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«: AGROCHEMICALS: FATE IN FOOD AND THE ENVIRONMENT PROCEEDINGS OF A SYMPOSIUM, ROME, 7 - 1 1 JUNE 1982 JOINTLY ORGANIZED BY IAEA AND FAO l^J I N T E R ...»

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(1 ) Many pesticides are lipophilic, apolar compounds. They are usually metabolized to more hydrophilic, polar compounds (see Refs [43—45]) or they may be stored 'forever' in the b o d y fat.

(2) Most metabolic transformation leads to less toxic compounds. There are some important exceptions to this principle, e.g. parathion may be transferred to paraoxon, aldrin to dieldrin and metals, particularly mercury, to methylmetals. Methylation pathways are discussed by Wood and Fanchiang [46] and a radioanalytical method to determine vitamin B-12, the main methylating enzyme, is described in Ref. [47]. There are even pesticides like dichlorbenil, which is formed out o f its inactive precursor chlorthiamide in the environment.

IAEA-SM-263/34 133 (3) Warm-blooded animals show a faster turnover than cold-blooded [48, 49].

In this context it should be mentioned that arctic fauna, with its long-living organisms, shows a particularly slow degradation [50].

(4) Korte [18] gives the following guidelines: alkyl groups are less persistent than branched groups, alkenes less than alkanes, and alkanes less than aromates.

The persistence o f rings increases with the number o f substituted hydrogen atoms.

Degradation of pesticides shows similarities to degradation o f natural organic compounds. The most prominent reaction is oxidation. Dagley [51] discusses various biochemical aspects o f the biological oxygen activation, the central problem of which is that the ground-state o f all organic compounds is a singlet, whereas dioxygen occurs in the triplet state. Thus, a straightforward reaction is forbidden.

In higher animals the mixed function oxydases o f the liver are the most important scavengers o f the organism [52]. Radiotracer methods are frequently applied in degradation studies. Most o f this work is devoted to soil rather than limnic systems. However, the results have relevance for aquatic systems too.

The differences are mainly:

(1 ) Primary producers occupy a much larger volume in the aquatic system (euphotic layer) (2) Protozoa constitute a larger part of the biomass o f the reducers (3) Transportation, particularly from oxidizing to reducing sites, is much faster owing to turbulence.

Often the complete degradation scheme of a pesticide may be t o o tedious to be followed in detail. In such cases the 14 C-labelled pesticide may be introduced into a bioassay and the I 4 C-dioxide evolving after complete mineralization is recovered (see Flores-Ruegg et al. [53]). Once the original material is transferred to 1 4 C-dioxide one can be sure that no potentially hazardous remnants are left, whatever the intermediate steps may have been. In such cases care must be taken not to take evaporated pesticides for 1 4 C-dioxide. A comprehensive collection of methods is compiled by Compaan [54]. Experiments on pesticide degradation are much too numerous to be reported on here.

The following papers have been selected to show some typical radiotracer

applications:

Metcalf [55] designed the classical approach: an aquarium-like model ecosystem in which radiolabeled pesticides can be administered. Klein et al. [56] report the fate o f 14 C-labelled aldrin, heptachlor and lindane in an outdoor model ecosystem containing soil and plants. Losses through evaporation and leaching water were checked and degradation products discussed. Sethunathan et al. [ 5 7 ] followed the degradation of 14 C-parathion in rice soils and observed that organic 134 ERNST matter enhanced its reduction. Carbon-14-dichlorbenil has been shown to be readily absorbed by fish and weakly hydroxylated; most o f it evaporated [58].

The fate o f 14 C-atrazine and 2,4D in bentic fresh water animals is described in Ref. [59]. A slight increase in 2,4D metabolizing microbes was observed after 2,4D treatment [60]. The fate o f 14 C-lindane, which is becoming o f major concern since the DDT ban, has been studied by Saha [61 ]. Degradation o f pesticides may follow remarkably different pathways, whether or not it occurs in oxidizing or reducing environments. The numerous metabolic pathways o f anaerobic organisms can sometimes attack compounds that are persistent in aerobic environments, e.g. DDT is converted to DDD in anaerobic soil [62] or lake water [63]. However, anaerobic conditions are disadvantageous for oxidative degradation reactions.

In Refs [64, 6 5 ] tracer experiments on thiobencarb and phorate show this effect.

Bioassays for such experiments require use o f gas-tight containers flushed with helium, carbon dioxide, hydrogen or nitrogen. The ultimate degradation product of 14 C-labelled compounds need not be 1 4 C-dioxide [66]. A review on the anaerobic degradation o f pesticides is given by Williams [67].

