|"Infinite Love is the Only Truth, - Everything Else is Illusion - David Icke, 2005|
Our transformation is well underway, click here for a non-scary update
Aerodynamic size influences how far a particle may travel when released into the environment. In essence, small particles of an aerodynamic size under 10 um, particularly sub-micron particles behave as gases and as such may travel long distances whilst large particles >10 um travel less than 1 km and quickly deposit to a surface. Thus, pollutant particle size at the point of exit into the atmosphere has considerable significance on the activity, concentration and spread of that pollutant into the local or far environment. Small particles with a rate of deposition of 1 mm s-1 are about 1 um aerodynamic diameter and are more affected by air turbulence than by gravity and as a result stay airborne for much longer periods. This results in widespread dispersal of the plume with a subsequent decrease in pollutant air concentration with distance from the source.
On the other hand, an increase in the aerodynamic diameter of particle size results in an increase in the rate of deposition due to gravitational effects. I define large particles as being > 10 um aerodynamic diameter which in effect, is purely operational; i.e. it is based on the 50% particle cut-off efficiency of the widely used Pm10 air samplers. Large particles by their very nature pose a different type of environmental threat when compared to small particles. Large particles that are released from tall stacks or waste tips do not travel far, possibly no more than 1 km from their source. This is very apparent to people living close to lead smelter and fossil fuel power stations who suffer the daily nuisance of large particle deposition in the form of dirt, grime and dust. More importantly, large particle deposition leads to the contamination of the food supply by direct deposition to the upper surfaces of grass and vegetation (Pinder et al 1990). This type of environmental threat i.e. the grass-cow-milk pathway can be assessed using models. Unfortunately, however, without knowledge of the association between pollutant particle size and activity the dose assessment from these models may be up to 10-1000 times too low.
On the bright side, large particle deposition leads to a rapid depletion of activity or concentration within the plume and a consequent decrease in the inhalation pathway. Local air concentrations remain low with however, subsequent high activity fluxes. An activity flux is defined as the amount of activity per unit area per second and for radioactive material is in Bq m-2 s-1. Atmospheric scientists use the aerodynamic behaviour of theoretical particle size to indirectly measure the particle size distribution of stack-released pollutants. In effect, high activity fluxes with low air concentrations suggest the activity is associated with large particle size and this forms the basis of the concept of Vg or dry particle deposition velocity. This is one way of assessing the size of particles emitted from stacks (and hence filtration efficiencies of the stack air filters) without having to take complex and expensive in-stack measurements. I suppose it is fair to say that in attempting to measure anything to do with airborne particles, whether it is size, shape or size/pollutant distribution, we inadvertently alter the very thing we are attempting to measure. The challenge is to recognise the practical problems and theoretical limitations of existing knowledge surrounding air sampling techniques and develop something which is much more rigorous i.e. size selective air and deposition sampling.
Large scale releases of anthropogenic radioactivity to the atmosphere first occurred during the period between 1945 to 1963 when nuclear weapons testing was carried out by a number of countries around the globe. During this period it was estimated that about 400 kCi of plutonium was produced and distributed globally with up to 20% of this material deposited close to the test site (Hardy 1980). Based on the 239Pu/90Sr activity ratio, Hanson (1975) calculated that by 1965 more than 98% of the plutonium produced during the weapons test era had deposited to ground surfaces.
Other sources of man-made radioactivity present in air include closely regulated chronic (operational or 24 hour duration) releases from nuclear power stations, gaseous operational releases and acute (accidental) releases such as the 1957 Windscale No. 1 Pile fire (BNFL 1994 :Chamberlain 1991) and of course, the Chernobyl accident of 1986. Of current major interest to legislators, regulators and dispersion modellers is the Chernobyl No. 4 reactor explosion of 1986 which resulted in a sudden release of up to 25% of the reactor core inventory (Gudiksen et al 1989). It was estimated that the Chernobyl reactor source term was several orders of magnitude higher than the Windscale reactor incident (ibid.). The amount of 137Cs released due to the Chernobyl incident was however, about 6% of that released during the U.S. and USSR weapons testing era of the 1950's and 1960's (Gudiksen 1989). Much of the radioactive material that was released to the troposphere from chronic and acute releases from nuclear-powered electricity generating plants and nuclear fuel reprocessing plants deposited in the northern hemisphere.
