2.4.2 Radio-ecotoxicology

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The term ecotoxicology has been defined elsewhere (167). This part of the review is concerned with toxicity or potential toxicity as a result of the ionizing radiation emitted by radionuclides contaminating soil, crops, and livestock. The term radio-ecotoxicoIogy is used to distinguish these effects from those of non-radioactive chemical residues or pollutants.

The detection of abnormal radioactivity in soil or environment implies the presence of one or more anthropogenic radionuclides as the free element or chemical compound While their biological significance will be due to the emitted radiation, their behaviour will be due entirely to their chemistry or biochemistry (when present in living organisms). In the context of radio-ecotoxicology it is important to recognize two features:

Firstly, the chemical weight of a fallout radionuclide will, relatively, be extremely small since the reactor fission products and transuranium products (actinides) will be largely emitted in the isotopically pure or 'carrier-free' state.

Secondly, possible radiation effects will depend very much upon the physical environment after fallout. Thus, a particular contamination level of soil (conventionally expressed in becquerels per square metre (Bq m^-2) can have very different implications according to physical environment and, therefore, according to the time after deposition.

The chemical weight or mass of an isotopically pure radionuclide is readily calculated (133) for a given level of radioactivity (Bq) on the basis of its radioactive half--life, mass number, and the 'Avogadro' number as illustrated in Table IX. Some non-radioactive (stable) fission products also accumulate in the reactor core (11) but not in sufficient quantities to affect significantly the carrier-free condition of the emitted radionuclides.

The low concentrations implied by Table IX have important implications in relation to the decontamination of freshly exposed and harvested crops as discussed earlier (Section 1.2.4). A level of 10,000 Bq m^-2 of iodine-131 per square metre of mature cabbage at a harvest yield of, say, 100 t ha^-1 (10 kg m^-2) would correspond to an average chemical concentration in the whole cabbage of some 2 x 10^-7 m g kg^-1.

Clearly, a fresh deposit by radioactive fallout will, initially, rest largely on the exposed soil and above-ground crop surfaces. Personnel working or walking over the soil would, therefore, be maximally exposed to the radiation emitted directly from such surfaces or, indeed, to exposure by inhaling re-suspended soil dust although this would normally represent a 'minor pathway' (see Section 1.2.2). Likewise, the consumption of such exposed and freshly harvested food crops, or pasture by grazing livestock, would lead to maximal ingestion of the fallout radionuclides. However, with time or through deliberate intervention the deposit will tend to move down the soil profile with natural leaching, irrigation, or admixture within the soil by weathering disturbances, ploughing, etc. (see Sections 2.5.5 and 2.7).

Radiation emitted from the deeper layers will be largely absorbed (self-absorbed) by the containing soil mass. Hence, with time there will be a continually diminishing fraction of the emitted radiation escaping from the surface (as distinct from any decline of radiation as a result of radioactive decay). Likewise, with time, diminishing fractions of the radioactive deposit will be taken up by the crop root system. The effects of vertical dispersion can be roughly indicated as follows:

TABLE IX
Mass of carrier-free radionuclide equivalent to a given level of radioactivity

Footnote: Radioactivity levels of 1 MCi and 1 kBq respectively typify the orders accumulating within the reactor core and as observed per square metre of soils from post-Chernobyl fallout. The first three naturally-occurring radionuclides are included for comparison.

If the significant deposit consisted mainly of cesium-137 and iodine131 (both beta-particle and gamma-ray emitters) then when, say, less than 20% remained in the top 10 cm of soil the entire beta emission would be effectively self absorbed by the soil and less than 10% of the gamma radiation would penetrate upwards through the surface (estimate based on ref. 168, p. 364).

It is for these and other (see below) reasons that a fresh deposit of, say, 5,000 Bq m^-2 of cesium-137 would represent a far more significant radioactive contamination level than the same deposit dispersed to a depth of 30 cm or more of soil. Indeed, while levels of this order prompted intervention in parts of Europe soon after 'Chernobyl' the pre-Chernobyl levels of cesium-137 accumulated since 1945 (mainly as the result nuclear weapons testing in the atmosphere) would be of the same order in many areas of Europe (see Table VII).

