In any aqueous environment, cations will be complexed, either by water molecules (hydration) or by other ligands present. The natural environment contains a range of common ligands, such as CO32 , OH, Cl and natural organic matter that can complex radionuclides. In addition, synthetic organic ligands can also be present as co-contaminants. For actinides the strength of inorganic ligand complexation decreases in the order: CO32 , OH >HPO43 , F, SO42 >NO з, Cl.109 Actinides interact with anionic ligand species by strong ionic bonding and so the complexation strength is related to the actinide charge; therefore the trend of complexation is: An4+>AnO22+>

An31 > AnO+.110 In natural systems the most important inorganic ligand is carbonate; it is ubiquitous in the environment, with concentrations ranging from 10 5 M in surface waters to 10 2 M in groundwater.64 Carbonates can form stable, negatively charged complexes with actinides. These complexes have a lower affinity for negatively charged mineral surfaces and so the com — plexed radionuclide will tend to remain in the solution phase, therefore increasing mobility in the environment. For example U(VI) has a high affinity for carbonate ligands, forming stable complexes such as UO2(CO3)22 or

Pathways of Radioactive Substances in the Environment и02(С03)34 as shown in equations (1) and (2).111

UO^ + 2HCO3- ! UO2(CO3)2^aq) + 2H+ (1)

UO2(CO3)2(aq) + HCO3- ! UO2(CO3)3(aq) + H+ (2)

The speciation of dissolved uranium in the presence of inorganic carbon is summarised in Figure 5. In oxidising and mildly reducing environments at pH >6, the negatively charged UO2(CO3)22 or UO2(CO3)34 dominate, between pH 5- 6 the less mobile UO2(CO3) species dominate and below pH 5 the uranyl ion (UO221) dominates.112

The presence of natural and synthetic organics may also increase the mobility of radionuclides. Natural humic substances are ubiquitous in the environment and concentrations can range from <1 to >200mg l 1 in wetlands.113 Humic substances are composed of three fractions: the humin fraction is insoluble under all pH conditions; humic acids are insoluble at <pH 2, and fulvic acids are soluble under all pH conditions.114 Humics and fulvics are important natural ligands that are able to complex with radionuclides and may increase mobility. Humic complexation is effected by pH; with increasing pH there is an

increase in the ionisation of humic functional groups (e. g. carboxylic and phenolic), thus increasing the humic complexation strength.115 The stability of humic complexes depends on the oxidation state of the complexed metal. The stability of radionuclide-humic complexes decrease in the order U(iv)>Th (iv)>Am(iii)>Eu(iii)>U(vi)>Co(II)>Sr(ii), as a result of strength of elec­trostatic interaction between radionuclide and functional group.42,116 Acti — nides-humic complexes can modify radionuclide oxidation states; the mediated reduction of Np(vi) to Np(iv) and Pu(vi) to Pu(iv) has been reported to occur.117 Complexation by humics also depends on the presence of other complexants. Moulin and Moulin118 investigated the effect of humics on acti­nide migration under conditions relevant to nuclear waste disposal (pH < 7). Only An(iii) were found to be complexed by humics, whilst An(v) and An(vi) were complexed with carbonates or hydroxide.

Organic acids can also be present in the environment as a result of microbial activity. These ligands can sequester cations from mineral surfaces or stabilise cations in the solution phase.119 The acids released during microbial metabolic processes are also potent mineral dissolution agents.119 Most minerals are stable at circumneutral pH and dissolve in the presence of acid101,120 releasing adsorbed or incorporated contaminants.

Organic complexing agents used in the processing of nuclear fuel or during decontamination (e. g., nitrilotriacetic acid (NTA), ethylenediaminetetraacetic acid (EDTA) and citric acid121) can be present as co-contaminants with radio­nuclides; for example the 149 radioactive waste storage tanks at the US Department of Energy contain an estimated 83 metric tons of EDTA.122 Com­plexation of radionuclides by organic ligands can reduce the reduction potential of the metal and several studies have found that bacterial reduction of radio­nuclides increases in the presence of organic complexants. Laboratory studies showed only minor reduction of Pu(iv), present as amorphous Pu(OH)4 to Pu(iii) by Shewanella oneidensis and Geobacter metallireducens.123 However, Pu(iv)- EDTA complexes were rapidly reduced to the more mobile Pu(iii)-EDTA by the same bacteria.123 in another study, the solubilisation of PuO2 by Fe-reducing bacteria was increased in the presence of NTA by approximately 90%; the pro­posed mechanism was reductive dissolution of Pu(iv) to Pu(iii).124 The co-dis­posal of radioactive 60Co(ii) and EDTA is also a concern. in the environment the Co(ii)-EDTA complexes can be oxidised by manganese(iv) and iron(iii) oxide minerals to the more stable and mobile Co(iii)-EDTA.47 Although metal-redu­cing bacteria can reduce Co(iii)-EDTA back to the less mobile Co(ii)-EDTA, in a natural system re-oxidation by oxide minerals may occur.125,126

