A review on cesium desorption at the freshwater-seawater interface

https://doi.org/10.1016/j.jenvrad.2020.106255Get rights and content

Highlights

  • We reviewed 32 studies on cesium desorption from particles in seawater.

  • Desorption ranges from 0 to 86% depending on the experimental design.

  • Desorption initiates at low salinity and reach a threshold above 10–15.

  • Initial distribution of Cs, amount sorbed and contact time influence the desorption.

Abstract

Understanding the processes governing the behavior of radiocesium in the sea is still essential to make accurate assessments of its potential impacts on marine ecosystems. One of the most important of this process is the desorption that may occur at the river-sea interface due to changes in physico-chemical conditions, including ionic strength and solution composition. It has been the subject of many studies using field measurements or laboratory experiments, but there was no global interpretation of these works and their results.

The present review summarizes relevant laboratory experiments studying desorption of Cs (stable or radioactive) from particles in sea or brackish waters. To date, 32 experimental studies have been carried out on 68 Cs-bearing samples since 1964. A wide range of desorbed fraction (0–86%) was observed, partly depending on the experimental design. For particles containing radiocesium issued from a contamination in the environment, the desorption ranges from 0 to 64% of the particulate activity, with a median at only 3%. Particles contaminated in laboratory show a range between 6 and 86% with a multimodal distribution. The desorption initiates at low salinity (3–4) and rapidly reaches a threshold around 10–15. Laboratory experiments show that two first-order reactions govern the kinetics of the process, with half-life reaction times of 1 h and a few days. These two reactions are probably linked to the adsorption of Cs onto different particles sites. Also, the dynamic of Cs desorption depends on its initial distribution on these different sites, in relation with the history of its contamination and an aging effect.

Introduction

The pollution of marine ecosystems by radionuclides is a major concern for society since the beginning of the nuclear era. It may be due to different sources including: in-situ releases from coastal nuclear power-plants, accidents associated with nuclear vessels (vessels, missils), Naturally Occurring Radioactive Materials (NORMs) related to oil or gas production and direct or indirect global releases associated with nuclear tests or accidents on installations. A direct input is provided by atmospheric fallout and/or offshore releases (Fukushima), while an indirect input is due to the transport by rivers, collecting radionuclides from watersheds consequently to atmospheric fallout. Finally, submarine groundwater discharge may constitute a very specific direct input (Sanial et al., 2017).

Cesium radioactive isotopes (134Cs and 137Cs) have been extensively monitored in the environment due to their significant radioecological hazard (Garnier-Laplace et al., 2011) and their persistence (half-life of 2.4 and 30.2 years respectively). They are produced through uranium fission within nuclear reactors and thus can be found in both accident or authorized releases.

In the case of accidents and atmospheric nuclear testing, radiocesium has been spread over large spatial scales by the way of atmospheric deposition (Mattsson et al., 1991), but rivers constitute an additional input to the sea (Trapeznikov et al., 1995; Yamashiki et al., 2014). These rivers bring radiocesium through the releases of reprocessing and power plants as well as by the runoff of contaminated watersheds like in Chernobyl or Fukushima (Walling and He, 1999; Garcia-Sanchez and Konoplev, 2009; Sakaguchi et al., 2018). Cesium exists in rivers as dissolved Cs+ with small tendency to form colloids (Onishi et al., 1981; Eyrolle and Charmasson, 2004) but is mainly transported in particular form (Takahashi et al., 2017) because of its high affinity for clay minerals (Torstenfelt et al., 1982; Fan et al., 2014).

At the river-sea interface, the important changes in physico-chemical conditions including ionic strength, solution composition and pH may induce the desorption of Cs from particles to the dissolved phase. A direct consequence is the shift of its distribution coefficient Kd (ratio between solid and liquid activities, L/Kg), decreasing from 6,66.103–1,35.105 in freshwater (Tomczak et al., 2019) to 4,5.102 -2.103 in seawater (IAEA, 2004; Tagami and Uchida, 2013).

This desorption has been highlighted in laboratory experiments and in-situ studies, through the monitoring of dissolved activities (Matishov et al., 2006; Kakehi et al., 2016) or sediment inventories (Kusakabe et al., 2013). It is well recognized that clay minerals are the most important solid phase for the adsorption of radiocesium, but a fraction may also be attached to organic material, at least in contaminated watersheds near Fukushima (Naulier et al., 2017). For clays, illite is the most important and strong absorbent but kaolinite, smectite and vermiculite are also active, as well as biotite for phyllosilicate (Okumura et al., 2018)).