Radio-ecotoxicological considerations here raise wider questions than human health. For example, are there likely to be adverse radioactive fallout effects upon soil-dependent fauna and flora per se? In particular, adverse effects upon the vital microbiological transformations in soil such as those of nitrogen biofixation and mineralization and, indeed, in the mobilization of radioactive contaminants themselves (73, pp. 552-555) or upon plant life itself? It is well established that sufficiently high radiation doses are lethal to all known forms of life. Indeed, irradiation by sealed gamma-ray sources is a well established method of sterilization or selective tissue destruction in vivo (169).

Some review of the immense range of studies of biological effects in organisms other than man (e.g. 21; 62; 73-75; 77; 170; 183; 213-215) indicates the following:

Because of the enormous complexity and variety in habitat, morphology, biochemistry, physiology, and natural longevity of living organisms there are correspondingly large variations in the biological effects of ionizing radiation exposure. Likewise, the significance of these effects differs in the context of human health and welfare.

External irradiation of the skin of pigs led to somatic lesions (radiation burns, etc.) as in man (73, pp. 884-886). Iodine-131 in rats caused thyroid cancer, and inhaled cerium-141 and 144 induced lung cancer in rats (170, Vol. II, pp. 21-44). The lethal dose-50's for animals tend generally to be higher than those for man (e.g., see ref. 170, Vol. I, p. 119) and those for the smaller forms of life are much higher still. Insects are much less dependent than mammals upon cell-renewal systems (170, Vol. I, p. 103) and would, therefore, be relatively immune to the delayed stochastic effects observed in higher animals. In any event, the relatively short life-spans of rodents, insects and microorganisms, for example, imply that many low-dose effects of possible 'late' significance in man would be of no significance to many classes of wildlife (170).

Irradiation can cause genetic damage and accelerate mutation rate in insects with their implied 'hereditary' effects. Some experiments have indicated that low-dose irradiation of plant seeds and seedlings could actually have beneficial effects (170, Vol. I, p. 141).

There was no evidence that present accumulated levels of fallout actinides (e.g. plutonium-239) as a result of nuclear weapons testing or from discharges by nuclear processing plants had any effect on soil microbiological nitrification. However, such effects might be expected at higher concentrations when inhibition could be due to chemical toxicity per se rather than due to the emitted alpha-radiation (73, pp. 667-673).

The relatively rapid emergence of strains of microorganisms, insects, and rodent pests, which have become resistant to previously effective chemical biocides is well recognized (135). Its increasingly serious implications for agriculture and human health now attract media attention (e.g., 171). The development of radiation-resistant microorganisms has been observed (76) which, likewise, suggests the ability of microbiological populations to adapt to higher than normal radiation levels.

When considering the possible effects of radiation on soil organisms and upon soil dependent flora and fauna it should be noted that all other living organisms than man and his immediate domestic animals are subject to the range of naturally inhibiting or lethal environmental stresses. Moreover, the gathering, hunting, cultivation and harvest of living organisms as human food have clearly been a condition of survival since the beginning of human evolution.

Earlier reviews (55, p. 28) have indicated that "the amount of radioactive contamination required to destroy or injure and so reduce the productivity of plants and animals is much greater than that which would render the resultant foodstuffs unsafe for human consumption; the ratio being several orders of magnitude in the case of radioactive iodine, strontium and caesium . . . . .".

Against this background and because of public sensitivity to possible or imaginary radiation harm to man a concensus has emerged that "it is the impact on man" that would be "important rather than the effects on other components of the biosphere" (77, Vol. I, p. 1).

2.4.3 Occupational and public health implications

The considerations above indicate that, in the contamination of soils, possible effects on soil-dependent fauna and flora can be ignored in the present context. In considering the contamination of soils in relation to human health three situations can be recognized (for units, terminology, etc. see Sections 3 and 4):

(a) Initial deposits on soil and exposed crop surfaces leading to unacceptable levels of external radiation exposure and consequent absorbed radiation doses for farm personnel working or walking over the ground ('occupational' exposure) - see Fig. 3 for illustration.

