Analysis of 129 I and 127 I in soils of the Chernobyl Exclusion Zone, 29 years after the deposition of 129 I

The Chernobyl Exclusion Zone (CEZ) represents a unique natural laboratory that received signi ﬁ cant 129 I contamination across a range of soils and land-use types in a short time period in 1986. Data are presented on 129 I and 127 I in soil samples collected from highly contaminated areas in the CEZ in 2015. The geometric mean (GM) total concentration of stable iodine ( 127 I) was 6.7 × 10 − 7 g g −


Analysis of
129 I and 127 I in soils of the Chernobyl Exclusion Zone, 29 years after the deposition of 129 I

H I G H L I G H T S
• Chernobyl Exclusion Zone is a unique natural laboratory for studying radionuclides.• Natural and anthropogenic iodine isotopes as well as 137 Cs were investigated.• 129 I not in complete equilibrium with 127 I 29 years after the Chernobyl accident • Native 127 I only a partial proxy for 129 I behaviour in soils even after three decades • Greater downwards leaching was observed for 129 I than for 137 Cs.

G R A P H I C A L A B S T R A C T
a b s t r a c t a r t i c l e i n f o

Introduction
In the aftermath of the Chernobyl accident in 1986 the short-lived radionuclide 131 I (half-life 8.02 days) was dispersed widely in the atmosphere, depositing to the land surface across the European continent and beyond.In some areas, notably in Belarus, radiation doses from 131 I to the human population were sufficient to induce excess thyroid cancers, predominantly in young people (WHO, 2016).These doses were received over a period of days to weeks, after which 131 I activities in the environment became undetectable.A much longer-term legacy of the Chernobyl accident, however, is the presence in the environment of 129 I (half-life 15.7 × 10 6 years), an estimated 1.3 kg of which were released in 1986 (Paul et al., 1987).Unlike 131 I, 129 I emits only a very weak gamma ray (39.6 keV) and, due to its long physical half-life, has a low specific activity (6.54 × 10 6 Bq g −1 , cf. 4.60 × 10 15 Bq g −1 for 131 I).Estimates of exact doses from 131 I after the Chernobyl and Fukushima accidents have been hampered by the narrow window of time available for 131 I measurements in the environment.Since 131 I and 129 I are emitted synchronously to the atmosphere during an accidental release, the lingering presence of 129 I in environmental media has proved useful in reconstructing 131 I deposition patterns and associated doses (Straume et al., 1996;Pietrzak-Flis et al., 2003;Muramatsu et al., 2015).
Despite its usefulness in radiation dose reconstruction, 129 I is a significant long-term source of radiation exposure due to its extreme persistence in the environment.For adults ingesting contaminated foodstuffs the committed effective dose per unit intake for 129 I (1.1 × 10 −7 Sv Bq −1 ) is 5 times higher than that of 131 I (2.2 × 10 −8 Sv Bq −1 ) (ICRP, 2012) and 129 I is of specific concern in the context of radioactive waste disposal.In safety assessment calculations, 129 I is treated as a poorly sorbed radionuclide which contributes significantly to long-term radiation dose in the biosphere above geological disposal facilities (Posiva, 2013;SKB, 2014).Understanding the long-term environmental behaviour of 129 I is, therefore, of considerable importance in quantifying and controlling risks from such facilities and some studies addressing this problem have used 127 I as a proxy for 129 I (Roulier et al., 2019).
Iodine is a relatively active and mobile element in the environment.Although 129 I and 127 I in soils have different sources, they are expected to behave in a very similar manner.The great majority of 129 I in the Chernobyl Exclusion Zone (CEZ) is expected to have originated from rapid, localised deposition of Chernobyl-released iodine in 1986, while 127 I has accumulated over the long-term (centuries to millennia) from deposition of atmospheric iodine originating from distant marine sources.Weathering of soil parent material can also contribute some 127 I to soils (though usually a minor source) and turnover of soil organic matter is a key part of the long-term accumulation and redistribution process of iodine in soils (Roulier et al., 2018).Following deposition on the soil, both 129 I and 127 I experience similar processes of fixation and leaching.It is expected that anthropogenic 129 I, deposited in the very short-term, should reach equilibrium with the numerous chemical species of naturally-occurring 127 I after some period of time through dynamic chemical and biochemical processes; thereafter the two iodine isotopes would be expected to behave in the same way.This paper aims to investigate if such an equilibrium has been achieved by speciation (fractionation) analysis of these two isotopes of iodine.
Our objectives in this study were to answer the following two key questions i) what are the physico-chemical forms of 129 I in soils over the long-term and ii) what is the time scale over which 129 I equilibrates with the stable isotope 127 I that is present naturally within the soil?To answer these questions contaminated soils within a 30 km radius of the Chernobyl nuclear power station were investigated.This area presents unique circumstances in which to make measurements of 129 I three decades after the world's worst nuclear accident.We hypothesised that, after residing for almost 30 years in the soils of the CEZ, 129 I would have achieved a physico-chemical equilibrium with the native, stable 127 I present in the soils: it should be possible to demonstrate such an equilibrium by comparing the relative extractabilities of both iodine isotopes from soils.The results of analyses of 127 I and 129 I in these samples are reported, with an emphasis on the comparative behaviour of both nuclides 29 years after the deposition of 129 I. Activity concentrations of 137 Cs and relevant soil properties, especially organic matter, are useful in interpreting the behaviour of iodine isotopes in soil and these are also reported.The data set provides a comprehensive analysis of the comparative extractabilities of 127 I and 129 I in soils of the CEZ, three decades after contamination with 129 I.

