Radionuclides in marine sediment

Introduction

Anthropogenic radioactive materials are introduced into the ocean from radioactive fallout from atmospheric nuclear weapons testing, accidents at nuclear power plants, radioactive waste dumping, nearshore discharges from nuclear facilities, and accidents from nuclear-powered ships. Moreover, river runoff and groundwater discharges carry to the sea some of the radionuclides deposited in the terrestrial environment, therefore becoming secondary sources of radionuclides.

Most contaminants including radionuclides sink and settle on the seabed. Marine sediments are considered a physical trap for chemicals released into the environment. Measuring the specific radioactivity and the total β-activity of sediments allows the determination of the magnitude of anthropogenic impacts, the localization of areas with a high content of natural or anthropogenic radionuclides, and the monitoring of contamination. Sedimentary natural radionuclides reflect geochemical features. The level of natural radioactivity is significantly impacted by mineralogy [1]. Anthropogenic radionuclides, such as cesium-137 (¹³⁷Cs) and strontium-90 (⁹⁰Sr), are of artificial origin.

Radionuclides in marine sediments can serve as radiotracers for studying oceanographic, sedimentological, and morpho-dynamic processes and for reconstructing the history of pollution incidents. Time-dependent changes in the concentrations of radionuclides from atmospheric nuclear weapons tests in the 1950s and 1960s and the accident at the Chernobyl Nuclear Power Plant in April 1986 have all created useful markers in the sediments of the Baltic Sea [2]. The presence of radionuclides in sediments may raise human background radiation exposures and brings crucial insights into the adverse impacts of natural radioactivity on ecosystems and human health [3], [4], [5].

The main objective of this work is to provide information on the occurrence, behavior, and fate of radionuclides in marine sediment with reference to the parameters controlling their deposition, remobilization, and transportation. In addition, the work describes the main inputs of radionuclides resulting from specific accidents with an emphasis on marine sediment. Finally, a description of the radioactivity in the sediments of the Mediterranean Basin is provided.

Fate of radionuclides in the marine environment

Once in the sea water column, radionuclides are transported by sea currents, which are considered to be the main vector for radionuclide transport in the sea, and mixed by turbulent diffusion [6]. Diffusion causes transport from regions of high to regions of low concentration. Some isotopes remain dissolved. Others are scavenged out of solution onto particulate matter such as organic material and other particles. Suspended matter sink because of gravity when its density is higher than the density of seawater and is deposited in sediments on the bottom of the sea. Settling of the particles is a slow process.

The accumulation of radionuclides on the sea floor sediments depends on various parameters including their vicinity to the coastline, the presence of aerial and fluvial input of particulates, the depth of the sea water column, the submarine topography, and the geochemical composition of the sediment. Radionuclides with a clear affinity for suspended particulate matter, like plutonium (Pu) and americium (Am), may deposit extensively in continental shelf areas [7].

The contaminants are enriched in the seafloor sediment by various physicochemical processes like absorption, ion exchange, co-precipitation, complexation, and chelation [8]. Sediment grain size, organic carbon content, clay content, pH, and cation exchange capacity are important factors affecting the kind of interaction between sinking chemicals and seabed sediment. The type of substrate which is defined based upon its physical properties is important. The sorption of metals and radionuclides onto sediment grains is influenced by grain size. Soft substrates (e.g., sand, mud, gravel, and mixed sediments) act mostly as sinks while hard substrates (i.e., bedrock and boulders) as transport bottoms for radionuclides [2]. Usually, contaminants tend to accumulate in the fine sediment as it has a high specific surface. Sediment with a high percentage of silt and clay has an affinity to bind metals and radionuclides because smaller grains have more sorption surface available to them [9]. Various studies have shown that finer sediments correlate with higher concentrations of radionuclides [4, 5, 10]. The radionuclide ¹³⁷Cs shows affinity for fine sediment and because of its relatively long radioactive half-life persists in sedimentary records. Studies of the distribution of radiocesium, derived from the Fukushima Dai-ichi Nuclear Power Plant accident on the seabed have shown good correlations between radiocesium concentration and grain sizes [11]. Since cesium has a strong affinity for clay, clay-bound Cs should behave in the same way that clay particles do. Desorption and ion exchange cause the release of sedimentary cesium. The rapid sinking of radiocesium bound particles and the secondary transport of particles attributed to turbulence near the seabed affect the particulate fluxes of radiocesium [12]. The ¹³⁷Cs desorption from the seabed has a significant effect on the reduction of the sedimentary ¹³⁷Cs [13].

