SCOPE 50 - Radioecology after Chernobyl, Chapter5, Radionuclide Aquatic Pathways

Aquatic environments occupy a major portion of the Earth's surface and, therefore, an understanding of the radionuclide and radiation exposure pathways within such systems is essential. The environments include rivers, lakes, estuaries, shelf seas, deep oceans, ice sheets and glaciers, groundwater and ground-ice. The aquatic environments all consist of an aqueous phase and a solid phase which is mainly sediment (particulates) in surface environments and the host bedrock in groundwaters. Living organisms, which make up the aquatic biota, can be involved in the geochemical cycle of radionuclides and play a role in their phase distribution. Radionuclides are present in the living and non-living components of each aquatic environment, both natural radionuclides of primordial and cosmogenic origin, and artificial radionuclides from nuclear and non-nuclear industrial wastes, accidental releases and nuclear weapons fallout (see sources in Chapter 1). Solid and liquid inputs of artificial radionuclides to the aquatic environments can be direct or indirect. Direct input mechanisms include fallout deposition onto the water surface, liquid discharges and releases from dumped solid wastes. Indirect inputs from secondary sources are also very important, as a result of the remobilization of contaminated material within an environmental compartment, e.g. erosion of soils contaminated by fallout radionuclides in a river catchment or of marine sediments contaminated by nuclear fuel reprocessing wastes. Indirect inputs also occur by transfer between environments of radionuclides in solution and on sediments.

Each radionuclide will be partitioned between the solid and solution phases. A variety of disparate processes may be involved in this partitioning, including sorption by an inorganic or organic environmental sediment particle, or the bedrock for groundwaters, precipitation

dissolution, colloid aggregation

disaggregation, microbial activity, and uptake into and release from the biota. The solid-solution partitioning of a radionuclide is a very important parameter describing its behaviour and, except for biological uptake, it can be defined by the distribution coefficient (K_d) or the numerically equivalent ratio (R_d) introduced by NEA/OECD (1983):

	Activity concentration of the solid phase (Bq kg^-1)
K_d =		(5.1)
	Activity concentration of the solution phase (Bq l^-1)

The widespread use of the K_d is due to the need for a simple parameter describing solid/solution distribution for the purpose of modelling the biogeochemical distribution of radionuclides. The concept of K_d implies true equilibrium and reversibility, and that the solid/solution ratio has no effect on K_d due to sufficiently low radionuclide concentrations. In practice, the first two requirements are frequently not demonstrated. Variations in K_d are likely to occur not only because of possible biological and colloidal effects but also due to changing solution and sediment chemistries and to the non-attainment of equilibrium.

K_d values quoted in the literature have been determined experimentally or from observation of the partitioning in the environment. Distribution coefficients vary by 9 orders of magnitude between different nuclides and 3 orders of magnitude for any particular one. This variation depends mainly on the solution composition and the nature of the solid substrate. It can be modulated by non-equilibrium conditions, resulting from slow kinetics, and by the effects of colloid or biological processes. Especially important in the aquatic environment are the differences in K_d due to the change in solution chemistry from seawater to freshwater, with K_d generally, but not always, higher in freshwater than seawater (Table 5.1). The fraction of the radionuclide present in each phase is often expressed in terms of the K_d and the solid (sediment) mass concentration (C_P), either as the ratio of the activity in the particulate fraction to that in the solution fraction (F_a):

F_a = K_d C_P

(5.2)

or by the percentage of radionuclide in the particulate fraction (P_P), per unit volume of suspension:

	100
P_P=		(5.3)
	[1+1/(K_d C_p)]

Sediment concentrations vary widely in aquatic environments, both spatially and temporally, e.g. from <1 to >1000 g m^-3. Characteristically, they are highest in high-energy environments such as shelf seas, tidal estuaries and rivers, especially during events which have low frequency of occurrence, such as storms. This variation means that the relative importance of the solid and solution phases in terms of radionuclide behaviour also varies widely in space and time. At one extreme, with high particulate concentrations, a large proportion of the radionuclides will be particle associated, even for low K_d elements. At the other extreme the converse will be true.


Element	Freshwater	Seawater

I	3 x 10²	10¹
Na	10²	10¹
Ru	10²	5 x 10³
Sr	10³	10²
Cs	10⁴	2 x 10³
Pu	10⁵	5 x 10⁴
Lanthanides	5 x 10⁶	5 x 10⁵

^aValues are the best estimates from Coughtrey et al. (1985).

The chemical characteristics of the water are important in determining the radionuclide behaviour and K_d. Radionuclide species in solution may depend on the ionic composition and ionic strength of the water, the presence of organic ligands, the redox state (Eh) and acidity (pH); an important contrast in nuclide behaviour is thus observed between aerobic and anaerobic waters. The major differences in ionic composition and ionic strength existing between saline water and freshwater, and within the latter between alkaline and acid waters, play an important role in determining radionuclide speciation and behaviour. Also, individual radioelements vary widely in the complexity of their solution chemistry, e.g. Pu may exist as three or four oxidation states and many ionic species, whereas Cs is a single monovalent ion.

The sediment grains (and bedrock grain surfaces) differ in their potential for adsorbing radionuclides, which results in a K_d variation with lithology, i.e. grain size and composition. In addition, the nature of the grain surface can be important, i.e. presence of organic and Fe/Mn oxide coatings. Minerogenic sediment grains, mainly silicate minerals or rock fragments, are derived from the crustal rocks and their derivative soils and sediments. These are eroded by rivers, in shallow seas and estuaries and are transferred from one aquatic environment to another. A proportion of this minerogenic matter is provided via atmospheric deposition. Another component is of biogenic origin, produced partly in the aquatic environment and partly introduced by terrestrial erosion. This includes skeletal minerals (mainly carbonates) and organic matter. Chemical precipitates form a hydrogenous component and can be locally important. Direct anthropogenic inputs of sedimentary material also occur, e.g. sewage solids. In addition, the solid phase may be primary, e.g. fuel debris. The size and density of these sediment grains must be taken into account for determining radionuclide behaviour. Typically, there is a wide range of grain sizes present in sediments. These include mud (silt + clay), which extends in size from an arbitrary boundary with the `solution' phase (usually 0.2 or 0.4 µm) up to 0.06 mm, sand (0.06

2 mm), gravel, etc. (> 2 mm). In addition, the presence of colloids of sub-micrometre size can be important for radionuclide pathways. Because of the increase in surface area per unit mass as grain size decreases, the fine-grained sediments have a higher adsorptive capacity for radionuclides than coarse, i.e. muds have higher activity concentrations than sands. Grain size also determines the potential mobility of the radionuclide particulate fraction. In terms of their dynamic behaviour, sediments are grouped into granular sediments (sand and gravel) and cohesive sediments (mud). The grain size is broadly related to the energy of the environment. In many situations, the fine-grained sediment is present not as individual grains but as biological aggregates (faecal pellets) and chemical flocs, which substantially change its dynamic behaviour. Occasionally, to facilitate comparison of activity concentrations of sediments with varying grain size, the concentrations are normalized to the Al, Sc or ⁴⁰Kcontent. This assumes that these elements are mainly contained in clay minerals and that their average concentration in these minerals remains roughly the same. These assumptions may not be valid in shallow marine and terrestrial aquatic environments, especially in glaciated terrains where Al- and K-bearing minerals may be common in coarse sediments (e.g. feldspars).

The characteristic features of radionuclide behaviour in the aquatic environment are the redistribution by transport of the solution and solid phases, the chemical interactions between the phases, and their biological cycling. Transport, in addition to leading to radionuclide redistribution, results in their dilution, fractionation and mixing, as well as vitally affecting the residence time in the aquatic environment. In addition, in the surface environments, accumulation of the particulate phase occurs by sediment deposition (sedimentation). This induces post-depositional chemical and physical changes, called diagenesis, which affect the radionuclide distribution.

The transport of the water results from a number of driving mechanisms, all ultimately a response to gravitational forces, modified by Coriolis and friction forces. Water velocities vary widely in space and time in aquatic environments, from as low as µm s^-1 in groundwaters to m s^-1 in rivers, tidal seas and estuaries and, also, in deep ocean turbidity currents. A basic contrast exists in surface waters between current regimes that, at least occasionally, are strong enough to erode sediment from the bed and those which are predominantly weak. The former include river flow, tidal currents and wave base oscillatory currents and the latter, ocean circulation, coastal and estuarine saline density currents and water surface elevation compensating currents, set up by wind shear and wave drift in coastal waters and lakes. Weak currents, however, can be important for transport of sediment introduced to the water column by another process. Transport of radionuclides occurs in these circulation/current/flow systems as a result of advection and dispersion. Advection is produced by the time averaged flow of water. Dispersion is due to a number of processes: molecular diffusion, turbulent eddy diffusion and dispersion due to velocity shear, i.e. the spreading that occurs in the direction of flow as a result of the vertical (and lateral) velocity gradients. The combined effects of the dispersion processes are described by dispersion coefficients applicable to the perpendicular axial directions: D_x, D_y, D_z. The magnitude of the dispersion coefficients varies with velocity, turbulence intensity and secondary characteristics of the different aquatic environments. Transport (flux) of a radionuclide by dispersion processes is related to the magnitude of the concentration gradient and the dispersion coefficient and takes place in the direction of the gradient.