4. BIOACCUMULATION

This term is used here in terms o f bioaccumulatory enrichment of environmental substances in organisms. Accumulation in sludge and similar problems are discussed in Section 5. Accumulation o f pesticides in the f o o d chain represents a major threat to man. This danger increases with the persistence o f the substance and with its affinity to fat tissue. Accumulation and degradation are intimately linked. Often in a radiotracer experiment certain activity is found in some compartment o f the bioassay and it is not known whether it is the parent pesticide or some degradation product. This will be dealt with under the term accumulation. This applies, for example, to the 'bound residues' in soil systems that have a certain analogue in limnic systems in the b o t t o m sediments. Examples o f bound residues and sediment accumulation o f a series o f pesticides are given in Ref. [68]. Many of the references discussed in Section 3 also deal with problems of accumulation.





The interdependencies o f all these questions lead to the concept o f ecotoxicological profile analysis [35]. The pesticide that is best investigated with respect to its bioaccumulation is DDT. It has been estimated that the world's biota contain as much DDT as was produced in 4 d in the mid-sixties [69].

Storage in fat tissue is the most important effect in bioaccumulation. Hence, lipophilicity is one o f the most important properties o f environmental chemicals.

For determination o f this property one needs a 'standard fat'. In pharmaceutical research n-octanol, although not a fat, has proved that it fits this purpose. The water-n-octanol partition coefficient is currently the universally accepted parameter to predict accumulation in lipids. The experimental procedure comprises shaking IAEA-SM-263/34 135 in a water-octanol mixture and determination of the ratio, usually by gas chromatography; details are given by Compaan [70]. If the substance in discussion is identified, the partition ratio can also be determined theoretically on the basis o f the chemical structure; for details see Ref. [70]. Data on the fat content o f different species are given by Morowitz [71 ] and particularly by Jorgensen [72].

Bioassays for accumulation tests differ from those for toxicity in so far as the concentration is kept well beyond the toxic level. Animals, with their fat tissues, are most prone to accumulation o f lipophilic pesticides and therefore they play a role in the human f o o d chain. Hence, most research has been carried out on crustaceans, clams and fish.

Uptake may be passive diffusion and sorption, active transport, or uptake via f o o d. The different forms are not always easy to discriminate. Gills are usually the most important organ o f uptake. The time courses for uptake and elimination have sometimes to be interpreted theoretically, particularly to get reasonable predictions on the times o f incubation or exposure. It is usual to apply first-order kinetics similar to the concept o f biological half-life applied in radiation biology; for theoretical details see Ref. [73] and for experimental evidence see Ref. [68]. An extended list o f half-life values is found in Ref. [72]. Streit and Schwoerbel [ 5 9 ] report hyperbolic kinetics. In the same paper bioaccumulation o f 10 pesticides in Chlorella fusca is reported. A case study on the accumulation o f mercury in a Mexican river system has been carried out by Baez and Nulmann [74].

Interesting background information for pesticide adsorption is given in Ref. [75].

Sorption o f 12 14 C-tagged pesticides to humic substances and two synthetic polymers were measured and interpreted in terms o f charge-transfer and hydrophobic interactions.

5. TRANSPORTATION A N D SOURCES OF PESTICIDES

Sources of pesticides in water are either intentional or unintentional contaminations. The latter comprise surface runoff, leaching, dumping o f pesticide remainders, washing o f spray facilities and wind drift during spray operations.

Particular problems arise in rivers that flow through heavily industrialized areas. In Ref. [76] the case histories o f the Hudson and Rhine rivers and o f Galveston Bay, Texas, are given. Intentional contaminations originate from chemical ditch clearing or eradication o f aquatic insects. A list o f the different sources o f pesticides is given in Ref. [17]. The problems o f ditch clearing are dealt with in Refs [12, 77]. These authors do not see any serious environmental threat through using these methods provided the herbicide is handled carefully. The poisons in use are not persistent and significant contamination o f waters downstream o f the spraying area has not been observed. It should, however, be pointed out that no observation on species composition after extended use has been made. Most 136 ERNST of the herbicides in use were strongly bound to a particular material, a fact which stresses the importance o f sedimentology in this field. The concentrations in use are given by K o c h and Hurle [78], i.e. approximately 1 ppm (assuming 1 m depth).

They assume that aeroplane spraying is only a minor source o f contamination, whereas Erne [ 7 9 ] claims that negligent dumping o f pesticide remainders is a major source.

