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Developmental disorders induced by pesticide degradetion products

Osano, O.F.

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Citation for published version (APA):Osano, O. F. (2002). Developmental disorders induced by pesticide degradetion products. Amsterdam.

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Developmenta ll disorder s induce dd by pesticid e degradatio nn product s

Odip oo Osano

Developmentall Disorders Induced by Pesticide Degradationn Products

ACADEMISC HH PROEFSCHRIFT

terr verkrijging van de graad van doctor aan de Universiteit van Amsterdam opp gezag van de Rector Magnificus prof. mr. P.F. van der Heijden ten

overstaann van een door het college voor promoties ingestelde commissie, in hett openbaar te verdedigen in de Aula der Universiteit op woensdag 17 april

2002,, te 12.00 uur

doorr Odipo Osano geborenn te Nairobi (Kenia)

PROMOTOR R

prof.. dr. W. Admiraal

CO-PROMOTOR R

dr.. M.H.S. Kraak

OVERIG EE LEDEN VAN DE COMMISSI E

prof.. dr. R.W.P.M. Laane prof.. dr. M.W. Sabelis prof.. dr. N.M. van Straalen prof.. dr. F.J. van Schooten dr.. P. de Voogt dr.. T.C.M. Broek dr.. J.W. Everts

Faculteitt der Natuurwetenschappen, Wiskunde en Informatica Universiteitt van Amsterdam

Developmentall Disorders Induced By Pesticide

Degradationn Products

n n Thiss study was conducted at the department of Aquatic Ecology and Ecotoxicology, IBED,, Kruislaan 320, University of Amsterdam, P. O. Box 94084, 1090 GB, Amsterdam,, The Netherlands (www.bio.uva.nl/onderzoek/aot)

andd at

Thee School of Environmental Studies, Moi University, P. O. Box 3900 Eldoret, Kenya. .

lÊfjklÊfjk Amsterdam Research Institute ^ ^ ** For Global Issues and Development Studies

AGIO S S

Thiss PhD dissertation is one of the results of the staff development project which is partt of the Joint Financing Program for Cooperation in Higher Education (MHO) programm 'Strengthening the School of Environmental Studies, Moi University, Eldoret,, Kenya', managed by The Netherlands Organization for International Co-operationn in Higher Education (NUFFIC) with funds from the Netherlands Minister forr Development Co-operation since 1991. The program is coordinated by the Amsterdamm Research Institute for Global Issues and Development Studies of the Facultyy of Social and Behavioral Sciences of the University of Amsterdam (scientificc co-ordinator Prof, dr Ton Dietz; project manager drs Annemieke van Haastrecht),, together with Moi University SES (current dean Dr Wilson Yabann, formerr deans Prof, dr Charles Okidi, Prof, dr Dismas Otieno and Prof, dr Mwakio Tole). .

Cover:: Adult Xenopus laevis, 3D molecular structures of 2,4-dimethylaniline and 2-ethyl-6-methylaniline,, an aircraft used for spraying pesticides and a drop of water. Cover design by Harmm van der Geest.

http://www.bio.uva.nl/onderzoek/aot

Contents s

Page e

Chapterr 1 General introduction 7

Chapterr 2 Fate and risk of chloroacetanilide degradation products inn the Nzoia basin, Kenya -- O Osano, D Nzyuko, PM Tole, W Admiraal SubmittedSubmitted for publication 23

Chapterr 3 Comparative toxic and genotoxic effects of chloroacetanilides,, formamidines and their degradation productss on Vibrio flscheri and Chironomus riparius -- O Osano, W Admiraal, HJC Klamer, D Pastor, EAJ Bleeker EnvironmentalEnvironmental Pollution (in press) 39

Chapterr 4 Developmental disorders in embryos of the frog XenopusXenopus laevis induced by chloroacetanilide herbicides andd their degradation products -- O Osano, W Admiraal, D Otieno EnvironmentEnvironment Toxicology and Chemistry Vol 21(2) (2002)ppp 375-379 55

Chapterr 5 Teratogenic effects of amitraz, 2,4-dimethylaniline and paraquatt on developing frog {Xenopus) embryos. -- O Osano, AA Oladimeji, MHS Kraak, W Admiraal ArchivesArchives of Environmental Contamination and ToxicologyToxicology (in press) 71

Chapterr 6 Concluding remarks 89

Summaryy 99

Samenvattingg 103

Acknowledgements s 107 7

ChapterChapter 1

Generall Introductio n

Currentt intensification of agriculture in Kenya and other developing countriess demands increased pesticide use, which may lead to pesticide contaminationn of ground and surface water. It has been estimated that less thann 0.1% of the pesticides applied to crops reach the target pests, thus more thann 99% of the applied pesticides have a potential to impact non-target organismss (Albert et al 1992). In the tropics, high levels of ultra-violet radiations,, high temperatures and high rainfall with subsequent runoffs modifyy these risks (Bossan et al. 1995; Abdullah et al. 1997).

Pesticidess degrade in the environment into transformation products thatt could alleviate, enhance and/or increase the diversity of their toxic effectss (Kraak et al. 1997; Bleeker et al. 1999; Admiraal et al. 2000). Besidess the acute toxic effects, teratogenic, genotoxic, mutagenic, carcinogenicc and other subtle toxic effects are important probable endpoints, whichwhich may differ between the parent compounds and their degradation products.. However, ecotoxicological studies of pesticides in the aquatic environmentt have hitherto underrated their degradation products. The invariablee persistence of the degradation products underlines the need to studyy their long-term effects in the environment. Most knowledge of pesticidess is derived from studies in the temperate regions. The tropical environment,, which is the focus of this thesis, has been sparsely studied, in spitee of the disparate physical and chemical environmental conditions prevailingg there.

Basedd on their extensive usage in Kenya to control weed in the cereall crops and tick in the dairy industry, chloroacetanilides, formamidines andd their resulting environmentally stable anilines degradation products are studiedd for their acute toxic, genotoxic, and teratogenic effects on the ubiquitouss Chironomus riparius, Vibrio fischeri and the locally abundant XenopusXenopus laevis.

Chloroacetanilides,, formamidines and their degradation products

Thee chloroacetanilides alachlor, butachlor (synthesized from 2,6-diethylaniline),, metolachlor and acetochlor (derived from 2-ethyI-6-methylaniline)) are some of the most intensively used herbicides worldwide

7 7

GeneralGeneral Introduction

(Hil ll et al. 1997), while the formamidine amitraz (derived from 2,4-dimethylaniline)) is an important acaricide used extensively in tick endemic areass of Australia and Southern Africa including Kenya (Partow 1995; Baxterr and Barker 1999). The use of amitraz in Kenya is currently on the increasee to alleviate the problems of increased tick resistance to alternate acaricides. .

Inn the environment the chloroacetanilides revert to the precursor aniliness as the environmentally stable degradation products (Tiedje and Hagedornn 1975; Allcock and Woods 1978; Kimmel et al. 1986; Knowles andd Hamed 1989; Wei and Vossbrinck 1992; Hill et al. 1997; Stamper and Tuovinenn 1998; Corta et al. 1999). Degradation of chloroacetanilides and formamidiness requires a consortium of bacteria for completion (De Schrijverr and De Mot 1999). This mostly occurs in aerobic conditions and is modestt under anaerobic conditions (Konopka 1994). Metolachlor is transformedd to a lesser extent than alachlor, while amitraz is quickly transformedd in the environment and in living organisms to the more acutely toxicc intermediate BTS27271 before further degradation (FAO/WHO 1985; Knowless and Hamed 1989; Pass and Mogg 1991; Konopka 1994; Pass and Moggg 1995; Corta et al. 1999).

Thee fate of the parent compounds and their breakdown products are influencedd by their solubility, stability, temperature, irradiation, biodegradability,, bioavailability, runoff, precipitation, management practices,, wind drift, erosion and chemical sorption (Becking et al. 1992; Allann 1994). High ambient temperatures, larger flux of the sun's irradiation andd torrential rains are characteristics of the tropics, which may enhance the ratee of pesticide degradation and dissipation.

Thee occurrence of degradation products of chloroacetanilides in the tropicss is not documented, while in the temperate zone the parent chloroacetanilidee compounds and their degradation products have been foundd in surface water and groundwater (Galassi et al. 1996; Thurman et al. 1996;; Albanis et al. 1998; Hostetler and Thurman 2000; Scribner et al. 2000b;; Scribner et al. 2000a). Indeed, the potential risks of the pesticides in thee environment should be a sum of the risks posed by the parent, intermediate,, and stable degradation products. The analysis of the stable transformationn products in water and the sediment is vital to evaluate the non-point-sourcee contamination of water by chloroacetanilides and to estimatee their long-term effects in the aquatic environment.

8 8

ChapterChapter 1

R1 1 CI I

R2 2

R1 1

Chloroacetanilides s

[O]] or hydrolysis

NH„ „

Anilines s

[01 1

R11 - . CC I NHOH"

[O] ]

J J

Hydroxylamines s

R1 1

r r

^ 1 1

C -- N R2 2

Formamidines s [O]] or hydrolysis

Anilines s

[«] ]

HOH H

J J

[O] ]

R1 1

Hydroxylaminess Nitrosobenzenes

\\ /

R1 1

Nitrosobenzenes s R11 Azoxybenzenes

Compoundd Substitute Rll R2

Compound d Substitute e Rll R2

Alachlorr CH2CH3 CH2OCH3

Metolachorr CH3 CH2(CH3)CH2OCH3

Acetochlorr CH3 CH2OCH2CH3

Butachlorr CH,CH3 CH2OCH2CH2CH2CH3

Chlordimeformm CI Amitrazz CH3

CH3 3

CH H

Figg 1: Designations and partial metabolic pathways of chloroacetanilide herbicides and formamidinee insecticides showing possible routes resulting in conversion of pesticidess to the mutagenic and carcinogenic nitrosobenzenes. Modified from Kimmell et al. (1986).

9 9


Testt Organisms

Xenopuss laevis larvae

Thee clawed frog Xenopus laevis is a native species in Sub-Saharan Africa.. It is grouped in the family Pipidae, all of whose members are exclusivelyy aquatic and tongue-less. They are easy to breed and raise in the laboratory.. Furthermore, they allow induction of artificial spawning at any timee of the year with the females laying 500 - 2400 eggs at each time (Dumpertt and Zietz 1984). Human Chorionic Gonadotropin (HCG) hormonee is used for breeding induction by injection into the dorsal lymph sacc (Fig 2).

Figg 2. Adult Xenopus laevis. The triangle shows the position of the dorsal lymph sac

Thee amphibian embryo remains a classical model for experimental embryologicall studies as it is an intact developing system, which undergoes

12 2

ChapterChapter 1

evolutionaryy conserved events of cleavage, gastrulation, and organogenesis, comparablee to those of other vertebrates, including mammals (Dumont et al 1983b).. Validation studies using compounds with known mammalian and humann developmental toxicity, or both, suggest that the predictive accuracy off the Xenopus laevis embryo test approximates 85%, so it can be used as an indicatorr of potential human developmental health hazard (Dumont et al. 1983a;; Courchesne and Bantle 1985; Sabourin et al. 1985; Dawson and Bantlee 1987; Sabourin and Faulk 1987; Bantle 1995). The test with the XenopusXenopus embryos is a standardized 96-h test using midblastula (stage 8) to gastrulaa (stage 9) stages, thereby exposing all the sensitive stages of primary organogenesiss (Nieuwkoop and Faber 1975; ASTM 1991; Bantle 1995). An embryonicc teratogenic index (TI; which is expressed as 96h-LC50/96h-EC50(malformation))) allows comparison of teratogenic risks of diverse compoundss and mixtures (Dumont et al. 1983b; Bantle 1995). The teratogenicityy of highly embryolethal compounds would obviously be less relevantt in the environment compared to that of less lethal compounds, whichh have a potential to cause malformation in a large number of surviving organisms. .

Chironomuss riparius larvae

Thee dipteran family Chironomidae are sediment-inhabiting organismss with a cosmopolitan distribution in freshwater ecosystems (Armitagee et al. 1995). Their first instar is planktonic, while the 2nd to 4th

instarr stages inhabit the upper layer of the sediment. The sediment is a major repositoryy for many persistent chemicals, making C. riparius larvae ideal organismss to evaluate adverse effects of toxicants. At 20°C a complete life cyclee of the C. riparius can be completed in three weeks, making the insect easyy to culture in the laboratory. The 1st instar larvae have been chosen for thee present study based on the their relatively higher sensitivity than the laterr stages of development (Williams et al. 1986). The larvae were obtained fromm a laboratory culture at the Department of Aquatic Ecology and Ecotoxicology,, which has been maintained at 20°C (16:7 h light-dark regime,, separated by 0.5 h twilight). To minimize the risk of inbreeding, egg massesmasses were regularly exchanged with other Dutch laboratory cultures of C. ripariusriparius and large insect populations were maintained.

13 3


VibrioVibrio fischeri

VibrioVibrio fischeri is a motile gram-negative bioluminescent marine bacteriumm with a rod shape. Luminescence in V. fischeri is initiated by a cell-density-dependentt activation of the gene lux regulon- 'Quorum Sensing'' (Dunlap and Kuo 1992). Each bacterium produces chemical signalss called autoinducers, which upon attaining a threshold concentration inducee the reactions leading to luminescence. A blue-green light at 480-490 nmm is produced after a luciferase-catalyzed oxidation of flavin mononucleotidee (FMNH2) and a range of long chained (8 - 14C) fatty aldehydess (RCHO), the luciferins (Meighen 1993). Fig 3. summarizes the VibrioVibrio fischeri light emitting reaction.

LuxABB (Luciferase)

FMNH 22 + 02+RCH O ^ ^ ^ ^ ^ FMN + H20 + RCOOH + hv

Figg 3. The light emitting reaction of V. fischeri

Cytotoxicc compounds or luciferase inhibitors will result in reduced luminescence,, the endpoint in the acute toxicity test using V. fischeri (Microtox®).. Microtox® is an inexpensive standardized short-term in vitro bioassayy that can screen complex industrial effluents, environmental mixturess and newly introduced compounds.

Sublethall concentrations of genotoxic compounds, restore luminescencee in a dark mutant (Ml69) of the V. fischeri in the Mutatox® test.. Different genotoxic agents, including base substitution, frame shift, DNAA synthesis inhibitors, DNA damaging agents, and DNA intercalating agents,, result in the appearance of light in the dark strain of Ml69 (Ulitzur 1982).. The multiple endpoint targets of the test system make it highly sensitive,, hence suitable as a rapid and cheap screening assay for suspect genotoxicants. .

14 4

ChapterChapter 1

Outlin e e

Thee fate of massively applied chloroacetanilides, formamidines and theirr degradation products in the tropical environment is not known and it is suspectedd that this may be different from that in the widely studied temperatee region. Therefore, river water and sediment samples from River Nzoia,, Kenya, were analyzed for alachlor, metolachlor and their stable anilinee degradation products (Chapter 2). An attempt was made to reconstructt their fate and to elucidate their potential risks. These risks were estimatedd by deducing the pesticide balance of the River Nzoia catchment. Basedd on the evidence that the pesticide degradation products are potentially moree mutagenic than the parent compounds, our first set of laboratory experimentss aimed to discriminate baseline toxicity from effects on specific biologicall endpoints. Toxicity of the pesticides, their degradation products andd chemically related compounds were investigated using Chironomus ripariusriparius and Vibrio fischeri (Chapter 3). Genotoxicity of the compounds wass explored using the MUTATOX® test. Potentially genotoxic compounds mayy affect especially developing embryos, given the inherent challenges of celll division, differentiation and rapid growth during this period. Therefore inn Chapter 4 and 5 we examined the developmental and teratogenic effects off the commonly used chloroacetanilides (alachlor and metolachlor), formamidinee (amitraz), their stable aniline degradation products and, additionally,, the herbicide paraquat on early embryos of a native frog species,, Xenopus laevis. The concluding remarks (Chapter 6) discuss the mainn findings of this thesis as well as the implication for risk assessment of pesticidess in the tropics.

15 5


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ChapterChapter 1

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21 1

ChapterChapter 2

Fatee and Risk of Chloroacetanilide Degradation Productss in the Nzoia Basin, Kenya

Abstract t

Alachlor,, metolachlor and their respective environmentally stable aniline degradationn products, 2,6-diethylaniline and 2-ethyl-6-methylaniline were analyzedd in water and sediment samples from 9 sites along River Nzoia, Kenyaa using gas chromatography. The degradation products were detected inn >90% of the sediment and water samples, while the parent compounds occurredd in <14% of the water samples. Much higher concentrations of the pesticidess and their degradation products occurred in the sediment than in thee water (1.4 - 10800 fold), indicating an accumulation of the compounds inn the sediment. The constant occurrence of the degradation products in the sedimentt during the study period infers a persistence of these compounds. It iss hypothesized that the prevailing tropical climatic conditions favor a quick breakdownn of the pesticides to their environmentally stable degradation products,, thereby making the latter more important pollutants than their parentt products in the study area.

23 3

FateFate and risks of degradation products

Introductio n n

Potentiall risks of pesticide use in the tropics may differ from those in thee widely studied temperate regions. It is generally assumed that the climaticc conditions in the tropics facilitate the breakdown of bioactive compounds,, thus avoiding much of the side effects of pesticides. On the contrary,, we hypothesize that degradation products of bioactive compounds playy a prominent role under the tropical conditions and we use the River Nzoiaa basin in Kenya as a case study. Kenya, like most other developing countries,, is at the crossroads of an environmentally sustainable agricultural practicee and ostensible economically sustainable practices, characterised by aa high usage of agricultural chemicals. The current rural to urban migration necessitatess the use of herbicides as an easier alternative to the age-old labourr intensive manual weed control. Also, the pressure of a fast growing populationn calls for more intensification of agriculture with an attendant increasedd use of agricultural chemicals, in Kenya as well as in other tropical countriess (Lacher and Goldstein 1997). Pesticide use in Kenya is already one off the highest in sub-Saharan Africa with a market share of approximately US$$ 40.4 million by 1992 (Partow 1995).

