Registration Dossier

Environmental fate & pathways

Endpoint summary

Administrative data

Description of key information

Additional information


Copper iodide is an inorganic compound with relatively low water solubility at environmentally relevant concentrations. Upon dissociating in aqueous media, copper iodide gives rise to cuprous copper ions (Cu+) and iodide ions (I-) that may then be subject to further transformation. The resulting ionic species will be more bioavailable than the parent compound and will therefore be independently responsible for any effects seen in the event that copper iodide enters the environment as a result of its production and/or use.

In view of this fact, please refer to the separate Chemical Safety Reports (CSR) previously produced in support of the REACH registrations of copper and iodine for detailed assessments of the following properties:

  • Degradation (Section 4.1).
  • Environmental distribution (Section 4.2).
  • Bioaccumulation (Section 4.3).
  • Secondary poisoning (Section 4.4).

These documents are attached in their entirety to Section 13 (Assessment Reports) of the copper iodide IUCLID dossier.

Summary information on the Environmental Fate of copper and iodine is also presented below.



General summary of the information on environmental fate and pathways

Copper is a natural element and transition metal with more than one oxidation state. Copper in its metallic form (Cu°) is not available. Copper needs to be transformed to its ionic forms to become available for uptake by living organisms

Stability and Biodegradation

The classic standard testing protocols on hydrolysis, photo-transformation, are not applicable to copper and copper compounds.

This was recognized in the Guidance to Regulation (EC) No 1272/2008 Classification, Labelling and Packaging of substances and mixtures (metal annex): ‘Environmental transformation of one species of a metal to another species of the same does not constitute degradation as applied to organic compounds and may increase or decrease the availability and bioavailability of the toxic species. However as a result of naturally occurring geochemical processes metal ions can partition from the water column. Data on water column residence time, the processes involved at the water – sediment interface (i.e. deposition and re-mobilisation) are fairly extensive, but have not been integrated into a meaningful database. Nevertheless, using the principles and assumptions discussed above in Section IV.1, it may be possible to incorporate this approach into classification.’ For a discussion on this please see Section 7.6.

Relevant fate aspects for copper in the environment have been included in the section ‘additionalinformation on fate and pathways’ and are summarized below.

As outlined in theguidance (2009 and 2012), the understanding of the transformation of copper into more or less bioavailable species is relevant to the environmental hazard assessments and this is described below.

- Transformation of Cu-ions released in the environment - Copper speciation

Once released to the environment, copper ions have more than one oxidation state and copper is thus characterized as transition metal. The principal ionic forms are cuprous (Cu(I), Cu+) and cupric (Cu(II), Cu2+). The trivalent form (Cu(III), Cu3+) occurs but is relatively unimportant in physical and biological systems. Cu+is unstable in aqueous media and soluble Cu1+compounds readily transforms into soluble Cu2+ions, compounds and/or insoluble Cu2+ions, compounds (e.g. copper sulphides) that precipitate. This transformation of Cu+to Cu2+is a result of a redox reaction initiated through atmospheric water vapour as well as in aqueous solution. However, monovalent copper cations are only susceptible to such transformation when they are not chemically bound in insoluble compounds or stabilised in complexed forms.

The transformation of Cu(I) to Cu (II) can be described by:

 (1) 2 Cu2O + 2H2O = 4Cu++ 4OH-


 (2) 4Cu++ O2+ 4H+= 4Cu2++ 2H2O

 Both sub-reactions are summarised as:

2Cu2O(s) + O2(g) + 4H+= 4Cu2++ 4OH-

Among the copper species released/transformed, Cu (II) is thus the most environmental relevant species. It is further recognised that Cu (II) ions - commonly named free cupric ions- are the most active copper species and that total Cu or Cu(II) concentrations are usually not directly related to ecological effects since exposure of biota may be limited by processes that render Cu unavailable for uptake (ICPS, 1998). Assessing the species of Cu (II) therefore has ecotoxicological relevance. After being released into the environment, the Cu(II) ions typically bind to inorganic and organic ligands contained within water, soil, and sediments. In water Cu(II) binds to dissolved organic matter (e. g. humic or fulvic acids). The Cu(II) ion forms stable complexes with -NH2, -SH, and, to a lesser extent, -OH groups in these organic acids. Cu(II) will also bind with varying affinities to inorganic and organic components in sediments and soils. For example, Cu(II) binds strongly to hydrous manganese and iron oxides in clay and to humic acids, but much less strongly to aluminosilicates in sand. In all environmental compartments (water, sediment, soil), the binding affinities of Cu(II) with inorganic and organic matter is dependent on pH, the oxidation-reduction potential in the local environment, and the presence of competing metal ions and inorganic anions.

