Sources of 127I and 129I in aerosols
Variation of 127I and 129I concentrations in aerosols against
meteorological parameters (i.e. wind direction, wind speed, and temperature)
during the sampling period shows that wind direction has a principle
influence (Fig. 3). Back trajectory model analysis shows that 127I and
129I concentrations in the aerosols were directly controlled by the
sources and pathways of air masses (Figs. 6, S1 and S2 in the Supplement).
The relatively high 127I and 129I concentrations were observed in
the aerosols collected in early April 2011, when air masses were mainly
transported from the Atlantic Ocean across the North Sea by prevailing
westerly winds. While, relatively low concentrations of iodine isotopes were
observed in the aerosols collected in later April, when air masses were
dominated by prevailing easterly winds and passed over the European continent
and the Baltic Sea (Figs. 6 and S1). Marine emission of volatile iodine
species (e.g. inorganic I2 and HOI, organic CH3I, CH2I2)
is a major source of iodine in the atmosphere (Prados-Román et al.,
2015). It results in relatively elevated 127I concentrations in the
marine atmosphere, as compared to the terrestrial atmosphere (Saiz-Lopez et
al., 2012). During the sampling period of 11–14 April, air masses were
transported by westerly winds from a vast area over the North Atlantic Ocean.
This caused the highest 127I concentration in the sampling period.
Distinct from source of 127I, 129I concentrations in the aerosols
were significantly influenced by aqueous and gaseous discharge of the two
European NRPs at Sellafield (UK) and La Hague (France) that contribute more
than 95 % of 129I inventory in the environment. The two European
NRPs are located at west and southwest of Denmark, respectively (Fig. 1). Of
the total release of 129I from the two European NRPs, about 99 % of
129I (200–300 kg year-1 since 1995) has been discharged as
aqueous form to the English Channel and Irish Sea, respectively (Hou et al.,
2007; Raisbeck et al., 1995). Only a small fraction of 129I (about
0.5–2 kg year-1 after 2004 and 3–10 kg yr-1 in 1981–2004)
has been released as gaseous form to the atmosphere and dispersed over a
large area, in particular in Europe (Schnabel et al., 2001; Ernst et al.,
2003; Persson et al., 2007; Jabbar et al., 2012). Aqueous 129I was
carried by ocean currents and transported to the North Sea, Kattegat, and
Baltic Sea, and continues to the Arctic along the Norwegian coast (Alfimov et
al., 2004a; Buraglio et al., 1999; Hou et al., 2007; Raisbeck et al., 1995;
Yi et al., 2012). Remarkably elevated 129I concentrations of up to
1010–1011 atoms L-1 have been found in the North Sea (with
129I / 127I atomic ratios of 10-7–10-6),
109–1010 atoms L-1 in Norwegian coastal waters and the
Kattegat (with 129I / 127I atomic ratios of
10-8–10-7), and 108–109 atoms L-1 in the Baltic
Sea (Aldahan et al., 2007; Alfimov et al., 2004b; He et al., 2014; Hou et
al., 2007; Michel et al., 2012; Yi et al., 2011). Besides the directly
atmospheric releases of 129I from European NRPs at La Hague and
Sellafield, the relatively high 129I concentrations in the aerosols
influenced by westerly wind might also be attributed to the secondary
emission of marine discharged 129I in the European seas (the North Sea,
Irish Sea, Norwegian coastal water). The back trajectories analysis (Figs. 6
and S1) shows that the aerosol samples collected in 31 March–4 April,
4–7 April and 11–14 April were directly influenced by westerly and
southwesterly air masses that passed over not only the high 129I
contaminated North Sea, Irish Sea and English Channel, but also above
Sellafield and La Hague NRPs. However, 129I concentration and
129I / 127I ratio in aerosol collected in 4–7 April were
2.6-fold and 1.7-fold higher than that aerosols collected in
31 March–4 April and 1.6-fold and 2.1 fold higher than that in 7–11 April, respectively. This might be attributed to
the fact that most of air masses during 4–7 April mainly passed over both
Sellafield and La Hague as well as the high 129I contaminated North Sea,
Irish sea, and English Channel, while air masses during 31 March–4 April
passed over only Sellafield, south part of the highly 129I contaminated
of the North Sea and low 129I level North Atlantic Ocean, and air masses
during 7–11 April passed over Sellafield, north part of the highly
contaminated North Sea and low 129I level Atlantic Ocean. However, the
129I directly atmospheric releases from the European NRPs and secondary
emission from the 129I
contaminated sea cannot be isolated and identified based on the trajectories
analysis of these three samples, because the air masses of these three
aerosols passed over both reprocessing plant(s) and high contaminated seas.