Afterr farm application, pesticides and their degradation products dissipatee into other environmental compartments including groundwater, surfacee water, and the atmosphere. Degradation of chloroacetanilide herbicidess after use has been shown to result in several products with 2,6-diethylanilinee and 2-ethyl-6-methylaniline being the environmentally stable endd products for alachlor and metolachlor, respectively (Kimmel et al. 1986; Tessierr and Clark 1995; Gonzalez-Barreiro et al. 2000). The rate and extent off degradation depends on the microbial composition of soil or sediment (Bollagg et al. 1986; Liu et al. 1995; Stamper and Tuovinen 1998), temperaturee (Hamaker 1972; Gerstl 1991), moisture content (Gerstl 1991), hydroxyll radical (*OH) generating processes (Webster et al. 1998), and uptakee and degradation by resistant plants (Feng 1991; Field and Thurman 1996;; Aga and Thurman 2001). A common soil fungus, Chaetomuin globosumglobosum (Tiedje and Hagedorn 1975) and sediment inhabiting chironomid larvaee (Wei and Vossbrinck 1992) have also been shown to degrade chloroacetanilides.. Variability in environmental half-lives of organic pesticidess at different climatic conditions is widely reported (Webster et al. 1998).. It is possible that the herbicides degrade faster in the tropics due to

24 4

ChapterChapter 2

thee higher ambient temperatures, conducive for microbial activities, metabolismm and/or photolysis. The chloroacetanilide degradation products havee been detected in surface and ground waters of the temperate agriculturall zones (Phillips et al. 1999; Scribner et al. 2000), and due to theirr persistence, they are important environmental contaminants. Moreover, thesee chloroacetanilide stable aniline degradation products are more or equallyy toxic compared to their parent pesticide compounds (Osano et al. 2002b;; Osano et al. 2002a) and are promutagens in the Ames test (Kimmel etet al. 1986). Hence, we investigated the extent of contamination of river Nzoiaa with the chloroacetanilide herbicides alachlor, metolachlor and their environmentallyy stable aniline degradation products 2,6-diethylaniline, and 2-ethyl-6-methylanilinee and reconstructed the fate of the pesticides in the basin.. The sequel of heavy use of the chloroacetanilides is collated with possiblee ecotoxicological effects of the compounds.

Thee Study Area

Thee Nzoia River basin is approximately 1 269 600 hectares and lies entirelyy within the lake Victoria basin in Kenya, East Africa. It is bounded byy the latitudes: 34° - 36° east, longitude: 0°03' - 1°15' north, and lies betweenn 1134 - 2700 m above sea level. It encompasses three geographical regions:: the highlands around Mount Elgon and the Cherangany Hills, the upperr plateau, which includes Eldoret, and the lowlands. The region receivess an average of 1350 mm/year of rain and is an important cereal and sugarcane-farmingg region of Kenya producing at least 30% of the national outputt of both maize and sugar. Potential major sources of pollution for the riverr are the agricultural chemicals, urban effluents of Eldoret (population 234000),, Kitale (pop. 88100), Bungoma (pop. 32900), Webuye (pop. 45100),, Kakamega (pop. 86500), Mumias (pop. 36200), industrial wastes of thee Panpaper pulp mills at Webuye, textile factories in Eldoret, coffee factoriess scattered in the higher regions, and sugar industries mainly at Mumias,, Kakamega and Bungoma districts. The polluting role played by anyy of these factors is unknown. Herein, we study the role played by agriculturall chemicals, specifically the chloroacetanilide herbicides, as agriculturee is still the predominant activity in the region (Fig 1). The total lengthh of the river is ca. 252 km with average fall of 4 per 1000 (JICA 1987). .

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26 6

ChapterChapter 2

Ninee stations were sampled along the river and its tributaries. Stationn 1 is a pristine upstream site, although it is located amidst an upland agriculturall land. It drains a protected natural forest reserve and the peak of Mt.. Elgon. Contamination of the river at this point is unlikely. Station 2 lies withinn large-scale maize farms in Endebes. Station 3 drains large-scale maizee and wheat farms of approximately 15000 ha., which use herbicides almostt exclusively to control weeds. The site drains the Agricultural Developmentt Corporation farm (approx 11000 ha.), the Kenya Seed farm andd Kitale Municipality. Station 4, drains approximately 100000 ha. watershed,, all the above stations and some additional large-scale and small-scalee maize farms. The large-scale maize farms probably use chemicals to controll weeds. Station 5 is a detour of the river at station 4 into a swamp. Stationn 6, draining approximately 1072700 ha. watershed, lies 1 km downstreamm of the Mumias Sugar factory effluent disposal point and drains alll the above stations and in addition, sugarcane cultures of Nzoia and Mumiass Sugar factories, and Eldoret Municipality. Station 7 is on a tributaryy (approx. 3 m wide and 1.5 m deep) within the nucleus farm (3400 ha.)) of the Mumias Sugar factory. Herbicides are used almost exclusively to controll weed in the nucleus farm. Station 8,1184900 ha. watershed, drains alll the above station in addition to the Kakamega municipality and other small-scalee farms, which do not use herbicides. This accounts for about 90% off the total Nzoia watershed. Station 9, 1296900 ha. watershed, is the river deltaa where the river joins the lake through a swamp.

Sampling g

Waterr (2.5 litres) and sediment (500 grams) samples were obtained fromm stations 1 -9 between March 1998 and August 1998. Each site was sampledd on at least 3 different days. The herbicides are applied pre-emergentt in the maize and wheat fields before the annual pulse of the long rainss occurring between April to June and in the sugarcane before the long orr short rains (occurring in ca. October). All sampling stations (except stationss 5 and 7) drain the preceding sampling stations on the river. Water wass collected in precleaned amber bottles with Teflon-lined caps and the sedimentt was carried in black polyethylene bags. At all sites, samples were collectedd from approximately 15 cm depth of the river. Sediment was obtainedd by scraping the upper ca. 10 cm of the sediment at the bottom of thee stream using a shovel. The samples were transported to the laboratory

27 7


withinn 24 h in cool-boxes. Sediment and acidified (HN03) water samples weree stored in a freezer before later use.

Chemicall analysis

Alachlorr (molecular weight: 269.77, purity: 99%), metolachlor (molecularr weight: 283.80, purity: 99%), 2,6-diethylaniline (molecular weight:: 149.24, purity: >98%) and 2-ethyl-6-methylaniline (molecular weight:: 135.21, purity: >97%) were obtained from Reidel de Haen Fluka (Seelze,, Germany). Analytical grade «-hexane, dichloromethane and acetonee were obtained from Sigma chemicals (Kenya). Anhydrous sodium sulphatee was purified by soxhlet extraction with dichloromethane.

Thee extraction and cleanup procedures for water and soil samples weree done according to Sanchez-Brunete et al. (1994) with some modification.. In brief, excess water was discarded, leaving 2000 ml in the amberr sampling bottles, to be subsequently mixed with saturated sodium sulphatee (50 ml). Thereafter, the mixture was shaken with a dichloromethane:«-hexanee (320 ml; 50/50 [v/v]) mixture for 5 min to extract thee pesticides. After settling down, the supernatant was decanted into a 1-litree separating runnel and allowed to settle. The bottom layer comprising waterr was decanted off and the pesticide containing supernatant was subsequentlyy cleaned and concentrated.

Thee sediment samples were dried at room temperature for 4 days andd sieved (mesh size 0.5 mm) to obtain small-sized particles. Pesticides weree extracted from the sediment (40 grams) by adding a mixture of acetone andd tf-hexane (40 ml, 50/50 [v/v]) (Funari et al. 1998). The extract was centrifugedd (1000 rpm) for 20 min. The procedure was repeated and the clearr supernatants were pooled in 200 ml flasks for a subsequent clean up andd concentration procedures. Moisture contents of the remaining sediment sampless were determined by deducing the differences between the oven dry (att 105°C for 2 h) and the room dry weights of the samples. The recoveries off the compounds under study were reported to be approximately 90% (Funarii et al. 1998) and were not verified here. Thus, concentrations reportedd in this study were not corrected for recovery.

Thee extracts were cleaned up by passing them through a column (10 mmm internal diameter) packed with preheated (210°C, 3 h) and cooled florisill (5 cm), and anhydrous granular Na2S04 (5 cm), respectively. The

28 8

ChapterChapter 2

columnn was eluted with diethyl ether (20 ml) and the extracts plus the eluate weree concentrated by drying under a vacuum in a round-bottomed flask at 45°C.. The products were reconstituted with «-hexane (5 ml) and stored in amberr vials in a deep freezer. Analysis of the pesticides was performed usingg a Varian1 Gas Chromatograph, model 3400CX equipped with an electronn capture detector and a 30 m by 0.2 mm column fused with 0.2 urn silicaa film. An isothermal temperature program of 200°C, 250°C and 250°C forr the column, injector and detector, respectively, was applied. Flow rate of thee carrier gas (N2) was maintained at 5 ml/min. Identification of alachlor, metolachlor,, 2,6-diethylaniline, and 2-ethyl-6-methylaniline was based solelyy on their retention times, and the quantities of the compounds were deducedd from the peak areas. The respective minimum detection limits for alachlor,, metolachlor, and degradation products in water were 0.016, 0.004, andd 0.008 ug/L and in sediment 0.198, 0.047, and 0.097 ug/Kg.

Concentrationss of the pesticides and their degradation products

Figg 2 (A - D) shows chromatograms with peaks of alachlor (RT = 2.166 0.002), metolachlor (RT = 2.63 0.003), 2,6-diethylaniline (RT -3.300 0.003), and 2-ethyl-6-methylaniline (RT - 3.30 0.003), respectively,, in standard preparations (Fig 2A) and field samples (Fig 2B -2D).. Both 2,6-diethylaniline and 2-ethyl-6-methylaniline eluted at the same time.. Thus, the test method proved not selective for the two degradation products,, in agreement with previous reports (Tadeo et al. 1996). Degradationn products occurred in all water and sediment samples except the stationn 1 sample (Fig 2B). Alachlor, metolachlor, and a probable combinationn of 2,6-diethylaniline and 2-ethyl-6-methylaniline were detected inn 12.5, 2.1, and 93.8%, respectively, of the water and in 13.9, 2.8, and 97.2%% of the sediment samples, respectively. The pesticides and their degradationn products occurred at higher concentrations in the sediment than inn the water samples (1.4 - 10800 fold) indicating an accumulation of the compoundss in the sediment, Table 1. The degradation products were present throughoutt the entire sampling period inferring a stability of the compounds inn the sediment or a continuous input of the pesticides into the river.

29 9


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30 0

ChapterChapter 2

Dissipationn of the pesticides and their degradation products

Thee relatively low detection frequency (<14%) of the parent compounds in thee examined samples and the high detection rates of the degradation productss (>90%) suggest that the parent compounds are metabolised quickly afterr application (Table 1). In a study conducted in a temperate agricultural catchment,, alachlor and metolachlor were recovered from a higher percentagee (>25%) of the water samples (Clark and Goolsby 2000). The movementt of the pesticides and their degradation products from the field of applicationn by leaching or runoff depends on their water solubility and partition-drivenn adsorption to the soil (Aga and Thurman 2001). Torrential rains,, a characteristic of the tropics, generally lead to soil erosion, which mayy produce major pulses of the pesticides and their degradation products intoo surface water as both the soil adsorbed and dissolved organic compoundss are moved along in the flash floods. Volatilizations of the compounds,, ambient atmospheric temperature, and wind will each contributee to movement of the pesticides from the point of application to remotee locations (Goolsby et al. 1997; Clement et al. 2000).

Thee solubility of alachlor, metolachlor, 2,6-diethylaniline and 2-ethyl-6-methylanilinee in water at 25°C is 242, 488, 670 and 538a mg/L, respectivelyy (Lyman 1982; Tadeo et al. 1996; Laabs et al. 2000), while their Kocc (octanol-carbon partitioning coefficient) are 312, 244, 357 and 197, respectivelyy (Fava et al. 2000). Therefore, the degradation products are generallyy more soluble than the parent compounds, even though they bind moree to the soil reducing their leaching potential (Gustafson 1989; Fava et al.al. 2000). The soil organic matter content and the nature of the organic matterr affect the adsorption of the organic compounds on the soil, aromaticityaromaticity of the humic material being the key structural parameter that regulatess the sorption of the nonionic pesticides (Ahmad et al. 2001). The Kocc (octanol-carbon partitioning coefficient) values are strongly dependent onn aromaticity and negatively correlated with alkyl carbon components (Ahmadd et al. 2001).

Inn presence of sunlight and nitrate or dissolved humic acid, water is photolysedd to liberate hydroxyl radical ("OH) which degrades organic pollutantss including the chloroacetanilides (Marbury and Crosby 1994; Brezonikk and Fulkerson-Brekken 1998) thus enhancing the quantities of the

estimatedd by Hyperchem

31 1


degradationn products. It can be deduced from the high concentration of degradationn products in the sediment (Table 1) that River Nzoia is already contaminatedd with the pesticide degradation products and that potential risk off higher contamination is substantial, especially with increased intensificationn of agriculture in the region. It is concluded that in our study thee environmentally stable aniline metabolites of the pesticides and not the parentt compounds were detected in the river water and sediment in consistencee with the expected environmental behavior of the compounds in thee area studied, given the higher ambient temperatures, the water regime andd the agricultural practice.

Riskss of the pesticides and their degradation products in the tropic s s

Thee tropical environmental conditions may enhance the toxicity of thee chloroacetanilides and the risk of exposure to toxicants may be greater withh higher temperatures, humidity and UV radiations all characteristics of thee tropics (Viswanathan and Krishna Murti 1989; Blaustein and Wake 1995;; Bossan et al. 1995; Abdullah et al. 1997; Heugens et al. 2001; Wiegmann et al. 2001). UV radiation, for instance, increases rate of frog malformationn (Blaustein and Wake 1995) and UV may also weaken bonds betweenn adsorbed chemicals thereby enhancing their bioavailability (Bossan etet al. 1995). The parent and degradation products of the chloroacetanilides weree acutely toxic to Vibrio fischeri, genotoxic to Vibrio fischeri (Mutatox* test),, teratogenic to Xenopus laevis, and promutagenic in the Ames test (Kimmell et al. 1986; Osano et al. 2002b; Osano et al. 2002a). Therefore, bothh the parent and the degradation products are toxicologically important, however,, in our study the degradation products proved more important pollutantss of the river than their parent compounds. The standards set for waterr quality analyses of pesticide contamination have hitherto underrated thee importance of the chemicals' degradation products. Our study showed valuess for the degradation products that exceeded the European Economic Commissionn limit of 0.1 and 0.5 u.g/L set for any individual compound and totall pesticides for drinking water, respectively (EEC 1998), and a lower valuee for the parent compounds in most samples. This finding differs from thosee in the temperate regions, where both the degradation products and parentt compounds are found as environmental contaminants (Galassi et al. 1996).. In view of toxic effects of the degradation products and apparent

32 2

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propensityy for their formation in the tropics, we propose that risk assessmentss and water quality analyses in the tropic should routinely include pesticidee degradation products.

References s

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Osanoo O, Admiraal W, Klamer HJC, Pastor D, Bleeker EAJ (2002b). Comparativee toxic and genotoxic effects of chloroacetanilides, formamidiness and their degradation products on Vibrio fischeri and ChironomusChironomus ripahus. Environ Pollut (in press).


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Weii LY, Vossbrinck CR (1992). Structure-toxicity relationships for the fatheadd minnow, Pimphales promelas to narcotic industrial chemicals.. JAgric Food Chem 40: 743-748.

Wiegmann S, Van Vlaardingen PLA, Bleeker EAJ, De Voogt P, Kraak MHS (2001).. Phototoxicity of azaarene isomers to the marine flagellate DunaliellaDunaliella tertiolecta. Environ Toxicol Chem 20: 1544-1550.

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ChapterChapter 5

Comparativee Toxic and Genotoxic Effects of Chloroacetanilides,, Formamidines and their Degradationn Products on Vibrio fischeri and

ChironomusChironomus riparius

Abstract t

Toxicc and genotoxic effects of alachlor, metolachlor, amitraz, chlordimeform,, their respective environmentally stable degradation products 2,6-diethylaniline,, 2-ethyl-6-methylaniline, 2,4-dimethylaniline, and two otherr related compounds, 3,4-dichloroaniline and aniline were compared. Acutee toxicity tests with C. riparius (96 h) and V. fischeri (Microtox ) and genotoxicityy tests with a dark mutant of V. fischeri (Mutatox®) were carried out.. Our results demonstrate that toxicity and genotoxicity of the pesticides aree retained upon degradation to their alkyl-aniline metabolites. In the case off the herbicides alachlor and metolachlor, the toxicity to V. fischeri was enhancedd upon degradation. Narcosis alone explains toxicity of the compoundss to the midge, but not so for the bacteria suggesting a disparity in thee selectivity of the test systems. All compounds showed direct genotoxicityy in the Vibrio test, but amitraz and its metabolite were genotoxic att concentrations 103 - 105 lower than all the other compounds. The observationss indicate that stable aniline degradation products of the pesticidess may contribute considerably to environmental risks of pesticides applicationn and that genotoxic effects may arise upon degradation of pesticides. .

39 9

ToxicityToxicity and genotoxicity of pesticide degradation products

Introductio n n

Thee chloroacetanilides alachlor and metolachlor are some of the mostt widely used selective herbicides worldwide in corn, soybean and other cropp cultures (Krause et al. 1985; Nesnow et al. 1995). Elevated concentrationss of these herbicides and their degradation products have been detectedd in surface and groundwater (Nielsen and Lee 1987; Wang et al. 1995).. In groundwater, the concentration of the degradation product 2,6-diethylanilinee can be more than two times that of the parent compound alachlorr (Potter and Carpenter 1995). The formamidines amitraz and chlordimeformm are used to control insects and mites. While the use of the latterr is banned worldwide there is a notable increased use of amitraz in the cattlee industry in Kenya. Both chloroacetanilides and formamidines are metabolisedd in intraperitoneally treated rats by the hepatic mixed function oxidasee systems to 2,4 and 2,6 disubstituted anilines (Kimmel et al. 1986). Thesee are further converted to the corresponding nitrosobenzenes, which are mutagenicc in the Ames test (Kimmel et al. 1986). 3,4-DichloroaniHne is a commonn degradation product of many herbicides including diuron and linuronn (Crossland 1990) used alongside the chloroacetanilides in agriculturee and aniline bears the basic unsubstituted aniline moiety of the abovee test compounds. The pesticides and their aniline metabolites are suspectt or confirmed oncogens (Weisburger et al 1978; USEPA 1990). Alachlorr use is banned in Canada because of its mutagenicity and potential carcinogenicityy (Hoberg 1990) while in Kenya its use is still unrestricted (Partoww 1995). Its genotoxic effects have been observed in plants, insects, yeastt and mammals, but not in bacteria (Kimmel et al. 1986). Previously we observedd that the aniline degradation products of the two chloroacetanilides inducedd developmental aberrations in the embryos of the frog Xenopus laevislaevis (Osano et al. 2002).

Thee 'target' theory (Hansch and Fujita 1963) states that bioactivities,, in our case toxicity, occur as a result of interactions between toxicantss and receptors in the organism. The toxicity will depend on the bioavailabilityy and chemical reactivity of the compounds. Based on this, lipophilicityy (Log KoW) and different parameters for chemical reactivity of thesee chloroacetanilides, formamidines, their corresponding degradation productss and two related compounds, 3,4-dichloroaniline and aniline are examinedd here to discuss specific mechanisms of toxicity to midges {Chironomus{Chironomus riparius) and bacteria (Vibrio fischeri). The concentration that

40 0

ChapterChapter 3

iss lethal to 50% of the organisms after 96 h (LC50 [96 h]), the concentration thatt causes 50% reduction in luminescence after 15 minutes (EC50 [15 min])) and lowest effect concentration (LEC) for induction of luminescence aree determined for C. riparius, V. fischeh (Microtox*), and a dark mutant of V.V. fischeri (Mutatox*), respectively. The selectivity in these different biologicall test systems is compared and the risk of the tested pesticides and theirr environmentally stable degradation products is discussed.

