IDENTIFICATION OF PBT AND vPvB SUBSTANCESecha.europa.eu/documents/10162/13628/tdm_pbt... ·...

TDM PBT/vPvB evaluation

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IDENTIFICATION OF PBT AND vPvB SUBSTANCES

Public Name(s): tert-Dodecanethiol

tert-Dodecyl mercaptan (TDM)

2,3,3,4,4,5-Hexamethylhexane-2-thiol

EC Number(s): 246-619-1

CAS number(s): 25103-58-6

SMILES: CC(C)(C)CC(C)(S)CC(C)(C)C - example

Chemical formula (2D- structure): CH3C(CH3)(CH3)CH2C(CH3)(SH)CH2C(CH3)(CH3)CH3 - example

Date: 22nd January 2014

Substance prioritised for screening/assessment based on: previous Existing Substances PBT EWG

Screened by: UK

SUMMARY OF PBT EG DISCUSSIONS (Please tick where appropriate) (Remarks)

1. PBT/vPvB properties fulfilled

2. Equivalent level of concern to PBT/vPvB

3. Further testing/information needed on: Compliance check / Substance evaluation

The registrant dossier was updated to include all available data in September 2013 but this was not clear to the evaluating Member State until October 2013, after this PBT evaluation was drafted. It has been noted where new data are available in the updated registration dossier but as the focus of this evaluation is on the bioaccumulation test these new data have not been evaluated in-depth.

The registrant needs to amend their RSSs for the

OECD 310 and bioaccumulation studies.

P:

B:

T:

4. Not PBT/vPvB The substance is assessed to be not B and not vB by the eMS

5. Further discussion necessary


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RESULTS OF EVALUATION OF PBT / vPvB PROPERTIES

This dossier covers the substance manufactured and supplied as detailed below.

Substance name: tert-Dodecanethiol

EINECS number: 246-619-1

EINECS name: tert-Dodecanethiol

CAS number: 25103-58-6

Registration number(s): 01-2119486132-42-0000, 01-2119486132-42-0001 and 01-2119486132-42-0002

Molecular formula: C12H26S

Structural formula: Example

Composition: UVCB substance

Evaluation Summary

Tertiary-dodecanethiol (tert-dodecyl mercaptan, TDM) is a transitional ‘existing’ substance which was discussed by the former EU PBT Working Group on a number of occas ions . As a result of these discussions the substance was included in Regulation (EC) No. 465/2008 of 28th May 2008, which required industry to conduct an enhanced biodegradation test and fish bioconcentration study and submit the results by November 2009. The data were provided in January 2013.

Based on the available information, TDM does not meets the Annex XIII criteria for either a ‘persistent, bioaccumulative and toxic’ (PBT) or a ‘very persistent and very bioaccumulative’ (vPvB) substance in the environment.


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CONTENTS

IDENTIFICATION OF PBT AND VPVB SUBSTANCES 1

RESULTS OF EVALUATION OF PBT / VPVB PROPERTIES 2

CONTENTS 3

RESULTS OF EVALUATION OF PBT/VPVP PROPERTIES 6

1 IDENTITY OF THE SUBSTANCE AND PHYSICAL AND CHEMICAL PROPERTIES 6

1.1 Name and other identifiers of the substance 6

1.2 Composition of the substance 7

1.3 Identity and composition of degradation products/metabolites relevant for the PBT assessment

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1.4 Identity and composition of structurally related substances (grouping approach) 8

1.5 Physico-chemical properties 9

2 CLASSIFICATION AND LABELLING 13

2.1 Harmonised Classification in Annex VI of the CLP 13

2.2 Proposal for harmonised classification in Annex VI of the CLP 13

2.3 Self classification(s) proposed by the registrant(s) 13

3 ENVIRONMENTAL FATE PROPERTIES 16

3.1 Degradation 16

3.1.1 Abiotic degradation 16

3.1.1.1 Hydrolysis 16

3.1.1.2 Oxidation 16

3.1.1.3 Phototransformation/photolysis 17

3.1.1.3.1 Phototransformation in air 17

3.1.1.3.3 Phototransformation in soil 17

3.1.2 Biodegradation 18

3.1.2.1 Biodegradation in water 18

3.1.2.1.1 Estimated data 18

3.1.2.1.2 Screening tests 18

3.1.2.1.3 Simulation tests 25

3.1.2.2 Biodegradation in sediments 25


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3.1.2.3 Biodegradation in soil 25

3.1.2.4 Summary and discussion on biodegradation 26

3.1.3 Summary and discussion on degradation 26

3.2 Environmental distribution 26

3.2.1 Adsorption/desorption 26

3.2.2 Volatilisation 27

3.2.3 Distribution modelling 27

3.3 Potential for long range transport 28

4.1 Bioaccumulation 32

4.1.1 Screening data 32

4.1.2 Bioaccumulation in aquatic organisms (pelagic and sediment organisms) 35

4.1.3 Bioaccumulation in terrestrial organisms (soil dwelling organisms, vertebrates) 54

4.1.4 Summary and discussion of bioaccumulation 54

5 HUMAN HEALTH HAZARD ASSESSMENT RELEVANT FOR THE PBT

ASSESSMENT 55

6 ENVIRONMENTAL HAZARD ASSESSMENT 56

6.1 Aquatic compartment (including sediment) 56

6.1.1 Toxicity data 56

6.1.1.1 Fish 56

6.1.1.1.1 Short-term toxicity to fish 56

6.1.1.1.2 Long-term toxicity to fish 57

6.1.1.2 Aquatic invertebrates 58

6.1.1.2.1 Short-term toxicity to aquatic invertebrates 58

6.1.1.2.2 Long-term toxicity to aquatic invertebrates 59

6.1.1.3 Algae and aquatic plants 60

6.1.1.4 Sediment organisms 61

6.1.1.5 Other aquatic organisms 61

6.2 Terrestrial compartment 61

6.2.1 Toxicity data 61

6.3 Atmospheric compartment 62

6.4 Microbiological activity in sewage treatment systems 62

6.4.1 Toxicity to aquatic micro-organisms 62

6.5 Non compartment specific effects relevant for the food chain (secondary poisoning) 62

6.5.1 Toxicity to birds 62

6.5.2 Toxicity to mammals 62

7 CONCLUSIONS ON THE PBT/VPVB ASSESSMENT 63

7.1 Assessment of PBT/vPvB properties – comparison with the criteria of Annex XIII 63

7.2 Persistence 63


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7.3 Bioaccumulation 64

7.4 Toxicity 65

7.4.1 Fulfilment of the T criterion based on human health classification 65

7.4.2 Fulfilment of the T criterion based on ecotoxicity data 65

7.5 Summary and overall conclusions on the PBT, vPvB properties 66

8 REFERENCES 67

INFORMATION ON USES AND EXPOSURE 70

1 MANUFACTURE AND USE(S) 70

1.1 Quantities 70

1.2 Identified uses 70

2 POTENTIAL EXPOSURE AND RISK(S) 72

2.1 Workers 72

2.2 Consumers 72

2.3 Environment 72

2.4 Man exposed via the enviromnent 73


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RESULTS OF EVALUATION OF PBT/VPVP PROPERTIES

Note: A detailed review of existing information on the properties of TDM was published by EA (2005) and also summarised in a fact sheet (substance ref. 98) for the PBT Working Group of the Technical Committee for New & Existing Substances in the period up to 2008 (PBT Working Group, 2008). Test reports submitted to the evaluating Member State (eMS) in response to Commission Regulation (EC) No 465/2008 are the focus of this evaluation. Additional data reported in the REACH registrations, previous PBT fact sheet and EA (2005) are summarised where relevant, but in general fewer details are provided. The REACH registration dossier was last updated in September 2013. It has been noted where new data are available in the updated dossier but as the focus of this evaluation is on the bioaccumulation test these new data have not been evaluated in-depth.

1 IDENTITY OF THE SUBSTANCE AND PHYSICAL AND CHEMICAL

PROPERTIES

1.1 Name and other identifiers of the substance

Although a UVCB substance, the registration dossier uses identifiers related to the major constituent.


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Table 1: Substance identity

EC number: 246-619-1

EC name: tert-Dodecanethiol

SMILES: CC(C)(C)CC(C)(S)CC(C)(C)C

(although a specific isomer is named in the registration dossier, no SMILES code is provided. This code is taken from EA (2005) for a representative structure with a high degree of branching: 2,2,4,6,6-pentamethyl-4-heptanethiol)

CAS number (in the EC inventory): 25103-58-6

CAS number: 25103-58-6

CAS name: tert-Dodecanethiol

IUPAC name: 2,3,3,4,4,5-Hexamethylhexane-2-thiol

Index number in Annex VI of the CLP Regulation Not applicable

Molecular formula: C12H26S

Molecular weight range: 202.4 g/mole

Synonyms: tert-Dodecyl mercaptan (TDM)

Structural formula:

1.2 Composition of the substance

The registration dossier indicates that the substance is a UVCB substance but little other information is given in the original registration dossier on the composition (information on impurities is confidential, and no additives are mentioned).

EA (2005) indicates that commercially supplied TDM is typically >95% pure, although this may refer to a mixture of isomers and carbon chain length fractions, which is consistent with the substance being a UVCB (purity in this sense is not relevant). Other constituents that may be present typically include olefins and light mercaptans and sulphides (EA, 2005).


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EA (2005) reported that the majority of TDM is produced using propylene tetramer as the feedstock and that typically such a foodstock would contain the following chain length distribution (although other compositions are also possible):

Component % by weight (typical) % by weight (range)

≤C10 3.5 <10

C11 13.7 15-20

C12 50.5 50-80

C13 18.9 2-20

C14 10.3 <15

≥ C15 3.1 -

The chain length distribution in the commercial TDM currently registered has recently been provided (personal communication to the evaluating Member State) and is included in the Confidential Annex. EA (2005) indicates that the alkyl chain in TDM can also be produced by the trimerisation of isobutylene. Whichever route is used to produce TDM, the final product is a highly branched alkyl chain, consisting of a mixture of isomers. EA (2005) considered 2,2,4,6,6-pentamethyl-4-heptanethiol to be the best-supported representative structure for TDM.

EA (2005) concluded that although alkylthiol constituents with different chain lengths present in the commercial product will have different physico-chemical properties, these differences are expected to be small. In particular, they were not expected to significantly affect the interpretation of the available data, or its use within models, and a similar approach is taken here for this evaluation.

1.3 Identity and composition of degradation products/metabolites

relevant for the PBT assessment

There is no information available on possible degradation products/metabolites in the registration dossier.

EA (2005) considered it theoretically possible that TDM could oxidise in the environment to form disulfides or sulfonic acids. However, the evidence available (e.g. see Section 3.1 and EA, 2005) suggests that this is not an environmentally significant degradation process.

1.4 Identity and composition of structurally related substances

(grouping approach)

The grouping approach is not applied to TDM. The registration dossier uses some data for a related substance (dodecane-1-thiol, CAS no. 112-55-0) as a supporting study for the algal toxicity endpoint.


1.5 Physico-chemical properties

The physico-chemical properties of TDM are summarised in Table 2. These data are taken from the registration dossier.


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Table 2: Overview of physicochemical properties1

Property Value Remarks

Physical state at 20°C and 101.3 kPa

Liquid Key study from registration dossier

Melting/freezing point <-20°C Key study from registration dossier

Boiling point 238°C (511K) at 101 kPa Key study from registration dossier

Vapour pressure 20 Pa at 25°C Key study from registration dossier

Surface tension No data Data waiving applied to the endpoint in the registration dossier

Water solubility 0.21 mg/l at 25°C (WATERNT v1.01)

0.28 mg/l at 25°C (WSKOW v1.41)

0.0039 mg/l (experimental value)

QSAR estimates using EPIWEB v4.0; taken from registration dossier. The experimental value is from a recent study by Baltussen (2013) obtained by the OECD 105 guideline slow- stirring method (see text).

Partition coefficient n- octanol/water (log value)

log Kow

Partition coefficient n- octanol/air (log value)

Log KOA

>6.2 Key study from registration dossier obtained using OECD Guideline 117 (HPLC method). A QSAR value obtained using EPIWEB v4.0 of log Kow=6.07 was used as supporting evidence. A higher value of 7.43 has recently been estimated by Comber and Thomas (2013) from the water solubility (see text).

No data No value reported in the registration dossier.

Partition coefficient air/water (log value)

Log KAW

0.38 to 3.93 (range of estimates at 20°-25°C)

2.62 (based on experimental vapour pressure at 25°C and water solubility)

No value reported in the registration dossier. The values reported are taken from EU (2005) or calculated by the eMS from the water solubility and vapour pressure (see Table 3 and text for details of the calculations).

Dissociation constant pKa = 9.95 Supporting study from registration dossier. Value estimated using SPARC computer calculation model v4.

A recent paper by Comber and Thomas (2013) provided by the registrant suggests that the water solubility of TDM could be lower than given in the registration dossier. The

1 The references of the values reported in Table 10 will be available in the technical dossier. In case references need to be included an additional column could be added manually to Table 5.


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Comber and Thomas (2013) paper refers to a water solubility for TDM of 0.00393 mg/l obtained in a slow-stir water solubility study.

Details of the new water solubility test (Baltussen, 2013) have recently been provided in a robust study summary. The study was a GLP compliant OECD Guideline 105 study using the slow-stirring method. The substance tested had an analytical purity of 99.1%. The test was carried out by preparing triplicate samples in double distilled water at 19.9±0.4°C and stirring at 40 rpm. At various time points samples were taken, centrifuged and prepared for analysis, taking care to avoid volatilisation of the test substance (no further details of how this was achieved are given). The concentration of TDM was determined by a validated analytical method involving derivatisation followed by analysis using HPLC/MS/MS (this was presumably a similar method to that discussed in relation to the biodegradation and bioaccumulation data (see Sections 3.1 and 4.1)). The pH of the water was in the range 6.5 to 7.1 throughout the test. Samples were analysed at 24, 48, 72, 96, 120 and 144 hours. For the first three samples the concentration was found to increase slightly with time (0.00139 mg/l at 24 hours, 0.00174 mg/l at 48 hours and 0.00217 mg/l at 72 hours). For the latter three sampling times the concentration was found to be more stable, although the maximum difference in the concentration at the three sampling points was >15%. The concentrations measured were 0.00467 mg/l at 96 hours, 0.00415 mg/l at 120 hours and 0.00296 mg/l at 144 hours. The test report concluded that the variability in the results at these sampling points probably reflected the difficulties in accurately determining very low concentrations of TDM rather than a continuing increase in the amount of TDM dissolved (in fact the concentrations declined slightly with time during this phase). The water solubility was therefore determined to be 0.00393 mg/l based on the mean concentration measured between 96 hours and 144 hours.

The robust study summary gives the study a reliability of 2 (reliable with restrictions) as the maximum difference between the measured concentrations at the last three sampling points was >15%. The eMS agrees with this reliability rating and also considers it likely that the variability seen in the measurements reflects the difficulties in measuring low concentrations of this substance rather than a continuing increase in the amount dissolved at the later sampling points. Therefore the actual water solubility of TDM can be taken to be around 0.00393 mg/l (3.93 µg/l) at 20°C.

Comber and Thomas (2013) estimated a log Kow value for TDM of 7.43 using a validated QSAR based on this water solubility. A Robust Study Summary and details of the QSAR used have been made available to the eMS. The linear regression model was proprietary and was developed using confidential data sets (details of these were not given), but it was reported that the substance fell within the applicability domain of the QSAR. It should be noted, however, that the types of chemical used to train the model did not appear to specifically contain thiols (although it is not possible to be certain about this as the specific substances used were not given). The applicability of this method to thiols has since been demonstrated for a set of four thiols (primary and secondary), although the log Kow values of these were lower than for TDM (experimental log Kow values of the validation set were between 1.5 and 3.7) (personal communication to the evaluating Member State, 6th December 2013).