For the aerosol collected in 14–18 April, the air masses passed over only
the north part of the highly contaminated North Sea and low 129I level
North Atlantic Ocean (Fig. 6c), the 129I / 127I ratio in this
aerosol is comparable with the samples collected in 11–14 April and
31 March–4 April ((4–5) × 10-7), and the 129I
concentration is higher than that in 31 March–4 April sample by a factor of
1.5. This result suggests that secondary emission of marine discharged
129I in the contaminated seawater (the English Channel, North Sea,
Kattegat and the Norwegian costal water) is the major source of 129I in
the westerly wind-influenced aerosols in Denmark in 2011 rather than the
directly gaseous release of the two European NRPs. A relative lower 129I
concentration (2.6 × 106 atoms m-3) and
129I / 127I ratio (2.8 × 10-7) were observed in
the aerosol collected in 7–11 April compared to the other four aerosol
samples influenced by westerly wind. This might be attributed to the fact
that the air masses during this period are mainly from the northwest and
mainly passed over to the lower 129I North Atlantic Ocean and small area
of the northern North Sea (Alfimov et al., 2004a).
The 129I / 127I atomic ratios in the studied aerosols
((1.8–8.6) × 10-7) fell well in the range of
129I / 127I ratios in the North Sea surface water in 2005 and
2009 ((1–10)) × 10-7) (Hou et al., 2007; Christl et al.,
2015). This also supports the above conclusion that the main source of
129I in the aerosols in Denmark in 2011 is re-emission of marine
discharged 129I in the surface water of the North Sea, Irish Sea, and
Kattegat.
There was a debate for a long time on either atmospheric releases or
re-emission of marine discharged 129I from the two European NRPs as the
dominant source of 129I in the European atmosphere (Ernst et al., 2003;
Hou et al., 2009a; Jabbar et al., 2011, 2012; Lopez-Gutierrez et al., 2004;
Michel et al., 2012; Persson et al., 2007; Reithmeier et al., 2010; Schnabel
et al., 2001). High 129I concentrations of
(3–445) × 108 atoms L-1 in rain water collected at
Zürich, Switzerland in 1994–1997, corresponding to estimated average
129I / 127I ratios of (3.2–4.0) × 10-7, have
been reported. Based on the estimated annual re-emission of 129I from
the contaminated European water (1.37 kg) in the 1990s and combined
atmospheric releases of 129I of about 6.5 kg in 1991–1996 from the two
European NRPs (Schnabel et al., 2001), as well as the relative positive
correlation of 129I concentration in the aerosols from Vienna, Austria
with the integrated monthly atmospheric releases of 129I from Sellafield
NRP (Jabbar et al., 2011), the direct atmospheric release of 129I was
proposed as the dominant source of 129I in the atmosphere in central
Europe (Schnabel et al., 2001; Jabbar et al., 2011).
An investigation on 129I in rain water (1994–2005) over Germany has
shown a high but slightly decreasing gradient of 129I concentrations and
129I / 127I ratios from coastal to inland areas (Krupp and
Aumann, 1999; Szidat et al., 2000; Michel et al., 2012). A high 129I
level (129I / 127I ratio of (2–8) × 10-7) was
also measured in air samples (gas and particle associated iodine) collected
at an island in the North Sea (Foehr) in April 2002, which is similar to the
129I / 127I ratio measured in the seawater surrounding the
Island (Ernst et al., 2003; Michel et al., 2012). This 129I level in the
aerosol is also comparable with those in precipitation in inland Germany. A
significant correlation of 129I with marine-derived chlorine
(R2 = 0.62) and bromine (R2 = 0.62) were measured in
precipitation in Denmark in 2001–2006 (Hou et al., 2009a). All these
investigations suggested that 129I in the European atmosphere, at least
in central and northern Europe, mainly originates from the re-emission of
marine discharged 129I in the European seawater (Krupp and Aumann, 1999;
Szidat et al., 2000; Michel et al., 2012; Ernst et al., 2003; Hou et al.,
2009a; Persson et al., 2007).