Material ss and Methods

Chemicals Chemicals

Thee test chemicals, alachlor (ALA), metolachlor (MET), amitraz (AMI) ,, chlordimeform (CHL) (purity: >97%), 2,6-diethylaniline (DEA), 2-ethyl-6-methylanilinee (EMA), 2,4-dimethylaniIine (DMA), 3,4-dichloroanilinee (DCA) and aniline (ANI) (purity: >99%), and analytical grade cosolvents,, methanol, acetone and dimethylsulfoxide (DMSO) were obtained fromm Fluka Riedel-de Haên. The Dutch Standard Water (DSW) medium for thee C. riparius tests was prepared from analytical grade salts as 0.2, 0.18, 0.1,, 0.02 g/L of CaCl2.2H20, MgS04.7H20, NaHC03, and KHC03, respectively,, all dissolved in deionised water. Microtox& and Mutatox* kits weree obtained from Azur Environmental (Carlsbad, CA, USA). Table 1 showss the Log K^ 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, highestt occupied molecular orbital (HOMO), lowest unoccupied molecular orbitall (LUMO). Chemical reactivity factors were calculated by the Hyperchemm (Hypercube Inc., Version 6.1) computer program. Prior to the calculations,, the structures of the compounds were drawn and AMI optimised.. For the Log K<,w, surface area and volume calculations these structuress were further optimised by MM+.

Chironomuss riparius acute toxicity test

Larvaee used in each of the tests were taken from a C. riparius culturee at the Department of Aquatic Ecology and Ecotoxicology (IBED, Universityy of Amsterdam). Triplicate 96 h static acute toxicity tests were

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carriedd out in 180 ml (6.5 cm diameters) glass vessels filled with 100 ml test mediumm and 50 first instar larvae of less than 24 h old. The larvae were fed 11 ml fish food suspension (containing 52.5 g/L Tetraphil®/Trouvit® 1:20 w/w)) and the medium was constantly aerated to maintain the O2 concentrationn at >40%. The vessels were covered with a perforated polythenee film to minimise evaporation of the medium. An incubation temperaturee of 20 1°C and a 16:8 h light:dark regime was maintained duringg the experiment. Test compounds were dissolved in DMSO and added too DSW to prepare a stock solution such that the final test concentration of DMSOO did not exceed 80 ul/L. The dissolution of the compounds was facilitatedd by sonication for 30 min. Test dilutions were made by addition of DSWW containing 80 ul/L DMSO to the various quantities of the stock solution.. The concentrations of the chemicals tested were verified by HPLC (highh performance liquid chromatography) at the start and at the end of the experiments.. The actual exposure concentration was deduced as the geometricc mean of start and end concentrations in the media. The LC50 and itss 95% confidence intervals for the compounds were deduced from Kaleidograph®® for windows (Synergy Software, version 3.08 for Windows, Reading,, UK) fittings using the logistic response model (Haanstra et al. 1985): :

YY = c/( l+expb ( X -a ))

wheree iY' = response (percentage survival), ' c' = control response (set to 100%),, Lb' = slope, 'X' = log of the exposure concentration and 'a' = log LC50. .

Vibrioo fischeri acute toxicity test (Microtox*)

Thee Microtox® tests were done according to the supplier's protocol (Azurr Environmental Corporation 2000a). In brief, the test compounds weree dissolved in the cosolvent methanol or acetone to make a stock solutionn with a 0.5% (v/v) cosolvent concentration and a 1:1 (v:v) dilution seriess of the stock solutions were made in Microtox* diluent (2% sterile NaCll solution). This mixture was incubated at 15 0.1°C for 15 min. Luminescencee was captured in the photoluminescence analyser at the beginningg and at the end of 15 min. The EC50 luminescence inhibition and

43 3


itss 95% confidence intervals after 15 min were deduced using an in-built programm in the Microbic model 500 program.

Vibrioo fischeri genotoxicity test (Mutatox*)

Mutatox** tests with and without S9 enzyme activation were done accordingg to the suppliers' protocol (Azur Environmental Corporation 2000a).. Guided by the results of the MicrotoxH tests above, we chose test dilutionss that included concentrations that were sublethal to V. fischeri. Ten 1:11 (v:v) serial dilutions of the stock solution, as in Microtox* above, were madee in Mutatox* Assay Medium (MAM) and in MAM plus rat S9 plus cofactorss for S9-mediated assays in cuvettes for each compound. The S9-mediatedd assays were incubated at 35 0.1 °C for 45 min to allow for S9 enzymee activation. The Mutatox8 tests were then incubated at 27 0.1 °C forr 24 h. The first luminescence reading was scored at 12 h and hourly until 244 h in the Microbic 500 analyser specially modified for the Mutatox* test. Inn this study 16-h readings were chosen for analysis. A positive genotoxic responsee was recorded when at least two dilution cuvettes showed a light outputt 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 lowestt effect concentration (LEC) was deduced by linear interpolation betweenn the dilutions with responses below and above 100 light units.

Resultss and Discussion

Chironomuss riparius acute toxicity test

Thee parent pesticide compounds; ALA, MET, and AMI were 1.6, 2.1 andd 42.9 times more acutely toxic to 1st instar larvae of C. riparius than theirr corresponding alkyl-aniline degradation products DEA, EMA, and DMA, respectivelyy (Table 2). For AMI the LC50 (11.2 uM) for C. riparius was obtainedd by extrapolation as its LC50 lies above its maximum water solubilityy of 3.4 jaM. However, we observed that when a higher concentrationn of DMSO (1% instead of 0.008% v/v) was used to facilitate dissolvingg of AMI, a 100% mortality of the larvae in 1.7 uM occurred, while noo lethal effect was observed in a corresponding control with a similar

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concentrationn of DMSO. Thus, the well-known specific toxicity of amitraz ass a monoamine oxidase inhibitor in invertebrates (Aziz and Knowles 1973) wass only apparent after a facilitation of its transport by a cosolvent. It is concludedd that the importance of this compound as an acaricide or an insecticidee is greatly dependent on its solvent carriers and that its aquatic ecotoxicityy is limited by its low solubility in water and high Log KoW

Degradationn of amitraz to DMA greatly enhances its water solubility, while decreasingg its lethal toxicity.

Theree was a significant correlation between acute toxicity to the midgess 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 quantitativee structural activity relationship (QSAR) for 'baseline toxicity' usingg relatively non-reactive chemicals (2-propanol, 2-methyl-2-propanol, 3-pentanol,, 1,1,2-trichloroethane, toluene, 1,4-dichlorobenzene, 1,2,3-trichorobenzene,, 1,2,3,4-tetrachlorobenzene, pentachlorobenzene) and 3rd

instarr C. hparius larvae in a 48-h test (Fig 1 A, broken line, (Roghair et al 1994)).. The slopes of Roghair et al. (1994) and our fitting were not differentt (p = 0.051). The relatively lower (p = 0.026) elevation in the presentt study may be attributed to a higher test sensitivity arising from the usee of earlier stages of C. riparius larvae (1st instar) and the longer test durationn (96 h). None of the molecular reactivity parameters tentatively exploredd (see Materials and Methods) yielded a significant correlation with toxicity.. Clearly, in the Chironomus acute toxicity test, the pesticides and theirr environmentally stable aniline degradation products tested exerted theirr effects through narcosis (Fig 1A) and not by a specific mode of action, exceptt that the toxicity of amitraz may have been underestimated because off its low solubility.

CosolventCosolvent effects in Vibrio fischeri assays

Thee cosolvent strength is the ratio between pollutant concentrations att saturation in cosolvent-water mix to pollutant concentration at saturation inn water. Therefore the stronger cosolvents would facilitate a higher availabilityy of the test compounds to the bacteria. In both the Microtox* and Mutatox11 assays all the compounds elicited a stronger effect when methanoll was used as a cosolvent instead of acetone (Table 2), in spite of

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Correlationn of toxicity data with Log K<,w. A. Chironomus riparius LC50 (96h) (B) VibrioVibrio fischeri EC50 (15 min Microtox*). ALA: alachlor, MET: metolachlor, AMI : amitraz,, DEA: 2,6-diethylaniline, EMA: 2-ethyl-6-methylaniline, DMA: 2,4-dimethylanilinee and AM: aniline. Bold line: linear regression fit and thick dotted lines:: 95% CLs for our test. Broken line: Roghair, (1994) fitting for QSAR of 'Baselinee toxicity' for C. riparius and thin dotted lines: its 95% CLs. AMI data was nott considered in our fitting in graph A. For toxicity of amitraz; see text

47 7

ToxicityToxicity andgeno toxicity of pesticide degradation products

acetone'ss higher cosolvent strength (Pastor i Rodrigues 1998). This was probablyy due to the higher toxicity of acetone (lowest observed effect concentrationn (LOEC) = 600 uM compared to methanol, LOEC = 1000 uM)) to V. fischeri (Azur Environmental Corporation 2000b), thereby reducingg the sensitivity of the test. Reported toxicity data in literature shouldd therefore take account of the cosolvents used. The toxicity values in methanoll cosolvent are considered hereinafter.

Vibrioo fischeri (Microtox* test)

Bioluminescencee of V. fischeri bacteria is controlled by a populationn density-responsive regulatory mechanism called quorum sensing (Dunlapp 1999). Cytotoxic compounds (bacteriocides) or inhibitors of any of thee enzymes in the bioluminescence mechanism will result in luminescence reductionn (Dunlap 1999). The degradation products DEA and EMA were 20.9 andd 6.3 times more acutely toxic than their corresponding parent compoundss ALA and MET (Table 2). These results contrast with those for C. ripariusriparius where the parent compounds showed higher toxicity than their respectivee degradation products (Table 2). Amitraz affected the Vibrio test onlyy at concentrations exceeding its solubility in water, while the degradationn product DMA was slightly more toxic in the Vibrio test than in thee midge test.

Thee role of lipophilicity (Fig IB) and other chemical reactivity parameterss (see Materials and methods) in toxicity in the Microtox* test couldd not be established, suggesting a more complex mechanism without anyy clear limiting step. The degradation products of the herbicides alachlor andd metolachlor were more toxic than the parent compounds despite their predictedd low affinity for uptake in the bacterial membranes by their low Logg 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

valuess for all the degradation products were below 3. Hermens et al. (1985) reportedd that Microtox* was not sensitive to very lipophilic compounds and Shustermann (1992) observed higher toxicities to Salmonella typhimurium of substancess with intermediate Log K^ values. Hence, lipophilicity alone doess not explain toxicity of the aniline based pesticides and their degradationn products in the Microtox test.

Oncee transported into the cytoplasm, the aniline based compounds (parentt compounds and degradation products) may have acted by inhibiting

48 8

ChapterChapter 3

essentiall bioluminescence enzymes. Hermens et al. (1985) suggested that molecularr refractiviry (a steric factor) of aniline and dichlorinated anilines couldd negatively influence the activity of the luciferase enzyme or reduce thee flavin mononucleotide (FMNH2) donating system of V. fischeri. The dependencee of the activity of the degradation products on lipophilicity as welll as the evidence of the hydrophobic site(s) on NADPH oxidase, an essentiall enzyme in the bioluminescence production (Nakata et al. 1997), furtherr suggest a possible enzyme inhibition mechanism of toxicity. This enzymee inhibition of bioluminescence may explain why toxicity of aniline compoundss to V. fischeri bacteria was found to be unrelated to their Log Koww values as opposed to E. coli (Jaworska and Schultz 1994). Toxicity of substitutedd nitrosobenzenes including N02, OH, NH2, OCH, methyl, and halogenn substituents to a river bacteria inoculate and V. fischeri was found too be controlled mainly by the electronic factors (Yuan et al. 1997). We suggestt that toxicity or bioluminescence inhibition depends on transport of thee toxicants to the cell membranes bilayer as well as transport from the membranee to luciferase enzyme receptors and agree with Huang et al. (1996)) that toxicity of the anilines to V. fischeri could be controlled by electronicc factors and Log K ^ combined. The limited set of our compounds hamperedd a search for correlations between toxicity and electronic factors.

Vibrioo fischeri genotoxicity (Mutatox® test)

Thee Mutatox* genotoxicity test is affected by sublethal concentrationss of substances that damage or intercalate DNA, inhibit synthesiss of new DNA, are direct mutagens causing base substitution or frame-shifts,, or are SOS inducing agents (Ulitzur and Weiser 1981; Weiser etet al. 1981; Ulitzur 1983; Johnson 1992). All the compounds were genotoxicc in the direct test (Table 2) and appeared nongenotoxic when S9 activatedd (results not shown). Lipophilicity, physico-chemical and reactivityy parameters of the compounds (see Materials and Methods) did nott correlate with toxicity in the Mutatox®.

Activationn of the compounds with the S9 enzyme system apparentlyy eliminated their genotoxicity or gave rise to highly bactericidal productss or bioluminescence inhibitors that prevented expression of any effectss in the present test. It is also possible that the products of S9 bioactivationn were less able to cross the membrane barriers into the site of photoluminescencee activity within the cytoplasm (De Maagd and Tonkes

49 9

ToxicityToxicity and geno toxicity of pesticide degradation products

2000).. However, genotoxicity of ALA after S9 activation at 30 ug/L (Canna-Michaelidouu and Nicolaou 1996), a value 10 times lower than our lowest testt concentration, suggest that the products of S9 activation which includes nitrosobenzeness and hydroxylamines (Kimmel et al 1986) may have been tooo 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 Microtox* EC50 values lower than our test concentrationss (Deneer et al 1989; Yuan et al 1997). Furthermore, nitrosobenzeness are potential inhibitors of NADPH-oxidase (Nakata et al 1997),, an essential enzyme in the bioluminescence production. So, the evidencee available suggests that the lack of response after S9 activation in thee present study may have been a false negative.

Thee direct genotoxicity of the test compounds in the present study contrastss with the findings in the Ames test (Kimmel et al 1986) where the samee compounds tested had to be S9 activated before they could become mutagenic.. Direct reactions with or modification of the DNA may have beenn responsible for the direct genotoxicity of the compounds in the present study.. For example ALA and its metabolites are capable of forming DNA adductss (Brown et al 1988; Nesnow et al 1995) and Ribas et al (1995) foundd ALA to be genotoxic when evaluated in the single-cell gel electrophoresiss assay both with and without S9 activation. The range of the concentrationss of ALA tested by Canna-Michaelidou and Nicolaou (1996) mayy have been too low for substantial interaction with DNA. It is possible thatt the compounds provided modification(s) to the DNA that never lead to mutagenicc events in the Ames test or that the Mutatox* test depends on regulatoryy events in the Vibrio that differ from that in Salmonella.

Inn conclusion all the three test systems indicate that toxicity is retainedd after degradation of the pesticides to their environmentally stable anilinee compounds. For the midges this effect is likely to be controlled by narcosiss (baseline toxicity), while the Vibrio test showed very different selectivity.. The Microtox3 test revealed relatively higher toxicities for the degradationn products of the herbicides than their parent compounds. All the pesticidess and degradation products in the present study gave positive genotoxicc response in the Mutatox* test and should be considered potential genotoxicantss capable of altering gene functions. Special attention should bee given to amitraz and its degradation product, which were genotoxic at veryy low concentrations (<0.005 |J.M). The widespread application of

50 0

ChapterChapter 3

anilinee based pesticides, both herbicides and acaricides, in Kenya and elsewheree might lead to accumulation of relatively stable degradation products.. Ecotoxicological effects of these compounds in the environment cann be diverse due to the different selectivities in different test systems as observedd in the present study.

References s


Azurr Environmental Corporation (2000a). Azur Mutatox®. Azur Corporationn 2232 Rutherford Road Carlsbad, California 92008, USA. .

Azurr Environmental Corporation (2000b). The Microtox* Chronic Toxicity Testingg System, www.azurenv.com/chron.htm.

Brownn MA, Kimmel EC, Casida JE (1988). DNA adducts formation by alachlorr metabolites. Life Sci 43: 2087-2094.

Canna-Michaelidouu S, Nicolaou A-S (1996). Evaluation of the genotoxicity potentiall (by Mutatox™ test) of ten pesticides found as water pollutantss in Cyprus. Sci Total Environ 193: 27-35.

Crosslandd NO (1990). A review of the fate and toxicity of 3,4-dichloroanilinee in aquatic environments. Chemosphere 21: 1489-1497. .

Dee Maagd PGJ, Tonkes M (2000). Selection of genotoxicity tests for risk assessmentt of effluents. Environ Toxicol 15: 81-90.

Deneerr JW, Vanleeuwen CJ, Seinen W, Maasdiepeveen JL, Hermens JLM (1989).. QSAR study of the toxicity of nitrobenzene derivatives towardss Daphnia magna, Chlorella pyrenoidosa and Photobacteriumphosphoreum.Photobacteriumphosphoreum. Aquat Toxicol 15: 83-98.

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http://www.azurenv.com/chron.htm


Dunlapp PV (1999). Quorum regulation of luminescence in Vibrio fischeri. J MolMol Microbiol Biotechnol 1: 5-12.

Haanstraa L, Doelman P, Oude Voshaar JH (1985). The use of sigmoidal dosee response curves in soil ecotoxicological research. Plant Soil 34:: 293-297.

Hanschh C, Fujita T (1963). p -a -n Analysis. A method for the correlation off biological activity and chemical structure. Biochem J 87: 1616-1626. .

Hermenss J, Busser F, Leeuwangh P, Musch A (1985). Quantitative structure-activityy relationships and mixture toxicity of organic chemicalss in Photobacterium phosphoreum: the microtox test. EcotoxicolEcotoxicol Environ Saf 9: 17-25.

Hobergg G (1990). Risk, science and politics - alachlor regulation in Canada andd the United States. CanJPolitSci 23: 257-277.

Huangg Q, Kong L, Wang L (1996). Applications of frontier molecular orbitall energies in QSAR studies. Bull Environ Contain Toxicol 56: 758-765. .

Jaworskaa JS, Schultz TW (1994). Mechanism-based comparisons of acute toxicitiess elicited by industrial organic-chemicals in prokaryotic andd eukaryotic systems. Ecotoxicol Environ Sqf 29: 200-213.

Johnsonn BT (1992). Potential genotoxicity of sediments from the Great-Lakes.. Environ Toxicol Water Quality 7: 373-390.

Kimmell EC, Cassida JE, Ruzo LO (1986). Formamidine insecticides and chloroacetanilidee herbicides: disubstituted anilines and nitrosobenzeness as mammalian metabolites and bacterial mutagens. JAgricJAgric Food Chem 40: 1695-1699.

Krausee A, Hancock WG, Minard RD, Freyer AJ, Honeycutt RC, LeBaron HM,, Paulson DL, Liu S-Y, Bollag J-M (1985). Microbial transformationn of the herbicide metolachlor by a soil actinomycete. JAgricJAgric Food Chem 33: 584-589.