Soil sampling
Soils were sampled within the CEZ on 1st and 2nd September 2015 at the locations shown in Fig. 1 and Table 1.Soil samples were taken close to the centre line of the plume of atmospheric deposition which emanated from the explosion in Reactor 4 on 26th April 1986 and extended due west (the 'Western Trace').The sampling line extended from the 'Red Forest', 2.5 km west of Reactor 4, where the heaviest radioactive deposition occurred, to an abandoned collective farm at Tolsty Les, 22 km west of Reactor 4. Samples were also taken on the 'North Trace', an area of high radionuclide deposition 8 km to the NNW of Reactor 4. Locations of sampling sites were recorded using a hand-held GPS unit (Garmin GPSMAP® 60CSx).Air kerma rates (μGy h −1 ) were recorded 1 m above the ground surface using a Mini Instruments Environmental Radiation Meter (type 6-80).
At each sampling site a soil pit was dug to a depth of 20 cm.Topsoil samples were taken with a trowel from a depth of 0-10 cm and subsoils from 10 to 20 cm.Soils were immediately placed in polythene bags, sealed and taken to the Chornobyl Center for Nuclear Safety, Radioactive Waste and Radioecology in Slavutych, 47 km NE of the Chernobyl power station.Here, the samples were stored in a cold room before processing.

Soil moisture content, pH, loss on ignition (LOI), and gamma ray analysis
Sub-samples of field moist soil (5-10 g) were used to determine moisture content (by oven drying at 105 °C), pH and loss on ignition (LOIthe mass lost after combustion at 550 °C as a proxy for total soil organic matter).Air-dried sub-samples (10 g) were used to determine total 137 Cs activity concentrations using a Canberra-Packard gammaspectrometer with a high-purity germanium (HPGe) detector (GC 3019).A standard source (OISN-16; Applied Ecology Laboratory of Environmental Safety Centre, Odessa, Ukraine) containing 152 Eu was used to calibrate the efficiency of the detector.Sample count rates were not decay corrected so represent activity concentrations just after the time of sampling.Sub-samples of the remaining field-moist soils were subject to a sequential extraction procedure, as described below.

