In the UK, HLWs are defined as materials that are heat generating due to their radioactivity, and this heat generation needs to be considered during the design of storage and disposal facilities.2 HLW in the UK is dominated by materials derived from spent nuclear fuel reprocessing (see Figure 1).2 Prior to reproces­sing, spent fuel is stored for several years in a holding pond to allow short lived radionuclides to undergo radioactive decay and to dissipate the resultant heat. At the Sellafield site, spent fuel from a range of reactor types is either taken out of its cladding or chopped up, dissolved in nitric acid and is chemically processed in the PUREX process, to separate uranium, plutonium and waste fission products. During reprocessing, the fission products are ultimately separated into a solution phase which is known as highly active raffinate (HAR). This fission product laden nitric acid solution contains the bulk of the radioactivity that was associated with the spent fuel. It will continue to generate heat for decades and is dominated by radioisotopes such as 90Sr and 137Cs. The HAR is converted to a

more stable wasteform by glassifying or vitrifying it prior to storage and ulti­mately, disposal.3 As extracted uranium and plutonium may have economic value, they are not currently classified as waste and are instead stored.

Radionuclide transfer to plants in the human food chain is often quantified using concentration ratios for different groups of plants and soil types. A recent IAEA handbook provides compiled concentration ratio values for the human food chain.27,28 There is an assumption that equilibrium exists between the plant and soil, which is not valid if there are large temporal spikes in releases.

Similarly, for estimation of exposure of plants themselves in environmental assessment models, the most common approach is the whole organism con­centration ratio (CRwo), where:

In human food chain models, CR is most usually defined on the basis of plant dry matter activity concentrations. The CR values, categorised by soil type, for human food chain assessments have been collated for many radionuclides in the IAEA handbook.28 More mechanistically based approaches enabling pre­dictions of radionuclides, such as radiocaesium and radiostrontium, which vary with soil properties are also available in some assessment tools for the human food chain. In contrast, current CR values for estimation of plant exposure generally do not distinguish between different soil types.

Whilst root uptake is a key pathway of plant contamination, for radio­nuclides which have a low root uptake, such as plutonium and americium, resuspension and adherence of contaminated soil on plant surfaces can con­stitute a significant proportion of the radionuclide content of plants as sampled in the environment.

As examples, the CR values for the selected radionuclide-wildlife group combinations as anticipated to be reported in the new IAEA handbook on transfer to wildlife are given in Figure 1.

The potential utility of plutonium as a component of MOX fuel depends on its isotopic composition. Thermal irradiation can produce isotopes from 239Pu to 242Pu inclusively (see Table 1). As irradiation time increases, so does the proportion of heavier isotopes, with mass number >239. Only Pu and Pu are fissile in a thermal reactor, so plutonium separated from high burnup fuels will have a lower fissile content than that separated from low burnup fuels. In addition, 241Pu has a relatively short half life so, after storage periods of years to decades, a significant proportion will have decayed to 241Am. This has

two effects. First, the fissile content of the plutonium has decreased and, second, the decay of 241Am is accompanied by a 59.5 keV gamma emission, which makes handling much more difficult because shielded facilities are needed to handle plutonium with a significant americium content. Thus, the Melox plant at Cadarache, which produces MOX fuel, is limited to using plutonium which is less than five years old.

The isotope 238Pu is also formed in irradiated uranium. This is formed by neutron capture in 237Np, itself produced through either an n,2n reaction in 238U, or by successive neutron captures in 235U.

Like the United Kingdom, the United States has accumulated a nuclear legacy from over 60 years of research, production, use and storage of nuclear mate­rials. Over this time frame, nuclear material was produced for use in both civilian power plants and in military weapons. The Department of Energy (DOE)’s 120 sites contain 40 million cubic meters of contaminated soil and debris and 1.7 trillion gallons of contaminated groundwater. Of this, at least 50% is contaminated with radionuclides, including caesium-137, plutonium — 239, strontium-90, technetium-99, uranium-238, and uranium-235, as well as heavy metal contamination including chromium, lead and mercury.23 The associated cleanup cost has been estimated to be in excess of a trillion dollars.24 Contaminated sites include former uranium ore processing facilities such as Rifle, Colorado and Moab, Utah. The clean-up of these sites was tasked to the DOE under the Uranium Mill Tailings Remedial Action (UMTRA). During the years of operation of the Moab site, approximately 10.5 million tons of tailings and contaminated soils accumulated in an unlined pile 750 feet from the Colorado River.25 In 2005, the DOE finalised the remediation strategy to be undertaken at Moab which included active groundwater remediation and off­site disposal of the tailings pile and other contaminated materials at the Crescent Junction disposal site.26 The Rifle site consists of two old uranium processing plants: Rifle Old Processing Site and Rifle New Processing Site. Tailings and tailings-contaminated material from Rifle were transferred to the Rifle disposal site approximately six miles north of the Rifle New Processing Site and surface remediation was completed in October 1996. Contaminants of concern in the groundwater at both sites include arsenic, molybdenum, selenium, nitrate, uranium and vanadium, with contamination at New Rifle extending approximately three miles west of the site. Groundwater remediation is being achieved through natural flushing of the groundwater in conjunction with contaminant monitoring.27 The in situ remediation of uranium was examined in a field scale study in 2003 in which acetate was injected into the subsurface over a three month period in order to stimulate microbial reduction of soluble U(vi) to insoluble U(iv) (ref. 28) and is discussed in detail later in this chapter. The Savannah River Nuclear Facility, South Carolina, was used to refine nuclear material for use in the United States defence program. The site used a system of canals and reservoirs to disperse heat from the reactors and consequently, various ponds connected to this system received cooling water discharges from the reactors. One such pond, Pond B, received discharges containing fission products such as Cs, Sr and Pu. Radionuclide input peaked in 1963 and 1964, believed to be caused by the leakage of fuel elements stored in a water-filled basin in the reactor.29 The vast majority of both 137Cs (98%) and 90Sr (85%) were found to be in the pond sediments.29

A number of other sites located in the United States have a more complicated environmental legacy left by the nuclear industry and are discussed below in more detail.