Co-contaminants can also stabilise radionuclides in the solution phase and therefore enhance migration in the environment. AlMahamid et al. (1996)127 investigated complexation of Pu(iii, iv, v and Vi) with NTA and EDTA. The predominant oxidation state at pH 5 to 8 was Pu(iv); Pu(iii) was oxidised and Pu(v/vi) were reduced under these conditions. Critically, the presence of NTA and EDTA stabilised Pu(iv) in solution. in another study, the co-contaminant citrate was found to decrease the adsorption of U(vi) onto iron-rich sand.128 The dominant mechanism was thought to be chemical alteration of the sand surface by citrate, decreasing the adsorption of U(vi), but aqueous complexa — tion of U(vi) by citrate may also be significant.128

A global bottleneck in the supply chain could derail the current plans for new nuclear build. These include a lack of skilled engineers, as well as a backlog in orders for machine parts and for reactors vessels. From an engineering standpoint, the larger the order book for new reactors the more cost effective it is for companies to invest in developing the skills of the workforce. In the USA, a 2009 survey by the Washington-based Nuclear Energy Institute showed that 38 percent of the current workforce in the nuclear industry will be eligible for retirement by 2014. This has led to tailored two-year education and training programmes being rolled out at technical colleges across the USA.120 However, such workers will be in demand globally; already a significant amount of American skilled nuclear engineers and managers work in the UK decommissioning sector, given that the UK ran down its indigenous nuclear skills base as a result of the stagnation of its nuclear industry.

Some countries are more susceptible to labour shortages than others, which may impact upon ambitious new build programmes. France faces an acute shortage of skilled workers with some 40 percent of EDF’s operators and maintenance staff retiring by 2015. This has led France to advertise for overseas students to study for masters degrees in nuclear engineering and related sub­jects in France. ‘‘The need for students in atomic energy is estimated at 1200 graduated students a year for the next 10 years, although, nowadays, the number of graduated students is 300 per year’’.121 If France can tempt overseas students to work in its nuclear power programme, then its potential foreign clients to whom it wants to sell its own nuclear technologies, won’t have a nuclear workforce of their own!122

In developing countries which are pursuing an aggressive nuclear expansion policy, the shortage of skilled labour is particularly severe. In China for instance, many students who enrol for nuclear engineering programme at university end up changing their majors, with only 30% staying in the field.123 This is a potential problem given that China will need at least 6000 nuclear engineers by 2020.124

Gaps and supply chain problems can lead, as they have in the past, to inflation in the cost of key parts and to delays in construction. A shortage of trained personnel can drive costs further up. Today, for example, only one facility in the world (Mitsubishi Heavy Industries, Ltd) has the forging capability to manufacture large reactor vessels, xlvi which raises questions about the ability of the firm to meet the increasing global demand for reactor vessels. More investment is necessary in plants that are capable of manufacturing reactor vessels. As more and more existing reactors are having their operating lifetime extended there is also an increasing demand for a range of replacement nuclear components that require advanced heavy machinery manufacturers.

1.3 Events Leading to the Accident

The accident at reactor number two at the Three-Mile Island (TMI) nuclear power station in Pennsylvania, USA, occurred on the early morning of the 28th of March 1979. Whilst leading to relatively little off-site environmental con­tamination, the accident caused a partial core meltdown and was considered to be the most serious accident at a civilian nuclear power station prior to Chernobyl.