Radionuclides have different behaviors in the marine environment. Uranium (VI) remains dissolved in water while uranium (IV) is insoluble. In the surface oxic waters of the Baltic Sea, uranium occurs in the +VI oxidation state. The bottom water is permanently hypoxic and periodically anoxic. Redox-driven mechanisms in the water column cause the reduction of U(IV) to U(IV) and its further deposition on the seabed as particulate matter. Investigations have shown that uranium is depleted in the sulfidic waters of the central Baltic Sea. In the anoxic sub-basins of the Baltic Sea, sediment acts as a net sink for natural and anthropogenic uranium isotopes causing the loss of ²³⁶U from the Baltic seawater [14]. Thorium is a particularly insoluble element in water, and it is usually found attached to solid material [15]. Cesium is a soluble radionuclide; it remains mostly in seawater and is transported by sea currents. On the contrary, Pu and Am mainly associate with suspended matter with Am having a greater affinity than Pu for suspended particles [16]. Since americium may occupy calcium positions within the calcite structure, sediment enriched in CaCO₃ may show a higher Am content that those enriched in SiO₂. Natural radioactivity content [radium-226 (²²⁶Ra), thorium-232 (²³²Th) and potassium-40 (⁴⁰K)] was reported to show a slightly increasing trend with increasing silt and Total Organic Matter (TOM) contents. A strong positive correlation was also revealed between ¹³⁷Cs activity concentration and TOM caused by the high cation exchange capacity of organic matter [17].

The mobility and bioavailability of radionuclides in soils is complex, depending on the clay-sized soil fraction, clay mineralogy, organic matter, cation exchange capacity, pH, and the quantities of competing cations [18]. Radionuclides bound either to clay minerals or organic material previously deposited on the seafloor may be remobilized and transported horizontally by physical processes, and/or redistributed and mixed vertically, mostly by bioturbation. A lateral transport of particulate radiocesium was reported as a possible explanation for fluctuations of ¹³⁷Cs concentrations in seabed sediments in Fukushima coastal waters [13]. Radionuclide kinetics depend upon grain size [19]. Finer-grained sediments are more easily resuspended and transported via currents. The resuspension of the fine-grained sediment forms a high turbidity layer a few meters above the seabed which influences the transport of suspended particulate matter. The resuspension causes the reintroduction of pollutants to the water column and their sorption onto the surface of particulate matter. A slow sorption process allows the ions to gradually penetrate the structure of the sedimentary solid material and the inter lattice spacing causing ion “fixation” to the sediment [20]. The rate of migration of contaminants varies. It is determined by parameters such as the chemical nature of the pollutants in the substrate and in the pore waters, and by the physicochemical conditions at the interface between the surface sediment and the deep seawater layer. As the pore water of seabed sediments has the potential to dissolve high concentrations of radiocesium, the migration of porewaters into the water column can act as a transport mechanism for water-soluble radionuclides. The high concentration of ¹³⁷Cs in the pore water, could biologically absorb and transport radiocesium to sediment fractions with high bioavailability [13].

In the course of time, long-lived nuclides that have sunk from the water phase to the seabed (surficial sediment), become buried into deeper sediment layers (sub-surface sediment) by the further deposition of suspended particles. Benthic fauna causes bioturbation at the sea bottom and pollutants are transported into deeper sediment layers. Nuclides at the water column are available for consumption by pelagic species. Nuclides in surficial sediment can be taken up by benthic species. Being buried in sub-surface sediment, nuclides are no longer easily available to the biota as contaminants are immediately accessible for microbial uptake only when in aqueous solution [21].