The planning and construction of a GDF will take several decades. Conse­quently, safe interim storage of HAWs is an integral part of their long term management.1,4 To ensure smooth transfer from storage to disposal facility, current interim storage plans must therefore be updated to meet a design
principle that facilitates the storage of HAWs for 100 years or more in a manner that protects both people and the environment.1 This represents a significant undertaking as large quantities of UK HAWs are stored in aging facilities. In addition, it is noteworthy that the current UK HAW stores are concentrated at the Sellafield site in west Cumbria, thus, any GDF imple­mentation plan will need to consider transport of these materials to the even­tual GDF location.

For radiological environmental risk assessments, the benchmark may be in the form of a dose rate or back-calculated using the available assessment tools to medium environmental concentrations for each radionuclide that would give rise to the predicted no effect dose rate. These environmental concentrations [Environmental Media Concentration Limits (EMCLs) in the ERICA Tool, or Biota Concentration Guides (BCGs) in the USDOE Graded Approach] can be compared directly to measured or model-predicted environmental media concentration values and subsequently used to determine a ‘‘risk quotient’’. Calculated environmental media concentration benchmark values are usually
applied at earlier tiers of a risk assessment for identifying (or screening out) sites where there is negligible risk of potential impact. The assumptions used in the calculation of environmental concentration benchmark values are usually conservative with respect to transfer to the organism, exposure scenario and in some instances geometry.

A risk quotient (RQ) provides a simple means of assessing risk by integrating the exposure and effects data to determine the likelihood of an ecological risk occurring. It is calculated from the quotient of the estimated exposure and a numeric benchmark (in the form of a dose rate or activity concentration). The benchmark dose rate is a dose rate which is assumed to be environmentally ‘‘safe’’. The RQ is defined as:

Where the resulting RQ is less than one, then no further effort or action would normally be required. Where the RQ is greater than one, then the assessment would likely need further work (such as collecting more data, refining the exposure assessment, or taking action to reduce the risk).

There are three methods commonly used to derive numeric criteria in ecotoxicology:

(i) deterministic — based on the application of assessment (or safety) factors to the most restrictive single sensitivity value observed;

(ii) probabilistic — based on Species Sensitivity Distribution (SSD) model­ling; and

(iii) a weight of evidence approach — typically using data from field expo­sures, such as in situ measurements of biodiversity indices co-occurring with stressor levels.

Over the last few years, the first two approaches have been applied to radiological assessment59 63 and are based on the guidance provided by the European Technical Guidance Document (TGD)64 for chemical risk assess­ment. The benchmark produced by both approaches is designed to ensure protection of ecosystem structure and function.

The third method has not been widely used to derive benchmarks for use in radiological assessments of the environment although there are examples for specific sites (e. g. uranium mining).65

The deterministic approach, takes the lowest dose rate observed to give a significant biological effect available for any tested species and divides it by a predefined assessment/safety factor ranging from 10 to 1000 (10000 for marine ecosystems) according to the quality and quantity of the data available. The assessment/safety factor is intended to account for uncertainty and guidance on what value to apply is set out in a technical document supporting EC Directive 93/67/EEC.64

In contrast, the probabilistic approach uses the available (quality-assured) ecotoxicological data to determine the dose rate, giving a 10% effect resulting distribution for chronic exposure in the ecotoxicological data (the so called ‘‘effective dose rate for a 10% effect’’; EDR10). The EDR10 value is used to compensate for the influence of experimental design. For instance, the lowest no effect concentration or the highest no effect concentration may be con­siderably different to the true no effects concentration dependent upon experimental replication. These EDR10 values are then plotted together for all species for which information exists and are used to identify (usually) the fifth percentile from the species sensitivity distribution (SSD). To account for any residual uncertainty, an assessment factor of between 1 and 5 is applied to the fifth percentile value based on the available quality and quantity of the data in the SSD to produce the predicted no effect dose rate. This approach, as applied to radiological assessment, is described in full by Garnier-Laplace et al.60 62

The SSD approach has been used to derive a screening dose rate of 10 pGyh-1 using different data selection criteria;59,60 this value is used as the default screening dose rate in the ERICA Tool.13 The screening dose rate is to be applied to incremental (i. e. above background) exposure. Currently, it has only been possible to derive a generic screening value applied to all ecosystems using this approach due to the lack of appropriate quantitative data across a sufficient number of different wildlife categories. Thus, it is not possible to derive screening values by the SSD approach sub-divided into different wildlife groups due to statistical constraints.

