²⁴¹Pu in the environment: insight into the understudied isotope of plutonium

Abstract

Studies of plutonium in the environment have focused on the α-emitting isotopes ²³⁸Pu, ²³⁹Pu and ²⁴⁰Pu, often overlooking the β-emitting isotope ²⁴¹Pu  because of its relatively short half-life (14.4 years). Here, we summarize sources of  ²⁴¹Pu and discuss its distribution and behavior in the environment. In the short term, ²⁴¹Pu, the largest contributor to the total plutonium radioactivity whereas in the long term its decay products, ²⁴¹Am and ²³⁷Np, are the major contributors as some 46% of current total ²⁴¹Am is attributable to ²⁴¹Pu decay. In this context, understanding the fate and transport of ²⁴¹Pu is crucial to assessing long-term radiological dose.

Introduction

The fate and transport of plutonium (Pu) isotopes, including ²³⁸Pu, ²³⁹Pu, ²⁴⁰Pu, and ²⁴¹Pu, in the environment are of particular interest because they are potentially hazardous to human health. The primary health risk of ²³⁸Pu, ²³⁹Pu, and ²⁴⁰Pu isotopes arise from the energetic α-particles they initially emit whereas the risk for ²⁴¹Pu is from its initial decay by the emission of β-particles. Alpha particles have a much higher relative biological effectiveness than the much lighter electrons that are the primary means of radiation damage by gamma and beta radiation. Nonetheless, these isotopes must enter the body in order to cause significant damage and inhalation is the primary pathway. In fact, lung doses from these transuranics mainly occur via environmental processes such as global dust storms, the resuspension of contaminated soils, and the transport of ash from burnt contaminated biomass. Studies of plutonium in the environment have focused mainly on the ²³⁸Pu, ²³⁹Pu and ²⁴⁰Pu, isotopes that initially decay by α-radiation. In contrast, ²⁴¹Pu, initially associated with β-radiation and later γ-radiation, has been overlooked, perhaps due to its relatively short half-life. Being the most abundant plutonium isotope released in the environment, ²⁴¹Pu is the largest contributor to the total plutonium radioactivity. In the long term, its daughters ²⁴¹Am (half-life = 432.2 years) and ²³⁷Np (half-life = 21.4 × 10⁶ years) become the major contributors to dose.

The primary source of plutonium isotopes in the environment has been atmospheric nuclear tests that were conducted between 1945 and 1980 [1]. Besides atmospheric nuclear tests, satellite accidents and accidents at nuclear power plants, such as Chernobyl, Ukraine in 1986 and Fukushima, Japan in 2011, have contributed to greater plutonium contamination. It is estimated that atmospheric nuclear weapons testing of 1950s and 1960s released about 6.52 PBq (1 PBq = 1 × 10¹⁵ Bq) of ²³⁹Pu, 5.35 PBq of ²⁴⁰Pu, and 142 PBq of ²⁴¹Pu globally [1]. The second major source of plutonium contamination was the SNAP-9A accident in April 1964, which released about 630 TBq of ²³⁸Pu into the atmosphere [2]. In addition, the B-52 accidents in Palomares, Spain in 1966 and at the Thule Air Base, Greenland in 1968 caused localized environmental contamination with weapons-grade plutonium [3, 4]. Currently, the most significant contributors of plutonium to the environment are nuclear fuel reprocessing facilities such as La Hague in France and Sellafield in the United Kingdom [5, 6].

Table 1 summarizes the various sources of plutonium released into the environment. Perhaps the most significant observation is that ²⁴¹Pu is the dominant isotope in five of the eight releases. The ²⁴¹Pu is produced by the neutron capture of lower atomic weight plutonium and increases its abundance as part of the Pu mixture relative to the irradiation exposure. Plutonium mixtures vary depending on their end use. Reactor-grade plutonium contains the most ²⁴¹Pu at ~ 11%, almost 30 times higher than weapons-grade plutonium (~ 0.4%) and about 400 times higher than heat-source grade Pu (~ 0,03%) [7]. Consequently, the ²⁴¹Pu activity is highest for reactor-grade plutonium.

Table 1 Sources of plutonium isotopes in the environment (Bq)

With a half-life of only 14.4 years, ²⁴¹Pu is the shortest-lived isotope of plutonium found in the environment, which may explain why it is the most understudied. Plutonium-241, unlike other plutonium isotopes, does not emit an α-particle on decay but decays with a primary beta minus (β⁻⁾ emission, as shown in Scheme 1, which is difficult to detect with standard field instruments.

Scheme 1

Decay of ²⁴¹Pu and the formation of ²⁴¹Am

However, ²⁴¹Pu decays to the α-emitter, ²⁴¹Am, which has a half-live of 432.2 years and is also much more radiotoxic and mobile than its parent [8, 9]. The ²⁴¹Am subsequently decays to ²³⁷Np, which is also an α-emitter with a half-life of 2.1 million years. Consequently, in the later years, both ²⁴¹Am and ²³⁷Np will be the major contributors to environmental radioactivity from the disposal of high-or intermediate-level radioactive waste because of their long half-lives and their constantly increasing inventory from both direct ²⁴¹Pu emissions and in-growth of ²⁴¹Am during decay.

Plutonium isotopes undoubtedly present a large risk for internal radiation exposure via inhalation or the ingestion of contaminated food. For dose estimation, either from atmospheric testing or severe reactor accidents, the contribution of ²⁴¹Pu is currently under estimated, not only because it is a β-emitting nuclide but because the risk presented by its progeny, ²⁴¹Am and ²³⁷Np, persist over a much longer term. In this context, understanding the distribution of ²⁴¹Pu in the environment is crucial for assessing the long-term radiological dose to the general public living in the vicinity of contaminated areas. Currently in the assessment of dose, bioassays are taken for ²³⁸Pu, ^239+240Pu and ²⁴¹Am, without considering ²⁴¹Pu. Moreover, bio-kinetic models and dose calculation codes, such as the Integrated Modules for Bioassay Analysis (IMBA^®), do not account for ²⁴¹Am in-growth unless the ²⁴¹Pu activity is specified so knowledge of ²⁴¹Pu activity is essential for quantifying ²⁴¹Am. In view of these issues, this paper discusses the sources, distribution and behavior of ²⁴¹Pu in terrestrial and aquatic environments. It also highlights the progress in analytical methods for detecting ²⁴¹Pu, discusses the main issues associated with the determination of ²⁴¹Pu in environmental samples, as well as the fate and transport in the various environmental compartments.

Occurrence of ²⁴¹Pu in the environment

Following ^239+240Pu and ²³⁸Pu, ²⁴¹Pu is the third most common isotope of plutonium found in the environment. Owing to the dependence of activity ratios (i.e., ²³⁸Pu/^239+240Pu, ²⁴¹Pu/^239+240Pu, ²⁴¹Am/^239+240Pu) and atomic ratios (²⁴⁰Pu/²³⁹Pu) on the source, isotopic ratios can act as a fingerprint to identify the origin of the plutonium from different sources released into the environment. For example, the activity ratio of ²⁴¹Pu to ^239+240Pu ranges from 12 to 16 for global fallout from nuclear tests (ref date: 1963–1972) [8]; 0.5–4 for weapon-grade plutonium (ref date: 1945–1974) [8]; 25 for releases from nuclear fuel reprocessing plants (ref date: 1970–1980) [8]; 70–100 for Chernobyl fallout (ref date: April, 1986) [9] and > 100 for Fukushima fallout (ref date: March 2011) [10]. In addition, ²⁴¹Pu is almost always present in uranium and plutonium-based nuclear weapons. Therefore, the activity ratios of ²⁴¹Pu to ^239+240Pu or ²⁴¹Am to ^239+240Pu and ²⁴¹Pu/²³⁹Pu serve as a good indicator to determine the origin and age of plutonium materials and plutonium contamination [3, 11, 12].

Yet, despite its ubiquity, the ²⁴¹Pu isotope has received less attention than the other the plutonium isotopes. Owing to the relatively few studies of its environmental and biological behavior, ²⁴¹Pu can be referred to as the “forgotten isotope of plutonium”. The paucity of studies is primarily because: (1) ²⁴¹Pu is a beta-emitter with maximum energy of only 20.8 keV, representing a lesser radiological risk compared with other alpha-emitting plutonium isotopes, and (2) the precise determination of ²⁴¹Pu requires additional separation and measurement techniques like liquid scintillation spectrometry or mass spectrometry [13,14,15,16].

The global inventory of ²⁴¹Am activity produced from the decay of fallout ²⁴¹Pu will be about 60% of ²³⁹Pu in approximately 70 years [17]. Allard [18] estimates that some 1.2 × 10¹⁶ Bq of ²⁴¹Am has in-grown from the decay of fallout ²⁴¹Pu. In-grown ²⁴¹Am from the decay of the ~ 6×10¹⁵ Bq of ²⁴¹Pu released from the Chernobyl Nuclear Power Plant (NPP) accident currently amounts to 1.5 × 10¹⁴ Bq (decay corrected to 2017). Therefore, accurate determination of ²⁴¹Pu and its distribution is of interest for long-term assessment of dose to the population, for the decommissioning of nuclear facilities, and for radioactive waste management [13, 18].