The disruption of sediments by dredging may cause pollutants to be released into the water column. This could alter the chemical composition of the sediment and lower the quality of the water at both extraction and disposal sites. Contaminants that are suspended have the potential to accumulate higher up the food chain and become accessible to marine life. Dredging activities from areas where the seabed is enriched in radioactive material leads to re-mobilization and transport of radionuclides [22]. Furthermore, the dredged material as a contaminated sediment needs careful management and treatment.

Sediment composition may affect the radionuclide content. The sorption capacity of sediments rich in iron (Fe) and manganese (Mn) oxides are significantly higher than that of sediments poor in these elements. Marine sediment samples from the port of Stratoni in the Ierissos Gulf located in the North Aegean Sea, were reported to have enhanced concentrations of ²²⁶Ra and ²³⁵U. This enrichment was attributed to the sediment composition that contains high amounts of Mn and Fe [9]. Organic matter can act as a major sink for labile ¹³⁷Cs. Radiocesium associated with organic matter could be desorbed and dissolved from the sediments [23]. Dissolution processes of radiocesium-bound biogenic components as well as exchanging surface-adsorbed radiocesium with alkaline metals would decrease sedimentary radiocesium content [24].

Accident at the TEPCO’s Fukushima Dai-ichi Nuclear Power Plant (FDNPP), East coast of Japan

The accident at the Fukushima Daiichi Nuclear Power Plant, operated by the Tokyo Electric Power Company (TEPCO), took place on 11 March 2011, due to tsunami waves triggered by a magnitude-9 earthquake. It released many anthropogenic radionuclides into the environment through mainly atmospheric fallout and to a lesser extent directed discharges, river runoff and submarine groundwater discharges [25]. Radionuclides have been transported long distances within the Pacific Ocean and are stored in marine sediments. The total release of radionuclides from Fukushima NPP was estimated to be approximately 520 PBq (PetaBecquerel), excluding noble gases. It includes 18–27 PBq of radioactive Cs (Cs-134, Cs-136, Cs-137), 150 PBq of ¹³¹I and 146 PBq of ¹³³I [26, 27]. Although because of their short half-life iodine isotopes are unlikely to become an important component of marine sediments.

Radiocesium was measured in sea sediment, to a core depth of 14 cm, from the area off the coast of Fukushima Prefecture to the area off the northern section of Ibaraki Prefecture, in February and July 2012. The concentration levels varied, they were greater in the upper layers and, in most of the surface sediment, surpassed 2.0 kBq/kg dry weight at the maximum. A narrow band of lowest concentration was detected along the 200-m isobaths, while high concentrations were seen in the region south of the FDNPP, particularly around the 100-m isobaths [11]. To evaluate the environmental harm, seabed sediment off the east coast of Japan, from the coastal regions of Ibaraki Prefecture collected in 2012–2014 were studied to detect changes in radioactivity concentrations of ¹³⁴Cs, ¹³⁷Cs, ⁹⁰Sr, and ²³⁸Pu (plutonium-238) and ^239,240Pu. The concentrations of sedimentary radiocesium were high for at least one year after the accident. The greatest concentration of ¹³⁷Cs recorded in 2012 was over 100 times greater than the pre-accident values (almost 1 Bq/kg) and then it decreased with time. In 2013 and 2014, the concentration of ¹³⁷Cs decreased reaching approximately the values before the accident. The decreasing tendency of ¹³⁷Cs concentrations was influenced by geographical features and the particle size of the sediment [28, 29]. The respective concentrations of radiocesium in seawater showed an exponential decrease with time indicating that radiocesium remains in the sediment for longer periods [24]. Mainly, ¹³⁷Cs inventories decrease with increasing distance from the FDNPP and increasing water mass. A correlation was not depicted between ¹³⁷Cs and ⁹⁰Sr concentrations. The accident affected the seabed sediments off the Ibaraki coast especially in the ¹³⁴Cs and ¹³⁷Cs concentrations and much less in the ⁹⁰Sr ones [29]. The inputs of Pu, Am, and ⁹⁰Sr were relatively low [25]. It was stated that there was no significant Pu contamination from the FDNPP accident to the Northwest Pacific [30].