Sediment and associated particulate phase radionuclides respond to the same advective and diffusive circulation processes as the solution phase, resulting in their transport and dilution. However, the sediment and particulate phase response is fundamentally different because a velocity-related threshold (bed shear stress) has to be exceeded before transport occurs, whereas this is not the case for the solution phase. Also, above this threshold, the concentration of mobile particles and their vertical distributions in the flow are dependent partly on velocity-related flow characteristics such as bed shear stress and turbulence intensity. These complex relationships to the flow depend on particle size and density. The state of aggregation of fine particles has a very important influence on this behaviour. As the capacity of the flow to transport sediment decreases, in space or time, above or below the erosion threshold, the excess sediment is deposited on the bed. However, because it takes a finite time for grains to settle through the water column there is a lag between the change in flow and deposition, which can have considerable importance for sediment behaviour. Settling rates vary from µm s^-1 to cm s^-1. Sediment can be transported either in suspension (suspended load), at velocities comparable to that of the water, or in contact with the bed (bedload) at a fraction of water velocity, as mobile bed forms such as ripples, dunes and bars. The sediment in suspension may be actively suspended by turbulence or may settle passively through the water after being introduced by another mechanism. The range of velocity regimes that can be found in aquatic environments means that either one mode of sediment behaviour may predominate, e.g. grain settling in oceans and lakes, or conditions may vary in space or time between various transport modes, e.g. in rivers. An important modifying factor is the availability of sediment: supply limitations may over-ride flow-imposed limits on sediment transport. Net deposition of sediment and associated radionuclides occurs where rates of sediment supply exceed those of transport. Such sediment deposits are extremely important as sinks or reservoirs for radionuclides. The radionuclide inventories preserved in aquatic sediment deposits vary widely, depending on radionuclide sources and on the deposit characteristics, such as lithology, sedimentation rates and residence times. The grain size dependence of sediment transport processes leads to the fractionation of sediment inputs and the production of deposits of different grain size and, hence, of different radionuclide concentrations. Certain environments are characterized by the lithological uniformity of their deposits, e.g. the mainly fine muds of oceans and lakes, whereas others, typically, have a more diverse range of sediment types depending on the local current regime, e.g. rivers and estuaries. Sediment accumulation (sedimentation) rates vary widely, from < mm y^-1 to > m y^-1 and usually correlate with grain size, with low rates in oceans and lakes and variable rates in other environments.

The sedimentation rates can be determined from radionuclide profiles. This can involve conventional radiometric methods based on the decay of an unsupported natural or artificial radionuclide such as ¹⁴C or an unsupported fraction, e.g. excess ²¹⁰Pb. Alternatively, artificial radionuclide profiles can be matched with the record of release of the nuclide, e.g. ¹³⁷Cs from weapons test fallout and, increasingly of use in the future, from the Chernobyl accident. The application of these methods can be affected by processes such as diagenetic, chemical or physical redistribution of the radionuclide and, for artificial radionuclides, the effects of processes which modify the relationship between the initial release to the environment and the signal received in the sediment, such as mixing of sediments labelled at different times. The residence times for sediment deposits also vary widely, from >100 to < 1 y. In short-lived deposits, sediment remobilization leads to reintroduction of particulate radionuclides to the water column. If subaqueous, the deposits are saturated and the pore waters will contain solution phase radionuclides.

In contrast to the high energy environment of coastal zones, transport of radionuclides in the open ocean is generally mediated by biological cycling. The various materials that have adsorbed or incorporated these nuclides, such as biogenic debris and clay, are packaged by zooplankton and can be transferred to the sediment in a few months. Other biological processes, such as aggregation of phytoplankton material, also lead to a rapid transfer of radionuclides to great depths. In the water column, and sediment and bedrock porewaters, the biogeochemical interactions listed earlier can lead to repartitioning of radionuclides between the solid and solution phases, in response to changes in the chemical environment and/or to biological mediation. Particularly important in surface waters are sorption

desorption reactions involving the sediment-bound radionuclides, caused by changes in salinity or redox state, or to the introduction of unlabelled or partially labelled sediment. In addition, physical change in the particulate phase can lead to the same result, e.g. by colloid aggregation/disaggregation or by organic particle degradation. Gain by the solution phase will increase the mobility of a radionuclide.

Transport by diffusion can occur also across sediment/water column interfaces, with loss or gain of activity by the deposits. Other post-depositional (diagenetic) processes in sediments can lead to the physical disturbance or modification of the original sediment and include the effects of biological disturbance by organisms, called bioturbation, and physical processes such as slumping. The effects of bioturbation depend on the number, size and habits of the organisms; they include random mixing of sediment or grain size sorting. These processes also affect the chemical behaviour of radionuclides, mainly via oxygenation of interstitial waters in the burrows and changes in alkalinity due, for instance, to carbon consumption.

Advective movement of porewaters and solution phase radionuclides can occur in sediments (in addition to bioturbation) as a result of consolidation due to the lithostatic stress generated by the overlying sediment. This results in an increase in sediment bulk density, decrease in pore volume and upward migration of porewater. In sediments, and importantly in groundwaters, porewater flow is also a response to regional hydrostatic gradients.

Radionuclide uptake by biota occurs by a number of mechanisms from both solution and particulate phases. Uptake by the primary producers, e.g. phytoplankton, occurs from solution by surface adsorption and metabolic processes. Surface contamination by particulates can also occur with macro-algae. The primary uptake mechanism for invertebrate and vertebrate organisms is ingestion of food. However, for the many invertebrates which are detrital filter- and sediment-feeders, this directly involves particulate radionuclides in general. Respiration also involves intake of solution phase (and particulate) radionuclides. Radionuclide contamination of terrestrial organisms can also occur by feeding on aquatic organisms. Uptake will depend not only on the organism concerned but also on the element involved and its activity concentration. The ratio between the activity concentration in the organism and aquatic environment is defined as the concentration factor.

These different processes affecting the behaviour of radionuclides in aquatic media are illustrated in the following sections.

5.2 FRESHWATERS

Solute transport in rivers is generally described in terms of a one-dimensional, partial differential equation, incorporating advection and eddy diffusion terms, although alternative statistical or systems approaches are receiving increasing attention (Young, 1990). Because of the complex nature of river channel geometry, transport models require detailed calibration. Artificial radionuclides can be used as tracers for studying river flows and for model calibration. Tritium is ideal for this purpose but is rarely used in practice. The main reason for this is the generally dispersed nature of the source term (bomb fallout), with a well-defined point source being the exception. An example of the latter is where tritiated water from the Grafenrheinfeld nuclear power plant has been used to define the dispersion characteristics of a 320 km section of the Main River, Germany (Krause and Mundschenk, 1989). Flow times, flow velocities and longitudinal dispersion (eddy diffusion) coefficients (20

200 m² s^-1) were determined as a function of river discharge from 0.4 to 4 times the mean runoff.

While there is an extensive literature on the relations between the hydraulic characteristics of flow and sediment particle behaviour, there is no coherent, mechanistically based approach to describing sediment transport and dispersion in river systems. A major contributing reason is that sediment transport by rivers is subject to non-hydraulic as well as hydraulic controls. Important non-hydraulic factors include the geology and soils present in the catchment, catchment topography, hydrology, land use and vegetation cover. A further complication is that many of the hydrological factors are stochastic in nature. These include storm duration and spatial effects, rainfall intensity, and antecedent discharge, all of which influence storm-period particle transport. There are also significant stochastic processes occurring within the river channel, e.g. an important source of sediment supply is through bank collapse.

Studies of artificial radionuclides in rivers have frequently highlighted the complex nature of sediment transport and the individuality of rivers. The Great Miami River, Ohio, has been studied over a 25-fold range of river flows and a 6-fold range of particle concentrations (Sprugel and Bartell, 1978). The activity concentrations of ^239,240Puin solution (Bq m^-3 ) and in suspended particles (Bq kg^-1) did not correlate with either river flow or suspended particle concentration. The total Pu inventory was largely a function of suspended particle concentration. The mean concentration of ^239,240Pu in riverborne suspended sediment was 2 to 3 times that in the source material (arable soils containing fallout Pu). The enhanced activities were attributed to size fractionation in the catchment-river system, resulting in a greater fine-grained component in the river. Studies of periodic, pulsed inputs of plutonium to the Great Miami River from the Mound Laboratory in Miamisburg, Ohio, have also given important insights to the dispersion behaviour of Pu (Muller et al., 1977). Mass balance calculations showed that under typical summer flow conditions, ~60 per cent of the effluent is lost through sedimentation within 10 km of the discharge point, and that resuspension of this material between pulses maintains a high `background' ²³⁸Pu flux in the river. The remaining ~40 per cent is transported downstream with each pulse.

Further understanding of the temporary nature of channel sediments comes from a study of the river Danube at Bezdan, near the Yugoslav

Hungarian border (Conkic et al., 1990). The concentrations of ¹³⁷Cs,¹³⁴Cs and ¹⁰⁶Ru in bottom sediments were studied following the Chernobyl accident. Decay-corrected activities showed exponential decreases with time that were the same for all three radionuclides, despite expected differences in chemical behaviour. This highlights the dynamic nature of the sediment deposits and suggests that the relative importance of physical processes was the same for all three radionuclides. Deployment of a sediment trap, higher up the Danube, at Vienna, provided evidence not only of a gradual reduction in Chernobyl activity with time in newly deposited sediments, related to the gradual downstream migration of the most contaminated sediments, but also of marked seasonal variations in activity (Maringer et al., 1989). As expected, the total activity content was influenced by the grain size composition of the sediments, which in turn was related to river discharge. The activity concentration of the < 20 µm fraction was between two and four times that in the bulk sediment. High discharges, related to the passage of spring meltwaters, resulted in a lowering of the ¹³⁷Cs activity in the newly deposited sediment, due in part to an associated coarsening in grain size but also due to a decrease with time in the radionuclide content of the fine-grained fraction (Figure 5.1). The subsequent fall in river flow produced a fining of the sediment in transit and a consequent increase in its radionuclide concentration.

Figure 5.1 Relation between river discharge and the seasonal trend in radiocaesium activities in total sediment and the <20 µm fraction. From Maringer et al. (1989).

Studies, using Chernobyl-derived radiocaesium as a tracer, have also begun to tackle some of the more intractable aspects of the dispersion of riverborne sediments, such as those related to bedload transport and the formation of floodplain deposits (e.g. Walling and Bradley, 1988).

