Two types of fuel are used in modern, commercial power reactors. Both are oxide materials, which are well suited to the demanding heat and radiation environment in the reactor. The more common is uranium, as UO2, with the 235U isotopic content typically enriched from the natural abundance to between 3 and 5 atom%. Occasionally, higher levels of enrichment, up to ca. 9 atom%, are used. Some reactors, notably in Japan, France and Switzerland, are par­tially loaded with mixed U/Pu oxide fuel (MOX). This typically comprises uranium with low 235U content, either the depleted uranium byproduct of enrichment, or from recycling of used (‘‘spent’’) uranium fuel, blended with plutonium to enhance its reactivity. MOX typically comprises a blend of 7% Pu/93% U. A modern pressurised water reactor (PWR) can accommodate core loadings of up to ca. 1/3 MOX without significant design implications.

Mayak, formerly known as Chelyabinsk-40 and later as Chelyabinsk-65, is one of the biggest nuclear facilities in the Russian Federation housing the former Soviet Union’s first industrial nuclear reactors. The facility was responsible for producing the material for the country’s first atomic bomb beginning in 1948.

Between the commencement of operations in 1948 through to September 1951, 78 million m3 of high-level nuclear waste containing a total of 1017 Bq of beta activity was discharged from the radiochemical plant directly into the Techa River six kilometres below its source.20 A radiation survey in 1951, revealed extensive contamination of the floodplain and bed of the Techa River and consequently excessive exposure to the inhabitants of the region. Of the total radioactivity discharged into the Techa, 99% was deposited in the first 35 km downstream. Four reservoirs were created along the Techa below Lake Kzyzyltash to isolate the most contaminated water. The final reservoir was completed in 1964 and including Lake Kyzyltash, now contains a volume of 380 million m3 and about 7141 TBq of 90Sr and 137Cs.20

Discharge of diluted HLW directly into the Techa stopped in 1951 and instead was diverted into Lake Karachay. In 1953, an intermediate waste sto­rage tank was put into operation but the excess supernatant (containing much of the caesium waste) was still discharged into the lake (29.6 PBq was added in 1992). A hot summer, followed by a dry winter in 1967, caused evaporation of

the lake and dust from the shoreline containing around 20 TBq of 90Sr and 137Cs (in a 1 : 3 ratio) was blown over an area of 1800 km2 and up to a distance of 75 km.21 Up to four million m3 of contaminated groundwater, containing in excess of 185 TBq has migrated 2.5 to 3 km away from Lake Karachay.20

As previously discussed, a waste storage facility became operational in 1953 and consisted of 20 stainless steel tanks utilising an external cooling system involving water flowing through a gap between the tank walls. The cooling system for one of these tanks failed and resulted in the evaporation, heat-up and ultimately explosion of the 70-80 tonnes of highly radioactive nitrate — acetate waste.22 Around 740 PBq of activity was ejected from the explosion with approximately 90% falling out in the immediate vicinity of the accident and the remaining 10% forming a cloud extending to a height of 1 km. The fallout from this cloud exposed the Chelyabinsk, Sverdlovsk, and Tyumen regions to contamination.22 In the immediate area, 1 to 2 km long by 0.5 to 1km wide, the soil contamination amounted to 5180 TBq km-2 with con­tamination of the wider area, 75 km long and 7 km wide, amounting to around 1 TBqkm-2.21

Planning for the UK GDF relies upon the accurate assessment of the types and quantities of HAWs destined for geodisposal. In response, the UK Government published a ‘baseline inventory’ of radioactive wastes in the 2008 MRWS White Paper.1 The baseline inventory (see Table 2), is underpinned by the 2007 UK Radioactive Waste Inventory2 and includes assessments of the likely volumes, activities, and material contributions of (i) current HAWs, (ii) future HAWs, and (iii) radioactive materials that are not currently classified as wastes (e. g. spent fuels and uranium and plutonium stockpiles). Accordingly, the baseline inventory defines the total volume (after packaging and conditioning) of geo­disposal managed waste as ~480 000 m3 and the activity ~8.7 x 1019 Bq. Of this, a minor volume (~0.3%) but a significant component of the activity (~41%) is designated as HLW (see Table 2). Spent fuels (that are currently not contracted for reprocessing) also comprise a minor volume (2.3%) but major activity (~52%) proportion of the inventory; however, along with HLW estimates, these figures should be considered indicative as the extent and life­span of current reactors and spent fuel reprocessing facilities is uncertain and the treatment of spent fuel may change (see section 3.3.5). With these caveats in mind, the estimated volume of ILW destined for geological disposal is 364 000 m3 (~76.3% of the total inventory volume), with an activity of 2.2 x 1018 Bq (2.5% of the total inventory radioactivity) (see Table 2). LLW is estimated to constitute ~3.6% of the potential waste volume, and <0.1% of the radio­activity (see Table 2). Lastly, the UK GDF generic design assessment must