Analytical techniques for determination of ²⁴¹Pu

Owing to its primary decay via a 20.8 keV β-emission, ²⁴¹Pu is difficult to detect with standard field instruments [7]. Traditionally, ²⁴¹Pu has been determined directly by liquid scintillation counting (LSC) or indirectly by alpha spectrometry to measure the in-grown ²⁴¹Am daughter. Both of these methods are capable of producing good results. Both methods require pre-concentration, separation of analytes, and long counting times (up to several days) to achieve high counting efficiency and the desired sensitivity.

The chemical separation procedure used for the α-emitting isotopes of plutonium can also be used for ²⁴¹Pu. Chemical separation is typically performed with anion exchange columns or TEVA^® chromatography columns (from Eichrom Technologies in the US or Triskem International in Europe), which have a high affinity for plutonium. Following separation, the stainless steel planchet used to electroplate plutonium isotopes for alpha spectrometric measurement is then subjected to LSC of ²⁴¹Pu, after alpha spectrometric measurement. A detection limit of 7–11 mBq has been reported using low background LSC [11].

Techniques have been developed to count alpha discs directly, with no further purification of the plutonium isotopes, by placing the planchet in the bottom of a liquid scintillation vial and adding the scintillant on top [19, 20]. However, variable deposition thickness on the alpha planchet may result in low and variable counting efficiencies for both ²⁴¹Pu and alpha isotopes of plutonium. Accounting for these low efficiencies can be difficult, even with conventional external-standard quench-correction techniques. One of the main challenges of the LSC method is the accurate determination of the counting efficiency [14]. The counting efficiency of ²⁴¹Pu in LSC usually varies from 30 to 43%, depending on the spectral quench parameter of the external standard (SQP[E]).

To maximize the counting efficiency of ²⁴¹Pu, a technique was developed whereby plutonium is leached from the alpha planchet with concentrated HNO₃ and then extracted into tri-octylphosphine oxide (TOPO)/toluene. The Pu-bearing organic extract is then mixed directly with scintillation cocktail and counted using LSC [21]. A detection limit of 105 mBq has been achieved with 5 g of LGC sediments [20]. This technique has also been used to determine ²⁴¹Pu in soils [22]. These authors compared several organic extractants (TOPO/cyclohexane, HDEHP/toluene, Aliquot 336/xylene, and DEDA/toluene) and observed that HDEHP/toluene can effectively extract ²⁴¹Pu from the sample. Lee and Lee [12] proposed a new approach in which they combined LSC with pulse-shape analysis (PSA) for the determination of ²⁴¹Pu co-existing with alpha-emitting plutonium isotopes and americium in soil samples. The indirect method uses alpha spectrometry to measure the in-grown ²⁴¹Am. In applications of this method, the separated plutonium planchets were kept for at-least six months for the in-growth of ²⁴¹Am. In the literature, the typical in-growth period for ²⁴¹Am ranged from 6 months to several years for environmental samples [20, 23]. The generated ²⁴¹Am was then separated from plutonium and measured by alpha spectrometry. The detection limit of the method varies with in-growth time. A typical detection limit with alpha spectrometry is 0.3 mBq, achieved with an in-growth period of 13 years [23]. There are two main disadvantages to this method: (1) it takes at least six months for measurable in-grown ²⁴¹Am to be produced and (2) an incomplete spectral separation of ²⁴¹Am from ²³⁸Pu interferes with the precise determination of ingrown ²⁴¹Am by alpha spectrometry. Reliable results can be achieved only if appropriate sample preparation is applied, which not only involves sample pre-concentrations but also effective matrix removal steps and multistage radiochemical separation processes. Livingston et al. [24] determined ²⁴¹Pu activities in environmental samples by measuring in-grown ²⁴¹Am produced by the decay of ²⁴¹Pu. Koide et al. [25, 26] also used this technique to measure ²⁴¹Pu activities in polar glaciers.

In recent years, several mass spectrometry techniques have also been used for the determination of ²⁴¹Pu. Inductively Coupled Plasma Mass Spectrometry (ICP-MS), Thermal Ionization Mass Spectroscopy (TIMS), Accelerator Mass Spectrometry (AMS), and Resonance Ionization Mass Spectroscopy (RIMS) have all been used to determine Pu isotopes, including ²⁴¹Pu [27]. Sturup et al. [28] used SF-ICP-MS for the determination of plutonium isotopes in sediment and seawater and obtained detection limits of 0.0115 mBq for ²³⁹Pu, 0.0084 mBq for ²⁴⁰Pu and 3.8 mBq for ²⁴¹Pu. Donard et al. [29] successfully applied SF-ICP-MS to the simultaneous detection of ²³⁹Pu, ²⁴⁰Pu and ²⁴¹Pu and ²⁴²Pu in sediments. Using the same technique and operating conditions, Varga et al. [15] reported detection limits of 0.034 mBq for ²³⁹Pu, 0.08 mBq for ²⁴⁰Pu, and 54 mBq for ²⁴¹Pu with 1-g soil and sediment samples. The formation of hydride and oxide interferences (e.g., ²³⁸U¹H⁺ or ²⁰⁷Pb¹⁶O₂⁺), usually encountered in the low-level determination of plutonium isotopes by ICP-MS, were minimized by using a desolvation sample introduction system with Ar-N₂ mixed gas plasma. Although the introduction of nitrogen might increase nitrogen containing polyatomic interferences, it decreases the UH⁺/U⁺ ratio to < 10⁻⁵. In the latter procedure, they reported improved detection limits of 0.021 mBq for ²³⁹Pu, 0.014 mBq/g for ²⁴⁰Pu, and 11.9 mBq for ²⁴¹Pu by using similar operating conditions, minus the Ar-N₂ mixed gas plasma [16]. A survey of the literature indicates that Multi Collector ICP-MS (MC-ICP-MS), coupled with ultrasonic nebulizers (USN) or MCN-6000, and Sector Field-ICP-MS (SF-ICP-MS), coupled with a microconcentric desolvating nebulizer (MCN-6000) are the most commonly utilized ICP-MS techniques for determination of the ²⁴¹Pu.

Among the mass spectrometric techniques, AMS is the most sensitive detection techniques for many long-lived radionuclides. Detection limits as low as 10⁶ atoms (ca. 0.001 mBq) can be achieved for ²³⁹Pu [30]. Additionally, matrix interferences are less in AMS than in ICP-MS and determination of the ²⁴⁰Pu to ²³⁹Pu ratio is readily accomplished. For Pu isotopes, the major advantage of AMS over the conventional MS (TIMS or ICP-MS) is the complete destruction of molecular isobars (e.g., ²³⁸UH⁺ for ²³⁹Pu) by stripping to highly positive charged states in the terminal of the tandem accelerator. The determination of ²⁴¹Pu can be achieved without additional sample processing steps, if the concentrations of ²⁴¹Pu in the sample are high enough. The major drawback of this technique in that the determination of ²⁴¹Pu has a low detection efficiency [30,31,32].

Table 2 compares the radiometric method and mass spectrometric methods for the determination of plutonium isotopes. Studies indicate that AMS, TIMS and RIMS are more sensitive than alpha spectrometry for the determination of ^239+240Pu, whereas the sensitivity of ICP-MS is comparable to LSC for the measurement of ²⁴¹Pu. The chemical separation methods for both radiometric and mass spectrometric methods can be similar; however, the counting time of mass spectrometric methods are much shorter than that of radiometric methods (a few minutes vs. a few hours and sometimes days). Lee et al. [33] compared alpha spectrometry, LSC, AMS, and ICP-MS techniques by analyzing a set of environmental reference materials. It was concluded that the alpha spectrometry obtained results for ^239+240Pu were in reasonably good agreement with those of ICP-MS and AMS, whereas values obtained by LSC and ICP-MS were only in fair agreement for ²⁴¹Pu. The determination of ²⁴¹Pu by ICP-MS is hampered by its low abundance and by polyatomic interference peaks, which can be expected at m/z of 239–242, up-mass tailing of ²³⁸U and non-spectral interferences resulting from matrix constituents [34]. This is particularly important in the analysis of environmental samples since the level of uranium is approximately 6–7 orders of magnitude higher than that of plutonium. Furthermore, the relative precision obtained by SF-ICP-MS has a mean value of 32%, which is lower than with alpha spectrometry (5 ± 1%) and liquid scintillation counting (19 ± 2%) [29].