Nakataa M, Nasuda-Kouyama A, Isogai Y, Kanegasaki S, Iizuka T (1997). Effectt of aromatic nitroso-compounds on superoxide-generating activityy in neutrophils. J Biochem (Tokyo) 122: 188-192.

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Nesnoww S, Agarwal SC, Padgett WT, Lambert GR, Boone P, Richard AM (1995).. Synthesis and characterization of adducts of alachlor and 2-chloro-Ar-(2,6-diethylphenyl)acetamidee with 2'-deoxyguanosine, thymidine,, and their 3'-monophosphates. Chem Res Toxicol 8: 209-217. .

Nielsenn EG, Lee LK (1987). The magnitude and costs of groundwater contaminationn from agricultural chemicals. A national perspective. USS Dept. Agricultural Staff, Washington D.C, USA.

Osanoo O, Admiraal W, Otieno D (2002). Developmental disorders in embryoss of the frog Xenopus laevis induced by chloroacetanilide herbicidess and their degradation products. Environ Toxicol Chem 21:375-379. .


Pastorr i Rodrigues MD (1998). Optimization of bacterial bioassays for directedd chemical fractionation of environmental samples, pp. 1-106.. National Institute for Coastal and Marine Management (RIKZ),, Haren, The Netherlands.

Potterr TL, Carpenter TL (1995). Occurence of alachlor environmental degradationn products in groundwater. Environ Sci Technol 29: 1557-1563. .

Ribass G, Frenzilli G, Barale R, Marcos R (1995). Herbicide-induced DNA damagee in human lymphocytes evaluated by the single-cell gel electrophoresiss (SCGE) assay. Mutat Res 344: 41-54.

Roghairr CJ, Buijze A, Yedema ESE, Hermens JLM (1994). A QSAR for base-linee toxicity to the midge Chironomus riparius. Chemosphere 28:: 989-997.

Shustermann AJ (1992). Predicting chemical mutagenicity by using quantitativee structure-activity relationships. In: Food Safety Assessment.Assessment. American Chemical Society, Portland, OR, USA.

Ulitzurr S (1983). A bioluminescence test for genotoxic agents. Methods EnzymolEnzymol 133: 264-274.

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Ulitzurr S, Weiser I (1981). Acridine dyes and other DNA-intercalating agentss induce the luminescense system of luminous bacteria and theirr dark variants. Proc Natl Acad Sci 78: 3338-3342.

USEPAA (1990). Drinking water regulations and health advisories. Lewis Pubishers,, Chelsea MI, USA.

Wangg WC, Liszewski M, Buchmiller R, Cherryholmes K (1995). Occurrencee of active and inactive herbicide ingredients at selected sitess in Iowa. Water Air Soil Pollut 83: 21-35.

Weisburgerr EK, Russfield AB, Homburger F, Weisburger JH, Boger E, Vann Dongen CG, (1978). Testing of 21 environmental aromatic aminess or derivatives for long-term toxicity of carcinogenicity. J. Environ.Environ. Pathol. Toxicol. Oncol. 2: 325-352.

Weiserr I, Ulitzur S, Yannai S (1981). DNA-damaging agents and DNA-synthesiss inhibitors induce luminescence in dark variants of luminouss bacteria. Mutat Res 91: 443-450.

Yuann X, Lu G, Lang P (1997). QSAR study of the toxicity of nitrobenzenes too river bacteria and Photobacterium phosphoreum. Bull Environ ContamContam Toxicol 58: 123-127.

54 4

Lhjptei-^-Lhjptei-^-

Developmentall Disorders in Embryos of the Frog XenopusXenopus laevis Induced by Chloroacetanilide Herbicides

andd their Degradation Products

Abstract t

Pesticidess are known to transform in the environment, but so far the study of theirr effects in the environment has concentrated on the parent compounds therebyy neglecting the effects of the degradation products. The embryotoxic, developmentall and teratogenic effects of chloroacetanilide herbicides and theirr environmentally stable aniline degradation products were investigated in thiss study in view of the massive application of alachlor and metolachlor. Embryoss at midblastula to early gastrula stages of a locally abundant African clawedd frog Xenopus laevis were used as test organisms. The embryos were exposedd to the test chemicals for 96 h in each experiment. Alachlor is more embryotoxicc (the concentration causing 50% embryo lethality, 96-h LC50 = 233 \iM [6.1 mg/L]) and teratogenic (teratogenic index, [TI] = 1.7) than metolachlorr (96-h LC50 - 48 uM [13.6 mg/L], TI = 0.2). The degradation productss of alachlor and metolachlor, respectively, 2,6-diethylaniline (96-h LC500 = 13 iM [19.4 mg/L], TI = 2.1) and 2-ethyl-6-methyaniline (96-h LC50 == 509 uM [68.8 mg/L], TI = 2.7) are less embryotoxic but more teratogenic thann their parent compounds. The most common teratogenic effects observed weree edema for alachlor as opposed to axial flexures and eye abnormalities forr 2,6-diethylaniline and 2-ethyl-6-methylaniline. Metolachlor is found to be ann example of a nonteratogenic herbicide that upon degradation loses toxicity butt gains teratogenicity, and both the herbicides, metolachlor and alachlor, aree potential sources of teratogenic transformation products.

55 5

ChtoroacetanilidesChtoroacetanilides induced developmental disorders

Introductio n n

Thee chloroacetanilides (acetamides), alachlor (2-chloro-2',6'-diethyl-A^methoxymethyl)) acetanilide) and metolachlor (2-chloro-iV-2 (ethyl-6-methylphenyl)-Ar-2-methoxy-l-methylethyll acetamide) are widely usedd herbicides in agriculture (Schottier and Eisenreich 1994; Partow 1995; Potterr and Carpenter 1995; Galassi et al. 1996). They are used extensively inn the cereal and sugarcane growing zones of the Lake Victoria basin in Kenyaa and are among the most detected herbicides in surface waters elsewheree (Gruessner and Watzin 1995; Muller and Buser 1995). Alachlor hass a potential to induce cancer in laboratory animals (Kimmel et al. 1986) andd is classified as group B2 carcinogen by U.S. Environmental Protection Agencyy (USEPA 2000). The herbicides, alachlor and metolachlor, and their stablee aniline degradation product 2,6-diethylaniline and 2-ethyl-6-methylaniline,, respectively, are converted in hepatic mixed oxidase systems off rats to the corresponding nitrosobenzenes, which are mutagenic in the Amess Assay (Kimmel et al. 1986). The compounds are genotoxic in Mutatox™™ assay (Osano et al. 2002), which uses a dark mutant of Vibrio fischerifischeri bacteria as the test organism. The bacteria revert to a fluorescent statee when exposed to genotoxic compounds. The present study aims to extendd the understanding of the role of the degradation products of the chloroacetanilidess in embryonic development.

AA genetic program guides early development of a fetus, entailing expressionn and repression of successive series of genes, and hence, genotoxicc agents could be developmental toxicants as well (Bantle 1995). Developmentall toxicants and teratogens producing developmental disorders havee often been analyzed using embryos of Xenopus laevis, a clawed Africann frog. The frog embryo is an intact developing system, which undergoess events comparable to those of other vertebrates, including mammals.. Validation studies using known human and mammalian developmentall toxicants show that the predictive accuracy of tests with XenopusXenopus embryos for developmental toxicant approaches or exceeds 85% (Courchesnee and Bantle 1985; Dawson and Bantle 1987; Sabourin and Faulkk 1987; Bantle et al. 1989). Xenopus laevis is a native species of the Lakee Victoria basin and is therefore particularly suitable to analyze the potentiall risk of the widely used chloroacetanilides and their degradation productss in this region.

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ChapterChapter 4

Inn this article, we report on the comparative embryotoxic, developmentall and teratogenic effects of two important chloroacetanilide pesticidess and two of their stable aniline degradation products on the embryoss of clawed frogX laevis.


TestTest chemicals

Thee pesticides alachlor (99%) and metolachlor (99%) were obtainedd from Riedel-de Haen (Seelze, Germany). The pesticides' degradationn products, 2,6-diethylaniline (>98%) and 2-ethyl-6-methylanilinee (>97%) were obtained from Fluka (Buchs, Switzerland). Analyticall grade solvent dimethylsulfoxide (DMSO) was obtained from Fluka.. Test medium, usually referred to as frog embryo teratogenesis assay-XenopusXenopus (FETAX) solution, was prepared as 625 mg NaCl, 96 mg NaHC03,, 30 mg KC1, 15 mg CaCl2, 60 mg CaS04.2H20 and 5 mg MgS04

alll analytical grade per liter of distilled water (ASTM 1991).

TestTest organisms

Adultt X. laevis, claw footed frogs, were obtained from the ponds at thee shores of L. Victoria (Kenya) and raised at the School of Environmental Studies,, Moi University. The frogs were acclimatized in the laboratory for 6 weekss before the first breeding. They were housed in glass aquaria with dechlorinatedd tap water and nourished on minced beef fillet fortified with Multivitamin®® (Norbook, UK) and vitamin C supplements. They were fed twicee a week and aquarium water was replaced at the time of feeding. The aquariaa were maintained at a diurnal light:dark cycle of approximately 12:122 h, which is the natural equatorial cycle, and at a temperature of 22°C duringg the study period.

Embryoss were obtained from at least three different male/female pairss for each bioassay. Human chorionic gonadotropin hormone, Pregnil&

(Organon,, Oss, Netherlands), was injected into the dorsal lymph sac of each off the males (500 IU) and the females (1000 IU). Each male/female pair wass placed in a different mating/laying cage (0.3 x 0.3 * 0.3 m) half filled withh test medium. The bottoms of the cages were separated from the frogs

57 7

ChloroacetanilidesChloroacetanilides induced developmental disorders

aboutt 3 cm by use of a perspex floor with 1-cm holes. The laid eggs sink to thee bottom through this floor out of reach from the adult frogs, who would otherwisee quickly feed on the eggs. After approximately 14 h the adult frogs weree removed and separately the embryos from each pair were de-jellied by gentlee swirling in 2% (w/v) L-cysteine (CAS 52-90-4, the pH of which had beenn adjusted to 8.1 by use of 1 N NaOH) for 1 to 3 minutes. They were thenn washed in copious amounts of test medium. Embryos of stage 8 (midblastula)) to stage 11 (early gastrula) were selected for the tests. Nieuwkoopp and Faber's (1975) Normal Tables was used in staging of the embryos. .

TestTest protocol

Thee standard protocol for the FETAX (ASTM 1991) was used as a guideline.. The FETAX is a static-renewal assay, and each test medium is renewedd every 24 h. For each dilution two 60-mm diameter glass petri dishess containing 10 ml test medium and 25 embryos were used. Concurrently,, four control dishes of which two had <1% (v/v) DMSO and twoo were without, were tested in each experiment. Dishes were randomly assignedd to their positions in an incubator, and the incubation temperature inn all cases was set at 24°C (range of 22 - 26°C). At least three definitive testss were conducted on each test chemical in a random block design.

Alachlor,, metolachlor, 2,6-diethylaniline, and 2-ethyl-6-methylaniline,, were first dissolved in DMSO before being dispensed in the testt medium solution to make a stock solution of each of the test chemicals. Nominall concentrations of 10.1, 20.1, 40.2, 80.4, 167.5, and 335.0 uM 2,6-diethylaniline,, 7.4, 37.0, 92.4, 184.9, 369.8, and 739.6 uM 2-ethyl-methylanilinee 3.9, 7.8 16.2 32.4, 37.1, and 64.9 uM alachlor and 3.5, 17.6, 35.2,, 88.1, 176.2, and 352.4 uM metolachlor were tested. The final concentrationn of DMSO was <1% (v/v) and was uniform at all the test concentrations.. The 25 (~1- mm diameter) eggs in each test were assumed nott to significantly sorb the dissolved chemical within the 24 h between the renewalss of the test solutions. This assumption was supported by earlier toxicityy tests in which 50 to 100% of the same test compounds were recoveredd even after 4 d of incubation (Osano et ah 2002).

Forr each of the dishes the number of surviving tadpoles was recordedd daily. The dead ones were removed and the surviving larvae were

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ChapterChapter 4

notedd after every 24 h until 96 h of incubation. The embryos were then fixedfixed in 3% formalin and examined. The 96-h 50% lethal concentration (LC50)) and 96-h EC50 (malformation) defined as the concentration causing malformationn in 50% of the surviving embryos, were determined by probit analysiss (Wardlaw 1985) after Abbott's (1925) adjustment for mortalities andd malformations in the control as Y = 100 x (C - T)/C, where Y = percent response,, C = percent not responding in the DMSO control, T = percent not respondingg in the test dilution, and the response was considered as growth retardation,, mortality, or malformation accordingly. The teratogenic index (TI)) was deduced by dividing the LC50 by the EC50. Head to tail lengths of thee tadpoles were measured at the end of each experiment under a binocular dissectingg microscope at a magnification of 10 x 2 by use of an ocular micrometer.. Growth inhibition was deduced by determining whether growth att a particular concentration was significantly different from that of the control.. Minimum concentration to inhibit growth (MCIG) was taken to be thee minimum concentration of test chemical that significantly inhibited growthh as determined by measurement of head (mouth)-to-tail length (end off the tail).

Dataa were statistically tested with the Students' Mest and ANOVA (withh Bonferroni adjustment matrix of pairwise comparison probabilities) in thee SYSTAT® (1996) statistical program at the a = 0.05 level to determine anyy significant difference in growth of tadpoles.

Al ll the tadpoles were examined for developmental aberrations underr the dissecting microscope. The Atlas of Abnormalities (Bantle et al. 1990)) was consulted to determine which embryos were normal, which were abnormal,, and in the identification of the different kinds of aberration. An embryoo was considered abnormal if it exhibited at least one of stunting, poorr or incomplete gut coil, pericardial edema, abdominal edema, facial edema,, cephalic edema, tail flexure, notochord flexure, undulating notochord,, optic cap rupture, microphthalmia, poor heart development, or anyy other gross abnormality. An abnormality was scored regardless of the seriousnesss of the abnormality, e.g., a slight facial edema, large-gut edema, orr extensive edema were all scored as edema and so on. The frequency of eachh type of malformation at each test concentration for each chemical was recorded.. The reported incidence of each type of malformation was an Abbott'ss (1925) adjusted incidence for the similar type of malformation

59 9


occurringg in the DMSO control. Tadpoles were preserved in rubber-sealed vialss with 3% formalin.

Results s

SurvivalSurvival of embryos

Thee parent compounds alachlor and metolachlor are more acutely toxicc than their respective aniline degradation products 2,6-diethylaniline andd 2-ethyl-6-methylaniline (Table 1). Alachlor and metolachlor are 5.7 and 10.66 times more embryotoxic than their stable aniline degradation products, respectively.. The environmentally stable degradation product of alachlor 2,6-diethylanilinee is 3.9 times more acutely toxic than 2-ethyl-6-methylanilinee (metolachlor's environmentally stable aniline degradation product).. The EC50 and LC50 values are shown in the Table 1.

EffectsEffects on growth

Alachlorr reduced the growth of the tadpoles in a dose dependent manner,, while its degradation product 2,6-diethylaniline did not inhibit growthh in the tadpoles at all the concentrations tested, not even in the highestt tolerated concentration (Fig. 1). Metolachlor and its stable aniline degradationn product reduced growth at all the concentrations tested (Fig. 1). Bonferronii adjustment matrix of pairwise comparison probabilities showed noo significant difference (p = 0.05) between growth in the concentrations 3.5,, 17.6 and 44.0 for uM metolachlor and 7.4, 37.0, 92.4, 184.9 and 739.6 |aMM for 2-ethyl-6-methylaniline.

TypesTypes of aberrations

Inn alachlor-exposed embryos, the most frequent symptoms of developmentall disorders were facial, pericardial, and gut edemas (46%) (Fig.. 2). Gut malformation, axial flexures and eye abnormalities comprised 13,, 29, and 17%, respectively, of all the observed abnormalities in the embryoss exposed to alachlor. Similarly, in metolachlor-exposed embryos, edemaa (48%) was the most frequently observed disorder. Gut malformation,

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Figg 1. Percent growth relative to the growth in the DMSO control of embryos of XenopusXenopus laevis after 96-h exposure to different concentrations of herbicides and their degradationn products. Panel A: filled symbol, Alachlor; open symbol, 2,6-diethylaniline.. Panel B: filled symbol, Metolachlor; open symbol, 2,6-diethylaniline. Errorr bars show 95% confidence limits.

axiall flexures, and eye abnormality comprised 19, 22, and 11% respectively,, of the observed abnormalities in metolachlor-exposed embryos. .

Ass opposed to the parent compounds, the degradation products 2,6-diethylanilinee (6%) and 2-ethyl-6-methylaniline (7%) induced edema to a muchh lower extent (Fig. 2). The predominant aberrations induced by the degradationn products were the axial flexures, which comprised 52 and 49% off all the observed abnormalities for 2,6-diethylaniline and 2-ethyl-6-methylanilinee exposures, respectively. Gut and eye abnormalities comprised

62 2

ChapterChapter 4

33 and 39%, respectively, of the observed abnormalities for 2,6-diethylanilinee exposures and 11 and 33%, respectively, of the observed abnormalitiess for 2-ethyl-6-methylaniline exposures.

IncidenceIncidence rates of the aberrations

Theree was no significant difference between the frequency of the abnormalitiess reported in the control with (17.7%, 95% confidence limits [CL]] - 12.5 - 22.9) and the control without (21.6%, 95% CL = 17.3 - 25.9) DMSO.. In deducing the EC50 (malformation) Abbott's adjustment (Abbott 1925)) was used to account for the deformities that occurred in the concurrentt control with DMSO. The abnormalities accounted for included minorr changes like slight curvatures in the tail. Correction for the control deformitiess was further indicated since these were of different types than thosee induced by the test compounds. Alachlor induced higher frequency of aberrationss for similar ranges of concentrations than metolachlor and the degradationn products showed similar dose-response relationships at low concentrationss (<70 \xM). However, the tadpoles survived more in higher concentrationss of 2-ethyl-6-methylaniline than 2,6-diethylaniline. The highestt tolerated concentrations for the degradation products were 167.5 andd 739.6 u.M for 2,6-dithethylaniline and 2-ethyl-6-methylaniline, respectively.. The character of our four test compounds was compared using theirr TIs. The TI of alachlor, metolachor, 2,6-diethylaniline and 2-ethyl-6-methylanilinee was 1.7, 0.2, 2.1, and 2.7, respectively. Thus, alachlor and its degradationn product are teratogenic according to their TI values, whereas thee teratogenincity of metolachlor is only apparent after degradation to 2-ethyll -6-methy lani 1 ine.