The negatively charged basal oxygen surfaces found on clay planar sites can form strong inner-sphere complexes with monovalent cations with low hydration energy, such as K+,NH4+, Rb+ and Cs+ (Sposito et al., 1999). These sites have usually low affinity due to their low selectivity (Wauters et al., 1996), but they adsorb Cs+ more efficiently than the other cations due to its lowest hydration energy (Nakao et al., 2014). This cation uptake can be effective within a few hours (Onishi et al., 1981). Other sorption occurred on edge sites, hydrated interlayer sites or frayed edge sites, corresponding to weathering fronts of micaceous minerals (Okumura et al., 2018). Finally, interlayer sites are not accessible to hydrated cations with large effective ionic radii, but are accessible to easily dehydrated cations such as Cs+ (Zachara et al., 2002). This binding could result in a strong fixation, similar to those of native stable Cs already present in the mineral (Yin et al., 2016). However, binding on both planar and interlayer sites are partially reversible, and the increase of competitive ions in seawater (K+, Na+ or NH4+) moves the equilibrium towards a release of Cs+, providing an additional input to the dissolved phase.

If desorption process at the freshwater-seawater interface has been already demonstrated, there is no consensus on the potential quantity of Cs that can be desorbed depending on the salinity. According to Sakaguchi et al. (2018): “the desorbed value remains open for discussion”. Furthermore, the influence of the major cations on the desorption efficiency is not completely clear. According Yin et al. (2016), questions remain on how each site is likely to desorb and on the various associated kinetic rates.

This paper provides a review on experimental laboratory studies conducted on Cs desorption in seawater. It aims to precise this process by identifying the values that can be expected according to the salinity, by characterizing the parameters of influence and by underlining unknowns for an eventual modelling work.

Section snippets

Literature search

All laboratory experiments exposing particles to a Cs (stable or radioactive) contamination in freshwater or seawater and then desorbing in seawater media were examined. The information required were the distribution of Cs between the solid and the dissolved phase.

Searches were performed to list all experimental studies carried out on stable and radioactive cesium desorption from particles samples in salt water. We included peer-reviewed and conferences papers, thesis and technical reports from

Worldwide and historic concerns on radiocesium desorption at river-sea interfaces

The 32 studies selected correspond to 68 samples and a total of 502 experimental results. 87% of these samples were sediments, 10% suspended particles and 3% soils. The sediments were collected mostly in river (56%), sea or bays (30%), or estuaries (14%). The majority of studies (14) were interested in the fate of radiocesium sorbed onto soils after atmospheric depositions due to global fallout or nuclear accidents. Also, most of the sediment or suspended particle samples were collected at the

Overview on experiments

Results of the RCs experiments follow a log-normal distribution (Fig. 2, Fig. 3). The percentages of desorption range between 0 and 65%, but the modal class 2–6% gathers more than 47% of the experimental results (93/197). Furthermore, it should be noted that values above 30% are issued from three studies only, and two of them used sediments from Bombay Harbour (Desai et al., 1994; Patel, 1978). These authors reported little information on their samples. The last and maximal value of 65% comes

Discussion

The objective of this review is to summarize data from literature in order to precise the efficiency of Cs desorption processes into seawater medium. We show that results issued from two main experimental designs may be used for that purpose. In the first design, RCs, particles have been exposed to 134Cs or 137Cs in the field, whereas the second one, RCS. Lab.Fw, concerns the desorption of radiocesium from particles spiked (134Cs or 137Cs) in laboratory and in freshwater media.

In the case of

Conclusion

Desorption of radiocesium from contaminated particles was observed at any salinity above 3, and the fraction desorbed ranges widely from 0 to 86%. For particles containing radiocesium due to an environmental contamination, this range extends from 0 to 64% with a peak at 3%. Particles containing radiocesium issued from a contamination in laboratory show higher values of Cs released and a reaction implying two first-order kinetics. The first one is really short with a half-life reaction time

Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

Acknowledgments

The authors are indebted to the Institute for Radiological Protection and Nuclear Safety (IRSN) and to Region Sud (Provence-Alpes-Côte d'Azur) authorities for the PhD funding. We thank the CRIS (Centre de ressources en informations scientifiques et techniques) of IRSN for their help to provide various documents. This study was conducted within the Rhône Sediment Observatory (OSR) program, a multi-partner research program funded through Plan Rhône of the European Regional Development Fund (ERDF)

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