Some key papers on copper speciation in freshwater, marine waters, sediments and soils are provided in the section ‘additional information on environmental fate

- Copper attenuation, removal from water column, geochemical cycling- Quantitative assessment

As described above, after the release of Cu(II) in the environment, further transformations occur thereby changing the potential for toxicity, induced by the free cupric ions. The concentrations of ‘active’ cupric ions, that remains available for uptake by biota depends on different processes: precipitation, dissolution, adsorption, desorption, complexation and competition for biological adsorption sites (ligands). These processes are critical for the fate of copper in the environment. This was recognized in the Guidance to Regulation (EC) No 1272/2008 Classification, Labelling and Packaging of substances and mixtures (metal annex):

‘Environmental transformation of one species of a metal to another species of the same does not constitute degradation as applied to organic compounds and may increase or decrease the availability and bioavailability of the toxic species. However as a result of naturally occurring geochemical processes metal ions can partition from the water column. Data on water column residence time, the processes involved at the water – sediment interface (i.e. deposition and re-mobilisation) are fairly extensive, but have not been integrated into a meaningful database. Nevertheless, using the principles and assumptions discussed above in Section IV.1, it may be possible to incorporate this approach into classification. ‘

The use of laboratory mesocosm and/or field tests for evaluating removal of soluble metal species through precipitation/partitioning processes over a range of environmentally relevant conditions are described in theguidance (2009) and for copper, such laboratory/mesocosm and/or field tests have therefore been assessed.

-In the water compartment,removal of soluble copper species through precipitation/partitioning processes over a range of environmentally relevant conditions, was assessed in Raderet al.,2013 and described in the section ‘additional information on environmnetal fate and pathways’.

The assessment relies on modeling simulations, based on the Tableau Input Coupled Kinetics Equilibrium Transport (TICKET) model(Farleyet al.,2008). The numerical engine of the model is a screening level model used to assess the fate and effects of chemicals through simultaneous consideration of chemical partitioning, transport, reactivity, and bioavailability (MacKay TICKET-UWM). The software includes metal-specific binding to inorganic ligands,and POC (using information from metal speciation models such as WHAM) and average-annual cycling of organic matter and sulfide production in the lake. 

The model was applied to a standard lake environment (EUSES characteristics), complemented with a sensitivity analysis on model parameters such as pH. The validity of the model outcome (removal rate) was assessed from mesocosm and field data The main conclusions are formulated as follows:

·   For a standard lake environment consisting of the EUSES model lake parameters and the Kd derived in the copper RA (Log Kd: 4.48), copper removal from the water column satisfies the criterion of rapid removal of 70% dissolved copper removal in 28 days;·   For a standard lake environment consisting of the EUSES model lake parameters but with pH varying between 6 and 8 (Kd estimated form the model), copper removal from the water column satisfies the criterion of rapid removal of 70% dissolved copper removal in 28 days;

·   For an experimental freshwater mesocosm study, carried out with a range of copper loadings (Schaeferset al.,2003), the measured data demonstrate a half life of 4 days and thus satisfy the criterion of rapid removal of copper (i.e. greater than 70% in 28 days);

·   For the whole-lake spike addition studies (LakeCourtilleand Saint Germain les Belles Reservoir), TICKET-UWM results, in concert with the measured data, indicate rapid removal of copper (i.e. greater than 70% in 28 days) for both lake systems;

·   Hypothetical TICKET-UWM simulations modeling the removal of copper in the MELIMEX limno-corrals following termination of copper loading demonstrate copper removal that does not meet the rapid removal benchmark because of a low settling velocity, low distribution coefficient, and low suspended solids concentration. 

Considering that the MILIMEX system is the only scenario that could not demonstrate ‘rapid removal’ it is critical to assess the environmental relevance of the MILIMEX system. The MILIMEX System was characterised by a setting velocity that is 10 times lower then the one in the EUSES system (0.2 versus 2.5 m/d) and a suspended solid concentration that is almost 3 times lower then the EUSES system (5.9 vesrus 16 mg/L). It is therefore concluded that the MILIMEX study was carried out under extreme conditions.