Discussion n

Thee embryonic TI, defined as LC50/EC50, has been developed to alloww for comparison between known and suspected teratogens (Dumont et ahah 1983). Based on the TIs derived from standard compounds, Dumont et ahah (1983) have set TI value >2 as one that indicates the need for further testingg of the chemical. The TI value of 1.5 - 2.0 indicates that the materials shouldd be treated as potential teratogen and tested further in other screening systems,, while TI value <1.5 reflects compounds that are more embryolethal,, suggesting that they are co-effective teratogens (Johnson

63 3


1981).. Their lethality may be more pertinent to risk assessment than teratogenicity.. Therefore, alachlor (TI = 1.7) is a potential teratogen, while itss stable aniline degradation product, 2,6-diethylaniline (TI = 2.1), is a teratogenn that requires further testing in other assays to establish its teratogenicity.. Metolachlor (TI = 0.2) is clearly not teratogenic, as opposed too its stable aniline degradation product, 2-ethyl-6-methylaniline, which is thee most teratogenic of the compounds in this study with a TI of 2.7. Three-foldd higher concentrations of 2-ethyl-6-methylaniline are needed to induce thee 50% of malformation as in 2,6-diethylaniline, even though at lower concentrationss (<60 uM) the two compounds induced similar percentages off aberrations.

DD Cut malformations Edema DD Axial flexures E2 Eye abnormalities

AL AA DEA MET EMA

Figg 2. The relative frequency of different types of malformations in Xenopus laevis afterr 96-h exposure to the chloroacetanilides and their stable aniline degradation products.. The frequency is expressed as the percent of the total number of observedd abnormalities in embryos accumulated for all test concentrations for eachh chemical. ALA : alachlor, DEA: 2,6-diethylaniline, MET: metolachlor, EMA:: 2-ethyl-6-methylaniline.

Alachlorr strongly induces excision repairable DNA lesions. Surralless et al. (1995) also found by use of fluorescence in situ hybridizationn and an antikinetochore antibody that alachlor is a clastogen

64 4

ChapterChapter 4

actingg in the S phase. It has been proposed that the genotoxicity of many organicc pesticides could be mediated by their alkylating potentials, and sincee alachlor has alkylating radicals, it may act as a DNA alkylating agent (Surralless et al. 1995). It is oncogenic (Hoberg 1990), and it forms adducts withh DNA (Nesnow et al. 1995). Consistent with earlier observations, the presentt study shows that alachlor is a suspect teratogen. Moreover, its stablee degradation product, 2-6-diethylaniline, is similarly teratogenic. Alachlorr and metolachlor are structurally related, their major difference beingg the methoxy alkyl chain attached to the nitrogen atom of the basic structure.. This relatively small difference in the molecular structure is apparentlyy sufficient to impart some desirable properties to metolachlor as ann herbicide. Some of these properties are high lipid (octanol-water partitionn coefficient) and water solubility for metolachlor (Chester et al. 1989).. Lipophilicity is often associated with chemical persistence in the environment.. Metolachlor is transformed in the soil to a lesser extent than alachlorr (Konopka 1994). It is more effective than alachlor and other chloroacetanilides;; therefore, its global demand has risen. Both alachlor and metolachlorr revert to their intermediates of manufacture, 2,6-diethylaniline andd 2-ethyl-6-methylaniline, respectively, in the mammalian metabolism (Kimmell et al. 1986) and upon environmental degradation (Tiedje and Hagedornn 1975; Wei and Vossbrinck 1992; Konopka 1994; Liu et al. 1995; Mullerr and Buser 1995). This study confirms that metolachlor has less adversee effects than alachlor, but caution in its use should be observed as its stablee aniline degradation product, 2-ethyl-6-methylaniline, showed the mostt prominent teratogenic capacity in the present study.

Alachlorr and metolachlor are some of the most widely used herbicidess in agriculture in Kenya (Partow 1995) and worldwide (Tessier andd Clark 1995), metolachlor being the most widely used herbicide in North Americaa (Muller and Buser 1995; Surralles et al. 1995) after banning of alachlorr in Canada (Hoberg 1990). We suggest that the criteria for establishmentt of safety of chloroacetanilides should include both acute and chronicc effects of the parent compounds and their degradation products. So far,, the present study indicates risk associated with the use of metolachlor, (i.e.. teratogenicity after breakdown), which before have been associated withh alachlor only and which led to banning of this latter product. Xenopus, whichh inhabits the fresh water ponds in the farming regions of Kenya, offer aa potential as an ideal organism for the study of the effects of suspect

65 5


teratogens.. This test organism is important because of the massive applicationn of the chloroacetanilides in Kenya and the potential liberation off the degradation product. Furthermore, the present study and related ones mayy provide additional clues on factors causing the decline in the populationss of amphibians worldwide (Wake 1991; Baustein and Wake 1995)) and explain the increased incidence of abnormalities in the natural populationss of frog species (Burkhart et al. 2000).

References s

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ASTMM (1991). Standard guide for conducting the frog embryo teratogenesis assay-Xenopusassay-Xenopus (FETAX) E 1439-91, pp. 826-836, Annual Book of ASTMASTM Standards, Philadelphia, PA, USA.

Bantlee JA (1995). FETAX- a developmental toxicity assay using frog embryos,, pp. 207-230. In: Rand, GM [ed.], Fundamentals of aquatic toxicology.toxicology. Effect, environmental fate, and risk assessment. Taylor andd Francis, North Palm Beach, Florida, USA.

Bantlee JA, Fort DJ, Dawson DA (1989). Bridging the gap from short-term teratogenesiss assays to human health hazard assessment by understandingg common modes of teratogenic action, pp. 46-58. In: Landis,, WG, Schalie, VD [eds.], Aquatic toxicology and hazard assessment.assessment. ASTM STP, Philadelphia, USA.

Bantlee J A, Dumont JN, Finch RA, Linder G (1990). 'Atlas of Abnormalities':: A guide for the performance of FETAX. Stilwater, OK,, USA.

Bausteinn AR, Wake DB (1995). The puzzle of declining amphibian populations.. Sci Am 272: 56-61.

Burkhartt JG, Ankley G, Bell H, Carpenter H, Fort D, Gardiner D, Gardner H,, Hale R, Helgen JC, Jepson P, Johnson D, Lannoo M, Lee D, Lary J,, Levey R, Magner J, Meteyer C, Shelby MD, Lucier G (2000).

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Strategiess for assessing the implications of malformed frogs for environmentall health. Environ Health Perspect 108: 83-90.

Chesterr G, Simsman GV, Levy J, Alhajjar BJ, Fathulla RN, Harkin JM (1989).. Environmental fate of alachlor and metolachlor. Rev Environ ContamContam Toxicol 110: 1-74.

Courchesnee CL, Bantle JA (1985). Analysis of the activity of DNA, RNA, andd protein synthesis Inhibitors on Xenopus embryo development. TeratogenTeratogen Carcinogen Mutagen 5: 177-193.


Dumontt JN, Schultz TW, Buchanan MV, Kao GL (1983). Frog embryo teratogenesiss assay: (Xenopus). A short-term assay applicable to complexx environmental mixtures, pp. 393-405. In: Waters, Sandhu, Lewtas,, Claxton, Chernoff, Nesnow [eds.], Short-term Bioassays in thethe Analysis of Environmental Mixtures III. Plenum, New York.


Gruessnerr B, Watzin MC (1995). Patterns of herbicide contamination in selectedd Vermont streams detected by enzyme immunoassay and gas chromatography/masss spectrometry. Environ Sci Technol 29: 2806-2813. .

Hobergg G (1990). Risk, science and politics - alachlor regulation in Canada andd the United-States. CanJPolit Sci 23: 257-277.

Johnsonn EM (1981). Screening for teratogenic hazards: Nature of problems. AnnAnn Rev Pharmacol Toxicol 21:417-429.

Kimmell EC, Cassida JE, Ruzo LO (1986). Formamidine insecticides and chloroacetanilidee herbicides: disubstituted anilines and nitrosobenzeness as mammalian metabolites and bacterial mutagens. / AgricAgric Food Chem 40: 1695-1699.


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Liuu D, Maguire RJ, Pacepavicius GJ, Aoyama I, Okamura H (1995). Microbiall transformation of metolachlor. Environ Toxicol Water QualityQuality 10: 249-258.

Mullerr MD, Buser H-R (1995). Environmental behavior of acetamide pesticidee stereo- and enantioselective degradation in sewage sludge andd soil. Environ Sci Technol 29: 2031-2203.

Nesnoww S, Agarwal SC, Padgett WT, Lambert GR, Boone P, Richard AM (1995).. Synthesis and characterization of adducts of alachlor and 2-chloro-A^-(2,6-diethylphenyl)acetamidee with 2'-deoxyguanosine, thymidine,, and their 3'-monophosphates. Chem Res Toxicol 8: 209-217. .


Osanoo O, Admiraal W, Klamer HJC, Pastor D, Bleeker EAJ (2002). Comparativee toxic and genotoxic effects of chloroacetanilides, formamidiness and their degradation products on Vibrio fischeri and ChironomusChironomus riparius. Environ Pollut in press.


Potterr TL, Carpenter TL (1995). Occurence of alachlor environmental degradationn products in groundwater. Environ Sci Technol 29: 1557-1563. .

Sabourinn TD, Faulk RT (1987). Comparative evaluation of a short-term test forr developmental effects using frog embryos, pp. 203-223. In Bradburyy [ed.], Report 26: Developmental toxicology: Mechanism andand risk. Cold Spring Harbor Laboratory, NY, USA.

Schottierr SP, Eisenreich SJ (1994). Herbicides in the Great-Lakes. Environ SciSci Technol 28: 2228-2232.

Surralless J, Xamena N, Creus A, Marcos R (1995). The suitability of the micronucleuss assay in human lymphocytes as a new biomarker of excisionn repair. Mutat Res 342: 43-59.

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Tessierr DM, Clark JM (1995). Quantitative assessment of the mutagenic potentiall of environmental degradative products of alachlor. J Agric FoodChemFoodChem 43: 2504-2512.

Tiedjee JM, Hagedorn ML (1975). Degradation of alachlor by a soil fungus ChaetomiumChaetomium globosum. J Agric Food Chem 23: 77-81.

USEPAA (2000). Drinking water regulations and health advisories, pp. 1-18. Lewiss Publishers: Chelsea, Washington D.C., MI, USA.

Wakee DB (1991). Declining amphibian populations. Science 253: 860-860.

Wardlaww AC (1985). How to deal with proportion data, pp. 107-110. PracticalPractical statistics for experimental biologist. Wiley Interscience, UK. .

Weii LY, Vossbrinck CR (1992). Degradation of alachlor in chironomid larvaee (Diptera, Chironomidae). J Agric Food Chem 40: 1695-1699.

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Teratogenicc Effects of Amitraz, 2,4-Dimethylaniline and Paraquatt on Developing Frog (Xenopus) Embryos

Abstract t

Developmentall effects of amitraz (acaricide), its metabolite (2,4-dimethylaniline),, and paraquat (herbicide) on embryos of a non-target organism,, Xenopus laevis, were investigated. Following the standard protocoll of the American Society for Testing and Materials (ASTM), the experimentss were carried out using native Xenopus frogs. There was a drasticc increase in mortality from 24 h to 96 h for paraquat, while 2,4-dimethylanilinee showed no mortality at the highest concentration tested (100 mg/L).. The 96-h LC50 values were 0.67, 3.27 and » 1 00 mg/L for paraquat, amitrazz and 2,4-dimethylaniline, respectively. At concentrations higher than 0.22 mg/L of paraquat all the embryos were malformed, while growth reductionn was apparent at all test concentrations (0.1 - 5 mg/L). The most commonn teratogenic effects were flexures of the notochord and stunting of growth.. Edema was the most common effect of amitraz on the embryos and 100%% of the surviving embryos in 5 mg/L were edematous. The 96-h EC50 (malformation)) values were 1.21 (95% CI 0.48 - 3.03) and 0.18 (95% CI 0.166 - 0.20) mg/L for amitraz and paraquat, respectively. The ratio of 96-h LC50LC50 to 96-h EC50 (malformation), i.e the teratogenicity index (TI), were 2.77 and 3.72 for amitraz and paraquat, respectively, while for 2,4-dimethylanilinee (TI>5) all the embryos in 25 mg/L showed observable pigmentt loss and encephalomegaly. This shows that paraquat and the degradationn product of amitraz, 2,4-dimethylaniline, should be classified as teratogens.. Teratogenic risks of massive application of these pesticides on Kenyann farms should therefore be considered.

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TeratogenicityTeratogenicity of amitraz, 2,4-dimethylaniline and paraquat

Introductio n n

Thee formamidine amitraz (Ar-2,4(dimethylphenyl)-JV-[((2,4-dimethylphenyl)-imino)methyl]-Ar-methanimidamine)) is currently the acaricidee of choice in Kenya. This compound and its biologically active metabolitee BTS27271 are a2-adrenoceptor agonists causing contractions of mammaliann cardiac and uterine muscles (Hsu and Kakuk 1984; Shin and Hsuu 1994). It is metabolised into 2,4-dimethylaniline, and other degradation products,, in arthropods and mammals (Schuntner and Thompson 1978; Knowless and Benezet 1981; Knowles and Gayen 1983; Kimmel et al 1986; Knowless and Hamed 1989). Paraquat (l,r-dimethyl-4-4'-bipyridium dichloride,, commercially available as Gramoxone®) is one of the most heavilyy used herbicides in commercial maize, sugar, and coffee farming in thee Lake Victoria basin. Like the formamidines and the chloroacetanilides, paraquatt is a chlorinated hydrocarbon. It is mainly photochemically degradedd on plants by the ultraviolet light from the sun into 4-carboxy-l-methylpyridiniumm chloride and methylamine hydrochloride (Slade 1965, 1966). .

Earlyy developmental stages of amphibians have been used to monitorr environmental contamination due to their sensitivity to a wide varietyy of toxic agents (Fort et al 1999b; Fort et al 1999a; Prati et al 2000; Tietgee et al 2000). Early stages are the most sensitive of all life stages to xenobioticss and the frog embryo teratogenesis assay-Xenopus (FETAX) providess information on mortality, malformation, and growth inhibition (ASTMM 1994). Susceptibility of various developmental stages of amphibianss to pesticides has been reported (Dial and Bauer 1984). Paraquat targetss the muscle cell cytoskeleton of amphibians leading to marked loss of actinn bundles, thus affecting the spatial organization of the cytoskeleton actinn structures (Vismara et al 2000). This explains the teratogenic effects likee growth stunting and flexures observed after exposures of frog embryos too paraquat in previous studies (Dial and Bauer 1984; Dial and Dial 1987; Vismaraa et al 2000). The effects of amitraz and its environmentally stable anilinee metabolite 2,4-dimethylaniline on amphibians are unknown.

Thee present study compares survival, growth, developmental and grosss teratogenic effects of amitraz, its metabolite (2,4-dimethylaniline) and paraquatt on Xenopus embryos. Xenopus laevis, a claw-footed and tongue-lesss frog, is native to the Lake Victoria basin. Ponds that drain the

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pesticides-ladenn farmlands in the Lake Victoria basin form the breeding groundss for the clawed frogs. Validation studies using compounds with knownn mammalian or human developmental toxicity, suggest that the predictivee accuracy of the FETAX approximates 85%, so it can be used to screenn potential human developmental health toxicants (Courchesne and Bantlee 1985; Dawson and Bantle 1987; Sabourin and Faulk 1987; Bantle et al.al. 1989). Xenopus laevis is relatively easy to breed under laboratory conditions.. Moreover, a standard protocol for conducting teratogenicity tests withh the species is available (ASTM 1994). The species is therefore a relevantt organism in teratogenicity and toxicity tests with pesticides applied onn the farmlands in the Lake Victoria basin.


Femalee and male Xenopus laevis used for this study were collected fromm the ponds at the shores of Lake Victoria, Kisumu district, Kenya. The adultt frogs were kept in laboratory glass aquaria in dechlorinated tap water. Dechlorinationn was achieved by addition of 1 ml saturated sodium thiosulfatee solution to 75 litres of tap water.

Amitrazz (analytical grade 97.5%), 2,4-dimethylaniline (98%), paraquatt (analytical grade 99%) and analytical grade solvent dimethylsulfoxidee (DMSO) were obtained from Fluka Riedel-de Haen®. FETAXX solution that comprised of 625 mg NaCl, 96 mg NaHC03, 30 mg KC1,, 15 mg CaCl2, 60 mg CaS04.2H20 and 5 mg MgS04 per litre of distilledd water was prepared and used as a medium in the assays (ASTM 1994).. All the salts were analytical grade quality.

Threee pairs (one per cage) of mating X. laevis were placed in 0.3 x

0.33 x 0,3 m cages half-filled with FETAX solution. The bottoms of the cagess were separated about 3 cm from the frogs by use of a perspex floor withh 1 cm-diameter holes. The holes allowed the eggs to sink below, out of reachh of the adult frogs, who would otherwise quickly consume them. The mediumm was aerated and maintained at 23°C. The frogs were induced to matee by injection of 500 (male) and 1000 (female) i.u. human chorionic gonadotropinn hormone (Pregnil®) into the dorsal lymph sac at 6.00 PM on thee eve of the start of the tests using a tuberculin syringe fitted with a 26-gaugee needle. Fourteen hours later the eggs from each pair were collected andd de-jellied by use of 2% w/v L-cystein prepared in FETAX solution. IN

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TeratogenicityTeratogenicity of amitraz, 2,4-dimethylaniHne and paraquat

NaOHH was used to adjust the pH of the cystein solution to 8.1. The eggs weree swirled in the L-cystein for 1 - 3 minutes. They were then washed in copiouss amounts of FETAX solution. Embryos of stage 8 (midblastula) to stagee 11 (early gastrula) were selected for the assay using the 'Normal Tables'' (Nieuwkoop and Faber 1975).

AA 96-h static-renewal whole embryo assay was conducted using the standardd protocol for the FETAX (ASTM 1994) as a guideline. The experimentall runs were conducted in a random block design with embryos fromm each pair forming a block (n = 3). Each block comprised duplicate testss and in each test 25 embryos were exposed to 10 ml of nominal concentrationss (0.1, 0.2, 0.5, 1.0, 2.0, 5.0, 10.0), (1.56, 3.25, 6.25,12.5, 25, 50,, 100), and (0, 0.05, 0.1, 0.2, 0.5, 1, 2, 5) mg/L of amitraz, 2,4-dimethylaniline,, and paraquat, respectively, in 60 mm diameter petri dishes. DMSOO was used to facilitate dissolution of the chemicals and the final concentrationn of the DMSO was 1% (v/v) in all the test concentrations. The controll group comprised 2 dishes of 25 embryos in 10 ml FETAX solution andd 2 dishes with FETAX solution plus DMSO (1% v/v). Dishes were placedd in random positions in an incubator set at 24°C (range of 23 - 25°C). Att 24 h, 48 h, and 72 h the test solutions were renewed and the dead embryoss were discarded. After 96 h the experiment was terminated and the livee larvae were counted and fixed in 3% formalin. The 96-h LC50 and 96-h EC500 (malformation) were determined by probit analysis (Wardlaw 1985) afterr adjustment for mortalities and malformations in the control as follows: YY = 100*(C - T)/C, where Y = % response; C = % not responding in the DMSOO control; T = % not responding in the test dilution and the response wass considered as percent growth retardation, mortality or malformation accordinglyy (Abbott 1925). The teratogenic index (TI) was deduced by dividingg the 96-h LC50 by 96-h EC50 and was used to quantify the degree off teratogenicity of compounds (Dumont et al. 1983).