A further measured water solubility value for TDM is reported in EA (2005). The water solubility was determined to be 0.25 mg/l at 20°C and the study used a non- guideline protocol (simple flask method) but was carried out according to GLP. This value was used in the EA (2005) assessment but only limited details are available


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(the registrants do not have access to the study) for this study and so the reasons for the discrepancy between this value and the value of 0.00393 mg/l given above are currently unknown2.

The physico-chemical properties of TDM have also been reviewed by EA (2005) and the data presented there are generally consistent with those from the registration dossier but, apart from the water solubility (see discussion above), the main exception is the vapour pressure, which is given as 4 hPa (400 Pa) at 20oC in EA (2005) based on a non-GLP study conducted according to Method A4 of Directive 92/69/EEC. The test report was not available for review by EA (2005) and the registrants do not have access to the study, so the influence of volatile impurities in the test substance is not known.

The value for the vapour pressure reported in EA (2005) is twenty times higher than the value reported in the registration dossier and the reasons for this discrepancy have not been investigated in detail for this evaluation. However, it is relevant to note that EA (2005) estimated a Henry’s law constant for TDM of around 3.24×105 Pa m3/mole at 20oC based on the water solubility and vapour pressure (EA (2005) assumed a water solubility of 0.25 mg/l for TDM) and commented that this was higher than the Henry’s law constant estimated using the bond contribution method in EPIWIN of 5,900 Pa m3/mole at 25°C. When the vapour pressure (20 Pa at 25oC) and water solubility (0.21-0.28 mg/l at 25oC) given in the registration dossier are used to estimate the Henry’s law constant the value obtained is in the region of 14,490- 19,230 Pa m3/mole at 25oC which is in closer agreement with the EPIWIN estimate than obtained using a vapour pressure of 400 Pa.

When the more recent and lower water solubility value (0.00393 mg/l) is considered the Henry’s law constant can be estimated as around 1.03×106 Pa m3/mole at 25°C using a vapour pressure of 20 Pa (and assuming the change in water solubility with temperature is minor between 20 and 25°C) or 2.06×107 Pa m3/mole at 20°C using a vapour pressure of 400 Pa.

The various estimates of Henry’s law constant, along with the equivalent dimensionless Henry’s law constants (Kaw) are summarised in Table 3. Clearly there is a wide range of values that can be estimated for TDM. The values all suggest that volatilisation from water to air will be an important process in the environmental distribution of TDM. The significance of the range of estimates in relation to long- range transport potential is considered in Section 3.3. Based on the currently available data the best estimate of the log Kaw is probably 2.62 based on the vapour pressure of 20 Pa at 25°C given in the registration dossier and the recent water solubility determination of 0.00393 mg/l at 20°C.

2 It could be speculated that the differences may be related to the analytical method used and/or the presence of more soluble constituents present in the test substance. The study cited by EA (2005) appears to have used a GC method whereas Baltussen (2013) used a method involving derivitisation of the test substance followed by HPLC/MS/MS analysis. It is not possible to conclude whether the difference is due to this or some other reason.


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Table 3: Summary of estimates of the Henry’s law constant for TDM

Method Henry’s law constant (Pa m3/mole)

Kaw log Kaw Remarks

Estimated from a water solubility of 0.25 mg/l and a vapour pressure of 400 Pa






Estimated using the bond contribution method in EPIWIN (HENRYWIN v3.1)

3.24×105 at 20°C 133 2.12 Value from EA (2005)

1.62×104 at 25°C 6.5 0.82 eMS estimate

1.93×104 at 25°C 7.8 0.89 eMS estimate

1.45×104 at 25°C 5.8 0.77 eMS estimate

1.03×106 at 25°C 416 2.62 eMS estimate (using most recent vapour pressure and water solubility data)

2.06×107 at 20°C 8,457 3.93 eMS estimate 5,900 at 25°C 2.4 0.38 Value from

EA (2005)


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2 CLASSIFICATION AND LABELLING

2.1 Harmonised Classification in Annex VI of the CLP

There is no harmonised classification in Annex VI of the CLP Regulation.

2.2 Proposal for harmonised classification in Annex VI of the CLP

None.

2.3 Self classification(s) proposed by the registrant(s)

The following self classification is proposed by the registrant(s) in the September 2013 update to the registration dossier.

• Skin Irrit. 2 H315: Causes skin irritation.

• Eye Irrit. 2 H319: Causes serious eye irritation.

• Skin Sens. 1 H317: May cause an allergic skin reaction.

• Aquatic Chronic 4 H4131: May cause long lasting harmful effects to aquatic life.

Hazard statements

• H315: Causes skin irritation.

• H319: Causes serious eye irritation.

• H317: May cause an allergic skin reaction.

• H410: Very toxic to aquatic life with long lasting effects.

The self-classification database contains 17 aggregated notifications. These are summarised below in Table 4. Entry number 11 corresponds to the self classification proposed by the registrant.

1 The original registration dossier proposed the following: • Aquatic Acute 1 H400: Very toxic to aquatic life (M-factor 1). • Aquatic Chronic 1 H410: Very toxic to aquatic life with long lasting effects (M-factor 1).


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Table 4: Overview of the aggregated notifications in the self classification database (as of August 2013)

Self classification and labelling Entry No. Number of notifications Hazard Class and

Category codes Hazard Statement codes

Classification Labelling

Skin Irrit. 2

Eye Irrit. 2

Aquatic Acute 1

Aquatic Chronic 1

H315

H319

H400

H410

H315

H319

H400

H410

1 280

Aquatic Acute 1

Aquatic Chronic 1

H400

H410

H400

H410

2 93


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Self classification and labelling Entry No. No. notific. Hazard Class and Category codes

Hazard Statement codes

Classification Labellling

Skin Irrit. 2 H315 H315 3 91 Eye Irrit. 2 H319 H319 Aquatic Acute 1 H400 H400 Aquatic Chronic 1 H410 H410

Not Classified 4 76

Skin Sens. 1 H317 H317 5 43

Aquatic Acute 1 H400 H400

Aquatic Chronic 1 H410 H410

Skin Irrit. 2 H315 H315 6 34

Eye Irrit. 2 H319 H319

Aquatic Acute 1 H400


Skin Irrit. 2 H315 H315 7 28



Skin Irrit. 2 H315 H315 8 23




Skin Sens. 1 H317 H317 9 19


Skin Irrit. 2 H315 H315 10 19



Skin Irrit. 2 H315 H315 11 3 (same as

Skin Sens. 1 H317 H317




lead registrant)

Skin Irrit. 2 H315 H315 12 3



Acute Tox. 2 H300 H300 13 3

Skin Irrit. 2 H315 H315

Skin Sens. 2 H317 H317


Acute Tox. 4 H332 H332

STOT SE 3 H335 H335

STOT RE 2 H373 H373



Skin Irrit. 2 H315 H315 14 2



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Self classification and labelling Entry No. Number of

Hazard Class and Category codes

Hazard Statement codes

Classification Labellling

notifications

Skin Irrit. 2 H315 H315 15 1


STOT RE 1 H372 H372



Aquatic Chronic 1 H410 H410 16 1

Acute Tox. 4 H302 H302 17 1


Skin Irrit. 2 H315 H315





3 ENVIRONMENTAL FATE PROPERTIES

3.1 Degradation

3.1.1 Abiotic degradation

3.1.1.1 Hydrolysis

No information is given in the registration dossier. EA (2005) considered it unlikely that TDM would undergo significant hydrolysis in the environment based on the chemical structure.

3.1.1.2 Oxidation

EA (2005) considered that, although abiotic degradation of thiols to disulfides or sulfonic acids by oxidation is reported in the literature, the significance of this process for TDM in the environment was unknown.

The registration dossier gives the results of a preliminary oxidation test carried out using the OECD 111 method (reliability rating 2). This test was considered a supporting study in the registration dossier. The TDM tested was a commercial sample with a purity of 99.3%. The test was carried out using both algal culture medium (prepared in accordance with the OECD 201 test guideline) with a pH of 8 and also buffer solution with a pH of 7. TDM was added to the media at 10 mg/l and incubated for up to 150 days at 20°C either under aerated (aerobic) conditions or non- aerated (anaerobic) conditions. A co-solvent (acetonitrile) at 10% v/v was used to maintain the substance in solution. The primary degradation of TDM was followed by parent compound analysis.


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TDM was found to degrade slowly under the aerated conditions, with a half-life of approximately 150 days in both algal medium and pH 7 buffer. Under non-aerated conditions the half-life for TDM was found to be approximately 30 days in algal medium and 100 days in pH 7 buffer.

Analyses were also carried out for di-tert-dodecyl disulphide, the anticipated oxidation product of TDM. This was detected at a concentration of 0.2-0.3 mg/l in the non-aerated algal medium experiment but was at or below the limit of quantification (~0.1 mg/l) in the other experimental systems. It was concluded that the levels of di- tert-dodecyl disulphide found did not account fully for the level of degradation of TDM seen implying that degradation mechanisms other that oxidation may also be occurring. It was also concluded in the registration dossier that the 30 day half-life measured in the non-aerated algal medium was probably falsely short owing to poor agreement between replicates for the later samples and that overall this test shows that TDM can be degraded slowly in solution but the route/mechanism of degradation is uncertain.

When considering this test, it should be noted that TDM is relatively volatile (vapour pressure 20 Pa at 25° C). The full test report of the study indicates that precautions were taken to avoid potential loss from volatilization (use of sealed vials and sampling via septa). Therefor it is unlikely that volatile loss would have contributed significantly to the removal of TDM seen. The other point worth noting is that, although the test was carried out using 10% v/v of acetonitrile as a cosolvent, the concentration of TDM used (10 mg/l) is well above the recently determined water solubility of 0.0039 mg/l. The solubility of TDM in an acetonitrile:water mixture is unknown but it is possible that not all of the TDM would have been in solution in this test.

In conclusion, the results of this study suggest that oxidation of TDM in the environment is likely to be only a minor loss process.

3.1.1.3 Phototransformation/photolysis

3.1.1.3.1 Phototransformation in air

No information is given in the registration dossier. EA (2005) estimated the atmospheric half-life for reaction with hydroxyl radicals as around 1.6 days based on an estimated rate constant for reaction with hydroxyl radicals of 36.5×10-12

cm3.molecule-1 s-1 for the substance 2,2,4,6,6-pentamethyl-4-heptanethiol and assuming an atmospheric hydroxyl radical concentration of 5×105 molecules cm-3. The rate constant used was estimated using the AOPWIN v1.91 programme.

3.1.1.3.2 Phototransformation in water

No information is given in the registration dossier.

3.1.1.3.3 Phototransformation in soil

No information is given in the registration dossier.


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3.1.2 Biodegradation

3.1.2.1 Biodegradation in water

3.1.2.1.1 Estimated data

No estimates are given in the registration dossier. Experimental data are available (see below).

3.1.2.1.2 Screening tests

A modified OECD 310 Test Guideline ready biodegradability test has been carried out with TDM (Davis et al., 2009). The test material used was a commercial sample with a purity of 99.9% and the test was carried out in accordance with GLP.

The substance was added to the test system coated on silica gel (as an inert support) in order to maximise its availability to the microbial inoculum in accordance with the ISO 10634 (1995) guidance. Two loading rates were used in the study. A nominal loading rate of 20.5 µmoles TDM/g silica gel (4.15 mg TDM/g) was firstly prepared by adding the test substance directly to the silica gel in a sealed bottle under argon atmosphere and mixing for three days. A nominal loading of 2.05 µmoles TDM/g silica gel was then prepared by mixing 1.1 g of the treated silica gel with 10.3 g of unspiked silica gel followed by mixing for 1 day. The loading rates, and uniformity of the spiked samples were confirmed by analysis of triplicate samples immediately after preparation of the silica gel and after preparation of the test microcosms (the mean loading rates determined were 16.9 µmol/g and 1.70 µmol/g at the two loading rates, respectively).

The inoculum used in the study was derived from activated sludge mixed liquor collected from a municipal waste water treatment plant treating predominantly domestic waste water (>90% from domestic sources). The mixed liquor suspended solids (MLSS) concentration of the activated sludge was 1,230 mg/l and appropriate volumes were added to mineral salts medium to give a nominal MLSS concentration in the test microcosm of either 30 mg/l or 4 mg/l4.

The tests were carried out using a series of sealed 160 ml glass serum bottles containing 75 ml of mineral salts media inoculated with MLSS at either 4 or 30 mg/l and containing TDM (adsorbed onto silica gel) at a nominal concentration of either 2 µM (~0.4 mg/l) or 20 µM (~4 mg/l). The 2 µM concentration was around twice the estimated water solubility for TDM (given as 1.4 µM, which is equivalent to a water solubility of 0.28 mg/l (the estimated water solubility given in the registration dossier). As discussed in Section 1.5 a much lower water solubility of 0.0039 mg/l has recently become available and so the 2 µM treatment may have been as much as 100 times higher, and the 20 µM treatment as much as 1,000 times higher than the actual water solubility of TDM. The significance of the new water solubility on the

4 A concentration of 30 mg/l is the highest concentration recommended in the OECD 301 ready biodegradation test guideline. A concentration of 4 mg/l is the concentration recommended in the OECD 310 test guideline, although a higher concentration of 30 mg/l can also be used.


19

bioavailability of TDM in this study is unclear but it is possible that the bioavailability may still have been limited even though the substance was adsorbed onto silica gel.

Viability controls (containing 25 mg/l of aniline and MLSS), toxicity controls (containing MLSS and both aniline and TDM) and inoculum blanks (containing MLSS only) were also prepared. In addition abiotic controls (containing heat sterilized MLSS and TDM) were also prepared in order to assess abiotic loss of TDM. The tests were carried out at 20°C.

The degradation was followed by monitoring the disappearance of TDM at various time periods (primary degradation). For this, replicate bottles (two or three per time point) were extracted with acetonitrile for 3 hours on a rotary shaker and the concentration of TDM determined. In addition, the formation of carbon dioxide (mineralization) was also determined at certain time points. The degradation of aniline was determined based on dissolved organic carbon measurements. The concentrations of TDM measured in the experiments using an initial TDM concentration of 20 µM are summarized in Table 5. The carbon dioxide measurements taken during the study indicated that little or no mineralization of TDM was occurring.

Table 5: Summary of the concentrations of TDM measured in the ready biodegradability test system

Time point

TDM Concentration (µM) 1, 2

(days) Inoculum concentration 4 mg/l MLSS

Abiotic control – 4 mg/l MLSS –

sterilized

Inoculum concentration 30 mg/l MLSS

Abiotic control – 30 mg/l MLSS -

sterilized

0 21.0, 15.1

[18.0±4.2]

23.6, 18.3

[20.9±3.7]

26.2, 31.6

[28.9±3.8]

30.7, 10.0

[20.3±14.6]

7 Not sampled Not sampled 8.8, 12.6

[10.7±2.7]

12.2, 15.6

[13.9±2.4]

14 16.2, 12.5, 11.7

[13.5±2.4]

16.4, 12.5

[14.4±2.8]

9.4, 10.8

[10.1±1.1]

8.4, 7.3

[7.9±0.8]

21 Not sampled Not sampled 3.30, 5.30

[4.3±1.4]

4.76, 3.20

[4.0±1.1]

28 15.1, 16.5

[15.8±1.0]

15.1, 19.2

[17.2±2.9]

3.35, 6.61

[5.0±2.3]

7.26, 6.58

[6.9±0.5]

48 Not sampled Not sampled 1.56, 3.00, 2.09

[2.22±0.73]

4.39, 4.36, 4.33

[4.36±0.03]

50 Not sampled Not sampled Not sampled 5.86, 5.50, 3.01

[4.79±1.55]

56 7.5, 11.0, 6.5

[8.4±2.4]

12.7, 10.2, 6.3

[9.7±3.2]

Not sampled

Note: 1) The values relate to individual replicates taken at each sampling point. 2) Values in square brackets are the mean ± standard deviation of the individual values.

The viability control with aniline showed >85% degradation within 14 days. The toxicity controls indicated that TDM was not inhibitory to the microbial inoculum (there was some evidence of an initial lag in the degradation of aniline at a MLSS


20

concentration of 4 mg/l but extensive degradation of aniline (89%) was still evident by day 21; there was no evidence of inhibition at a MLSS concentration of 30 mg/l).