Other less important sources of radioactivity include the fallout of SNAP 9A satellite fuel which burned up over the Indian Ocean in 1964 (Harley 1980). Radioactive fallout from this source consisted primarily of 238Pu and up to 75% of this material was deposited in the southern hemisphere (Hanson 1975). Both routine and accidental releases still occur to the atmosphere providing the opportunity for the study of the short and long range dispersal of this material into the environment using atmospheric dispersion modelling. In order to run these models a large amount of environmental data is required as is a thorough knowledge of particle transport processes and the chemical behaviour of pollutants both in the plume and on the ground. This is discussed more fully below.
Stack-released aerosols containing a mixture of moisture and radioactive pollutants deposit out of the atmosphere to soils or ground surfaces by dry and wet deposition processes (Sehmel 1980). Dry deposition processes include impaction or interception and the mode of particle transport is crucially dependent on particle size and density. Dry deposition includes all processes other than precipitation for the removal of gases and particulates to the earth's surface (Dasch 1983). Sub-micron particles exhibit gas-like behaviour and deposit to a surface via diffusion processes (Sickles 1983). Intermediate sized particles between 2-4 um deposit to plant and grass surfaces via impaction and interception (Chamberlain 1967). Large particles greater than 10 um aerodynamic diameter deposit to a surface rapidly through gravitational settling (Foster 1982).
Wet deposition processes include the removal of material by rain, hail, sleet or snow (Weiss et al 1983). Material is also scavenged out of the atmosphere by either below-cloud or in-cloud washout mechanisms (Isaac et al 1983). Below-cloud scavenging and 'rainout' occurs when material under a cloud base is effectively 'rained out' to a surface and this process is only significant for local deposition. In-cloud 'washout' refers to the incorporation and mixing of particulate material within the cloud base which moves away from a source by the action of wind and releases or 'washes-out' its pollutant load some distance away from the source. This process was especially significant for the transport and eventual washout of Chernobyl debris over many parts of Europe including high rainfall areas of the Lake District and N. Wales (Livens et al 1992).
The above explains some of the processes of the atmospheric transport of stack-released material and the next stage is to look at some of the processes of particle movement when this stuff actually hits bare ground or vegetation!!
By virtue of its surface area soils act as a sink for deposited radionuclides and particularly with regard to the environmental behaviour of the actinides in soils, the downwards migration of this group is limited to the surface layers i.e. ~10 cm (Livens 1985; Volchok et al 1972 ;Cawse 1980). Because of low soil to plant transfer factors for the actinide group, much of the plutonium present on vegetation is due to foliar contamination via direct atmospheric deposition or from resuspended soil with subsequent dose implications for the terrestrial food-chain pathway (Romney & Davis 1972). Plutonium is strongly adsorbed to soil particles and its pattern of re-distribution and environmental dissemination is largely governed by soil erosion processes. The dominant factors of soil erosion are the action of wind and water (Dreicer et al 1984)
Resuspension includes all processes capable of suspending particulate material into the near-surface air. Traditionally, the term resuspension also includes the term suspension because it is difficult to distinguish between an aerial source and that source some time after it has deposited to a surface and is subsequently resuspended (Sehmel 1980 ;Nicholson 1988). The resuspension of deposited or ground-surface material occurs through saltation, suspension and surface creep and particle transport via these mechanisms are strongly particle-size dependent (Chepil 1945: Bagnold 1954).