From Raderet al.,2013, it can therefore be concluded that under ‘environmental relevant’conditions, copper-ions are rapidly removed from the water-column.

This information is relevant to the environmental classification.

-In the sediment compartment,copper binds to the sediment organic carbon (particulate and dissolved) and to the anareobic sulphides, resulting in the formation of CuS. CuS has a very low stability constants/solubility limit (LogK=-41 (Di Toroet al.,1990) – see sectionadsorption/desorption) and therefore the ‘insoluble’ CuS keeps copper in the anaerobic sediment layers, limiting the potential for remobilization of Cu-ions into the water column.

To examine the potential for remobilization of copper from sediments, a series of 1-year simulations were performed, using the TICKET-UWM. These focused on re-suspension, diffusion, and burial to/from the sediment layer, their net effect on copper concentrations in the water column and changes in speciation in the sediment. Simulations were made with an oxic sediment layer as well as with an anoxic sediment layer (with varying concentrations of Acid Volatile Sulfides (AVS)) and varying re-suspension rates (up to 10 times the default EUSES model lake value).

In simulated sediments with AVS present in excess of copper, essentially all copper in sediment was present as copper sulfide because the affinity of copper for sulfides is much larger than the affinity for Organic Carbon.  CuS has a very low solubility product constant (Kps) and therefore, full copper sulfide precipitation was generally demonstrated : in all cases where AVS >1 µmol (reasonable worst case AVS concentration in European surface waters) and at environmentally relevant copper concentrations (< 0.1mg/L). As a result of this strong binding, the sediment log Kd greatly exceeded the water column log Kd and the net diffusive flux of copper was directed into the sediment. For anoxic sediments devoid of AVS and for oxic sediments, the net diffusive flux was small and directed out of the sediment. However, for all cases considered, the pseudo steady-state total and dissolved copper concentrations were at least 8 times lower than the concentration corresponding to conditions of 70% removal from the water column (see conditions detailed above).

Simpson et al (1998) and Sundelin and Erikson (2001) (see sectionadsorption/desorption)provide field evidence on the stability of the CuS binding :

·       Simpson et al (1998) investigated the oxidation rates of model metal sulfide phases to provide mechanistic information for interpreting the observations on natural sediments. CuS phases were kinetically stable over periods of several hours.

·       Sundelin and Erikson (2001) provide further evidence that, after long term oxygenation of sediment cores (3 to 7 months) Cu remains comparatively unavailable.

Last but not least, the assessment of 2 field experiments with intermittent copper dosing (LakeCourtille and the Saint Germain les Belles Reservoir lakes, yearly dosed with copper), assessed in Raderet al.,2013, provides further support for the absence of re-mobilization. Since both waterbodies are shallow, polymictic lakes, wind-driven resuspension is expected to play a role in copper dynamics in the water column. Neverteless, even if long-term resuspension does in fact occur, for both waterbodies, > 70% removal in less then 28 days was observed. The information therefore validates the results from the model simulations and absence of remobilization from the water column (Rader et al., 2013).

-In soils,decreases in copper solubility and in copper bio-availability are observed following copper spiking in the laboratory and from long-term field copper exposure experiments. Short term attenuation and long term ageing of copper, spiked in soluble forms to soils was demonstrated from laboratory and field experiments (Maet al.,2006a and 2006b) and reported in the section ‘adsorption/desorption’.

The soil environmental factors governing short term attenuation and ageing rates are soil pH, organic matter content, incubation time and temperature with soil pH being the key factor for ageing of Cu added to soils. From a range of laboratory and field experiments an ageing factor of 2 was derived as a reasonable worst case when considering field exposure data. This information is relevant to the soil PNEC derivation.

Transport and distribution

Relevant partitioning coefficients are available from literature.