consider the inclusion of radioactive wastes from any new build nuclear power stations.1,15 However, the current baseline inventory does not include estimates of radioactive wastes arising from any new nuclear build program. Clearly, the development and operation of new UK reactors will lead to significant amounts of waste, and the government is thus committed to updating the baseline inventory as the GDF planning and any new nuclear build programmes pro­ceed.[38] It is important to note that the HAWs (including the spent fuel) from the potential reactor designs submitted to the evolving new build program are consistent with a geodisposal management pathway.16,17

3.3.2 Conditioning and Packaging of Geodisposal Wastes Conditioning is the immobilisation of radioactive waste in a suitable medium to produce a stable or solid waste form usually within a packaged container. The containerised waste can then be stored for several decades and ultimately will be disposed as the GDF is implemented. UK conditioning practices vary according to waste type:

vitrification plant. Here, the HAL is initially calcinated to remove water and nitrate, and then heated with crushed borosilicate glass to form molten glass.3 The molten glass is then encapsulated and cooled in 150 litre stainless steel containers. This vitrified waste constitutes the high level waste form destined for geodisposal, but it is important to note that fuel reprocessing gives rise to large quantities of ILW (see section 2.3). As of 2007, approximately one third of existing UK HLW had been conditioned2 and there is a commitment to reprocess all MAGNOX type fuel prior to closure of the Magnox reprocessing facility in ~2016 (ref. 18). The lifespan of the THORP reprocessing plant and hence the extent of oxide fuel reprocessing is uncertain; however, the NDA is con­tractually committed to reprocess approximately half of the UKs spent oxide fuel.18 The fate of the remaining spent fuel is currently under consideration, but the NDA expect that a proportion may not undergo reprocessing and thus will be directly managed via geodisposal.18 Finally, spent fuel from Sizewell B and from potential new build reac­tors, is currently considered unlikely to undergo reprocessing.1,18

(ii) Spent fuel: The eventual outcome of the UK policy toward oxide fuel reprocessing (see above), and decisions regarding any UK new nuclear build, will ultimately determine the quantity of spent fuel that is destined to be managed via geodisposal. As the UK does not currently declare spent fuel as waste, a bespoke UK geodisposal conditioning concept is not available. However, international best practice (see section 3.3.2 and Table 1) will likely inform future UK spent fuel strategies; indeed, the NDA explicitly consider the Swiss NAGRA SF disposal concept (see Table 1) in their latest geodisposal implementation document.14

(iii) ILW: For UK ILW waste forms (see section 2.3 and Figure 1), immo­bilisation in a cement based matrix and encapsulation in steel drums is the standard (existing) conditioning approach2 and this reflects inter­national practice (see Table 1). To date, approximately one fifth of UK ILW has been conditioned in this way2 and pragmatically, the existence of these waste packages will likely inform future conditioning practices and UK GDF design; for example, a cavern system (see Table 1) is likely to be necessary to accommodate the large ILW volume, further, che­mical compatibility issues between cementitous ILW and likely HLW/ SF waste forms and buffer materials must also be considered. It is also important to note that a small but significant quantity of UK ILW has been conditioned in an organic polymer matrix2 and that graphite from AGR reactors is a large volume, difficult ILW material.