The reported 129I / 127I ratios in the atmosphere (aerosol and
precipitation) in central and northern Europe remain
constantly high from about
(3–4) × 10-7 at Zurich in 1994–1997 (Schnabel et al., 2001),
(3–14) × 10-7 in 1988–1995 (Krupp and Aumann, 1999; Bachhuber
and Bunzl, 1992) and (2–11) × 10-7 in 1997–2005 (Michel et
al., 2012; Szidat et al., 2000) in Germany, (0.8–5.2) × 10-7
in 2001–2002 in south Sweden, to (0.5–8) × 10-7 in 2001–2006
in Denmark. The 129I / 127I ratios measured in the investigated
aerosol samples in 2001 ((1.8–8.7) × 10-7) are also comparable
to the value in the precipitation in 2001–2006 in Denmark and other
locations in central and northern Europe in 1988–2005. However, the
atmospheric releases of 129I from the two NRPs at La Hague and
Sellafield have significantly reduced since 2002, from peak value of
5–10 kg yr-1 in 1981–2000 to less than 2 kg after 2004 (Michel et
al., 2012), if the contribution from another European NRP at Marcoule
(France), which was operated in 1960–1997, is included, the atmospheric
releases of 129I were 9–17 kg yr-1 in 1980–1997 (Reithmeier et
al., 2006). The insignificant influence of the remarkably reduced atmospheric
releases of 129I from the European NRPs since 2002 on the
129I / 127I ratios in the atmosphere in central and northern
Europe confirms that the atmospheric releases 129I from the two European
NRPs at La Hague and Sellafield is not the dominant source of 129I in
the European atmosphere, at least after 2002. However, this does not conflict
with the suggestion that 129I in the atmosphere in Zürich in
1994–1997 and Vienna in 2001 mainly originated from the atmospheric releases
of the European NRPs due to the relative high atmospheric releases of
129I from the European NRPs in 1980–2002, and
relatively lower marine discharges of
129I from the two European NRPs before 1990. The marine discharges of
129I from the two European NRPs has significantly increased from less
than 60 kg yr-1 before 1990 to more than 100 kg yr-1 in 1995
and keeping relative constant at about 250 kg yr-1 from 1998 until now
(Reithmeier et al., 2006). The 129I discharged to the English Channel
from the La Hague NRP and to the Irish Sea from the Sellafield NRP has
dispersed to a large area in the North Sea, Kattegat, Norwegian Sea, and
further to the Arctic by sea currents, and mainly remains in the surface
water (less than 100 m) in this area. Different from the atmospherically
released 129I which is quickly dispersed in the atmosphere and deposited
on the land in a short time (a few weeks/months), the marine discharged
129I is integrated in the marine system with much less and slow removal
from the water body to the sediment as particle associated form and to the
atmosphere in gaseous forms. The huge amount and constantly increased inventory of 129I in the
marine system has made the marine discharged 129I the major source of
129I in the present environment, therefore, the continuous re-emission of 129I from the contaminated
seawater to the atmosphere makes the marine discharged 129I the
dominated source in the atmosphere in Europe, and probably all over the world
after 2002. The very well matched trend of 129I / 127I ratio in
the atmosphere in central and northern Europe with the marine discharges of
129I from the two European NRPs (Michel et al., 2012) and constantly
high 129I level in the precipitation in the Northern Hemisphere outside
Europe (e.g. in Japan before 2011 and
China) also confirmed this assumption (Toyama et al., 2013; Zhang et al.,
2011). In addition, the measurement of 129I in the aerosol collected in
the high altitude Alps sites in 2001 and back trajectories analysis of air
masses reached to these locations have also confirmed that re-emission of
marine discharged 129I from the contaminated European seawater was the
major source of 129I in these aerosols (Jabbar et al., 2012).
Possible pathways of formation of iodide by reduction of sulfur
compounds.
Phase
Reactions
Eq.