The accident was initiated by a failure in the secondary (non-nuclear) elec­tricity generating part of the plant which caused a pressure increase in the cooling system of the primary (nuclear) part of the plant. The pressure was released by a valve but, once released, the valve failed to close, causing loss of cooling water.26 Had proper action been taken at this stage, this would have been a relatively minor incident, but subsequent operator errors, caused by poor information and training27 led to a major reactor failure. There was no indication to the operators that the pressure release valve had failed to close, nor was there a clear indication of the level of the coolant in the reactor core. With ‘‘more than 100’’ (ref. 27) alarms going off in the control room, the plant operators were not aware that coolant was being lost from the core. Thus, when an automatic safety system began to inject cooling water into the core, the operators overrode the system, drastically reducing the rate of injection of cooling water. The loss of coolant led to exposure and the resulting meltdown of a significant part of the reactor core.

Low level waste is defined as material having a radioactive content between not exceeding 4GBqtonne-1 of alpha, or 12GBqtonne-1 beta/gamma activity.2 Most LLWs consist of contaminated paper, plastic and metal products, and these are currently disposed of via shallow burial at the UK national LLW Repository Site, Drigg. However, a small proportion of LLW is unsuitable for shallow disposal due to the concentration of certain long lived radioisotopes (e. g. Np, Pu and Am) in the waste, as well as chemical incompatibility issues. These LLWs will be managed via geodisposal and treated in the same way as ILWs.2

The degree of absorption from the gastrointestinal tract is an important factor in determining the radionuclide content of animal tissues.

For mammals, fractional absorption has been quantified as the true absorption coefficient, At, which can be determined as the difference in dietary intake and faecal output (corrected for the endogenous secretion of radio­nuclides into the gastrointestinal tract),29 expressed as a proportion of dietary intake. Absorption of most essential elements is controlled by dietary supply and the animal’s requirement (absorption tending to decrease with increasing dietary concentrations when requirement is met) and, in some instances, other essential elements such as interactions between calcium and phosphorous.

Fractional absorption values for both monogastric mammals and ruminants can be found in papers by the ICRP30 and Howard et al.31 The absorption of essential elements tends to be relatively high. In contrast, elements with high atomic weights which are not essential elements or analogues of essential elements are generally poorly absorbed.

Most forms of iodine are rapidly reduced to iodide within the digestive tract and absorption is complete regardless of the source of radioiodine or amount of dietary stable iodine.

In the case of radiocaesium, the source ingested is a major factor determining subsequent concentrations in tissues, with the true absorption coefficient ran­ging from <0.10 to >0.80.29 Absorption of particle or soil-associated radio­caesium is considerably lower than that of radiocaesium incorporated within plants.

The variation in At for radiostrontium, with recorded values ranging from 0.05-0.7, does not appear to be related to source but is strongly influenced by its analogue, calcium, which is a homeostatically controlled essential element. The extent of calcium absorption in the gastrointestinal tract is governed by the animal’s calcium requirement which depends on factors such as age, growth rate and milk yield. At a given calcium requirement, there is an inverse rela­tionship between the absorption of calcium and the amount of calcium in the diet.29 Under normal levels of calcium intake, the source of radiostrontium ingested is unlikely to influence either the extent of absorption or the con­centrations in tissues.

The bioavailability of transuranic elements, such as plutonium and americium, is low compared with that of many elements. Fractional absorption values are generally less than 0.0001.

Gastrointestinal fractional absorption decreases with age which may be due to the lower permeability of the membranes of the gastrointestinal wall of mature animals compared with young animals (most notably from birth to a few weeks old) which have a greater need to absorb a wide range of nutrients and essential elements.

1.1.1 Uranium/Plutonium Fast Reactors

The high energy (‘‘fast’’) neutrons produced in fission have an energy > 1 MeV and can fission 238U effectively. In addition, the high energy fission events triggered by fast neutrons produce more neutrons (e. g. in 239Pu, 4.9 for fast versus 2.6 for thermal) so that there is a substantial surplus of neutrons which can be used either to transmute a fertile material (a breeder reactor) or to destroy problematic waste isotopes. A thermal reactor is specifically designed to ‘‘moderate’’ neutron energies by allowing the neutrons to collide with

light atoms. It is, however, also possible to design reactors where reaction is sustained by fast, not thermal, neutrons. While such reactors could be sup­ported by variants on the present U/Pu fuel cycle, there are major engineering difficulties associated with them due, for example, to the very high energy density in the cores, and the need to use corrosive materials such as liquid metals as coolants. Technologically, fast reactors are much less well developed than thermal reactors.