Table 2 Typical detection limit of ²⁴¹Pu in the environmental samples by various analytical methods

Another detection technique that has been used for the determination of ²⁴¹Pu is gas-flow proportional counting. This technique uses a gaseous ionization detector to measure the energy of incident α and β ionizing radiation by generating a detector output proportional to the radiation energy. The major advantage of this technique is that the background β count rate is often lower that of LSC [27, 35]. However, the efficiency of the gas-flow proportional counter method is low due to the absorption of low-energy β particles emitted by ²⁴¹Pu into the counter window. A detection limit of about 10 mBq per sample has been reported by Rosner et al. [35], which is comparable to LSC.

These studies show that the main disadvantage of ICP-MS techniques is the relatively low abundance sensitivity that results in relatively large measurement uncertainties for ²⁴¹Pu reported at low signal intensity. In addition, there is a risk of interferences from polyatomic species and tailing of ²³⁸U, which overlap the peaks for ²³⁸Pu, ²³⁹Pu, ²⁴⁰Pu, ²⁴¹Pu and ²³⁷Np isotopes. The ICP-MS technique, therefore, may not be a very useful method for low level determination of ²⁴¹Pu generally found in the environmental samples. Environmental concentrations of ²⁴¹Pu are expected to be low because most of the ²⁴¹Pu released from atmospheric nuclear weapons testing has decayed to ²⁴¹Am. In fact, the ²⁴¹Pu/²³⁹Pu atom ratio in the atmospheric fallout samples has decreased from ~ 0.014 in 1963 to ~ 0.00194 in 2000 [36]. Owing to its short half-life and correspondingly high specific activity, radiometric determination of ²⁴¹Pu is the preferred technique. On the other hand, the mass spectroscopy technique is preferred if high concentrations of ²⁴¹Pu are present in the sample e.g., Chernobyl contaminated samples.

²⁴¹Pu in the environment: distribution and behavior

There have been several studies of plutonium in the environment since the early 1960s but most have investigated the concentration, distribution, and migration behavior of ^239+240Pu and ²³⁸Pu in the various environmental compartments with little attention paid to ²⁴¹Pu. One of the earliest studies of ²⁴¹Pu is that reported by Livingston et al. [24] in which ²⁴¹Pu concentrations and ²⁴¹Pu/^239+240Pu activity ratios were measured in several environmental samples (sediment core, beach sand, harbor sediment, seaweed, and starfish) collected from the Cape Cod area (41.34°N, 70.42°W) to identify the source of ²⁴¹Pu contamination. They found only low activity concentrations of ²⁴¹Pu, ranging from 0.28 to 18.8 mBq/g, mostly with the signature of global fallout. One of the few subsequent studies was that of Gasco et al. [3] who measured ²⁴¹Pu, other alpha isotopes of plutonium, and their isotopic ratios in soil samples collected from Palomares to evaluate the impact of the B-52 accident. Struminska-Parulska and Skwarzec [37] studied ²⁴¹Pu occurrence and its distribution in seabirds from northern Eurasia to identify its source and to quantify bio-accumulation. There was a non-uniform distribution of ²⁴¹Pu with the highest concentrations in the digestive organs and feathers and the lowest in muscle tissue. The internal dose from the accumulated ²⁴¹Pu in seabirds showed no significant effects. Moreno et al. [38] studied the spatial distribution of ²⁴¹Pu and the ²⁴¹Pu/^239+240Pu ratio in Irish Sea Plankton. The data were used to estimate transit times from Sellafield to locations farther away and the age of plutonium in Plankton. The estimated mean age in this study was 17 ± 2 years (n = 10) for plutonium and 18.6 ± 0.8 years (n = 13) for phytoplankton and zooplankton. The spatial distribution was reasonably homogeneous across the Irish Sea.

Another major source of plutonium contamination in the environment is the Chernobyl NPP accident, which occurred on April 26, 1986. It is estimated that the Chernobyl accident released about 6.1 PBq of plutonium isotopes. It increased the concentration of ^239+240Pu, ²³⁸Pu and ²⁴¹Pu in surface air during the 1986–1987 period, particularly over Europe, and contributed slightly to the global plutonium inventory [1]. An important characteristic of the radioactive material released to the atmosphere by Chernobyl was the presence of highly irradiated fuel particles known as “hot particles” or “Chernobyl dust”. Most of the plutonium was associated with larger particles (10–15 μm) causing the majority of the released plutonium to be deposited near the plant.

The accident at the Fukushima Daiichi NPP on March 11, 2011, also released a small amount of plutonium. Releases were confirmed by the detection of plutonium in environmental samples at a site 1.7 km from the Fukushima NPP and at several other sites within a 20–30 km zone around the plant [10, 39]. Although several studies have attempted to estimate the total released fission products, such as ¹³¹I and ¹³⁷Cs, very few have tried to estimate the total amount of plutonium released. Schwantes et al. [40] estimated a release of only 0.002% (± 0.003%) of the overall plutonium inventory (5.6 kg in Units 1 and 3). Recent analyses by Zheng et al. [10] reported an even lower release, totaling only 0.00002% of the core inventory (about 1.0 × 10⁹ to 2.4 × 10⁹ Bq of ^239+240Pu and 1.1 × 10¹¹ to 2.6 × 10¹¹ Bq of ²⁴¹Pu.)

The baseline data of dispersal of fallout plutonium from nuclear weapon tests are provided by Hardy et al. [2]. The minimum baseline level of ²³⁹Pu contamination in the northern hemisphere was determined to be between 0.04 and 0.15 Bq/m², based on 65 soil samples collected around the world. The global integrated deposition of ²⁴¹Pu was about 440 Bq/m² with an air concentration of about 0.8 Bq/m³ [1]. The majority of plutonium released into the atmosphere as a result of nuclear weapon tests is now associated with sub-surface soils, bottom sediments, or particulates suspended in the water column. Table 3 lists ²⁴¹Pu concentration levels measured in various environmental samples.

Table 3 The ²⁴¹Pu concentrations in environmental samples contaminated from different sources

²⁴¹Pu in surface air

Data on ²⁴¹Pu in air are very sparse because of its short half-life and difficulty in quantification by conventional radiometric techniques. Concentrations of ²⁴¹Pu in surface air were not systematically monitored during 1959–1964, the period with the heaviest contributions from global fallout. The annual average concentration of ²⁴¹Pu in surface air in the mid-latitudes of the northern hemisphere resulting from nuclear weapons testing between 1950s and 1980s was estimated using an atmospheric transport model and the known amount of ²⁴¹Pu released to the stratosphere as a result of the weapons testing. The results show a rapid increase in ²⁴¹Am surface air concentrations, which reached a maximum of 0.60 mBq/m³ in 1963–1964. Since 1973, levels have been around 0.01 mBq/m³ or less and are continuing to decline. Figure 1 shows the predicted air concentrations of ²⁴¹Pu from global fallout [41]. Aerosol concentrations of ²⁴¹Pu were measured in the lower stratosphere (10.1–14.2 km altitude) over Switzerland for the periods 1973–1977, 2007–2008 and 2010 [42]. Concentrations of ²⁴¹Pu in the stratosphere show a steady decrease from 1.34 ± 0.150 mBq/m³ in 1977 to 0.004 ± 0.001 mBq/m³ in 2010 [42]. The ²⁴¹Pu levels were more than one order of magnitude higher than those for ^239+240Pu during the period 1973–1977. These authors used a ²⁴¹Am/²⁴¹Pu age-dating model to determine the deposition history of plutonium into the atmosphere. Their calculations suggest plutonium contamination dates between 1964 and 1982.

Fig. 1

Data from Bennet [41]

Predicted surface air concentrations of ²⁴¹Pu and ²⁴¹Am (mBq/m³) during 1951–1980.

Thomas and Perkins [43] calculated ²⁴¹Pu concentrations in surface air at Richland, Washington, USA, as a function of time. They used measured concentrations of ²³⁹Pu in the atmosphere and an estimated ²³⁹Pu/²⁴¹Pu ratio of 0.015, based on the November 1952 “Ivy Mike” nuclear explosion at Enewetak Atoll, to derive concentrations of ²⁴¹Pu in surface air. Figure 2 shows the calculated surface-air concentrations of ²⁴¹Pu at Richland for the period 1963 through 1973, together with the measured concentrations of ²⁴¹Am. The surface-air concentration of ²⁴¹Pu ranged from 0.02 to 1.83 mBq/m³ with a ²⁴¹Pu/^239+240Pu ratio of about 15 [43]. As expected, the concentration of ²⁴¹Pu was highest (~ 1.83 mBq/m³) in 1963, after which it decreased steadily to 0.02 mBq/m³ by 1973. Salminen and Paatero [44] measured ²⁴¹Pu concentrations in archived filters collected in 1963 at Sodankylä, Finland. The ²⁴¹Pu concentrations in this study, which ranged from < 0.016 to 1.73 mBq/m³ with a mean ²⁴¹Pu/^239+240Pu ratio of 18, agree well with the ²⁴¹Pu concentrations estimated by Thomas and Perkins [43]. The ²⁴¹Pu/^239+240Pu ratio indicates that the source of ²⁴¹Pu in the surface air is largely due to the atmospheric weapons tests of 1950s and 1960s. During these decades, plutonium concentrations in surface air varied widely due continued contributions from nuclear weapons testing with concurrent depositions of fallout materials on the earth surface and resuspension of contaminated soil. However, due to the short half-life of ²⁴¹Pu, concentrations in the surface air have fallen below the conventional analytical detection level since the mid-1970s.