Mass transport of a solute through and within a lake may also be described in terms of advection and eddy diffusion, but the complexities of water motions are far greater than those in rivers. Dynamic processes can be classified according to origin: wind-induced currents including seiches, inflow-induced currents, and convective currents. For large lakes there are, in addition, significant currents produced by Coriolis force, gravitational forces, frictional forces, and the effects of meteorological variations. The need to take account of the various types of current likely to be discernible and the distinction between the advection and eddy diffusion terms are very much dependent on the time and length scales of interest. A particularly important feature of lakes is the degree and temporal dependency of vertical density stratification, which generally is related to temperature but can be related to differences in dissolved salt content (i.e. in meromictic lakes). Transport by molecular diffusion is rarely important in the water column, except possibly associated with the boundary layer (~1 mm) overlaying the bottom sediments and vertically in the thermocline region.

Artificial radionuclides have rarely been used in studies of water movements in lakes. A contributing factor is that the complex temporal variation of the input function of fallout ³H, the radionuclide most frequently used in such studies, occurs on similar timescales to typical mixing times in lakes, thus greatly complicating interpretation. In a study of Lake Tahoe, no significant variation with depth in fallout concentrations was observed, despite it being a deep lake (max. 500 m) with a long hydraulic residence time (Imboden et al., 1977). Far greater success has been achieved through employing experimental injections of tritiated water (Quay et al., 1980).

A more widely applicable approach is to measure the parent

stable daughter tracer pair, ³H

³He (Torgersen et al., 1977). The method is based on the fact that during periods of isolation, water bodies build up an excess of 'He, relative to its equilibrium concentration with respect to the atmosphere, due to the in situ decay of ³H. Consequently, an effective water mass age can be determined and this has been used to calculate (a) gas exchange rates between surface waters and the atmosphere or, more specifically, the piston velocity and hence the surface boundary layer thickness; (b) the extent of degassing or gas renewal at turnover; (c) vertical diffusivities in the well-mixed surface waters; (d) transport across the thermocline during stratification; and (e) renewal times for water masses at different depths. Vertical eddy diffusivities of 0.6

3.6 cm² s^-1 have been estimated for the well-mixed surface waters of Lakes Erie, Huron and Ontario, and an apparent vertical diffusivity of 1 cm² s^-1 for the thermocline region in Lake Constance. The data indicate also that, whereas degassing of the deep waters is virtually complete during turnover in the three Great Lakes, a degassing rate of only ~50 per cent per year occurs in the deep waters of Lake Constance.

Accidental releases of elevated ³H from nuclear power plants have been used as tracers. For example, releases from Douglas Point and Bruce A power stations in Ontario have been used to study the counter-clockwise circulation of Lake Huron (Veska and Tracy, 1986).

In contrast to rivers, lakes are generally regarded as being efficient and permanent sediment traps, related principally to their greater depths, smaller currents and longer hydraulic residence times. Thus lakes are short-lived on the geological timescale, disappearing from the landscape principally through the action of sedimentation. In all but very shallow lakes, bottom sediments away from the high-energy shoreline region are normally subjected to low current velocities and negligible wave action. The transport and fate of sediments within lakes, however, is far from simple because of the many competing processes regulating erosion, transport and deposition. In areas dominated by river action, grain size and the rate of sedimentation generally decrease logarithmically with distance from the river mouth. The distance over which river flow is important varies greatly, depending on factors such as basin topography, stratification, and river discharge. At a distance from the river mouth, wind and wave action become the dominating influence. The associated rate of net sedimentation increases with water depth from zero at some intermediate depth, due to sediment focusing, and involving the resuspension of shallow-water sediments and their transport to deeper waters.

The transport of sediment particles in lakes is studied predominantly through an examination of bottom sediments, rather than direct observation in the water column. Artificial radionuclides with high particle affinities have proved to be useful as tracers in such studies, especially as dating tools providing a measure of sedimentation rates. The first appearance of fallout ¹³⁷Cs, and to a lesser extent ^239,240Pu, in the early 1950s and peak fallout in 1963 have become particularly important stratigraphic markers in lake sediments. In Europe, ¹³⁷Cs resulting from the Chernobyl accident has become a useful additional time marker (1986). Edgington and Robbins (1975) employed the bomb fallout markers, together with ²¹⁰Pb dating, to determine sedimentation rates in Lake Michigan and found good agreement between the three methods. The sedimentation rate in the southern part of the lake varies by more than an order of magnitude from virtually zero in large areas of the western side to >50 mg cm² y^-1 in areas close to the eastern shore, where several large rivers discharge (Figure 5.2a). Radionuclide fluxes at the sediment surface were normalized to fallout fluxes at the air

water interface, which represents the predominant source term. A map of the distribution of this function provides a measure of the relative sediment loss or gain at any point (Figure 5.2b). Particularly high sedimentation occurs offshore to the north and west of the major rivers, whereas bottom topography slopes continuously down to the central axis of the lake. No sedimentation occurs at water depths <50 m, even near the river mouths. This pattern of deposition, which is very different from the classic picture of focusing and delta formation described above, was attributed to the hydrodynamic characteristics of the lake and in particular the pattern of lake-wide currents.

More than 95 per cent of the ¹³⁷Cs and ^239,240Pu that has entered southern lake Michigan as fallout now resides in the sediments, indicating that the efficiency of sediment trapping is very high. This is not always the case, however, and studies with artificial radionuclides have provided data to assess the efficiency of trapping and its main controlling factors. Fallout^239,240Pu has been studied in the sediments of seven lakes in eastern Ontario, Canada, with a large range of mean depths (2 to 19 m) and hydraulic residence times (0.03 to 9 y) (Cornett and Chant, 1988). From 28 to 100 per cent of the ^239,240Pu input to the lakes was retained in the sediments, with the percentage retained being correlated with the hydraulic residence time. In other words, in the lakes with short hydraulic residence times, a significant fraction of the Pu and hence sediment inputs must have remained in the water column long enough or have been resuspended frequently enough to be lost via the outflow. The data also indicated the absence of any sediment focusing, which the authors linked to the shallow nature of the lakes.

Figure 5.2 Sediment characteristics in southern Lake Michigan: (a) mass sedimentation rate (mg cm^-2 y^-1); (b) flux normalization factor for ¹³⁷Cs (sediment flux/flux at lake surface). From Edgington and Robbins (1975); reproduced by permission of the IAEA.

Similar situations exist in Lake St Clair, located between Lakes Huron and Erie, and in a number of man-made reservoirs in the Susquehanna river valley. The latter study (Donoghue et al., 1989) was based on a range of particle-associated tracers (⁶⁰Co, ¹³⁴Cs, ¹³⁷Cs and ⁶⁵Zn), derived from the nuclear power plants at Three Mile Island and Peach Bottom. It is apparent, therefore, that shallow lakes with short hydraulic residence times may be regarded as being intermediate in character between river channels and more typical lake basins in terms of their sediment dynamics.

Artificial radionuclides have also been used as sediment tracers to study other, diverse sedimentological problems. One example involves the sedimentological regime that exists at the eastern end of Lake Erie, which is a high-energy region with a substrate comprising bedrock and coarse-grained sediment. Fine-grained sediment is thought to be transported northwards to the Niagara river, and subsequently Lake Ontario, by the dominant advective flow in this part of the lake. An important radionuclide source, the Western New York Nuclear Service Center (WNYNSC), is located about 50 km south of Buffalo, NY. Water from the site eventually drains into Cattaraugus Creek and thence Lake Erie. Radionuclide inventories and ratios, involving ²³⁸Pu, ^239,240Pu, ²⁴¹Am, ⁵⁴Mn,⁶⁰Co, ⁶⁵Zn,¹⁰⁶Ru,¹³⁴Cs and ¹³⁷Cs have been used to demonstrate the presence of WNYNSC radionuclides in Lake Ontario sediments off the mouth of the Niagara River (e.g. Joshi, 1988). In this region, about 36 per cent and 80 per cent of the 1982 sediment inventories of ^239,240Pu and ²⁴¹Am respectively, are derived from WNYNSC, the remainder coming from atmospheric fallout. The studies also show that nearly all the Pu, ²⁴¹Am and ¹³⁷Cs activity, derived from WNYNSC, ends up in Lake Ontario sediments around the mouth of the Niagara River. These findings are an important illustration of the highly specific nature of artificial radionuclides as tracers. More particularly, they demonstrate that transport processes in surface waters can result in locally concentrated radionuclide activities at sites remote from source, even when separated by a lake basin, which would normally be associated with efficient removal of particle-associated radionuclides.

Another example involves the use of ¹³⁷C s and ²¹⁰Pb profiles in sediments to understand the processes operating in a series of small arctic lakes with depths < 10 m (Hermanson, 1990). Interpretation is based mainly on the inventory in each core relative to decay-corrected fallout. Sediment redistribution processes are shown to vary widely over short distances and include ice rafting effects.

The occurrence of groundwater may be divided into zones of aeration and saturation. The former zone consists of pore spaces occupied partially by water and partially by air. In the latter zone all pore spaces are filled with water under hydrostatic pressure. Over most of the land masses of the earth a single zone of aeration overlies a single zone of saturation and extends upward to the ground surface. In the absence of overlying impermeable strata, the upper surface of the zone of saturation is the water table. This is defined as the surface of atmospheric pressure and would be revealed by the level at which water stands in an open well. Water within the ground moves downwards through the unsaturated zone under gravity, whilst in the saturated zone it moves in a direction determined by the local hydraulic situation. Most natural groundwater discharge occurs as flow into surface water bodies but also occurs via the land surface, including transpiration from vegetation.

Due to advection, non-reactive solutes are carried at an average rate equal to the average linear velocity of the groundwater. However, there is a tendency of the solute to spread out from the path expected due to the advection hydraulics of the flow system. This is called hydrodynamic dispersion and causes dilution of the solute. It occurs because of mechanical mixing during fluid advection and because of molecular diffusion. The latter is significant only at low velocities. Dispersion caused entirely by the motion of the fluid is known as hydraulic dispersion. Dispersion is a mixing process and, qualitatively, has a similar effect to turbulence in surface-water regimes. For porous media, the concepts of average linear velocity and longitudinal dispersion are closely related. Longitudinal dispersion is the process whereby some of the water and solute molecules travel more rapidly than the average linear velocity and some travel more slowly. Thus, the solute spreads out in the direction of flow and declines in concentration.