For the experiments carried out at a TDM concentration of 2 µM and 30 mg/l MLSS the concentration of TDM was found to be below the limit of quantification (~0.2 µM) by day 14 in both the biotic samples and abiotic control (no experiments were carried out at this TDM concentration with an inoculum concentration of 4 mg/l MLSS). This indicates >90% primary degradation occurred within 14 days in this system.

Assuming that the primary degradation of TDM followed first order kinetics, Davis et al. (2009) estimated the degradation rate constant for the experiments carried out using a TDM concentration of 20 µM by least squares curve fitting a plot of the mean Ct/C0 versus time5. The degradation rate constants obtained by this method are summarized below (R2 = the correlation coefficient of the curve fitting; no other statistics were reported).

Inoculum 4 mg/l MLSS – k = 0.0123 day-1; R2 = 0.59.

Abiotic control 4 mg/l MLSS – k = 0.013 day-1; R2 = 0.61.

Inoculum 30 mg/l MLSS - k = 0.0923 day-1; R2 = 0.90.

Abiotic control 30 mg/l MLSS - k = 0.0534 day-1; R2 = 0.74.

Based on these data, Davis et al. (2009) estimated the primary degradation half-life to be 7.5 and 13.0 days in the biotic system and abiotic control respectively at the 30 mg/l MLSS concentration, and 56.4 and 53.3 days respectively at the 4 mg/l MLSS concentration. These results are reported in the robust study summary in the registration dossier for this test (the study is considered a supporting study in the registration dossier).

Although this approach to estimating the degradation rate constants is valid, it should be noted that each time point is dependent on the C0 value and thus any uncertainty in the C0 value will translate throughout the entire dataset. As can be seen from Table 5, the uncertainty (as indicated by the standard deviation around the mean) is quite high for some of the C0 values, particularly the abiotic control at the 30 mg/l MLSS loading, where the standard deviation is around 72% of the mean.

An alternative, equally valid, approach to estimating the degradation rate constant is to plot ln [concentration of TDM] against time. The slope of this plot gives the degradation rate constant and the intercept gives an estimate of the C0 value. Such plots have been constructed by the eMS using the individual data points (rather than the mean concentrations) and the resulting degradation rate constants are summarized below, along with the resulting statistics from the linear regression analysis6.

5 For first order degradation, Ct/C0 = e-kt,where Ct = the concentration at time t, C0 is the concentration at time=0, and k is the first order degradation rate constant.

6 R2 = correlation coefficient of the regression. Standard error is the standard error in the slope of the regression plot (degradation rate constant). A p value <0.05 indicates that there is a 95% probability that the slope of the plot is different from zero.


21

Inoculum 4 mg/l MLSS – k = 0.0131 day-1; Standard Error ± 0.0035; R2 = 0.64, p=0.0054.

Abiotic control 4 mg/l MLSS – k = 0.0132 day-1; Standard Error ± 0.0040; R2 = 0.60, p=0.0138.

Inoculum 30 mg/l MLSS - k = 0.0484 day-1; Standard Error ± 0.0065; R2 = 0.84, p=1.21×10-5.

Abiotic control 30 mg/l MLSS - k = 0.0236 day-1; Standard Error ± 0.00558; R2 = 0.56, p=. 0.00085.

As can be seen, when analysed in this way the degradation rate constants for the two 4 mg/l MLSS experiments are similar to those estimated by Davis et al. (2009) but the rate constants estimated for the 30 mg/l MLSS experiments are around half of those estimated by Davis et al. (2009).

Two important points are evident from the degradation data. Firstly, it is clear that the primary degradation of TDM occurred in both the biotic and abiotic systems and secondly the degradation rate constant obtained at the 30 mg/l MLSS concentration was higher than at the 4 mg/l MLSS concentration.

In order to investigate the possible mechanism of degradation further, Davis et al. (2009) carried out further series of experiments as follows:

a) A study was carried out to investigate whether the decline in the TDM concentration with time was a result of increased sorption of the test substance to sludge biosolids as the experiment progressed. For this, an abiotic control mixture prepared with 30 mg/l MLSS was incubated for 50 days and then subjected to a more stringent, prolonged, solvent extraction regime prior to analysis (extraction with acetonitrile for 72 hours at room temperature versus 3 hours in the main test). This did not result in an increased recovery of the test substance compared with the standard extraction method used in the main test, suggesting that the decline in concentration was not a result of sorption. However, discussion between the eMS and Industry has indicated that this extraction method would not have removed any covalently bound substance.

b) A study was carried out with TDM (20 µM) incubated in a) heat sterilized mineral salts medium without addition of the inoculum and b) a filtered abiotic control mixture (prepared by adding 30 mg/l MLSS to a mineral salts medium, heat sterilizing the mixture followed by filtration through a 0.45 µm filter to remove the suspended biosolids; the purpose of this was to evaluate the role of soluble matter derived from the inoculum in the degradation). The incubations were carried out for 26 days with triplicate samples sacrificed for analysis on days 0, 6, 17 and 26. After 26 days there was around a 14% decline in the TDM concentration in the mineral salts medium whereas there was around a 52% decline in the TDM concentration in the filtered abiotic control experiment. The concentrations measured are summarized in Table 6.


22

Table 6: Summary of the concentrations of TDM measured in mineral salts medium and filtered abiotic control experiments

Time point (days)

TDM Concentration (µM) 1, 2

Mineral salts medium Filtered abiotic control derived from 30 mg/l MLSS

0 23.2, 24.4, 12.7

[20.1±6.4]

17.4, 48.8, 29.9

[32.0±15.8]

6 20.2, 22.6, 11.0

[17.9±6.1]

22.3, 24.9, 20.4

[22.5±2.3]

17 13.6, 25.1, 16.9

[18.5±5.9]

23.5, 23.0, 15.8

[20.8±4.8]

26 19.6, 14.6, 17.8

[17.3±2.5]

14.3, 16.6

15.5±1.6

Note: 1) The values relate to individual replicates taken at each sampling point. 2) Values in square brackets are the mean ± standard deviation of the individual values.

No degradation rate constants were derived by Davis et al. (2009) for the experiments in sterilized mineral salts medium and the sterilized filtered abiotic control. The eMS has plotted ln [concentration] versus time, which reveals the following degradation rate constants:

Sterilized mineral salts medium: k = 0.0030 day-1; Standard Error ± 0.0083; R2

= 0.013, p=0.73.

Sterilized filtered abiotic control prepared from 30 mg/l MLSS: k = 0.0220 day-1; Standard Error ± 0.0089; R2 = 0.41, p=0.035.

The rate constant obtained in the mineral salts medium is not statistically significantly different from zero (p>0.05) indicating that minimal primary degradation occurred in this test system. The degradation rate constant obtained in the filtered abiotic control is of a similar order to that obtained in the abiotic control that was not filtered (see above).

Overall, the results of this test show that although TDM cannot be considered to be readily biodegradable (little or no mineralization was seen), extensive primary degradation of TDM was occurring in the test. Davis et al. (2009) concluded that the presence of biological matter in the form of (or derived from) microbial suspended solids, either viable or heat inactivated, was the primary factor contributing to the removal of TDM in these test systems. This interpretation is consistent with the reanalysis of the test results carried out by the eMS as part of the current evaluation.

When considering the results of the test, it is relevant bear in mind that TDM itself is a UVCB substance. Although the substance tested had a stated purity of 99.9% this relates to the mercaptan content of the substance. The substance tested would likely have contained different constituents with varying position of the sulphur group in the molecule and different combinations of alkyl groups attached to the sulphur (see Section 1.2 for further information on the likely composition of TDM). In terms of mineralization, the UVCB nature of TDM would not significantly affect the interpretation of the results as the theoretical amount of carbon dioxide that would be


23

released from the main constituents of the substance would be predicted to be similar (and in any case little or no evolution of CO2 was evident).

The significance of this in terms of primary degradation is more difficult to assess as it is dependent on which constituents of the substance are detected by the analytical method used. The method involved derivatisation of the sulphur group in TDM after extraction followed by HPLC with MS/MS detection (negative ion electrospray ionization with multiple reaction ion monitoring). The method has been validated for the analysis of TDM in aqueous media (Perala and Markham, no year). The derivative had an expected molecular weight of 386 and the molecular ions found between m/z=385 (corresponding to the deprotonated derivative) and m/z=216 (corresponding to the derivative where all alkyl chains have been lost) were monitored. This method should, as far as can be established, therefore identify the majority, if not all, of sulphur containing components present in the TDM tested.

A further point worth noting (and not considered by Davis et al. (2009)) is that the primary degradation of TDM appears to be well described by a first order degradation process. Normally in a static biotic system such as that used here it would be expected that the microbial community would increase during the test, and that the degradation would not then necessarily follow first order kinetics. Thus the fact that first order kinetics appears to be followed provides some, albeit circumstantial, support for an abiotic removal mechanism in this test system. It is also relevant to note that other process, such as binding/adsorption or partitioning could also theoretically follow first order kinetics and so the kinetics would not distinguish between true degradation of TDM and some of these other loss processes. Although some tests were carried out using a more prolonged solvent extraction period prior to analysis, the extraction method used would not have determined any TDM that was covalently bound to the solids present. Therefore, this remains an uncertainty with the study. It could be argued that if covalent bonding did occur this could be considered a removal process in itself (as the covalently bound substance would have different properties to TDM and such binding may or may not be reversible)

In terms of estimating the potential for degradation of TDM in the environment, the results of this test are difficult to interpret. The test clearly shows that primary degradation of TDM is occurring and that this degradation depends on the presence of sludge biosolids. However the actual mechanism of the degradation is unclear and so it is difficult to relate the conditions used in this test to the conditions likely to be found in the environment.

The experiments were carried out at 20oC and for REACH purposes it is standard practice to extrapolate to ambient environmental conditions (typically a value around 12°C is assumed). The variation of the primary degradation rate constant with temperature is not known from the current study and so the default temperature correction in EUSES program has been used to adjust the experimental half-lives to a temperature of 12°C. This has been done using the rate constants as presented by Davis et al. (2009) and the rate constants obtained by reanalysis of the data. The results are summarized in Table 7.


24

Table 7: Summary half-lives estimated from the Davis et al. (2009) study

Test system Based on analysis carried out by Davis et al. (2009)

Based on the reanalysis of the data by the eMS

Rate constant derived at 20°C

Half-life at 20°C (days)

Half-life extrapolated

to 12°C (days)

Rate constant

derived at 20°C

Half-life at 20°C (days)

Half-life extrapolated

to 12°C (days)

Inoculum 30 mg/l MLSS

Abiotic control 30 mg/l MLSS

Inoculum 4 mg/l MLSS

Abiotic control 4 mg/l

MLSS

0.0923 7.5 14.2 0.0484 14.3 27.1

0.0534 13.0 24.7 0.0236 29.4 55.8 0.0123 56.4 107 0.0131 52.9 100

0.013 53.3 101 0.0132 52.5 99.6

Filtered abiotic control 30 mg/l MLSS

Not derived

- - 0.0220 31.5 59.7

It should be noted that the half-lives determined in this study should be used with caution as the test system used is designed to establish whether or not a substance is readily biodegradable rather than to obtain a half-life for degradation that is applicable to the general environment. The MLSS concentrations used in the study (4 or 30 mg/l) are in fact similar to the suspended matter content that is assumed in freshwater environments under REACH (e.g. a suspended matter concentration of 15 mg/l is usually assumed) but the similarity of the MLSS (in terms of physical properties, microbial populations, organic carbon etc.) to the suspended matter found in natural waters is unknown. Therefore, these half-lives should be considered as indicative values rather than precise values for the likely environmental half-life of TDM.

The Davis et al. (2009) study was included in the registration dossier as a supporting study (and was given a reliability rating of 2). The robust study summary is generally consistent with the original analysis of the data in the study report. However, the registrant could update the robust study summary to take account of some of the further discussion of the data by the eMS as outlined above.

The registration dossier contains two further screening studies for biodegradation (Jenkins (1990) and Thiebaud (1994)). A brief summary of these studies is given below based on the robust study summaries (the full test reports for these studies have not been reviewed by the eMS for the purposes of this evaluation but the study by Thiebaud (1994) has been reviewed previously for the EA (2005) report and was given a reliability rating of 2).


25

The key study (reliability rating 2) in the registration dossier was an OECD 301D test (ready biodegradability: closed bottle test) referenced as Jenkins (1990) in the registration dossier. The inoculum used was non-adapted activated sludge from a predominantly domestic source and the TDM concentration tested was 2 mg/l. Oxygen consumption during the test was found to be negligible compared with the control flasks and it was concluded that the substance was not readily biodegradable. The reference substance (sodium benzoate) achieved 57% degradation by day 5 and 63% degradation by day 28. Toxicity controls showed that TDM was not inhibitory to the test system at a concentration of 2 or 10 mg/l. Overall this test shows that TDM is not readily biodegradable but it should be noted that the concentration tested was in excess of the water solubility of TDM and so not all of the substance added may have been bioavailable in this test.

A second OECD 301D test (reliability rating 2) was presented in the registration dossier as a supporting study (Thiebaud, 1994). Few details of the test are available but it was indicated in the registration dossier that owing to the low solubility of the test substance the method described in the Annex to the OECD 301D method was applied. The degradation was monitored by dissolved organic carbon (DOC) removal and 10.4% degradation of TDM was found after 28 days. The reference substance (sodium benzoate) reached 81.9% degradation after 14 days and the toxicity control showed that TDM was not inhibitory to the test system. This test again shows that TDM is not readily biodegradable but it is not clear if the substance was fully bioavailable in this test system (there is no information on the concentration tested).

EA (2005) contains the results of two non-GLP tests conducted according to the OECD 301D test guideline. Both of these studies resulted in 0% degradation. Full details of these tests were not available for EA (2005). The references to the studies given in EA (2005) were ATOCHEM (1990) and Bayer AG internal study (1973). It is possible that one of these studies is the same as the Jenkins (1990) study reported above.

The previous PBT factsheet (PBT Working Group, 2008) for TDM prepared by the eMS briefly reports the results of a screening test conducted with TDM according to OECD TG 301B in the presence of a non-biodegradable emulsifier. These data do not appear in the registration dossier. Biodegradation reached 30-40 per cent within the first 28 days, suggesting that TDM is not persistent. However, a follow-up test was being planned (i.e. the OECD 310 study by Davis et al. (2009) reviewed above).

3.1.2.1.3 Simulation tests

No data are given in the registration dossier.

3.1.2.2 Biodegradation in sediments


3.1.2.3 Biodegradation in soil



26

3.1.2.4 Summary and discussion on biodegradation

The available data show that the substance cannot be considered to be readily biodegradable in standard ready biodegradation test systems. However, many of the tests were carried out at concentrations in excess of the reported water solubility of the test substance, which may have limited the bioavailability of TDM in the test system used (i.e. it is possible that biodegradation might have been more extensive at lower concentrations). Primary degradation has been observed in an enhanced test system.

3.1.3 Summary and discussion on degradation

TDM is predicted to degrade rapidly in the atmosphere by reaction with hydroxyl radicals and the half-life estimated for this reaction is 1.6 days.

Oxidation of TDM to the corresponding disulphide could theoretically occur in the environment but the available screening experimental data suggest that this is not likely to be a major removal process in aquatic systems. Similarly hydrolysis is not considered to be a major removal process based on structural considerations.

The substance is not readily biodegradable in standard test systems; however the concentrations tested in some of the studies are well above the water solubility of TDM which may mean that the bioavailability of the substance could have been limited in the test systems used.

A recent study has been carried out to investigate the biodegradation of TDM using an enhanced test system. The results of this test appear to show that TDM undergoes substantial primary biodegradation in the presence of sewage sludge (or substances derived from sewage sludge). However, there are some uncertainties remaining over whether at least some of the removal seen could have resulted from covalent binding to solids. The significance of these results is considered further in Section 7.2.