Fig. 2

Data from Ref. [43]

Surface air concentrations of ²⁴¹Pu and ²⁴¹Am (mBq/m³) during 1962–1973 in Richland, Washington, USA.

While only limited data exist on the atmospheric dispersion of ²⁴¹Pu resulting from weapons testing, the existing surface-air concentrations of ²⁴¹Pu, like ^239+240Pu, show a seasonal dependence. The 1963 data set from Sodankylä, Finland, reported by Salminen and Paatero [44] show the highest concentrations of ²⁴¹Pu occurred during the spring and summer, with the lowest concentrations in the winter (Fig. 3a). The seasonal cycle of plutonium concentrations in air has been observed since the early days of atmospheric monitoring. Seasonal variations in ^239+240Pu concentrations have been confirmed by much of the data collected in the northern hemisphere. For example, Kierepko et al. [45] showed seasonal variations of ^239+240Pu concentrations in surface air at Krakow and Bialystok, Poland for the period 1991–2008. Arnold and Wershofen [46] observed a similar seasonal cycle for ^239+240Pu in surface air over Germany during the period 1991–2003.

Fig. 3

The ²⁴¹Pu concentrations (from Ref. [44]) in surface air of Sodankyla, Finland during 1963 (top) and the estimated ²⁴¹Am concentrations in surface air of Sodankyla, Finland during 2017 (bottom)

Figure 3b shows the estimated concentrations of ²⁴¹Am in surface air of Sodankylä, Finland in 2017, produced from the decay of ²⁴¹Pu measured in 1963 by Salminen and Paatero [44]. The ²⁴¹Am concentrations obtained were in the range from 0.001 to 0.050 mBq/m³. As with ^239+240Pu, a seasonal cycle of ²⁴¹Am concentrations in air has also been observed. In the United States, surface-air concentrations of ²⁴¹Am at the Waste Isolation Pilot Plant (WIPP), a transuranic waste repository site in southeastern New Mexico, have been monitored since 2002 [47, 48]. The activities of ²⁴¹Am range from 2.0 to 61.8 nBq/m³ (Fig. 4). Approximately 50% of the ²⁴¹Am activity concentration was found in the PM₁₀ particle size fraction. Concentrations of ²⁴¹Am in surface air closely tracked those of ^239+240Pu. Although the concentrations are quite low, strong springtime peaks in ²⁴¹Am activity concentrations are evident.

Fig. 4

The ²⁴¹Am concentrations in ambient air in the vicinity of WIPP site during 2001–2013

The seasonality in surface-air concentrations are attributed to the seasonal variation in the elevation of the tropopause, the boundary between troposphere and the stratosphere. In the winter, the tropopause is at a lower elevation thereby enhancing transport of radioactive aerosols into the stratosphere and lowering surface-air concentrations. In the summer months, the tropopause is at a higher elevation and transport of radioactive aerosols into the stratosphere decreases thereby increasing surface-air concentrations. Data from the WIPP suggest another mechanism may be active. Using recent measurements of ^239+240Pu concentrations obtained in the early 1980s, and the estimated mean residence times for plutonium in the troposphere and stratosphere, it seems more likely that the observed ²⁴¹Am concentrations are due to a combination of the resuspension of contaminated soil particles and local contamination. At the WIPP, the windier part of the year occurs from January to June with April being the windiest month.

The ²⁴¹Pu in surface air were detected at several locations in Europe following the Chernobyl accident. Surface air samples taken during May 1–15, 1986 from Belgrade showed ²⁴¹Pu levels in the range 0.2–7.8 mBq/m³ [49]. In Finland, the activity concentrations in air were ~ 0.15–13.3 mBq/m³, which is about 80 times higher than the corresponding value measured for ^239+240Pu in April, 1986 [49]. Concentrations measured in Austria were in the range 0.011–0.089 mBq/m³ for ^239+240Pu, 0.0042–0.041 mBq/m³ for ²³⁸Pu and 0.0002–0.0061 mBq/m³ for ²⁴¹Pu [50]. Average surface air concentration of ²⁴¹Am in Roskilde, Denmark were in the range from 5.2 to 11.0 μBq/m³ during April–May, 1986 [51]. The ²⁴¹Pu concentrations in airborne dust samples from Gdynia (northern Poland) reached a maximum of 3.50 mBq/g during April–May time frame and then decreased gradually to 0.001 mBq/g by December, 1986 [52].

The ²⁴¹Pu/^239+240Pu activity ratio in atmospheric samples was 85 ± 20 in south Sweden, 74.6 in Austria, and 86 ± 8 in Denmark [13]. Russian researchers reported ²⁴¹Pu/^239+240Pu ratios in the range 67–82 at the height of the Chernobyl event [13]. Paatero and Jaakkola [9] reported a ²⁴¹Pu/^239+240Pu activity ratio of 94.8 in Finland compared to a ratio of 83 estimated from Chernobyl core inventory. Prior to the Chernobyl NPP accident, the ²⁴¹Pu/^239+240Pu activity ratios in airborne dust collected from Gdynia ranged from 33 and 39. In contrast, at Tsukuba, Japan, a ratio of 14.5 was reported and attributed mostly to Chinese nuclear tests in the 1980s [52].

There is report of the detection of airborne plutonium at a sampling station located 120 km from the Fukushima NPP [53]. The highest levels detected were 0.00028 mBq/m³ of ^239+240Pu and 0.025 mBq/m³ of ²⁴¹Pu (decay corrected to March 15, 2011). The authors noted that the concentrations of plutonium, and predominantly anthropogenic ²³⁶U, increased in the environment after the incident. However, levels detected were still very low and would have contributed negligibly to the total dose at the time of the incident. Furthermore, the ²⁴⁰Pu/²³⁹Pu ratios of 0.318 ± 0.10 (global fallout ratio of 1.7–0.25) and ²⁴¹Pu/²³⁹Pu ratios of 0.117 ± 0.032 (global fallout ratio of 0.00255–0.00314) in five Pu-rich samples, analyzed by AMS, were clearly different from the global fallout ratios in Japan and correspond to the ratio observed in the litter sample near the Fukushima NPP. The discharge of actinides from Fukushima was expected to be low due to their low volatility. The detection of atmospheric ^239+240Pu at a concentration of 44.5 ± 2.5 nBq/m³ after the Fukushima accident, which is four orders of magnitude lower than after Chernobyl, further confirmed that Fukushima was a negligible source of plutonium, at least outside Japan [54].

Figure 5 shows predicated increase in ²⁴¹Am/^239+240Pu activity ratio from the global fallout, Chernobyl fallout and Fukushima fallout as a function of time. The ²⁴¹Pu/^239+240Pu activity ratios from the global fallout, Chernobyl fallout, and Fukushima fallout were used to estimate the in-growth of ²⁴¹Am as described previously [10]. The ²⁴¹Am in-growth from the global fallout was based on the ²⁴¹Pu/^239+240Pu ratio of 12.8 reported by Livingston et al. [24]. For the Chernobyl ratios, the value of 96 from Paatero and Jaakkola [9] was used, while for the Fukushima NPP, initial ratio 107.8 reported by Zheng et al. [10] was used. Currently, the ²⁴¹Pu/^239+240Pu activity ratio is about 0.83 for global fallout from nuclear tests, approximately 28 for the Chernobyl fallout and about 86 for Fukushima. As shown in Fig. 5, the ²⁴¹Am/^239+240Pu ratio for global fallout never reached 1.0 but is projected to reach a maximum of 0.36 by the year 2042. For Chernobyl fallout, a ratio of 1.0 occurred in 1994 and a peak of 2.83, almost an order of magnitude higher than global fallout, is expected to occur in the year 2058. The ratio reached 1.0 for the Fukushima fallout in 2018 and is expected to peak at 3.18, more than one order of magnitude higher than global fallout in 2081, followed by a gradual decrease. Owing to the much longer half-life, ²⁴¹Am is expected to remain in the environment for a long time, potentially contributing to internal radiation doses.