Generally, tritium can be used to indicate the presence of young groundwaters (less than 30 years). At high continental latitudes, waters having more than about 4 TR fall in this category, while in low and middle marine latitudes the limit is nearer l TR. As a general guideline, it can be said that waters containing over 10 TR contain a thermonuclear test contribution, while 20 TR or more would suggest a component of groundwater recharged since 1961 (IAEA, 1983). On a local scale, ¹⁴C can be used as a tracer of transfer processes from surface systems to either plants or groundwater. A large number of ¹⁴C groundwater studies exist in the scientific literature (e.g. Fontes and Garnier, 1979).³⁶Cl has virtually ideal properties as a tracer for solutes in groundwater and soil water. ³⁶Cl from the Gnome event near Carlsbad, New Mexico (the first nuclear detonation of the Plowshare series, detonated 10 December 1961) illustrates how ³⁶Cl can be used to study the redistribution of radionuclides in the soil profile (Phillips et al., 1990). Atmospherically derived ⁸⁵Kr activity is used in hydrology to identify admixtures of young groundwaters recharged during the last 35 years. The use of ⁸⁵Kr is particularly attractive in combination with ³H, which has a similar half-life but a different history of release.

There are a number of examples where tritium contamination from a localized source has been used to trace groundwater movement and to study surface water

groundwater interactions. Liquid-waste effluents are routinely discharged to the ground at the Hanford Site in Washington State (see site description in Chapter 2). Adsorption, chemical precipitation, and ion exchange attenuate or delay the movement of some radionuclides, such as ⁹⁰Sr, ¹³⁷Cs, and ^239,240Pu. Other radionuclides, such as tritium, ⁹⁹Tc, and ¹²⁹I, are not as readily retained by the soil. These radionuclides move through the soil column at varying rates and eventually enter groundwater. Radionuclide concentrations are reduced by dilution when they reach groundwater. The more mobile constituents move downgradient in the same direction and at a rate nearly equal to groundwater flow.

The maximum extent of radionuclide contamination in the groundwater beneath the Hanford Site can be defined using tritium. Figure 5.3 shows the distribution of tritium in the unconfined aquifer during 1989 (Evans et al., 1990). The highest groundwater tritium concentration measured in 1989 was over 5 x l 0⁶pCi l^-1 in a well in the 200-East Area. The large tritium plume from the 200-East Area moves generally east to the Columbia River, where it discharges. Separate tritium pulses, associated with two major episodes of the Plutonium

Uranium Extraction Plant (PUREX) operations, can be distinguished as lobes with concentrations exceeding 2 x 10⁵ pCi l^-1. The lobe near the Columbia River is a result of discharges to ground during the operation of PUREX from 1956 to 1972. The narrow lobe extending several kilometres south-east of the 200-East Area represents discharges from PUREX between 1983 and 1989.

The travel time of contaminated groundwater from source areas to the Columbia River and the mass of radioactive contamination that is reaching the river have not been determined with a high degree of accuracy. A recent estimate by the US Geological Survey (1987) stated the average travel time is in the range of 10 to 20 years. Contaminants also enter the river along the Hanford Reach as direct effluent discharges. In 1989, the average tritium concentration upstream of the Hanford Site was 63 pCi l^-1, while the average downstream concentration was 129 pCi l^-1, thus substantial dilution occurs when groundwater, containing tritium at concentrations greater than 2 x 10⁵ pCi 1^-1 discharges into the river.

Seasonal rises in the Columbia River stage result in bank storage, which affects groundwater levels and contaminant concentrations in wells near the river. In the eastern part of the site, where tritium exceeds 2 x 10⁵ pCi l^-1, the concentration of tritium in groundwater from wells near the river has been documented to fluctuate as much as a factor of four over the course of a year. The highest concentrations correspond to periods of low river stage. The tritium concentrations are lower when the river stage is higher, during which times bank storage effects from the river dilute the tritium concentrations in the groundwater system.

Figure 5.3. Distribution of tritium on the Hanford Site. From Evans et al.

Another radionuclide migration project was started in 1974 at the Nevada Test Site to determine the potential for movement of radioactivity away from underground nuclear explosions (Ogard et al., 1988). The first field experiment in this project was a long-term single-well pumping test in which the activity of the nuclear explosion was treated as a slug-injection point and the explosion products as tracers. The radionuclides detected in the pumped water from the saturated zone of an alluvium aquifer were ³H and ³⁶Cl. The elution curve of ³⁶Cl preceded that of ³H, and the observed phenomenon was ascribed to an anion exclusion process (Thomas and Swoboda, 1970). Anions, being of the same charge as clays and zeolites in the soil, are repelled by these surfaces and are effectively prevented from entering into the intragranular porosity of the matrix.

A pulse of tritiated water, which was discharged accidentally from an isotope processing plant in the Glatt River Valley, northern Switzerland, was traced through a sewage treatment plant and various rivers and groundwater wells (Santschi et al., 1987). Tritium concentrations were used to test predictions for the transport of conservative anthropogenic trace contaminants accidentally discharged into the sewage system. Mass balance calculations indicated that about 2

10 per cent of the tritium pulse infiltrated the groundwater and about 0.5 per cent of the total pulse reached eight major drinking-water wells in the area. In spite of the complex hydrology of the lower Glatt River valley, tritium breakthrough curves could be simulated effectively by a model developed from an experimental well field.

The accident at the nuclear power plant at Chernobyl also provided an opportunity to investigate the infiltration and migration behaviour of radionuclides in the Glatt River valley. The radionuclides ^99mTc,¹⁰³Ru, ¹³¹I, ¹³²Te, ¹³⁴Cs and ¹³⁷CSwere measured several times during May 1986 in the river and an adjacent shallow aquifer (Waber et al., 1987). The main radioactivity (> 75 per cent) of the river water was found to pass through a 0.05 µm filter. Iodine, Ru and Te were found in the groundwater as a result of river water infiltration, being subject to only slight sorption on particulates and other solid surfaces. This was explained to be the consequence of having formed anionic or neutral species. In contrast, Cs was retained completely by river sediments. Particulate (> 0.05 µm fraction) infiltration from the river into the groundwater system was found to be a negligible process.

Determination of groundwater residence time from radionuclide data is possible only if the initial concentration is known and the radionuclide contents do not change as a result of mixing with isotopically different waters. Increases in content and changes resulting from solution/deposition processes normally preclude estimation of residence times. The estimation of the initial content in the groundwater is difficult although, in principle, a plot of radionuclide content as a function of distance along flow lines in a piston-flow system may permit deduction of the initial concentration and thus, interpretation of changes in terms of decay of the radionuclide through time. The closed system requirement in a groundwater system, however, is the most difficult to satisfy.

The measurement of the parent

stable daughter pair, ³H/³He allows the calculation of groundwater age. Tritiogenic ³He is added to the natural ³He content of the groundwater, and if it is assumed that no tritiogenic ³He is lost by diffusion across the groundwater table, the ³H/³He groundwater age is given by:

= t_1/2/ln 2

ln(2 + [³He]/[³H])

(5.4)

where

is the age in years and [³He]/[³H] the concentration ratio.This determination is independent of the initial tritium concentration. This is an apparent age of a water parcel if ³He sources other than ³H decay can be excluded or corrected for and mixing of isotopically different waters is negligible. An example of the application of the method is given by Schlosser et al. (1988). For the bomb ³H peak, the deviation of the ³H/³He age from the age determined by identifying the groundwater layer recharged between 1962 and 1965 was about 3 years (15 per cent). The deviation was explained by diffusive ³He loss across the water table and by flow dispersion.

In practice ¹⁴C dating can be applied in similar fashion to that of ³H dating but in a longer time range because of its longer half-life. Applications have included identifying ancient recharge, calculating velocity of groundwater movement by dating the water at different horizons in the aquifer, and calculating contributions of different components to groundwater blends. The age obtained is again an apparent age only and is subject to substantial corrections due to exchange of carbon in solution with carbon of rock minerals and due to solution and precipitation of carbonate compounds during the transit of water through the aquifer. Nevertheless, a groundwater containing no ³H and 30 PMC (per cent modern C) (apparent age 1 x 10⁴ y) can be designated as ancient recharge with some confidence. Furthermore, differences in apparent age at different points in an aquifer system can be used to calculate minimum groundwater flow rate even though the absolute ages may not be known precisely.

If there are no internal sinks or sources for dissolved chloride and ³⁶Cl in an aquifer system, apart from loss of ³⁶Cl to decay, the groundwater may be dated using the equation:

t = -1/

ln((RR_se)/(R_o

R_se))

(5.5)

where R is the measured ³⁶Cl/Cl ratio, R_othe initial ³⁶Cl/Cl ratio, and R_sesecular equilibrium ³⁶Cl/Cl ratio due to hypogene production. The method was comprehensively tested on old groundwaters from the Great Artesian Basin, Australia, against a hydraulic model (Bentley et al., 1986). The ³⁶Cl groundwater ages ranging from less than 1 x 10⁵ years to over 1 million years were obtained in excellent agreement with a hydrodynamic model, but only in areas of the aquifer in which no groundwater mixing occurred.

One of the principal reasons for the common usage of the distribution coefficient (K_d) is the need, in radiological and transport/dispersion models, for a simple term describing the biogeochemical behaviour of radionuclides. The K_dprovides a reasonable first approximation of this behaviour, but its associated limitations (see Section 5.1) must be recognized in any subsequent modelling exercise. More sophisticated solid-solution models (e.g. surface complexation modelling) do exist and have recently been applied to the behaviour of artificial radionuclides under environmental conditions (Turner et al., 1991).

The importance of suspended particle concentration, when assessing solid-solution partitioning, should not be overlooked. The partitioning of three representative radionuclides is used to illustrate this point, based on typical particle concentrations (Table 5.2). Suspended particle concentrations, in lakes are generally lower than those in rivers, so that a concentration of 1 mg l^-1 is commonplace. At this concentration, even a K_das high as 10⁵ would result in less than 10 per cent of a radionuclide being associated with the solid phase.