References
Gas
DMS+OH→SO2
(1)
Chatfield and Crutzen (1990)
DMS+NO3→SO2
Gas/Aerosol
SO2+H2O→HSO3-
(2)
SO2+H2O→SO32-
Gas-Aerosol interface
I+HSO3-→I-+SO42-
(3)
I+SO32-→I-+SO42-
Aerosol
HOI+HSO3-/SO32-→I-+SO42-
(4)
Saiz-Lopez et al. (2012)
HOI+SO32-→I-+SO42-
HOI2+HSO3-/SO32-→I-+SO42-
HOI2+SO32-→I-+SO42-
The low 129I concentrations,
(11–13) × 105 atoms m-3, were observed in the aerosol
samples collected in 18–26 April and 26 April–2 May 2011. Back trajectory
analysis shows that in this period the air masses at the sampling site were
mainly transported by easterly or northwesterly winds, i.e. from the European
continent (Fig. S1). Terrestrial emissions of iodine occur through vegetation
and terrestrial microorganisms (Bewers and Haysom, 1974). The
relatively lower
129I / 127I ratios in the terrestrial samples in Europe
(10-8–10-7) (Osterc and Stibilj, 2013; Ezerinsk et al., 2016)
compared to that in the contaminated European seawater (10-7–10-6)
have been reported. This is reflected in these two aerosol samples by their
relatively low 129I concentrations. An elevated 210Pb level
(249–253 µBq m-3) (Table 1) for this period is also
consistent with a continental origin (210Pb in the air is a decay
product of 222Rn which is mainly released from the soil in the
continental area). However, it should be mentioned that the 129I level
in the European soil is still much higher than that in the uncontaminated
area such as in China and Chile (129I / 127I ratio of
10-10–10-9) (Zhang et al., 2011; Daraoui et al., 2012); this is
attributed to the high 129I level in the atmosphere in Europe.
Consequently, the 129I level in the aerosols with air masses from
terrestrial area is lower than that from the contaminated European seas, but
it is still much higher than that in other places with less contamination. It
should be mentioned that the conclusion of re-emission of the NRPs marine
discharged 129I in the contaminated seas as the dominant 129I
source in the atmosphere in Europe and other locations in the Northern
Hemisphere suggested in this work does not exclude the contribution of the
direct atmospheric releases of 129I from the NRPs. In the local area
surrounding the NRPs, the atmospheric releases might become the dominant
source.
The highest 129I concentration
(97.0 × 105 atoms m-3) found in 8–15 December 2014
might be related to the increased 129I re-emission from the North Sea
since 2011, where the air masses passed over (Fig. S2), as well as probably
increased liquid releases of NRPS,
which needs further investigations.
Species of 129I and 127I in aerosols
WSI is virtually pure iodide in the aerosols investigated, with iodate and
water-soluble organic iodine accounting for less than 3 % of total
iodine, and these are only measurable in two samples. Iodate was once
considered to be the only WSI species in aerosol (Vogt et al., 1999). This
was supported by earlier field observations demonstrating that iodate was
dominant in size-segregated aerosols from the tropical Atlantic (Wimschneider
and Heumann, 1995). However, this iodate-dominant feature was not found in
other aerosol samples, e.g. in the northwest Atlantic Ocean and in tropical
atmospheric aerosols (Baker, 2004, 2005). In these cases, iodide was the
dominant water soluble iodine species in the aerosols, as observed in this
study. Significant amounts of soluble organic iodine, accounting for
83–97 % of WSI, has been reported in aerosols collected at the Mace Head
atmospheric research station on the west coast of Ireland (Gilfedder et al.,
2008). Water-soluble organic iodine accounting for 4–75 % of WSI were
also measured in aerosols collected from a cruise from the UK to the Falkland
Islands in 2003 (Baker, 2005). This suggests that the proportion of soluble
organic iodine in aerosols varies regionally and depends on particular
aerosol sources and formation processes. Some of this variability might also
be related to the analytical methods employed for speciation analysis (Zhang
et al., 2015).
It is not clear how iodide is formed in the atmosphere, in an oxidizing
environment containing oxygen and ozone. Early models predicted a negligible
iodide concentration in particle phases based on the assumption that the
iodide in aerosols only originates from the low levels of gaseous HI
(McFiggans et al., 2000; Vogt et al., 1999). This work in combination with
the previous reports (Baker, 2004; Xu et al., 2015) suggests that there must
be other primary pathways that contribute to iodide formation at the observed
levels. It is generally accepted that iodine atoms are photochemically
produced by photolysis of gaseous iodinated compounds, and oxidized by ozone
to form reactive iodine oxides (Carpenter, 2003; Saiz-Lopez et al., 2012;
Vogt et al., 1999). The formation of iodide from iodine atoms and other
reactive iodine compounds must rely on electron-donors that are capable of
reducing high valence iodine species to iodide. One possibility is the
involvement of sulfur compounds (Chatfield and Crutzen, 1990). Possible
reaction pathways are given in Table 3. Gaseous SO2 can be formed in
nature by reactions of dimethyl sulfate (DMS) with hydroxide and nitrate.