3.2 Environmental distribution

3.2.1 Adsorption/desorption

The registration dossier contains the result of an OECD 121 test guideline study (reliability rating 1; the full test report has not been seen for this factsheet). The method used was the HPLC method and the TDM used had a purity of 99.1%. The organic carbon-water partition coefficient for TDM was determined to be 3,980 l/kg (log Koc = 3.6).

EA (2005) estimated a higher log Koc value of >5.12 (Koc >131,825 l/kg) using a standard regression equation (log Koc = 0.81 × log Kow + 0.1), and assuming the log Kow value to be >6.27. However, a log Koc value of 3.50 (Koc = 3,180 l/kg) was estimated using PCKOCWIN v1.66 for the specific structure of 2,2,4,6,6-

7 The log Koc estimated using this equation assuming a log Kow of 7.43 is log Koc = 6.11 (Koc = 1,288,250 l/kg).


27

pentamethyl-4-heptanethiol, which is in very good agreement with the experimental value reported above.

Based on the available experimental data, the log Koc for TDM is around 3.6. This indicates that TDM will adsorb strongly onto sediment and soil in the environment and would not be expected to leach readily from soil. The Koc is lower than might be predicted from the log Kow alone.

3.2.2 Volatilisation

The Henry’s law constant of the substance is estimated to be around 14,490 - 19,230 Pa m3/mole, with other estimates being around 5,900 Pa m3/mole and 3.24×105 Pa m3/mole at room temperature (see Section 1 for discussion). This indicates that volatilisation from water to air could occur, although this is likely to be moderated to some extent by adsorption of the substance to particulate matter.

3.2.3 Distribution modelling

No distribution modelling was included in the registration dossier. EA (2005) estimated the likely environmental distribution of TDM using a Mackay level III fugacity model (Table 8). A water solubility of 0.25 mg/l and a vapour pressure of 400 Pa, both at 20°C, were used in these estimates.

Table 8: Summary of Mackay Level III fugacity modelling carried out by EA (2005)

Compartment Distribution (percentage mass)

100% release to air 100% release to water

100% release to soil

Air 99.9 0.036 1.66

Water 2.9×10-4 9.53 0.011

Sediment 2.7×10-3 90.4 0.11

Soil 0.11 4×10-5 98.2

EA (2005) concluded that the results indicate a general lack of movement between compartments, owing to high affinity for both air and solids.

Since the EA (2005) report was completed further information on the physico- chemical properties of TDM have become available (see Section 1) in particular, more recent determinations indicate that both the vapour pressure (20 Pa at 25°C) and water solubility (0.0039 mg/l at 20°C) may be lower than assumed in EA (2005) and the log Kow (7.43) may be higher. In order to investigate the possible significance of this, the Mackay Level III model (EQC V2.02) has been re-run using these properties (assuming an atmospheric half-life of 38.4 hours and nominal degradation half-lives of 2,400 hours in water and 10,000 hours in sediment and soil). The results are summarised in Table 9. The modelling using the new physico-chemical properties is broadly in line with that carried out in EA (2005).


28

Table 9: Summary of Mackay Level III fugacity modelling carried out by eMS

Compartment Distribution (percentage mass)

100% release to air 100% release to water

100% release to soil

Air 92.6 0.012 0.56

Water 7.04×10-4 2.43 2.01×10-3

Sediment 0.0282 97.6 0.081

Soil 7.38 99.4

3.3 Potential for long range transport

The long range transport potential of TDM was not considered in the registration dossier. For the purposes of this evaluation, the long range transport potential has been explored by the eMS using the OECD Pov and LRTP screening tool8 version 2.2.

In order to try to assess the effects of the uncertainties in several parameters, notably the Henry’s law constant, the log Kow and the primary degradation half-life in water, the modelling was carried out several times using different assumptions for these parameters. The inputs used and the resulting modelled outputs are summarised in Table 10. For all estimates the molecular weight was set at 202.4 g/mole, the degradation half-life in air was set at 38.4 hours (1.6 days), the half-life in soil was set at a nominal value of 10,000 hours (417 days) and the log Kow was set at either 6.07 or 7.43. The key outputs for the simulations are displayed graphically in Figure 1.

As can be seen from Figure 1 most of the simulations result in the substance appearing in the lower left hand quadrant, which signifies a low potential for long range transport. The main exceptions to this are the three simulations using the lowest value of the log Kaw (0.38) for both log Kow values and the three simulations using the log Kaw of 0.82 and a log Kow of 7.43, which are just in the lower right hand quadrant (indicating a higher potential for long range transport). When considering these results is should be noted that the lowest value of the log Kaw is derived from a QSAR estimate and the values obtained using the water solubility and vapour pressure are higher than this value and probably reflect better the actual volatility of the substance. When the most likely value for the log Kaw value of 2.62 is used (see

8 http://www.oecd.org/env/ehs/risk-assessment/oecdpovandlrtpscreeningtool.htm


29

Section 1.5), the simulations all indicate a low potential for long range transport. In addition, as discussed in Section 3.2.1 the adsorption of TDM to sediment and soil appears to be lower than would be expected based on the log Kow of 7.43. Also, it is important to note that for several of the simulations the longest overall persistence is predicted to occur for the soil compartment which is dependent on the assumptions made over the volatility and adsorption of the substance and the degradation half-life in soil. The actual degradation half-life in this compartment is unknown (a relatively long half-life of > 1 year has been assumed for the simulation here).

Overall, based on the available information and the simulations carried out, it can be concluded that TDM has a low potential for long-range transport.


30

Table 10: Summary of long range transport potential estimated using the OECD Pov and LRTP screening tool

Input assumptions 4 Modelled outputs1

Pov (days)2 CTD (km)3 TE (%) 4

log Kaw = 0.38

log Kow = 6.07

[log Kow = 7.43]

log Kaw = 0.82

log Kow = 6.07

[log Kow = 7.43]

log Kaw = 2.12

log Kow = 6.07

[log Kow = 7.43]

2.62

log Kow = 6.07

[log Kow = 7.43]

log kaw = 3.93

log Kow = 6.07

[log Kow = 7.43]

Half-life in water = 648 hours (27 days)

Half-life in water = 1,200 hours (50 days)














244 (soil)

[512 (soil)]

244 (soil)

[512 (soil)]

244 (soil)

[512 (soil)]

126 (soil)

[468 (soil)]

126 (soil)

[469 (soil)]

126 (soil)

[469 (soil)]

31 (water)

[140 (soil)]

49 (water)

[140 (soil)]

74 (water)

[140 (soil)]

31 (water)

[55 (soil)]

49 (water)

[58 (water)]

74 (water)

[97 (water)]

31 (water)

[35 (water)]

49 (water)

[58 (water)]

74 (water)

[98 (water)]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

797

[797]

0.00069

[0.0014]

0.00069

[0.0014]

0.00069

[0.0014]

0.00048

[0.00077]

0.00048

[0.00077]

0.00048

[0.00077]

0.00037

[0.00038]

0.00037

[0.00038]

0.00037

[0.00038]

0.00037

[0.00037]

0.00037

[0.00037]

0.00037

[0.00037]

0.00037

[0.00036]

0.00037

[0.00036]

0.00037

[0.00036]

Note: 1) Upper estimate assumes a log Kow of 6.07. The lower estimate [ ] assumes a log Kow of 7.43.

2) Pov is an estimate of the overall persistence of the substance in the environment. The compartment to which the persistence relates is given in brackets. The values for soil depend on the assumptions made over the half-life in soil, which is currently uncertain, and the assumed volatility of the substance.

3) Characteristic travel distance which is an estimate of the distance from a point source at which the chemical’s concentration has dropped to 38% of its initial concentration. For all the simulations here the CTD relates to transport by air and so will be dependent on the assumptions made over the half-life in air.


31

i5

"'

()

w

"c'

4) Transport efficiency. This is an estimate of the percentage of emitted chemical that is deposited to surface media after transport away fi·om the region of release

Figure 1 Long range transport potential of TDM

1.0E+07

Ê

oi 1.0E+06 ()

c ro ûi

1.0E+05

Q; > r..o. 1.0E+04

()

:;:: · ;: 1.0E+03 G>

ü r.o.. ro

a.

...J

1.0E+02

1.0E+01

1.0E+OO 1.E-01 1.E+01 1 E+03 1.E+05

POV {days)

1 OE+OO

1.0E+05

1.0E+04 ;? " 1 OE+03 :>;

1.0E+02 c •Gt>; 1.0E+01

E 1.0E+OO ... .!

r..o. 1.0E-01

1 OE-02

t:. 1 OE-03 a. 0::

1 OE-04 ...J 1 OE-05

1 OE-06

1.0E-07 1.E-01 1.E+01 1 E+03 1.E+05

POV (days)


32

4.1 Bioaccumulation

4.1.1 Screening data

The BCF has been predicted by the eMS using the BCFBAFv3.01 model within the EPIWIN programme. This gives estimates for the BCF using a regression equation and also a mass-balance food chain model. The mass-balance model also gives an estimate of the metabolism rate constant. The regression equations were developed using a training set of 466 non-ionic substances and a validation set of 158 substances was used to test the regressions. The regression equations themselves consist of two equations, one for substances with log Kow values in the range 1 to 7 (derived from 396 data points) and one for substances with log Kow values >7 (derived from 35 data points). Of the validation set, 11 substances had a log Kow of 7 or above. The data sets used for the regression contain several sulphur-containing substances but do not appear to contain an alkyl thiol specifically. The mass-balance model is a theoretical model that should apply to substances generally. The approach used for the development of the data base of metabolism rate constant values used in the BCFBAF was based on a training set of 421 substances and a validation set of 211 substances. The help file for the mass balance model indicates that the model can be broadly applicable to diverse chemical structures but notes that this may not reflect the entire domain of possible structural fragments and that it is difficult to define precisely the domain of applicability. The applicability of the methods for TDM should be seen in this context.

The estimates were carried out using two possible structures, differing in their degree of branching, and taking into account the various estimates of the log Kow. The results are summarised in Table 11 below.

Table 11: Summary of estimated BCFs using the BCFBAF model

Structure assumed (smiles notation)

log Kow BCF (l/kg) –

regression

BCF (l/kg) – mass-balance model

Metabolism rate

constant equation Lower

trophic level

Middle trophic

level

Upper trophic

level

(day-1) – normalized

to a 10 g fish

SC(CCC)(CCCC)CCCC 7.43 2,074 760 688 496 0.082

CC(C)(C)CC(C)(S)CC(C)(C)C [highly branched structure]

7.43 8,190 1,093 988 709 0.056

SC(CCC)(CCCC)CCCC 6.071 198 1,570 1,430 1,046 0.22

CC(C)(C)CC(C)(S)CC(C)(C)C [highly branched structure]

5.841 3,338 2,076 1,901 1,406 0.17

Note: 1) Estimated within the BCFBAF model.

As can be seen from Table 11, the regression equation predicts that the more linear structure generally has a BCF close to or below 2,000 l/kg, particularly when the lower log Kow estimate is used, but that the more highly branched structure would have a BCF above 2,000 l/kg and possibly above 5,000 l/kg based on some of the estimates. The regression equation does not necessarily take account of metabolism of


33

TDM directly, but the method includes group correction factors that take account of the potential (which can include metabolism potential) for certain functional groups to either increase or decrease the bioconcentration factor. It should also be noted that the BCF estimates are dependent on the log Kow. As indicated in Section 1.5, the available experimental value is a limit value (log Kow <6.2) and so estimates of the log Kow value have to be used in the calculations. The log Kow of 7.43 is estimated from the water solubility but the uncertainty around this value (or the other estimates) is unclear.

The mass balance model takes into account metabolism directly. The metabolism rate constants estimated for TDM in the model compare well with the metabolism rate constants estimated above from the study by Davis and Erhardt (2009) (summarised in Section 4.1.2). The BCF values obtained from the BCFBAF model are not normalised to a standard lipid content of 5% but reflect the lipid contents of the fish assumed in each trophic level (i.e. 5.98% for the lower trophic level, 6.85% for the middle trophic level and 10.7% for the upper trophic level). If the estimates from the BCFBAF mass balance model are normalised to a standard lipid content of 5% then the highest BCFs are obtained for the lower trophic level and the lipid normalised BCF would be 1,300 l/kg for the more linear structure and 1,740 l/kg for the upper trophic level using the log Kow values estimated within the BCFBAF model or 635 and 914 l/kg respectively using a log Kow of 7.43. It is not possible to lipid normalise the BCF from the regression equation. These lipid normalised values predicted using the BCFBAF mass balance model suggest that the BCF value for TDM is below 2,000 l/kg. However, it is important to note that the BCFBAF model estimates the BCF on a total water concentration basis rather than a dissolved water concentration basis by taking into account possible adsorption of the substance onto particulate organic carbon or association with dissolved organic carbon; both of these are assumed within the model to reduce the bioavailability. As TDM has a relatively high log Kow, it would be expected to be associated with these components in the environment but it does imply that the BCF based on a truly dissolved concentration obtained from the BCFBAF model would be higher9 than the figures given in Table 11, and the same would apply to the estimates given below.

Several other estimates of the BCF for TDM have been undertaken and details of these are given below:

a) Davis and Erhardt (2009) used the Gobas model (similar to the mass-balance model included in the BCFBAF programme). When the model was run without including metabolism, the BCF for TDM was predicted to be 24,144 l/kg but when the rate constant for metabolism of 0.112 or 0.165 day-1 was included the predicted BCF reduced to 862 and 592 l/kg respectively. The uptake rate constant predicted by the model was 100 l/kg/day, which is higher than those estimated in Section 4.1.2 and probably reflects a larger size of fish assumed in the model (trout rather than zebrafish).

9 For example, for a log Kow of 6.07 the model predicts that that the freely dissolved fraction would be around 80% of the total concentration. Expressed on a freely dissolved concentration basis the predicted BCFs would therefore be around 1.25 times larger than given. Similarly for a log Kow of 7.43 the freely dissolved fraction would be around 15% of the total concentration and the BCF expressed on the freely dissolved concentration would be around 6.7 times larger than given.


34

b) Davis and Hancock (2009) used two different models to assess the bioaccumulation potential of TDM, the BCFBAF program and the base-line model of Dimitrov et al. (2005). For the BCFBAF program three example structures were used to represent TDM and based on these the predicted BCF assuming no metabolism was in the range 19,880-20,990 l/kg. When metabolism was taken into account the predicted BCF was in the range 883- 2,076 l/kg using a metabolism rate constant of 0.112 day-1. The highest BCF values were obtained for the lower trophic level within the BCFBAF model but the values were not normalised to a standard 5% lipid content; normalising to a 5% lipid content would reduce all predicted BCFs to <2,000 l/kg (see above).

The Dimitrov et al. (2005) method derives a BCFmax value from a regression of log Kow with known BCF values and then adjusts the BCFmax according to various factors, including molecular size/diameter, ionization and potential for metabolism. Significant metabolism was predicted and the corrected BCF value was predicted to be in the range 651-1,220 l/kg. These estimates are in general agreement with the findings above.

c) Comber and Thomas (2013) carried out further predictions for the BCF of TDM using the BCFBAF model and a log Kow of 7.43 (which was estimated from a validated QSAR from the water solubility of TDM of 3.93 µg/l obtained from a slow stiring method (see Section 1.5)). Using the BCFBAF model BCFs of 496, 688 and 760 l/kg were obtained for the three trophic levels included in the model.

As noted in Section 1, TDM is a UVCB substance containing constituents (all mercaptans) with different carbon chain lengths (typically between C10 and C14) and different degrees and patterns of branching. In order to investigate the effect of carbon chain length and branching on the expected BCF, a series of predictions have been carried out by the registrant using the BCFBAF model (personal communication to the evaluating Member State, 6th December 2013). For these predictions the log Kow value of a range of structures was firstly predicted using a regression-fragment model (this model predicted the BCF for four representative C12 isomers to be in the range 7.0 to 7.6, which is in good agreement with the log Kow of 7.43 for TDM estimated from the validated QSAR from the water solubility). The log Kow values predicted for other carbon chain length mercaptans were in the range 5.5 to 5.8 for C9 isomers, 5.9 to 6.5 for C10 isomers, 6.5 to 7.0 for C11 isomers, 7.5 to 8.2 for C13 isomers, 7.9 to 8.7 for C14 isomers and 8.6 to 9.3 for C15 isomers. The BCFBAF model predicted that biodilution would occur. The BCFs predicted for the three trophic levels within the model are shown below.