Environment		Particle concentration	Particulate fraction (%)

		(mg l^-1)	3 x 10²	10⁴	10⁵

Fine-grained bottom sediment		2.5 x 10⁵	99^a	>99.9^a	>99.9^a
Rivers
	extreme range	<l>10⁵	<197	<1>99.9	<9>99.9
	common range	10¹10³	0.323	1090	5099

^aBased on a sediment porosity of 90 per cent and a sediment dry density of 2.5 g cm^-3.

Thus, what seem to be contradictory statements in the literature can be explained in terms of sediment concentration. A high particle affinity is widely attributed to ¹³⁷Cs and, by way of illustration, its extensive use as a sediment tracer has already been discussed in Section 5.2.2.2. In a study of ¹³⁷Cs released to the Susquehanna River during operation of the nuclear power plants at Three Mile Island and Peach Bottom, the nuclide budget was converted to a sediment budget by assuming all the ¹³⁷Cs was in a sorbed state on sediment particles, which was justified by the high affinity of caesium for particles in freshwater (Donoghue et al., 1989). In contrast, following the Chernobyl accident, more than 90 per cent of the atmospherically derived ¹³⁴Cs in various European lake waters was found to be in solution (Santschi et al., 1987), with the relatively low sediment concentrations in these lakes undoubtedly being an important factor.

The solid-solution partitioning of radionuclides is also a function of the environmental history of particles, since this is likely to influence the form of the radionuclide in the solid phase and, consequently, the reversibility of the solid-solution reaction. Again this is well illustrated by reference to Cs. Sorption by micaceous clays (illite), when present in significant amounts, dominates the environmental behaviour of Cs. This association is attributed to the presence of frayed edge sites, corresponding to partially weathered 1 nm interlayers, which are highly selective for Cs⁺ because of its size and charge. With time, a significant fraction of the Cs becomes `irreversibly' bound by the illite, probably due to fixation at interlayer sites through collapse of the frayed edges and migration along the interlayers. For instance, less than 25 per cent of the ¹³⁷Cs is leached by 1 M HNO₃ from stream channel sediments, contaminated by effluents from the Savannah River Plant (Brisbin et al., 1974).

A number of approaches have been adopted to provide insights into the related issues of the nature of solid-solution reactions, the chemical forms of sediment bound radionuclides, and their remobilization characteristics under a range of environmental conditions. Some of the unique advantages of artificial radionuclides, as chemical tracers in the environment, are well demonstrated by the following examples. One approach has been the application of sequential extraction procedures. Förstner and Schoer (1984) employed the technique in a study of various river sediments, subjected to discharges from nuclear installations in Europe and the USA. They compared the leachability of the associated radionuclides with their naturally occurring stable counterparts and with artificial radionuclides added to the sediments in the laboratory. The results indicated that the extent and ease of leachability for each element decreased in the order: laboratory spiked radionuclides > environmental radionuclides > stable isotopes. This suggests an ageing process, perhaps involving incorporation into lattice positions and recrystallization, and emphasizes the need to consider the environmental history of sediments when predicting solid-solution behaviour. A sequential extraction procedure has also been applied to bottom sediments and sediment trap material from Lake Michigan (Alberts et al., 1989). ²³⁸Pu, ^239,240Pu and ²⁴¹Am were associated mainly with a citrate-dithionate extract, corresponding to the nominal hydrous oxide fraction and probably occurring as mineral coatings. In contrast, ¹³⁷Cs was almost totally associated with the most inert fraction of the sediment, attributed to fixation by clay minerals.

The specific role of Fe

Mn coatings has been examined by a different approach, involving a series of novel field experiments (Cerling and Turner, 1982). The study involved placing contaminated gravels in uncontaminated streams, and vice-versa, within the White Oak Creek watershed, Tennessee. Fe

Mn coatings were known to form on stream gravels throughout much of the watershed, part of which is contaminated by several nuclear plants and a low-level radioactive waste disposal site. ⁶⁰CO was rapidly (e.g. days) sorbed from solution, principally by the Fe

Mn coatings.

Under oxidizing conditions, the sorption process was not reversible over a period of months, although some ⁶⁰Co was lost by abrasion of the Fe

Mn coatings. Under mildly reducing conditions, ⁶⁰Co dissolution accompanied reductive dissolution of the Mn oxide fraction of the coatings, while the Fe oxide component continued to form. Thus the behaviour of ⁶⁰Co was dominated by its well-known geochemical association with Mn oxide, probably through co-precipitation. ⁹⁰Sr was held primarily as an exchangeable cation, although some was associated in a non-exchangeable form with newly precipitated Mn oxide. Adsorption, together with desorption of the exchangeable fraction, was rapid (days), whereas only about half the remaining fraction had been lost after 8 months. ¹³⁷Cs was adsorbed irreversibly by illite, the dominant mineral, as discussed previously.

The examples described are only illustrative, so that the patterns of behaviour that emerge for ¹³⁷Cs and ⁹⁰Sr should not be regarded as being universal. Under the variety of conditions that exist in freshwater systems, different solid-solution behaviours must be expected. There is evidence, for instance, that the solid-solution behaviour and sedimentation of ¹³⁷Cs in marl lakes is dominated by scavenging, associated with calcite precipitation (Lindner et al., 1989). Any process leading to a change in the solid phase composition is likely to affect K_d, and surface water systems are particularly prone to such changes.

The integration of ideas on solid-solution partitioning and other important processes into whole-lake studies is rare. This deficiency has most successfully been addressed by Santschi and co-workers through the use of experiments in which limnocorals and whole lakes have been spiked with artificial radionuclides (e.g. Santschi et al., 1986). Such studies have highlighted that radionuclides are removed from lakes not only by hydraulic flushing and scavenging by sedimenting particles but also through direct adsorption to surface sediments. The latter pathway is dominant for less particle-reactive elements, and it is interesting to note that Cs can fall into this category at the low suspended particle concentrations normally present in lakes (see also above discussion). Whole-lake studies, employing Chernobyl ¹³⁷Cs as a tracer, have confirmed the potential importance of direct adsorption as a removal pathway (Santschi et al., 1990).

The chemical speciation of an element in aqueous solution is important in determining its solid-solution partitioning and general biogeochemical behaviour. As in the case of Cs, which occurs predominantly as the hydrated Cs⁺ ion, speciation may vary only slightly in natural waters and consequently not require special consideration. On the other hand, knowledge of speciation, and more importantly how it varies, might be critical to understanding and predicting the behaviour of an element. Pu is a good example and is therefore discussed at some length.

Pu can exist in aqueous solution in any of four oxidation states

III, IV, V, and VI. A number of novel methods have been developed specifically to determine the oxidation state, the most widely used being that based on the separation of `reduced' Pu(III+IV) from `oxidized' Pu(V+Vl) by co-precipitating the former with neodymium fluoride. Dissolved Pu in freshwaters is invariably a mixture of the two forms. Consideration of the available thermodynamic data led to the suggestion that the oxidized form occurred mainly as Pu(VI). More recently, there has been a growing consensus, based in large part on studies involving the selective adsorption of Pu(VI) by silica gel and on the selective co-precipitation of Pu(V) with calcium carbonate (Orlandini et al., 1986), that Pu(V) is the stable `oxidized' form in natural waters, principally as PuO₂⁺. A comparable situation exists in the case of reduced Pu. Pu(IV) is generally regarded as being of greater environmental significance, although thermodynamic data suggest that Pu(III) may be stable under anoxic conditions, particularly at low pH.

We saw in Section 5.2.4.1 that Pu has a high sediment affinity. There are a number of features of the sorption process that are linked to Pu speciation. Firstly, adsorption to natural particles is almost exclusively via the reduced state, and secondly, natural particles appear to take an active role in redox reactions. A series of experiments with Lake Michigan waters (Nelson et al., 1987) showed that when natural particles were present redox equilibrium was attained within two days and that the same oxidation state distribution was achieved whether Pu was added initially in a reduced or oxidized state. In the absence of the natural particles, equilibrium was not achieved after four weeks. With particles present, not only did the oxidation state distributions converge, irrespective of initial oxidation state, but the final equilibrium distribution was the same as that in situ in the lake, associated with fallout Pu. The role of natural particles in the reduction of Pu(V) to Pu(III,IV) has been studied in more detail by Penrose et al. (1987). Rapid reduction occurred only in the presence of natural sediment (from Lake Michigan), with the rate appearing to be nearly first order with respect to Pu(V) and being proportional to sediment concentration. A small number of sites with homogeneous properties seemed to be responsible for the reaction. Direct microbiological mediation is unlikely, since reduction was unaffected by preheating the sediments to 105°C. Ashing the sediments at 500°C, however, decreased the rate of reduction by 98 per cent, suggesting that organic matter may be involved in the reduction process.

The other dominant factor in terms of Pu speciation is dissolved, or more correctly colloidal, organic matter (COC), comprising mainly humic substances derived from decaying organic matter of terrestrial and aquatic origin. Field measurements and laboratory experiments indicate clearly that Pu(III,IV) is complexed by COC. Figure 5.4 shows the K_d of reduced Pu as a function of COC concentration, determined in laboratory experiments with spiked Pu, a natural sediment and a variety of freshwaters. The critical COC concentration in each curve, above which K_d starts to decrease significantly, ranges from about 0.1 mg l^-1 to about 3 mg l^-1 The other, less well-defined effect of COC is that it appears to influence the redox behaviour of Pu, probably by acting directly as a reducing agent. Field and laboratory studies indicate that the concentration of oxidized Pu is high only when colloidal organic matter is low. The effects of Pu(III,IV)

COC complexation, Pu(V,VI) reduction and the contrasting adsorption affinities of the two oxidation forms work in combination to produce a fairly coherent pattern of behaviour (Figure 5.5). Up to about 10 mg l^-1 COC, dissolved concentrations and hence the K_d of total Pu stay more or less constant, since the effects of Pu(III,IV) complexation and Pu(V,VI) reduction seem to cancel each other out. Above 10 mg l^-1 COC, Pu(V, VI) has all but disappeared from solution so that Pu(III,IV) complexation dominates and the total Pu K_d begins to decrease markedly (by up to two orders of magnitude, as has been observed in freshwater systems).