A screening dose rate is for application in regulatory assessments of planned releases and is not a useful benchmark for use in accidental situations such as that ongoing at Fukushima. Consideration of effects on wildlife after the Chernobyl accident with reference to Fukushima is presented by Beresford and Copplestone.70

Many different solvent extraction processes have been devised to produce particular product and waste streams. For example, it may be attractive to avoid separation of a pure plutonium stream in order to limit proliferation risks, or it may be useful to separate long-lived isotopes from high level liquid waste for separate treatment, so that the radiotoxicity of the remaining high level waste decreases more rapidly through decay. In some cases, separations are designed to fit with specific national regulatory requirements and many of these processes could be used in combinations to give particular, desired out­comes. Some examples are given in Table 4.

4 Waste Management Options

The critical decision in the ‘‘back end’’ of the nuclear fuel cycle is whether or not to reprocess (in other words, whether the cycle is closed or open). If an open fuel cycle is chosen (e. g. as in Sweden), then waste management is essentially confined to the management of spent fuel and reactor decommissioning wastes,

with the former dominating the radioactive content of the wastes. Inter­nationally, deep geological disposal, usually preceded by some decades of cooling to limit heat and radiation load on the host rock, is always assumed to be the management route for spent fuel. There is currently no operational disposal facility for spent fuel.

In a closed fuel cycle, as described above, there are many more options for the management of different streams. As a country which has operated a closed nuclear fuel cycle for over 50 years, the UK’s approach to managing these different waste streams is fairly typical. Conventional Purex processes will produce uranium and plutonium product streams, and a liquid high level waste stream, which in current thinking will be vitrified for deep geological disposal. Concepts for disposal of vitrified high level waste are generally quite similar to those for spent fuel, because the heat and radiation loads are similar. However, as outlined above, removal of most of the actinide inventory would allow the hazard from vitrified high level waste to decrease faster than that from spent fuel.

Operations at all stages of either a closed or open fuel cycle will generate lower activity wastes associated with uranium mining, fuel fabrication, energy generation and spent fuel management. These are generally classified on the basis of their radioactive content (in the UK, classification is in decreasing order of radioactivity: intermediate level, low level and very low level wastes; see Chapter 6). In a closed fuel cycle, a wide variety of process wastes, for example 14C — or 85Kr-containing gases from fuel shearing and dissolution, or water from storage ponds, is also created. These are decontaminated where necessary (for example stripping of 14CO2 from gases by precipitation as BaCO3, or removal of 90Sr and 137Cs from aqueous effluents by ion exchange) prior to discharge to the environment under regulatory authorisa­tion. The radioactive wastes from an open fuel cycle tend to be smaller in volume and less diverse than from a closed one. Obviously, the total activity is not changed.

As discussed previously, the Rifle UMTRA site in Colorado is an old uranium processing facility which suffers from various contamination issues, including uranium. Uranium is predominantly found in the mobile U(VI) form in the subsurface due to an insufficient supply of electron donors to stimulate anaerobic respiration and/or consume dissolved oxygen. Laboratory studies have demonstrated the potential of microbes to reduce U(VI) to immobile U(IV) in an aquifer system,110 and the in situ treatment of U(vi) using the same method was tested at the Old Rifle site. Contaminated soil has been removed from the site, leaving only groundwater contamination within the local aquifer. Concentrations of uranium in this area range from 0.4 to 1.4 mM, above the maximum UMTRA contamination limit of 0.18 mM.28

The method used in this field-scale test was describe in detail by Anderson et al.,28 and is summarised briefly here. Injection wells were installed in two rows of ten, perpendicular to groundwater flow (which is typically towards the Colorado River). Each well contained three injection points positioned at different depths in the subsurface. A storage tank was filled periodically with native groundwater and was amended with sodium acetate as an electron donor to stimulate uranium-reducing bacteria and potassium bromide as a conservative tracer at concentrations of 100 and 10 mM, respectively. Oxygen was removed from the groundwater through nitrogen sparging. During operations, the injections were set to provide 1 to 3 ml of the solution from the storage tank per minute corresponding to 1 to 3 mM acetate and 100 to 300 pM bromide per day. Monitoring wells were installed at intervals downgradient corresponding to groundwater travel of approximately 4, 9, and 18 days with a further three wells placed upgradient to serve as controls. Acetate was injected continuously over a three month period from June to October 2002, with groundwater samples collected at regular time intervals from all monitoring wells. Groundwater conditions were monitored, including pH, conductivity, redox potential and dissolved oxygen, with further samples taken for U(VI), anion (bromide, nitrate, and sulfate), Fe(n), sulfide and acetate analysis. A second round of acetate injections were made over the same months in 2003, after which no further amendments were made.111

Bromide, added as a groundwater tracer, was not detected in any of the upgradient wells but was detected after 4, 9, and 18 days at each of the corresponding downgradient wells confirming the injection solution had reached the targeted area. After the first set of injections, U(VI) concentra­tions were observed to decrease 9 days after the injections began with concentrations dropping to or below 0.18 pM within 50 days at some wells.28 The decrease in U(VI) was concurrent with the accumulation of Fe(II) and prior to any sulfate reduction. After 50 days, the U(vi) concentration began to increase, coincident with a decrease of Fe(II) and acetate falling to non-detectable levels. Bromide levels were still detected at wells where acetate levels had fallen suggesting that an increase in consumption of acetate was occurring near the point of injection. This correlated with observations following the second injection stage of a depletion of reducible iron oxide near the injection point and an accumulation of sulfide111 suggesting all the available Fe(III) had been consumed and that sulfate reducers were now actively consuming acetate at the injection point.