Particular precaution has to be applied if large water bodies are treated for insect eradication programmes, e.g. problems of blackfly (Simolinum arcticum and S. luggeri) eradication with methoxychlor in large Canadian rivers have been reported by Haufe [80]. The insecticide was best administered at several points distributed across the river and with careful timing concerning non-target taxa.

Absorption to particulate matter is discussed. Another large project is the Onchoceriasis Control Programme in the Volta river. In Ref. [81 ] the effects o f aquatic pesticides are reported; these differ as to whether or not target taxa are emergent, submerged or floating weeds and whether or not they are rooted plants. Water soluble pesticides are usually degraded or evaporated within days or weeks. Vaporization is less important if the material is absorbed by sediments;

this was shown with 1 4 C-thiobencarb by Ishikawa [64]. Diquat absorption to the sediments and its incorporation to the bottom fauna are shown in Ref. [82].

Sludge absorption o f PCP, PCNB and HCB are described by Klein and Korte [68].

Problems similar to pesticide transport by sediments occur by surface runoff Metribuzin, trifluralin and MSMA have been studied in this concern by Wauchope et al. [83]; Wauchope has also reviewed the problem [84]. It has, for example, also been found that about 0.5% o f applied commercial pesticides is lost through runoff. White et al. [ 8 5 ] also found insignificant runoff losses for 2,4D, whereas Vrochinskij [86] maintains that runoff is the commonest cause o f surface water contamination.

The properties o f sediment, such as grain size and consolidation (squeezing o f water out o f clay particles), and their relation to erosion and deposition and metal ion absorption are treated by Goltermann [ 1 ] and further details are given in Ref. [87]. Vaporization is another important feature o f pesticide transportation, e.g. in the 14 C-aldrin experiment reported in Ref. [68] about one-half o f the pesticide evaporated. In many cases the water-air partition coefficient for sparingly soluble compounds may be calculated theoretically by

К = 62.4 * С * T/P

where К is the partition coefficient (mol/1), С is the water solubility, T is the temperature (in kelvin), and P is the vapour pressure (mm Hg) [70].

However, one has to be aware that concentration o f organic molecules near the surface may not correspond to the bulk concentration. Wu et al. [88] report a.

three-fold concentration o f atrazine in the top 100—150 цm water layer. Similar results are reported by Bidleman and Olney [89] and Moriaty [90].

IAEA-SM-263/34

–  –  –

[ 11 GOLTERMANN, H.L., Physiological Limnology, An Approach to the Physiology of Lake Ecosystems, in the series: Developments in Water Science 2, Elsevier, Amsterdam (1975).

121 HARRIS, G.P., Photosynthesis, productivity and growth: the physiological ecology of phytoplankton, Arch. Hydrobiol., Erg. Limnolog. I—IV (1978) 1.

(3¡ OHLE, W., Die hypolimnische Kohlendioxydakkumulationals produktionsbiologischer Indikator, Arch. Hydrobiol. 46 (1951) 153.

[4] CAPPENBERG, Th.E., "Methanogenesis in the bottom deposits of a small stratifying lake", Microbial Production and Utilization of Gases (SCHLEGEL, H.G., et al., Eds), Akademie der Wissenschaften zu GSttingen (1976) 125.

[51 OLAH, J., "Metalimneon function in shallow lakes", Symp. Biol. Hungar. (SALANKI, J., PONY, J.E., Eds) 15 (1975) 149.

[6] FELLENBERG, G., Umweltforschung, Springer Verlag, Berlin (1977) 25.

[7] SHAPIRO, J., The Current Status of Lake Trophic Indices - a Review, Interim Report No. 15, Limnological Research Center, University of Minnesota (prepared for EPA under No.04J IP 01520) (1975).

[8] RODHE, W., "Crystallization of eutrophication concepts in northern Europe", Eutrophication: Causes, Consequences, Correctives, Natl. Acad. Sci. Symp. Washington (1969).

[9] SCHRODER, R., SCHRODER, Hanne, Ein Versuch der Quantifizierung des Trophiegrades von Seen, Arch. Hydrobiol. 82 (1978) 240.

[10] Umweltgutachten 1978, Der Rat von Sachverstândigen für Umweltfragen, Kohlhammer Verlag, Stuttgart (1978) 17.

[11] FISHER, N.S., Chlorinated hydrocarbon pollutants and photosynthesis of marine phytoplankton: a reassessment, Science 189 (1975) 463.

[ 12] MULLA, M.S., GIANCARLO, M., ARATA, A.A., Mosquito control agents in aquatic ecosystems, Res. Rev. 71 (1979) 121.



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