-Aquatic compartment

Partition coefficient in freshwater suspended matter    Kpsusp= 30,246 l/kg (log Kp (pm/w) = 4.48) (50thpercentile)

Partition coefficient in freshwater sediment           Kpsed = 24,409 l/kg (log Kp(sed/w) = 4.39) (50th percentile)

Partition coefficient in estuarine suspended matter      Kpsusp= 56,234 l/kg (log Kp (pm/w) = 4.75) (50thpercentile)

Partition coefficient in marine suspended matter       Kpsusp= 131,826 l/kg (log Kp (pm/w) = 5.12) (50thpercentile)

-Terrestrial compartment

Partitioning coefficient                         Kd value soil: 2120 L/kg(log Kp (pm/w) = 3.33) (50thpercentile)


Because copper is an essential nutrient, all living organisms have well developed mechanisms for regulating copper intake, copper elimination and internal copper binding. The information in the accumulation section demonstrates that copper is well regulated in all living organisms and that highest/ BAF values are noted when copper concentrations in water, sediments and soils are low and for organisms/ life stages with high nutritional needs. The/ BAF values therefore have no ecotoxicological meaning. It should be mentioned that the non-applicability of BCFs for metal and especially for essential metals was already recognized in the regulatory framework of aquatic hazard classification (OECD,2001).

Importantly, the literature review demonstrates that copper is not biomagnified in aquatic or terrestrial ecosystems.

The section further includes critical data related to (1) the accumulation of copper on critical target tissues (e.g. gills in aquatic organisms); (2) the influence of environmental parameters (e.g. Organic Carbon, pH, Cationic Exchange Capacity) as well as food intake on the accumulation of copper. This information is relevant to the understanding of the accumulation as well as the mechanism of actions, described in the sectionecotoxicological information

More detailed summaries on respectively aquatic and terrestrial bioaccumulation are available from the aquatic and terrestrial bioaccumulation summary sections

Information relevant to assessing copper toxicity from dietary exposure - of relevance to secondary poisoning assessments - is included in the section ‘ecotoxicological information’.

The summary record “ecotoxicological information” further provides an overall summary of the rational for the absence of bio-accumulation and no-concern for secondary poisoning.


Iodine is a naturally occurring inorganic element which is found in all compartments of the environment. Muramatsu and Yoshida reported that the valence states of the major iodine species in the environment are: iodide, I-(-1), iodate, IO3-(+5), elemental iodine, I2(+0), and methyl iodide, CH3I (+1) (Muramatsu, 1999). Furthermore, several inorganic or organic iodine containing species are known depending on the compartment and prevalent conditions. Without consideration of the influence of organic substances and the biosphere on the speciation of iodine the dominant species in the hydrosphere, atmosphere and pedosphere can be estimated on basis of the Eh- and pH- values with molecular iodine and iodide playing the major role under common environmental conditions (Rucklidge, 1994).

Most of the elemental iodine (I2) is produced synthetically from iodate and/or iodide salts. Aerobic and anaerobic biodegradation can be considered as negligible, and in principle only abiotic degradation processes are relevant for the environmental fate and transportation processes. Thus, hydrolysis in the aquatic compartment and photolysis in the atmosphere are key transformations.

Elemental iodine in contact with water is rapidly disproportionated into iodide and hypoiodite (IO-) with the latter being further disproportionated to iodide and iodate. Although the second step is significantly slower a transformation of the major partition of molecular iodine can be assumed within the first hour of the degradation process (Truesdale, 1994). In addition to this inorganic transformation products with iodide being the major component (up to 90 % of total dissolved iodine, FOREGS database) iodine also rapidly forms organic species by reaction with dissolved organic matter, like humic acid etc.. Hypoiodous acid is considered as a key intermediate in the hydrolysis of iodine (Truesdale, 1994), and to be an important species for the reaction with organic matter in aqueous solutions due to observations in the marine environment (Truesdale, 1995).

In the ocean iodide and iodate are the dominant species with varying ratios depending on the depth and the geographic location. Although iodate is thermodynamically more stable under oxic conditions in a slightly basic solution like seawater, there seems to be a kinetic barrier which prevents a direct oxidation from iodide to iodate so that iodide persists in seawater (Truesdale, 1974, Sugawara, 1958). Aside from this kinetic barrier the speciation is affected by adsorption/scavenging onto metal oxides/hydroxides (Price, 1973; Price, 1977; Neal, 1976), and since iodine is a biophilic element also biologically mediated reduction and oxidation processes as well as incorporation into biogenic particles play a major role. Finally, exchange processes between bottom waters and sediments are very important for the geochemical fate of iodine (Wakefield, 1985; Bojanowski, 1970; Pedersen, 1980; Kennedy 1987a; Kennedy, 1987b; de Luca Rebello, 1990; Malcolm, 1984; Neal, 1976).