Particles moving in saltation have diameters of between 50-500 um and they are small enough to move by direct wind action but large enough to have settling velocities higher than the upward eddy velocity of the wind (Anspaugh et al 1975). Saltating particles move along a surface in a series of 'hops'. When saltating particles hit the ground they transfer momentum to other particles to initiate their suspension into the near-surface air. Suspended particles have a much smaller aerodynamic size range and have settling velocities less than the turbulent eddy velocities of the wind. The resuspension and transport of small particles are of concern because small suspended particles are able to travel long distances to contaminate previously clean areas and may lead to elevated air concentrations for an extended period after a contaminating event (Paretzke & Garland 1991). Secondly, suspended particles less than 80 um also contribute to inhalation dose (Healy et al 1966).
A third mechanism of resuspension is surface creep and this involves the movement of large particles up to 2 mm diameter along the ground (Anspaugh et al 1975). The quantification and attribution of one or any of these mechanisms of resuspension to field study measurements however, is difficult to determine. Chamberlain (from Prupacher 1983) however, summarised some basic conditions favourable to soil erosion and subsequent particle resuspension by wind;
Much of this historical work, however, was carried out in arid and desert environments with a deficiency of soil moisture (Bisal & Hsieth 1966) and factors which govern the initiation and sustainability of any of the above mechanisms of soil erosion and subsequent wind-driven resuspension may not be strictly applicable to soils in the more northern temperate latitudes (Garland from Pruppacher 1983).
A fourth mechanism of soil erosion and subsequent resuspension of particles relevant to this work occurs via rain-soil splash. Dreicer et al (1984) found that the impact of falling raindrops resuspended soil particles to a maximum height of 40 cm above ground and the majority of resuspended particles were less than 125 um diameter. These workers also found linear relationship between the intensities of rainfall from four natural rainfall events and the resuspension of particle size less than 125 um diameter up to a height of 40 cm above the ground. This mechanism of resuspension however, is episodic and unusually difficult to quantify under field conditions because of the range of environmental variables involved. These variables include the initial surface moisture before the last rainfall event, soil type, vegetative cover, vegetative type, wind speed and particle size-activity characteristics of the depositing or deposited nuclides.
On a local scale and with particular reference to nuclear fuel reprocessing plants it is probably the mechanical processes of resuspension which are more important than wind borne processes of particle resuspension (Hakonson et al 1980). Deposited material is re-entrained and resupended into the near-surface air from soils and concrete areas via a number of mechanical or physical transport processes. Resuspension of surface deposits by physical transport processes include ground-intrusion machinery such as earth diggers, pile-driving equipment, vehicular tyre movement and pedestrian traffic.
Agricultural activities particularly ploughing and tilling are capable of raising soil-bound plutonium and other nuclides within the near-surface air (Shinn et al 1983) although the distribution of activity with particle size will be limiting factors with regard to inhalation dose. Core analyses for the migration of Pu, Am and Cs down the soil profile for West Cumbrian soils show that over 80% of activity was limited to the first 6 cm depth (Evans 1991). The particle-reactive behaviour of plutonium suggest that under most environmental conditions this nuclide is effectively adsorbed and held within the first 5 cm of surface soil. The ability of these nuclides to be resuspended back into the near-surface air is chiefly dependent on physical transfer processes (Hartmann et al 1989).
There is little in the way of literature on the resuspension characteristics of caesium and plutonium around the site itself. This is partly due to the complicated nature of the site which is under constant construction with numerous point sources emitting material at different effective heights above ground. The deposition history of the site (see Gray et al 1995) is further complicated by historical marine releases of Pu that are transported back onto site via sea-to-land transfer processes. The efficacy of using the resuspension factor within this study is limited to:
(a) assessing the effects of on-site and off-site winds on resuspension;
(b) assessing the effects of different wind speeds on resuspension; and
(c) identifying 238Pu, 239+240Pu and 137Cs relationships between dry deposition velocities and resuspension factors for each run.