Upper trophic level Middle trophic level Lower trophic level

C9 878-1,125 l/kg 1,191-1,517 l/kg 1,303-1,656 l/kg

C10 976-1,616 l/kg 1,341-2,209 l/kg 1,476-2,424 l/kg

C11 778-1,204 l/kg 1,076-1,660 l/kg 1,188-1,830 l/kg

C12 388-1,130 l/kg 5,38-1,571 l/kg 594-1,736 l/kg

C13 158-590 l/kg 218-826 l/kg 241-914 l/kg


35

C14 66-352 l/kg 91-490 l/kg 100-542 l/kg

C15 27-148 l/kg 37-206 l/kg 40-228 l/kg

As can be seen from the predictions all chain lengths, with the exception of some C10 structures, are predicted to have BCF values <2,000 l/kg in all three trophic levels. For the C10 structures, BCF values above 2,000 l/kg were predicted for the mid and lower trophic levels for two of the four structures tested and these structures had high degrees of branching. The mean BCF predicted for all of the C10 isomers was 1,287 l/kg for the upper trophic level, 1,758 l/kg for the mid trophic level and 1,930 l/kg for the lower trophic level. Based on these predictions it would be expected that the highest BCF values will occur for some of the C10 constituents that may be present in TDM and that the BCF would be expect to decrease with increasing carbon chain length. However the mean BCF value for each chain length group and each trophic level is predicted to be <2,000 l/kg.

4.1.2 Bioaccumulation in aquatic organisms (pelagic and sediment

organisms)

No experimental BCF value is available for TDM.

A dietary accumulation study has been carried out for TDM (Egeler et al., 2013) using a draft version of the OECD 305 test guideline (unpublished draft update of TG 305 of 31st August 2010; this was the version of the test guideline available at the time of the study protocol design). The study was carried out in accordance with GLP.

The substance tested was named tertiododecylmercaptan (CAS No. 25103-58-6) and had a purity of 99.7%. The identity of substance tested was stated to be the same as the registered substance. Hexachlorobenzene (14C-labelled; radiochemical purity 99.31%) was used as a positive control. The method used to analyse TDM involved derivatisation of the TDM followed by HPLC/MS/MS analysis. The method was similar to the method outlined in Section 3.1.2 but utilised a different chemical for the derivitisation. The range of masses monitored mean that most sulphur-containing constituents of TDMwould be likely to be detected by the method. Procedural recovery samples for fish indicated that the recovery was generally within the range 74% to 98% (11 samples) with one sample only giving a lower recovery (66%). Mean recoveries from procedural food samples were 103% and 107%. As the recoveries were generally in the range 70-110% the analytical method can be considered acceptable.

The test was carried out using zebrafish (Danio rerio) and the test consisted of a 15 day uptake period followed by a 42 day depuration period.

The food used in the test was ground fish food flakes. TDM was added to the food as a solution in cod liver oil to give a nominal concentration of 100 mg/kg food dry weight. The food for the positive control contained 14C-hexachlorobenzene at a concentration of 10.3 mg/kg food dry weight (in this case the hexachlorobenzene was added as a solution in cyclohexane to the cod liver oil before adding to the food). A negative control food, containing cod liver oil only, was also prepared. The amount of cod liver oil present in the treated and control foods was 0.1 ml oil per g food.

The concentration of TDM in the food was determined prior to the start of the test


36

(day -1; four samples) and at the end of the uptake phase (day 15; four samples). Each sample was analysed in duplicate and the range of concentrations found were between 87 and 112 mg/kg dry weight prior to the start of the test and between 96 and 117 mg/kg dry weight at the end of the uptake phase. These results demonstrate that the test substance was well mixed within the food and that the concentration in food was stable over the experimental period. The overall mean concentration in food was 103.3 mg/kg food dry weight (standard deviation of ±8.7 mg/kg food dry weight). TDM was not detectable in the control food (<0.0041 mg/kg food dry weight). The lipid content of the food was determined to be 18.6% of wet weight based on measured concentrations at the start.

Four replicates were carried out for the TDM and negative control exposure, each replicate consisting of 15 fish at the start of the exposure. One replicate of 15 fish was used for the positive control (14C-hexachlorobenzene). Prior to the start of the test the average body weight of the fish was 252.6 mg/fish (range 206.3-306.0 mg/fish; measured 8 days prior to the start of the exposure). The fish were fed at a daily rate of 3% of the fish body weight (fed in three portions separated by 2-5 hours; one portion at weekends). The temperature was in the range 23.6 to 26.0°C during the test.

Groups of four fish were analysed for the presence of TDM on days 0, 1, 3, 8 and 15 (eight fish sampled) of uptake and days 0.210, 1, 4, 7, 14, 28 and 42 of depuration. The weights of the fish were also recorded on each sampling date. Lipid contents of the fish were determined on day 0 of uptake (three fish) and day 42 of depuration (six fish). The data are summarised in Table 12. All fish showed normal feeding behaviour and ingested the administered food within a regular observation period. There were single incidents of individual fish in the controls and positive control being reluctant in accepting food but the food was readily consumed in the treatment group. Other than that an absence of food remnants in the tanks was generally noted.

The concentration of the positive control (14C-hexachlorobenzene) was determined on day 1 and day 15 by radiochemical analysis. Based on the day 1 measurements the mean assimilation efficiency of hexachlorobenzene was estimated to be 51.9%. This demonstrates that the substance was taken up from the food used. The mean BMF (not growth or lipid corrected) at day 15 was estimated to be 0.134.

The following parameters were estimated from the data in the Egeler et al. (2013) test report (see below for further discussion).

• Rate constant for growth dilution – 0.00397 day-1.

• Lipid correction factor (ratio of fish lipid/food lipid) = 0.105/0.186 = 0.56.

• The uptake and depuration rate constants were obtained simultaneously by nonlinear regression using all from the uptake and depuration phase. Where

10 The samples were taken between three and five hours after the start of depuration.


37

the concentration in the fish was below the limit of quantification (i.e. <0.01 mg/kg) a value of ½ of the quantification limit (i.e. 0.005 mg/kg) was used in the curve fitting. This lead to the following values:

o Uptake rate constant – k1 = 0.00152 kg food/kg fish/day11

o Overall depuration rate constant – k2 = 0.6065 day-1.

o Growth corrected depuration rate constant – k2g = 0.6026 day-1.

• The concentration in fish at the start of the depuration phase (C0,d) was determined as 0.3649 mg/kg based on the mean concentrations measured in the fish at the end of the uptake phase (day 15 values).

• The assimilation efficiency (α) was determined to be 0.0714 (i.e. 7.14%) using the feeding rate of 0.03 kg food/kg fish/day and the concentration in food (Cfood) of 103.3 mg/kg food dry weight.

• The following BMF values were derived.

o Kinetic BMFK = 0.0025 (calculated as k1/k2).

o [BMF based on day 15 concentrations = 0.0035].

o Growth-corrected kinetic BMFKg = 0.00356 (calculated as I×α/k2g).

o Lipid-corrected kinetic BMFL = 0.00447.

o Lipid- and growth-corrected kinetic BMFLg = 0.00635.

11 Note: in the feeding study k1 effectively equals I×α, where I is the food rate (kg food/kg fish/day) and α is the fractional assimilation efficiency.


38

Table 12: Summary of concentration, lipid and weight measurements from the dietary bioaccumulation study

Time point (days) TDM Concentration (mg/kg)1

Uptake

Fish weight (g)1 Mean fish lipid content (% of wet

weight)

0 (Control) <LOQ, <LOQ, <LOQ, <LOQ

3 (Exposed) 0.159, 0.024, 0.061, 0.299

8 (Exposed) 0.0359, 0.902, 0.107, <LOQ

0.3041, 0.3838, 0.4195, 0.5371

0.4827, 0.4069, 0.5804, 0.4971

0.4517, 0.3383, 0.2524, 0.2938

8.4%

15 (Control) <LOQ 0.5481, 0.3475, 0.2305, 0.2719

10.7%

15 (Exposed) 0.370, 0.220, 0.324, <LOQ, 0.963, 0.431,

0.281, 0.325

0.3627, 0.3568, 0.2995, 0.2161, 0.3567, 0.5211,

0.6434, 0.4845

10.4%

[Mean of control and exposed = 10.5%]

Depuration

0.2 (Exposed) <LOQ, 0.0633, 0.0863, 0.158

1 (Exposed) 0.013, 0.157, 0.0616, 0.189

4 (Exposed) 0.0172, 0.0347, 0.0309, 0.0779

7 (Exposed) <LOQ, <LOQ, 0.0227, 0.029

14 (Exposed) 0.0386, <LOQ, <LOQ, 0.0136

28 (Exposed) <LOQ, <LOQ, <LOQ, <LOQ

42 (Control) <LOQ, <LOQ, <LOQ, <LOQ

42 (Exposed) <LOQ, <LOQ, <LOQ, <LOQ

0.2982, 0.7376, 0.4008,

0.2996

0.2789, 0.5545, 0.5545, 0.2801

0.6297, 0.3065, 0.5647, 0.5415

0.5771, 0.3930, 0.4468, 0.4902

0.5175, 0.1494, 0.4301, 0.5877

0.3808, 0.6013, 0.3955, 0.3127

0.6230, 0.3575, 0.5055, 0.3260

0.5085, 0.4848, 0.5148, 0.4612

16.7%

Note: 1) Values relate to individual fish sampled at each time point. LOQ = Limit of quantification, which was 0.01 mg/kg.

The study was well carried out. There are, however, a number of minor issues with the study, and these are considered further below.

• The original target TDM test concentration was 10 mg/kg food dry weight. However an error during preparation of the food resulted in a concentration of 100 mg/kg food dry weight being tested instead. A pre-study had been carried out to show there were no effects on the palatability of the food up to a concentration of 10 mg/kg food dry weight but higher concentrations were not tested. The study report says that all food was eaten, and no effects on the weight of the fish compared to the control fish were evident. Therefore this deviation is considered to have not adversely affected the results of the test.


39

• The concentrations in food used in the calculations were on a dry food weight basis but the lipid content of food used in the calculations was based on a wet food weight basis. The concentrations in food and lipid contents should be on the same basis for the calculations, preferably on a wet food weight basis.

• The curve fitting carried out by Egeler et al. (2013) resulted in a k1 value of 0.00152 kg food/kg fish/day. This value is slightly lower than would be expected based on the assimilation efficiency (α) derived from the C0,d value and the feeding rate (I). That is, k1 should effectively equal α×I = 0.0714×0.03= 0.0021 kg food/kg fish/day. However in the Egeler et al. (2013) analysis the k1 value obtained directly is only used in the calculation of the BMFK and BMFL values. The remaining growth corrected kinetic BMF values are estimated from the assimilation efficiency and feeding rate.

• The C0 value was determined based on analysis of the exposed fish at day 15 of the uptake phase. These data could have been influenced by the presence of undigested food in the guts (given the relatively high concentration tested). This may have led to an overestimation of the assimilation efficiency (and may explain the apparent discrepancy noted above). An alternative way to estimate the C0,d value, that may be less influenced by potential undigested food, would have been to estimate the C0,d from the intercept of a plot of ln [Concentration] versus time for the depuration phase.

• The inclusion of concentrations below the limit of quantification in the curve fitting (they were assigned a value of half of the limit of quantification) is not ideal, especially as the substance was not quantifiable in any of the day 28 and day 42 samples. Furthermore, as both the k1 and k2 were determined by Egeler et al. (2013) by non-linear regression curve fitting techniques, any uncertainty introduced by the inclusion of these concentrations in the analysis would affect both the k1 and k2 values obtained. It would have been preferable, in the opinion of the eMS, to derive the k2 value directly from the depuration data alone and then derive the uptake parameters separately from the C0 value. This is considered further below.

• The kinetic BMFK of 0.0025 (not growth corrected or lipid normalised) is smaller than the BMF of 0.0035 estimated at day 15 based on the ratio of measured concentrations in fish and food. This may suggest that the k1 may have been underestimated and/or the k2 overestimated in the Egeler et al. (2013) analysis, although it should be noted that both methods give BMF values of a similar order (and importantly both well below 1).

• The rate constant for growth dilution (kg) was derived as 0.00397 day-1 in the test report. This was apparently based on the pooled control and treatment groups (no statistically significant differences were reported between the two groups individually). The statistics of the regression analysis from the report suggest that the correlation coefficient (R2) for the regression was 0.058 but that the correlation was statistically significant (p=0.015). The growth of the fish was clearly limited over the time course of the study, but the eMS has not been able to recreate the above rate constant from the data reported in the study (see below: it is not entirely clear which data points were included in the regression analysis by Egeler et al., 2013). Given the very low (compared with


12 A p-value <0.05 indicates that the slope of the line (i.e. k2) is statistically significantly different from zero at the 95% confidence level.

38

the overall depuration rate constant) rate constant for growth dilution it is questionable if growth correction is actually needed for this study (not including growth dilution would decrease the reported BMF slightly).

• The assumed lipid content of the fish was based on the lipid concentration measured in the control and exposed fish at day 15 of the uptake phase. The lipid content increased during the depuration phase and so the actual lipid content over the entire depuration period could be higher than assumed in the calculations. If a higher fish lipid content was assumed the lipid normalised BMFs would all be lower than given in the Egeler et al. (2013) paper and so the approach taken can be considered to be a worst case.

In order to investigate these issues further and the influence, if any, they may have on the derived bioaccumulation parameters, the raw data from Egeler et al. (2013) have been re-analysed by the eMS for the purposes of this evaluation as follows.

i) The overall depuration rate constant has been re-determined from the slope of a plot of ln [Concentration] versus time. This has been carried out firstly using only the concentration values that were above the limit of quantification. The plot is shown in Figure 2 below. The best fit line was obtained by linear regression. The slope and intercept of this plot gives the following values for k2 and C0,d:

• k2 = 0.104 day-1; standard error = 0.045 day-1; p-value = 0.03712; 95% Confidence interval 0.007 to 0.200 day-1

• C0,d = 0.0716 mg/kg; 95% confidence interval 0.040 to 0.13 mg/kg.

Figure 2 Plot of ln [Concentration] versus time for the depuration phase using only concentration above the limit of quantification

This method of analysing the data results in lower values for the k2 and C0,d than determined in the Egeler et al. (2013) of the data. However the above re-analysis


13 A p-value <0.05 indicates that the slope of the line (i.e. k2) is statistically significantly different from zero at the 95% confidence level.

39

effectively ignores that fact that at day 0.2, 7 and 14 the concentration in some fish was below the limit of quantification (and also in all samples at day 28 and day 42). To try to take this into account the above analysis has been repeated assuming a value of half of the limit of quantification (i.e. a concentration of 0.005 mg/kg) for the concentrations below the detection limit at day 0.2, 7 and 14 (as all of the data points at day 28 and day 42 are below the limit of quantification the eMS believes that including these data points in the analysis as well will introduce further uncertainty into the derived parameters). The relevant plot is shown in Figure 3.

Figure 3 Plot of ln [Concentration] versus time for the depuration phase using half of the limit of quantification for missing data points on days 0.2, 7 and 14

Using this approach the following parameters can be derived:

• k2 = 0.133 day-1; standard error = 0.048 day-1; p-value = 0.01313; 95% confidence interval 0.032 to 0.23 day-1

• C0,d = 0.053 mg/kg; 95% confidence interval 0.026 to 0.11 mg/kg.

These values are again lower than determined in the Egeler et al. (2013) analysis. It is important to note that, although the regressions are generally poor (as evidenced by the low R2 values), the slopes of the plots are significantly different from zero. The low R2-values probably reflect the scatter in the available data and the difficulties inherent in measuring small concentrations of TDM in fish.