Sequential extraction of soils
The extraction method is based on one published by Zhao and McGrath (1994) for sulphate extraction that was subsequently adapted by Martens and Suarez (1997) for Se extraction.The method is designed to extract soluble iodine species using 0.01 M KCl, then specifically adsorbed species using 0.18 M KH 2 PO 4 and finally organically-bound iodine using 2.5 M NaOH.A mass of each field moist soil sample equivalent to 6 g oven-dry material was weighed into a polythene centrifuge tube.After adding 30 mL of 0.01 M KCl to each tube, the tubes were shaken for approximately 16 h then centrifuged at 3500 rpm for 30 min.The supernatants were carefully removed using a syringe and filtered through a 0.22 μm PTFE filter.A 19.8 mL aliquot of filtered KCl extract was pipetted into a polythene tube and 0.2 mL of 10% w/v NaOH added to stabilise the iodine in the extract.The excess supernatant was carefully removed and 30 mL 0.18 M KH 2 PO 4 added to the remaining soil pellets.The tubes were shaken vigorously by hand to disaggregate the pellets and then shaken gently for approximately 16 h using an orbital shaker.The soil suspensions were centrifuged at 3500 rpm for 30 min before removing and filtering the supernatants, as for the first extraction step.An aliquot (19.8 mL) of filtered KH 2 PO 4 extract was pipetted into a polythene tube and stabilised with 0.2 mL of 10% w/v NaOH.After removing the excess supernatant, remaining soil pellets were subjected to a third and final extraction in 15 mL of 10% w/v (2.5 M) NaOH, added to each centrifuge tube.The tubes were shaken to disaggregate the pellet then loosely capped, placed in a rack and incubated at a maximum temperature of 90 °C for approximately 16 h.After incubation, the tubes were centrifuged at 3500 rpm for 30 min then 2 mL of NaOH extract removed and transferred to a clean polythene tube.The NaOH extracts were diluted with 18 mL of ultrapure water to give a solution of 1% w/v NaOH.After each extraction step the tubes containing the respective extracts were capped and sealed with Parafilm® prior to transport to the UK.A flow diagram illustrating the sequential extraction procedure is provided in Supplementary materials (Fig. S1).  1. order of 2 min.These aliquots were used to analyse 127 I by ICP-MS (Thermo-Fisher Scientific, iCAP-Q) with Rh and Re (10 mg L −1 ) as internal standards for drift correction.Stock standards for 127 I calibration were prepared at iodine concentrations of 1000 mg L −1 from ovendried KI and KIO 3 , and stored at 4 °C in 1% tetra methyl ammonium hydroxide (TMAH).Two vials containing 20 mL of 0.5 M (2% w/v) NaOH were left open in the laboratory for 48 h to assess the (worst case) contamination from 129 I which had previously been used as a tracer in this laboratory; these were used as reagent blanks when preparing AgI targets for 129 I analysis (see below).The remaining soil extracts and reagent blanks were sealed in the polythene tubes and transported to the Technical University of Denmark for separation of 129 I from each extract and preparation of targets for AMS measurement of 129 I.

Preparation of targets for AMS measurement of 129 I
Aliquots of 5.0-18 mL of each extract were transferred to a beaker and the mass of solution recorded.A carrier solution of 127 I (1.802 mg mL −1 ) was prepared by dissolving iodine crystal (Woodward Company, USA; 129 I/ 127 I ratio b 2 × 10 −14 ) in a solution consisting of 0.4 M NaOH and 0.05 M NaHSO 3 .A 0.5 mL aliquot of this carrier was added to the sample, followed by 8 M NaOH to achieve a NaOH concentration of 0.5 M. The beaker was covered with a watch glass then heated at 60 °C for 10-12 h to convert all organic iodine to an inorganic form.After cooling, 0.5 mL of 1 M NaHSO 3 solution was added and the pH adjusted to b2 using 3 M HNO 3 , to convert all iodine to iodide.The sample solution was then transferred to a separation funnel to which 20-30 mL of CHCl 3 were added, followed by addition of 0.5-0.8mL of 1 M NaNO 2 to oxidize iodide to I 2 .Iodine as I 2 was extracted into the organic phase which was then separated and collected in a beaker.The organic extraction step was repeated by adding another 15 mL aliquot of CHCl 3 to the separation funnel, plus 0.1 mL of 1 M NaNO 2 , and the two organic phase extracts were combined.The CHCl 3 phase solution was then transferred to another separation funnel to which were added 0.2 mL of 0.05 M NaHSO 3 and 15 mL H 2 O to reduce I 2 to iodide (I − ) and back extract it into aqueous phase, after which the organic phase was discarded.These organic solvent extraction and back-extraction steps were repeated, leaving an aqueous phase containing iodide.This was transferred from the separation funnel to a 10 mL centrifuge tube.The funnel was then washed three times with water and the washings combined in the centrifuge tube.
A 1 mL aliquot of 3 M HNO 3 was added to the centrifuge tube, followed by 0.5 mL of 1 M AgNO 3 to form a AgI precipitate which was then separated by centrifugation.After washing the AgI precipitate with water it was centrifuged again, then transferred with 0.5 mL water to a 1.4 mL centrifuge tube.The original centrifuge tube was washed and the washings combined with the AgI precipitate.This was centrifuged and the supernatant discarded.Finally, the AgI precipitate was dried at 60 °C for 2-3 h in readiness for AMS measurement of 129 I.