Between January 2014 and August 2015, the suspended matter of the coastal waters around the FDNPP contained high concentrations of ¹³⁴Cs and ¹³⁷Cs, up to two orders of magnitude greater than those found in the sediment. It also contained highly radioactive cesium particles. A possible cause for the noticeable delay in radioactive Cs depuration from benthic species was the resuspension of highly radioactive particles [31]. Radioactivity monitoring studies in seafloor sediments off Fukushima and nearby prefectures indicated that the geometric mean concentration declined steadily from 47 Bq/kg in September 2011 to 13 Bq/kg in February 2016. Nevertheless, even in 2016, ¹³⁷Cs concentrations in the surface sediments were higher than the pre-accident level. At shallower water depths, ¹³⁷Cs penetrated more deeply into the sediments, most likely because wave action and tidal currents severely stirred up the top sediment [32].

A portion of the radiocesium was emitted as glassy, water-resistant cesium bearing microplastics (CsMPs). Their presence may account for the large fluctuations and slow decline of particulate radiocesium concentrations [33]. The total amount of cesium-rich microparticle (CsMPs) in the marine environment remains unknown [27].

Sedimentary radiocesium originated from the FDNPP accident was redistributed through desorption and lateral offshore transport. The vertical transport towards deeper sedimentary layers was limited [24]. Nearshore sediments off Japan will remain a significant long-term source of radio cesium for years to decades, depending on location and sediment type [25] Plutonium from the FDNPP was not found in the South China Sea [34].

Accident at the Chernobyl Nuclear Power Plant (ChNPP), Ukraine

The accident at the Chernobyl Nuclear Power Plant took place on 26 April 1986. The Number Four reactor went out of control during improper testing at low power, leading to an explosion and fire that demolished the reactor building and released massive amounts of radioactive material into the atmosphere. According to the International Atomic Energy, 1–2 × 10¹⁸ Bq of radioactive material was discharged [2]. Radioactive elements including plutonium (Pu), iodine I (Ι), strontium (Sr), and cesium (Cs) were scattered over a wide area. In addition, the graphite moderator blocks caught fire at high temperature, which contributed to the emission of radioactive materials into the environment [35].

The Chernobyl fallout was particularly rich in ¹³⁷Cs and ¹³⁴Cs. The Baltic Sea received cesium through its dispersion directly onto the sea surface during the acute fallout event, from the entire drainage area because of runoff and river discharges, and from the adjacent coastal area via sea currents. The total input of ¹³⁷Cs from the Chernobyl into the Baltic Sea has been estimated at 4100–5100 TBq (terabecquerel), and of this, one half of it has accumulated in the seafloor [2]. Furthermore, the Chernobyl accident caused an increase of ruthenium-103 (¹⁰³Ru), ruthenium-106 (¹⁰⁶Ru), silver-110 (¹¹⁰Ag), and antimony-125 (¹²⁵Sb) in Baltic Sea sediment [35]. These nuclides showed a clear enrichment in 1987 and then started to decrease by 1988/1989 because of their relatively short half-lives. Studies on the Baltic sediments showed that the ²³⁶U Chernobyl fallouts were very limited and only a small amount of Chernobyl ²³⁶U was dissolved in the Baltic seawater [36].

Baltic Sea sediment studies, for the period 2000–2005 revealed that the ¹³⁷Cs enrichment was strongest in the first 5–6 years after the fallout. The drainage area continued to act as a ¹³⁷Cs source to the seawater column and finally the seafloor. Because of variations in the atmospheric deposition, the sedimentation conditions, the geomorphological character of the sea bottom, and the sedimentation rates among the various sites of the Baltic Sea, the distribution of Chernobyl-derived ¹³⁷Cs in the Baltic Sea seabed varies. Important enrichments were identified at the northern parts of the Bothnian Sea, the southern parts of the Bothnian Bay, and the eastern parts of the Gulf of Finland. During the 2000–2005 monitoring study, the maximum highest concentration of radiocesium in Baltic Sea sediments (125 000 Bq m⁻²) was recorded in the northernmost part of the Bothnian Sea [2].