The Oak Ridge Reservation in east Tennessee consists of about 357 250 acres of land and contains the Y-12 Plant which played a historical role in the production of nuclear weapons.33 Oak Ridge also houses one of the three DOE gaseous diffusion plants (K-25 Plant) used to enrich uranium. The Y-12 facility itself encompasses 324 hectares and is associated with mercury contamination arising from its use in nuclear weapons production through the 1950s and 1960s. An estimated 108 000 to 212000 kg of mercury was released into the headwaters of the East Fork Poplar Creek over this time frame from the Y-12 plant.34

Further studies have showed that 77 180 kg of mercury are contained in the sediments and floodplain soils of a 15 mile stretch of the East Fork Poplar Creek which has its headwaters at Y-12. Some 227 kg are thought to leave this watershed annually.35 The concentration of mercury in soils from the flood­plains ranges up to 2400 mgg-1 with an average of greater than 70% found to be metallic mercury and mercuric sulfide.36

Between the years 1951 to 1983, liquid acidic wastes (pH <2.0) containing metals (including uranium) dissolved in nitric acid were discharged into four seepage pits on the Y-12 complex known as the S-3 Ponds. The ponds were neutralised in 1983 by the addition of limestone, quicklime and sodium hydroxide until the pH reached greater than 9.0 resulting in the precipitation of calcium, iron and aluminium compounds.37 Leakage from the S-3 ponds has created a contaminated groundwater plume in the underlying shale bedrock which extends more than 2 km both east and west of the ponds. Analysis of soils from this site has revealed sorbed and precipitated uranium concentrations of up to 800 mg kg-1.38 A summary of groundwater contaminants in the plume extending from the S-3 ponds is given in Table 3.

Uranium and technetium are likely to exist as the mobile U(vi) (as UO221) and Tc(vii) (as TcO4 ) forms, although in the low pH conditions and in the presence of high nitrate and sulfate concentrations the uranium may be associated with nitrate (as UO2NO3+) or sulfate (as UO2SO4 or UO2(SO4)22-).39 These condi­tions present problems for remediation strategies such as in situ bioremediation as the groundwater requires pre-treatment so that conditions are favourable for

microbially mediated reduction, discussed later in this chapter. High nitrate concentrations can potentially present further problems for remediation as in addition to being a competing electron donor to U(vi), it is also an effective oxidant of U(iv), leading to the oxidation and remobilisation of U(vi) through a variety of possible mechanisms.40 2.3.3 Hanford

The Hanford Nuclear Reservation, located on the Columbia River in Washington State, USA, was a plutonium production site which began operation in 1945. As a consequence of its former activities, a number of contamination issues have arisen. It has been estimated that more than 436 TBq of 239Pu, 1065 TBq of 241Am and 2 TBq of 237Np were disposed of as liquid waste to the near surface at Hanford,41 with 86% of the 239Pu, 97% of the 241Am and 77% of the 237Np released into the Plutonium Finishing Plant (PFP) zone of the site. Despite the release of large quantities of radionuclides into the Hanford Vadose zone, only negligible amounts have entered the groundwater. Filtered samples from on site wells have so far shown no activity above the DOE-derived guide for 239Pu (1.11 BqC1) and only two unfiltered samples taken in 2006 exceeded this limit (1.3 and 1.5 BqC1).41

Although the majority of plutonium on site is reported to be immobile, there are areas where vertical migration of plutonium and americium has occurred. The Z-9 trench is considered a ‘‘worst case’’ representation of the disposal area due to the acidic (pH 2.5), high salt waste solution containing nitrate (~5M), aluminium (~0.6 M) and organic solvents. As much as 140 kg of plutonium was disposed of in this waste and although ~ 58 kg of plutonium was recovered in 1978 , 239 1240Pu has been found to be concentrated (up to 9.25 MBqkg-1) in silt layers 15-20 m below the surface correlating with the occurrence of co-disposed tributylphosphate (TBP).42 Americium-241 was also found to be accumulated at this point, and also concentrated at a second horizon at the bottom of the underlying fine-grained Cold Creek Unit (~40m below ground surface) at levels greater than 11.1 MBqkg-1 with no accompanying TBP.42 A number of possible conditions may have contributed to this vertical migration including the acidic nature of the waste, formation of soluble complexes and suspension and transport of colloids or nanoparticles.