The ICRP outlined its framework for radiological protection of the environ­ment in ICRP Publication 108 and described its use of Reference Animals and Plants.4 Within Publication 108, literature review and expert judgement have been used to produce Derived Consideration Reference Levels (DCRLs) for each of the RAPs (see Figure 2). The DCRLs are defined as a band of dose rate (in mGyd-1) within which there is likely to be some chance of deleterious effects of ionising radiation occurring to individuals of that type of RAP

(derived from a knowledge of expected biological effects for that type of organism) that, when considered together with other relevant information, can be used as a point of reference to optimise the level of effort expended on environmental protection, dependent upon the overall management objectives and the relevant exposure situation. The DCRLS refer to additional dose rates above that from exposure to background radionuclides (mean background dose rates from 238U and 232Th series radionuclides and 40K are typically below 2 mGyh-1 for all of the ICRPs RAPs).57,58 The ICRP is now working on guidance to show how to use the DCRLs in actual assessments.

The DCRLs varies for the RAP reflecting the differing radiosensitivities. In general, mammals, birds and pine trees are more sensitive to radiation exposure than other organisms.

In practice, Purex, like any chemical separation process, generates multiple waste streams ranging in activity from high level liquid waste to trace active washings. Solid wastes are also produced, comprising fuel cladding material (e. g. stainless steel, graphite and zirconium), as well as contaminated process plant components and miscellaneous industrial waste. Prior to reprocessing, the irradiated fuel is stored for several years (in the UK, usually in water-filled ponds) so the pond waters also form low level effluents.

Preparation for separation involves shearing and dissolution of fuel, which will lead to the release of volatile radionuclides, principally 3H, 14C, 106Ru and 129I. These may be trapped for subsequent immobilisation and disposal, or they may be released. Thus, nuclear fuel reprocessing produces a diverse range of solid, liquid and gaseous waste streams, all of which will impact on the environment; for example, the current UK Radioactive Waste Inventory identifies over 180 different waste streams currently being produced at Sella — field.8 All of these waste streams have potential environmental impacts, and all have to be managed to ensure that their impacts are tolerable.

A number of the remediation methods described above have been utilised in various field tests at sites suffering with radionuclide contamination. Three of these case studies are discussed below.

2.2 Hanford Case Study

Bench — and field-scale studies were performed by the Pacific Northwest National Laboratory along with a number of collaborators in order to test the remediation potential of using polyphosphate injections to reduce uranium concentrations in groundwater beneath the contaminated 300 Area of the Hanford Site. A detailed study is provided in a number of PNNL reports105 107 and will be summarised briefly in this section. The concept of polyphosphate injections works by the formation of stable and insoluble uranium phosphate minerals (autunite) and phosphate precipitates (apatite) for uranium sorp­tion.71,108 As autunite sequesters uranium as U(vi) rather than reducing it to U(iv), the issue of re-oxidation and consequent remobilisation is nullified offering a potential advantage over bioreduction methods. Phosphate minerals precipitate when phosphate-containing compounds degrade in water, due to hydrolysis, and hence rapid mineral formation can occur in an aquifer resulting in a reduction in permeability. However, the longer the phosphate chain, the slower the hydrolysis and consequently the use of long-chain polyphosphate compounds results in a lower change in hydraulic conductivity.109

The test site chosen for the field scale study was located in the 300 Area of the site and involved a three-stage approach to the polyphosphate injections. Water was routed from an extraction well located 190 m from the injection well. Sampling pumps were installed in all site monitoring wells, capable of deli­vering flows up to 7.57 litres per minute. The sample tubing from these wells was routed directly into a mobile laboratory and connected to a sampling manifold which monitored field parameters (Eh, pH, temperature and dissolved oxygen) and collected samples for anion, cation and trace metal analysis.

Based on previous laboratory studies summarised by Vermeul,105 a three — phase injection strategy was identified in order to generate both the uranium­bearing autunite and uranium sorping apatite. An initial injection of poly­phosphate was delivered to the subsurface to initiate the formation of autunite, followed directly by an injection of calcium chloride to allow the formation of calcium phosphate, apatite. The process was concluded with a final injection of polyphosphate following on directly from the CaCl2 injection. A contribution of 25% orthophosphate, 25% pyrophosphate, and 50% tripolyphosphate made up the phosphorous in each polyphosphate injection.