Fig. 5

The calculated activity ratios of ²⁴¹Am/^239+240Pu from the global fallout, the Chernobyl and the Fukushima plutonium with time

²⁴¹Pu in soils and sediments

The fallout-based concentrations of plutonium isotopes, including ²⁴¹Pu, in soil samples collected during the period 1969–1977 in Japan were reported by Yang et al. [55]. Concentrations of ²⁴¹Pu in this study were estimated using the ²⁴¹Pu/²³⁹Pu atom ratio detected in fallout reference materials. Estimated concentrations of ²⁴¹Pu in these soils were in the range 0.06–21.24 mBq/g (decay corrected to January, 1964), whereas the measured concentrations of ^239+240Pu were in the range 0.004–1.46 mBq/g. The plutonium deposition was higher in the northern prefectures of Japan than in the southern prefectures. The spatial distribution of measured ^239+240Pu and estimated ²⁴¹Pu concentrations were highly correlated with a ²⁴¹Pu/^239+240Pu activity ratio of 14.8 [55], which is slightly higher than the global fallout ratio of 12.1 reported for lake sediments. Plutonium in these soils is mainly derived from nuclear weapons testing, with a minor contribution from Chinese nuclear tests.

There have been several studies of the vertical distribution of plutonium in soil profiles, contaminated as a result of nuclear fallout and/or by accidental releases. Generally, these studies show higher concentrations in surface soils than in subsurface soils. Depth-profile studies of ²⁴¹Am, ²³⁸Pu and ^239+240Pu in soil at the Atolls of Mururoa and Fagataufa, where French nuclear weapons tests were conducted over the period 1977–1974, show that more than 95% of the plutonium and americium remained in the top 5-cm layer. The ²⁴¹Pu concentrations in alpine soils from France and Switzerland show depth profiles similar to those observed at other sites. Maximum ²⁴¹Pu concentrations were generally found in the top 10 cm of the soil profiles. A similar depth distribution profile has been reported for ²⁴¹Am and plutonium isotopes (²³⁸Pu, ^239+240Pu) at these sites [56]. However, in the Mercantour wetlands where soils have high organic matter content (69–70%), significant differences in the vertical distributions of plutonium and americium have been observed [57]. This difference is probably due to the different migration behavior of these radionuclides in fully-saturated wetlands or to the complexation of these radionuclides with organic matter in the soil. While americium is strongly retained in the organic-rich layers of the wetlands, plutonium seems to be slightly mobilized, likely due to the formation of colloids. Hence, a slight enrichment of americium, with respect to plutonium, was observed at depths with high organic contents. Another important parameter affecting the distribution of ²⁴¹Pu in soil is particle size distribution. Lee and Clark [58] observed that majority of both ²⁴¹Pu and ²⁴¹Am in soil at the McGuire Air Force Base and Boeing Michigan Aeronautical Research Center (BOMARC) accident site in New Jersey, USA, was associated with the small size fraction (75–147 μm; very fine sand to fine sand) [58]. In a sub-surface oxic soil near Los Alamos National Laboratory, New Mexico, USA, plutonium is relatively mobile and has been transported by soil particles in the 25–450 μm (medium silt to medium sand) size range [58]. In contrast, in wet anoxic soils near Sellafield, most of the plutonium was immobilized in the sediments although a small fraction remains mobile. These differences in mobility of plutonium is attributed to differences in the oxidation state, i.e., Pu(IV) versus Pu(V), as well as the humic content of the soils. Several authors have also reported exponentially decreasing relationships between plutonium concentration and soil depth [59]. The concentration profile of ^239+240Pu in the sandy loam soil at the Trinity nuclear test site in New Mexico shows that about 50% of the plutonium initially residing in the top 0–5 cm layer of soil in 1953 had moved downward to 5–20 cm by 1973 [59]. The authors noted that the mobility of plutonium in soil profiles is greatest when plutonium is present as PuO₂, or when the soil is low in clay content, high in soluble organic material, and covered with vegetation. Most of the radioactivity was found on the < 100 µm particle size fraction, which corresponds to very fine sand and smaller. In the WIPP’s arid environment in southeastern New Mexico, even though the surface soils are quite coarse, leaching and colloidal transport are not major factors affecting the vertical migration of plutonium. Lateral movement of soil by wind erosion is the dominant mechanism in the redistribution of the radionuclides in this ecosystem.

Several studies have reported deposition of plutonium from Chernobyl NPP accident, particularly in the 30-km exclusion zone surrounding the NPP [60, 61]. The total ground deposition of plutonium in Chernobyl soil was reported to be around 2540 Bq/m² for ²⁴¹Pu and 77.4 Bq/m² for ^239+240Pu. About 99% of the released plutonium accumulated within the top 0–4 cm layers where there was significant humic acid content [60]. However, some of the plutonium remained in the surface soil layers because the radionuclides were released as hot particles that remain in the near-surface weathered zone. The 34.6 Bq/m² of ²⁴¹Am deposited in the soil profiles, as of 2000, is attributed to decay of ²⁴¹Pu. A 1998 study of soil profiles at Kapachi (7 km south of the Chernobyl NPP) showed total activities of 177 kBq/m² for ²⁴¹Pu, 17.1 kBq/m² for ^239+240Pu, and 13.7 kBq/m² for ²⁴¹Am in the top 5 cm layer, indicative of very slow vertical movement of the actinides in that soil type [61]. The authors noted that ²⁴¹Am appeared to migrate faster than plutonium. A study of forest soils from the Chernobyl 30-km exclusion zone suggests that plutonium was retained very effectively by the soil organic layers. On the basis of the almost equal distribution of plutonium between organic and mineral layers, it was suggested that plutonium is redistributed in forest soil by migrating from the organic to the underlying mineral layers. The majority of plutonium in sandy and podzolic soils was confined to the surface horizons whereas it migrated to a depth of 10–15 cm in peat soils [62].

Outside of 30-km exclusion zone, the deposition of Chernobyl fallout plutonium has been much less. For example, northeast Poland received about 1.0 kBq/m² of ²⁴¹Pu and 25 Bq/m² of ^239+240Pu [63, 64]. Southern Finland received up to 430 Bq/m² of ²⁴¹Pu [9] while Romania received between 156 and 385 Bq/m² of ²⁴¹Pu [63]. The ²⁴¹Pu level in coniferous forest soils taken from Poland was about 254 mBq/g with an average ²⁴¹Pu/^239+240Pu ratio of 86 ± 47, which is characteristic of Chernobyl fallout. No measurable Chernobyl fallout plutonium has been reported for soils from Western Europe.

Chernobyl plutonium exhibits higher abundances of ²⁴⁰Pu and ²⁴¹Pu. Activity ratios of 0.30 ± 0.03 for ²³⁸Pu/^239+240Pu and 115 ± 14 for ²⁴¹Pu/^239+240Pu in 1986 are representative to the Chernobyl fallout [64]. The ²⁴¹Am/^239+240Pu ratio of 1.6 ± 0.2 (in 2000) reported by Carbol et al. [60] is consistent with measurements in other environmental samples contaminated with Chernobyl fallout. In forest soils, ratios of ²⁴⁰Pu/²³⁹Pu and ²⁴¹Pu/²³⁹Pu co-vary and range from 0.186 to 0.348 and 0.0029 to 0.0412, respectively. In hot particles collected in soil from the Ukraine, the ²⁴¹Pu/^239+240Pu activity ratio ranged from 54.7 to 73.1 with a mean of 60.2 [65] whereas the range in Polish soils was 86–87 [63, 64].

A few studies from Japan provided solid, mass-spectrometric, isotopic evidence for the detection of Fukushima-derived plutonium in the soil and litter samples in the 20–30 km around the Fukushima NPP [10]. These authors found only low soil concentrations of ^239+240Pu in the Fukushima prefecture with values ranging from 0.019 ± 0.003 to 1.4 ± 0.023 mBq/g. These values are within the range of 0.15–4.31 mBq/g typically measured in Japanese soil samples and attributed to global fallout from nuclear weapons testing. In another study [66], soil samples collected outside of 20 km exclusion zone had ^239+240Pu levels in the range 0.0067–0.347 Bq/kg. Furthermore, the ²³⁸Pu/^239+240Pu activity ratio in these soil samples was close to 0.03 (range 0.028–0.034), indicative of global fallout. However, high activities of ²⁴¹Pu, ranging from 4.52 to 34.8 mBq/g, and ²⁴¹Pu/^239+240Pu activity ratios > 100 were detected in surface at the J-Village (about 20 km south of Fukushima NPP) and in litter samples collected from Namie Town and Iitate Village northwest of the Fukushima NPP. The release of ²⁴¹Pu is known to have occurred during the time of atmospheric nuclear weapons testing, but due to a half-life of 14.4 years, residual activity in Japanese soils is very low (²⁴¹Pu/^239+240Pu activity ratio of ~ 1.2, ²⁴¹Pu decay corrected to March 15, 2011). Thus, the detection of high ²⁴¹Pu activity in these samples suggested an additional source of plutonium. The levels detected were extremely low and ²⁴¹Am produced from the decay of Fukushima-derived ²⁴¹Pu will reach only 3% of the already low levels of soil ²⁴¹Pu. At sites most contaminated with ²⁴¹Pu, this will result in a maximum ²⁴¹Am concentration of about 1 Bq/kg by the year 2081, which is about the same found in many soils in Japan, and worldwide, as a result of nuclear weapons testing in the 1950s and 1960s. However, based on a reported ²⁴¹Pu/^239+240Pu activity ratio of > 100 for Fukushima (compared to 83 ± 5 from the Chernobyl accident), the authors emphasized the need for a long-term dose assessment of ²⁴¹Pu and ²⁴¹Am, which has a high radiotoxicity. Recently, Ikeuchi [67] confirmed the widespread distribution of Fukushima-derived ²⁴¹Pu in surface soils from the Fukushima prefectures. Based on their findings, these authors also suggested the need for a long-term dose assessment of ²⁴¹Pu and ²⁴¹Am, which has a high radiotoxicity.