Analysis of 129 I by accelerator mass spectrometry
After preparation, the AgI targets were transported to the Scottish Universities Environmental Research Centre (UK) where 129 I concentrations were determined using a tandem accelerator mass spectrometer (AMS), as previously described by Xu et al. (2013).The AgI precipitates from each sample were mixed with high purity Ag powder (100 mesh, 99.95%, Assure) with a mass ratio 1:2 for AgI:Ag and then pressed into an aluminium target holder (1 mm ∅).Iodide (I − ) ions were extracted using a Cs-sputtering ion source.A terminal voltage of 3 MV was used and I 5+ ions selected for detection.The 127 I 5+ ions were detected using a Faraday cup mounted at the exit of a high energy analyzing magnet, while 129 I 5+ ions were counted using an ionization chamber detector with a 100 nm thick SiN detector window.Dissociation of MoO 2 − can produce 97 Mo 4+ which may interfere with 129 I 5+ measurement due to a similar magnetic rigidity (ME/q (Bowley et al., 2016)); however, the two ions can be completely separated in the detector.The measured 129 I/ 127 I ratios were corrected against a standard material with 129 I/ 127 I ratio of 1.138 × 10 −10 prepared by 127 I addition to the NIST 4949B standard.The measured 129 / 127 I ratios in the prepared targets ranged from 2.9 × 10 −11 − 3.5 × 10 −9 , which are higher than the mean of the reagent and laboratory blanks (1.1 × 10 −11 ) and the estimated pre-atomic ratio (1.5 × 10 −12 ) (Moran et al., 1999).I-129 concentrations in the samples were calculated based on the amount of 127 I carrier added to the samples before chemical separation and the measured 129 I/ 127 I atomic ratios after subtracting the background of the reagent and laboratory blanks.Uncertainties on measured 129 I/ 127 I ratios in samples ranged from 0.4 to 2.8%.After subtracting 129 I/ 127 I ratios attributable to the reagent blank, a limit of detection (LoD) of 2.0 × 10 5 129 I atoms per sample was calculated.I-129 in all samples was above this LoD, ranging from 7.7 × 10 7 -5.3 × 10 10 atoms per sample.