A substantial shift in the microbial community was observed throughout the injection trials. Organisms in the family Geobacteraceae (which includes the known U(VI)-reducing genus Geobacter) became dominant early on,28 with the greatest enrichment of Geobacteraceae correlated to the greatest proportion of U(iv) detected.111 As reducible Fe(m) became depleted and sulfide accumulation occurred, the dominance of the Geobacteraceae decreased as they were replaced by species related to known sulfate-reducers.111 After the second round of injections in 2003, U(vi) continued to be removed from the groundwater for over a year after the cessation of acetate injections.112 This casts doubt on the suggestion following on from the first round of injections that U(VI) removal is acetate dependent. Flow-through column experiments suggested that the continued decrease in groundwater U(VI) levels could be linked to increased sorption to soils in a reduced environment.112

This series of field studies suggest that the stimulation of metal-reducing bacteria is an effective method for the removal of U(vi) from groundwater. However, when the supply of reducible Fe(III) oxides runs out, sulfate-reducers become dominant and do not appear to be as effective at reducing U(vi) to U(iv). Promising data from the second round of injections indicates that in sufficiently reduced soils, U(vi) removal may continue, without the continued need for acetate injections, via sorption to soils.

For redox-sensitive radionuclides such as uranium, technetium, neptunium and plutonium, oxidation state is one of the primary controls on mobility, affecting precipitation, complexation, sorption and colloid formation behaviour. The dominant oxidation states for some key radionuclides at pH 7 are shown in Figure 4. Microbial metabolism can drive a wide range of redox transformations, utilising a succession of terminal electron acceptors (TEA), including some redox-active radionuclides, for the oxidation of organic matter. The amount of

energy gained from the use of each TEA influences the rate and sequence of TEA utilisation. The classical TEA sequence is reduction of: O2, NO3 Mn(IV), Fe(iii), SO42 , followed finally by methanogenesis,17 but radionuclides such as uranium and technetium can also be used as TEAs. Under Fe(iii) and SO42 reducing conditions, U(vi) (as UO221) and Tc(vii) (as TcO4 ) can be reduced by a wide range of microorganisms to less mobile U(iv) and Tc(iv), respectively.1,82,83 There have also have been a few limited studies reporting microbial reduction of neptunium and Pu.84 85 The Fe(iii)-reducing bacteria Geobacter sulfurreducens and Shewanella oneidensis have been reported to slowly reduce Pu(iv), as amorphous Pu(OH)4, to Pu(iii); for S. oneidensis, the rate of reduction was increased by the presence of riboflavin as an endogenous redox mediator.84 Shewanella oneidensis and a mixed consortium of sulfate-reducing bacteria have been found to reduced soluble NpO2+ to insoluble Np(iv).85,86

in addition to direct microbial reduction, the oxidation state of redox-active radionuclides will also be influenced by presence of microbially-generated redox-active species, reactive mineral phases and microbial alteration of mineral phases. Redox-active ions exposed at the surface of a mineral, such as sulfur or iron in mackinawite (FeS) can reduce an adsorbed radionuclide such as U(vi) and Tc (vii).87 This changes the controls on subsequent remobilization processes, and makes oxidation the dominant re-suspension pathway rather than the presence of competing cations in solution or pH fluctuations. Livens et al.87 found that the reduced uranium was readily reoxidised and desorbed upon introduction of oxygen. Reduced iron sediments within a soil profile can also immobilise Tc(vii) by similar surface-mediated reduction to Tc(iv).88 in contrast to uranium, however, technetium does not remobilise as readily with oxygen when it is in association with mackinawite.89

Bacteria can reduce transition metals (notably iron and manganese) locked within mineral structures90,91 and this could alter the characteristics of the reactive surface, generate new reactive mineral phases or release redox-active species into solution.15,74 in particular, iron-bearing minerals can play a crucial role as mediators between microbial anaerobic respiration and redox sensitive radionuclides. Ferrous iron released by microbial reduction of iron-bearing phases can react with a number of ligands and phases present in solution to form a range of new iron phases: oxyhydroxides (magnetite, goethite),92 car­bonates (siderite) or phosphates (vivianite)93 and others. The mineral phases formed in the environment can be hard to predict, but are likely to be domi­nated by carbonates and hydroxides due to the abundance of those ligands in solution. Wildung et al.88 investigated technetium reduction in shallow aquifer sediments from the US Atlantic Coastal Plane. The primary control on the reduction of Tc(vii) was the amount of readily extractable (and so more reac­tive) Fe(ii) present in the sediments. Other studies have also found that Tc(Vii) can be reduced to Tc(iV), as TcO2 by biogenic Fe(ii) , with TcO2 associated with the biogenic Fe(ii) mineral phase.94,95 Upon reoxidation of reduced sediments there can be significant reoxidation and remobilization of technetium, but it is dependent on the nature of the oxidant. When sediments are reoxidised with air, a significant (50-80%) fraction of the reduced and immobilised technetium can be reoxidised and remobilised; however, when the oxidant present is nitrate, there is much more limited (o 10%) reoxidation of technetium.96,97

Bioreduction may also lead to dissolution of the mineral phase and loss of sorption sites.98 Bacteria can reduce iron within a number of iron oxyhydr — oxides of varying crystallinity: hematite, goethite, lepidocrocite and schwert — mannite,99,100 and even micas such as biotite,101 smectite102,103 and illite.99,104 This can cause release and remobilization of radionuclides which have been adsorbed or incorporated into the mineral phase. Langley et al. (2009)105 investigated the impact of microbes on strontium sorbed to bacteriogenic iron oxides. Microbial reduction of the ferric iron within the iron oxides remobilised strontium (increased its concentration in solution), most likely due to loss of sorption sites. The authors suggest, however, that in a natural system remo­bilised strontium would be transported upwards by advection and recaptured within newly-formed bacteriogenic iron oxides near the surface of the water body, once again retarding its transport.