Perturbations of plutonium isotopic and atom ratios in various environmental samples were also significant after the Fukushima NPP accident. For example, the ²⁴¹Pu/²³⁹Pu atom ratio (²⁴¹Pu decay corrected to March 15, 2011) of 0.122 ± 0.038, the ²⁴⁰Pu/²³⁹Pu atom ratio of 0.319 ± 0.047 in soil [10], and 0.381 ±  0.046 in vegetation samples [68] were much higher than the corresponding ratios of 0.180 ± 0.007 and 0.00194 ± 0.00014, respectively for the global fallout. Shinonaga et al. [53] found similarly high plutonium atom ratios (0.32 ± 0.10 for ²⁴⁰Pu/²³⁹Pu and 0.117 ± 0.032 for ²⁴¹Pu/²³⁹Pu) in aerosol samples collected 120 km from the Fukushima NPP. The ²³⁸Pu/^239+240Pu activity ratios (range 1.64–2.64) reported in black roadside dust, collected from the highly contaminated area of Fukushima prefectures (Minamisoma and Namie), were also higher than the global fallout value of ~ 0.03.

No ²⁴¹Pu activity was detected in surface soils at Mito, Kamagaya, and Chiba, located 100–200 km from the Fukushima NPP. Additionally, the ²⁴⁰Pu/²³⁹Pu atom ratios in samples from these cities were similar to that of the global fallout ratio, indicating that the plutonium released from the Fukushima NPP was deposited within 20–30 km around the plants. The atom ratios of ²⁴⁰Pu/²³⁹Pu and ²⁴¹Pu/²³⁹Pu found in the surface soil of J-Village were slightly lower than those in litter samples in Namie Town and Iitate Village in the NW direction of the Fukushima NPP. Using a simple two-term mixing model, Zheng et al. [10] estimated that the contribution of Fukushima-derived ^239+240Pu in the J-Village soil was 87% and the remaining 13% of ^239+240Pu was of a global fallout origin.

²⁴¹Pu in biomass

Several studies have been made on the occurrence of ²⁴¹Pu in lichens. Lichens have been identified as an important environmental medium in understanding the time deposition of plutonium. Fallout-based concentrations of ²⁴¹Pu were measured in several lichen samples taken between 1967 and 1976 in Finland [69, 70]. Lichen ²⁴¹Pu concentration ranged from 2.3 to 93 mBq/g (ref date: July 1 of sample year), some ten times higher than ^239+240Pu. In another study by Hakanen et al. [71] lichen samples taken from same region over the period 1960–1977 showed ²⁴¹Pu concentrations in the range 22–123 mBq/g, which is consistent with the levels (22–118 mBq/g) detected in lichens sampled in Sweden during 1966–1975 period [72]. Higher depositions of ²⁴¹Pu were generally observed at the bottom of the lichen than at the top. A similar distribution pattern was reported for ^239+240Pu, suggesting downward migration of deposited plutonium. This may be because lichens, which are the first to colonize bare rock, secrete carbonic and organic acids as products of metabolism. These metabolites collect at the rock surface (base of the lichen colony) where they start the process of rock weathering. Radionuclides that penetrate the colony and those transported downward by metabolites would accumulate at the low permeability rock surfaces. Activity ratios for ²⁴¹Pu/^239+240Pu in lichens were consistent with the typical global fallout ratio observed during 1960s and 1970s. Activity ratios for lichen samples collected during the period 1967–1976 ranged from 7.3 to 12.8. Samples collected during 1966–1971 period showed ratios ranging from 7.2 to 9.1. Ratios for samples collected over the period 1966–1975 ranged from 7.1 to 14. There was no difference in ²⁴¹Pu/^239+240Pu activity ratios at the top and bottom sections of these lichens. Studies also concluded that biological mean residence time of ²⁴¹Pu in lichens is on the order of 4–6 years. The distribution pattern of ²⁴¹Am in the lichen carpet was different from that of ²⁴¹Pu. No pronounced peak in activity was observed over the study period. The average ²⁴¹Am/^239+240Pu activity ratios measured were in the range from 0.12 to 0.18 [73].

The distribution of ²⁴¹Pu in lichens provide further insight into behavior of Chernobyl-derived plutonium. Lichen samples from Finland in 1986 had ²⁴¹Pu concentrations in the range 30–686 mBq/g with ²⁴¹Pu/^239+240Pu activity ratios of 37–120 (reference date July 1986), which are typical of the Chernobyl fallout ratio [69]. Since these samples were collected immediately following the Chernobyl accident, higher levels of ²⁴¹Pu were generally detected at the top part of the lichen than at the bottom. The downward migration of ²⁴¹Pu, as observed in the lichens contaminated due to global nuclear testing, was not seen in these samples. In another study, the lichen samples taken during 1987–1988, from the same regions, show ²⁴¹Pu levels in the range10–204 mBq/g and ²⁴¹Pu/^239+240Pu activity ratios between 37 and 115 (ref date: May 1986). A non-uniform distribution of ²⁴¹Pu was also reported in these lichens. The ²⁴¹Am levels in lichens after the Chernobyl accident were also determined. The highest concentration measured was 1.8 mBq/g [74]. The ²⁴¹Am/^239+240Pu activity ratio of 1.6 (in 2000) was characteristic for Chernobyl-derived plutonium in the environment.

There is no reported detection of Fukushima-derived plutonium in lichens. Since the Fukushima NPP accident seemingly did not release significant amounts of plutonium, no widespread distribution of plutonium in the environment samples is expected. However, positive detection of plutonium (^239+240Pu) has been reported in only two vegetation samples sampled at different hot spots in Japan, at the level of 0.49 and 0.17 mBq/g. The detection of Fukushima-derived plutonium in these samples indicates a very non-uniform distribution of plutonium, most probably in particulate form [68].

²⁴¹Pu in ice cores

Polar glaciers typically exhibit a detailed historical record of anthropogenic radionuclides, including fallout from nuclear weapons testing. Therefore, it is not surprising that ice cores have been used to reconstruct atmospheric transport and distribution of fallout from nuclear weapons testing. A few of the studies aimed at quantifying the plutonium recorded in Antarctic and Arctic ice cores have focused on ²⁴¹Pu distribution [26]. For the 1950s and 1960s, a good chronology exists from Arctic and Antarctic ice cores [26]. The ²⁴¹Pu in the glacier deposit of South Dome, Greenland showed two distinct fallout maxima for the periods 1950–1960 and 1963–1965 (Fig. 6a). The maximum activities measured in South Dome were 30 mBq/kg in 1955 and 27.8 mBq/kg in 1963. The first set of thermonuclear tests was conducted in the 1950s and included the “Ivy Mike” test at the Enewetak Atoll in 1952 and the “Bravo Castle” test at the Bikini Atoll in 1954. These tests were reflected in both Arctic and Antarctica ice cores with increased activity concentrations of ²⁴¹Pu from 1955 to 1959. A second peak from 1960 to 1963 reflects the post-moratorium weapon testing conducted by the former Soviet Union (USSR) at Novaya Zemlya (Russian Arctic) and Semipalatinsk (Kazakhstan) in the autumn of 1961. However, unlike the Arctic core in which the ²⁴¹Pu concentration peaked again in the early 1960s, the ²⁴¹Pu activity in the Antarctic core remained relatively low after 1958, with only a slight increase in the early 1960s. Although the tests conducted in the 1960s were large, there was minimal transport of fallout plutonium from the Russian Arctic to Antarctica, resulting in lower ²⁴¹Pu levels during the post-moratorium period. In contrast, the 1970s ice core from Dome C, Antarctica showed a large increase in ²⁴¹Pu activity, to about 72.2 mBq/kg in 1956, and a significantly lower value in the 1960s [26]. No significant activity was measured in the Dome C core prior to 1955. A similar depositional pattern has also been reported for ²³⁹Pu at Dome C. The highest ²³⁹Pu level, 9.4 mBq/kg, was observed in 1956, followed by a decreasing trend during the 1960s [75]. At the J-9 site, also in Antarctica, ²⁴¹Pu levels between 1955 and 1962 were too small to be measured by the LSC technique (Fig. 6b) The Partial Test Ban treaty of 1963 resulted in a decline in ²⁴¹Pu activity concentrations, both in the Arctic and Antarctic, but the concentrations remained above the baseline because of limited atmospheric testing by the French and Chinese through late 1970s.

