2005 International Nuclear Atlantic Conference - INAC 2005
Santos, SP, Brazil, August 28 to September 2, 2005
ASSOCIAÇÃO BRASILEIRA DE ENERGIA NUCLEAR - ABEN
ISBN: 85-99141-01-5
GEOCHEMISTRY AND ENVIRONMENTAL MOBILITY OF IODINE-129
Eduardo Figueira da Silva
Divisão de Rejeitos Radioativos (DIREJ / CNEN - RJ)
Comissão Nacional de Energia Nuclear
Rua General Severiano, 90
22260-001 Rio de Janeiro, RJ
[email protected]
ABSTRACT
Iodine-129 is the longest lived of the iodine isotopes and is considered to be one of the largest potential
contributors to the doses resulting from nuclear waste repositories. It is generated in significant quantities in
nuclear reactors and its importance is due to its very long half-life (1.6 × 107 years) and relative mobility under
environmental conditions. The present paper reviews the important aspects concerning the generation, emission
and global circulation of 129I in the geosphere. Following, geochemical aspects relevant to the mobility of iodine
in the environment are analyzed within the framework of the safety of nuclear waste repositories. Iodine species
usually found in nature are identified, in addition to species expected to be found in the near-field of
repositories. Previous studies concerning the precipitation and sorption behaviors of the major species of iodine
are critically reviewed. Finally, some aspects of the microbial activity influences on the mobility of 129I are
presented. Although a reasonable amount of work has been done on trying to understand the geochemistry of
129
I over long time frames, there is still a large amount of uncertainty in characterizing its behavior under
environmental conditions and in the near-field of engineered repositories. In order to be able to predict the longterm behavior of a nuclear waste repository with respect to 129I, it is first necessary to have a better
understanding of the geochemistry and mobility of this radionuclide, especially with respect to sorption and
microbial interactions.
1. INTRODUCTION
Iodine-129 is the longest-lived of the volatile radionuclides produced in the nuclear fuel cycle
[1, 2], with a half-life of approximately 1.6 × 107 years, formed as a product of fission of
uranium or plutonium and accumulates in reactor fuel in proportion to the fission in that fuel.
The decay of 129I requires a very long period (1.1 × 108 years) to reach 1% of the original
amount. Together with 14C, it persists for a period often considered to perpetuity for mankind,
and will eventually enter the environment unless geological processes can be exploited to
ensure its isolation [2].
129
I emits soft beta particles, a weak gamma ray and an X-ray, representing an ingestion and
inhalation hazard [2,3]. When people breathe of ingest iodine it first moves to the thyroid
gland because Thyroxin, a growth and metabolism regulating hormone produced in the
thyroid, is rich in iodine. Therefore, the thyroid is the organ at the higher risk from exposure
to radioiodine, which is of especial concern for fetus and children [4,3].
If released from engineered or geological barriers, 129I will enter the global circulation of
iodine, mixing with the much larger quantity of stable iodine and following it in its
movement through the environment. This circulation involves the accessible iodine
reservoirs, particularly the deep ocean and its sediments. The geochemical behavior of 129I
over the long periods of time until its decay to an innocuous isotope will determine its
circulation patterns, the environmental impact, and ultimately the dose to humans.
2. GENERAL PROPERTIES, OCCURRENCE AND RADIOLOGICAL IMPACT
Iodine is an essential element for many organisms and is often added to dietary salt, because
deficiency of I causes goitre. Natural sources are most important environmental sources for
the stable isotope 127I, while nuclear bomb tests and accidents have provided most of the
radioactive isotopes. Iodine is naturally associated with Li, Na, K, B, P, W, F, Br, Cl, SO4,
CO3 (brines, evaporites) [5] and its environmental mobility is usually very high.
Iodine-129, the longest lived of the iodine isotopes, decays by the route [2]:
IT (1.0 ns)
β- 16 × 106
129
129m
129
I
Xe
Xe (stable)
150 keV
39.6 keV
Only 7.52% of the 129Xe transitions are by gamma ray emmission. The remainder yield
conversion electrons (mainly 5 to 10 keV) and xenon X-rays [2].
Iodine has a complex chemistry. The element is volatile and reacts to form many volatile
organic compounds. It may undergo oxidation or reduction to yield very soluble salts. It is
intimately involved with the life cycle, and its concentration in the environment is frequently
the result of biochemical processes. In the human it is concentrated in the thyroid gland, and
the concerns of radioiodine largely involve its effect on this gland [2].