Results and discussion
3.1.Total concentrations of 127 I and 129 I in CEZ soils Total concentrations of 127 I and 129 I in our samples were calculated by summing the concentrations determined in KCl, KH 2 PO 4 , and NaOH extracts (see Supplementary material, Tables S1 to S4); concentrations were expressed as either mass concentrations (for both 127 I and 129 I) or activity concentrations (for 129 I) per unit dry weight of soil.Total concentrations of stable iodine ( 127 I) in the CEZ soils ranged from 1.54 × 10 −7 g g −1 in the B1 subsoil to 2.64 × 10 −6 g g −1 in the RF3 topsoil, with a geometric mean (GM) of 6.7 × 10 −7 g g −1 (Fig. 2(i), Table S4).This range is below the average soil concentration of 3.94 mg kg −1 (3.94 × 10 −6 g g −1 ) for iodine in European soils (Salminen et al., 2005) and at the lower end of the range for world soils (Fleming, 1980).Sahoo et al. (2009) took soil samples in 1994/95 and 2001 from sites in the CEZ which covered a very similar geographical range to the samples taken in 2015; soils and vegetation types at the sampling sites are almost identical between the two studies.They also reported low 127 I concentrations in CEZ soils (mean = 5.4 × 10 −7 g g −1 ) similar to the concentrations determined here.Shiraishi et al. (2006) found that dietary intake of iodine by residents in NW Ukraine was below the WHO recommended value of 150 μg d −1 for adults and was linked to increased occurrence of goitre; this probably reflects low mean regional iodine concentrations that are likely to increase exposure to environmental sources of radioactive iodine.Total mass concentrations of 129 I in the CEZ soils ranged from 2.40 × 10 −14 g g −1 in the B1 subsoil to 3.83 × 10 −12 g g −1 in the RF3 topsoil, with a GM of 2.39 × 10 −13 g g −1 (Fig. 2(ii), Table S4).When expressed as activity concentrations this range was 0.16 to 25 mBq kg −1 with a GM of 1.56 mBq kg −1 (Fig. 2(iii), Table S4).Sahoo et al. (2009) reported a range of 1.7 to 168 mBq kg −1 in soils from the CEZ, with the highest activity concentrations in surface soil samples including forest litter and organic horizons.The slightly lower activity concentrations in our study are consistent with the fact that our samples were averaged across soil depths of 0-10 and 10-20 cm.For the same reason, the range of 129 I/ 137 Cs atom ratios in our samples (0.011-0.566,Table S5) is lower than the ranges reported by Hou et al. (2003) and Sahoo et al. (2009) (overall range from 0.1 to 7.2), though all three ranges overlap.Mironov et al. (2002) determined an average pre-Chernobyl activity concentration for 129 I of 0.094 (± 0.014) mBq kg −1 in subsoil samples collected 400 km north of Chernobyl in 1985; 129 I activity concentrations in all our samples were significantly higher than this and it can be concluded that the 129 I measured in the soil samples originated from the Chernobyl reactor.A further indication of the source of 129 I in our study is given by the strong and highly significant correlation between 129 I and 137 Cs activity concentrations in the samples (Fig. 3); a similar relationship has previously been proposed as a means of estimating 131 I deposition densities in areas of Russia, Belarus, and Sweden contaminated by the Chernobyl accident (Hou et al., 2003).

Relationships between 127 I, 129 I and organic matter in CEZ soils
The RF3 site is a peaty meadow (Table 1) at the western extremity of the Red Forest and the soil here had the highest organic matter content (LOI) of all the samples taken (Table S5), as well as the highest 127 I and 129 I concentrations (Table S4).Overall, there was a significant positive relationship (Pearson r = 0.73; Spearman rank p = 0.037) between 127 I and soil organic matter (Fig. 4(i)), as expected from previous studies on soil iodine (e.g.Yamada et al., 1999;Xu et al., 2016;Söderlund et al., 2017).Total concentrations of 127 I and 129 I were positively and significantly correlated (Fig. 5) resulting from a similar tendency for retention of both isotopes in soils with similar organic matter contents.However, the relationship between 129 I and soil LOI (not shown) was much weaker (Pearson r = 0.44) than for 127 I versus LOI.Spearman rank correlation indicated an insignificant relationship between 129 I and soil LOI (p = 0.22); when the RF3 data points were removed the slope of the relationship was effectively zero, suggesting that the variations in 129 I concentrations in the soil are due primarily to the initial deposition pattern in 1986 rather than the properties of the soil on which the 129 I was deposited.