In order to generate the 239Pu needed to produce nuclear weapons, a very large quantity of uranium (either as metal or UO2) was irradiated. The sub­sequent retrieval of plutonium from the matrix resulted in a large volume of aqueous waste containing high concentrations of uranium. This waste (also containing other fission products) is stored in 177 underground steel tanks in different areas of the Hanford site referred to as ‘‘Tank Farms’’ which are subdivided into Waste Management Areas. A large number (68 out of 149) are known or suspected to have leaked to date with the largest release occurring in 1951 in the 241-BX-Tank Farm. Nearly 3.5 x 105 litres of highly radioactive waste containing more than 7000 kg of uranium was released into the sub — surface.43 Between the years 1944 and 1988, almost two million cubic meters of tank waste was generated with subsequent evaporation, discharge, chemical treatment and leakage reducing this volume to 200000 cubic meters. This makes up to 60% of the current tank waste which contains around 7.03 x 106 TBq of radioactivity and 170 000 tonnes of chemicals with each cubic meter of tank waste containing nearly 37 TBq of radioactivity.44

The mobility of uranium in the contaminated sub-surface beneath the tank farms has been shown to vary depending on the surface phases present at different depths. Uranium silicate precipitates were found in relatively shallow sediments, whereas the uranium was found adsorbed to sediment surfaces at intermediate and deeper depths, both in the form of U(vi).45 Migration of uranium in the shallow sub-surface may therefore be slow as it relies on the slow process of mineral dissolution. In contrast, migration may be relatively fast in deeper conditions as surface desorption processes occur over a faster time frame. Work conducted on sediments taken from boreholes near to the storage tanks at Hanford revealed that the uranium is again predominantly found in the U(VI) state, with approximately 51% to 63% labile and therefore potentially mobile, with the remaining portion locked up in mobilization — resistant phases.46

A history of liquid waste disposal activities is provided by Gephart44 and is briefly summarised here. Liquid waste has been dealt with via a number of methods during the operations of the Hanford site. In 1944, during fuel reprocessing, liquids which were only mildly contaminated were dumped into depressions on the ground, contaminating both the sandy sediments and eventually the groundwater. Some of these contaminants were blown downwind thus contaminating an even greater area. When dumping of these liquid wastes was halted they were instead pumped down reverse wells which lead to contaminants being injected closer to, or directly into, the underlying hydrology, bypassing the overlying sediment which could otherwise have acted, via sorp­tion, as a sink. When this process was stopped after only a few months, liquid wastes were pumped directly into shallow buried box structures, gravel-filled tile fields, buried concrete pipes and open trenches later backfilled with gravel.

These processes, compounded by tank leakages, have led to the con­tamination of up to 28 300 m3 of soil,47 which along with the contaminated groundwater contains around 8325 TBq of 137Cs, 6660 Tbq of 3H, 1924 TBq of 90Sr, 1850 TBq of Pu and 25.9 TBq of 99Tc (ref. 44). Groundwater underlying around 12% of the Hanford site contains carbon tetrachloride, chromium, nitrate, 90Sr, 99Tc, 129I and uranium at levels above the drinking water stan — dard.48 Although the site groundwater is not a source of public drinking water and does not significantly affect off-site water resources, contaminants such as 99Tc and 129I are mobile in groundwater and thus can migrate deep into the vadose zone and could potentially enter aquifers.

Earlier work estimated that up to 4 x 1016 Bq of 137Cs had leaked into the vadose zone from the tank farms with measured activity from contaminated sediments as high as 105 Bqg-1.49 The waste stored in these tanks also typically contained a significant concentration of high ionic strength solutions including NaNO3 (> 0.5 moll-1).49 In the presence of high salt concentrations, caesium was only found to absorb to high-affinity, frayed edge sites of mica minerals with sodium being an effective competitor for such sites.50 The high sodium released in the leaked Hanford waste may therefore prevent the retardation of 137Cs at the site. Borehole data also suggests that caesium is not undergoing significant sorption as peak 137Cs activities were detected between 20 and 26 m, reaching up to around 40 m, beneath the SX tank farm responsible for the majority of the caesium release.51

The global legacy of radioactive wastes combined with the urgent need for low carbon power generation means that we are now at a pivotal time for the implementation of geological disposal facilities for radioactive wastes. In the UK, we have a large and complex legacy of materials from over half a century of nuclear power plant operation and weapon production. Indeed, the fact that the UK was the first nation to implement nuclear power generation, that we have been active in creating a bespoke research and development programme in nuclear power, and that we have always had an extensive reprocessing programme, means that we have a waste legacy that is both diverse and highly challenging. With this background in mind, it is useful to define our view on some of the key environmental chemistry challenges facing successful HAW geodisposal:

(i) For ILW (assuming a cementitous waste form and repository envir­onment), there are challenges in understanding and predicting the corrosion and evolution of the waste form and host rock environment during storage and disposal.