(b) Deposits where external radiation can be neglected but where levels of direct crop or pasture contamination could lead to unacceptable levels of crop or livestock contamination for harvest as food, and for drink such as milk from dairy cattle grazing on directly contaminated pasture (problems of local and public health).

(c) Significant levels remaining in or on the soil which could indicate the need for constraints on future crops or products as a result of plant uptake from the soil (problems of Local and public health?.

Significant radiation exposure as a result of a passing cloud of radioactive emissions or through the inhalation of radioactive dusts, aerosols, etc. would be unlikely to arise except in the vicinity of an extremely serious accident or because of some more distant but unusual weather condition (75). For example, in Sweden after 'Chernobyl' although ground deposits of 100,000 Bq m^-2 and more were recorded in some cases, doses from these external radiation sources were reported as "negligible" (82).

External radiation exposure from the soil and floral cover will, of course, depend upon a number of factors (see above), especially radionuclide composition of the fallout (and, therefore, upon the time of exposure after the fallout), upon the precipitation conditions at the time of deposition and afterwards, and upon the nature of the aerial biomass at the time of deposition.

These factors probably account for the different external radiation dose rates per becquerel per square meter implied in some of the post-Chernobyl reports. For example, in the Federal Republic of Germany (150) "top level" ground deposits of 100,000 Bq m^-2 representing an "absorbed dose" equivalent to 0.15 mrem per day were indicated for persons "permanently on such ground". This would correspond to approx. 0.06 m Sv hr^-1. Data from Sweden, on the other hand, implied (82, pp. 4 & 10) that in the "regions of highest contamination" local levels of the order of 1 MBq m^-2 (i.e., for iodine-131 and cesium-137) were reached and involved external irradiation dose rates of the order of 10 m Sv hr^-1, i.e., approx. 1 m Sv hr^-1 for a deposit of 100,000 Bq m^-2 (compare above).

Absorbed radiation dose equivalent rate for a person standing on ground contaminated by a significant deposit of gamma-emitting radionuclide can be measured with a suitable monitor (usually at the conventional height of one metre above ground level). Alternatively, it can be estimated by the application of an empirical equation to an already measured deposit in terms of Bq m^-2 for one or more radionuclides. For estimating effective dose equivalent (EDE) commitments as a result of exposure to external ground or surface sources, dose rates must, of course, be integrated over an appropriate time into the future to indicate the total dose likely to be accumulated in that future (88, p. 16 and appendix; 80, pp. 27 et seq). The following equation relates dose rate and deposit:

DR = SD_M x CF (Eq. 1)

where DR is rate of absorbed dose equivalent in m Sv hr^-1, SD_M the ground surface deposit in MBq m^-2 (mega Becquerels per square metre), and CF the appropriate conversion factor for the particular radionuclide (80, pp. 27 et seq) as indicated below:

Thus, equation (1) indicates a dose rate of 0.075 m Sv hr^-1 for a surface deposit of 100,000 Bq m^-2 (0.1 MBq m^-2 of iodine-131 which can be compared with the data quoted at the bottom of the previous page. Integrated doses have been tabulated for continuous periods of 1 week, 1 year, and up to 50 years spent on the ground (81, pp. 56-57).

For the protection of personnel exposed to both significant external sources and as result of radionuclide intake by food and drink the total dose equivalent commitments might, of course, have to be considered additively (157) and associated derived intervention levels, or their application, modified accordingly (see above and Section 3).

2.5. Behaviour in soils and movement into foodwebs

2.5.1 Effective 'disappearance' and redistribution
2.5.2 Uptake by crops
2.5.3 Scenario for movement into food
2.5.4 Movement in foodwebs
2.5.5 Leachin
2.5.6 Application of models to soil-food chain transfer

2.5.1 Effective 'disappearance' and redistribution

Effective 'disappearance' from within the soil or from the soil surface deposit will depend upon radioactive decay on the basis of known half-life (see Table I and footnotes), surface erosion and possible volatilization, leaching and other transport down the soil profile, and removal by plant uptake and harvest. There is an immense literature on these aspects which can be no more than illustrated here (e.g., see 62; 77; 59).