Figure 5.4 Variation of the K_d of Pu(III/IV) as a function of colloidal organic carbon concentration in a series of laboratory experiments, using natural freshwaters, as indicated, and sediment. The natural COC concentrations are indicated by arrows. From Nelson et al., 1987.

Less attention has been paid to other artificial radionuclides with regard to the direct measurement of speciation. Nelson and Orlandini (1986) have shown experimentally that ²⁴¹Am, which invariably occurs as Am(III) in natural waters, is complexed by colloidal organic matter in the same manner as Pu(III, IV). Orlandini et al. (1990) studied Lake Trawsfynydd in North Wales, which receives waste discharges from a Magnox reactor plant, and found that all the dissolved actinides measured (^239,240Pu(IV), ²⁴¹Am, ²⁴⁴Cm and ²³²Th) were associated mainly with the colloidal fraction. Pu(V) made up less than 25 per cent of the dissolved Pu and, in contrast, appeared to be in true solution. A common approach to the question of speciation in environmental settings is to rely on thermodynamic predictions, e.g. with ¹³¹I and ^103,106Ru.Iodine may occur in the

I, 0 and V oxidation states and Ru as Ru(III) or Ru(VI). Both elements can occur as oxyanions (IO₃^- and RuO₄^2-) and I as iodide. Their frequently observed mobility in surface and groundwater systems is generally attributed to the low adsorbance of anionic species, since natural particles themselves generally have a negative surface charge, especially in the presence of organic coatings.

Figure 5.5 Concentrations of dissolved forms of plutonium as a function of colloidal organic carbon concentration in a laboratory experiment, using data from ANL Pond experiment shown in Figure 5.4. Pu concentrations are normalized to a constant plutonium concentration in the solid phase of 1 pCi g^-1. From Nelson et al., 1987.

5.2.4.3 Redox boundaries as critical features in the biogeochemical behaviour of artificial radionuclides

As we have seen, redox reactions in part determine the speciation and hence the mobility and fate of many artificial radionuclides in freshwater systems. The most important, but by no means only, driving force for such reactions is the microbiologically mediated, oxidative breakdown of organic matter by various oxidizing agents. The radionuclides may themselves undergo redox transformations or may be affected indirectly (e.g. by being associated with redox sensitive phases, such as Fe and Mn oxyhydroxides). The redox conditions in freshwater systems depend largely on the rates of reaction and the mixing processes affecting the availability of reactants. Reducing conditions are more likely to occur in soil waters and sediment porewaters, than in surface waters, due to high net respiration rates and restricted oxygen supplies. In river waters, reducing conditions are generally restricted to badly polluted stretches of river, in which the net rate of respiration exceeds reaeration. In lakes, reducing conditions exist mainly in bottom waters isolated by stratification, and in groundwaters they may be linked to high rates of respiration or the long hydraulic residence times. It is not uncommon for the redox potential in groundwaters to decrease with distance along the flow path, in other words with the length of time since the water was last in contact with the atmosphere.

The best-studied situation is the redox boundary that generally occurs in the vicinity of the sediment

water interface in lakes and reservoirs, principally because it is relatively stable and easy to observe with minimal disturbance to the ambient redox conditions. The results of such studies, involving artificial radionuclides, not only can be applied to other environmental settings but can be used as models for understanding better the behaviour of other elements. One of the best examples is ¹³⁷Cs. Although Cs itself does not undergo any redox transformations, elevated concentrations of¹³⁷Cs have been observed in the isolated bottom waters of Par Pond, a seasonally anoxic reservoir on the site of the Savannah River Plant in South Carolina. Initially, it was tempting to link the mobilization behaviour of ¹³⁷Cs to the well-known reductive dissolution of Fe oxyhydroxides. However, more detailed studies indicated that mobilization was due to ion-exchange displacement of Cs⁺ from sediments (principally clay minerals) by a range of cations, produced as a result of anaerobic diagenesis (Evans et al., 1983). Combined field and laboratory experiments identified three types of ¹³⁷Cs binding sites: (1) surface sites on clays and other minerals, where binding is relatively non-selective being related to the charge density of the ions; (2) sites at the edges of clay interlayers of 1 nm spacing (frayed edges of partially weathered micas), where Cs⁺ is displaced only by cations of similar charge and size; (3) true interlayer sites from which Cs⁺ is not readily exchanged. NH₄+ appears to be particularly efficient at displacing Cs⁺ from the frayed edge sites, and this ties in with the fact that its concentration varies seasonally and with depth in the sediment pore waters and bottom waters of productive lakes. The fraction of ¹³⁷Cs present in each of the three forms identified by Evans and co-workers will vary, depending on the mineral composition of the sediments and to a lesser extent the time since release of¹³⁷Cs, due to the slow fixation in interlayer sites. The former explanation, or more specifically the higher ratio of kaolinite to weathered micaceous clays, has been used to account for the greater leachability of ¹³⁷Cs from Par Pond sediments than from lake and river sediments from Oak Ridge, Tennessee. That is not to say, however, that ¹³⁷Cs significant remobilization, associated with seasonal anoxia, occurs only in situations where micaceous clays are unimportant. One year after the Chernobyl accident, the same process represented the major source of ¹³⁷Cs in the bottom waters of Esthwaite Water, a lake in Cumbria (UK), the sediments of which have a clay mineral composition dominated by illite and chlorite (Davison et al., 1992).

The effect of the same redox boundary on the biogeochemical behaviour of Pu is more controversial, there being disagreement as to whether significant release from bottom sediments actually occurs under reducing conditions (e.g. Sholkovitz et al., 1982; Alberts et al., 1986). This situation highlights the fact that determining the remobilization behaviour of elements under such conditions is far from being straightforward, and that the problem is best tackled by several complementary approaches. More specifically, it is inadequate to examine water column profiles alone, since this approach is open to misinterpretation.

The study of lake sediments as a means of providing information on the sedimentary processes operating throughout a lake basin is considered in Section 5.2.2.2. Of more general interest is the way in which fine-grained sediments in the deeper waters of a lake may represent a more or less continuous historical record of the processes operating in a lake basin, its catchment and beyond. Thus, lake sediments have become the principal natural medium for providing historical records of environmental change. The ideal requirements are that: (a) the environmental change is reflected in some way in the nature or composition of the sedimentary particles; and (b) the record is not altered by the various diagenetic processes typically operating in a sediment, such as remobilization or degradation due to biogeochemical transformations, compaction, physical disturbance due to resuspension and reworking by bottom currents or benthic organisms. Within limits, the historical records can be corrected for the effects of diagenesis by mathematically modelling the constituent processes. Such diagenetic models have become an end in themselves and are capable of providing quantitative information on the various processes operating in a sediment (Berner, 1980).

Artificial radionuclides may act either as the `pollutant', for which the discharge or emission record needs to be reconstructed or, more commonly, as the `tracer' by which the diagenetic model is calibrated. The latter approach applies particularly to studies of the vertical distribution of ¹³⁷Cs in sediments, with the source term, due to fallout, being well defined. One of the earliest and most frequently quoted studies of this type is that of Robbins and Edgington (1975), who looked at sedimentation rates in Lake Michigan by means of ²¹⁰Pband ¹³⁷Cs measurements. Figures 5.6a-c show the data and modelling results for the core with greatest surficial mixing. Firstly, a best fit to the ²¹⁰Pb profile was obtained for a model incorporating sedimentation and compaction only (Figure 5.6a). This provided a timescale for converting the known historical record of ¹³⁷Cs atmospheric fallout in Lake Michigan to a depth scale with respect to the sediment surface (Figure 5.6b). The ²¹⁰Pb results for all eight sites examined indicate that the sedimentation rates have remained unchanged at each of the sites for at least 100 years and perhaps for as long as 7 x 10³ years. Assuming this to be true, there should be a straightforward correspondence between the profile in Figure 5.6b and that of ¹³⁷Cs activity concentration in the sediment (Figure 5.6c), which is clearly not the case. The first appearance and the 1963

64 maximum of ¹³⁷Cs fallout are deeper and the activity concentrations in recent years are greater than predicted. This is accounted for by additionally incorporating a sediment mixing term into the model. Best-fit values for the sedimentation rate at the sediment surface (R_o in cm y^-1)and the mixing depth (S in cm), together with the calculated profile, are shown in Figure 5.6c. The model, with the mixing term, was then applied to the ²¹⁰Pb profile, resulting in best-fit values of R_o = 0.28 cm y^-1 and S = 4.0 cm. Thus fitting the model to the two independent profiles yielded the same values for sedimentation rate and mixing depth, providing supporting evidence for the validity of the model. Edgington and Robbins (1975) have also successfully applied this approach to ^239,240Pu profiles in Lake Michigan sediments.

Figure 5.6 Vertical profiles in a sediment core from southern Lake Michigan. (a) Measured values and calculated profiles of ²¹⁰Pb. (b) The annual flux of ¹³⁷Cs to the lake surface from bomb fallout, plotted using a timescale established from the ²¹⁰Pb measurements. (c) Measured ¹³⁷Cs activities and the calculated profile, assuming rapid surface sediment mixing. (Reprinted with permission from Robbins and Edgington 1975; copyright (1975) Pergamon Press.)

The above approach yields the mixing depth and indicates implicitly that mixing is rapid in comparison with the sedimentation rate. When mixing is insufficient to produce a homogeneous mixed layer, the distribution of activity becomes sensitive to the details of the mixing process. Thus concepts, such as `conveyor belt' mixing or `biotransport', have been developed in addition to the more common approach of random mixing (see Berner, 1980).