Topsoil versus subsoil concentrations of 127 I and 129 I in CEZ soils
Total topsoil concentrations of 129 I and 127 I were generally higher than subsoil concentrations, but not at all sampling sites; 129 I concentrations were greater in topsoils at more sites than 127 I and the differences between top-and subsoil concentrations were higher for 129 I than for 127 I (Table S4).Furthermore, ratios of total 129 I/ 127 I were higher in topsoils than in sub-soils at all sites except RF1 (Table S4), indicating that proportionally more of the 129 I deposited to the soil surface in 1986 was still present in the topsoil.Sahoo et al. (2009) reported generally declining 127 I and 129 I concentrations from the soil surface to a depth of 15-20 cm in samples taken on the North Trace in 2001.However, some downwards migration of 129 I must have occurred in the 30 year period following the Chernobyl accident.Independent estimates of ~30 years have been made for the residence half-times of 129 I in the upper 30 cm of soil profiles at Savannah River (Boone et al., 1985) and the Karlsruhe reprocessing plant (Robens et al., 1989) suggesting that significant downwards migration of 129 I in CEZ soils is possible since initial deposition in 1986.This is supported by the observation that the 129 I/ 137 Cs atom ratios in topsoils are lower than those in subsoils for 7 of the 9 soils sampled in this study indicating that 129 I has moved faster than 137 Cs from the topsoil to the subsoil.Leaching of a portion of the originally deposited 129 I from the upper 20 cm of the soils sampled cannot be excluded, which would partially explain the slightly lower total 129 I activity concentrations we measured compared with previous studies, as discussed in Section 3.1.Approximate estimates of the magnitude of 129 I leaching below a depth of 20 cm may be possible by comparing 129 I activity concentrations measured by Sahoo et al. (2009) and in our study.Furthermore, if the rate of 137 Cs leaching were known than the comparative rate of 129 I could be calculated using the 129 I/ 137 Cs atom ratios in Table S5.However, the only reliable way to quantify the degree of 129 I leaching deep into the soil profile is to take deeper soil samples for further analysis by AMS (beyond the scope of this study).

Sequentially-extractable concentrations of 127 I and 129 I in CEZ soils
In this study 10% w/v (2.5 M) NaOH was used to recover humic matter and associated iodine isotopes from soil samples.Studies by Watts and Mitchell (2009) and Shetaya et al. (2012) have shown that strongly alkaline extractants such as TMAH and NaOH are effective in quantitatively extracting iodine from soils.Total concentrations of both 127 I and 129 I (defined above) were dominated by the NaOH-extractable fraction that represents iodine associated with soil organic matter (Fig. 2 (i) and (ii); see also Supplementary material).In the case of 127 I, the organically-bound concentrations were significantly higher than both KCl and KH 2 PO 4 extractable concentrations (Fig. 2(i)), but not significantly different from the total 127 I concentration.The organicallybound concentrations of 129 I were significantly higher than the KCl extractable concentrations, but not the KH 2 PO 4 extractable concentrations (Fig. 2(ii)).Extraction of soils with a weak electrolyte such as 0.01 M KCl accesses the readily soluble pool within the soil; the dominant iodine species expected to reside in this pool are iodide (I − ) and iodate (IO 3 − ).Yuita (1992) found that iodate accounted for N80% of the iodine in soil solutions from a brown forest soil under aerobic conditions, but only ~14% when the soil was flooded.In contrast, iodide accounted for ~6% of soluble iodine in non-flooded forest soil, increasing to N80% in flooded soil.At the time of sampling (September 2015) the CEZ soils were generally dry (gravimetric moisture contents ranged from b0.2% in the sandy B1 subsoil to 61% in the highly organic RF3 subsoil).Even though the soils of the CEZ are generally free-draining, they frequently flood after snow melt during early spring or for longer periods (e.g.2013-2014) so it is probable that soluble inorganic iodine species vary throughout the year.Laboratory studies by Shetaya et al. (2012) and Duborska et al. (2019) indicated that iodide and iodate were lost from soil solutions due to interactions with reactive surfaces such as hydrous oxides and humic molecules on time scales of min to h and h to days, respectively, suggesting that equilibration of inorganic iodine species in soil solution is complete on these time scales.Longer-term kinetics under environmental conditions cannot be ruled out, however.
1.E-07 Iodide interacts with anion-exchanging surfaces in soils by electrostatic interaction, iodate by chemisorption principally to metal hydrous oxides (Yoshida et al., 1995); extraction of soils with KH 2 PO 4 provides a measure of iodide and iodate specifically adsorbed to sesquioxides of Fe, Al and Mn.The GM adsorbed 127 I concentration was significantly different (higher and lower, respectively) from GM soluble and organic 127 I concentrations (Fig. 2(i)).However, the GM adsorbed 129 I concentration was significantly higher than soluble 129 I, but not significantly different from organic 129 I concentrations (Fig. 2(ii), (iii)), indicating a difference in the physico-chemical distribution of native 127 I and Chernobylderived 129 I.