Uptake of soil radionuclides will depend upon the radionuclide, its chemical concentration (not upon the level of radioactivity), its chemical form, the presence and availability of natural isotopes of the same element, its distribution down the soil profile, its availability to the plant root system, plant, and plant growth status.

As explained above the chemical concentration of the radionuclide will be extremely low. Therefore, initially, it is likely to be relatively firmly held at abiotic adsorption sites on the basis of the well-known mechanisms of adsorption (172; 166) and movement by solution into soil water and loss by volatilization from the surface will likewise tend to be relatively slow. However, effective chemical concentration of the radionuclide will also depend on the presence and availability of naturally-present isotopic element, but those of the fallout radionuclides are also relatively small as illustrated in Table X (from data of ref. 77). Natural elements of the actinides (plutonium, americium, etc.) do not exist. 'Availability' and movement of the radionuclide will also be greatly influenced by the presence of chemically-related elements and their 'availability', e.g. Cs-134 and -137 by available potassium, Sr-89 and -90 by available calcium (see Tables I X and XI).

TABLE X
Reported concentrations in soils and plants of naturally occurring elements of fallout radionuclides m g.g^-1 dry wt (from ref. 77)

Note: A useful and comprehensive review of micronutrient elements and their distributions in soils worldwide is available (173).

However, raising the effective concentration of the fallout radionuclide by these mechanisms of 'isotopic dilution' and chemical mixing will require time, availability, and the necessary mobility in solution or by dry diffusion.

In short, raising the 'effective' chemical concentration (other things being equal) will tend to reduce the proportion of radionuclide taken up by the plant root system, and increase its availability to move down the soil profile by leaching as a result of de-adsorption. The behaviour of radionuclides in the soil-plant system as a function of concentration, soil pH, etc. has been reviewed by Scott Russell and his colleagues (53, pp. 89-124).

For the reasons indicated above, except for short lived radionuclides such as iodine-131, the "disappearance" from, and movement within soils, tend to be slow. The persisting levels of much earlier fallout illustrated in Table VII are not, therefore, surprising. Nor would they be more than marginally accounted for by the continuing and detectable fallout before 'Chernobyl' (see Table VI).

Studies in Switzerland (150) indicated that cesium-137, deposited as a result of the atmospheric testing of some 500 atomic and hydrogen bombs during the 1950s and '60s, sank "only a few millimetres . . . . . each year". Strontium-90 is "bound in the earth more weakly and is now up to 30 cm deep". This slow movement does have the advantage of lessening the possibility of groundwater pollution.

Many studies have been made of the "availability" of radionuclides deliberately added to representative soil types. Some data are illustrated in Table XI.

TABLE XI
"Available" fractions of radionuclides in different soils after 6 months (from ref. 73, pp. 660-662)

Note: Other isotopes of the same element (e.g., Cs-137, Sr90) would, of course, give the same results.

Although ruthenium-106 can exist in anionic or cationic forms in soil the major fraction tends to be "strongly" fixed and, like cesium, moves downwards only a few millimetres per year (77, Vol. I, pp. 280-281). Strontium-90 is largely held at exchange sites in the soil so that movement tends to be greater in soils which are high in organic matter or in the chemically related calcium (77, Vol. I, p. 98; see also Section 7.3). Cesium-134 (and 137) are influenced by the "form of vegetation overlying the soil". In the presence of a litter horizon a high proportion would be bound, and movement by leaching or erosion would be small (77, Vol. I, p. 326; see also Table XI).

The major source of plutonium-239 (and its other isotopes) in contemporary soils is from the earlier atmospheric testing of nuclear weapons. For example, no plutonium-239 was detected in Switzerland from the Chernobyl fallout (150). Plutonium isotopes can show relatively high mobility in acid or alkaline soils depending on clay and chalk content. Physical adsorption of plutonium in soils can be reversed by uranium addition. More than 90% of plutonium in soils would be expected to become rapidly adsorbed by clay particles (77, Vol. IV, pp. 1-7); see also americium data in Table VIII).

A major factor in the movement of deposited radionuclides into soils will be ploughing (see Section 2.7.3), hence the difference observed in the distributions of earlier deposits between pasture and cultivated land. However, ploughing will not, necessarily, affect plant uptake (77, Vol. I, pp. 17-18).

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