A Study on the Transport of ¹³⁷Cs and ⁹⁰Sr in Marine Biota in a Hypothetical Scenario of a Nuclear Accident in the Western Mediterranean Sea

Abstract

A Lagrangian model which simulates the transport of radionuclides released from nuclear accidents in the western Mediterranean Sea was recently described. This model was developed in spherical coordinates and includes three-dimensional mixing due to turbulence, advection by currents, radioactive decay, and radionuclide exchanges between water and bed sediments. Water circulation was downloaded from the HYCOM global ocean model. Water–sediment interactions were described using a dynamic model based on kinetic transfer coefficients. Mixing, decay, and water–sediment interactions were solved using a stochastic method. Now, a dynamic biological uptake model consisting of four species (phytoplankton, zooplankton, non-piscivorous fish, and piscivorous fish) has been integrated within the transport model to be able to assess the effects of a potential accident in biota and fishery regions. The model has been set up for ${}^{137}$Cs and ${}^{90}$Sr due to the radiological relevance of these radionuclides. Several hypothetical accidents were simulated, resulting in ${}^{137}$Cs concentrations in biota significantly higher than background levels. In contrast, ${}^{90}$Sr accumulates in the food chain to a considerably weaker extent.

1. Introduction

Models that simulate the transport of radionuclides in the marine environment have been developed since the pioneering works starting in the 1980s [1,2,3,4]. This topic has attracted more attention since the accident in Fukushima NPP (Nuclear Power Plant) in 2011, as can be seen in references [5,6,7,8,9], among many others.

More recently, several models that simulate the transport of radionuclides in the marine environment have been developed for areas that are potentially exposed to nuclear accidents, due to intense shipping activities or to the presence of nuclear facilities. These are the cases, for instance, of the Red Sea [10], the Arabian/Persian Gulf [11,12,13], the Indian Ocean [14], the eastern Mediterranean [15,16] as well as the western Mediterranean Sea [17]. All these works were motivated by projects launched by the International Atomic Energy Agency (IAEA), MODARIA, and MODARIA-II [18,19], which have highlighted the need to have numerical models ready to run in marine areas that could be potentially affected by a nuclear accident. The purpose is to be able to carry out a rapid assessment of an accident’s aftermath.

Most of these models are Lagrangian in nature, because of their many advantages over Eulerian models when dealing with accidental releases into the sea as described in detail in the review in [20]. Essentially, Lagrangian models better handle the large concentration gradients that are involved in the case of accidental spills. Some earlier examples of Lagrangian models can be seen in [3,5,21,22]. Most models focus on transport processes in the water column and radionuclide interactions with sediments. Nevertheless, the uptake of radionuclides by marine biota is recently being considered and some models already include a biological uptake model (BUM) integrated within the marine transport model [13,23,24,25,26,27,28,29,30,31].

A radionuclide transport model for the western Mediterranean Sea is described in [17]. The model is Lagrangian and developed in spherical coordinates. It includes three-dimensional mixing due to turbulence, advection by currents, radioactive decay, and interactions of radionuclides between water and bed sediments. It was applied, as examples, both to hypothetical accidents in a coastal NPP and to releases due to an accident in a nuclear vessel or a ship transporting nuclear waste.

The objective of the present work is to expand the radionuclide transport model for the western Mediterranean Sea described in [17] in such a way that contamination of marine biota can also be assessed. Thus, a BUM has been integrated into the marine transport model. It includes four species: phytoplankton, zooplankton, non-piscivorous and piscivorous fish; and parameters for ${}^{137}$Cs and ${}^{90}$Sr have been implemented. These radionuclides were selected due to their high bioavailability and hence radiological relevance: Sr chemistry is similar to Ca, thus accumulates in bones, and Cs is similar to K, thus accumulates in the flesh [29]. This is the fist time that this type of model is applied to the western Mediterranean.

The description of the model is summarized in Section 2 and some applications are presented in Section 3.