Castledine and Davis analyzed the iodine concentration in several seaweed species observing a high bioaccumulation which can be assumed to contribute to the iodine content of marine sediments (Castledine, 1973; Brehler, 1974). Further findings give indications that iodine is also incorporated by algae and transformed inter alia into iodinated volatile organic compounds (VOCs) (Burreson, 1976; Carpenter, 1999; McFiggans, 2002). The exact physiological function as well as the mechanism of uptake and metabolism is still unsolved. However, the release of this iodinated VOCs is one of the key processess in the geochemical lifecycle of iodine as the ocean is the major reservoir of available iodine and therefore being the major supplier of atmospheric iodine (Lovelock, 1973; Fuge, 1986; Ullman, 1990).

In addition, Amachi et al. (Amachi, 2001) found indications that a wide variety of terrestrial and marine bacteria are also capable of methylating iodine at environmental level of iodide (0.1 µM).

Released iodinated VOCs or molecular iodine are rapidly photolysed in the marine atmosphere. While the lifetime of molecular iodine is less than 10 seconds for overhead sun conditions (Saiz-Lopez, 2004; Jenkin, 1985), the lifetime of iodinated VOCs varies from 200 seconds to 90 hours (Roehl, 1997). Furthermore Duce et al. estimated that the mean lifetime of gaseous iodine before adsorption to a surface is about 30 minutes (Duce, 1965). In the marine boundary layer the transformation products of iodine and iodinated VOCs play a key role in the particle formation and the ozone depletion (McFiggans, 2000; Mäkelä, 2002; O'Dowd, 2002). Principally in these aerosols inorganic iodine oxides as well as soluble iodocarbons can be observed (Baker, 2005). Pechtl et al. found that iodate is increasingly depleted in acidic aerosols by inorganic reactions with decreasing pH value. Furthermore he assumed that HOI is formed in aqueous chemistry and then reacts with dissolved organic matter originated from the ocean. By combination of these organic and inorganic reaction cycles he was able to reproduce field observations, but constrained that each aerosol particle seems to be an "individual laboratory" with unique properties that may constantly change during its lifetime (Pechtl, 2007).

Gas-to-particle conversion is estimated to extend the atmospheric lifetime of iodine by about two weeks. Important parameters for the atmospheric aerosol lifetime are inter alia particle size, coagulation and scavenging processes, sedimentation, winds, etc.. In general, dry and wet deposition are the major loss processes of atmospheric iodine, and accordingly, these are main sources of iodine in soil. Another source are the decomposition processes of organisms. From a vast number of studies it is apparent that in general the iodine content varies between soil type and locality as it is influenced by a great number of factors (Ernst, 2003; FOREGS database; Muramatsu, 1999; Sheppard, 1992; Johnson, 1980), e.g. it can be observed that while in acidic soils and prevalent oxidizing redox potential iodide is being transformed in iodine which is readily volatile, in more basic soils iodine exists as non-volatile iodate. Based on the estimation of Rucklidge iodide is the major existent species of iodine in soil which is available for interactions with soil components.

Johnson analyzed about 200 soil samples for organic content and total iodine, and found a good correlation of high organic matter content and enrichment of total iodine for topsoil samples (Johnson, 1980). These findings are also supported by a great many of other studies concluding that soils rich in organic matter/humus are rich in iodine (Gallego, 1959a; Gallego, 1959b; Sinitskaya, 1969; Irinevich, 1970; Sazonov, 1970).

In addition, studies give evidence that soil iodine also binds to hydrous oxides of iron and aluminium. The extent of this process is depending on the pH of the soil and decreases at pH values above 7. Hence, it is assumed that hydroxyl ions are blocking the corresponding binding positions in soil and iodide could not be adsorbed (Whitehead, 1973; Whitehead, 1974). In general, findings indicate that iodine in soil will predominantly exist as soluble species and the content of organic matter as well as the presence of iron and aluminium hydroxides play the decisive role for the retention of iodine. In this context Muramatsu analyzed the influence of soil microorganisms on the behavior of iodine. It can be observed that microorganisms or their products play an important role in the accumulation of iodine as well as in the loss process from soil by evaporation. In saturated soils microorganisms support the development of reducing soil conditions. Within this process iodine is transformed into methyliodide which is readily volatile and barely soluble in water (Muramatsu, 1999; Amachi, 2001).