Thee effects of the pesticides and the degradation product of amitraz, 2,4-dimethylaniline,, on survival, growth, and malformation were observed withh a binocular-dissecting microscope. Head (mouth) to tail lengths (end of thee tail) were measured at the end of each run by use of an ocular micrometerr at a magnification of x20. Separately, the tail length (end of the taill to the region of the hind limb bud) and body length (the difference betweenn the tail length and the whole tadpole body) were recorded for each larva.. For the flexed larvae, measurements were taken along the curvature of

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thee notochord. The tadpoles were examined for developmental abnormalitiess under the binocular-dissecting microscope. The "Atlas of Abnormalities"" was consulted as a guide to distinguish between normal and abnormall embryos and to determine the different kinds of abnormalities (Bantlee et al. 1990). An embryo was considered abnormal if it exhibited at leastt one type of malformation. In this investigation an abnormality was scoredd regardless of its seriousness. For example, a slight facial edema, largee gut edema, or extensive edemas were all scored as edema. Tadpoles weree preserved in rubber-sealed vials with 3% formalin.

Statisticall analyses were done with the Student's '?'-test and ANOVAA (with Bonferroni adjustment matrix of pairwise comparison probabilities,, a = 0.05 level) in the SYSTAT® (1996) statistical program to determinee any significant difference in growth of the tadpoles.

Results s

EmbryolethalEmbryolethal effects

Thee survival in the DMSO control was 93.3% after 96 h. The 96-h LC500 were 3.27 (95% CI 2.56 - 4.16) and 0.67 (95% CI 0.57 - 0.81) mg/L forr amitraz and paraquat, respectively. There was no mortality for the highestt tested concentration (100 mg/L) of 2,4-dimethylaniline. The embryoss in media that contained > 0.5 mg/L paraquat appeared moribund at 966 h, as they were unable to swim, but could be seen to be alive from the pulsationn of their hearts. For paraquat drastic increased lethal effects from 244 h to 96 h of incubation were observed (Fig 1, Table 1), but not for amitrazz or 2,4-dimethylaniline in this study, nor for the chloroacetanilides; alachlor,, metolachlor and their aniline degradation products in our previous studyy (Osano et al. 2002a). Table 1 gives the LC50 values of the compounds afterr different periods during the test.

GrowthGrowth effects

Thee mean size of the tadpoles in the DMSO control after 96h incubationn was 7.51 0.048 mm (SE) (n = 144). The minimum concentrationn to inhibit growth (MCIG) was 0.1 mg/L (p<0.05) for both amitrazz and paraquat. 2,4-Dimethylaniline caused subtoxic stimulation of

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TeratogenicityTeratogenicity ofamitraz, 2,4-dimethylaniline and paraquat

Tab lee 1: 24-h, 48-h, 72-h, and 96-h LC50(s); 96-h EC50 (mal format ion); and T\(s)

off the test chemica ls for Xenopus laevis embryos.

Chemicall Ami t raz 2,4-Dimethylanilinee Paraquat

24-hh LCs. 5<LC5o<10 mg/L >100mg/L 00 mg/L

48-hh LC50 5<LC5(i<10 mg/L >100mg/L 14.555 mg/L (95%% CI 9.80-21.60)

72-hh LC50 5<LC5I.<10 mg/L >100mg/L 3.300 mg/L (95%% CI 2.30-4.55)

96-hh LC50 3.27 mg/L >100mg/L (95%CII 2.56-4.16)

96-hh EC50 1.21 mg/L *20 mg/L (95%% CI 0.48-3.03)

0.677 mg/L (95%% CI 0.57-0.81)

0.188 mg/L (95%% CI 0.16-0.20)

I I I 2.70 0 (95%% CI 0.84-8.67)

>5 5 3.72 2 (95%% CI 2.26-6.40)

Estimated d

1000 r

-3 3 > > 'E E = =

Contro l l 0.11 1

Paraquatt concentration (mg/L)

10 0

Figg 1. The % survival of the Xenopus laevis tadpoles against concentrations of paraquat after incubationn for different periods.

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growthh between 3.13 - 100 mg/L (p<0.001) concentrations (Fig 2). Amitraz causedd a small but significant growth retardation of up to 10% in the maximumm tolerated concentration (5.0 mg/L) (Fig 2). Paraquat caused substantiall growth retardation with the tadpoles mean length of 5.27 (0.0677 SE, n = 74) mm in 1 mg/L medium. This represents a 30% reduction off the growth relative to the control. There was no significant difference (p == 0.31) between the lengths at this concentration and the lengths at 2 mg/L (5.166 4 SE) mm. Relative to the full lengths of the tadpoles the tail lengthss were proportionally shorter in the 4-day-old tadpoles incubated in thee higher concentration of paraquat than those in the control (p = 0.027). In 1.00 mg/L the tail was 68.9% 1 SE, n = 19) of the total lengths of the tadpolee compared to 70.8% 9 SE, n = 20) in the control. Paraquat inducedd a general reduction of the lengths of both the deformed and the non-deformedd tadpoles in this study. However, the tadpoles that showed other deformitiess (other than stunting) exhibited more severe growth retardation thann those that were not deformed (Fig 3).

TeratogenicTeratogenic effects

Thee 96-h EC50 malformation in the embryos of X. laevis amounted too 1.21 (95% CI 0.48 - 3.03) and 0.18 (95% CI 0.16 - 0.20) mg/L for amitraz andd paraquat, respectively (Table 1). The TI were 2.70 (95% CI 0.84 - 8.67), 3.722 (95% CI 2.26 - 6.40) and >5 for amitraz, paraquat, and 2,4-dimethylaniline,, respectively (Table 1).

Amitrazz caused edema and axial flexures as the main types of abnormalitiess (Fig 4). At 5 mg/L all the surviving embryos were edematous. Axiall flexures were identified as curvature of the notochord or bending of thee tail (Plate B). Edemas comprised edema of the face, heart and/or abdomenn (Plate C). The significant effect of 2,4-dimethylaniline was a progressivee loss of pigment together with encephalomegaly that was observablee from 25-mg/L test media. At 100 mg/L there was total loss of the pigmentt leading to loss of colour contrast between the eyes and the rest of thee body, and the bifurcation at the forebrain was indistinguishable as a resultt of swelling of the brain (Plate D). For paraquat medial flexures of the notochordd and stunting were the main types of abnormalities observed on thee embryos (Fig 4). The frequency of deformities increased with increase in concentrationn of paraquat. At 0.5 mg/L, 98% of the embryos were deformed,, out of which 45% were stunted and 50% had their notochords

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flexedd compared to 11, 0, and 7%, respectively, for the same types of deformitiess in the control. Plates A & E show a normal (control) tadpole, andd a flexed and stunted specimen, respectively. Table 1 summarizes the LC50s,, EC50s and TIs in the present study.

Discussion n

Paraquatt was highly toxic to the embryos of X. laevis (LC50 = 0.67 (95%% CI 0.57 - 0.81) mg/L; MCIG = 0.1 mg/L (p = 0.05)). It is persistent in thee soil environment with a reported half-life of >1000 days (Wauchope et al.al. 1992; Weed Science Society of America 1994). It is however degraded byy UV light to 4-carboxy-l-methylpyridinium chloride and methylamine hydrochloridee as the major decomposition products (Slade 1965, 1966). Amitrazz is quickly (half-life < 1 day) degraded to 2,4-dimethylaniline and otherr metabolites (Kimmel et al. 1986; Kidd and James 1991). It was found too be relatively less toxic than paraquat to the embryos of X. laevis (3.27 [95%% CI 2.56 - 4.16]). Moreover, its major use as an acaricide on livestock restrictss its contamination to focal points such as around the spray races, cattlee plunge dips, and veterinary clinics. However, careless disposal of the usedd mixtures of the acaricide could expose the environment to the risks of thee degradation products. The environmentally stable aniline degradation productt of amitraz (2,4-dimethylaniline) proved non-lethal to the embryos off X. laevis at our tested concentrations of 0 - 100 mg/L.

Thee impairment of growth of Xenopus in the present study was proportionall to the paraquat concentration. This corroborates earlier findings off growth retardation in tadpoles of Rana pipiens and fingerlings of OreochromisOreochromis niloticus (Dial and Bauer 1984; Babatunde 1997). Also the tail abnormalities,, flexure and shortness, have been observed in Rana (Dial and Diall 1995), but the present study demonstrates that this aberration is induced byy very low paraquat concentrations. X. laevis embryos are potentially more sensitivee test organisms than the previously used vertebrates, although in the presentt study, the tests were initiated with eggs and not older larvae like in thee tests using fish (Babatunde 1997) and other frog species (Dial and Dial 1995).. Shortening of the embryos were more apparent in the assay with paraquatt than those with amitraz, in spite of the latter compound being an a2-adrenoceptorr agonist that causes contraction of muscles in pigs (uterus)

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andd rats, (Hsu and Kakuk 1984; Shin and Hsu 1994). Paraquat causes loss of actinn bundles thus affecting the spatial organization of the muscle cytoskeletonn actin structures (Vismara et al 2000). Apparently this leads to aa marked reduction in the lengths of the tadpoles and more especially the taill that comprises mostly of the notochord and tail muscle (Dial and Bauer 1984;; Dial and Dial 1987; Vismara et al. 2000). In this study the prominent effectt of amitraz of generalised edema may have been due to a disruption of osmoregulationn resulting from cell membrane lipid bilayer disruption. Similarly,, related aniline based compounds alachlor and metolachlor exert theirr toxicity to the chironomid larvae through narcosis (Osano et al 2002b). .

Thee exposure of the embryos in the present study was through the waterr medium rather than through the food as by Dial and Dial (1987). The apparentt drastic increase in toxicity of paraquat to the tadpoles from 24 h to 966 h post hatch suggests that the toxicity in the early stages of embryogenesiss may be decreased through physical protection of any remainingg jelly or a different physiological process. A similar observation wass made in amphibians and crayfish (Leung et al. 1980; Dial and Bauer 1984).. Paraquat and amitraz have short half-lives in aquatic environment, thee former being adsorbed and concentrated by sediment, suspended solids orr aquatic plants while the latter breaks down to BTS27271 and subsequentlyy to 2,4-dimethylaniline and other metabolites (Way et al 1971; Calderbankk 1972; Kosinki and Merkle 1984; Bernal et al 1997). Paraquat is nott easily degraded chemically or microbiologically and demonstrates a longg half-life (>56 d) in river water medium (Wang et al 1994). In the wild thee larvae that survive paraquat contamination at the early stages would be additionallyy challenged by oral route of exposure at the later stages when theyy feed on contaminated algae. A significant number of tadpoles fed three dayss on paraquat treated Myriophylum were observed to have abnormal tails (Diall and Dial 1995). In the present experiment the fertilised Xenopus eggs weree kept in dilute media without organic additions and nourishment, other thann the embryonic yolk sac, was absent. Such an effect could be enhanced byy the degradation of paraquat in association with plants (Slade 1965, 1966).. The resulting degradation products (4-carboxy-l-methylpyridinium chloridee and methylamine hydrochloride) have not been tested for teratogenicc effects.

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TeratogenicityTeratogenicity of amitraz, 2,4-dimethylaniline and paraquat

Contro ll 0.1 1 10 100

Compoun dd concentration s (mg/L )

Figg 2. Growth expressed as % of DMSO control mean length of the Xenopus laevis embryoss after 96-h incubation in different concentrations of amitraz, 2,4-dimethylanilinee and paraquat. The error bars represent standard errors of mean. .

Nonn deformed

Paraquatt concentration (mg/L)

Figg 3. Growth expressed as % of DMSO control mean length of the Xenopus laevis nonn deformed tadpoles and tadpoles with deformities other than stunting of XenopusXenopus laevis after 96-h incubation in paraquat. The error bars represent standardd errors of mean.

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ChapterChapter 5

Amitra z z

Paraquat t

© -- eye abnormality $4$4—— stunting H—— one or more flexures Q—— gut abnormality

—— edema ll abnormal

II ) )

Chemicall concentration (mg/L)

Figg 4. The % occurrence of normal and abnormal tadpoles of Xenopus laevis afterr 96-h incubation in different concentration of amitraz (top) and paraquatt (bottom).

Effectss of amitraz on developing amphibians have not been reported inn literature. Our TI of 2.7 indicates teratogenicity of the compound, which iss increased after its degradation into 2,4-dimethylaniline (TI>5). All the embryoss grown in the media of 2 5 - 1 00 mg/L 2,4-dimethylaniline were deformedd (showed loss of pigment) but survived the test concentrations (100%).. The encephalomegaly observed in this study confers neural

81 1


Plates:: Four-day-old tadpoles of Xenopus laevis, A: in the FETAX DMSO control solution, B: in 0.55 mg/L amitraz. The tadpole is flexed, C: in 0.5 mg/L amitraz. The tadpole is edematous; D:: in 100 mg/L solution of 2,4-dimethylaniline. The tadpole has enlarged brain and is depigmentedd making it difficult to contrast between the eyes and the rest of the body. Noticee the lack of pigmentation on the left eye, E: in 1 mg/L solution of paraquat. The tadpolee is stunted and flexed.

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deformity.. Similarly, a high teratogenicity upon degradation was observed forr other aniline based pesticides, alachlor and metolachlor, in our previous studyy (Osano et al. 2002a). Teratogenic effects of the degradation products off paraquat need study.

X.X. laevis used in our study are native species of Lake Victoria basin andd inhabit the ponds that drain agricultural farmlands where paraquat is appliedd while amitraz is the current acaricide of choice in the same region. Thee wastes of the used acaricide are often drained into the environment withoutt regard of their potential toxicity. We suggest that the sensitivity of XenopusXenopus embryonal development to the environmentally stable 2,4-dimethylanilinee and paraquat in conjunction with the observed effects of paraquatt sorbed in plants is likely to cause detrimental effects on natural populationss of this species in the Lake Victoria basin. Further studies on levelss of the pesticides and their degradation products in water, sediment andd biota of Lake Victoria basin water bodies are required as well as observationss on teratogenic effects in nature.

References s

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ASTMM (1994). Standard guide for conducting the frog embryo teratogenesiss assay-Xenopus (FETAX) E 729-80, pp. 272-296, AnnualAnnual Book of ASTM Standards, PA, USA.

Babatundee MM (1997). Toxicity of paraquat (Gramoxone®) to Oreochromis niloticus.niloticus. Ahmadu Bello University, Zaria, Nigeria.

Bantlee JA, Fort DJ, Dawson DA (1989). Bridging the gap from short-term teratogenesiss assays to human health hazard assessment by understandingg common modes of teratogenic action, pp. 46-58. In: Landis,, WG, Schalie, VD [eds.], Aquatic toxicology and hazard assessment.assessment. ASTM STP, Philadelphia, USA.

Bantlee J A, Dumont JN, Finch RA, Linder G (1990). Atlas of Abnormalities':: A guide for the performance of FETAX, Stilwater. OKK USA.

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Bernall JL, DelNozal MJ, Jimenez JJ (1997). Influence of solvent and storagee conditions on the stability of acaricide standard stock solutions.. J Chromatogr A 765: 109-114.

Calderbankk A (1972). Environmental consideration in the development of diquatt and paraquat as aquatic herbicides. Outlook Agr 7: 51-54.

Courchesnee CL, Bantle J A (1985). Analysis of the activity of DNA, RNA, andd protein synthesis inhibitors on Xenopus embryo development. TeratogenTeratogen Carcinogen Mutagen 5: 177-193.


DialDial CAB, Dial NA (1995). Lethal effects of the consumption of field levels off paraquat- contaminated plants on frog tadpoles. Bull Environ ContamContam Toxicol 55: 870-877.

DialDial NA, Bauer CA (1984). Teratogenic and lethal effects of paraquat on developingg frog embryos {Rana pipiens). Bull Environ Contam ToxicolToxicol 33: 592-597.

Diall NA, Dial CAB (1987). Lethal effects of diquat and paraquat on developingg frog embryos and 15-day-old tadpoles, Rana pipiens. BullBull Environ Contam Toxicol 38: 1006-1011.

Dumontt JN, Schultz TW, Buchanan MV, Kao GL (1983). Frog embryo teratogenesiss assay: {Xenopus). A short-term assay applicable to complexx environmental mixtures, pp. 393-405. In: Waters, Sandhu, Lewtas,, Claxton, Chernoff, Nesnow [eds.], Short-term Bioassays in thethe Analysis of Environmental Mixtures III. Plenum, NY, USA.

Fortt DJ, Rogers RL, Copley HF, Bruning LA, Stover EL, Heigen JC, Burkhartt JG (1999a). Progress toward identifying causes of maldevelopmentt induced in Xenopus by pond water and sediment extractss from Minnesota, USA. Environ Toxicol Chem 18: 2316-2324. .

FortFort DJ, Propst TL, Stover EL, Helgen JC, Levey RB, Gallagher K, Burkhartt JG (1999b). Effects of pond water, sediment, and sediment extractss from Minnesota and Vermont, USA, on early development

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andd metamorphosis of Xenopus. Environ Toxicol Chem 18: 2305-2315. .

Hsuu WH, Kakuk TJ (1984). Effect of amitraz and chlordimeform on heart ratee and pupil diameter in rats: mediated by a2-adrenoreceptors. ToxicolToxicol Appl Pharm 73:411-415.

Kiddd H, James DR (1991). The Agrochemical Handbook, pp. 10-12, Royal SocietySociety of Chemistry Information Services, 3rd ed, Cambridge, UK.

Kimmell EC, Cassida JE, Ruzo LO (1986). Formamidine insecticides and chloroacetanilidee herbicides: disubstituted anilines and nitrosobenzeness as mammalian metabolites and bacterial mutagens. JJ Agric Food Chem 40: 1695-1699.

Knowless CO, Benezet HJ (1981). Excretion balance, metabolic fate and tissuee residues following treatment of rats with amitraz and A"-(2,4-dimethylphenyl)-jV-methylformamidine.. J Environ Sci Health B 16: 547-555. .

Knowless CO, Gayen AK (1983). Penetration, metabolism and elimination of amitrazz and Ar-(2,4-dimethylphenyl)-A^-methylformamidine in Southwesternn corn borer larvae (Lepidoptera: Pyralidae). J Econ EntomolEntomol 76: 410-413.

Knowless CO, Hamed MS (1989). Comparative fate of amitraz and iV-(2,4-dimethylpheny^-^-methylformamidinee (BTS-27271) in bollworm andd tobacco budworm larvae (Lepidoptera, Noctuidae). J Econ Entomoni:Entomoni: 1328-1334.

Kosinkii JR, Merkle MG (1984). The effect of four terrestrial herbicides on thee productivity of artificial stream algal communities. J Environ Qual\3:15-$2. Qual\3:15-$2.

Leungg T-S, Naqvi SM, Naqvi NZ (1980). Paraquat toxicity to Lousiana Crayfishh {Procambrarus clarkii). Bull Environ Contam Toxicol 25: 465-469. .