Fig. 6

Date from Ref. [27] for ²⁴¹Pu and from Ref. [68] for ²³⁹Pu

The ²⁴¹Pu in the glacier deposit of South Dome, Greenland and Dome C, Antarctica during 1950–1973 (a) and annual average ²³⁹Pu in Arctic and Antarctic during 1945–1980 (b).

Overall, ²⁴¹Pu levels in the Antarctic were lower than those observed in the Arctic, which is consistent with the distribution pattern of fallout plutonium from nuclear weapons testing. Global fallout from these tests shows the highest deposition in the mid-latitudes of the northern hemisphere and lowest in the southern hemisphere. A similar distribution profile has also been observed for ²³⁹Pu and ²⁴¹Am in polar glaciers. For example, ice-core data from Greenland sites Belukha Glacier, Colle Gnifetti, and Colle du Dome, and from the Antarctic all showed increased ²³⁹Pu activity concentration from 1955 to 1960 with the highest activity recorded between 1963 and 1965 [76].

Tables 4 and 5 summarize ²⁴¹Pu/^239+240Pu activity ratios and ²⁴¹Pu/²³⁹Pu atom ratios originating from various sources. The early 1950s, fallout in the Antarctic had a ²⁴¹Pu/^239+240Pu ratio of 25–30 but this decreased to about 10 in the late 1950s [26]. In the Arctic cores, the ratio showed essentially the same trend. Sources other than atmospheric nuclear weapons testing are not a significant source of plutonium contamination in the Arctic and Antarctic and are therefore not discussed here.

Table 4 The ²⁴¹Pu/^239+240Pu activity ratio in environmental samples

Table 5 The ²⁴¹Pu/²³⁹Pu atom ratio in environmental samples

²⁴¹Pu in the marine environment

The World’s oceans are a major repository of plutonium released to the environment. The three major sources of plutonium in oceans are atmospheric nuclear weapons testing, the Chernobyl accident, and the discharge of effluents from nuclear reprocessing plants [77]. Both global as well as regional fallout contributed to the present levels of plutonium in seawater and sediment. The oceans contain an estimated 30 PBq of ²⁴¹Pu [77], attributed to weapons testing (decay corrected to Jan, 2000). In the marine environment, the majority of plutonium released is found in sediment layers. Marine sediments then become a source of plutonium to seawater. Sediment mixing by actions of benthic organisms and physical processes, such as waves and currents, further facilitate re-introduction of radionuclides to seawater. Plutonium exchanged from contaminated sediment to seawater is then available to be transported via ocean currents. A variety of biogeochemical processes such as changes in oxidation state, dissolution, hydrolysis, complexation, sorption, colloid formation, and microbial activity all influence the dispersion, mobility, and long-term behavior of plutonium in the marine environment [77]. The most common oxidation state of plutonium in marine waters are Pu(IV) and Pu(V). The Pu(IV) state is highly particle-reactive and is therefore easily scavenged by suspended matter and colloids, whereas Pu(V) is relatively soluble and can be transported in the dissolved phase over long distances. However, in anoxic marine waters, Pu(III) is the dominant oxidation state [78].

The first ocean-wide study of plutonium in seawater originated with the Pacific GEOSECS sampling program in the early 1970s [79]. At present, seawater concentrations of ²⁴¹Pu originating from nuclear weapons testing vary between 20 and 40 mBq/m³ (ref. date 2000) whereas the concentration of ²⁴¹Am is less than 1–2 mBq/m³ [80, 81]. The ^239+240Pu activity concentration in surface seawater varies from a few mBq/m³ in the open ocean to greater than 100 mBq/m³ in semi-enclosed waters close to the source, e.g., in the Irish Sea [82]. However, elevated concentrations in the water column have been observed in some locations as a result of local or regional fallout. For example, surface seawater concentrations of ^239+240Pu in the vicinity of the Enewetok Atoll were around 3000 mBq/m³ (measured in 1997) [80]. Localized elevated seawater concentrations of plutonium, especially near the bottom water, have also been observed at sites like Chernaya Bay in the Arctic sea, the Irish Sea near Sellafield, the nuclear weapon test site near Novaya Zemlya, and the Thule accident site [80,81,82,83]. Owing to their particle-reactive nature, plutonium isotopes and ²⁴¹Am are easily attached to particles and effectively scavenged from surface water to be released again at medium depths. Typical profiles of these radionuclides in the open ocean showed minimum values at the surface and maxima at medium water depths [82].

Since April 1986, Chernobyl has been a new and significant source of plutonium in Eastern European and the Baltic Sea. Following a 1986 peak in plutonium isotopes from Chernobyl, radionuclide concentrations decreased continuously [84]. The fallout from Chernobyl to the marine environment was reasonably small and localized, e.g., in the Baltic Sea and the Black Sea [84]. The total inventory of ²⁴¹Pu in the Baltic Sea in 1986 was estimated at 138 TBq, of which about 64% (88 TBq) is believed to have originated from the Chernobyl accident [84]. Approximately 99% of the inventory is assumed to have been rapidly transferred to the bottom sediments. Being the closest marine body to the Chernobyl site, the Black Sea has also received, and continues to receive, an additional input of Chernobyl-derived radionuclides through runoff from the Danube and Dnieper Rivers. The most significant radionuclides carried into the Black Sea from these sources were ¹³⁷Cs and ⁹⁰Sr. However, plutonium isotopes (²³⁸Pu, ^239+240Pu and ²⁴¹Pu) and ²⁴¹Am have also been detected in Black Sea sediments at concentrations ranging from 1.4 to 15 mBq/g for ²⁴¹Pu and 0.07 to 0.75 mBq/g for ^239+240Pu [85, 86].

An initial survey of ²⁴¹Pu in marine sediments and biological samples was provided by Livingston et al. [24]. The ²⁴¹Pu concentrations in marine sediment samples were in the range 21–950 mBq/g, whereas the range in biological samples was 17–58 mBq/g. Analysis of the vertical migration of ²⁴¹Pu in marine sediments show that most of the fallout ²⁴¹Pu still remains in the first few centimeters of sediment cores [24]. The ²⁴¹Pu/^239+240Pu activity ratio in these sediments averaged 9.4 at the time of collection (12.8, if decay corrected to 1962), which agrees with the fallout ratio of 13–14 for the 1961–1962 USSR/USA tests and the 1967 Chinese tests. Koide et al. [17] reported representative ratios of 14–16 (12.4, if decay corrected to July 1, 1962) for the fresh fallout debris in sediments from coastal basins off California and Mexico.

Several studies have reported the distribution of ²⁴¹Pu in the Baltic Sea ecosystems [37, 87, 88]. The most significant source of plutonium in the Baltic Sea is the Chernobyl NPP accident. The second most important source is fallout from nuclear weapon tests. Surface waters of the southern Baltic, sampled during 1986–1988, contained 170 mBq/m³ of ²⁴¹Pu in Gulf of Gdańsk samples and 90 mBq/m³ in Gdańsk Deep samples with ²⁴¹Pu/^239+240Pu activity ratios of 74 and 89, respectively [88]. In a separate campaign, surface waters collected from four different regions of the Baltic Sea (Gulf of Gdańsk, Słupsk Bank, Bornholm Deep and Pomeranian Bay), during the period 1997–2001, show an increase in concentration of ²⁴¹Pu to about 2210–3350 mBq/m³. The study showed that colloidal plutonium (> 1 kDa) is only a minor fraction of the total plutonium in Baltic waters. The largest percentage of plutonium (50–80%) in these waters was found in the dissolved fraction [89]. The highest total activity concentration of ²⁴¹Pu was found in Słupsk Bank and lowest in Pomeranian Bay. The ²⁴¹Pu/^239+240Pu activity ratios are higher than the global fallout ratios and reflect a mixture of Chernobyl and global fallout plutonium [89].

The vertical distribution of ²⁴¹Pu in sediments from these four regions of the Baltic Sea was reported by Struminska-Parulska [87]. A non-uniform depth distribution was reported for ²⁴¹Pu, with most ²⁴¹Pu occurring at shallower depths that did not exceed 10 cm. There was no evidence of the downward migration of ²⁴¹Pu. The sandy sediments from the Gulf of Gdansk appear to contain less plutonium (average ²⁴¹Pu 7.92 mBq/g; range 0.02–20.0 mBq/g) than the silty and organic-rich sediments from the Gdańsk Deep (average ²⁴¹Pu conc. 22.3 mBq/g; range 1.5–62 mBq/g). The ²⁴¹Pu/^239+240Pu activity ratio in the Gulf of Gdańsk was characteristic of Chernobyl fallout (44 ± 13) in the top 1-cm sediment layer, whereas the ratio (< 10) at the bottom layer (2–8 cm) was reflective of global fallout. In the Gdańsk deep, the ratio obtained was similar to that of global fallout.