Ashworth et al. found evidence in a long-term adsorption/desorption experiment indicating that iodine within oxic environments is less mobile and, presumably, less bio-available than in anoxic environments (Ashworth, 2006).

These findings are supported by results of a one-year long-term laboratory migration-test in which an accumulation of iodine in the transition zone between anoxic and oxic soil conditions indicating a mobility of iodine only in the saturated/ low redox zone was observable. In contrast to the marine environment uptake by terrestrial plants can be considered as low, since only low uptake by ryegrass was found (Ashworth, 2003).

Newton and Shacklette also could not find a relationship between iodine concentration in plants and iodine content of the soils on which they grew (Newton, 1951; Shacklette, 1967).

Further loss mechanisms from soil are still quite unclear. Beside from evaporation it is assumed that already absorbed iodine is lost from soil by leaching of the binding material (organic matter, iron/aluminium hydroxides) into the groundwater or by decomposition of the organic matter which enables transformation and evaporation of iodine. Additionally, it is supposed that in case of soil saturation with iodine additional iodine will not be retained and will readily pass the soil compartment (Fuge, 1986).




Amachi S, Kamagata Y, Kanagawa T, Muramatsu Y (2001). Bacteria mediate methylation of iodine in marine and terrestrial environments, Appl. Environ. Microbiol., 2718 -2722.

Ashworth DJ, Shaw G, Butler AP, Ciciani L (2003). Soil transport and plant uptake of radio-iodine from near-surface groundwater, J Environ Radioactivity, 70, 99-114, 2003

Ashworth DJ, Shaw G (2006). Effects of moisture content and redox potential on in situ Kdvalues for radioiodine in soil, Sci Total Environ, 359, 244-254

Baker AR (2005). Marine Aerosol Iodine Chemistry: The Importance of Soluble Organic Iodine, Environ Chem, 2, 295-298.

Bojanowski R, Paslawska S (1970) On the occurrence of iodine in bottom sediments and interstitial waters of the Southern Baltic Sea, ActaGeophys.Pol., 18, 277 -286.

Brehler B, Fuge R (1974). Iodine, in Handbook of Geochemistry, vol. II/4, K.H. Wedepohl, ed., Springer-Verlag, Berlin.

Carpenter LJ, Sturges WT et al. (1999). Short-lived alkyl iodides and bromides at Mace Head, Ireland: Links to biogenic sources, J. Geophys. Res., 104, 1679– 1689.

Castledine SA, Davies DRA, The analysis of Jersey seaweeds, J Assoc Publ Analysts, 11, 97-103, 1973

de Luca Rebello A, Herms FW, Wagener K, The cycling of iodine as iodate and iodide in a tropical estuarine system, Mar Chem, 29, 77-93, 1990

Duce RA, Winchester JW, van Nahl TW, Iodine, bromine and chlorine in the Hawaiian marine atmosphere, J. Geophys.Res., 70, 1775-1799, 1965

Ernst T, Szidat S et al. (2003). Migration of iodine-129 and iodine-127 in soils, Kerntechnik, 68, 155-167.

Fuge R, Johnson CC (1986). The geochemistry of iodine - a review, Environ Geochem Health, 8(2), 31-54.

Gallego R, Oliver S (1959a). Iodine in soils, Anales de edafologia y fisiologia vegetal (Madrid), 18, 207-238. (in Spanish)

Gallego R, Oliver S (1959b). Relations between the iodine content and composition of soils, Anales de edafologia y fisiologia vegetal (Madrid), 18, 275-288. (in Spanish)

Irinevich AD, Rabinovich IZ, Fil'kov VA (1970). Iodine in Moldavian soils, Pochvovedenie, 58-68. (in Russian)

Jenkin ME, Cox RA (1985). Kinetics study of the Reaction IO + NO2+ M ->IONO2+ M, IO + IO -> Products, and I + O3-> IO + O2, J Phys Chem, 89(1), 192-199.

Johnson CC (1980). The geochemistry of iodine and a preliminary investigation into its potential use as a pathfinder element in geochemical exploration, Ph.D. thesis, University College of Wales, Aberystwyth.

Kennedy HA, Elderfield H (1987a). Iodine diagenesis in pelagic deep-sea sediments, Geochim Cosmochim Acta, 51, 2489 -2504.

Kennedy HA, Elderfield H (1987b). Iodine diagenesis in non-pelagic deep-sea sediments, Geochim Cosmochim Acta, 51, 2505 -2514.