The fact that the C0,d values derived by this method is, in both cases, lower than the C0,d value determined based on the measured concentration in fish at day 15 is suggestive, but not conclusive, that the fish concentrations measured at day 15 may have been influenced by the presence of undigested food in the fish and, as a


15 A p-value <0.05 signifies that there is a 95% probabili ty that the slope is dif ferent from zero.

40

consequence, the assimilation efficiency and overall depuration rate constant may have been overestimated in the original analysis (i.e. the depuration curve is assumed to start from a higher concentration than is the case in reality)14. Alternatively the difference in C0,d values could imply that the analysis carried out here may have underestimated the actual depuration rate constant.

ii) The assimilation efficiency has been re-calculated based on a C0,d concentration of either 0.0716 mg/kg or 0.053 mg/kg and assuming a feeding rate of 0.03 kg food/kg fish/day and a k2 value of either 0.104 day-1 or 0.133 day-1. In the original Egeler et al. (2013) analysis, the concentration in food was taken to be 103.3 mg/kg food dry weight for the calculations. As explained above, it would be preferable for the concentration in food to be expressed in terms of mg/kg food wet weight. The Egeler et al. (2013) report indicates that the dry weight of food was 96.3% of the wet weight. Therefore the equivalent concentration in a wet food basis would be 99.4 mg/kg wet weight and this value has been used in the calculations here. This leads to a calculated assimilation efficiency of around 0.003 (i.e. 0.3% in both cases). This is very low indicating that the substance is not readily taken up from food. For comparison, the assimilation efficiency for HCB was estimated to be around 51.9% based on measurements carried out on day 1 of the uptake.

Using the derived assimilation efficiency a kinetic BMFK (not lipid normalised or growth corrected) of either 0.00091 or 0.00062 can be estimated.

iii) The available fish weight data have also been reanalysed in order to determine the rate constant for growth dilution. Plots of ln [1/fish weight] versus time have been constructed for the control population, the exposed population and the combined population over the entire experimental period (data as reported in Table 12; there are insufficient data points to allow this to be done meaningfully for the uptake and depuration phases separately). The rate constants derived are summarised below, along with the standard error and p-value15 derived for the slope from the regression:

• Control population: kg = 0.00259 day-1; standard error = 0.00376 day-1; p = 0.506.

• Exposed population: kg = 0.00256 day-1; standard error = 0.00331 day-1; p = 0.444.

• Combined population: kg = 0.00249 day-1; standard error = 0.00248 day-1; p = 0.320.

Comparing the estimates obtained with that from the original Egeler et al. (2013) analysis, it can be seen that the kg values are smaller than estimated by Egeler et al. (2013) (kg = 0.00397 day-1) and that there is no significant difference between the exposed population and control population. Importantly for the overall interpretation of the study, the rate constant for growth dilution is small compared with the overall depuration rate constant (k2) and is in fact not statistically significant from zero. This means that the effects of growth dilution on the study

14 The eMS estimates that if around 14% of the applied dose was present in the gut at analysis this would be sufficient to result in an apparent concentration in the fish of 0.42 mg/kg.


41

can, in the view of the eMS, be effectively ignored, i.e. BMFKg ≈ BMFK = 0.00091 or 0.00062.

iv) The final stage in re-analysis of the data is to consider the lipid normalisation. For this, Egeler et al. (2013) assumed the lipid content of the food to be 18.6% on a wet weight basis and the lipid content of fish to be 10.5% on a wet weight basis (the mean level in the fish at the end of uptake). Using these lipid contents the lipid normalised BMFL (which is approximately the same as the growth-corrected and lipid normalised BMFLg) would be 0.0016 or 0.0011.

However, the lipid content of the fish increased during the study and the lipid content of the fish reached 16.7% on a wet weight basis in the control fish at the end of depuration. If the higher lipid content of 16.7% is used, then the BMFL

would be lower at 0.0010 or 0.00069.

Overall, the study shows that TDM has a low BMF (clearly below 0.01) indicating a very low potential for accumulation via dietary routes. The low BMF is a combination of a low assimilation efficiency and rapid depuration. The derivation of the overall depuration rate constant is dependent on the method used, possibly a result of the presence of undigested food in the gut of the fish at certain sampling points, and estimates of the k2g range between 0.104 day-1 and 0.6026 day-1. The registrant could update the robust study summary to take account of some of the further discussion of the data by the eMS as outlined above. Although a range of k2g values can be estimated from the available data (dependent upon how the data are analysed) in all cases the information shows rapid depuration of TDM was occurring in the study, which may be indicative of extensive metabolism of TDM.

The potential for metabolism of TDM in fish has been investigated by Davis and Erhardt (2009). The method used was an in vitro trout liver S9 metabolism assay following the approach outlined in Cowan-Ellsberry et al. (2008). The substance used in the test was a commercial sample with a stated purity of 99.9%. The test was carried out using rainbow trout hepatic S9 fraction. Replicate incubations were prepared containing approximately 2 mg/ml protein in 0.1 M Tris buffer, MgCl2, NADPH generating system16 and TDM (approximately 10 µM; prepared in ethanol) in a final volume of 1 ml. The system was incubated at 15°C for up to 2 hours and triplicate sub-samples were removed at 0, 30, 60, 90 and 120 minutes of incubation and analysed for the presence of TDM. Incubations using heat deactivated S9 samples were used as controls. The report indicates that care was taken to limit loss from volatilisation of TDM during incubation and sample processing.

No significant loss of TDM from the heat deactivated controls was evident. The average TDM concentration in the biologically active system (mean±standard deviation) was 8.86±2.63 µM, 9.47±2.83 µM, 4.71±2.23 µM, 7.33±1.42 µM and 6.15±0.39 µM after 0, 30, 60, 90 and 120 minutes incubation respectively. Assuming the loss followed first order kinetics, Davis and Erhardt (2009) estimated the first order rate constant for the loss of TDM from the test system as 0.0033 min-1.

16 Consisting of iso-citrate, NADP+ and iso-citrate dehydrogenase


42

Davis and Erhardt (2009), following an approach outlined in Cowan-Ellsberry et al. (2008), extrapolated the in vitro data obtained in this laboratory study to an equivalent in vivo, whole body metabolism rate constant for rainbow trout. Using a number of assumptions, the whole body metabolism rate constant (km) was estimated to be

0.112 day-1 assuming arterial hepatic blood flow alone in the extrapolation and 0.165 day-1 assuming both portal and arterial blood flow. These estimated values of km are consistent with the overall depuration rate constants obtained in the above feeding study (taking into account that the method used to estimate the km values was parameterised for rainbow trout and the fish species using the dietary study was zebrafish). Other estimates of the km can be obtained from the BCFBAFv3.01 model and values in the range 0.056 day-1 to 0.22 day-1 normalised to a 10 g fish have been estimated by the eMS (see Section 4.1.1). These are a similar order to the values obtained above by extrapolation of the in vitro data.

If the estimated km values are correct, then these suggest that metabolism is a significant, if not the predominant, depuration process as the km is likely to be close to the overall depuration rate constant obtained experimentally. However it should be noted that the km values available are all estimates or extrapolations and will be subject to unquantifiable uncertainties.

It is relevant to note that the overall depuration rate constant cannot be lower than the rate constant for metabolism. As other elimination process will also be occurring in the fish (particularly faecal egestion and respiratory elimination), it could be argued that the depuration rate constants of 0.104 day-1 and 0.133 day-1 obtained in the in vitro study with zebra fish are too low compared with the metabolism rate constant extrapolated from the in vitro data for rainbow trout. Although it is recognized that read-across of data from one species to another is difficult, it is relevant to note that it is also difficult to derive a precise depuration rate constant from the zebrafish study owing to the rapid depuration seen, and the analysis above suggests a value somewhere in the range 0.104 day-1 and 0.6026 day-1. To investigate this further, the eMS has estimated the depuration rate constant from the dietary study using the measured concentrations in fish on day 15 of uptake (taken as the start of depuration) along with the day 0.4, 7 and 14 of depuration data. Carrying out the analysis this way results in a k2 value of around 0.19 day-1 (95% confidence interval 0.085 to 0.29 day-1) when the ‘not detected’ values are omitted and around 0.20 day-1 (95% confidence interval 0.089 to 0.30 day-1) when the ‘not detected’ values are included as half the limit of quantification17. The plots for these are shown in figures 4 and 5. These values appear to be reasonably consistent with the metabolism rate constants extrapolated from the in vitro information on metabolism with rainbow trout. When considering this it is important to recognize that the metabolism rate constant derived for rainbow trout considers only one depuration process (other processes such as respiratory elimination and fecal elimination will also contribute to the overall depuration in rainbow trout) whereas the depuration rate constant for zebrafish includes all processes. Thus the available data suggest that the depuration of TDM is likely to be rapid, and of a similar order, in both zebrafish and rainbow trout. The significance of the relatively rapid depuration of TDM, most probably by metabolism, is considered below in relation to the likely BCF for TDM.

17 Estimated from the slope of a plot of ln [concentration of TDM in fish] versus time. The correlation coefficients (R2) of the plots were 0.42 and 0.35 respectively


43

Figure 4 Plot of ln [Concentration] versus time for the depuration phase

including the uptake day 15 values and using only concentration above the limit of quantification

Figure 5 Plot of ln [concentration] versus time for the depuration phase including the uptake day 15 values and using half of the limit of quantification for missing data points on days 0.2, 7 and 14

y = -0.1901x - 1.8655

R² = 0.4171

-5

-4.5

-4

-3.5

-3

-2.5

-2

-1.5

-1

-0.5

0

0 2 4 6 8 10 12 14 16

Ln [

Co

nce

ntr

ati

on

(m

g/k

g)]

Time (days)

y = -0.1972x - 2.2881

R² = 0.3525

-6

-5

-4

-3

-2

-1

0

0 2 4 6 8 10 12 14 16

Ln [

Co

nce

ntr

ati

on

(m

g/k

g)]

Time (days)


44

When considering the various kinetic parameters that are derived in the study it is relevant to note that the assimilation efficiency derived is a function of the C0 value (the concentration in fish at the start of the depuration phase), which can either be obtained directly from measurement or indirectly from the intercept of the plot of ln [concentration in fish] versus time that is used to determine the depuration rate constant. If this latter method is used the C0 value, and hence assimilation efficiency is related to the depuration rate constant in that a steeper slope (hence higher depuration rate constant) will lead to a higher assimilation efficiency (C0). For the current study the rapid depuration, and the occurrence of not detectable values in the depuration data set means that various values for the assimilation efficiency and depuration rate constant can be derived, depending on how the data are analysed. This is exemplified below for the various approaches used in the current evaluation. Method Co

(mg/kg) α k1

a (kg food/kg fish/day)

k2g (day-1) BMFL

As derived in the original test report

0.3649b 0.0714 0.00152 (or 0.0021)

0.6026 0.00635

Depuration rate constant from depuration phase only, omitting non-detects

0.0716c 0.003 0.00009 0.104 0.0016

Depuration rate constant from depuration phase only, assuming non-detects = detection limit/2

0.053c 0.003 0.00009 0.133 0.0011

Depuration rate constant from day 15 uptake onwards, omitting non-detects

0.15c 0.0098 0.00029 0.19 0.0027

Depuration rate constant from day 15 uptake onwards, assuming non-detects = detection limit/2

0.10c 0.0068 0.00020 0.20 0.0018

Note: a) The k1 value represents the rate constant for uptake from food and is estimated here as assimilation efficiency (α) × feeding rate (I). The feeding rate was 0.03 kg food/kg fish/day. See the text for a discussion of the k1 value derived in the original study report.

b) Based on the measured concentration in fish at the end of the uptake phase. c) Estimated from the intercept of the ln [concentration] versus time plots.


45

As can be seen from this analysis, there is a general trend in the assimilation efficiency increasing when the estimate of the depuration rate constant increases (this is for the reasons stated above). However it is important to note that although this results in some uncertainty over the precise value of the assimilation efficiency, when these two trends tend to cancel themselves out somewhat in the final biomagnification factor. Thus, although the estimates for the assimilation efficiency cover the range 0.003 to 0.07 (a factor of over 20) the derived BMFL values covers a much smaller range (0.0011 to 0.00635, a factor of 6), with most values in the range 0.0011 to 0.0027. Importantly all methods for analysing the data result in a BMFL value <0.01. A number of points have been raised on the available experimental bioaccumulation data for TDM by the ECHA PBT Expert Group and a detailed response to these comments has been produced by the evaluating Member State. Some of the technical issues raised are discussed below.

• Comment 1: The BMF value of HCB in the test is below 1, yet this substance is known to biomagnify and generally BMF values in the literature exceed 1. The value is also in contradiction of the test guideline validation criteria. The BMF depends on fish size so might this low BMF be affected by the small fish that were used? The concern raised here relates to the fact that the BMF reported in the test for the reference substance HCB was <1. The OECD test guideline does not itself give any validity criteria based on the results obtained for HCB but, as HCB is known to biomagnify (for example the mean lipid normalised and growth corrected BMF for HCB found in the OECD 305 ring test (OECD, 2012) at a 3% feeding rate was 3.10 (relative standard deviation 37%) in rainbow trout and 1.45 (relative standard deviation 14%) in carp), the information results for HCB warrant further investigation, both in terms of the BMF and the assimilation efficiency (see below).

It is important to note that the feeding study was carried out before the OECD 305 Test Guideline was finalised and, although the method follows closely the recommended approach for TDM, the accumulation of the reference substance (HCB) was only determined on day 1 and day 15 of the uptake phase. Thus the BMF determined for HCB in the study is not a steady state value but rather was determined by dividing the concentration in the fish measured on day 15 by the concentration in the food. The values for the BMF obtained using this approach are 0.134 (not lipid normalized) or 0.246 (lipid normalized and corrected for the dry weight of food). Although the lipid normalized value is still below 1 it is important to note that this is not a steady state value. No kinetics were derived for HCB during the depuration phase and so it is not possible to estimate a kinetic BMF for HCB. However the depuration of HCB would be expected to be slow, particularly as the fish were not growing or growing only slowly (meaning that growth dilution would only be a minor “loss” process for HCB in this study).

According to the OECD 305 test guideline the expected time to steady state can roughly be estimated by the following equation.

t95 = 3.0/k2

where t95 = time to 95% of steady state.

k2 = depuration rate constant.


46

Using this equation, it can be calculated that for 95% of steady state to be reached within 15 days, the depuration rate constant would need to be 0.2 day-1 or greater.

Unfortunately, as noted above, the depuration kinetics of HCB were not studied in this test. However we know from the validation report for the OECD 305 test guideline (OECD, 2012), that the mean (±relative standard deviation) growth-corrected depuration rate constant for HCB was 0.016 day-1 (±38%) in rainbow trout and 0.022 day-1 (±27%) in carp. So, for HCB to have reached steady state in the current study, the depuration rate constant in zebrafish would need to be an order of magnitude larger than was found in both rainbow trout and carp, which seems unrealistic. For illustration, a depuration rate constant of around 0.02 day-1 would result in an estimated time to 95% of steady state of 150 days. It is therefore considered likely that the true, steady state BMF for HCB in this study would be over 1 if a kinetic value could be calculated.

With regards to fish selection and fish size, it is relevant to note that the OECD 305 test guideline also says:.

Selection of fish species

“123. Fish species as specified for the aqueous exposure may be used (cf. paragraph 32 and Annex 3). Rainbow trout (Oncorhynchus mykiss), carp (Cyprinus carpio) and fathead minnow (Pimephales promelas) have been commonly used in dietary bioaccumulation studies with organic chemicals before the publication of this TG. The test species should have a feeding behaviour that results in rapid consumption of the administered food ration to ensure that any factor influencing the concentration of the test substance in food (e.g. leaching into the water and the possibility of aqueous exposure) is kept to a minimum. Fish within the recommended size/weight range (cf. Annex 3) should be used. Fish should not be so small as to hamper ease of analyses on an individual basis. Species tested during a life-stage with rapid growth can complicate data interpretation, and high growth rates can influence the calculation of assimilation efficiency“.