Several natural processes contribute to the 129I global inventory, such as cosmic rays
interactions with xenon isotopes, which makes an equilibrium contribution of 250 kg, and
spontaneous fission processes in natural 238U and 235U , resulting in an equilibrium inventory
contribution of 113 kg [2]. However, anthropogenic 129I has been produced by nuclear
programs in a significant amount. Nuclear explosions in the atmosphere have added about
170 kg of 129I, most of this amount deposited on the surface of the Earth and in the oceans.
But it is the 129I that comes from reprocessing of spent nuclear fuel that is of greatest concern.
It has been estimated [2] that the world nuclear program can potentially produce from
reprocessing of spent nuclear fuel about 8 × 107 kg of 129I over the next thousand years, if the
amount of uranium readily available in ores is utilized. Although the dilution factors of
129 127
I/ I for the atmosphere and oceans are similar (2.2 × 10-12 and 4.4 × 10-12, respectively
[2]), the amounts of stable iodine in these environmental compartments differ markedly.
There are 108 kg of stable iodine in the atmosphere and 9.5 × 1013 kg in the oceans, and
consequently the ocean iodine has more than 106 times the capacity for isotopic dilution than
that in the atmosphere [2]. More recently, Kabata-Pendias and Pendias [6] claimed that the
129 127
I/ I ratio has increased in recent times due to nuclear weapons tests, from values of the
order of 10-12 to about 10-8.
Iodine-129 is the anthropogenic nuclear fission product that may present the greatest potential
hazard in the long-term exposure following the ingestion of food contaminated from
accidental releases from facilities from the nuclear fuel cycle [7]. Nevertheless, when
compared to other natural volatile radionuclides, such as 14C or radon, anthropogenic 129I is
considered to contribute much less (i.e., about one millionth) to the potential collective dose
committed than the background due to those natural radionuclides [2], if 129I is assumed to be
uniformly dispersed globally. However, release of 129I might lead to important local
concentrations if provisions for adequate dispersion are not taken. Dispersion of 129I into the
environment should preferably be associated with a simultaneous effect of adequate dilution
by stable iodine 127I. Dispersion media which already contain some significant amount of
natural iodine, like the ocean, would therefore be advantageous in terms of dose [2]. Yet,
recovery and concentration for storage or disposal offers advantages in reducing doses for the
near term (about 104 years), and complies with international agreements prohibiting ocean
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disposal of radioactive wastes. Most of the work done on 129I management has dealt with its
recovery, concentration and immobilization in a suitable form for long term storage.
3. GENERATION, EMMISSION PATHWAYS AND CIRCULATION MODELS
The radioactive isotopes from iodine belong to a larger family of fission products (Z = 30-66)
of U and Pu. The majority of the raioisotopes of iodine is obtained by β--decay from Sn, Sb
and Te. Two important facts should to be noticed [8]. First, the production of 129I in terms of
activity is lower, when compared to other isotopes, generally with shorter life. The average
iodine concentration is of 106 Ci/t, while the 129I concentration is only of 0.02 Ci/t. Second,
the production of radioactive iodines is more important in fast reactors, where the fission
yield is more important, than in the thermal reactors.
The net production of 129I from 235U fission in a thermal reactor is about 1 µCi/MW.d,
depending on the operation conditions [1]. It is estimated that the world production of the
long-lived iodine, i.e. 129I, will reach 2 × 106 Ci by the year 2060. The amount of 129I present
in the environment by then will be around 2 × 104 Ci, i.e., 120 tones of radioactive iodine [8].
The emission of 129I from reactors is very small when compared to the shorter lived isotopes
[8]. On the other hand, emissions from reprocessing plants are much larger [8] and enter the
environment as releases to the atmosphere and hydrosphere [9]. Although smaller, the
emissions of 129I from reactors shall not be totally neglected, due to its long half-life.
Two sources of emission of 129I from nuclear reactors are possible: the effluents stack and the
liquid effluents. From reprocessing plants, the atmospheric emissions are originated from the
effluents dissolution and degassing units, and can be retained with appropriate absorbing
filters and storage. For the adequate long-term storage of 129I, it shall be maintained in sterile
medium and not in contact with air [8,15]. A fraction of iodine is also lost by volatilization in
the process of vitrification of wastes [10].