Under circumneutral conditions, abiotic or biotic oxidation of Fe(n) or Mn(n) leads to the formation of new oxyhydroxide phases.106,107 These sec­ondary mineral phases are characterised by large surface area and small crystal size, and have very high sorption capacity.107,106 At low concentrations, oxi­dised uranium has been reported to form an inner sphere complex on biomi­neralising manganese oxides.108 At high concentrations U(vi) was sequestered very efficiently and was incorporated into the oxide structure. The large cation caused distortion of the manganese oxide lattice and the formation of a mineral with tunnel-like structures. The results highlight the significance of the solution chemistry during mineral formation and the sequence of biomineralisation processes and the presence of radionuclides in solution.

All of the situations discussed in the preceding section apply to planned ‘nor­mal’ exposures. On rare occasions, however, abnormal situations arise, the most recent being that at Fukushima in Japan. This happened some twenty five years after the Chernobyl accident, which occurred on 26 April 1986 during a low power engineering test of their Unit 4 reactor. At the time of writing, little is known about the events at Fukushima, but it is useful to review what has been learned from Chernobyl.

The Chernobyl site is located in present-day Northern Ukraine, some 20 km south of the border with Belarus and 140 km west of the border with the Russian Federation. The accident was caused by the improper operation of the reactor, which itself had severe design flaws, allowing an uncontrollable power surge to occur. This resulted in successive explosions that severely damaged the reactor building and completely destroyed the reactor. The accident caused the uncontrolled release of large quantities of radioactive substances into the air for about 10 days. The radioactive cloud dispersed over the entire northern hemisphere and deposited substantial amounts of radioactive material. At the site itself, two workers died from injuries, and approximately 600 workers responded within the first day to the immediate emergency, including staff at the plant, firemen, security guards, and staff of the local medical facility. The dominant exposures for these personnel were external irradiation of the whole body at high dose rates, and beta-irradiation of the skin. Internal contamina­tion was of relatively minor importance, and neutron exposure was insignif­icant. As was to be expected, a very considerable effort has since been expended to follow up on the human consequences of this major nuclear disaster, and the latest findings are those of UNSCEAR.24

Cases of acute radiation syndrome (ARS) occurred among the plant employees and so-called ‘first responders’ but not among the evacuated populations or the general population. The diagnosis of ARS was initially considered for 237 persons, based on symptoms of nausea, vomiting and diarrhoea. The diagnosis was confirmed in 134 persons. There were 28 early deaths (first four months), primarily (95%) where whole-body doses were in excess of 6.5 Gy. Underlying bone marrow failure was the main contributor to all deaths during the first two months, in spite of attempts to save them with bone marrow transplants. Skin doses exceeded bone marrow doses by a factor of 10 to 30, and many ARS patients received skin doses in the range of 400-500 Gy. Radiation damage to the skin aggravated other conditions, and this was considered to be a major contributor to at least 19 of the deaths. Such damage significantly increased the severity of the ARS, especially when skin burns exceeded 50% of the body surface area and led to major infections. Since then, 19 ARS survivors have died (up to 2006), but their deaths have been attributed

to various causes, and usually not associated with radiation exposure. Skin injuries and radiation-induced cataracts are, however, major lasting clinical impacts for the ARS survivors.

In 1986 and 1987, some 440 000 recovery operation workers were used at the Chernobyl site, and more ‘recovery workers’ were involved in various activities between 1988-1990. Collectively, about 600 000 persons (civilian and military) received special certificates confirming their status as recovery operation workers (unfortunately also known as ‘‘liquidators’’). About 240 000 were military servicemen. The average effective dose received by these recovery operation workers between 1986-1990, and mainly due to external irradia­tion, is estimated to have been about 120 mSv. The recorded worker doses varied from > 10 mSv too 1 Sv, although about 85% of the recorded doses were in the range 20-500 mSv. (Uncertainties in the individual dose estimates vary from > 50% up to a factor of 5, and the estimates for the military per­sonnel are suspected to be biased towards high values.) To date, there is some evidence of a detectable increase in the incidence of leukemia, primarily based upon results from the Russian Federation, and of cataracts among those who received higher doses, but there is no evidence of other health effects than can be attributed to radiation exposure.

With regard to the public, the number of evacuees was about 115 000, consisting of about 25 000 persons from Belarus, 200 from the Russian Fed­eration and 90 000 from the Ukraine. The areas from which people were evacuated form what is called the ‘‘exclusion zone’’, which includes not only the 30 km zone, which is the area within a 30 km radius centred on the location of the Chernobyl reactor, but also highly-contaminated areas adjacent to the 30 km zone and more distant areas where high levels of radionuclide deposition density were measured.