2. The Model

The Lagrangian transport model, developed in spherical coordinates, is described in detail in Periáñez and Cortés [17]. The model includes three-dimensional turbulent mixing, advection by three-dimensional currents, radioactive decay, and exchanges of radionuclides between water and seabed sediments. Decay, turbulent mixing, and water–sediment interactions are solved using a stochastic method. Water–sediment interactions are simulated through a dynamic kinetic model consisting of a single reversible reaction. Two kinetic coefficients, denoted as $k_1$ and $k_2$, describe the adsorption and release reactions, respectively. Additionally, the model calculates the ages of particles discharged to the sea and can also deal with both instantaneous and continuous releases of radioactivity. Full details can be found in [17] and are not repeated here.

The model domain covers the region from −5.80° to 9.00° in longitude and 34.50° to 44.50° in latitude (Figure 1). The water currents over this area were obtained from the HYCOM ocean model [32]. Several HYCOM application examples are presented on the model web page (https://www.hycom.org/) and specific applications to the Mediterranean Sea can be seen for instance in [33,34]. HYCOM has 32 vertical levels, and its horizontal resolution is 0.08° in longitude. Horizontal resolution is not uniform in latitude and ranges between 0.0659° and 0.0571°.

As described in detail in [17], the $k_1$ coefficient is deduced from the $k_d$ distribution coefficient of each radionuclide and the value of the desorption coefficient $k_2$. Distribution coefficients, $k_d$s, are taken as the recommended values for coastal waters given by the IAEA [35]. The value of the desorption rate $k_2$ was obtained from Nyffeler et al. [36]: $k_2=1.16\times {10}^{-5}$ s${}^{-1}$. It was used in many other previous modeling works carried out for different radionuclides (see for instance references included in the review in [20]).

The basic dynamic foodweb model used in this work consists of four types of marine organisms [25,28,29]: phytoplankton, zooplankton, non-piscivorous and piscivorous fish. The equation connecting the radionuclide concentration in predator $C_{pred}$ (Bq/kg wet weight [WW]) with radionuclide concentration in food $C_f$ (Bq/kg WW) is:

$$\frac{\partial C_{pred}}{\partial t}=a{K}_1{C}_f+b{K}_w{C}_w-K_{0.5}{C}_{pred},$$

(1)

where $K_1$ ($s^{-1}$) is food uptake rate, a is the transfer coefficient through food, $K_w$ is water uptake rate ($s^{-1}$), b is the transfer coefficient from water and finally $C_w$ is activity concentration in water (Bq/m${}^3$). The radionuclide elimination rate from the body of fish, $K_{0.5}$, is given by

$$K_{0.5}=\frac{ln2}{T_{0.5}}.$$

(2)

Here, $T_{0.5}$ is the biological half-life of the considered radionuclide. All species take radionuclides from water, phytoplankton is food for zooplankton, zooplankton is food for non-piscivorous fish and, finally, non-piscivorous fish is food for piscivorous fish. Phytoplankton takes radionuclides only with the water through adsorption and desorption processes. Since radioactivity presents a short retention time and rapid uptake, the radionuclide concentration in phytoplankton is obtained from an equilibrium approach [13,28,29]:

$$C_{phyto}=C{R}_{phyto}{C}_w,$$

(3)

where $C{R}_{phyto}$ (m${}^3$/kg, WW) is the concentration ratio for phytoplankton. It must be mentioned that fish movement in the sea is not considered in the present work. Only pelagic fish are considered in the model, defined as those living in a 50 m thick surface water layer, as in the models in [23]. A scheme representing the model can be seen in Figure 2.

Please note that the Lagrangian transport model provides the radionuclide concentration in the water surface layer, $C_w$, in each grid cell from the number of particles within such cell at every time step during the model integration, as described in detail in [11,12,14] for instance. Then, the differential Equation (1) is solved (in a Eulerian framework) for each grid cell and each biota organism.