Formation of apatite is affected by the mixing time between the polyphosphate and calcium species which proved to be variable throughout the site. Phosphate data indicated that wells in a radial distance of 23 m from the injection site received between 40% and 60% of the injection con­centration. This suggests that a relatively large lateral area could be treated via the formation of autunite although apatite formation may be problematic.

Uranium concentrations in the targeted treatment zone were typically between 60 and 80 mg l-1 prior to the injections. Uranium monitoring data from wells inside the target area showed an initial decrease in the concentration of uranium, to below the drinking water standard of 30 mgl-1, but a significant rebound was observed about two months after the treatment. At a well outside the treatment zone, uranium concentrations were not observed to decrease until one month after treatment and then displayed a slower rebound. This suggests that uranium concentrations were effectively decreased through the formation of uranyl-phosphate mineral phases (autunite) and were then recharged by the uranium plume on site. However, it is also possible that the uranium decrease was partially due to displacement by the injections of large volumes of high ionic strength solutions. It was thought that long-term remediation would occur via the sequestration of uranium through adsorption to apatite and subsequent conversion to stable uranyl-phosphates (autunite) but these data suggest that this is not the case at this site.

Retardation of radionuclides in the subsurface primarily results from their interactions with minerals.61 Sorption of radionuclides to mineral surfaces is controlled by the structure and charge at the mineral surface.42 There are a number of mechanisms by which radionuclides can interact with the mineral phase, including ion exchange, chemisorption and physisorption.45 With ion exchange, the sorbing ion exchanges with another similarly charged ion within the mineral structure. For Cs1, cation exchange with clay minerals is one of the key processes controlling the mobility of the ion.44. When the sorbing cation binds through covalent bonding to the mineral surface, forming inner sphere complexes, the reaction is termed chemisorptions.45,66 Bonding occurs only at specific sites on the mineral surface and the strength of binding depends on the metal ion. For example U(VI) can bind to iron oxyhydroxides (goethite and hydrated ferric oxide) and micas (muscovite and chlorite) through the for­mation of inner sphere complexes at the mineral surface.66 68 Radionuclides also interact with the mineral surface through weak van der Waals’ forces (physisorption),69 forming an outer sphere complex. Such bonding is relatively weak and so radionuclides are readily desorbed from the surface by small variations in the geochemistry of the environment. This interaction is typical of Sr21 with many mineral surfaces including ferrihydrite,66 bacteriogenic iron oxides,70 kaolinite71 and calcite.72

Irregularities on the mineral surface, such as kink and step sites or etch pits, tend to be more reactive than other parts of the crystal surface and may therefore be the preferred site of adsorption66 or microbially mediated dis­solution of the mineral and any related processes.73,74 The steric environment of the adsorption site in combination with the chemical composition of the ligands at the surface affects the affinity of the site for particular radionuclides. Micaceous minerals like illite,75 montmorillonite and vermiculite76 and biotite73 have a high affinity for Cs1 at their edge and step sites. Caesium(I) can diffuse into the interlayers of these sheet silicate minerals over time, making the adsorption irreversible.77

Sorption processes are affected by the biogeochemistry of the solute.61,78 There is typically a strong dependence on pH, with most mineral surfaces being most efficient sorbents at circumneutral pH.79 In more acidic environments, the large concentration of H1 ions in solution causes protonation of mineral surfaces, altering their overall surface charge. In alkaline environments com — plexation of cationic radionuclides by hydroxyl ions (OH ) decreases the positive charge of the cation. Both of these effects lower the electrostatic affinity between the radionuclide and mineral surface and so reduce sorption.

The ionic strength and cation concentration in the solution phase will also affect adsorption to mineral surfaces. Increasing ionic strength will reduce the activities of the radionuclides in solution and will also alter the effective charge at the mineral surface, in both cases decreasing complexation to the mineral surface.62 Changes to the ionic strength of the solution phase can also cause desorption of radionuclides. Standring et al.9 investigated the remobilisation potential of 137Cs, 60Co, 99Tc and 90Sr associated with sediments within Reservoir 10 along the Techa River system at the Mayak site. The sorption of caesium and technetium to sediments within the Techa River system was found to be relatively irreversible, however significant proportions of the sediment — bound strontium and cobalt could be remobilised upon mixing with fresh water or seawater. The desorption effect was significantly increased in the presence of seawater; this effect was attributed to the higher pH and ionic strength of the seawater.