Two radionuclides, the short-lived iodine-131 (with a half-life of 8 days) and the longer-lived caesium-137 (with a half-life of 30 years), were particularly significant for the radiation dose they delivered to members of the public. In the former Soviet Union the contamination of fresh milk with iodine-131, and the lack of prompt countermeasures, led to high thyroid doses, particularly among children. The thyroid doses received by the evacuees varied according to their age, place of residence, consumption habits, and date of evacuation. For many pre-school children the doses to the thyroid were well in excess of 1 Gy. It is therefore not surprising that there has been a substantial increase in thyroid cancer incidence amongst those exposed as children or adolescents in Belarus, the Russian Federation, and the Ukraine since the Chernobyl accident, and this increase has shown no signs of diminishing (up to 20 years after exposure). Amongst those under the age of 14 years in 1986, 5127 cases (for those under the age of 18 years in 1986, 6848 cases) of thyroid cancer have been reported between 1991-2005 for the whole of Belarus and Ukraine and the four more affected regions of the Russian Federation. By 2005, 15 cases had proved fatal.

In the longer term, mainly due to caesium-137, the general population was also exposed to radiation externally from radioactive deposition and internally from consuming contaminated foodstuffs. The resulting radiation doses were relatively low, however, partly because of the countermeasures taken. Excluding doses to the thyroid, the mean effective doses due to external irra­diation were estimated to have been about 30 mSv for the Belarusian evacuees, about 25 mSv for the Russian evacuees, and about 20 mSv for the Ukrainian evacuees. These values were at least 10 times smaller than the corresponding numerical values of thyroid doses resulting from internal irradiation. The mean effective doses due to internal irradiation were estimated to have been about 6 mSv for the Belarusian evacuees, about 10 mSv for the Ukrainian evacuees, and about 10 mSv for the Russian evacuees. These values were at least half of the corresponding effective doses due to external irradiation.

Among those exposed in utero and as children, no persuasive evidence has apparently accrued to suggest that there is a measurable increase in the risk of leukemia due to radiation exposure. This is not unreasonable, because the doses involved were generally very small, and therefore epidemiological studies would lack sufficient statistical power to observe any effect, had there been one. Overall, therefore, the average effective doses, due to both external and internal exposures, received by members of the public during 1986-2005 were about 30 mSv for the evacuees, 1 mSv for the residents of the former Soviet Union, and 0.3 mSv for the populations of the rest of Europe.

More recently, on 11 March 2011, an earthquake and accompanying tsunami struck the coastal area of Japan and caused major damage to the Fukushima Dai-ichi nuclear power plant, which consists of six boiling water reactors, three of which were operating at the time. Three staff were killed as a result of these events — not related to radiation exposure. When the earthquake struck the reactors automatically shut down and the emergency cooling systems were activated but one hour later these were all damaged by a wall of water some 14 m high as a result of the accompanying tsunami. (The tsunami itself was responsible for the deaths of over 26 000 local residents.) Hydrogen explosions subsequently badly damaged the control rooms of the three operating reactors (Units 1, 2 and 3) and there were problems with the spent fuel pool of Unit 4, which subsequently led to a fourth hydrogen explosion. There have been no recorded cases of ARS amongst the staff dealing with the emergency, and none are expected.

Local residents were evacuated out of the area in a staged manner up to a radius of 20 km around the site, the evacuation being compounded by evacuees from the tsunami. The principal nuclides of concern were again those of iodine and caesium. Residents within a 20-30 km radius were instructed to shelter indoors. In contrast to Chernobyl, the radionuclides released were not widely distributed and considerable precipitation subse­quently occurred due to snowfall. Protective actions were immediately implemented with regard to the consumption of contaminated water and foodstuffs and the screening of children, in particular, for iodine con­centrations in the thyroid gland was undertaken. More detailed information is still awaited but clearly the major long-term impact for the local popu­lation, having experienced a severe earthquake, tsunami, and a nuclear accident, will be psychological.

Radiological Protection of Workers and the General Public

3 Conclusions

The current system of radiological protection for people has been developed over a long period of time, and has involved an enormous body of scientific, medical and cultural information. All of these areas are still actively pursued, and the system reviewed and revised. In terms of application, an enormous amount of experience has now been gathered over many decades. Exposures of people to ionizing radiation may be through medical diagnostic or therapeutic exposures, of which there must be a vast number undertaken daily throughout the world; through exposures at work in all forms of industry that may involve radioactive or radiation sources; or through public exposures arising from releases from both nuclear and non-nuclear establishments. All of these exposures, and the sources leading to them, are controlled on the same scientific basis and interpretation, and on the advice of the same international committee-the ICRP, and its extensive support. If there was something ser­iously amiss with this system, then it would by now have come to light. Not that there is any reason to be complacent, as the recent incident at Fukushima, and the 25th anniversary of Chernobyl serve to remind us, accidents can happen, as they can in any industrial endeavor. But it should provide a high degree of assurance to anyone that is concerned about radiation safety and our ability to manage it safely that, whatever the source of exposure or the category of people exposed, the actions taken to safeguard human health are based on a wealth of experience that is unequalled in any other field.