In the case of fish, the organ fraction f of the tissue that bioaccumulates each radionuclide must be considered. This target tissue depends on the radionuclide, being bones for ${}^{90}$Sr and flesh for ${}^{137}$Cs [29]. These organ fractions are given in

Table 1. Organ fractions and biological half-lives for fish [13,28,30].

	${}^{90}$Sr (Target Tissue: Bone)	${}^{137}$Cs (Target Tissue: Flesh)
Organ fraction f	0.12	0.80
$T_{0.5}$ prey (days)	128	75
$T_{0.5}$ predator (days)	257	200

(see for instance [28,30]). Thus, in the case of fish Equation (1) provides the radionuclide concentration in the target tissue, $C_{tt}$, and concentration in the fish is: 0

$$C_{fish}=f{C}_{tt}$$

(4)

Standard literature values for all parameters in Equation (1) can be seen in

Table 2. Parameters for the BUM [13,27,28,29].

	Zooplankton	Non-Pisc. Fish	Pisc. Fish
$T_{0.5}$ (day)	5	Table 1	Table 1
a (Cs–Sr)	0.2–0.2	0.5–0.29	0.7–0.29
b (Cs–Sr)	0.001–0.0001	0.001–$3\times {10}^{-5}$	0.001–$3\times {10}^{-5}$
$K_1$ (day${}^{-1}$)	1.0	0.035	0.0055
$K_w$ (m${}^3$/kg day)	1.5	0.1	0.075

. The concentration ratio for phytoplankton is $C{R}_{phyto}=20\times {10}^{-3}$ m${}^3$/kg in the case of ${}^{137}$Cs and $1\times {10}^{-3}$ m${}^3$/kg for ${}^{90}$Sr [35]. In the case of ${}^{90}$Sr, $b=1.0\times {10}^{-4}$ for zooplankton and $b=3.0\times {10}^{-5}$ for fish [13,29]. Food and water uptakes depend on the metabolic rate of each species [27], thus some parameters are the same for ${}^{137}$Cs and ${}^{90}$Sr. As commented in the introduction, these radionuclides were considered due to their high radiological relevance, but BUM parameters for other radionuclides can be seen in [13,27].

Some information must be provided to the model for each specific simulation and each considered radionuclide: location of the accident (geographical coordinates), depth, date, duration, and magnitude of the release; simulation time; radionuclide $k_d$ and decay constant and appropriate values for the BUM. All parameters are automatically defined after specifying the radionuclide of interest. The model can deal with either instantaneous or continuous releases. Output files are generated at the end of the simulated time. They consist of radionuclide concentration maps in the surface water layer, bed sediments, and radionuclide concentration maps in zooplankton, non-piscivorous and piscivorous fish. FAO (Food and Agricultural Organization of the United Nations) has divided the world ocean into several fishery areas. These fishery regions for the Mediterranean Sea are shown in Figure 3. The model provides the temporal evolution of radionuclide concentrations in surface water and biota in regions 1.1 and 1.2 (which are within the model domain) to assess the effects of potential accidents in these fisheries. It must be mentioned that averages over these relatively large regions cannot be used to assess the impact of an accident on fisheries during the acute phase of the releases, since very high concentrations are locally produced in the region close to the accident site and very small values far from it. However, averages can provide valuable information on the general trends of radionuclide concentrations at a longer temporal scale, when a high homogenization of the radionuclide patch is expected [14].

The number of particles used in the simulations is $2\times {10}^6$, but it can be changed. With this number of particles and for a one-year-long simulation, running time is about two hours in the case that the release is 90 days long.

3. Results

As mentioned in [17], it is not possible to make comparisons between model outputs and field measurements since hypothetical accidents were simulated. Nevertheless, the model formulation was tested in the frame of IAEA modeling intercomparisons exercises [19]. These exercises consisted of a comparison between models applied to simulate a hypothetical release (constant in time) from Fukushima NPP in the Pacific Ocean, to investigate the reasons for differences between models. Initially, a conservative radionuclide (not interacting with sediments) was considered, and, later, models were applied to ${}^{137}$Cs including water–sediment interactions. Models were also compared using water circulation fields from different ocean models and the same circulation data; and using different and the same model parameters (diffusion coefficients, water–sediment interaction description, seabed topography, etc.). It was concluded that the main reason for discrepancies between model outputs was due to the description of water circulation. Full details of these exercises may be seen in [19].