Sorption properties may be modified by the presence of coatings on the surface of the mineral but the effect will depend on the radionuclide.78 Adding an aluminium coating to illite, kaolinite and montmorillonite surfaces enhanced Sr21 sorption compared to uncoated surfaces.80 Coatings of the minerals with humic substances did not affect Sr21 sorption significantly, but decreased Cs1 sorption compared to uncoated surfaces, and the magnitude of the effect was different for each clay mineral (strongest for illite and weakest for kaolinite), reflecting the specificity of adsorption systems.80 Biological coatings, such as bacterial biofilms, on mineral surfaces may also modify the reactivity of the mineral surfaces by masking existing sorption sites. Anderson et al. (2007; ref. 81) found that the presence of a biofilm over a granite ‘‘fracture surface’’ reduced adsorption of Am(iii), Pu(iv) and Np(v) and interpreted this as the biofilm preventing adsorption by decreasing the diffusion of the radionuclides near the mineral surface.

In recent years, radioactivity in the environment has come from several sources. These include natural radiation, residues from the Chernobyl accident and from the atmospheric testing of weapons, plus radioactive discharges and emissions from nuclear and non-nuclear sites (so-called “authorised pre­mises’’). Nuclear licensed sites, which are subject to the Nuclear Installations Act21 may also be authorised to dispose of radioactive wastes under the Radioactive Substances Act.22 These discharges are primarily liquid, and made into rivers, estuaries or coastal waters. Discharges of radioactive wastes from other sites, such as hospitals, industrial sites and research establishments, are also regulated under this Act but are not subject to the Nuclear Installations Act. Small amounts of very low level solid radioactive waste are routinely disposed of from some non-nuclear sites, and there is also a significant radiological impact due to the legacy of past discharges of radionuclides from non-nuclear industrial activity in the UK. These involve radionuclides that also occur naturally in the environment. Discharges from terrestrial non-nuclear sites are generally considered insignificant, and as such environmental mon­itoring of their effects is usually not required for the purposes of protection of public health in the UK. This situation is, however, reviewed from time to time.

The discharge limits are set through an authorisation assessment process which can be initiated by either the operator or the relevant environment agency. In support of the assessment process, prospective assessments of doses to the public are made assuming discharges are kept within the authorised limits. Authorisations are then set so that doses to the public from the site will be below the dose constraint of 0.3 mSv per year (or 0.5 mSv per year if discharges occurred actually at the authorised limits) for that source — the dose limit for the public from all sources being 1 mSv per year.

The Environment Agencies set limits and regulate the discharges and emissions of radioactive waste from authorised premises. Operators of nuclear sites are required both to monitor their discharges and the effects on the environment. In England, Wales and Northern Ireland, the Food Standards Agency, the Environment Agency and the Northern Ireland Environment Agency conduct their own monitoring programmes, whereas in Scotland the Scottish Environment Protection Agency incorporates the requirements of the Food Standards Agency within its own programme. These programmes are important because they provide an independent assessment of the potential harm resulting from authorised releases of radioactive discharges, and act as an additional check to the monitoring programmes conducted by site operators.

The assessments are based on a collection of data relating to the radionuclide concentrations of foodstuffs, external dose rates and information on the habits of people living near the sites. Changes in doses received do occur from year to year, usually because of variations in concentrations of radionuclides in food and in the external dose rates, but in some years doses are affected by changes in people’s habits, in particular their consumption of food, which are identified by carrying out regular food habits’ surveys.