In areas contaminated by the Thule accident, concentrations of ^239+240Pu, ²³⁸Pu, ²⁴¹Pu and ²⁴¹Am were one to three orders of magnitude higher than the fallout level. In sediments, the highest concentrations of ²⁴¹Pu and ^239+240Pu were 5900 and 7600 mBq/g, respectively. The vertical distribution of Thule sediments showed the downward migration of radionuclides to 20 cm, somewhat deeper than reported by Smith et al. [90] where radionuclides reached down to 15 cm with maximum concentrations at a depth of 3–6 cm.

Another major source of ²⁴¹Pu in marine environment has been the discharge of waste from nuclear fuel reprocessing facilities such as La Hague, France and Sellafield, United Kingdom. The estimated release of total plutonium into the Irish Sea from Sellafield for the period 1951–1992 amounts to 0.12 PBq of ²³⁸Pu, 0.61 PBq of ^239+240Pu, and 21 PBq of ²⁴¹Pu with a ²⁴¹Pu/^239+240 ratio between 30 and 40 [5]. Most of the discharge occurred from the mid-1960s through the mid-1980s. The ²⁴¹Pu/^239+240 ratio peaked at around 50 in the mid-1970s and has remained between 30 and 40 ever since. Plutonium release from the La Hague into the north coast of France has been small, about 0.4% of that of Sellafield, with a total plutonium release of less than 0.005 PBq [6]. More than 95% of the plutonium released from Sellafield to the Irish Sea has been rapidly transported to the seabed in association with particulate matter. It has also been demonstrated that the spatial distribution of ²⁴¹Pu is controlled by the tidal currents and the clay contents in the sediments. Particle transport is considered to be the most important transport mechanism for the dispersion of plutonium to intertidal and offshore sediments in the Irish Sea [20].

The depth profile of ²⁴¹Pu and ²⁴¹Am in sediment cores from the Irish Sea indicates that both ²⁴¹Pu and ²⁴¹Am concentrations increase with depth, reaching a peak at about 19–20 cm (Fig. 7). The ²⁴¹Am/²⁴¹Pu, which varies from 0.06 to 0.11, reached a maximum near the surface and at the depth of the ²⁴¹Pu and ²⁴¹Am peaks. The ²⁴¹Am in the profile is due to in-growth from the decay of ²⁴¹Pu, but the ratio does not necessarily reflect this. It is believed that mixing of sediments largely contributes to the changes in isotopic ratio of radionuclides. Studies have shown that the sediment mixing rate for the plutonium isotopes was approximately 90%.

Fig. 7

Data from Ref. [20]

Depth distribution profile in Irish Sea sediments: a distribution of ²⁴¹Pu, b distribution of ²⁴¹Am and c ²⁴¹Am/²⁴¹Pu.

The ²⁴¹Pu, like ^239+240Pu, is accumulated by marine organisms and plants. Plutonium in fine sediments is slowly remobilized by benthic organisms through sorption onto their cell surfaces or by accumulation within cells. This plutonium is ultimately transported from the sediments up through the food chain. The redistribution of plutonium from sediments back to the water column has been observed in several investigations of contamination from global fallout at the Enewetak and Bikini Atolls [80].

Studies have shown that in the Baltic ecosystems approximately 80% of plutonium can be found in zoobenthos, 10% in phytobenthos and the rest in phytoplankton, zooplankton and fish [88]. In general, plutonium bio-accumulation in marine organisms is higher compared to that in terrestrial organisms. The target organs and tissues for ²⁴¹Pu bio-accumulation are mainly the digestive gland, gill, and skeleton [89]. However, sorption and retention by benthic organisms is quite variable. The average ²⁴¹Pu concentrations in Baltic plankton, benthos and fish varied in the range 0.006–9.88 mBq/g [88], whereas those in Baltic seabirds varied between 0.01 and 0.228 mBq/g. Struminska-Parulska [89] measured bio-accumulation factors of 0.01–3.90 for Baltic fish and 0.8–99 for seabirds, algae and plankton. The distributions of ²⁴¹Pu in the marine environment follow the order: marine birds < fish < zooplankton < phytoplankton < zoobenthos < phytobenthos [87, 88]. Therefore, the uptake of ²⁴¹Pu into marine biota is dependent on the species. Sediment-dwelling species play a significant role in the remobilization process as they may redistribute plutonium deposited in the sediments by mixing and agitation. In general, levels of radionuclides in marine biota are related to the corresponding levels in seawater and sediment via accumulation through food chains. The distribution of ²⁴¹Pu in marine organisms generally follows the same trend as ^239+240Pu.

The Fukushima NPP accident introduced contamination into the marine environment through the deposition of the radionuclides released to the atmosphere and also through the direct discharge of thousands of tons of radioactive fluids into the western North Pacific Ocean. However, the amount of plutonium isotopes released directly into the marine environment remains unknown. Several studies have been conducted to find Fukushima-derived plutonium in the marine environment [10, 91]. However, none of these studies found elevated levels of plutonium that could be attributed to the Fukushima NPP accident. Studies of plutonium in marine sediments, sampled 30 km from the Fukushima NPP, suggest negligible contributions of the Fukushima NPP accident to plutonium in marine sediments. Both activities and isotopic composition in these samples reflect the presence of global fallout and Pacific Proving Ground (PPG) close-in fallout rather than plutonium from the Fukushima NPP accident. Similar findings were reported by Bu et al. [92] who have continued their marine monitoring work in the search for Fukushima-derived plutonium in the Pacific. Sediment cores collected from the western North Pacific had a ²⁴¹Pu activity of 7.42 mBq/g, which is typical of the background levels of ²⁴¹Pu detected in this region. Furthermore, the ²⁴¹Pu/^239+240Pu activity ratio (2.5) is comparable to the background ratio reported before the accident and was significantly lower than derived ratios associated with Fukushima. Several other studies had measured ²⁴¹Pu levels in marine sediments off the coast of Japan prior to the Fukushima NPP accident. The ²⁴¹Pu activity concentrations and ²⁴¹Pu/^239+240Pu activity ratios in these studies were in the range 3.3–8.4 and 1.1–1.9 mBq/g, respectively (ref. date: March 11, 2011) [34, 93] suggesting the presence of global fallout plutonium rather than Fukushima fallout. Furthermore, ²⁴¹Pu activities as high as 33.4 mBq/g (range 19.3–33.4 mBq/g), and ²⁴¹Pu/^239+240Pu activity ratios of 2.2–2.7 (ref. date: March 11, 2011) have been reported in sediments near the Bikini Atoll due to PPG close-in fallout [94]. However, contamination within the 30 km exclusion zone around the plant remains unknown, suggesting a need for further study of the few “hotspots” close to the Fukushima NPP. Additional research is required to fully understand the long-term effects of Fukushima-derived plutonium in the marine environment.

A few studies dedicated to actinides in environmental waters were published by Sakaguchi et al. [95] and Hain et al. [32]. These authors studied river water, paddy field water, and Pacific Ocean water for plutonium and uranium isotopes. Both the concentrations and isotopic composition of plutonium in sea water reflect the presence of global fallout and Pacific Proving Ground (PPG) close-in fallout rather than plutonium from the Fukushima NPP.

Summary and conclusion

Despite being the most abundant plutonium isotope released from the atmospheric nuclear tests and from the other sources, ²⁴¹Pu is still the most understudied isotope of plutonium. This review provides a summary of the sources, distribution, and behavior of shortest-lived, yet important, isotope of plutonium (²⁴¹Pu) found in the environment. The ²⁴¹Pu inventory in low-level nuclear waste and in environmental samples is of interest because ²⁴¹Pu is a precursor of other transuranium nuclides (²⁴¹Am and ²³⁷Np) that have longer half-lives, greater environmental mobility, and greater toxicity. In addition, the global inventory of ²⁴¹Am activity produced from the decay of fallout ²⁴¹Pu will reach about 60% of ²³⁹Pu in approximately 70 years. Vertical profiles of ²⁴¹Pu in soils and sediments exhibit distribution patterns similar to those of ^239+240Pu and ²⁴¹Am. However, ²⁴¹Am is observed to migrate faster than plutonium isotopes in some soil and sediment profiles.

Liquid scintillation counting remains a powerful technique for determination of ²⁴¹Pu because of its simplicity, high selectivity, and acceptable sensitivity and is expected to remain a frequently used technique in ²⁴¹Pu analyses for a long time. Over the last decade, because of the increasing sensitivity of ICP-MS and the improvement in sample introduction systems, the detection of ²⁴¹Pu has been achieved. Techniques like AMS, TIMS, and RIMS are very sensitive, but expensive, techniques that can be utilized for ²⁴¹Pu determination. A major advantage of mass spectroscopy techniques is that ²⁴¹Pu isotope can be measured together with other isotopes of plutonium without any additional sample preparation step. However, the short half-life and correspondingly high specific activity of ²⁴¹Pu favor its determination radiometrically.