(ii) Microbial processes can affect the solubility of radionuclides;20 how­ever, the fundamental biogeochemistry of radionuclides (especially the long lived radionuclides of importance to geological disposal including the transuranics) is poorly defined, especially under geological disposal settings. The presence of electron donors such as organic matter in the waste, H2 from radiolysis of water and anaerobic corrosion of iron metal, and iron metal itself within the GDF, combined with the increasing recognition that “extremophile” microbes can tolerate high pH conditions and extreme radiation fluxes, means that understanding the whole biogeochemistry of a evolving GDF will be critical in underpinning the safety case.

(iii)For HLW/spent fuel, there are challenges in co-locating the waste within the same GDF as ILW as it is likely that a large fraction of ILW will be grouted and backfilled with cement. This hyperalkaline ILW concept is chemically incompatible with international HLW and spent fuel concepts although separate ‘‘sub-chambers’’ for ILW and HLW that take into account the predicted hydrogeology of the subsurface may offer a route forward.

(iv) There is a poor understanding of the long term waste form performance of UK specific fuels such as advanced gas cooled reactor oxide fuels.

(v) Colloids have the potential to influence radionuclide transport both within and outside of an evolved GDF environment.

(vi) It is important to note that computational modelling, from an atomistic scale to a regional scale, is required to underpin the predictive model(s) for GDF performance assessment. This is because the scientific com­munity cannot hope to perform experiments on all the systems of interest (particularly with highly radiotoxic transuranics) over all of the time and spatial scales of relevance.

(vii) Research with radionuclides presents a significant safety hazard and the UK infrastructure and capability to perform these experiments has been eroded in the last decades. As a consequence, successful GDF implementation will require innovation, collaboration, and investment across a range of scientific disciplines.

In conclusion, it is important to realise that the safety case for the successful implementation of a UK GDF will involve environmental chemistry challenges that encompass a range of scales from sub micro-second reactivity to millions of years; from molecular scale understanding of biogeochemical processes to field scale transport modelling; and from essentially pure radionuclide phases such as UO2 fuel to sub 10~12 molar concentrations of radionuclides that are likely to be transported from the GDF into the host rock. Implementation of the UK GDF is a unique and exciting challenge and it requires innovation across the scientific disciplines and a commitment to engage with the wider public to deliver the reward of safe and credible management of the nuclear legacy. Thus our view is that the management of radioactive wastes forms a real focus for environmental chemistry research that will enable both man­agement of a difficult and challenging legacy and potentially allow con­sideration of a new fleet of nuclear reactors to power the future energy needs of the UK.

Acknowledgements

We thank Prof. Francis Livens for helpful discussions and for proof reading the manuscript. We acknowledge support from NERC for ongoing research on nuclear legacy management, particularly grant numbers NE/H007768/1, NE/ D00473X/1, NE/D005361/1.

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This sophisticated approach would not be possible were it not for the fact that there is sufficient information on the effects of radiation on humans to be able to differentiate between levels of dose received by different tissues and organs from both external and internal sources, and the different types of effect that result. These relationships are complex, but in order to manage the exposure of people to radiation, it is useful to consider the principal effects of radiation as being those that result in either:

(i) deterministic effects, which are due in large part to the killing or malfunction of cells following high doses of exposure to radiation; and those that cause

(ii) stochastic effects, such as cancer and heritable effects, which involve either cancer development in exposed individuals due to the mutation of somatic cells, or heritable disease in their offspring owing to mutation of reproductive (germ) cells.

These two categories do not of course cover every adverse health effect, and thus consideration is also given to effects on the embryo and foetus, and to diseases other than cancer.