Biogeochemical reactions may also be an important influence on radionuclide profiles, especially where concentration gradients are created in the sediment porewaters. Under such conditions, diffusional transport occurs even in the absence of mixing, i.e. by molecular diffusion. Lerman and Lietzke (1975) were amongst the first to consider the role of porewater diffusion on sediment profiles of artificial radionuclides. They developed a model to describe the vertical profiles of ¹³⁷Cs and ⁹⁰Sr in the sediments of Lakes Erie and Ontario. The cores were collected in 1969

70, so that the 1963

64 maxima of ¹³⁷Cs and ⁹⁰Sr were preserved at or close to the sediment surface. The model, which incorporated adsorption, diffusion in the porewaters and sedimentation, produced a good fit to the data, suggesting indirectly that sediment mixing was not important at the coring sites. The need for a porewater diffusion term was manifest most clearly in the much greater penetration of ⁹⁰Sr in the sediment cores, compared to ¹³⁷Cs, attributable principally to the two order of magnitude lower K_d for ⁹⁰Sr.

Later studies have extended the above concepts and have included the use of other radionuclides e.g. ¹⁰⁶Ru (Sickle et al., 1983). In addition, an increasing number of studies have focused attention on major inconsistencies between sediment profiles of ¹³⁷Cs and local fallout histories for a large number of disparate and geographically dispersed lakes. This has raised questions as to the validity of ¹³⁷Cs dating, at least under certain conditions, and emphasizes the need to adopt more than one dating technique when a reliable chronology is required. Such behaviour has been observed in lakes in North America, Europe and Australia. Explanations for the inconsistencies include mixing effects, changes in sedimentation rate, delayed but significant inputs from the catchment, recycling in the water column, sediment focusing, and redistribution due to porewater diffusion. The latter appears to be important at low pH conditions and where there is a lack of micaceous clay minerals. Under such conditions, ¹³⁷Cs may be associated mainly with organic matter and consequently be more susceptible to biogeochemical recycling.

The case study focuses on the behaviour of ¹³⁷Cs and ⁹⁰Sr and a feature of the two radionuclides is the greater geochemical mobility of ⁹⁰Sr, as described in Sections 5.2.4 and 5.2.5. Pre-Chernobyl measurements in the large rivers of Finland show an average removal by runoff of 41 per cent for ⁹⁰Sr and 7 per cent for ¹³⁷Cs of the total amount deposited in the catchment areas (Salo, 1983). As the water surface is about 9 per cent of the Finnish area, these results indicate that ⁹⁰Sr is significantly removed from land areas by runoff, while ¹³⁷Cs is not. Measurements of meltwater runoff during spring 1989 in a Norwegian mountain area (Haugen et al., 1991), combined with estimates of the total amount of snow, indicate that runoff of mainly Chernobyl-derived radiocaesium from the drainage area was 0.01

0.1 per cent.

Large variations have been observed in the relative activity concentrations of ¹³⁷Cs in solution and associated with suspended sediment (Carlsson, 1976; Haugen et al., 1991). These probably reflect real differences in conditions through the seasons. In particular, sediment bound activity is relatively greater when suspended sediment concentrations are high, due to spring meltwaters (see Section 5.2.4.1). Finnish measurements showed that lake sedimentation resulted in a sharp decline in the surface waters concentrations of Chernobyl-derived caesium by the autumn of 1986. There was a further decline in 1987, and by the end of the year concentrations in the two drainage areas, where Chernobyl fallout was highest, were only 3

15 per cent of the maximum values in May 1986. A slightly more modest decrease (down to 10

25 per cent) was measured in the larger rivers. In Sweden, Hammar et al. (1989) observed a similar rapid decrease in concentrations in surface waters, together with a more than two-fold increase in bottom sediment concentrations during the period October 1986 to September 1988. The amount retained in lake sediments varied, e.g. with clay mineral content, but was invariably greater for ¹³⁷Cs than ⁹⁰Sr. In addition to the effects associated with spring meltwaters, there appear to be many other complicating factors, related to winter conditions at high latitudes, that are important in terms of the atmospheric deposition

runoff characteristics of radioactivity. Although poorly understood, some insight into these factors was gained from a study using weapons fallout as a tracer (Lund et al., 1962).

From a population exposure point of view, radioactivity ⁱⁿsurface waters and bottom sediments is unlikely to be important. Direct exposure of bottom sediment may occur, especially during dry spells. The critical areas will be the shores of lakes and rivers. Use of these areas can lead to radiation exposure directly from the radioactive materials on the ground and via inhalation of resuspended sediment. The activity content of the sediment may vary strongly, even quite locally. Variation is particularly marked in rivers, where high concentrations are often found on the inner side of river bends. Such locations may be used for swimming, sunbathing and camping. However, both the size of the potentially exposed population and the exposure time over a year will be quite limited, at least in a North European climate. Boating and fishing activities are also unlikely to pose any significant risk. Pre- and post-Chernobyl studies of ¹³⁷Cs have shown that irrigation is an insignificant exposure pathway unless it takes place in an area otherwise unaffected by contamination.

Post-Chernobyl measurements in drinking water were performed in Finland at a large number of waterworks from 2 May 1986. The dominating radionuclide was ¹³¹I until about the beginning of June. Later on the main contaminants were ¹³⁴Cs and ¹³⁷Cs, but smaller amounts of several other radionuclides were also found. The concentrations of ¹³¹I were up to around 30 Bq kg^-1,but rapidly decreased to about 1 Bq kg^-1 at the end of May. ¹³⁴Cs and ¹³⁷Cs were rarely above 1 Bq kg^-1 and decreased toward autumn. The 1987 values were lower than the 1986 values, and the decrease continued through the summer. The effect of water purification plants upon strontium and caesium concentrations in both pre- and post-Chernobyl studies was found to be of moderate importance. Iodine and strontium concentrations are hardly affected, and the reduction of caesium concentration is at most 50 per cent.

It is easy to show by simplified, pessimistic calculations that the concentrations in water resulting from deposition upon a water surface will not even in the worst circumstances lead to doses via drinking water which could rival those via other exposure pathways, unless the drinking water is used in areas otherwise not affected by contamination.

The total amount of weapons fallout ¹³⁷Cs contained in fish in a typical lake is reported to be as low as 1 per cent of the total amount present in the lake water, although the concentration in fish is much higher (Carlsson, 1976). Typical concentration factors for ¹³⁷Cs, in fish relative to water are 10³

10⁴. The usefulness of concentration factors of this type, however, has been questioned after Chernobyl, since uptake does not take place directly from water but indirectly via numerous nutrition pathways. A direct correlation between concentration in fish and the average fallout in the drainage area (in Bq m^-2) seems to be just as relevant and useful. For trout in the mountain lakes of Norway this 'transfer factor' [(Bq kg^-1 fish)/(Bq m^-2 deposited ¹³⁷Cs)] )] ranges from 0.01 to 0.2 (Tveten, 1991).

Post-Chernobyl measurements both in Norway (Ugedal and Blakar, 1991) and Sweden (Andersson et al., 1990) indicate that the radiocaesium concentration in fish is closely linked to the content in sediment of the same lake. Predictive models perform quite well, in which concentration in fish is expressed as a function of the amount of radiocaesium in sediment (Bq m^-2 ) and various sensitivity parameters are used to characterize the lake. The sensitivity parameters may be height above sea level, area of the lake, residence time of water in the lake, and the potassium concentration in lake water. A certain division into type classes of lakes is necessary before it is possible to determine valid predictive models, however. In particular, mountain lakes seem to be in a class by themselves.

Both from earlier measurements on weapons fallout and from post-Chernobyl studies, it was found that strontium in an exposure context is less important than caesium, as strontium concentrates in bone. Most of the freshwater fish consumed are of species where most of the bones are easily removed, and it can be assumed that less than 10 per cent of the fish bone will be consumed, making the contribution from strontium negligible in most cases. The concentration of ¹³⁷Cs in fish depends, among other factors, upon the feeding habits of the fish, which are different for different species, but which also change with age and size of fish of the same species. Saxen and Aaltonen (1987) found post-Chernobyl that when fish were divided into three classes (benthic, predators and intermediate), the activity rose first in benthic and intermediate classes. In July to September 1986, activity in fish rose sharply, but was in autumn still lowest in predators. By 1987 the activity was highest in predators and was still rising, while for non-predatory fish it had started to decrease. Similar results were found in Sweden.

It was found in Finland in 1986 that for the same species (perch in particular), the smaller fish had the higher concentrations (Saxen and Rantavaara, 1987), and this was explained by differences in diet for different size and age of fish. The youngest perch eat mainly plankton. Later on they turn to insects, worms etc., and even later they start eating small fish. This inverse correlation between concentration and size did, however, not continue to be valid in 1987, when it was found both for perch and pike that the concentration was higher in larger fish. The difference between the years is explained by the fact that in 1986 the caesium had not had time to propagate through the food chain to the larger fish. Swedish investigations (Andersson et al., 1990) in 1987 also found that the concentration was larger in larger fish.

Another correlation, observed in Finland (Saxen, 1990) and in Norway (Ugedal and Blakar, 1991) is that in the same drainage area the concentration in fish is higher in smaller lakes. The concentration is also higher in lakes where the residence time of the water is longer, when the lakes are otherwise similar (Andersson et al., 1990). These authors found this factor to be the most important next to fallout levels. They also found that concentrations in fish are lower in hard water or water with high concentrations of phosphorus or potassium. Experiments on dumping large quantities of potassium in lakes to decrease caesium uptake to fish have been performed in Norway (a slight decrease could perhaps be seen) and Sweden (no decrease found).

Experiments to determine the `biological caesium half-life' have been performed in Norway. It was found that excretion of caesium is very dependent upon water temperature, and one basin-type experiment (Tveten, 1990b; Christensen, 1989), performed over a whole winter season, showed no decrease at all. Long-term measurement of caesium half-life has been performed in one mountain lake in Norway continuously since Chernobyl (Brittain, 1991), showing that the half-life decreased from about 8 to about 2 years from 1986/87 to 1988/89. For the season 1989/90, however, it had increased once more to more than 9 years. The half-life measured in this experiment is a `half-life' of the whole environment, as reflected in fish concentrations, since new activity from the drainage area will enter the lake, particularly during the spring melt. In the basin-type experiment, the fish had been transferred from its original contaminated environment to an environment free of radioactive contamination.