As zebrafish are listed in Annex 3 of the test guideline they can be used for the test. The recommended size is given in terms of fish length, rather than weight and for zebrafish the recommended fish length is 3.0 ± 0.5 cm. Furthermore, the guidance on fish selection and size given in the OECD 305 guideline relates mainly to possible analytical difficulties from using fish that are too small and complications that arise from use of rapidly growing fish, rather than to any perceived effect of smaller fish resulting in a smaller BMF. Thus the use of zebrafish, or small fish in general, is not precluded by the OECD 305 test guideline.

For the TDM study, the fish used were in the range 3.27 to 3.67 cm (mean 3.47 cm) 8 days prior to the start of the test. Thus the fish used are consistent with the requirements of the test guideline and are towards the upper end of the recommended fish size. The mean fish weight 8 days prior to the start of the test was 0.253 g (range 0.206-0.306 g). The individual fish weights measured during the course of the study were in the range 0.2149 to 0.747 g.

For comparison, the mean initial fish weights used in the various studies that


47

formed part of the OECD 305 ring test (OECD, 2012 and 2013) were 1.25, 8.41, 1.95, 1.17, 6.77, 0.72, 1.20, 1.24, 1.49 and 1.97 g for rainbow trout and 5.42 g in carp. So, although the fish size used in the TDM study were smaller than used in the OECD ring test, this was by a factor of only around 3 when compared with the smallest fish used in the ring test.

OECD (2012) investigated the possible relationship (if any) between initial fish weight and some of the derived bioaccumulation parameters. The comparison was complicated by the fact that apparent correlation between the mean fish lipid content and mean initial weight was found (evidence was found for increasing fish lipid content with increasing mean initial weight) and so it was not possible to distinguish unambiguously between factors that were dependent on the initial fish weight and factors that were dependent on the fish lipid content. In addition many of the apparent trends were not statistically significant. The general findings from this analysis are summarised below (taken from OECD, 2012):

o No clear trend was apparent in the growth rate constant in relation to the fish lipid content.

o There was a trend towards a decrease in the growth rate constant with an increase in the fish lipid content.

o There was a trend towards a decrease in the growth rate constant with an increase in the initial fish weight.

o There was a correlation between the food lipid content and the fish lipid content, particularly at the end of the study period.

o There was a trend towards a decrease in the overall depuration rate constant with increasing lipid content of the fish (or initial fish weight) for hexachlorobenzene, musk xylene, o-terphenyl and methoxychlor. This may relate, at least in part, to the apparent decrease in the growth rate constant component with increasing lipid content of the fish.

o No trend was apparent in the growth-corrected depuration rate constant with fish lipid content in the studies carried out using a 3% feeding rate (OECD, 2012), although a later analysis using the data for both the 3% and the 1.5% feeding rate studies found an inverse relationship (statistically significant for o-terphenyl and methoxychlor, not statistically significant for hexachlorobenzene and musk xylene) between the growth-corrected depuration rate constant and the fish lipid content.

o The assimilation efficiency shows a general increasing trend with increasing fish lipid content and initial fish weight.

o There was no discernible trend in the variation of the assimilation efficiency with food lipid content.

o The BMFg generally showed an increasing trend with increasing fish lipid content.

o No overall trend was evident between the BMFg and initial fish weight or food lipid content.


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o An apparent trend towards increasing BMFL with increasing fish lipid content was observed for hexachlorobenzene, musk xylene, o-terphenyl and methoxychlor.

o No overall trend was evident between the BMFL and initial fish weight.

o An apparent trend towards increasing BMFL with increasing food lipid was observed for all substances.

The effect of the differences between species in the OECD 305 ring test has been considered in more detail in Part II of OECD (2012). Here it was recognised that gill elimination rates, biotransformation rates and fish growth rates are a function of organism size, lipid content and the system temperature. For example transfer of chemicals across the fish gills (either uptake from water or respiratory elimination) is thought to be dependent on the ventilation rate of the fish, which may be higher, relative to fish weight, for smaller fish than larger fish. Similarly metabolic capacity may vary with fish size (age).

The analysis carried out in Part II of OECD (2012) found that, for the ring test, the assimilation efficiency and lipid- normalised and growth- corrected BMF were generally larger in trout than in carp and that depuration rate constants showed a tendency to decrease with increasing body size and lipid content (as noted above). It was also noted that the ring test data showed considerable inter-species variability and that a key contribution to this was probably from differences in the rate of biotransformation between the two species (carp versus rainbow trout). The evidence for the ring test suggested that, of the two species tested, carp had a greater capacity for biotransformation of some of the substances tested than rainbow trout but the report cautioned that the available historical data provide some conflicting evidence (where rainbow trout may have a greater capacity for metabolism than carp). Overall it was concluded that the natural variability and individual differences and metabolic properties and lipid contents probably had more impact on the inter-species variability seen than differences in body size and temperature. Furthermore, the report concluded that many of these factors may be chemical-specific (as well as species- and condition-specific) and it is not possible to generalise over the possible implications of selection of one species over another in the test.

It is also important to remember that when the BMF data are converted to the equivalent BCF (see below), the estimate of the uptake rate constant from water (k1) is also dependent on the fish size in many of the methods available, and increases with decreasing fish size. Thus, even if the k2 value is higher in smaller fish than larger fish, this trend will be balanced by a corresponding increase in the k1 value for smaller fish than larger fish. One final point that also needs to be considered is that although zebrafish are small fish, the lipid content of the fish used in the TDM study was relatively high (around 10.4% at the end of the uptake period and 16.7% at the end of depuration). As it is thought that the growth corrected depuration rate constant is inversely related to lipid for some substances (see OECD, 2013), this means that depuration may have been slower in the zebrafish than would be expected in fish with lower lipid contents. The REACH guidance recommends that where


49

possible data are normalised to a 5% lipid content when considering the BCF. Overall, although it is agreed that it is possible that both fish size and temperature may affect the bioaccumulation of TDM (and substances in general), the effects of these should be considered within the overall uncertainties in measuring the BMF (or BCF). In particular, inter-species differences in metabolic capacity will be equally (if not more) important. It is not possible, or ethical, to test substances in a range of species unless there is specific evidence that supports this. For TDM there is evidence from the dietary study with zebrafish that the substance may be metabolised and this is supported by evidence from the trout liver S9 study that the metabolism is also likely in another species. Therefore further bioaccumulation testing in other species is not warranted for TDM.

• Comment 2. An assimilation value of 52% is considered by some experts to be

“not that high” for HCB, and they would “expect values of 80-90%”. Therefore they claim the value does not necessarily show that food spiking was adequate particularly given the very low assimilation for TDM.

The assimilation efficiency obtained in the study for HCB of 52% is actually consistent with the available experimental data for HCB. The reasons for this are outlined below.

Firstly, it should be noted that the assimilation efficiency was estimated for HCB using the data from day 1 of uptake only and assumes that the uptake is still in the linear phase. The reason for this is that the depuration of HCB was not studied in the test. The approach used is consistent with the method in the OECD 305 test guideline and outlined below, and is considered appropriate for slowly depurating substances. However, it is possible that this approach will slightly underestimate the actual assimilation efficiency for the reason discussed below.

The method used for HCB in the TDM test report essentially calculates the assimilation efficiency for HCB based on the ratio of the amount of HCB determined in the fish at day 1 to the amount of HCB given to the fish in the first feeding period. Thus this reflects the percentage of the applied dose that was absorbed by the fish. An important assumption in estimating the assimilation efficiency in this way is that there is no significant elimination of HCB from the fish over the timeframe of the calculation. This is likely to be a reasonable assumption for HCB although it is possible that even for slowly depurating substances some elimination will still occur leading to a (slight) underestimation of the assimilation efficiency.

The OECD 305 test guideline indicates that assimilation efficiency can be estimated from the linear phase of uptake for the fish while still near the beginning of the exposure period using the following equation.

α = Cfish (t) /(Cfood × I × t)

Where Cfish(t) = the concentration of test substance in the fish at time t.

Cfood = concentration in food.

I = feeding rate.

t = time.


50

The calculation for HCB was carried out in a way that is consistent with the above equation and took into account the following:

Mean concentration in food = 1,587,470 dpm (disintegrations per minute)/g (dry weight).

Amount of food per tank per day = 0.1781 g/day (day 0 to day 1).

Number of fish per tank: 15.

Amount of food per fish per day: 0.0119 g/fish/day (day 0 to day 1).

Amount of 14C-HCB fed per fish per day: = 18,850 dpm/fish/day (day 0 to day 1).

The amount of food per tank per day of 0.178 g/day corresponds to a feeding rate of 0.03 wt/wt for 15 fish of mean initial weight 0.4 g (the mean fish weight of the control fish on day 0 was 0.41 g).

The assimilation efficiency can then be estimated from the ratio of the amount of 14C-HCB measured in the fish on day 1 (in units of dpm/fish; this can be estimated by multiplying the individual fish weight (g) by the concentration measured in fish (in units of dpm/g)) and the amount of 14C-HCB fed to fish between day 0 and day 1 (18,850 dpm/fish/day). The assimilation efficiencies calculated this way are given in table 13. This results in a mean assimilation efficiency (±standard deviation) of 0.52±0.15 (or expressed as a percentage 52%±15%).

Table 13 Concentrations of 14C-HCB.

Update day

Replicate Fish weight (g)

Concentration in fish (dpm/g wet fish)

BMF at day 15 (based on dry weight food concentration)

Lipid normalised BMF at day 15 (based on wet weight food concentration)

Assimilation efficiency at day 1

1 A 0.475 13,548 / / 0.34

B 0.359 36,520 / / 0.70

C 0.516 17,120 / / 0.47

D 0.150 71,593 / / 0.57

Mean (±standard deviation)

0.375± 0.164

34,659±26,589 / / 0.52±0.15

15 A 0.513 282,861 0.178 0.328 /

B 0.396 111,482 0.070 0.129 /


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C 0.511 249,149 0.157 0.289 /

D 0.359 205,725 0.130 0.238 /

Mean (±standard deviation)

0.445± 0.079

212,304±74,261

0.134±0.047

0.246±0.086 /

For comparison, using the equation above from the OECD 305 guideline, and assuming a feeding rate of 0.03, a mean concentration in fish of 34,659 dpm/g on day 1 and a mean concentration in food of 1,587,470 dpm/g (dry weight) the assimilation efficiency can be estimated to be 0.73 (73%). It should be noted, however, that one individual value obtained using this method is above 1 (the individual values for the four replicates are 0.28, 0.77, 0.36 and 1.50).

It is not appropriate to estimate the assimilation efficiencies from the day 15 data using this approach owing to the requirement/assumption that the uptake should still be in the linear phase.

The assimilation efficiency determined for HCB in the current TDM study should also be considered in relation to the data available for HCB from other studies. For this, a list of studies present in REACH registration dossiers has been made available by EChA, and the available data taken from the publicly available dossiers available on the ECHA website. Unfortunately, for many of the studies there was no information on HCB. However, the assimilation efficiencies were reported for HCB in two of the studies with rainbow trout as 47% and 50.7%, which is a similar order as (and indeed slightly lower than) the assimilation efficiency for HCB of 52%±15% measured in the TDM study.

The OECD ring test validation report (OECD, 2012 and 2013) also contains data for the assimilation efficiency of HCB. For the studies carried out using a feeding rate of 0.03 wt/wt the mean (±standard deviation) of the assimilation efficiency determined for HCB was 0.60±0.17 (i.e. 60%±17%) in rainbow trout and 0.38±0.04 (i.e. 38%±4%) in carp. The mean (±standard deviation) obtained in the studies using a lower feeding rate of 0.015 wt/wt was 84±23%. The value for HCB of 52%±15% measured in the TDM study is in agreement with the values obtained in the studies using a feeding rate of 0.03 wt/wt (as was used in the study with TDM) but is slightly lower (but not statistically significantly so as the standard deviation ranges overlap) than obtained in the studies using a lower feeding rate.

Further data on the assimilation efficiency of HCB are given in the published literature [note: an in-depth search of the literature to find all available data has not been undertaken and the various studies have not been evaluated for their reliability for this response to comments document].

o Fisk et al. (1998) reported assimilation efficiencies for HCB of 50%±7% and 34% in rainbow trout (Oncorhynchus mykiss) for HCB concentrations of 14 ng/g and 103 ng/g, respectively. Fisk et al. (1998) remarked that these values were similar to those obtained in goldfish (Gobas et al., 1993) (see below) but were lower than values obtained by Niimi and Oliver (1988) in rainbow trout (assimilation efficiency of 73–88%). Fisk et al. (1998) thought that this latter difference may be a result of differences in digestability of the food


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used as the trout in the Niimi and Oliver (1988) study were exposed to HCB in herring oil by gavage and this may have resulted in higher assimilation than would be expected from food.

o Gobas et al. (1993) determined assimilation efficiencies of 50%, 46% and 46% for HCB in goldfish (Carassius auratus) using diets with lipid contents of <0.2%, 6.3% and 13.5%, respectively.

o Inuoue et al. (2012) reported the assimilation efficiency for HCB in carp to be 86, 76, 78, 73 and 75% in five studies where the fish were exposed to a single test substance and HCB in the diet and 49% in a study where the fish were exposed to a mixture of three test substances and HCB in diet. The mean (±standard deviation) of these determinations is 73%±13%.

Therefore the assimilation efficiency for HCB measured in zebrafish in the TDM study is comparable with that measured in rainbow trout, carp and goldfish in other studies. Indeed the available evidence from dietary studies does not appear to support the assumption that the assimilation efficiency for HCB would be expected to be in the range 80-90% as stated in the comment.

It is also relevant to consider the food spiking method requirements specified in the OECD 305 test guideline. The food used in the TDM test was a commercially available fish food (ground fish food flakes) and the substance was added to the fish food as a solution in cod liver oil (0.5 ml was added to 5 g of food, i.e. 0.1 ml oil/g food; a similar amount of oil was added to the control diet) and analysis showed that the concentration in food was homogeneous (four samples were analysed in duplicate prior to the start of the test and the concentrations were in the range 87 to 112 mg/kg dry weight). The lipid content of the food was 18.6%.

This is in general agreement with the OECD 305 test guideline which recommends a commercially available fish food with a uniform pellet size (with the size being appropriate to the size of the fish) and a lipid content in the range 15-20%. Spiking using a solution in fish oil is one of the recommended methods for addition of the test substance to the food.

Overall, the HCB data from the TDM study are consistent with the data available from other feeding studies.

Comment 3: The k2 values derived in the test for TDM are considered by some experts to be “not that high” particularly given the small fish size.

As indicated above the evidence for the relationship of k2 and fish size is limited.

It is important to remember that when the BMF data are converted to the equivalent BCF, the estimate of the uptake rate constant from water (k1) is also dependent on the fish size in many of the methods available (see below), and increases with decreasing fish size. So, even if the k2 value is higher in smaller fish than larger fish, this trend will be balanced by a corresponding increase in the k1 value for smaller fish than larger fish.

A previous analysis on available k2 data (from all sources, not just dietary studies) in order to evaluate if the k2 value alone could be used as an indicator of a high bioaccumulation potential has been carried out. This is available in an Environment Agency report (Environment Agency, 2012). This analysis concluded that the critical value of k2 for identification of a substance with a BCF >2,000 l/kg is k2 ≤ 0.178 day-1; substances with k2 values above this critical


53

value have a low probability of possessing BCF values of 2,000 l/kg or above (a similar analysis for a BCF of 5,000 l/kg identified a critical k2 ≤ 0.085 day-1). When normalisation of the data to a standard lipid content of 5% was considered, the critical k2 for a BCFlipid normalised of >2,000 l/kg was determined to be k2 ≤ 0.181 day-1.

In this context the depuration rate constant for TDM was around 0.20 day-1. This was determined in fish with lipid contents above 5% (10.5% at the end of uptake and 16.7% at the end of depuration). The equivalent value normalised to a lipid content of 5% would be significantly higher (approximately 0.42-0.67 day-1 depending on which lipid content is used); however lipid normalisation of the depuration rate constant may not be appropriate for substances for which metabolism is a significant depuration process. In all cases the depuration rate constant obtained for TDM is above the critical value identified in the Environment Agency (2012) analysis.