Lovelock JE, Maggs RJ, Wade RJ (1973). Halogenated Hydrocarbons in and over the Atlantic, Nature, 241, 194-196.

Malcolm SJ, Price NB (1984). The behavior of iodine and bromine in estuarine surface sediments, Mar Chem, 15, 263-271.

Mäkelä, JM, Hoffmann T et al. (2002). Biogenic iodine emissions and identification of end-products in coastal ultrafine particles during nucleation bursts, J Geophys Res, 107(D19), 8110,doi:10.1029/2001JD000580.

McFiggans G, Plane JMC et al. (2002). A modeling study of iodine chemistry in the marine boundary layer, J Geophys Res, 105(D11), 14371-14385.

Muramatsu, Y., Yoshida, S. (1999). Effects of microorganisms on the fate of iodine in the soil environment, Geomicrobiol J, 16:85-93.

Neal C, Truesdale VW (1976). The sorption of iodate and iodide by riverine sediments: its implications to dilution gauging and hydrochemistry of iodine, J Hydrol., 31, 281-291.

Newton HP, Toth SJ (1951). Iodine content of some soils and plants in New Jersey, Soil Science, 71, 175-179.

O'Dowd CD, Jimenez JL et al. (2002). Marine aerosol formation frombiogenic iodine emissions, Nature, 417, 632-636.

Pechtl S, Schmitz G, von Glasow R (2007). Modelling iodide-iodate speciation in atmospheric aerosol: contributions of inorganic and organic iodine chemistry, Atmos. Chem. Phys, 7, 1381-1393.

Pedersen TF, Price NB (1980). The geochemistry of iodine and bromine in sediments of the Panama basin, J Mar Res, 38, 397 -411.

Price NB, Calvert SE (1973). The geochemistry of iodine in oxidized and reduced recent marine sediments, Geochim Cosmochim Acta, 37, 2149 -2158.

Price NB, Calvert SE (1977). The contrasting geochemical behaviours of iodine and bromine in recent sediments from the Namibian shelf,Geochim. Cosmochim. Acta, 41, 12, 1769 -1775.

Roehl CM, Burkholder JB et al. (1997). Temperature dependence of UV absorption cross sections and atmospheric implications of several alkyl iodides, J. Geophys. Res., 102(D11), 12819-12829.

Rucklidge J, Kilius L, Fuge R (1994). 129-Iodine in moss down-wind from the Sellafield nuclear fuel reprocessing plant, Nuclear Instruments and Methods in Physics Research B 92, 417-420.

Saiz-Lopez A et al. (2004). Absolute absorption cross-section and photolysis rate of I2, Atmos. Chem. Phys. 4, 1443-1450.

Sazonov NN (1970). Contents and migration of iodine in central Yakutia soils, Referatwnyi Zhurnal Geologiia V, Abstract No. IV36. (in Russia)

Shacklette HT, Cuthbert ME (1967). Iodine content of plant groups influenced by variation in rock and soil types, United States Geological Society Special Paper No. 90, 30-46.

Sinitskaya GI (1969). Iodine content in the Zeya-Bureya Plain soils, Uchenye Zapiski, Dal'nevost Gos Universitet Khimii, 27, 72-88. (inRussian)

Sugawara K, Terada K (1958). Oxidized iodine in seawater, Nature, 182, 251 -252.

Truesdale VW (1974). The chemical reduction of molecular iodine in seawater, Deep-Sea Res, 21, 9, 761 -766.

Truesdale VW, Canosa-Mas C, Luther GW (1994). Kinetics ofDisproportionation of Hypoiodous Acid, J Chem Soc Faraday Trans, 90(24), 3639-3643.

Truesdale VW, Luther GW (1995). Molecular iodine reduction by natural and model organic substances in seawater, Aq. Geochem., 1, 89 -104.

Ullman WJ, Luther GW et al. (1990) Iodine chemistry in deep anoxic basins and overlying waters of the Mediterranean Sea, Mar Chem, 31, 153 -170.

Wakefield SJ, Elderfield H (1985). Interstitial water iodine enrichments in sediments from the Eastern Pacific, J Mar Res, 43, 951 -961.

Whitehead DC (1973). Studies on iodine in British soils, J Soil Sci, 24, 260 -270.

Whitehead DC (1974). The sorption of iodide by soil components, J Sci Food Agri, 25, 73-79.