Comparison of extractabilities of 127 I and 129 I in CEZ soils
The relationships between 127 I and 129 I in the soil extracts are summarised graphically as 129 I/ 127 I ratios in Figs.2(iv) and 6.Fig. 2(iv) shows that 129 I/ 127 I ratios in KCl and KH 2 PO 4 extracts were not significantly different from each other, with GMs of 1.74 × 10 −6 and 1.66 × 10 −6 , respectively.The GM 129 I/ 127 I ratios in NaOH extracts were significantly lower (2.37 × 10 −7 ) and, since this was the dominant fraction (Fig. 2(i) and (ii)), the GM of the total 129 I/ 127 I ratio was also significantly lower than the KCl and KH 2 PO 4 extracts (2.94 × 10 −7 ).Fig. 6 shows that the 129 I/ 127 I ratios in all soil extracts were positively and significantly related to the total 129 I concentrations in the soils.Since 129 I/ 127 I ratios for KCl and KH 2 PO 4 extracts were not significantly different, a single regression line was fitted to these data in Fig. 6.It is clear that 129 I/ 127 I ratios in KCl and KH 2 PO 4 extracts are higher than in the NaOH extracts at all 129 I concentrations, indicating a preponderance of 129 I in more 'labile' soil extracts compared with soil organic matter (extractable with NaOH) that can be considered to be the long-term sink for iodine in soils.Bowley et al. (2016) observed that mixtures of I − and IO 3 − in suspensions of humic acid were partially transformed to organic iodine over the course of 77 days.Kinetic modelling suggested a short half-time (~15 min) for conversion of IO 3 − to organic iodine, though this was balanced by a slower reverse reaction which led to a 'pseudo-steady-state' over a time scale up to 250 days.Size selective analysis of humic acid by Xu et al. (2012) showed proportionally less 129 I than native 127 I in the larger molecular weight humic fraction, attributable to insufficient reaction time for full isotopic equilibration and to partial exclusion of 129 I from the complex humic acid structure.Our results from samples collected almost 30 years after 129 I deposition in the CEZ suggest that full isotopic equilibration between 129 I and 127 I in soil humic molecules has not been achieved on a decadal time-scale.Schmitz and Aumann (1995) found that 129 I was proportionally more water-soluble than 127 I in soils around the Karlsruhe reprocessing plant that they explained by the much shorter residence time of 129 I in the soil compared with native 127 I. Surprisingly, only 4-15% of 129 I was found in the organic fractions of the Karlsruhe soils.In contrast, Hou et al. (2003) found approximately 40% of 129 I bound to organic matter in a soil sampled in Belarus (Gomel) in September 1999.Englund et al. (2008) found 50-85% of 129 I bound to organic matter in a Swedish lake sediment, while the proportion of organically associated 129 I in our CEZ soils ranged from 24 to 94%.The Englund et al. (2008) study found 5-8% of 129 I in water-soluble, exchangeable and carbonate fractions, combined.In our study, the combined KCl and KH 2 PO 4 extractable fractions of 127 I and 129 I can be considered to represent 'labile' iodine.For 127 I, the magnitude of these combined fractions relative to the total iodine concentration ranged from 0.4% (in RF3 subsoil) to 20% (in RF1 topsoil).For 129 I, the range was from 6% (in RF1 and RF3 subsoils) to 76% (in RF2 topsoil).In the case of 127 I, the ratio was significantly and negatively correlated with LOI (Fig. 4 (ii)) whereas for 129 I there was no significant relationship with LOI another indication of differences in behaviour of the two nuclides in the same soils.