Osanoo O, Admiraal W, Otieno D (2002a). Developmental disorders in embryoss of the frog Xenopus laevis induced by chloroacetanilide

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herbicidess and their degradation products. Environ Toxicol Chem 21:375-379. .

Osanoo 0, Admiraal W, Klamer HJC, Pastor D, Bleeker EAJ (2002b). Comparativee toxic and genotoxic effects of chloroacetanilides, formamidiness and their degradation products on Vibrio fischeri and ChironomusChironomus riparius. Environ Pollut (in press).

Pratii M, Biganzoli E, Boracchi P, Tesauro M, Monetti C, Bernardini G (2000).. Ecotoxicological soil evaluation by FETAX. Chemosphere 41:: 1621-1628.

Sabourinn TD, Faulk RT (1987). Comparative evaluation of a short-term test forr developmental effects using frog embryos, pp. 203-223. In: Bradburyy [ed.], Report 26: Developmental toxicology: Mechanism andand risk. Cold Spring Harbor Laboratory, NY, USA.

Schuntnerr CA, Thompson PG (1978). Metabolism of [l4C]amitraz in larvae off Boophilus microplus. AustJ Biol Sci 31: 141-148.

Shinn DH, Hsu WH (1994). Influence of the formamidine pesticide amitraz andd its metabolites on porcine myometrial contractility involvement off cb-adrenoceptors and Ca2+ channels. Toxicol Appl Pharmacol 128:: 45-49.

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Sladee P (1966). The fate of paraquat applied to plants. Weed Res 6.

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Vismaraa C, Battista V, Vailati G, Bacchetta R (2000). Paraquat induced embryotoxicityy on Xenopus laevis development. Aquat Toxicol 49: 171-179. .

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Wardlaww AC (1985). How to deal with proportion data, pp. 107-110. In: PracticalPractical statistics for experimental biologist. Wiley Interscience, UK. .

Wauchopee RD, Buttler TM, Hornsby AG, Augustijn Beckers PWM, Burt JP (1992).. SCS/ARS/CES Pesticide properties database for environmentall decision making. Rev Environ Contain Toxicol 123.

Wayy JM, Newman JF, Moore NW, Knaggs FW (1971). Some ecological effectss of the use of paraquat for the control of weeds in small lakes. JApplEcolJApplEcol 8: 509-532.

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Concludingg Remarks

Pesticidess are designed to chemically control undesired organisms in agriculturall ecosystems and hence their behavior in all steps of intended use onn the fields and unintended distribution (e.g. in water) has to be considered. Itt was hypothesized in this thesis that degradation products provoke adverse effectss on aquatic biota that differ from specific effects of the parent pesticides.. The aim of this thesis was therefore to analyze the effects of environmentallyy stable pesticide degradation products on fundamental processess in a living organism, like growth and development. Here, I review thee findings on a limited series of compounds, the chloroacetanilides, the formamidiness and their degradation products, and a limited series of biologicall tests in order to analyze the chain of effects from pesticide applicationn to adverse environmental effects. A risk to the environment and too man has been indicated and this has prompted a remark on the regulation off pesticides.

Pesticidee degradation products

Thee pesticide's environmental persistence (half-life), transformation pathwayss and the transformation products vary with environmental conditionss (Webster et al. 1998; Graham et al. 2000). The massively applied chloroacetanilidess and formamidines break down to compounds that are persistentt in the environment, indicating a probable long-term hazard (Konopkaa 1994; Stamper and Tuovinen 1998; Scribner et al. 2000). The degradationn products are generally more water-soluble and are easily translocatedd into other environmental compartments, thus extending the areass that are potentially at risk of the pesticides. This latter problem could bee mitigated by improved soil adsorption properties and agricultural managementt practices. For instance, reduced soil erosion and increased organicc matter content would reduce dissipation of the pesticides and their degradationn products (Farenhorst and Bowman 2000). In the temperate region,, due to the longer lapse of time after application of the pesticides, surfacee waters receive pesticide degradation products, which get distributed ass far as in the river mouth (Steen et al. 1999). The tropical environment conferss ecotoxicological risks to pesticides that differ from that in the temperatee regions. The role of the degradation products is more prominent inn the tropics than in the temperate regions (Lalah and Wandiga 1996; Karlssonn et al. 2000). The relatively higher ambient diurnal temperatures

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ConcludingConcluding remarks

andd UV irradiation in these regions favor microbial degradation, hydrolysis andd photolysis of the pesticides, thereby decreasing their persistence, while increasingg the concentrations of their degradation products in the environment.. It is concluded that although degradation products of pesticidess receive a growing scientific attention, their role in ecotoxicology iss still inadequate.

Specificityy of Pesticide degradation products

Toxicityy of pesticides is governed by their bioavailability and reactivityy at the target sites (Hansch and Fujita 1963). Both the bioavailabilityy and the interaction with the target change upon the degradationn of the pesticides. Bioavailability of most organic compounds to thee cells of living organism is dependent on their lipophilicity, expressed usuallyy as the Log KoW and their baseline toxicity (narcosis) can be predicted byy this factor (Roghair et al. 1994; Cronin and Dearden 1995; Vaes et al. 1998).. The toxicity of aniline based pesticides and their degradation productss to Chironomus riparius larvae in the present study (Chapter 3) also mett this relationship.

Duringg the manufacture of pesticides, carriers are incorporated to enhancee the pesticides availability or to slow down their release. In the environmentt bioavailability of pesticides and other toxicants is influenced byy several factors, including their water solubility; for example in the presentt study the lower water solubility of amitraz prevented its toxicity to thee aquatic C. riparius (Chapter 3). Bioavailability of pesticides is also influencedd by the presence and nature (aromaticity) of the organic matter (Farenhorstt and Bowman 2000; Ahmad et al. 2001), UV irradiation-aided weakeningg of adsorption bonds (Bossan et al. 1995; Wernersson et al. 2000) andd temperature-governed increased plant uptake (Kiflom et al. 1999), whichh could enhance exposure of grazers through ingestion, for example in thee case of paraquat toxicity to frog larvae (Dial and Dial 1995). All in all, transformationn of the pesticides lowers their Log K w and hence their lipophilicity.. Can this change be construed with diminished hazard or does thee transformation alter the specificity of the compounds?

Specificityy of pesticide's reactivity on the target organisms is exploitedd in their design and manufacture. The formamidines mimic monoaminee oxidase in invertebrates, thereby inhibiting the enzyme (Aziz andd Knowles 1973), while the chloroacetanilides inhibit photosynthesis in

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ChapterChapter 6

selectedd species of plants. Upon degradation this selective toxicity to target speciess is lost and their lethal toxicity to non-target species is also generally reduced.. By applying a battery of 4 test systems, this thesis has shown that transformationn of the pesticides give them new toxic properties. The more polarr aromatic amine degradation products elicit toxicity by a mechanism thatt is different than narcosis in the V.fischeri (Hermens et al. 1985; Vaes et al.al. 1998). In the present study the degradation products were more toxic thann their corresponding parent compounds to the bacteria (procaryote), evenn though their Log K<,w values were lower. The opposite was true in the animall species (eucaryote) (Chapter 3). This thesis shows that while the simplee QSAR approaches could be a useful tool in prediction of toxicity of untestedd compounds that act by narcosis, it can be seriously in error in predictingg toxicity that is governed by mechanisms different than narcosis (Chapterr 3). In the frog developmental studies the predominant type of aberrationn caused by the parent compounds were edema, which is typical of celll membranes and hence osmoregulation disruptions, while the degradationn products exhibited diverse types of aberrations (Fig 2, Chapter 4;; Chapter 5). The teratogenicity expressed as the ratio between LC50 to EC500 of the pesticides was enhanced upon degradation; for example the concentrationss of 2,4-dimethylaniline that induced 100% malformation provedd non lethal to embryos (Chapter 5). In view of their diversified specificc toxic effects established in the present study, the seemingly innocuouss degradation products could underlie long-term hidden risks of the pesticidess in the environment.

Developmentall disorders

Couldd the role of the pesticide degradation products in the etiology off developmental malformation help to understand the epidemiology of increasedd frog malformation and the dwindling of amphibian species (Wake 1991;; Blaustein and Wake 1995; Burkhart et al. 2000)? The long list of environmentall factors identified as possible cause of frog abnormalities has underratedd the pesticide degradation products (Chapter 3-5). This list includess antithyroid compounds, mineral (Ca and Mg) depletion, surfactants,, nitrate fertilizers, parasitic trematodes, petroleum products, metalss and UV radiation, besides other numerous anthropogenic environmentall teratogens (Davis et al. 1981; De Zwart and Sloof 1987; Blausteinn and Wake 1995; Fort et al. 1999; Schuytema and Nebeker 1999; Burkhartt et al. 2000; Mann and Bidwell 2000). The widespread use of

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pesticides,, especially the chloroacetanilides, in the agricultural fields makes themm important contaminants of the amphibian habitat, which could be teratogenicc as was demonstrated in this study (Chapter 4-5).

Thee fact that the pesticides studied here, chloroacetanilide and formamidineformamidine degradation products, were found to be genotoxic (Chapter 3) andd form DNA adducts (Bonfanti et al. 1992; Nesnow et al. 1995) raises the questionn on the possible ramification of this scenario in a developing embryo.. The compounds in the present study exhibited a predilection for specificc organs or system of the embryos. 2,4-dimethylaniline, for instance, causedd depigmentation and encephalomegaly, both representing the nervous systemm effects and paraquat retarded tail growth of the larvae (Chapter 5). Theree are several human developmental disabilities, whose etiologies are nott well understood and there are also a number of chemicals and pesticides thatt have not yet been evaluated for their teratogenic potential (Goldman andd Koduru 2000). The assay with the Xenopus laevis embryos is reported too be able to predict 85% of the human teratogens (Dumont et al. 1983; Courchesnee and Bantle 1985; Sabourin et al. 1985; Dawson and Bantle 1987;; Sabourin and Faulk 1987; Bantle 1995). Therefore, the degradation productss of chloroacetanilides and formamidine pose potential health risks too man.

Environmentall Regulation

Thiss study underlines the need to consider pesticide degradation productss in regulating pesticide use. I have demonstrated the presences of diversee toxicity endpoints of pesticides and their degradation products and thereforee propose a more diverse evaluation of toxicity in the current regulatoryy process. This thesis has shown that degradation of a pesticide couldd attenuate its lethal toxicity while conferring on it risks of genotoxicity orr teratogenicity (Chapters 3-5). Such risks could be hidden and on a long-termm scale they could even bear far-reaching effects on a species leading to nearr extinction as was experienced with the effects of organochlorines on topp predators in the sixties. These risks may be underrated in the current admissionn policies that characterize the pesticides based on the toxicology off the parent compounds mostly. Therefore, the extra costs required in identifyingg and quantifying toxic effects and in accommodating degradation productss is justified and a revision of environmental monitoring and quality criteriaa is appropriate.

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ChapterChapter 6

Settingg up risk assessment for application in regulations on pesticidess is currently dogged by inadequate toxicological data (Steen et ai 1999),, and this is worse for the degradation products (Belfroid et al. 1996). Thee transformation of pesticides follows varied pathways and results in diversee products, most of which cannot be detected with sufficient reliability (Belfroidd et al. 1996). The coupled processes of sorption and degradation of thee pesticides and their degradation products are indeed varied making it difficul tt to come up with a simple fate model to be used in the risk evaluationn procedures. Certainly, not all degradation products are environmentallyy stable. Thus a first step could be to identify the environmentallyy stable degradation products and determine their adverse biologicall effects, as in the present study, in order to incorporate them in the chainn of effects assessment.

Formulationn of environmental protection legislation is still at its inceptionn stage in Kenya (GOK 1999). Environmental assessment and monitoringg mostly rely on World Health Organization or European Union standards.. These are inadequate given the unique environmental conditions inn Kenya. To give an example from the present study: the absence of herbicidee residues from water and sediment would suggest that the applicationn was safe, while the abundance of the degradation products indicatedd the contrary. To achieve the global 2020 vision of sustainable developmentt and nutrition for all, formalized by Kenya (Republic of Kenya 1997),, it is imperative that Kenya regulates pesticides according to a new advancedd standard and avoid the mistakes that marked earlier agro-industriall revolutions in the North and produced unwholesome environments. .

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References s

Ahmadd R, Kookana RS, Alston AM, Skjemstad JO (2001). The nature of soill organic matter affects sorption of pesticides. 1. Relationships with carbonn chemistry as determined by 13C CPMAS NMR spectroscopy. EnvironEnviron Sci Technol 35: 878-884.


Bantlee JA (1995). FETAX- A developmental toxicity assay using frog embryos,, pp. 207-230. In: Rand, GM [ed.], Fundamentals of Aquatic Toxicology.Toxicology. Taylor and Francis, North Palm Beach, Florida,USA.

Belfroidd AC, van Drunen M, Beek MA, Schrap SM, van Gestel CAM, van Hattumm B (1996). Relative risks of transformation products of pesticidess for aquatic ecosystems. Sci Total Environ 222: 167-183.

Blausteinn AR, Wake DB (1995). The puzzle of the declining amphibian populations.. Sci Am 272: 56-61.

Bonfantii M, Taverna P, Chiappetta L, Vill a P, D'lncalci M, Bagnati R, Fanellii R (1992). DNA damage induced by alachlor after in vitro activationn by rat hepatocytes. Toxicology 72: 207-219.


Burkhartt JG, Ankley G, Bell H, Carpenter H, Fort D, Gardiner D, Gardner H,, Hale R, Helgen JC, Jepson P, Johnson D, Lannoo M, Lee D, Lary J,, Levey R, Magner J, Meteyer C, Shelby MD, Lucier G (2000). Strategiess for assessing the implications of malformed frogs for environmentall health. Environ Health Perspect 108: 83-90.

Courchesnee CL, Bantle JA (1985). Analysis of the activity of DNA, RNA, andd protein synthesis inhibitors on Xenopus embryo development. TeratogenTeratogen Carcinogen Mutagen 5: 177-193.

Croninn MTD, Dearden JC (1995). QSAR in Toxicology. 1. Prediction of aquaticc toxicity. Quant Struct-Act Relat 14: 1-7.

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Daviss KR, Schultz TW, Dumont JN (1981). Toxic and teratogenic effects of selectedd aromatic amines on embryos of the amphibian Xenopus laevis.laevis. Arch Environ Contam Toxicol 10.

Dawsonn DA, Bantle JA (1987). Development of a reconstituted water mediumm and preliminary validation of the Frog Embryo Teratogenesis Kss&y-XenopusKss&y-Xenopus (FETAX). J Appl Toxicol 7: 237-244.

Dee Zwart D, Sloof W (1987). Toxicity of mixtures of heavy metals and petrochemicalss on Xenopus laevis. Bull Environ Contam Toxicol 38: 345-351. .

Diall CAB, Dial NA (1995). Lethal effects of the consumption of field levels off paraquat-contaminated plants on frog tadpoles. Bull Environ ContamContam Toxicol 55: 870-877.

Dumontt JN, Schultz TW, Buchanan MV, Kao GL (1983). Frog embryo teratogenesiss assay: {Xenopus). A short-term assay applicable to complexx environmental mixtures, pp. 393-405. In: Waters, Sandhu, Lewtas,, Claxton, Chernoff, Nesnow [eds.], Short-term Bioassays in thethe Analysis of Environmental Mixtures III. Plenum, NY, USA.

Farenhorstt A, Bowman BT (2000). Sorption of atrazine and metolchlor by earthwomm surface castings and soil. J Environ Sci Health B 35: 157-173. .

Fortt DJ, Rogers RL, Copley HF, Burning LA, Stover EL, Heigen JC, Burkhartt JG (1999). Progress toward identifying causes of maldevelopmentt induced in Xenopus by pond water and sediment extractss from Minnesota, USA. Environ Toxicol Chem 18: 2316-2324.

GOKK (1999). Environmental coordination and management act. Nairobi, Kenya a

Goldmann LR, Koduru S (2000). Chemicals in the environment and developmentall toxicity to children: a public health and policy perspective.. Environ Health Perspect 108 Supplement 3: 443-448.

Grahamm DW, Miley MK, DeNoyelles F, Smith VH, Thurman EM, Carter R (2000).. Alachlor transformation patterns in aquatic field mesocosms underr variable oxygen and nutrient conditions. Water Res 34: 4054-4062. .

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Hanschh C, Fujita T (1963). p-a-n Analysis. A method for the correlation of biologicall activity and chemical structure. Biochem J 87: 1616-1626.

Hermenss J, Busser F, Leeuwangh P, Musch A (1985). Quantitative structure-activityy relationships and mixture toxicity of organic chemicalss in Photobacterium phosphoreum: the microtox test. EcoEco toxicol Environ Saf 9: 17-25.

Karlssonn H, Muir DCG, Teixiera CF, Burniston DA, Strachan WMJ, Hecky RE,, Mwita J, Bootsma HA, Grift NP, Kidd KA, Rosenberg B (2000). Persistentt chlorinated pesticides in air, water, and precipitation from thee Lake Malawi area, Southern Africa. Environ Sci Technol 34: 4490-4495. .

Kiflo mm WG, Wandiga SO, Ng'ang'a PK, Kamau GN (1999). Variation of plantt p,p '-DDT uptake with age and soil type and dependence of dissipationn on temperature. Environ Int 25: 479-487.


Lalahh JO, Wandiga SO (1996). Distribution and dissipation of carbofuran in aa paddy field in the Kano plains of Kenya. Bull Environ Contam ToxicolToxicol 56: 584-593.

Mannn RM, Bidwell JR (2000). Application of the FETAX protocol to assess thee developmental toxicity of nonylphenol ethoxylate to Xenopus laevislaevis and two Australian frogs. Aquat Toxicol 51: 19-29.

Nesnoww S, Agarwal SC, Padgett WT, Lambert GR, Boone P, Richard AM (1995).. Synthesis and characterization of adducts of alachlor and 2-chloro-7^-(2,6-diethylphenyl)acetamidee with 2'-deoxyguanosine, thymidine,, and their 3'-monophosphates. Chem Res Toxicol 8: 209-217. .

Republicc of Kenya (1997). National Development Plan, 1997-2001. Governmentt Printing office, Nairobi, Kenya.

Roghairr CJ, Buijze A, Yedema ESE, Hermens JLM (1994). A QSAR for base-linee toxicity to the midge Chironomus riparius. Chemosphere 28:: 989-997.

Sabourinn TD, Faulk RT (1987). Comparative evaluation of a short-term test forr developmental effects using frog embryos, pp. 203-223. In:

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Bradburyy [ed.], Report 26: Developmental toxicology: Mechanism andand risk. Cold Spring Harbor Laboratory, NY, USA.

Sabourinn TD, Faulk RT, Goss LB (1985). The efficacy of three non-mammaliann test systems in the identification of chemical teratogens. J ApplAppl Toxicol 5: 225-233.

Schuytemaa GS, Nebeker AV (1999). Comparative toxicity of ammonium andd nitrate compounds to Pacific treefrog and African clawed frog tadpoles.. Environ Toxicol Chem 18: 2251-2257.

