oude Voshaar, J.H., 1985. The use of sigmoidal dose response curves in soil ecotoxicology research. Plant and Soil 84, 293–297. Hansch, C., Fujita, T., 1964. r-s-p-Analysis. A Method for the Correlation of Biological Activity and Chemical Structure. Journal of American Chemical Society 86, 1616–1626. Hermens, J., Busser, F., Leeuwangh, P., Musch, A., 1985. Quantitative structure–activity relationships and mixture toxicity of organic chemicals in Photobacterium phosphoreum: the Microtox1 test. Ecotoxicology and Environmental Safety 9, 17–25. Horberg, G., 1990. Risk, science, and politics. Alachlor regulation in Canada and the United States. Canadian Journal of Political Sciences 23 (2), 257–277. Huang, Q., Kong, L., Wang, L., 1996. Applications of frontier molecular orbital energies in QSAR studies. Bulletin of Environmental Contamination and Toxicology 56, 758–765.

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Jaworska, J.S., Schultz, T.W., 1994. Mechanism-based comparison of acute toxicities elicited by industrial organic chemicals in prokaryotic and eukaryotic systems. Ecotoxicology and Environmental Safety 29, 200–213. Johnson, B.T., 1992. Potential genotoxicity of sediments from the Great Lakes. Environmental Toxicology and Water Quality 7, 255– 268. Kimmel, E.C., Casida, J.E., Ruzo, L.O., 1986. Formamidine insecticides and chloroacetanilide herbicides: disubstituted anilines and nitrosobenzenes as mammalian metabolites and bacterial mutagens. Journal of Agriculture and Food Chemistry 34, 157–161. Krausse, A., Hancock, W.G., Minard, R.D., Freyer, A.J., Honeycutt, R.C., LeBaron, H.M., Paulson, D.L., Liu, S.Y., Bollag, J.M., 1985. Microbial transformation of the herbicide metolachlor by a soil actinomycete. Journal of Agriculture and Food Chemistry 33, 584– 589. Nakata, M., Nasuda-Kouyama, A., Isogai, Y., Kanegasaki, S., Lizuka, T., 1997. Effect of aromatic nitroso-compounds on superoxide-generating activity in neutrophils. Journal of Biochemistry 122, 188–192. Nesnow, S., Agarwal, S.C., Padgett, W.T., Lambert, G.R., Boone, P., Richard, A.M., 1995. Synthesis and characterisation of adducts of alachlor and 2-choro-N-(2,6-diethylphenyl)acetamide with 20 -deoxyguanosine, thymine and their 30 -monophosphate. Chemical Research in Toxicology 8, 209–217. Nielsen, E.G., Lee, L.K., 1987. The Magnitude and Costs of Groundwater Contamination from Agricultural Chemicals. A National perspective (US Dept. Agricultural Staff report AGES870318). Washington DC USA. Osano, O., Admiraal, W., Otieno, D., 2001. Developmental disorders in the embryos of the frog Xenopus laevis induced by chloroacetanilide herbicide and degradation products. Environmental Toxicology and Chemistry (in press). Partow, H., 1995. What Prospects for Pesticide Use in Kenya? (A WWF country report). Nairobi, Kenya. Pastor i Rodrigues, M. D., 1998. Optimization of Bacterial Bioassays for Directed chemical Fractionation of Environmental Samples (Rikz report). National Institute for Coastal and Marine Management (RIKZ), Haren. The Netherlands, pp. 1–106. Potter, T.L., Carpenter, T.L., 1995. Occurrence of alachlor environmental degradation products in groundwater. Environmental Science and Technology 29, 1557–1563. Ribas, G., Frenzilli, G., Barale, R., Marcos, R., 1995. Herbicideinduced DNA damage in human lymphocytes evaluated by the single-cell gel electrophoresis (SCGE) assay. Mutation Research 344, 41–54. Roghair, C.J., Buijze, A.B., Yedema, E.S.E., Hermens, J.L.M., 1994. A QSAR for baseline toxicity to the midge Chironomus riparius. Chemosphere 28, 989–997. Shusterman A. J., 1992. Predicting chemical mutagenicity by using quantitative structure activity relationships. In: Food Assessments. ACS Symposium Series 484. American Chemical Society ed. Washington, USA (18) pp. 181–189. Ulitzur, S., Weiser, I., 1981. Acridine dyes and other DNA-intercalating agents induce the luminescence system of luminous bacteria and their dark variants. Proceedings of the National Academy of Sciences 78, 3338–3342. Ulitzur, S., 1983. Bioluminescence test for genotoxic agents. Methods in Enzymology 133, 264–274. US Environmental Protection Agency, 1990. Drinking water regulations and health advisories. Lewis Publishers, Chelsea, MI, USA. Wang, W., Liszweski, M., Buchmiller, R., 1995. Occurrence of active and inactive herbicide ingredients at selected sites in Iowa. Water, Air and Soil Pollution 83, 21–35. Weisburger, E.K., Russfield, A.B., Homburger, F., Weisburger, J.H., Boger, E., Van Dongen, C.G., Chu, K.C., 1978. Testing of 21 environmental aromatic amines or derivatives for long-term toxicity

202

O. Osano et al. / Environmental Pollution 119 (2002) 195–202

of carcinogenicity. Journal of Environmental Pathology, Toxicology and Oncology 2, 325–352. Weiser, I., Ulitzur, S., Yannai, S., 1981. DNA-damaging agents and DNA-synthesis inhibitors induce luminescence in dark variants of luminous bacteria. Mutation Research 91, 443–450.

Yuan, X., Lu, G., Lang, P., 1997. QSAR study of the toxicity of nitrosobenzenes to river bacteria and Photobacterium phosphoreum. Bulletin of Environmental Contamination and Toxicology 58, 123–127.