Radionuclide behavior in the aquatic environment is determined by transport of the liquid and solid phases as well as the chemical interactions between phases and their biological cycling. Once the radioactive debris released by a nuclear weapons test or from a nuclear power plant accident in the atmosphere enters the terrestrial ecosystem, the debris can infiltrate into the deep soil layer to contaminate the groundwater in the aquatic ecosystem. Various radio-nuclides, whether they are naturally or non-naturally made or leaked from a nuclear reactor, that have appeared in the groundwater have been of most concern. Mobility of radionuclides in the groundwater involves several processes: precipitation, dissolution, adsorption, desorption, and ion exchange. Propagation of a radioactive plume through groundwater is a dynamic event in which all of these processes may occur simultaneously.
To evaluate potential risks of radioactive contamination in soil and groundwater, fate and transport modeling were used to calculate and predict the migration of site contaminants through the ecosystem. Commonly used fate and transport models in the groundwater transport are inherently simplistic because of the complexities of the biogeochemical processes. Most common models assume an equilibrium state, that is, linear sorption isotherm, which assumes a reversible adsorption of masses between the solid and liquid phases in the ground such that a constant distribution coefficient (Kd), which is the ratio of the amount of a solute sorbed onto solid to the concentration of the solute in the liquid solution, is used. But, variations in Kd are likely to occur not only because of possible biological and colloidal effects but also due to changing solution and sediment chemistries becoming a nonequilibrium state. In many cases, the injection/extraction of groundwater can cause mass transfer processes among the phases to be in a nonequilibrium state, which is an irreversible process.
It has been shown that the ion-exchange process including chemisorption has a profound impact on the model calculations of underground contaminants, such as uranium plume in the aquifer. The processes of uranium sorption to iron oxides and iron oxyhydroxides are not completely reversible. These oxide phases act as irreversible sinks for uranium in soil and groundwater. This irreversible process leads to attenuation of the solute.
The nonequilibrium model of groundwater transport for the liquid and solid phase of a radionuclide resolved in contaminated water can be described as follows:
where C(x, y, z, t) is the mean concentration of a radionuclide in water in the liquid phase, v(x, y, z, t) is a vector of the mean groundwater velocity, k(kx ky, kz) is a diagonal matrix of the dispersion coefficient, S(x, y, z, t) is the mean concentration of sorbing radionuclide in the solid phase, Sm represents the maximum amount of radionuclide that can be absorbed in the solid phase, C0 represents the solubility limit ofradio-nuclide, Sc(x, y, z, t) is the mean concentration of chemisorbed radionuclide, pb is the density of adsorbed radionuclide in the solid phase, Q is the adsorption rate, Q2 is the desorption rate, Q is the chemisorption rate, and B(x, y, z, t) indicates the biological and colloidal or chemical processes other than the processes of precipitation, dissolution, adsorption, desorption, and chemisorption. The sorption rates Q^ Q2, and Q can be determined by performing a sorption experiment of soil samples in the laboratory.
Nonequilibrium sorption for mass transfer between liquid and solid phases has been examined and applied to studies of the transport of the underground uranium plume at various nuclear production sites where the soil has been contaminated by leaks and spills from processing activities. The nonequilibrium sorption model is still required for more examination of quantifying the various sorption rates to improve the accuracy of the calculation in reality.
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