Issues in Environmental Science and Technology, 32 Nuclear Power and the Environment Edited by R. E. Hester and R. M. Harrison © Royal Society of Chemistry 2011

Published by the Royal Society of Chemistry, www. rsc. org

“At present there are over 440 commercial nuclear power reactors operating in 30 countries, with 376 000 MWe of total capacity. In total, they provide about 15% of the world’s electricity. ш The only country that developed nuclear reactors with no military link was Canada, whose ZEEP (Zero Energy Experimental Pile) formed the basis of Canada’s indigenous nuclear reactor design — CANDU — which used natural as opposed to the more expensive, enriched uranium. However, the first reactor which formed part of the Manhattan projects’ attempt to produce plutonium for the atomic bomb, involved scientists from Canada, Britain and France. Although Canada did not develop its own nuclear weapons programme after the war, it did sell plutonium to the UK in order to fund the Canadian civilian reactor programme. ivChicago Pilel. The term ‘‘nuclear reactor’’ was not used until 1952.

[3]The focus on fast breeder reactors (FBR) in these early years reflected a concern that sourcing all of the uranium to power the world’s nuclear reactors was going to prove extremely difficult. However, they all turned out to be too costly to operate and were beset by technological diffi­culties, as well as the heightened proliferation risk that would accompany a ‘‘plutonium econ­omy’’. Subsequently large uranium deposits were discovered in Canada and Australia negating the original rational for FBRs.

[4] Winston Churchill opposed the idea, suggesting to Roosevelt that he stop Bohr travelling to the Soviet Union to make his case, even suggesting at one point that he should be put under house arrest.4

™The plan failed for a number of reasons, including the refusal of the USSR to allow inspections on its territory, as well as the US position that it would not destroy its nuclear arsenal until it was convinced of the efficacy of international control and monitoring procedures. The talks collapsed two years after they had begun and the UN AEC abandoned. It would be a decade before a replacement body, the IAEA, was conceived and then shorn of any pretensions to global over­sight of nuclear matters envisaged by Oppenheimer and Ascheson.5

[6]Before the Second World War, France had invested the most money of any country in the world in the attempt to develop the first nuclear reactor, but German invasion and dispersion of its scientists meant that this honour was to become Enrico Fermi’s, as part of the Manhattan Project The lack of financial resources in the immediate aftermath of the Second World War meant that French nuclear research fell well behind that of the British and Americans.

[7]The UK’s most serious nuclear accident occurred as a result of a fire in Windscale Pile 1 in 1957. Even today, it remains a decommissioning headache in both financial and technological terms. The fire received very little media and public attention at the time, reflecting the tight security and secrecy that enveloped the nuclear industry in these early years of IES development. Unlike today, reactor designs were not subject to public scrutiny and/or parliamentary oversight.

[8] This was a forerunner of the RBMK reactors, the same design as the reactors at Chernobyl.

[9]Rickover became known as the ‘‘Father of the Nuclear Navy’’.

x111 The pressure on reactor designers to keep the costs down, it is claimed, led to compromises on safety, especially given the intense competition from coal and oil-fired power stations.16

[11]A prototype BWR, Vallecitos, ran from 1957 to 1963.

[12]Not all reactors were light water reactors. Canadian reactor development headed down a quite different track, using natural uranium and heavy water as both a moderator and coolant. The first of these ‘‘CANDU’’ units started up in 1962 and they were the first reactors to not have a military connection. Along with Canada, Germany and Sweden followed this heavy water/ natural uranium route given their desire not to have to rely on foreign states for costly enrich­ment services.

[13] It was not until 1995 that plutonium production ceased.

[14] Eventually it estimated that a programme of twelve nuclear power stations with a total capacity of between 1400-1800 MW would be on line by 1965 (ref. 22).

[15] There were technical problems during operation, much longer construction times than planned and as a result a much greater cost of electricity than budgeted.

[16] Led by the then chief economist of the NCB, Fritz Schumacher, who went on to penn the environmentalist classic Small is Beautiful.

** The Prime minister at the time, Harold Wilson, and his Energy Secretary, Tony Benn, were both pro nuclear. Benn was convinced of the case for civil nuclear power based on the ‘‘beating swords into ploughshares” sentiment. A position Benn has since retracted, arguing that he was misled when Minister of Technology about the costs of nuclear energy, ‘‘I was told, believed and argued publicly that civil nuclear power was cheap, safe and peaceful and it was only later that I learned that this was all untrue since, if the full cost of development and the cost of storing long-term nuclear waste is included in the calculations nuclear power is three times the cost of coal when the pits were being closed on economic grounds’’.26

[18] The regulatory structure was also more permissive with regard to nuclear power than exists today, reflecting in part the deferential culture toward experts and scientists in the 1950s. This scrutiny as it existed was carried out by the UK Atomic Energy Authority internal safety branch. It relied in essence on a ‘‘staged operating experience to demonstrate that if the reactor worked, then it must be safe after all’’ which is in stark contrast to the risk-based approach adopted by contemporary regulators.28

[19]This was increased to $7 billion in 1988.

mnAs Laurent Striker, senior vice president at Electricite de France, commented ‘‘France chose nuclear because we have no oil, gas or coal resources’’.

[21]Indeed, in December 2009 the United Arab Emirates accepted a bid from a South Korean consortium to construct four APR1400 reactors by 2020. China will reportedly invest $175 billion over the next ten years on developing the 130 square-kilometre Haiyan ‘‘Nuclear City’’.

[22] During this period of expansion the uranium-based thermal reactor were seen by the nuclear industry as very much the first ‘‘primitive” form of reactor,37 in comparison with more advanced fast breeder reactors which are designed to ‘‘breed’’ more plutonium than they can consume as fuel (some breeders can produce 30% more fuel than they use). India, Russia, Japan and China currently have operational fast breeder reactor programmes. The UK, France and Germany have effectively shut down theirs.