The location of fission and transmutation products in fuel is dependent on the chemical nature
of these products and on the reactor operating conditions, which can control the ability of the
fission products to migrate through the fuel. Gaseous products, such as I, collect as small
bubbles in the fuel matrix and on the grain boundaries; some of the gases escape from the
body of the fuel pellet and accumulate in the pellet-cladding gap [11]. Canadian HWR
CANDU reactors generally have higher fuel temperatures and experience larger fission gas
segregation and release [11]. The importance of defining well were the 129I is located in the
spent nuclear fuel rests on a better prediction of the release rates. A rapid release is expected
for the radionuclides from the gap region of the fuel (cracks and cladding-pellets interstice),
which occurs in laboratory experiments on a time scale of hours or days. A slower release
occurs from grain boundaries in the fuel, on a time scale of tens to several hundreds of years.
Stroes-Gascoune et al. [12] measured the release of 129I located on the gaps and grain
boundaries of the fuel, and concluded that it may be considered as an instantaneous release on
a geological time scale. Therefore, the waste fuel may not be regarded as an adequate
engineered barrier for the long term containment of iodine, confirming the importance of
better understanding the geochemical behavior of this element.
Several models have been developed to predict the concentration of 129I in the many parts of
the geosphere. Most of them are based on measured concentrations of naturally occurring
stable iodine within the different compartments of the geosphere and quantification of the
fluxes between these compartments, using these values to predict the 129I concentrations and,
ultimately, the dose to man. Deposition of 129I from the atmosphere occurs on vegetation and
soil. Some resuspension of 129I from the soil appears to occur [9], as well as releases from
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plants undergoing the process of rotting [13]. Kocher [14] presented a dynamic model of the
global iodine cycle developed for the purpose of estimating global radiological impacts from
releases of 129I to the environment. The global iodine cycle was described by means of a
linear, time-invariant environmental compartment model, based on measured concentrations
for naturally occurring stable iodine and on data on the global hydrologic cycle (Figure 1).
The model assumed that the behavior of the 129I released to the environment is the same as
that for naturally occurring stable iodine, and estimated individual dose rates and global
population doses based on the inventories in each compartment as a function of time. A
parameter sensitivity analysis indicated that only a few parameters are important for
determining global radiological impact from 129I. For times up to about 104 years after
release, the mean residence time in the surface soil region is the most important parameter
according to his calculations. For times beyond 106 years, the doses are sensitive only to the
mean residence time in the ocean sediments, because this determines the fraction of the
released 129I that becomes inaccessible to man. Still, for the first few thousand years after
release, the model developed is strictly applicable only for releases which occur relatively
uniformly throughout one or more of the global environmental compartments. Consequently,
more research should be focused on time periods of few tens of thousands of years after
release, since it is for this time period that the released 129I circulates primarily in the
environmental compartments to which man is exposed.
Figure 1. Global inventories and fluxes for naturally occurring stable I [14].
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4. GEOCHEMISTRY
The geochemistry of iodine, a biophile element, is strongly connected to its involvement in
biological processes. The high I content of sediments and soils is mostly due to uptake of I by
plankton or is due to fixation of I by organic matter [6]. Iodine content is reported to be
closely related with organic carbon content of sediments. Sediments of reducing
environments contain greater amounts of I than do oxidized sediments.
The influence of soil reaction on the I status in soil is diverse. Soil acidity favors I sorption by
soil components such as organic matter, hydrous oxides of Fe and Al, and illitic clays. On the
other hand, in alkali soils of arid and semiarid regions, I is known to be greatly accumulated
[6]. This is, however, due both to salinity processes and to the low degree of oxidized I
mobilization occurring under alkaline pH conditions.
The oxidation of iodide to iodate and further alteration to elemental I may occur in soils and
the exchange of volatile I compounds between soil and atmosphere is reported to be possible
[6]. Several ionic forms (I-, IO3-, I3-, IO-, IO63-, H4IO6-) may occur in the aquatic phase of soil
of which iodide, I- is most probably the dominant species in underground environments.
However, significant amounts of iodate (IO3-) are expected to form in the near-field of high
level nuclear waste repositories due to radiolytic oxidation of iodine 1. The geochemical
behavior is quite different between iodide and iodate, as analyzed later in this paper.
Due to the long half life of 129I, adequate engineered barriers must be selected which possess
long-term stability against chemical, mechanical, and thermal stresses. Iodine, however,
forms relatively few stable and insoluble inorganic compounds that would be suitable for
long-term storage or ultimate disposal. Of the possible compounds, the iodides (I-) and
iodates (IO3-) are the most likely candidates.