The induction of deterministic effects (or what are also called ‘‘tissue reactions’’) is generally characterized by a threshold dose of radiation, because radiation damage to a critical population of cells, in a given tissue, needs to be sustained before injury is expressed in a clinically relevant form. Above the threshold dose, the severity of the injury, including impairment of the capacity for tissue recovery, increases with increasing dose. In cases where the threshold dose has been exceeded, early tissue reactions (days to weeks) may be of an inflammatory type, resulting from the release of cellular factors, or there may be reactions resulting from cell loss. Late tissue reactions (months to years) can be of a generic type if they arise as a direct result of damage to that tissue. Essentially, in the absorbed dose below about 100 mGy, tissues are judged generally not to express clinically relevant functional impairment, and this judgement applies both to single acute doses, and to situations where these low doses are experienced in a protracted form as repeated annual exposures.

With regard to stochastic effects, epidemiological and experimental studies provide evidence of radiation risks of cancer, albeit with uncertainties, at doses of about 100 mSv or less. As to be expected, a very large amount of effort has been expended on trying to understand, and quantify, this cancer risk. As far as the mechanisms are concerned, the accumulation of cellular and animal data over decades of study lead to the view that DNA damage response processes within single cells are of critical importance to the development of cancer after radiation exposure. Of particular importance are effects such as the induction of complex forms of DNA double strand breaks, the problems experienced by cells in correctly repairing these complex forms of DNA damage, and the consequent appearance of gene or chromosomal mutations.

Due to its stochastic nature, great reliance is necessarily placed on epide­miological information relating to the incidence of cancer. These data bases are continually growing, and a lot of information on the risk of organ-specific cancer following exposure to radiation has come from the continuing follow-up of survivors of the 1945 atomic bomb explosions in Japan — the so-called Life Span Study (LSS). These data relate to both cancer mortality and cancer incidence, the latter providing more reliable estimates of risk. In addition, epidemiological data from the LSS provide further information on the temporal and age-dependent pattern of radiation cancer risk, particularly the assessment of risk amongst those exposed at an early age.

The LSS is not, however, the sole source of information, and data from medical, occupational and environmental studies are also considered and evaluated. For cancers occurring in some parts of the human body, there is reasonable compatibility between the data from the LSS and those from other sources. There are also differences in radiation risk estimates among the various data sets, however, and most studies on environmental radiation exposures currently lack sufficient data (on dosimetry and tumour ascertainment) to contribute directly to risk estimation.

Part of the problem of relating the cancer risk to radiation exposure is the large range of doses and dose rates over which observations have been made. With low doses the risk of developing cancer is much less, and thus the data bases need to be much larger in order to obtain statistically significant infor­mation. A dose and dose-rate effectiveness factor (DDREF) has therefore been used by UNSCEAR to project cancer risks determined at high doses, and high dose rates, to the risks that would apply at low doses, and low dose rates. From a combination of epidemiological, animal and cellular data, cancer risks at low doses and low dose rates are judged, in general, to be reduced by the value of the factor ascribed to the DDREF. And although, in reality, different dose and dose rate effects may well apply to different organs and tissues, the ICRP judges that, for the general purposes of radiological protection, a DDREF of 2 should be applied, and this is used to derive its nominal risk coefficients for all cancers.

One approach that has been used to manage cancer risk at low doses and low dose rates is to assume that, at doses below about 100 mSv, a given increment in dose will produce a directly proportional increment in the probability of incurring cancer (or heritable effects) attributable to radiation. This dose — response model is generally known as the ‘‘linear-non-threshold’’ or LNT model. It is a view that accords with that given by UNSCEAR,10 but other estimates have been provided by various national organisations. Some of these are in line with the UNSCEAR view11,12 whilst others are not — such as the report from the French Academies,13 which argues in support of a practical threshold for radiation cancer risk.

As far as the ICRP is concerned, although it recognises that there are exceptions, it judges that the weight of evidence supports the view that, for the purposes of radiological protection, it is scientifically plausible to assume that the incidence of cancer or heritable effects will rise in direct proportion to an increase in the equivalent dose in the relevant organs and tissues at doses below about 100 mSv. This is, of course, a practical judgement. In arriving at it, the potential challenges associated with information on cellular adaptive responses, and on the relative abundance of spontaneously arising and low-dose-induced DNA damage, have also been considered. Indeed, there are other factors to consider, such as the existence of the post-irradiation cellular phenomena of induced genomic instability and bystander signalling. All of these biological factors, together with possible tumour-promoting effects of protracted irra­diation and immunological phenomena, may influence radiation cancer risk.14 Because the estimation of nominal cancer risk coefficients is based upon direct human epidemiological data, however, any contribution from them would, in any case, be included in that estimate.