[23] Policy changes meant that there was a shift from ‘‘military uses first’’ to ‘‘combining military and civilian uses’’, this led to a state Ministry being reorganized and renamed to become China National Nuclear Corporation (CNNC) in 1989 (ref. 44).

[24] Some industry insiders suggest that the problem was exacerbated by the introduction post­privatization of a non-technical management, whose focus was on short-term profits, ‘‘with no understanding of the need to technically maintain the assets and the skill base or the long-term needs of the business, which in turn led to the massive shareholder losses’’.62

xxvin The ageing Magnox stations with less than 10 years’ lifespan could not be sold and were given to BNFL.

[26] Exacerbated by the lack of a long-term waste disposal route and a regulatory regime that hindered rather than facilitate new nuclear plants.

[27] However, even a doubling of existing nuclear capacity will only reduce GHG emissions by 8% given that electricity is only a third of total energy production.

““Whilst it is clear that nuclear power is not a completely carbon-free energy source (e. g. both uranium mining and the construction of the nuclear plant relies on fossil fuel energy), it is substantially better than either coal or gas.

[29] In some countries (such as Finland), potential increases in natural gas prices played a key role in the decision to proceed with new nuclear. In addition, nuclear power was portrayed as the cheapest low carbon option.73 In May 2002, the Finnish parliament voted on the new reactor and decided in favour, becoming the first OECD country to decide to build a new nuclear reactor for several years. The vote was very close, however, with 107 votes for and 92 against.

xxxm Bickerstaff et al. describe the British public attitude toward new nuclear as ‘‘reluctant accep­tance”, i. e. when presented against the impending danger of climate change, the risks of nuclear power seem acceptable even to people who are a priori hostile to nuclear power. While science itself is trusted, the government institutions are seen as ‘‘unreliable, secretive and failing to execute their proper duties (or functions) to serve the public interest’’.84

[31]Whilst a recent report from the International Atomic Energy Agency (IAEA) International Status and Prospects of Nuclear Power suggested that there are some 65 countries currently without nuclear power plants who ‘‘are expressing interest in, considering, or actively planning for nuclear power’’ there are technical barriers to at least 17 of those proceeding, due to the fact that they have electricity grids of less than 5 GW which are ‘‘too small to accommodate most of the reactor designs on offer’’. Moreover, many of these countries do not have the ‘‘necessary nuclear regulations, regulators, maintenance capacity, or the skilled workforce to run a nuclear plant. The head of France’s Nuclear Safety Authority has estimated that it would take at least 15 years to build the necessary regulatory framework in countries that are starting from scratch’’.85 There are also doubts as to whether grids of up to 10 GW could cope with nuclear power generation.

[32]For many developing countries the language of national economic development is often invoked as a rationale for investing in nuclear energy, even in countries with no history of the technology.

mvi Russia’s neighbour, Ukraine, is currently building two reactors and planning as many as 11 more by 2030 as it seeks to reduce its dependence on energy from Russia, particularly in light of the disputes over gas in 2006 and 2009. The strategy also envisages completing the construction by 2017 of two reactors at Khmelnitsky, work on which has been halted since 1990. mvnA recent MIT report on the Future of Nuclear Power pointed out that given increasing demand, to increase nuclear powers’ share from its present 17% of world electricity to just 19% by 2050 would require a near-trebling of nuclear capacity: 1000-1500 large nuclear plants would have to be built worldwide.

[34] Prompt decommissioning means that site knowledge is retained and can be used to assist in the decommissioning process.

• Availability of money to perform decommissioning may be limited.

• Financial depreciation makes it attractive to defer large expenditure.

• Radioactive decay means that decommissioning may be easier, and therefore cheaper, if it is deferred. For some wastes, the radioactive decay may result in material being reclassified as a less onerous waste form, so requiring less treatment and incurring reduced costs.

• A final disposal route for wastes may not be available. This may require construction of intermediate facilities incurring costs and using natural resources that would not otherwise be required; these facilities themselves will require decommissioning at the end of their lives. It may therefore be desirable to postpone decommissioning until a final disposal route is available.

Time

[35] Fuel recycle plants at Sellafield

• Fuel enrichment plant at Capenhurst

[36] The waste form. Wastes are conditioned (see section 3.3.5) prior to disposal to make them more stable. For example, highly active raffinate

[37] Site selection for GDF construction will be based upon community volunteerism and the siting process will take several years (see section 3.3.3).

(ii) The GDF will be tailored to the UK baseline inventory which is both large by volume and radioactivity, and complex in character due to the

[38] HLW: The highly active raffinate (HAR) solution from fuel reprocessing (see section 2.2) produces excessive heat and radioactivity and is highly unstable. Consequently, HAR is evaporated to reduce its volume to form highly active liquor (HAL). The HAL is then stored in water cooled tanks to permit heat dissipation and radioactive decay. After storage, the HAL is homogenised, immobilised, and conditioned in a

[39] Planned exposure situations, which are situations involving the planned introduction and operation of sources, and include situations that were previously categorised as ‘‘practices’’. These include situations that are anticipated to occur (in other words, ‘‘normal’’ exposures) as well as exposures that are not anticipated to occur but may occur (‘‘potential’’ exposures), such as accidents. In the latter case, although the situation was not planned to occur, the situation itself can be planned for, although not necessarily in great detail. These days such potential exposures can include a variety of possibilities, from accidents that may