In recent years, a group of people in Cumbria that consume a large amount of fish and shellfish have received the highest dose of radiation due to discharges from two different sources. Their dose was estimated to be 0.52 mSv in 2007 (ref. 23). This was due to the effects of authorized current and past liquid discharges from the reprocessing plant at Sellafield into the Irish Sea, and from past liquid discharges from a phosphate processing plant at Whitehaven a few miles up the coast from Sellafield. The Sellafield discharges were estimated to have contributed 0.24 mSv to this dose in 2007, primarily due to the accumulation of caesium-137, plutonium isotopes and americium-241 in seafood, from past liquid discharges, as well as external exposure from contaminated sediment. The phosphate plant’s discharges (of what are known as “technologically enhanced naturally-occurring radioactive material’’, where there is an increase in con­centrations of some radionuclides that occur naturally due to industrial operations) resulted in the people who consumed seafood also receiving

0. 28 mSv from that source. This was due to polonium-210 concentrations in seafood, which occur naturally anyway, but which also partly arise from the decay of radium-226 and lead-210 in past discharges from the phosphate plant.

Doses to people who had consumed crops grown on land fertilised by seaweed from around Sellafield were also assessed and their estimated dose for 2007 was 0.012 mSv. Doses to people using the beaches and other intertidal areas in the vicinity were less than 0.02 mSv.

People living around operating nuclear reactors generating electricity within the UK receive doses that are typically less than 0.1 mSv per year in 2007. Such low doses often then raise the question of how they compare with natural background radiation. In doing so, however, it must be recalled that radiological protection is based on the premise that an increment in dose results in an increase in risk, the increment being on top of whatever the existing dose rate may be. Nevertheless, it is sometimes useful to note that the background dose rate in the UK can typically vary from about 1.5 to 7.5 mSv per year, with an average of about 2.2 mSv per year, the variation being due primarily to radon. Where radon exposure levels in homes are high, action to reduce them is encouraged. For comparison, it may also be noted that a typical single chest X-ray would give a dose of about 0.02 mSv and a chest CT scan about 8 mSv.

Nuclear power depends upon a finite resource: uranium. Uranium is ubiquitous on the Earth. It is a metal approximately as common as tin or zinc, and it is a constituent of most rocks.97 Central to assessing to what extent the current expansion of nuclear power is sustainable is an assessment of the given reserves of uranium. Current usage is about 68 000 tonnes of uranium per year, with current resources of uranium estimated at 5.4Mt. At current rates of consumption this will last 80 years. However, further exploration and improvements in extraction technology are likely to at least double this estimate over time, particularly as rising prices incentivise mining firms to increase exploration. Uranium exploration has occurred in cycles: initially driven by military needs from 1945 to 1958, then by the needs of civilian plants between 1974 and 1983, with next to no uranium exploration occurring between 1985 and 2003. As a result of growing global demand due to new build programmes in Asia, a new exploration cycle is underway with uranium mining companies embarked on renewed exploratory work, drawing upon new technology and new geological analyses.98,xxxix As a result of this increase in investment and exploration between 2005 and 2006, the world’s known uranium reserves increased by 17%. Moreover:

“The price of a mineral commodity also directly determines the amount of known resources which are economically extractable. On the basis of ana­logies with other metal minerals, a doubling of price from present levels

mx“In the third uranium exploration cycle from 2003 to the end of 2009, about US$ 5.75 billion was spent on uranium exploration and deposit delineation on over 600 projects’’.100

could be expected to create about a ten-fold increase in measured economic resources, over time, due both to increased exploration and the reclassifi­cation of resources regarding what is economically recoverable ’’.99

The Australian Uranium Information Center suggests that if the price of uranium were to double we could expect to see a ten-fold increase in known resources: that is an increase from 3 to 30 million tonnes.101 At various times the nuclear industry has used concerns about the long-term availability of fissile material to justify the development of Fast Breeder Technology. This type of reactor can be started up on plutonium derived from the spent fuel from conventional reactors and operated in closed circuit with its reprocessing plant. Such a reactor, supplied with natural or depleted uranium for its ‘‘fertile blanket’’, can be operated so that each tonne of ore yields 60 times more energy than in a conventional reactor. Breeder reactors could match today’s nuclear output for 30 000 years. It is estimated that electricity from FBRs would cost around three times the amount per kilowatt as that from conventional nuclear power plants; a great deal of investment in R & D is necessary before com­mercialization is a possibility. India, along with a number of other countries, is currently undertaking extensive R & D in this area. However, this cannot be relied upon in either the near or medium term. xl Moreover, a recent report from MIT suggested that the rationale for breeders was based on an out-of-date understanding of uranium scarcity, given that it would take an LWR 30 years to provide the plutonium to start one such breeder reactor which had proven uneconomic.