There are however, a number of problems in calculating K within the immediate vicinity of the Sellafield reprocessing plant. The resuspension factor assumes that the measured air concentration is related solely to ground contamination immediately from that area with a negligible contribution of material which may be transported via wind stress or other particle transfer processes from upwind source areas (Horst 1976). The resuspension factor also assumes that the contaminant is homogeneously distributed within the resuspendable layer.
Other limitations to the general applicability of the resuspension factor relate to the characterisation of the resuspendable fraction of the soil surface layer A number of studies do not clearly define the resuspendable layer whilst (Shinn et al 1983 ; Krey et al 1973) have defined the first 5 cm as being resuspendable. Sehmel (1980) argued that perhaps the first centimetre of surface soil more adequately represented the resuspendable fraction whilst other workers (Nicholson pers. comm. 1995) suggested using a small portable vacuum cleaner to 'grab' the more easily resuspended material. Healy (1971) argued that account must be taken of the variation of activity with particle size and the nature of the material i.e. whether the deposit is 'fresh' and heterogeneously distributed within the surface layer or aged and intimately associated with soils where soil attachment leads to a non-uniform size distribution of activity.
In recent years a considerable amount of interest has focused on the inhalation dose from the resuspension of plutonium around nuclear plants and old weapons test sites located in Australia (Johnston et al 1976 ;Fry 1983 ;Langer 1983 ;Nicholson & Fulker 1994) via wind-stress . A number of resuspension studies have focused on the inhalation dose of plutonium and americium to Aboriginal tribes-people from single-shot nuclear explosions and radionuclide dispersal experiments conducted between 1955 and 1963 at the desert environment test site in Maralinga W. Australia (Johnston et al 1993).
These authors found an average K of 10-10 with a three orders of magnitude increase in the resupension factor during dust storms with average wind speeds in excess of 10 m s-1 In an earlier study based on the immediate environs of this site, Haywood & Smith (1992) determined the total annual effective dose equivalents from a range of radionuclides including caesium and strontium to the semi-traditional lifestyle of an Aboriginal average population of adult, child and infant. They found the most limiting dose of 470 mSv to the 10 year old child group with the principal pathway being inhalation. This relatively high annual effective dose equivalent is roughly 15 times higher than a similar group some 25 km north east of the test site see also (Johnston et al 1992 ;Stradling et al 1992). The findings of many of these studies, however, may not be strictly applicable to the more temperate latitudes of the northern hemisphere where ground and soil conditions are substantially wetter during the year with average rainfalls of 1100-1300 mm y-1 (Geiss 1993 ;Playford et al 1992).
A number of workers (Garland and Nicholson 1991: Nicholson 1993 :) have noted that large particles greater than 20 um diameter are more easily resuspended from a surface than smaller particles. The least resuspendable size fraction of soils are the less than 1 um diameter particles (Gillette & Walker 1977). The major limiting hazard from the resuspension of plutonium is via an inhalation pathway, i.e. breathing in this material va the nose or mouth (Nicholson 1992).
Particles up to 100 um in diameter are generally thought of as being inhalable with the less than 4 um diameter size fraction defined as the respirable fraction (Lippman & Harris 1962). Inhaled particles deposit to either the nasopharangeal regions, the tracheo-bronchial region or the pulmonary or alveolar regions of the body (Geiss 1993). The fraction of inhaled particles which is retained in the respiratory system and the depth to which the particles penetrate before deposition is sensitively related to particle size (Brown et al 1950). Inhaled particles less than 0.5 um aerodynamic diameter may be carried deep into the lung to irradiate sensitive alveolar tissue (Burkart 1989). Particles greater than 5 um deposit out of the inhaled air stream via impaction or they are intercepted by mucus layers within the nasopharangeal cavity
Mucociliary clearance from the nasopharangeal cavity at the back of the nose transfers the deposited material into the gastrointestinal tract where it is eventually excreted from the body. Transfer factors for plutonium and americium across the gut wall for primates are very small, 10-3 and 10-4 respectively (Ham et al 1994). There is little work, however, on changes in speciation of plutonium or americium which may affect the solubility and hence bioavailability of inhaled actinides as they pass down the gastrointestinal tract.