The effects of the Chernobyl disaster are still observed in the marine sediments of numerous European countries. In the Gulf of Bothnia, which is the northernmost arm of the Baltic Sea between Finland’s west coast and Sweden’s east coast, elevated ¹³⁷Cs contents compared to pre-Chernobyl levels were found in subsurface sediments from the southern Bothnian Sea, Kvarken archipelago, and southern Bothnian Bay [37]. The study of sediment cores from the Baltic Sea indicated that the main source of ¹³⁷Cs was the Chernobyl accident, while ^239+240Pu and ²⁴¹Am were mainly introduced via global nuclear weapons tests fallout. Regarding ²³⁸Pu and ²⁴¹Am, the Chernobyl disaster had a substantial effect on the direct fallout into the Baltic Proper [38]. Concerning ¹³⁷Cs, the Chernobyl accident has impacted the Mediterranean Sea as well (see Section 7). More than 30 years after the Chernobyl accident, the Norwegian fjord Vefsnfjord was found to be heavily affected by ¹³⁷Cs-contamination with contents in surface sediment two orders of magnitude higher than in open Norwegian sea area sediments [5]. In sediments from the Eastern Black Sea (Turkey), ¹³⁷Cs showed a declining trend from 1993 to 2015 apart from the areas that receive river discharges enriched in anthropogenic radionuclides from the Chernobyl accident, nuclear power plants within Black Sea countries and global nuclear weapons tests fallout [39].

Discharges from the fuel reprocessing plant at the Sellafield Works of British Nuclear Fuels Ltd (BNFL), UK

The nuclear fuel reprocessing plants at the Sellafield Works of British Nuclear Fuels Ltd (BNFL) on the Sellafield site, on the NW coast of the UK is the major source of artificial radionuclides in the NE Irish Sea. Authorized radioactive discharges from Sellafield to the Irish Sea started in 1951 with a maximum total annual release of ¹³⁷Cs, ²⁴¹Am, ^239,240Pu in 1970 [16]. The effects were more significant in the 1970s and 1980s with ¹³⁷Cs being the most noticeable component. Furthermore, considerable amounts of cesium-134 (¹³⁴Cs), strontium-90 (⁹⁰Sr) and technetium-99 (⁹⁹Tc) as well as detectable amounts of plutonium and americium have been carried from Sellafield, via the Irish Sea and Northern Atlantic Ocean to the Arctic [40].

In November 1983, a radioactivity discharge to sea, higher than normal, happened because of operations at the BNFL plant [41]. The implementation of enhanced waste treatment techniques has led to a gradual decrease in the quantity of radionuclides discharged from Sellafield. Conversely, the operation of the Thermal Oxide Reprocessing Plant starting in 1995 has led to a rise in Sellafield’s ¹²⁹I discharges. The annual discharge by 1997 was around 2.5 times more than it was in the early 1990s [42]. Continuing remobilization of ¹³⁷Cs from sediment lying at the immediate tidal area around Sellafield acted as a secondary source for ¹³⁷Cs for the period 1989–2009; about 300 TBq of ¹³⁷Cs was remobilized. Afterward the remobilization rate decreased [43]. Although plutonium has a high affinity for particulate matter, it has been transported in solution from Sellafield to the Arctic Ocean [44]. In addition, plutonium was released from sediments, introduced into the overlying water column, and transported through the North Channel of the Irish Sea. The sediment repository was considered the principal source of plutonium exported from the Irish Sea to the Arctic [42].

Sediment of Loch Etive on the west coast of Scotland showed a broad peak of ¹³⁷Cs concentration dated to the mid-1970s which was attributed to the maximum Sellafield Cs discharges from ∼1970 to the mid-1980s. It was suggested that there was a direct solution input of Sellafield-derived Cs to Loch Etive until the mid-1980s while a Sellafield input of Cs to Loch Etive continued caused by ¹³⁷Cs redissolution from sedimentary material or associated with supply of fine grained, resuspended particulates. Sellafield-derived ²³⁶U was found also to occur in Loch Etive sediments [16].