Scribnerr EA, Battaglin WA, Goolsby DA, Thurman EM (2000). Changes in herbicidee concentrations in Midwestern streams in relation to changes inn use, 1989-1998. Sci Total Environ 248: 255-263.

Stamperr DM, Tuovinen OH (1998). Biodegradation of the acetanilide herbicidess alachlor, metolachlor, and propachlor. Crit Rev Microbiol 24:: 1-22.

Steenn RJCA, Leonards PEG, Brinkman UAT, Barceló D, Tronczynski J, Albaniss TA, Cofino WP (1999). Ecological risk assessment of agrochemicalss in European estuaries. Environ Toxicol Chem 18: 1574-1581. .

Vaess WHJ, Ramos EU, Verhaar HJM, Hermens JLM (1998). Acute toxicity off nonpolar versus polar narcosis: Is there a difference? Environ ToxicolToxicol Chem 17: 1380-1384.

Wakee DB (1991). Declining amphibian populations. Science 253: 860-860.

Websterr E, Mackay D, Wania F (1998). Evaluating environmental persistence.. Environ Toxicol Chem 17: 2148-2158.

Wernerssonn A-S, Dave G, Nilsson E (2000). Assessing pollution and UV-enhancedd toxicity in Torsviken, Sweden, a shallow bay exposed to contaminatedd dredged harbor sediment and hazardous waste leachate. AquatAquat Ecosyst Health Manage 3:301-316.

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Summary y

Deducingg the potential risks of pesticides by considering the parent compoundss alone is inadequate, because pesticides do transform in the environmentt after application. This thesis addressed the adverse effects of pesticidess arising from their environmentally stable breakdown products and exploredd the probable long-term effects of these degradation products in comparisonn to those arising from the parent compounds.

Basedd on the massive usage of pesticides and potential for increased pesticidee use in its catchment, the river Nzoia (Kenya) was studied for contaminationn with chloroacetanilides and their stable aniline degradation productss (Chapter 2). Alachlor, metolachlor and their respective environmentallyy stable aniline degradation products, 2,6-diethylaniline and 2-ethyl-6-methylanilinee were analyzed in water and sediment samples from 99 sites along the river, using gas chromatography. These tests revealed a comparativelyy higher concentration and frequency of detection of the degradationn products than the parent compounds in both sediment and water samples.. A widespread occurrence of the degradation products during the studyy period indicated their persistence in the environment. It was concludedd that the prevailing tropical climatic conditions favor a quick breakk down of the pesticides into their environmentally stable degradation products,, thereby making the latter more important pollutants than their parentt products in the study area. The toxic risk of these degradation productss were therefore investigated in the succeeding research steps (Chapterss 3-5).

Itt is generally believed that degradation products of pesticides are lesss toxic than their parent compounds. This belief is, however, invariably basedd on the acute lethal effects. Long term risks and other toxicity endpointss are seldom considered. Therefore in Chapter 3 the relative toxicologicall properties of the degradation products were investigated in differentt biological systems. Toxic and genotoxic effects of alachlor, metolachlor,, amitraz, chlordimeform, their respective environmentally stablee degradation products 2,6-diethylaniline, 2-ethyl-6-methylaniline, 2,4-dimethylaniline,, and two other related compounds, 3,4-dichloroaniline and anilinee were compared. Acute toxicity tests with C. riparius (96 h) and V. fischehfischeh (Microtox®) and genotoxicity tests with a dark mutant of V. fischeri (Mutatox®)) were carried out. Our results demonstrated that toxicity and

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SummonSummon7 7

genotoxicityy of the pesticides are retained upon degradation into their alkyl-anilinee metabolites. In the case of the herbicides alachlor and metolachlor, thee toxicity to V. fischeri was enhanced upon degradation. Narcosis alone explainedd the toxicity of the compounds to the midge, but not so for the bacteria,, suggesting a disparity in the selectivity of the test systems. All compoundss showed direct genotoxicity in the Vibrio test, but amitraz and its metabolitee were genotoxic at concentrations 103-105 lower than all the other compounds.. These observations indicate that stable degradation products of pesticidess may contribute considerably to the environmental risks of pesticidee application and that genotoxic effects may arise upon degradation off pesticides.

Thee fact that the pesticides studied here, chloroacetanilide and formamidinee degradation products, were found to be genotoxic (Chapter 3) raisedd the question on the possible implications for developing embryos. Duringg the embryonic development the living organism undergoes several gene-guided,, cellular and molecular processes to generate a complex multicellularr organism from a zygote making the embryos the 'weak link' in aa life cycle of an organism. A substantive gene alteration will lead to death off the embryo, while a benign change may be carried on in life leading to developmentall disorders, cancers, mutations or loss of fertility. Embryos at midblastulaa to early gastrula stages of a locally abundant African clawed frogg Xenopus laevis were used as test organisms (Chapter 4 and 5). The embryoss were exposed to the test chemicals for 96h. The lethal toxicity, growthh and teratogenic effects of the compounds were investigated. An embryonicc teratogenic index (TI; which is expressed as 96h-LC50/96h-EC50(malformation))) allows comparison of teratogenic risks of diverse compoundss and mixtures. The teratogenicity of highly embryolethal compoundss would obviously be less relevant in the environment compared too that of less lethal compounds, which have a potential to cause malformationn in a large number of surviving organisms (Chapter 4-5). In additionn paraquat was tested (Chapter 5). Like the formamidines and the chloracetanilides,, paraquat is a chlorinated hydrocarbon. It is mainly photo-chemicallyy degraded on the plants by the ultraviolet light from the sun into 4-carboxy-l-methylpyridiniumm chloride and methylamine hydrochloride as thee major decomposition products.

Thee parent chloroacetanilides and formamidine were more acutely toxicc than their stable aniline degradation products, however, the latter

100 0

provedd more teratogenic (Chapter 4). The most common teratogenic effects off the parent compounds were edema as opposed to axial flexures and eye abnormalitiess for 2,6-diethylaniline and 2-ethyl-6-methylaniline and depigmentationn with encephalomegaly for 2,4-dimethylaniline. The edema iss a symptom of osmoregulatory disruption resulting from cell membrane disruptionn while the aberrations caused by the degradation products are diverse.. Therefore, the chloroacetanilides and the formamidines are potentiall sources of teratogenic transformation products. Paraquat was foundd to cause growth retardation and flexures of the notochord similar to earlierr observations in the larger embryos of Rana pipiens.

Inn the concluding remarks (Chapter 6) an overview of the chain of effectss from pesticide application to adverse environmental effects is given. Thee implication of the present results for the environment and in regulatory practicess of the pesticides in Kenya is addressed. After application pesticidess degrade through several intermediary stages to the environmentallyy stable degradation products. Environmental conditions in thee tropics favor quick breakdown of the pesticides, but the resultant transformationn products, which are persistent and more water-soluble, dissipatee more easily from the point of application to other environmental compartments.. Even though the acute lethal toxicity of the pesticides is attenuatedd upon degradation, the resulting transformation compounds presentt new types of risks, as was revealed in a battery of 4 test systems in thiss study. The transformation products retained genotoxicity, while teratogenicityy was enhanced compared to their corresponding parent compounds.. Hitherto, the toxicity of the degradation products of commonly usedd pesticides has been ignored. My findings underline the need to identify,, quantify and study pesticide degradation products in order to incorporatee them in pesticide admission and regulatory processes.

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Samenvatting g

Hett inschatten van de risico's van pesticidengebruik voor het milieu alleen opp basis van de toegediende stoffen volstaat niet, omdat pesticiden na toedieningg in het milieu worden omgezet. Dit proefschrift richt zich daarom opp de schadelijke effecten van stabiele afbraakproducten van pesticiden en maaktt een vergelijking tussen de mogelijke langetermijn effecten van deze afbraakproductenn en de eigenlijke pesticiden.

Gebaseerdd op het massale gebruik van pesticiden en het vooruitzicht opp een verdere toename daarvan in het stroomgebied van de Nzoia rivier in Keniaa is in dit gebied de verontreiniging met chloroacetanilide pesticiden en hunn stabiele aniline afbraakproducten onderzocht (hoofdstuk 2). Alachlor, Metolachlorr en hun respectievelijke stabiele aniline afbraakproducten, 2,6-di-ethylanilinee en 2-ethyl-6-methylaniline zijn gemeten in water- en sedimentmonsterss van 9 locaties in de Nzoia rivier met behulp van gaschromatografie.. Deze analyses lieten hogere concentraties en een frequenterr voorkomen zien van de afbraakproducten dan van de moederstoffen,, zowel in het water als in het sediment. Het wijd verspreide voorkomenn van de afbraakproducten gedurende de gehele studieperiode bevestigdee hun persistentie in het milieu. Geconcludeerd werd dat het heersendee tropische klimaat de omzetting van pesticiden tot hun stabiele afbraakproductenn bespoedigt, waardoor het belang van afbraakproducten voorr de milieuverontreiniging in de tropen groter is dan dat van de toegediendee pesticiden. De mogelijke schadelijke effecten van deze afbraakproductenn zijn vervolgens in de resterende hoofdstukken (3-5) onderzocht. .

Err wordt algemeen aangenomen dat afbraakproducten van pesticidenn minder toxisch zijn dan hun moederstoffen. Deze aanname is echterr alleen gebaseerd op acute letale effecten. Langetermijn effecten en anderee toxicologische eindpunten worden zelden in overweging genomen. Inn hoofdstuk 3 zijn daarom de toxiciteit en de genotoxiciteit van moederstoffenn en afbraakproducten onderzocht in verschillende biologische systemen.. Toxische en genotoxische effecten van Alachlor, Metolachlor, Amitraz,, Chlordimeform, hun respectievelijke stabiele degradatieproducten 2,6-diethylaniline,, 2-ethyl-4-methylaniline, 2,4-dimethylaniline en twee gerelateerdee stoffen, 3,4-dichloroaniline en aniline zijn vergeleken. Hiertoe zijnn acute toxiciteittesten met larven van de dansmug Chironomus riparius enn de bacterie Vibrio fischeh (Microtox®) en genotoxiciteittesten met een

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Samenvatting Samenvatting

donkeree mutant van V. fischeri (Mutatox*) uitgevoerd. Deze testen toonden aann dat de toxiciteit en de genotoxiciteit van de pesticiden gehandhaafd bleeff na omzetting in hun alkyl-aniline metabolieten. In het geval van de herbicidenn Alachlor and Metolachlor was de toxiciteit voor V. fischeri zelfs verhoogdd na omzetting. Een narcotisch werkingsmechanisme bood een goedee verklaring voor de toxiciteit van de geteste stoffen voor de dansmug. Ditt gold echter niet voor de bacteriën, hetgeen duidde op een verschil in selectiviteitt tussen de testsystemen. All e geteste stoffen bleken direct genotoxischh in de (MutatoxR) test, en Amitraz en de bijbehorende metaboliett waren al genotoxisch bij concentraties die 103-105 lager lagen dann voor de andere stoffen. Deze waarnemingen toonden aan dat stabiele afbraakproductenn van pesticiden in belangrijke mate kunnen bijdragen aan dee risico's van pesticidengebruik voor het milieu en dat omzettingen van pesticidenn in de natuur tot genotoxische effecten kunnen leiden.

Hett feit dat de hier onderzochte pesticide degradatieproducten genotoxischh bleken te zijn (hoofdstuk 3) riep de vraag op naar de mogelijke effectenn van deze stoffen op zich de embryonale ontwikkeling. Gedurende dee embryonale ontwikkeling ondergaan levende organismen genetisch bepaaldee cellulaire en moleculaire veranderingen die leiden tot het ontstaan vann een complex multicellulair organisme uit een zygote. Hierdoor wordt hett embryo wel gezien als de zwakke schakel in de levenscyclus van een organisme.. Een substantiële verandering in de genen zal de dood van het embryoo tot gevolg hebben, terwijl subtielere veranderingen kunnen doorwerkenn tijdens het verdere leven, leidend tot ontwikkelingsstoornissen, kanker,, mutaties of onvruchtbaarheid. Embryo's in de midblastula tot vroegee gastrula stadia van de in Kenia voorkomende klauwpaddensoort XenopusXenopus laevis zijn gebruikt als testorganisme (hoofdstukken 4 en 5). De embryo'ss werden gedurende 96 uur blootgesteld aan de teststoffen, waarna mortaliteit,, groeiremming en teratogene effecten werden vastgesteld. De teratogeniteitt van zeer toxische stoffen is van ondergeschikt belang, maar de teratogeniteitt van minder toxische stoffen kan tot een hoog percentage misvormingenn in de overlevende organismen leiden (hoofdstukken 4 en 5). Behalvee de eerder genoemde stoffen is in hoofdstuk 5 ook het herbicide Paraquatt getest. Paraquat is, net zoals de overige geteste pesticiden, een gechloreerdee koolwaterstof. Het wordt voornamelijk fotochemisch afgebrokenn door het ultraviolette licht van de zon, met 4-carboxy-l-methylpyridiniumchloridee en methylaminehydrochloride als de

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belangrijkstee afbraakproducten. Eerder was aangetoond dat paraquat groeiremmingg en misvormingen in de kikker Rana pipiens veroorzaakt.

Hoewell de pesticiden toxischer waren dan hun stabiele aniline afbraakproducten,, bleken deze laatste teratogener te zijn (hoofdstukken 4 en 5).. Het meest voorkomende teratogene effect van de moederstoffen was oedeem,, in tegenstelling tot ruggengraat krommingen en oogafwijkingen veroorzaaktt door 2,6-diethylaniline and 2-ethyI-6-methylaniline en pigmentatieverliess en hoofdvergroting (waterhoofd) veroorzaakt door 2,4-dimethylaniline.. Oedeem is symptomatisch voor verstoring van de osmoregulatie,, veroorzaakt door celmembraan verstoringen, terwijl de ontregelingenn door de afbraakproducten veel specifieker en meer uiteenlopendd waren. De chloroacetanilide- en formamidine pesticiden zijn duss potentiële bronnen van teratogene afbraakproducten. Paraquat bleek groeivertragingg en krommingen van de ruggengraat tot gevolg te hebben, conformm eerdere waarnemingen in de grotere embryo's van de kikker Rana pipiens. pipiens.

Inn het concluderende hoofdstuk (6) wordt een overzicht gegeven vann het traject van pesticidentoepassing tot schadelijke effecten op het milieu.. Tevens worden de gevolgen de bevindingen van dit proefschrift voor wetgevendee maatregelen voor pesticidentoepassing in Kenia bediscussieerd. Naa toepassing breken de pesticiden af tot stabiele producten en het heersendee tropische klimaat versnelt deze afbraak. De aldus ontstane afbraakproductenn zijn persistent en beter wateroplosbaar dan de moederstoffenn en verspreiden zich daardoor makkelijker over andere milieucompartimentenn dan degene waarin ze zijn toegepast. Hoewel de acutee letale effecten van de pesticiden afnemen na omzetting, vertegenwoordigenn de nieuw ontstane afbraakproducten andersoortige risico's,, zoals in deze studie, mede door het gebruik van 4 verschillende testsystemen,, naar voren kwam. De afbraakproducten waren vergelijkbaar genotoxisch,, terwijl hun teratogeniteit zelfs hoger was dan die van de moederstoffen.. Geconcludeerd wordt dat de schadelijke effecten van afbraakproductenn van algemeen toegepaste pesticiden genegeerd worden. Dee bevindingen van dit proefschrift onderstrepen de noodzaak tot het identificeren,, kwantificeren en onderzoeken van afbraakproducten van pesticidenn om ze op te nemen in pesticide toelatingseisen en wetgevende maatregelen. .

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Acknowledgements s Thiss study benefited from the generous grants of the Nuffic-MHO

thee Netherlands, through the Moi University School of Environmental Studiess and University of Amsterdam (MUSES/UvA) Program. I relied on thee administrative support of Moi University, Amsterdam University and MUSES/UvAA project. For this, I would like to express my special gratitude too Annemieke van Haastrecht for her able leadership of the MUSES/UvA project;; Dr Yabann of MUSES and Prof dr Dietz of AGIDS, UvA for coordinationn of the project; Prof Davies of MUSES, and Prof W. Admiraal off Aquatic Ecology and Ecotoxicology (AEE, UvA), also my supervisor. I amm also grateful to Helen Bergman for her administrative role at the AEE.

Thee research work within this thesis was done in Kenya and the Netherlands.. I am very grateful to all the people that made this possible. I amm especially thankful to Mr Lewela and Mr Ndwiga of MUSES laboratory, andd the AEE (UvA) team. To mention just a few they cared for the test organisms,, ensured that the laboratories operated well and passively or activelyy showed me some tricks in the lab and on my desk (the list is inexhaustible).. I would like to thank Prof Oduor-Okello, Prof Owiti and Dr Mwangii of Department of Vet Anatomy, University of Nairobi for their pricelesss support and pieces of advice.

Thee patient and firm support from my supervisor Prof dr. W Admiraall prevented any disorders in the development of this thesis. I cannot thankk Dr MHS Kraak enough for his dual supervisory and friendship roles duringg the development of this thesis.

Awayy from home this endless list of people provided me with the neededd moral support: Harm, Eric, Ze, Frank, Gerdit, Elske, Saskia, Nuria, Christiane,, Bas, Reanne, Evelyn, Dick, Jaap, Jenny, Cees, Esther, Tineke, Peter,, Heather, Sam, Matthijs, Maarten, Thank you all for your contributions. .

Myy family and friends have been my pillar of strength at the difficultt moments. Thank you mum for molding me during my formative lif ee stages. Jane, your patience and round the clock support where and when thee rest of the world would not was great. And lastly to those who would caree to know and myself: 'kinda piny emanyalo gimoro'.

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ABOUTABOUT THE A UTHOR

OdipoOdipo Osano graduated from Maseno high School (1983) and

taughttaught at Rangenya Primary School (1984-1985). He thereafter joined

UniversityUniversity of Nairobi for Bachelor of Veterinary Medicine (1985-1989). He

waswas awarded a post-graduate scholarship for a Master of Veterinary Public

HealthHealth by Norwegian Agency for Development (NORAD) in 1989 which

culminatedculminated in a MVPH thesis (1992), excerpts later published as: O. Osano

andand M. Arimi, Retail Poultry as sources of Campylobacter jejuni, East Afr

MedMed J (1999) 76: 141-143. He doubled as a Veterinary officer incharge of

publicpublic health in Busia (1990-1992) and Kakamega (1992-1996) districts

(Kenya),(Kenya), and as a veterinary clinician. He is currently affiliated to the

SchoolSchool of Environmental Studies, Moi University. This thesis is a result of

studiesstudies and research conducted mainly at Moi University (Kenya) and

UniversityUniversity of Amsterdam, (The Netherlands) and partly at other research

institutions.institutions. Dr. Osano aspires to consolidate his varied experience of

workingworking at different environments to stimulate research in ecotoxicology in

thethe tropical regions.

Welcomee home: www,osano. freeservers.com

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