Typical freshwater fish consumption in Finland is reported to be about 4 kg y^-1. It is concluded (Saxen and Rantavaara, 1987) that the average intake of radiocaesium in 1986 via fish was of the same order of magnitude as via beef. In 1987, the relative contribution from fish was larger (about 65 per cent of the total average intake via foodstuffs) than in 1986; mainly because the activity levels in agricultural products decreased more rapidly than in freshwater fish. The radiocaesium concentration in cultivated rainbow trout in Finland was quite low (average of about 40 Bq kg^-1 in the most affected areas). Norwegian cultivated trout and salmon had no increase in radiocaesium content after Chernobyl, as they are cultivated in ocean waters.

The Scandinavian case study provides a representative picture of the state of knowledge concerning radionuclide behaviour in the freshwater environment. The main radionuclide and exposure pathways are reasonably well defined, especially for common nuclides such as ¹³⁷Cs and ⁹⁰Sr, but a detailed knowledge is frequently lacking. Therefore the effects of short- and long-term environmental perturbations on the fate of radionuclides in the future are often difficult to predict.

5.3 ESTUARIES AND INTERTIDAL ENVIRONMENTS

Radionuclides enter estuaries as both solution and particulate solid phases, together with the water and sediment, from any or all of four source environments: sea, river, atmosphere and the estuary itself. The magnitude and radionuclide composition of these inputs varies widely between estuaries (Table 5.3), as well as varying with time in any one. For many estuaries the dominant sources of artificial radionuclides is fallout, introduced by rivers from the land and by currents from the sea, as well as by direct deposition on the estuary. Where nuclear industry discharges occur, they can be the dominant source, as river, estuarine or marine inputs, depending on the location of the discharge point.

Within an estuary, the physical and chemical pathways followed by the radionuclides in solution and particulate phases are determined by three main groups of processes. Firstly, transport in the water column of both phases by the estuarine circulation results in dispersion, dilution, mixing and, for the particulate phase, fractionation of the radionuclides, as well as enabling chemical interactions between the two phases. Secondly, deposition of the particulate phase in the estuarine sediment deposits results in the long-term accumulation of radionuclides in the estuary. In these deposits further chemical interactions between phases are possible. Thirdly, uptake by biota of radionuclides can occur. The last category is not considered in this section.

In essence, an estuary is a narrow coastal inlet in which the sea is in contact with a river, resulting in a variation in salinity between seawater and freshwater. The seaward boundary is nominally set by the opening out of the coastline, although the geochemical effects associated with the plume of estuarine water extend the influence of the estuary into the shelf seas. The landward boundary can be put at either the limit of penetration of salt or at the limit of tidal changes in water level. The distinction is important, since there can be a considerable difference in the relative position of the two limits, e.g. in the Amazon, tidal rise and fall of water levels can be detected 850 km upstream whereas salt penetration is limited to the area of the plume outside the mouth. An estuary, therefore, can include three zones which have been given a variety of names: freshwater, also limnetic or riverine; brackish or mixohaline; and seawater or euhaline zones.

In addition to morphology, the other major determinant of the estuarine environment is the water circulation system. This has two basic components, a tidal and a non-tidal or residual circulation. The tidal circulation consists of reversing, landward and seaward tidal currents generated by the tidal change in sea level beyond the estuary mouth. These reversing currents are accompanied by changes in water level in the estuary itself (high and low tide). This results in a cyclical change in the volume of water stored in the estuary, the tidal prism, and a cyclical inundation of the marginal intertidal parts of the estuary.

The non-tidal circulation, in turn, includes two components: an outflowing current due to the river discharge and a saline density current generated by the density contrast between freshwater and saltwater. Together, these contribute to a circulation in a vertical plane, with a seaward current at the surface and a headward current at the bed. A horizontal non-tidal circulation can exist also, especially in wide estuaries, generated by the Coriolis force due to the Earth's rotation, i.e. anticlockwise in the northern hemisphere. These non-tidal circulations can be affected and even reversed by wind stress. For further discussion of estuarine circulation and hydrodynamics see, for example, Bowden (1980).

The relative importance of the three circulation components varies between estuaries, principally due to the variation of tidal amplitude around the globe, the size of the river discharge and the estuary morphology. Several different types of estuarine circulation can occur, defined by the longitudinal distribution of salinity in the estuary averaged over a tidal cycle, as shown by the isohalines for the three types in Figure 5.7.

Figure 5.7 Estuarine circulation types based on salinity distribution (contours in ppt) and showing mean flow vectors over a tidal cycle (arrows) and suspended sediment distribution (dots). From Postma, 1980; reproduced by permission of John Wiley.

It is important to remember that the division of estuaries between these circulation types is not rigid, and a single estuary may exhibit variations in its circulation regime, both spatially and temporally, e.g. a well-mixed estuary may be stratified in its upper reaches and this zone may become more extensive on the early ebb tide.

The tidal and non-tidal components of the water circulation system transport the radionuclides within the estuary by advection and diffusion. This results in changes in the distribution in space and time of radionuclide activity concentrations in the water column of both the solution and sediment phases.

The dynamic behaviour can be considered over various timescales, i.e. on timescales appropriate to either the tidal cycle or to the non-tidal circulation (tidally averaged), or longer.

5	Radionuclide Aquatic Pathways
	Co-ordinators:	J. Hamilton-Taylor, M. Kelly, P. Kershaw and C. E. Lambert
	Contributors:	A. Aarkrog, D. P. Calmet, S. Charmasson, R. Carpenter, S. Fowler, M. Ivanovich,
		G. Kuznetsov, S. P. Luttrell, S. J. Malcolm, P. I. Mitchell, H. Nies, B. Patel,
		G. G. Polikarpov, I. Rjabov, D. Swift and U. Tveten

5.1 Introduction
5.2 Freshwaters
	5.2.1 Transport and Dynamics in Rivers
		5.2.1.1 Solution phase
		5.2.1.2 Sediments
	5.2.2 Transport and Dynamics in Lakes and Reservoirs
		5.2.2.1 Solution phase
		5.2.2.2 Sediments
	5.2.3 Transport and Dynamics in Groundwater Systems
		5.2.3.1 Introduction
		5.2.3.2 Tracing applications
		5.2.3.3 Dating applications
	5.2.4 Biogeochemical Transformations and Chemical Species in Freshwaters
		5.2.4.1 Solid-solution partitioning
		5.2.4.2 Chemical speciation
		5.2.4.3 Redox boundaries as critical features in the biogeochemical behaviour of artificial radionuclides
	5.2.5 Lake Sediments and Postdepositional Change
	5.2.6 A Case Study Dealing with Exposure Pathways in Scandinavia
		5.2.6.1 Runoff characteristics and surface water and sediment activity
		5.2.6.2 Drinking water
		5.2.6.3 Freshwater fish
5.3 Estuaries and Intertidal Environments
	5.3.1 Introduction
	5.3.2 The Estuarine Environment
	5.3.3 Transport of Radionuclide in Solution
	5.3.4 Transport and Deposition of Radionuclide Particulate Phase
	5.3.5 Particle-Solution Reactions
	5.3.6 Diagenesis
	5.3.7 Radionuclide Budgets and Inventories
	5.3.8 Other Coastal Intertidal Environments
	5.3.9 The Future of Radioactive Contaminated Coastal Environments
5.4 Coastal, Semi-enclosed Basins, Shelf and Continental Margins
	5.4.1 Introduction
	5.4.2 Biomediated Pathways
		5.4.2.1 Biomediation in the water column
		5.4.2.2 Biomediation in the seabed
	5.4.3 Diagenesis
	5.4.4 Transport and Enhanced Boundary Scavenging
		5.4.4.1 Coastal mechanisms
		5.4.4.2 Cross-shelf transport and boundary scavenging
		5.4.4.3 Transport in marginal basins and semi-enclosed inland seas
		5.4.4.4 Long distance transport pathways
		5.4.4.5 Radioactive tracers and transport modelling
5.5 Disposal in the Deep Ocean
	5.5.1 Introduction
	5.5.2 Radioactive Waste Disposal
		5.5.2.1 Control and Assessment
		5.5.2.2 Advection and dispersion
		5.5.2.3 Biogeochemistry
5.6 Recommendations


	Solution phase			Particulate phase
	(Bq/m³)			(Bq/kg)				Dominant
Estuary	¹³⁷Cs	^239,240Pu	Others	¹³⁷Cs	^239,240Pu	Others	Origin	source

Connecticut, 1983		0.002			0.006		Weapons	Fluvial
(n = 710)							test fallout	+ marine
Savannah, 1986)		0.003		41.8	0.439		Weapons	Fluvial
(n = 5)							test fallout
							+ reprocessing
Seine, 1979	12.6	0.003		39.2	2.4	¹⁰⁶Ru134	Weapons	Fluvial/
(n = 1113)							test fallout	marine
							+ reprocessing
Dnieper, 1988	116		⁹⁰Sr 410				Chernobyl	Fluvial
(n = 3)							fallout
Esk, 1981	12660	18.5	²⁴¹Am 53	16260	12790	²⁴¹Am22150	Reprocessing	Marine
(n = 17171)			¹⁰⁶Ru 1050			¹⁰⁶Ru 790909

Refs: Connecticut (Sholkovitz and Mann, 1987); Savannah (Olsen et al., 1989);Seine (Jeandel et al,, 1980); Dneiper (Polikarpov et al,, 1991);
Esk (Assinder et al,, 1985)

SCOPE 50 - Radioecology after Chernobyl

5

Radionuclide Aquatic Pathways

5.1 INTRODUCTION

5.2 FRESHWATERS

5.3 ESTUARIES AND INTERTIDAL ENVIRONMENTS