One further point in relation to the LNT model needs to be noted, and that is whilst it remains a scientifically plausible element in a practical system of

radiological protection, any biological or epidemiological information that would, unambiguously, verify the hypothesis that underpins the model is unlikely to be forthcoming.10,11 Due to this uncertainty, the ICRP advises that it is not appropriate, for the purposes of public health planning, to calculate the hypothetical number of cases of cancer or heritable disease that might be associated with very small radiation doses, received by large numbers of people, over very long periods of time.

In addition, contrary to the commonly held view that often appears in the popular press, there is currently no direct evidence that the exposure of human parents to radiation leads to an excess of heritable disease in their offspring. There is compelling evidence, however, that radiation does cause heritable effects in experimental animals. The ICRP therefore prudently includes the risk of heritable effects in its system of radiological protection. This risk is based on the concept of the doubling dose (DD) for disease-associated mutations, for which the present estimate, up to the second generation, is about 0.2% per Gy.

With regard to the embryo and foetus, it is recognized that there is a susceptibility to the effects of irradiation in the pre-implantation period of embryonic developments, and there are gestational age-dependent patterns of in utero radiosensitivity, with maximum sensitivity being expressed during the period when organs are being formed. At doses under 100 mGy, lethal effects are very infrequent and, on the basis of animal data, it is judged that there is a true dose threshold of around 100 mGy for the induction of malformations. For practical purposes, therefore, it is assumed that risks of malformation after in utero exposure to doses well below 100 mGy would not be expected. With regard to cancer risk following in utero irradiation, the largest case-control studies of in utero (medical) irradiation provide evidence that there is an increased risk of childhood cancer of all types. There are particular uncer­tainties regarding the risk of radiation-induced solid cancers following in utero exposure, but it is prudent to assume that life-time cancer risk following in utero exposure will be similar to that following irradiation in early childhood which is, at most, about three times that of the population as a whole.

4.1 Growth in Demand

As more countries pursue industrial and technological development, with the associated increases in energy demand, and it becomes increasingly unac­ceptable to use carbon-based fuels in large quantities, alternative sources of energy are required. Nuclear power has been out of favour for a generation in North America and much of western Europe, partly as a result of the Three Mile Island (1979) and Chernobyl (1986) accidents, but is now being re­evaluated and new reactors are proposed in the UK. Current UK proposals are for these new reactors to operate an open fuel cycle. The impact of Fukushima on these proposals is not yet clear. In Germany, there is public and political opposition to new nuclear build, and proposed life extensions for existing reactors have been thrown into doubt following the Fukushima accident. Sweden’s government changed a 30-year old policy of phasing out nuclear

The majority of the information in this section is taken from the World Nuclear Association (http://www. world-nuclear. org/info/).

energy in 2010, and the existing reactors are being uprated and their lives extended. Switzerland plans to retain nuclear as part of its energy mix and build replacements for existing reactors as they reach end of life. Italy, having closed its last reactors in 1990, has imported large amounts of nuclear electricity from France and had started to consider new build, though these plans are now delayed by the Fukushima accident. France, by contrast, has consistently developed nuclear electricity generation. It currently operates 58 reactors, providing 75% of its electricity, and intends to replace these from 2020, which implies construction of about one new reactor per year for 40 years. France is also actively pursuing more advanced nuclear technologies (fast reactors and Generation IV systems). In the USA, after 30 years in which no new reactors were built, 16 applications to build a total of 24 new reactors have been made since 2007. However, the existing 104 reactors, many of which have had life extensions, continue to provide about 20% of US electricity.

In other regions, notably Asia, some nations such as Japan and South Korea have pursued nuclear energy over many years. China currently has 12 reactors, with 24 under construction, and further plans for large scale expansion. Japan aims to expand nuclear generation from 30% of electricity production today to 40% by 2017, and also sees nuclear power as an important part in a 90% reduction of carbon emissions by 2100. South Korea is now positioned to export reactor technology. In the Middle East, several nations are exploring nuclear generation. Iran is in the process of fuelling a Russian-built reactor, while the United Arab Emirates have ordered four South Korean reactors.

Worldwide, there are currently 440 reactors operating, and projections of future growth in nuclear capacity, while they are obviously very uncertain, range from a 2-4 fold expansion by 2030, 3-10 fold by 2060 and 6-30 fold by 2100.