Human activity is a major source of atmospheric SO2, globally about
three times as much SO2 as natural processes (Galloway, 1995). By
associating with H2O, these reactions produce HSO3- and
SO32- (Eqs. 1 and 2 in Table 3). Native iodine and other reactive
species (not shown) can be reduced to I- on gas-aerosol interfaces
(Eq. 3 in Table 3). Other iodine species in aerosols can be also reduced by
reductive sulfur compounds to form iodide (Eq. 4 in Table 3). The
electron-donors are not limited to sulfur compounds either, for example,
nitrogen in the form of gaseous ammonia (NH3 → NO2 / NO3)
(McFiggans et al., 2000; Saiz-Lopez and Plane, 2004) and elemental mercury
(Hg0 → HgO / HgX, where X is a halogen, I-, Br- or
Cl-) (Lindberg et al., 2002; Simpson et al., 2007) are also candidates
responsible for iodide formation.
We note that the percentage of WSI 129I and 127I in marine-sourced
aerosol from the North Sea is relatively lower than that in the continental
aerosol, as compared to the European continent-sourced aerosols (Figs. 4 and
5). This is consistent with the findings drawn from an iodine speciation
study of coastal aerosol samples from England (Baker et al., 2001), where the
concentrations of total water-soluble iodine from continental aerosols were
significantly higher than those from marine aerosols.
A large proportion of 129I and 127I in our aerosol samples were
NaOH-soluble, which is consistent with the results of aerosol from Tsukuba,
Japan, collected shortly after the Fukushima nuclear accident in March 2011
(Xu et al., 2015). Abundant NaOH-soluble 129I (32–44 % of total
129I) in Fukushima-derived aerosols was attributed to coarse
vegetation-related organic particles concentrated during spring. The measured
NaOH-soluble iodine (NSI) fractions of 129I and 127I during the
entire sampling period in the spring of 2011 are similar. This indicates that
NSI is relatively stable and less affected by the source and pathways of air
masses than WSI. NaOH leaching is often used to extract organic substance in
fractionation analysis of soil and sediment (Englund et al., 2010a; Hou et
al., 2003) based on the high solubility of organic matter, such as humic
substances, as well as on nucleophilic substitution and decomposition of
organic matter. Organic compounds are important contributors to aerosols,
such as lipidic, saccharides, and proteinaceous materials (O'Dowd et al.,
2004; Quinn et al., 2014). A significant portion of atmospheric aerosols was
found to be humic-like substances (HULIS), named for their strong structural
similarity to humic and fulvic acids (Havers et al., 1998). Most of these
organic compounds are water-soluble, but a significant water-insoluble
fraction of the HULIS material is hydrophobic and acidic in character, and
can be dissolved in an alkaline solvent, like NaOH (Feczko et al., 2007;
Havers et al., 1998). On the other hand, the hydroxide anion can also
initiate a nucleophilic substitution or elimination of iodine-containing
organic compounds, which releases iodine from the organic substances.
NaOH-soluble iodine is therefore suggested to be likely associated with
organic substances in aerosols.
RII in aerosols has received less attention than WSI (Gilfedder et al., 2010;
Tsukada et al., 1987). The early report on water-insoluble iodine fraction in
aerosol particles showed that water-insoluble iodine accounted for
27–58 % of total iodine bound in aerosols from Tokyo, Japan, collected
in 1983–1984 (Tsukada et al., 1987). Another similar result of 17–53 %
of total iodine as insoluble species was reported for aerosols from the west
coast of Ireland in 2007, and from a ship transect from China to Antarctica
in 2005–2006 (Gilfedder et al., 2010). Taking the alkaline-leachable iodine
in aerosols into account, these results are very consistent with our
observations from Risø (Fig. 5). The residual insoluble 129I
fractions were reported to be 4–23 % of total 129I in
Fukushima-derived aerosol particles (Xu et al., 2015), less than the
proportion in the aerosols collected in Denmark. This discrepancy might
reveal the different formation processes of RII species for the NRPs-derived
129I in this study as compared to those from Fukushima. A significant
difference is the timing of the 129I releases. NRPs have releasing
129I into the European environment for about 50 years, allowing
129I to follow geochemical pathways on timescales ranging from days to
decades. In contrast, RII in Fukushima-derived aerosols had only days to
react with their environment prior to sampling, 15–22 March 2011.