Among the iodides, the silver and monovalent mercury salts (mercurous iodide, Hg2I2) are,
by far, the most insoluble, followed by cuprous iodide, CuI. However, the cost of silver and
mercury makes such compounds economically unattractive. Lead iodide, although much
more soluble, might be an acceptable substitute. Among the iodates, the mercurous salt
(mercurous iodate, Hg2(IO3)2), is again the most insoluble. However, the addition of mercury
to a repository may be a problem [2]. Lead, silver, and barium are the next most attractive,
although all are reasonably soluble. The high volatility of iodine compounds above 400oC
also seriously limits the selection of candidate waste matrices, due to the high processing
temperatures of most solidification and encapsulation processes [15].
Iodine occurs as a minor constituent of various minerals, but does not form many separate
minerals. The known I minerals include iodides of some metals such as AgI, CuI,
Cu(OH)(IO3) and polyhalides, iodates, and periodates [6]. Usually soils derived from igneous
rocks contain 10 to 20 times more iodine than the unweathered virgin rocks from which they
originate [6,16] Similarly, soils overlying sedimentary rocks are richer in iodine than the
parent rocks which generate them. Sedimentary rocks contain 2 to 4 times more iodine than
igneous rocks. In gleyed water-rich soils, higher I concentration in lower soil horizons can be
expected [6].
Several factors contribute to the amount of iodine retained in soils. Among the enriching
agents are vegetation, colloids, acidity, fossils and atmospheric precipitation. The depleting
factors are leaching, alkalinity, cropping and catalytic action [16]. Since solubility of iodine
compounds is usually not a limiting factor in iodine transport in environmental systems, it is
necessary to investigate the possibility of relying upon sorption to solid surfaces in order to
immobilize iodine in the geochemical environment.
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Sorption of iodide (I-) by minerals that are commonly found in nature is very low under
conditions that are not acidic. However, high sorption ratios of I- by lead and copper
compounds have been found [1]. Two different I- sorption mechanisms on soil have been
identified: one is the competitive anion adsorption on positive charges; another is enzyme
catalyzed iodination of humic substances. Microbes seem to play an indirect role in iodide
adsorption, by producing enzymes to catalyze the iodination of humic acids. On the other
hand, iodate (IO3-) may be easily adsorbed by geological materials. Fabryka-Martin et al. [17]
investigated the natural analogues in the Kongarra uranium deposit in Australia, observing
considerable enrichment relative to U under slightly oxidizing conditions, which was
attributed to sorption of 129I to Fe oxide or clay surfaces. The estimated high sorption is in
contrast with the very low sorption of I- by geological materials. Thus, the enhanced sorption
of iodine may be due to radiolytic oxidation of I- and formation of IO3-, which is sorbed more
strongly by geological materials. Hence, IO3- is expected to be an important species of iodine
in the near field of a high-level nuclear waste repository [1].
Studies have indicated that adsorption rate of radioiodine by soils is influenced by several
soils factors, (e.g. moisture, pH) and by the chemical forms of I, being higher for I- than for
IO3-. Although I is scarcely solubilized under most soil conditions due to sorption to fixed
forms of humic matter [6], and perhaps to the oxidized state of IO3-, in extreme (reducing)
conditions, such as submerged soils, I can be very mobilized by a factor of ten thousand, as
compared to regular soil conditions [6]. Although reducing conditions are usually not
expected to be found in the near field of a geological disposal site, this kind of environment is
commonly present in aquifers and under environments microbially active. The association
between I and organic matter, hydrous oxides of Fe and Al, and clay of the chlorite-illite
group has been noted [6]. However, since organic matter is most responsible for I sorption in
soil, I is accumulated mostly in topsoil horizons. Soil microorganisms are believed to play a
significant role in the I cycle for their great capability to accumulate this element. It is
calculated that the microorganism biomass contains from 0.012 to 3.24% of the I present in
surface soil layers, and some fungi are known to accumulate even much higher amounts [6].
Koch and Kay [28] studied the transportability of the long-lived and potentially hazardous
nuclide 129I in organic soils. The proportion of I losses from the solution phase differ widely
with the kind of organic matter used in the experiment. The authors concluded that the longterm processes that operate in the field to immobilize I are highly variable between soils.
Iodine in groundwater will be present mainly as iodide ions, I-. Since most minerals have a
negative surface charge in groundwater environment, it will not strongly sorb to minerals.
However, a small but significant sorption of iodide is found on certain minerals [19].