Consideration of the initial physico-chemical form of 129 I deposited in 1986
The initial physico-chemical form of radionuclides deposited from the atmosphere can have a major impact on their long-term behaviour in the environment, but we can only speculate about the exact physicochemical form of 129 I when it first contacted the CEZ soils.Radioactive deposits in the highly contaminated western and northern traces of the CEZ were mostly in the form of irradiated reactor fuel fragments, referred to as 'hot particles' (Sandalls et al., 1993).Even though samples in this study were taken almost 30 years after deposition there was clear autoradiographic evidence (Supplementary materials, Fig. S2) of the existence of hot particles in CEZ soils.When Vapirev et al. (1990) analysed a hot particle from Chernobyl they found it to be depleted of 131 I.At the very high temperatures prevalent during the accident in 1986 radioactive iodine would have been released to the atmosphere in the gas phase; indeed, in less severe accidents in which only volatile radionuclides were released from reactor cores (Windscale, Three Mile Island) radioisotopes of iodine constituted major proportions of the release inventories (Crick and Linsley, 1984;Toth et al., 1986).Volatile radioiodine released to the atmosphere from Chernobyl is thought to have been progressively sorbed to aerosol particles on a time scale of approximately 45-50 days (Maryon et al., 1991); this is too long for this process to have influenced radio-iodine deposition in the CEZ.Weather records from Chernobyl indicate that, apart from very slight (0.4 mm) rainfall on 27th April, there was no significant precipitation at the time of the accident and in the following days up to May 3rd when records stop (NOAA, 2018).In the vicinity of the reactor, gaseous 129 I would have been subject to dry deposition to vegetation surfaces, followed by weathering and transfer to the underlying soil over the following weeks.In forests close to Fukushima, Xu et al. (2016) have shown that, in comparison to rainwater, 127 I is concentrated in throughfall and stemflow and that it is entirely in organic form in these sub-canopy waters.Although we do not know the exact physico-chemical form of 129 I deposited on CEZ soils in 1986 it is highly probable that it was not 'bound' within the uranium oxide particles which are a unique characteristic of the CEZ.

Conclusions
Contamination of the CEZ can be precisely dated to the 10 day period from 26th April to 5th May 1986.Any delay in the introduction of 129 I to soils as deposits were weathered from vegetation canopies would have been over a period of weeks.The CEZ therefore provides the opportunity to quantify the long-term behaviour of radioactive iodine under real-world conditions following a well-defined pulse injection.The observation that full isotopic equilibrium between the native 127 I and 129 I has not been achieved after 29 years indicates that complete mixing between the nuclides is a very long-term process.This contrasts with the observations of most laboratory incubation experiments that suggest that interaction of iodine species with organic and inorganic soil components is complete within a period of days (Duborska et al., 2019) to months (Bowley et al., 2016).Our data show that 129 I remains more mobile and more bioavailable than 127 I three decades after the Chernobyl accident, which has implications for the use of native, stable iodine as a proxy for assessments of the radiation doses attributable to 129 I, certainly on a decadal time scale and perhaps longer.

Acknowledgments
This work was carried out within the TREE (Transfer-Exposure-Effects) consortium, one of three projects within the Radioactivity and

129 IFig. 1 .
Fig. 1.Sampling locations (○) within the Chernobyl Exclusion Zone (base map from De Cort et al., 1998).The scale is shown by the 30 km radius from the reactor, visible along the western edge of the map.The grey circle in the inset map shows the general location of Chernobyl.Codes for sampling sites are shown in Table1.

Fig. 2 .Fig. 3 .
Fig. 2. I-127 and 129 I concentrations (i, ii and iii) and 129 I/ 127 I ratios (iv) in soil extracts from CEZ. Points represent geometric means, boxes represent ±1 geometric standard deviation and whiskers represent maxima and minima.Points with different letters are significantly different based on Tukey pairwise comparisons following ANOVA of Ln-transformed data.

Fig. 5 .
Fig. 5. Relationship between TOTAL 127 I and 129 I (g g-1 DW) concentrations in soils from the CEZ.Both Pearson (r = 0.81) and Spearman Rank correlations were highly significant (p b 0.001 and p = 0.001, respectively).

Table 1
Sampling locations and characteristics.
a Air kerma rate determined 1 m above ground level at the time of sampling.
Fig. 6.Relationships between 129 I/ 127 I ratios in soil extracts and total 129 I concentrations in soils from CEZ.The dashed regression line is for 'labile' (KCl + KH 2 PO 4 ) extracts; the solid regression line is for NaOH extracts.Spearman Rank Correlations: p = 0.003 for labile (KCl + KH 2 PO 4 ), p b 0.001 for NaOH.