The report suggests that a more effective and efficient plant would be an enriched uranium-initiated breeder (with a unitary conversion rate). In this design, natural or depleted uranium could be added to the reactor core at the same rate enriched uranium is burned up, and producing no excess nuclear material. This, suggests a recent report from MIT ‘‘is a much simpler and more efficient self-sustaining fuel cycle’’. They suggest there is plenty of uranium to sustain even the most optimistic worldwide nuclear power scenarios ‘‘for much of this century at least’’.103

There are additional technical solutions. For example, using more enrich­ment work could reduce the uranium needs of LWRs by as much as 30 percent per tonne of low enriched uranium (LEU). And separating plutonium and uranium from spent LEU and using them to make fresh fuel could reduce requirements by another 30 percent. Taking both steps would cut the uranium requirements of an LWR in half.

Known reserves of uranium are found in relatively stable industrialised countries (Australia 23%), Kazakhstan (15%), Russia (10%), Canada (8%), South Africa (8%) the USA (6%). As Montgomery indicates, ‘‘This distribu­tion may not please everyone in New Delhi or Qinshan, but it does at least promise a reliable supply from stable nations’’.104 Regardless of whether the countries are ‘‘friendly’’, the very fact that countries have to import uranium

xlThe extraction of uranium from seawater would create 4.5 billion tonnes of uranium — a 60000- year supply at present rates but it is currently uneconomic.102

deals a blow to claims that nuclear energy enables one to achieve “energy independence”, particularly for countries like France where energy indepen­dence was the primary rationale for the headlong rush into new nuclear during the 1970s and 1980s.

The investment bank RBC Capital has recently said that the uranium market has moved from oversupply to undersupply in just a few months as China has begun to purchase long-term supplies for new reactors. The rise in price is also due to reports in the Chinese media that the country is going to build 60% more megawatts of nuclear power by 2020 than previously thought. xl1 Some countries face more difficulties than others in sourcing uranium, for example under a treaty with Russia, the United States currently receive 40% of its reactor fuel from decommissioned Russian nuclear weapons. The treaty runs out in 2013. Uranium mines only provide two thirds of worldwide uranium requirements with the rest coming from military sources (stores of uranium/plutonium and uranium, from decommissioned nuclear missiles).

Initially, radiation doses to people were primarily from external gamma radiation, but after nine months, ingestion of contaminated products became

Figure 3 Declines in total radioactivity in the environment during the years after the Kyshtym and Chernobyl accidents. (From data in Nikipelov et al.18 and Smith and Beresford).48

the most important dose pathway, as relatively short-lived beta/gamma emit­ters decayed away and (the pure-beta emitting) 90Sr began to dominate.17 In the first years after the accident, ingestion of 90Sr was primarily due to consump­tion of contaminated bread and milk as well as water from local reservoirs.17 It was also reported17 that, eight years after the accident, foodstuffs contributed to ingestion doses in the following order: milk (50%); vegetables (15%), potatoes, 12%, eggs (8%), meat (7%) and bread (4%). Note that these values are likely to be strongly influenced by countermeasures to control and reduce activity concentrations in foodstuffs. Since strontium has a similar uptake chemistry to calcium (both being alkaline earth metals), products rich in calcium tended to have high 90Sr concentrations. In the human body, strontium is incorporated into bone and teeth, so doses to the bone marrow can be high.

In 1987, thirty years after the accident, it was reported,17,18 that the 90Sr content in farm products had declined significantly compared to one year after the accident. Following atmospheric nuclear weapons testing, radiostrontium activity concentrations in plants were observed typically to decline with an effective half life in the range 8-14 years19,20 due to physical decay, loss of 90Sr from the upper soil layers and changes in soil adsorption. At Kyshtym, in addition to these natural processes, Alexakhin et al.11 report an extensive system of countermeasures which played an important role in reducing 90Sr activity concentrations in foodstuffs. These countermeasures included deep — ploughing of contaminated land to reduce radioactivity in the rooting layer of plants, liming of acid soils and addition of calcium to the feed of any calcium — deficient farm animals.