Minerals containing cations capable of forming iodides of low solubility (Hg, Ag, Bi, Cu,
Pb,…) may be expected to sorb iodide by chemisorption. Cinnabar (HgS), chalcopyrite
(CuFeS2), silver chloride, and lead hydroxide have Kd up to 1 m3/kg. Precipitates with high
surface-to-volume ratio such as ferric hydroxide and aluminum hydroxide also exhibit a
significant sorption, as well as some silicate minerals with high anion exchange capacity
(olivine and serpentine) [19].
Iodide (I-) and iodate (IO3) are known to react with certain minerals, soils, and sediments.
Iodide is sorbed by soil organic matter under basic conditions and by iron oxide under acidic
conditions. Iodate is sorbed quickly by Fe(OH)3 added to soil and is also slowly reduced and
sorbed by soil. Aerobic lake sediments and and Fe(OH)3 freshly formed from dissolved
ferrous iron remove I- from solution, but anaerobic sediments do not. Uptake of I- by
planktonic seston was demonstrated under oxidizing conditions in continental shelf
sediments. I- and IO3 are also sorbed by several sediments and sedimentary minerals in
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seawater, river water, and distilled water, being strongly adsorbed by hematite in seawater
and somewhat weakly adsorbed in distilled water and river water [20].
Additional information on sorption of iodine species by minerals, especially iron oxides, as a
function of pH, water composition, and iodine concentration would be helpful for
understanding the environmental behavior of iodine. Most work has been on sorption of I-,
although IO3- occurs in surface waters and seawater and is stable with respect to I- under
oxidizing conditions [20].
Coutoure and Seitz [20] investigated the sorption of iodide, iodate and periodate by hematite
and kaolinite, as well as the sorption of iodate by pelagic red clay. Iodate ions (IO3-) are
strongly adsorbed by hematite at pH values up to 9. The reaction is rapid and reversible and
apparently occurs by replacement of OH- ions on the surface. Periodate ions (IO4-) are more
strongly sorbed by Fe2O3 than IO3- ions. Iodide (I-) ions appear to be less strongly sorbed. At
pH 4, hematite-free, goethite-free kaolinite sorbs IO3- only sightly and I- not at all. IO3- is
stable relative to I- in the presence of oxygen, and occurs in surface waters and in seawater.
Small amounts of finely divided iron oxides, which are present in many types of sediments,
may play a major role in the geochemistry of iodine by absorbing IO3-, such as pelagic red
clay. This result suggested that pelagic red clay may be an appropriate host for disposal of
radioactive 129I.
Experiments on sorption of radioactive 131I- by sedimentary minerals reported in the literature
[20] may provide dubious results if the oxidation state and the mineralogical purity are not
adequately established. Some of the iodine from these experiments is likely to have been in
the IO3- or IO4- forms, which would have been sorbed by iron oxide impurities. In acid
solution I- ions can be rapidly oxidized by air; moreover, even strongly alkaline solutions of
131 I have been found to contain iodate or periodate, apparently as a result of oxidation of
iodide by radiolysis products of water.
Sorption of IO3- and IO4- by iron oxides is potentially useful for disposal of radioactive 129I
produced by nuclear reactors. Iodine may be disposed in the IO3- form, I- may be oxidized to
iodate by oxygen, or IO3- and IO4- may be formed as results of radiolysis. The use of iron
oxide as backfill material has been suggested [20]. Red sediments, such as red sandstones and
red shales may be appropriate hosts for a repository on land.
Sazarashi et al. [21] reviewed the adsorptivities of various minerals (cinnabar, alophane,
attapulgite, chalcopyrite, and montmorillonite) to I- ions, in order to identify materials
capable of adsorbing iodine ions for long periods. As a result, the cinnabar was found to have
fairly high adsorptivity to I- ions, which was considered to be due to the reaction between
iodide ions and thermally unstable components in cinnabar. Nevertheless, it is supposed to be
difficult to use the natural minerals as the adsorbents of I- ions in the geologic disposal,
because of their low adsorptivities to iodine ion and toxicity. Hence, they prepared modified
minerals for adsorbing I- ions, by intercalating Ag(I)-thiourea(th) complexes within interlayer
regions of the montmorillonite. They found that the modified minerals had a point of zero
charge at a pH approximately equal to 4. This means that the surface of the modified minerals
were negatively charged for pH above this value, and I- selective adsorption would not be
expected. Yet, they could measure considerable I- selective sorption of the modified mineral
in distilled water containing Cl- and I-. Since the I- selective adsorptivity could not be
explained by the traditional ion exchange mechanism for pH values above 4, they suggested
that the I- selective adsorptivity is attributed to the formation of [Ag(tu)]I complexes, that
precipitate within the interlayer region of montmorillonite. Notwithstanding, a previous study
[22] had concluded that the use of Ag was unlikely to significantly improve the
immobilization properties of the near field for radioiodide. This difference is perhaps due to
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the fact that Sazarashi et al. [21] used [Ag(tu)]I complexes, while Atkins et al. [22] used AgI,
which was more unstable under their experimental conditions. Nevertheless, the results of
Sazarashi et al. [21] are very interesting in terms of indicating a potential means for
immobilizing 129I- by adsorption to modified minerals in the near-field of a nuclear waste
repository.