However, the present formulation was also applied to real discharges in which measurements in water and sediment could be compared with model results, such as the real releases from Fukushima in the Pacific Ocean [23,37], releases from European nuclear fuel reprocessing plants (Sellafield, UK, and La Hague, France) into the Atlantic Ocean [38,39] and deposition due to the Chernobyl accident in the Baltic Sea [40].

Vandellós II NPP is located on the east Spanish coast, at coordinates 40°57′ N 0°52′ E (the location is indicated in Figure 1). The NPP consists of a 1087 MW power pressurized water reactor which started its operation in 1988. Some hypothetical releases of ${}^{137}$Cs and ${}^{90}$Sr from this NPP have been simulated.

Releases of ${}^{137}$Cs have been studied first. All radionuclides are released in dissolved form. The releases were supposed to occur at the surface, with a magnitude equal to 1 PBq, lasting 90 days and being uniform in time. This is the same order of magnitude as the ${}^{137}$Cs released from Fukushima NPP into the Pacific during the first three months after the nuclear accident in March 2011 [41]. Four hypothetical accidents have been analyzed, starting, respectively, on 21 March, 21 June, 21 September, and 21 December, to observe any seasonal effects in calculated transport pathways. It must be mentioned that atmospheric releases often occur in the cases of nuclear accidents (as described in [13] for instance). Radionuclides released into the atmosphere are later deposited on the surface of the sea, as in the Fukushima accident [41,42]. The model can also handle this atmospheric deposition, but only direct-release cases are presented.

Unlike what was carried out in [17], where 3-month simulations were carried out, in this paper the simulations have been extended to one year since the main goal is to observe the accumulation of radionuclides in biota.

Typical background levels of ${}^{137}$Cs in water in the western Mediterranean, from measurements compiled in the IAEA MARIS database [43], are between 1 and 4 Bq/m${}^3$. In the case of sediments, ${}^{137}$Cs concentrations in the western Mediterranean range are 0.4–11 Bq/kg in the central sea and about 2 Bq/kg offshore northeast Spain. There are few data on ${}^{137}$Cs in fish, but in the Northwestern Mediterranean concentrations are in the range of 0.1–0.2 Bq/kg (WW). In the Arabian Gulf, for instance, these background levels are lower, about 0.034 Bq/kg (WW) [44], a difference probably due to the influence of radioactive discharges through the Rhone River from the Marcoule nuclear complex in France. In the case of ${}^{90}$Sr, concentration in water is about 1.7 Bq/m${}^3$ [43], similar to that of, for instance, the Arabian Gulf [44].

Results for the ${}^{137}$Cs release starting on 21 March are presented in Figure 4, which shows the calculated concentration in water, zooplankton, non-piscivorous and piscivorous fish. Due to the dynamic nature of the used BUM, radionuclide distribution in biota does not reflect that of water: there is a delay in the adsorption of radionuclides with respect to the concentration increase in water and biota remains contaminated during some time after a patch of contaminated water has traveled through a given area. It has been shown [45] that a dynamic model is more appropriate than an equilibrium model, based on concentration ratios, to describe radionuclide build-up in biota after a nuclear accident. In addition, concentrations resulting from this accident, both in water and fish, are significantly higher than background levels, reported above, along the northeastern Spanish coastline.

Time series of ${}^{137}$Cs concentrations in water in FAO regions 1.1 and 1.2 can be seen in Figure 5 and do not show a clear temporal reduction after the accident since the western Mediterranean essentially is an enclosed sea. In Fukushima simulations [31] a rapid decrease in radionuclide concentrations in the coastal area was obtained, since releases took place in a dynamic coastal environment where radionuclides were quickly captured by the Kuroshio Current and efficiently exported to the open Pacific Ocean. Releases from Vandellós II occur in Region 1.1, thus concentrations in water are above zero from the initial time while it takes two months for radionuclides to reach Region 1.2. The delay in radionuclide build-up in biota with respect to water, due to the dynamic nature of the BUM commented above, can be seen in Figure 5.