The collision of a B-52 US bomber with a KC-135 tanker aircraft, at Palomares, Spain

On 17^th January 1966, a B-52 US bomber collided with a KC-135 tanker aircraft during its inflight refueling. The bomber carried four unarmed B 28 hydrogen bombs. Three of them fell to the ground near the small fishing village of Palomares in Almeria South Spain. The explosive of two of the bombs detonated upon impact with the ground causing the spread of transuranic elements. Large volumes of radioactive aerosol with a high transuranics content were carried to the adjacent marine area by the wind. The fourth bomb fell into the Mediterranean Sea and was recovered intact after two and a half months. Although the top 10 cm of the soil has been removed, there are still high contamination levels in some areas. Both ground weathering and runoff of the Almanzora and Aguas rivers caused contamination of the coastal area with ^239,240Pu, ²⁴¹Am, and ²³⁵U. The narrow continental shelf was the area particularly contaminated [45].

Radioactivity in the Mediterranean Sea

The Mediterranean is a midlatitude semi-enclosed sea bordered by the continent of Europe in the North, Africa in the South, and Asia in the East. It is connected to the Atlantic by the shallow and narrow Gibraltar Strait. In the northeast, it is connected to the Black Sea by the Dardanelles Strait, the Marmara Sea, and the Bosporus Strait. In the southeast, it is connected to the Red Sea via the Suez Canal.

The Mediterranean Sea is composed of two nearly equal size basins, the western basin and the eastern one which are connected by the narrow and shallow Straits of Sicily. The Western basin includes the Tyrrhenian and Balearic Seas. The Eastern basin includes the Adriatic, the Ionian, the Aegean, and the Levantine Seas. Mediterranean Sea exchanges water, salt and heat with the North Atlantic Ocean and provides a laboratory basin for circulation and marine process research. The Mediterranean Sea has a high biodiversity hosting about 10 000 marine species including some endangered and threatened ones. Since the Industrial Revolution, it has received various anthropogenic contaminants that caused the degradation of its environmental status.

Major sources of anthropogenic radioactivity in the Mediterranean Sea are global fallout that mostly derives from atmospheric nuclear weapon testing and the Chernobyl accident. Releases from nuclear industry plants located at the coastal zone or on rivers are small. The Rhône River plays an important role transporting artificial radionuclides into the Mediterranean Sea. The Rhône area is enriched in ¹³⁷Cs originated form atmospheric fallout from nuclear bomb testing and the Chernobyl disaster, wash-out from the contaminated catchment area of the Rhône River as well as wastes from the nuclear industry [46]. The most important contribution of the nuclear industry results from the discharges of spent fuel reprocessing plants. The coastal Mediterranean area is impacted from the Marcoule reprocessing plant, on the banks of the Rhône River in South France. At the end of 1977, the plant was been shut down entered into a decommissioning phase.

The main radionuclides present are fission and activation products with relatively long half-life, namely ¹³⁷Cs, ⁹⁰Sr, ^238,239,240Pu and small amounts of ²⁴¹Am which derives from the decay of ²⁴¹Pu. The Chernobyl accident caused a significant input of ¹³⁷Cs which was transported via plume. The initial input of ¹³⁷Cs over the whole Mediterranean Sea was estimated as 2.5 PBq, with the deepest areas and mainly the eastern basin receiving most. This is associated with the deep-water formation in the Aegean and Adriatic Seas [47]. Marine sediment from the Amvrakikos Gulf of the Ionian Sea in northwestern Greece show increased levels of ¹³⁷Cs caused by the Chernobyl disaster [48]. In contrast, less volatile radionuclides including strontium and plutonium isotopes were deposited in the vicinity of the Chernobyl Nuclear Power Plant, and did not reach the Mediterranean Sea [45]. The major source of Pu in the Mediterranean Sea is nuclear testing fallout which was introduced to the sea via atmospheric deposition [47].