The resuspension and inhalation of particles greater than about 5 um aerodynamic diameter are radiologically less damaging to lung tissue than particles with an aerodynamic size range less than 1 um because this size fraction is less respirable (Geiss 1993).
The figure below show the probability of deposition in various parts of the lung relative to particle size
Size-dependent deposition probabilities of particulate radioactivity in the different compartments of the human lung (ICRP 1975)
Dry particle deposition velocity is defined as;
activity flux Bq m-2 s-1
air concentration Bq m-3
(Chamberlain and Chadwick 1953)
Deposition samples and airborne concentrations are usually collected at a reference height of ~1 m above ground level. This reference height above ground is thought to be representative of atmospheric material in suspension (see chapter 1.4) unaffected by surface resuspension or re-entrainment processes (Sehmel 1980). The derivation of dry particle deposition velocity is subject to much environmental variability and Vg values can vary by over 3-4 orders of magnitude under similar field test studies (Sehmel 1980). Much of this variability is attributable to constantly changing environmental variables particularly wind speed, and the heterogeneous nature of deposited material. It is also difficult to compare air and deposition experimental results even from the same latitude due to non-standardised sampling protocols for collecting airborne and deposited material with respect to particle size.
The height of the dry deposition collectors used in this study were 1 m above ground level. The Pm10 air sampler was 1.6 m above ground whilst the large particle impaction rod sampler and the BNFL HVAS was 1 m above ground level. Particle deposition was also derived for total (wet + dry ) processes and these were determined to assess differences in dry and total particle deposition velocities. There are a number of limitations in using the ratio of downward flux divided by the air concentration as a means of expressing dry particle deposition velocity. These limitations relate to non-representative air sampling of the ambient aerosol due to ill-defined inlet collection efficiencies for large particles greater than 10 um aerodynamic diameter (Nicholson 1988). Secondly, it is not clear if the material collected by air samplers is representative of particles in suspension or whether this material has been resuspended off the surface by wind stress or other means such as mechanical intrusion (Sehmel 1980).
A detailed chronology of atmospheric discharges from a number of stacks within Sellafield between 1951 - 1992 can be found in Gray et al (1995). Data relating to the nature and size distribution of emitted material is limited and it is assumed that the High Efficiency Particulate in Air (HEPA) filters give a decontamination factor of at least 100 at the most penetrating particle size of 0.3 um (Fulker pers. comm. 1994).
The main source of plutonium emissions from 1952-1965 was B204 stack and some of these waste-streams were unfiltered. Past releases of plutonium emissions from unfiltered waste-streams will have different particle size distributions prior to discharge compared to the majority of present day releases. It is clear that the particle size distribution of emitted material will influence its atmospheric behaviour and transport. Early 137Cs discharges were dominated by B38 stack emissions from the decanning areas of the Magnox swarf silos.
The evidence produced at the Black Inquiry which was set up to look at the increased incidence of leukaemia in the Sellafield area (NRPB R171 1984) detailed some incidents involving abnormal releases of plutonium at the Sellafield reprocessing plant.
(a) **October 1957 - Windscale fire in Pile no.1
(b) June 1960 - Cooling tower plutonium release
(c) July l969 - High activity liquor release
(d) September 1979 - Atmospheric plutonium release from accidental spillage
Other sources of plutonium isotopes which may not be typical of current Magnox reprocessing include radioactive material which is ejected off the surfaces of the Piles and Magnox cooling ponds (Felstead & Woollam 1979) and possibly material which may become available for re-entrainment during decommissioning of the AGR, the Windscale Piles and older remnants of the original plutonium reprocessing plant, B204. A further environmental source of historic discharges relate to the sea-to-air to-land transfer of marine effluents (Eakins et al 1981 :Eakins et al 1982 :Pattenden et al 1989).