Iodine-129 in groundwater discharging from a geologic disposal vault could accumulate in
wetlands by chemical sorption onto low pH, highly organic solid surfaces or by direct or
indirect microbial processes. Although the iodine is present in the anion form as either I- or
IO3-, these anionic forms may be quickly complexed organically or inorganically [23].
Anionic competition with Cl- was shown to affect soil sorption of iodine. Higgo et al. [24]
reported that microorganisms play a role in the sorption of both I- and IO3-. The study of
Behrens [25] points to the involvement of micro-organisms in the loss of I- from freshwater
aquatic systems, suggesting that the reactions were extracellular, possibly enzymatic
oxidation of I- to I2, which then reacts with organics, probably proteins. Many authors have
inferred microbial effects on iodine sorption through soil treatments, however, few have
quantified these effects.
The results from Sheppard and Hawkins [23] suggest that microbes may only play a minor
and indirect role in iodine sorption through the decomposition of organic matter. More
significantly, they have shown that microbial populations can be drastically reduced in the
presence of iodine of about 100 mg/l of iodine in groundwater. Later, the same authors [26]
showed that iodine sorbs to soils rich in organics and hydrous oxides, and suggested that the
oxidation of I- to I2 and complexation to organic functional groups or oxides are the major
processes for I retention in Shield soils.
Direct microbial reduction of iodate was demonstrated with anaerobic cell suspensions of the
sulfate reducing bacterium Desulfovibrio desulfuricans and of the anaerobic dissimilatory
Fe(III)-reducing bacterium Shewanella putrefaciens [27]. Also, both ferrous iron and sulfide
as well as FeS were shown to abiologically reduce iodate to iodide. These results indicate that
ferric iron or sulfate reducing bacteria are capable of mediating both direct, enzymatic as well
as abiotic reduction of iodate in natural anaerobic environments. These microbially mediated
reactions may be important factor in the transport of 129I in natural systems.
5. CONCLUSIONS
The migration of 129I in the subsurface appears to be an important factor in determining the
dose to man in the first 10,000 years after the release. The geochemistry of this element
seems to be an important aspect in the quantification of its dispersion in the subsurface, and
ultimately in the calculation of the dose to the human being. Unlike other radionuclides,
iodine is very mobile, especially if in the most abundant reduced species, iodide (I-).
However, there is evidence that iodide may undergo oxidation due to radiolysis in the nearfield of a nuclear waste repository. The relatively high sorption attributed to I- in alkaline
conditions, as found by some authors, may have been caused by oxidation to iodate (IO3-),
which is more commonly sorbed to mineral surfaces.
There still seems to be a large controversy of how microbial activity influences the mobility
of iodine in the subsurface. Some authors claim that extracellular enzymatic reactions are
responsible for the oxidation of iodide and subsequent immobilization by complexation to
soil organic matter. Others have shown the reduction of iodate to iodide by sulfate and iron
reducing bacteria, thus enhancing the mobility of 129I.
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Immobilization by precipitation is not likely to be considered for a repository, since the
simple insoluble compounds are either very expensive or toxic. Sorption or complexation to
organic matter seem to be more promising in terms of long term immobilization.
Although a reasonable amount of work has been done on trying to understand the
geochemistry of 129I over long time frames, there is still a large amount of uncertainty in
characterizing its behavior under environmental conditions and in the near-field of engineered
repositories. In order to be able to predict the long-term behavior of a nuclear waste
repository with respect to 129I, it is first necessary to have a better understanding of the
geochemistry and mobility of this radionuclide. For this reason, the intricate aspects related to
the speciation and interactions of 129I with natural soils shall be better studied, in addition to
the influence of microbial activity on these phenomena.
ACKNOWLEDGMENTS
The author would like to acknowledge the financial support by CNPq – Brazil.
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