The origin of the RII fraction is not well understood at present. It is
possible that part of the RII fraction is derived from suspended soil
particles (Xu et al., 2013). It has been demonstrated that iodine can be
associated with metal oxide (notably iron and manganese). A relatively large
fraction of iodine (about 38 %) in soil and sediment has been observed in
Fe / Mn oxides associated form (Hou et al., 2003). Our data show that RII
fraction is as high as 67 % of total aerosol iodine. In addition to metal
oxides associated iodine, speciation analysis of 129I in soil shows that
residual iodine after leaching with NaOH and weak acid accounts for less than
10 % of the total, and this component is assumed to be associated with
minerals (Hou et al., 2003; Qiao et al., 2012). As stated above, the aerosols
collected in early April 2011 were mainly marine-derived aerosols with
relatively higher RII percentage than those continental-derived aerosols
(Fig. 5). This might be attributed to the fact that some marine components facilitate the
association of gaseous iodine with oxides and minerals.
131I radioactivity (red, Nielsen et al., 2011), 129I
concentrations (blue) in aerosols from Risø, Denmark after the Fukushima
accident. The Fukushima-derived 129I concentrations are calculated based
on 129I / 131I atomic ratio of 16.0 ± 2.2 deduced from
Fukushima-affected aerosol samples (Xu et al., 2015).
Fukushima-derived 129I signal in the European atmosphere
The Fukushima Dai-ichi nuclear power plant accident on 11 March 2011 released
radioiodine to the atmosphere, primarily as 131I and 129I, which
was mainly transported eastwards by prevailing westerly winds. Based on
129I levels in the Fukushima offshore seawater, the released 129I
from this accident was estimated to be 1.2 kg (Hou et al., 2013).
Radioactive iodine in the air dispersed across the Pacific Ocean, American
continent and Atlantic Ocean, and some fraction arrived in the European
continent after 1–2 weeks (Clemenza et al., 2012; Leon et al., 2011;
Manolopoulou et al., 2011; Pittauerová et al., 2011). Anthropogenic
129I has been reported from a variety of environmental samples in Japan,
including soil, seawater, precipitation, and aerosols (Buesseler et al., 2012;
Hou et al., 2013; Muramatsu et al., 2015; Xu et al., 2013, 2015). The level
of 129I in aerosols collected in Tsukuba, about 170 km from the
Fukushima Dai-ichi NPP, reached 5 × 108 atoms m-3 (Xu et
al., 2015). While the Fukushima-derived 129I in environmental samples
was less presented outside of Japan. 131I in the aerosol samples
collected at Risø, Denmark, 10 days after the Fukushima accident have been
observed (Fig. 7) (Nielsen et al., 2011). The radioactivity of 131I
reached the peak on 24–30 March 2011 (763 µBq m-3 in
aerosol), then fell to below detection limits for 131I in the middle of
May. Based on the measured 131I radioactivity in the aerosol samples and
an 129I / 131I atomic ratio of 16.0 ± 2.2 deduced from
the aerosol samples collected at Tsukuba, Japan shortly after the Fukushima
accident (Xu et al., 2015), the Fukushima-derived 129I signal in Denmark
can be reconstructed (Fig. 7). The highest 129I concentration in the
aerosols in Denmark from the Fukushima accident is estimated to be
6.3 × 104 atoms m-3 on 30–31 March 2011, which accounts
for less than ∼ 6 % of total 129I
(1.1–9.7 × 106 atoms m-3) in Denmark when the Fukushima
131I peak was measured. Considering the rapid decline of 129I
levels in aerosols and precipitation in Japan to nearly pre-accident levels
within 2 years (Xu et al., 2013), the contribution of Fukushima-derived
129I to the 129I level and inventory in the Europe is now
negligible in comparison to NRPs-derived 129I in the European
atmosphere.