Because releases from the NPP last 90 days and simulations are carried out over one year, final distributions of ${}^{137}$Cs in water, sediments and biota do not show significant differences if the accident occurs in March, June, September, or December (seasonal differences were investigated at a shorter temporal scale in [17]). All results are presented in Supplementary Figures S1–S3, which show that as far as we move forward in the foodweb, seasonal differences become less pronounced. The time series of mean concentrations in FAO regions for the accident starting on 21 December are presented in Figure 6 as an additional example. In this case, nearly five months are required to achieve significant ${}^{137}$Cs mean concentrations in Region 1.2.

In the case of sediments, radionuclides are distributed along the coastal sediments of the whole domain (Figure S4). Concentrations of the order of ${10}^5$ Bq/kg, significantly larger than the background, are found along the northeast Spain coastline. In addition, results are essentially non-depending on the moment when the accident occurs since sediments act as radionuclide buffers and hence integrate the releases in time [46].

The transport of ${}^{90}$Sr was simulated for the same hypothetical accidents, although the total discharged amount was fixed as 100 TBq, which is the order of magnitude of the releases of this radionuclide from the Fukushima accident, as estimated in [47]. Calculated concentrations one year after the accident in water, zooplankton, non-piscivorous, and piscivorous fish are presented in Figure 7 for the accident starting on 21 June as an example. In the case of water, background level concentrations are obtained over most of the domain (about 1 Bq/m${}^3$ as reported above), except along a segment of the Spanish coastline southwards from the NPP. Activity concentrations in biota are some orders of magnitude below those of ${}^{137}$Cs, especially in the case of piscivorous fish. Indeed, Maderich et al. [13] have also noted in their model that ${}^{90}$Sr does not significantly accumulate in the food chain. This can also be seen in Figure 8, where the temporal evolution of mean radionuclide concentrations in FAO regions are presented: mean concentrations in piscivorous fish are below the scale. It is also interesting to note that some oscillations in concentrations in water appear in Region 1.2. Since zooplankton takes radionuclides both from water and phytoplankton, which is at equilibrium with water, radionuclide concentrations in zooplankton follow the same oscillations as water. These oscillations are already lost in the case of fish, which shows a smooth increase in radionuclide concentrations.

It is interesting to note that, both for ${}^{137}$Cs and ${}^{90}$Sr, high concentrations (significantly higher than reported backgrounds) persist along the northeastern Spanish coastline even nine months after the releases have finished (Figure 4 and Figure 7). This is due to the effects of sediments buffering radionuclides which are later released back to the water column [46]. Thus, sediments are acting as a long-term delayed source of radionuclides.

Calculated concentrations in bed sediments are significantly lower (up to three orders of magnitude) than those of ${}^{137}$Cs, which is due to the low affinity of ${}^{90}$Sr to be fixed to sediments (it has a smaller $k_d$ than ${}^{137}$Cs [35]). These results are presented in Figure S5, where it can be seen that only the Spanish coastline close to the NPP and the Balearic Islands are reached by radionuclides.

Generally speaking, time series in Figure 5, Figure 6 and Figure 8 show variability in concentrations of Region 1.2 which is not apparent for Region 1.1. The reason is that Region 1.2 is much smaller than Region 1.1 (Figure 3) and is located north of the accident location. Thus, the calculated mean concentration is highly affected by circulation patterns, leading to abrupt increases as contaminated water comes into the region or decreases if water circulation makes such water leave Region 1.2. Time series variations in biota are softer than in water due to the dynamic nature of the BUM, as already commented.