Other sources of radionuclides in the Mediterranean Sea are exchanges through the Gibraltar and the Turkish Straits [45]. Part of the significant ¹³⁷Cs input to the Black Sea because of the Chernobyl accident was exported to the Mediterranean Sea through the Bosporus Strait. The radionuclide ⁹⁰Sr remains the main anthropogenic radionuclide of Chernobyl origin being transferred from the Black Sea to the Mediterranean Sea [49]. In the period 1997–1999, mean activity concentrations of naturally occurring polonium-210 (²¹⁰Po) and lead-210 (²¹⁰Pb) in the Atlantic water entering the Mediterranean Sea were double those in the Mediterranean outflowing water. The reverse trend was detected for ^239,240Pu. The net outflow of ^239,240Pu in Gibraltar was not balanced by the Black Sea input [45]. Studies of radionuclide activity concentrations of marine sediments of Akkuyu nuclear power plant in southern Turkey at the eastern edge of the Mediterranean Sea showed an increase from east to west that could be attributed to the general circulation pattern of the Mediterranean Sea [17]. In the Spanish Mediterranean margin, the abundance sequence of anthropogenic radionuclides in coastal sediments collected in 1992 is: ¹³⁷Cs > ^239+240Pu > ²⁴¹Am > ²³⁸Pu. Since the Spanish continental margin accounts for 1 % of the total surface of the Mediterranean seabed, it was reported that the accumulated deposit of ^239+240Pu would correspond to 4 %, ²⁴¹Am to 3.5 % and ¹³⁷Cs to 1.4 % of the total inventory of these radionuclides in the Mediterranean margin [7]. Marine sediment from the Port of Stratoni on the northwest corner of Ierissos Bay, N. Aegean Sea, showed enhanced radium-226 (²²⁶Ra) activity concentration in comparison with other coastal areas of Greece and the Mediterranean Sea. This was attributed to mining activities located in Stratoni that includes the silver-lead-zinc mines, the metal processing factory, the load out pier and solid wastes [9]. The significant role of heavy metals to the natural radioactivity concentration level in beach sediment was also reported for the Adriatic Sea coastline in Albania. Beach sands there are enriched in potassium-40 (⁴⁰K), radium-226 (²²⁶Ra) and thorium-232 (²³²Th) because of heavy metal enrichment [50].

Along the Northern Cyprus coast, from Karpaz to Girne, the average activity concentration of ⁴⁰K in sediment samples is higher than the worldwide average. On the contrary, average activity concentrations of other natural radionuclides (²²⁶Ra and ²³²Th) are within acceptable limits [51]. In Egypt, the Nile supplies black sand which accumulates mostly on Rashid beach. Part of this is transported westwards via currents and accumulates on Alexandria beaches. This black sand is the major source of natural radionuclides (²³⁸U, ²³²Th, ⁴⁰K) that are found in sediments across Alexandria and Rashid coasts [10].

Conclusions

Bottom sediments are essential to radioecological studies of the marine environment since a large portion of radioactive substances entering the sea are adsorbed onto suspended particulate matter and accumulated in sediments. The anthropogenic radionuclide content of sediments is primarily determined by particle fluxes.

The dispersal of radionuclides depends on several parameters including the amount up taken or released from particulate material, the sinking process, the horizontal transport, and the accumulation in underlying sediments. Radionuclides prefer being bound by sediment that has a high percentage of silt and clay. Speciation studies of sedimentary radionuclides are needed.

Particularly for radiocesium, resuspension and subsequent lateral transport may play a critical role in reducing the ¹³⁷Cs inventory in the sediment. Redissolution of ¹³⁷Cs from contaminated sediment may also occur. Organic material is considered an important reservoir of labile ¹³⁷Cs.

The Chernobyl and Fukushima accidents caused marine radionuclide contamination across the northern hemisphere. Similarly, accidents and the operations of the nuclear industry have resulted in significant contamination of marine sediments.