The radiation dose to the population varied greatly, being dependent to a large extent on the contamination density. Average doses (received before evacuation) were 520 mSv to the 600 people living within the 18.5 MBq m 2 zone who were evacuated within 20 days. The 3100 people living in areas of contamination density of 0.12-0.33 MBq m 2 received 23mSv in the 610 days before their evacuation was completed.18 In areas with a 90Sr contamination density of 0.031 MBq m 2 which were not evacuated, Alexakhin et al.11 reported an effective equivalent dose over a 30 year period of 12mSv. Relatively recently, it was reported21 that ‘‘At present, doses to members of the public living in the area of the East-Ural radioactive trace are significantly lower than 1 mSv/year’’.

The overall average radiation dose to the 210 000 people who were living within the EURT at the time of the accident was reported to be 1 mSv, with a collective dose of approximately 2000 person-Sv.22 Assuming a fatal cancer risk factor of 0.05 Sv 1, this implies an average individual risk of ca. 1 in 3000 and an expected 100 fatal cancers in the affected population.

The doses and dose rates to the population were therefore likely to give rise to a significant increase in fatal cancers, though owing to the relatively small number of people exposed to the highest doses and incomplete follow-up, this may not have been possible to observe in the epidemiological evidence. A summary of this evidence18 suggests no observable increases in deterministic symptoms (radiation sickness) or serious illness, though a reduction in leukocyte blood count was observed in the exposed population. The radiation doses to the exposed population at Kyshtym (at least, the reported average dose to the most exposed people) was significantly below lethal levels, though it is possible that some deterministic effects would have been seen in some of the most exposed of this group. Kossenko22 summarised the epidemiological data as follows:

“The cancer mortality rate was estimated for 7800 people exposed due to the Kyshtym accident and moved from the EURT territory, and for 8000 people left to reside in that area. In the four groups studied, with doses averaging from 0.54 to 0.006 Gy [presumably approximately equal to 0.54 to 0.006 Sv for this beta/gamma radiation], no increase in the incidence of lethal leukaemia and solid cancers was found. In addition, the investigation did not reveal disorders of the reproductive function in the exposed people (according to birth rate data) nor any increase in the incidence of devel­opmental defects among the offspring of people exposed to radiation due to the Kyshtym accident’’.

Workers at the Kyshtym site during and after the accident received higher doses than the population. According to Kruglov (2002), 5000 workers at the site at the time of the accident received doses up to 1 Sv and 30,000 clean-up workers received doses above 250 mSv during 1957—59.74

There were a number of long-term studies on the effects of Kyshtym on biota. In the most contaminated ‘‘head’’ of the EURT, acute radiation syndrome was observed in farm animals, in many cases leading to death. Within 12 days of the release, estimated whole-body doses to some farm animals reached 2.9 Gy and doses to the gastrointestinal tract (GIT) were in the region 20-50 Gy.17

Longer term studies also showed evidence of effects on organisms, though this evidence is not always conclusive and radiation exposures to animals were not always estimated, making interpretation difficult. A study on dandelion seeds in 1991 and 1993 (ref. 23) showed some increase in chromosome aber­rations in contaminated compared to control areas, though this was not con­sistently observed. Germination and viability did not differ between contaminated and control areas.23 In studies conducted from 1961-1991, Krivolutsky24 observed that the number of species of soil-dwelling insects was lower by a factor two in a contaminated plot of birch forest compared to two control plots, but no significant difference was found between a contaminated agricultural system compared to a control one.

In aquatic systems, dose rates to benthic (bottom-dwelling) fish in the highly contaminated Uruskul’ and Berdyanish lakes were up to 0.1 Gy d 1 during the first few months after the accident.25 There were reported25 “disturbances in the reproductive process of carp and goldfish… [which] manifested itself predominantly in the first few years following the accident’’. No effect on the number of fish species was found, attributed by the authors25 to ‘‘the fact that the radiation effect was compensated for by a ban on commercial catches of fish’’.