A final example is described to show that the model also works for long simulations. For this purpose, the simulation of the accident described by Figure 4 and Figure 5 was repeated but extending the simulated time to 20 years. The ${}^{137}$Cs distribution in piscivorous fish after 20 years is presented in Figure 9. Maximum calculated concentrations occur at several spots mainly in the southwestern part of the Mediterranean and along the coastlines of Africa and Spain, close to the NPP. These persisting concentrations result in the same order of magnitude as reported background levels of the Persian Gulf, as mentioned above [13]. Minimum concentrations are found in the northern region.

Finally, the time series of concentrations in FAO fisheries are presented in Figure 10. In the case of Region 1.1, there is a rapid reduction of mean concentration, and then a quite steady value is maintained. This also implies a very slow decrease in concentrations in biota, after the initial decreases which are delayed with respect to the decrease in water. The graphic for Region 1.2 shows an essentially constant concentration of water after the initial increase, which indicates an almost closed water circulation in this area. In this region, the mean concentration of fish reaches values above the aforementioned background levels.

The effective half-life, $T_{1/2}$, of a radionuclide in a marine region is defined as the time required for the mean radionuclide concentration in water to decrease in a factor 2 [14,48]. As depicted in Figure 10, concentration in water in Region 1.1 seems to exponentially decrease from day 120 onwards (in Region 1.2 it remains almost constant). This part of the curve was fitted to an exponential decay curve, resulting in a value $T_{1/2}=0.47$ year with $r^2=0.9816$. This relatively long half-life, compared to that of the Indian Ocean [14], can be attributed to the fact that the western Mediterranean is a quite closed basin.

4. Summary

A Lagrangian model that simulates the transport of radionuclides in the western Mediterranean Sea was presented in a previous publication [17]. Details of the physical model (advection by currents, three-dimensional mixing, radioactive decay, and exchanges of radionuclides between water and sediments) are described in such reference.

A dynamic biological uptake model, consisting of four types of marine organisms, has now been integrated within the marine transport model and setup for ${}^{137}$Cs and ${}^{90}$Sr. The model can be used to simulate either instantaneous or continuous radionuclide releases and it gives maps of radionuclide concentrations in water, sediments, and biota, as well as provides age distribution of radionuclides which can be useful to infer oceanographic information. Also mean radionuclide concentrations in FAO fisheries are calculated, both in water and in biota.

The model can be used as a rapid-response tool for decision-making after a radioactive spill in the western Mediterranean. Several radionuclides are released in a nuclear accident, but ${}^{137}$Cs and ${}^{90}$Sr were chosen to provide results of the biota uptake model due to their radiological relevance. As mentioned, only the radionuclide decay constant, $k_d$ and release data must be modified in the model to make a particular simulation. In addition to the use of the model as a rapid-response tool for the initial phase after an accident, long-term simulations can be carried out as well. The output of this kind of model, providing radionuclide concentrations in different environmental compartments (water, sediment, phytoplankton, zooplankton, non-piscivorous fish, and piscivorous fish), can be complemented with specific tools for dose calculations, as for instance ERICA 2.0 software [49]. The model output can (1) support civil protection authorities to protect humans in the case of such a scenario, and (2) support research groups to better identify areas of interest for performing surveillance methods integrating in situ instrumentation onto fixed and/or mobile marine platforms.

Some application examples of the model are presented. Since simulations are carried out over one year, there are no significant differences in resulting concentration maps depending on the season when the hypothetical accident occurs. These differences are reduced as moving forward in the foodweb, i.e., the minimum differences are apparent for piscivorous fish. Calculated concentrations of both radionuclides in water are significantly higher than reported background values along the northeastern Spanish coastline, even one year after the accident. This must be attributed to the fact that sediments act as radionuclide buffers, these radionuclides are later released back to the water column. Thus, these sediments behave as long-term delayed radionuclide sources. In agreement with previous work, it is found that ${}^{137}$Cs is more effectively accumulated in the foodweb than ${}^{90}$Sr. Thus, calculated ${}^{137}$Cs concentrations in fish result higher than reported background values over a significant area of the western Mediterranean.