Therefore the available evidence suggests that the depuration rate constant for TDM is consistent with other substances that have a low bioaccumulation potential within the meaning of PBT/vPvB assessment.

Inoue et al. (2012) carried out a comparison between lipid normalised BMF and lipid normalised (to 5% lipid) BCF for several poorly water soluble substances in carp (Cyprinus carpio). This analysis found that a lipid normalised BMF of 0.31 (95% confidence interval 0.11-0.87) corresponded to a lipid normalised BCF of 5,000 l/kg. The available data for TDM suggests that the BMF for TDM is well below this value.

Estimating a BCF

In order to estimate a BCF from the data obtained in the dietary accumulation study the REACH guidance recommends that an estimate for the rate constant for uptake from water is obtained using the approach outlined by Sijm et al. (1995). This can then be combined with the growth corrected depuration rate constant (for TDM the growth correction is minimal) to obtain an estimate of the kinetic BCF value:

BCF = kuptake/k2g = 520×W-0.32/k2g

where kuptake = uptake rate constant (l/kg/day).

k2g = growth corrected depuration rate constant (day-1). For TDM this can be approximated to the overall depuration rate constant k2.

W = weight of fish (g wet weight).

Crookes and Brooke (2011) and Brooke et al. (2012) identified several other methods for estimating the uptake rate constant (mainly from fish weight but some methods also take into account other factors such as the log Kow – see Crookes and Brooke (2011), including:

Barber, 2003 observed Hayton and Barron, 1990

Barber, 2003 calibrated Omega/Hendriks, 2001

BASS/Barber, 2001 QEAFDCHN/Thomann, 1989

Erickson and McKim, 1990a Spacie and Hamelink, 1982


54

Erickson and McKim, 1990b Streit and Sire, 1993

FGETS/Barber et al., 1991 Tolls and Sijm, 1995

These methods have been used to estimate the kuptake and, when combined with the k2g, the growth-corrected BCF from the available feeding study data. The relevant data from are shown in Table 14 (estimation of the kuptake) and Table 15 (estimation of the growth-corrected BCF) for a number of different assumptions over the fish weight, overall depuration rate constant and lipid contents.

As can be seen from Table 15, the BCF obtained depends crucially on the assumption made over the growth-corrected depuration rate constant. The predicted BCFs above 5,000 l/kg are shaded in red, those between 2,000 l/kg and 5,000 l/kg are shaded in light orange and those below 2,000 l/kg are unshaded.


44

Table 14: Estimation of uptake rate constant values from dietary accumulation studies with fish

Identifier Log Kow

Fish weight assumed

(g)a

Assumed k2g (day-1)b

Fish lipid content

(%) c

Estimated uptake rate constant (l/kg/day)e

1 2d 3 4 5 6 7 8 9 10 11 12d 13d

a 6.07 0.41 0.2 8.36 692 503 1000 800 801 774 698 600 470 530 970 745 856

b 6.07 0.41 0.2 10.54 692 503 1000 800 801 774 698 600 470 530 970 745 856

c 6.07 0.41 0.2 16.7 692 503 1000 800 801 774 698 600 470 530 970 745 856

d 6.07 0.41 0.1 8.36 692 503 1000 800 801 774 698 600 470 530 970 745 856

e 6.07 0.41 0.1 10.54 692 503 1000 800 801 774 698 600 470 530 970 745 856

f 6.07 0.41 0.1 16.7 692 503 1000 800 801 774 698 600 470 530 970 745 856

g 6.07 0.41 0.61 8.36 692 503 1000 800 801 774 698 600 470 530 970 745 856

h 6.07 0.41 0.61 10.54 692 503 1000 800 801 774 698 600 470 530 970 745 856

i 6.07 0.41 0.61 16.7 692 503 1000 800 801 774 698 600 470 530 970 745 856

j 6.07 0.47 0.2 8.36 662 486 966 783 781 757 677 584 460 516 917 745 856

k 6.07 0.47 0.2 10.54 662 486 966 783 781 757 677 584 460 516 917 745 856

l 6.07 0.47 0.2 16.7 662 486 966 783 781 757 677 584 460 516 917 745 856

m 6.07 0.47 0.1 8.36 662 486 966 783 781 757 677 584 460 516 917 745 856

n 6.07 0.47 0.1 10.54 662 486 966 783 781 757 677 584 460 516 917 745 856

o 6.07 0.47 0.1 16.7 662 486 966 783 781 757 677 584 460 516 917 745 856

p 6.07 0.47 0.61 8.36 662 486 966 783 781 757 677 584 460 516 917 745 856

q 6.07 0.47 0.61 10.54 662 486 966 783 781 757 677 584 460 516 917 745 856

r 6.07 0.47 0.61 16.7 662 486 966 783 781 757 677 584 460 516 917 745 856

Note: a) 0.41g is the mean fish weight measured at the start of uptake and 0.47 g reflects the fish weight at the end of depuration in the dietary accumulation study. Thus the estimates reflect the range of weights throughout the study period. b) The k2g reflecting the range found from different ways of analysis of the data (see text). For TDM the effects of growth dilution can e ffectively be ignored in the dietary study. c) The values represent the range of values obtained in the dietary accumulation study (see text). d) These estimates depend on the log Kow. When a log Kow of 7.43 is used, the value of the estimated rate constants increases slightly to 506 (1), 1,181 (12) or 1,254 l/kg/day (13) for the 0.41 g fish and 489 (1) 1,181 (12) or 1,254 (13 l/kg/day) for the 0.47 g fish. e) Uptake rate constant estimated using the following methods.


45

1 - Sijm et al., 1995 (method given in the REACH Guidance) 2 - Omega/Hendriks, 2001 3 - QEAFDCHN/Thomann, 1989 4 - BASS/Barber, 2001 5 - FGETS/Barber et al., 1991 6 - Erickson and McKim, 1990a 7 - Erickson and McKim, 1990b 8 - Hayton and Barron, 1990 9 - Streit and Sire, 1993 10 - Barber, 2003 observed 11 - Barber, 2003 calibrated 12 - Spacie and Hamelink, 1982 13 - Tolls and Sijm, 1995


46

Table 15: Estimation of growth corrected BCF values from dietary accumulation studies with fish

Identifier a Estimated growth corrected BCF (l/kg)c – not lipid normalised 1 2b 3 4 5 6 7 8 9 10 11 12b 13b

a 3,458 2,515 4,999 4,000 4,003 3,868 3,490 2,999 2,348 2,652 4,849 3,726 [5,904]

b 3,458 2,515 4,999 4,000 4,003 3,868 3,490 2,999 2,348 2,652 4,849 3,726 [5,904]

c 3,458 2,515 4,999 4,000 4,003 3,868 3,490 2,999 2,348 2,652 4,849 3,726 [5,904]

d 6,917 5,029 9,998 7,999 8,005 7,735 6,980 5,999 4,696 5,304 9,697 7,452 [11,809]

e 6,917 5,029 9,998 7,999 8,005 7,735 6,980 5,999 4,696 5,304 9,697 7,452 [11,809]

f 6,917 5,029 9,998 7,999 8,005 7,735 6,980 5,999 4,696 5,304 9,697 7,452 [11,809]

g 1,134 824 1,639 1,311 1,312 1,268 1,144 983 770 870 1,590 1,222 [1,936]

h 1,134 824 1,639 1,311 1,312 1,268 1,144 983 770 870 1,590 1,222 [1,936]

i 1,134 824 1,639 1,311 1,312 1,268 1,144 983 770 870 1,590 1,222 [1,936]

j 3,311 2,430 4,831 3,913 3,904 3,786 3,383 2,920 2,298 2,582 4,585 3,726 [5,904]

k 3,311 2,430 4,831 3,913 3,904 3,786 3,383 2,920 2,298 2,582 4,585 3,726 [5,904]

l 3,311 2,430 4,831 3,913 3,904 3,786 3,383 2,920 2,298 2,582 4,585 3,726 [5,904]

m 6,621 4,860 9,662 7,826 7,809 7,571 6,766 5,840 4,596 5,163 9,169 7,452 [11,809]

n 6,621 4,860 9,662 7,826 7,809 7,571 6,766 5,840 4,596 5,163 9,169 7,452 [11,809]

o 6,621 4,860 9,662 7,826 7,809 7,571 6,766 5,840 4,596 5,163 9,169 7,452 [11,809]

4,281 [6,272] 4,281

[6,272] 4,281

[6,272] 8,561

[12,545] 8,561

[12,545] 8,561

[12,545] 1,403

[2,057] 1,403

[2,057] 1,403

[2,057] 4,281

[6,272] 4,281

[6,272] 4,281

[6,272] 8,561

[12,545] 8,561

[12,545] 8,561

[12,545]


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Identifier a Estimated growth corrected BCF (l/kg)c – not lipid normalised p 1,085 797 1,584 1,283 1,280 1,241 1,109 957 753 846 1,503 1,222

[1,936] q 1,085 797 1,584 1,283 1,280 1,241 1,109 957 753 846 1,503 1,222

[1,936] r 1,085 797 1,584 1,283 1,280 1,241 1,109 957 753 846 1,503 1,222

[1,936]

1,403

[2,057] 1,403

[2,057] 1,403

[2,057] Identifier a Estimated growth corrected BCF (l/kg)c –normalised to a standard 5% lipid content

1 3 4 5 6 7 8 9 10 11

a 2,068 1,504 2,990 2,392 2,394 2,313 2,087 1,794 1,404 1,586 2,900 2,229

[3,531] 2,560

[3,751] b 1,641 1,193 2,371 1,897 1,899 1,835 1,656 1,423 1,114 1,258 2,300 1,768 2,031

[2,801] c 1,035 753 1,497 1,198 1,198 1,158 1,045 898 703 794 1,452 1,116

[1,768] d 4,137 3,008 5,979 4,784 4,788 4,626 4,175 3,588 2,809 3,172 5,800 4,457

[7,063] e 3,281 2,386 4,743 3,795 3,798 3,670 3,311 2,846 2,228 2,516 4,600 3,535

[5,602] f 2,071 1,506 2,993 2,395 2,397 2,316 2,090 1,796 1,406 1,588 2,903 2,231

[3,536] g 678 493 980 784 785 758 684 588 460 520 951 731

[1,158] h 538 391 777 622 623 602 543 467 365 412 754 580

[918] i 339 247 491 393 393 380 343 294 230 260 476 366

[580] j 1,980 1,453 2,889 2,340 2,335 2,264 2,023 1,747 1,374 1,544 2,742 2,229

[3,531]

[2975] 1,282

[1,878] 5,120

[7,503] 4,061

[5,951] 2,563

[3,756] 839

[1,230] 666

[976] 420

[616] 2,560

[3,751] k 1,570 1,153 2,292 1,856 1,852 1,796 1,605 1,385 1,090 1,225 2,175 1,768 2,031

[2,801] [2,975]


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Identifier a Estimated growth corrected BCF (l/kg)c – normalised to a standard 5% lipid content l 991 728 1,446 1,171 1,169 1,133 1,013 874 688 773 1,373 1,116

[1,768] m 3,960 2,907 5,779 4,680 4,670 4,528 4,047 3,493 2,749 3,088 5,484 4,457

[7,063] n 3,141 2,306 4,583 3,712 3,704 3,592 3,210 2,771 2,180 2,449 4,350 3,535

[5,602] o 1,982 1,455 2,893 2,343 2,338 2,267 2,026 1,749 1,376 1,546 2,745 2,231

[3,536] p 649 477 947 767 766 742 663 573 451 506 899 731

[1,158] q 515 378 751 609 607 589 526 454 357 402 713 580

[918] r 325 239 474 384 383 372 332 287 226 253 450 366

[580]

1,282

[1,878] 5,120

[7,503] 4,061

[5,951] 2,563

[3,756] 839

[1,230] 666

[976] 420

[616]

Note: a) See Table 14 for key to values used b) The estimates for methods 2, 12 and 13 depend on the log Kow assumed. The main values given in the Table are for a log Kow value of 6.07. When a log Kow of 7.43 is used, the estimates for method 2 are essentially the same as those given. The estimates for methods 12 and 13 using a log Kow of 7. 43 are shown in [ ]. c) Using the uptake rate constant estimated using the following methods: 1 - Sijm et al., 1995 (method given in the REACH Guidance) 2 - Omega/Hendriks, 2001 3 - QEAFDCHN/Thomann, 1989 4 - BASS/Barber, 2001 5 - FGETS/Barber et al., 1991 6 - Erickson and McKim, 1990a 7 - Erickson and McKim, 1990b 8 - Hayton and Barron, 1990 9 - Streit and Sire, 1993 10 - Barber, 2003 observed 11 - Barber, 2003 calibrated 12 - Spacie and Hamelink, 1982 13 - Tolls and Sijm, 1995


49

There is insufficient experience and knowledge to prefer one estimate over another. When the BCF values that have not been lipid normalised are examined (top half of Table 15), it can be seen that all but one of the methods gives predicted BCF values that are below 2,000 l/kg when a depuration rate constant of 0.61 day-1 is used (as reported in the original test report of the dietary study) (method 13 results in a BCF >2,000 l/kg only when a log Kow of 7.43 is assumed). When an overall depuration rate constant of around 0.2 day-1 is assumed, the vast majority of the predicted BCF values are greater than 2,000 l/kg but below 5,000 l/kg (only predictions using methods 12 and 13 and a log Kow of 7.43 result in BCF values >5,000 l/kg). Finally when an overall depuration rate constant of around 0.1 day-1 is used (the approximate lower limit obtained from re-analysis of the available depuration data), the majority of the predicted BCF values are above 5,000 l/kg.

The REACH guidance document recommends that, where possible, all the BCF values should be normalised to a standard lipid content of 5%. For TDM this is an important consideration as the lipid content of the fish used in the dietary study was significantly higher than 5%. The effect of lipid normalisation in this case is to reduce the BCF. The predicted lipid normalised BCF values are shown in the lower half of Table 15.

When lipid normalisation is taken into account, almost all of the predicted BCF values are below 5,000 l/kg, the few exceptions being a limited number of estimates using the lowest depuration rate constant. Taken overall, these estimates provide a strong weight of evidence to conclude that the lipid normalised and growth-corrected BCF for TDM is below 5,000 l/kg.

Again, when the lipid normalised BCF is considered, the BCFs obtained using the highest estimate of the depuration rate constant are all well below 2,000 l/kg. However, the situation with the lipid normalised BCF obtained using an estimate for the depuration rate constant of 0.2 day-1 is mixed, with a significant number of predictions being above 2,000 l/kg but also a significant number of predictions being below 2,000 l/kg. When the depuration rate constant is assumed to be 0.1 day-1, the majority of the predictions are above 2,000 l/kg but some predictions are below 2,000 l/kg.

Based on these predictions, it is clear that the bioaccumulation assessment is critically dependent on the choice of depuration rate constant. As discussed above, there are various different methods for estimating the depuration rate constant and these lead to estimates in the approximate range 0.1 to 0.61 day-1 for the same data set. When deciding on the most appropriate value for the overall depuration rate constant it is also relevant to consider the available data on metabolism, since the overall depuration rate constant cannot be smaller than the metabolism rate constant. As indicated above, an estimate for the metabolism rate constant for trout of around 0.112 day-1 assuming arterial hepatic blood flow alone and 0.165 day-1 assuming both portal and arterial blood flow have been obtained by extrapolation of in vitro data. Assuming that these metabolism rate constants are also applicable to zebrafish, it is probable that the actual depuration rate constant from the dietary study would not be as low as 0.1 day-1 and would probably be higher (e.g. 0.2 day-1 and above).

When considering the uncertainty in the depuration rate constant, it is relevant to note that much of this is likely to result from the fact that only limited uptake was seen in the dietary study. The levels in the fish were generally low compared to the detection


50

limit of the analytical method, making it difficult to determine the relatively small changes in concentration occurring during the depuration. Thus, although the data clearly show that the substance was rapidly depurated, and has a low biomagnification factor, it means that there is some uncertainty in the precise value of the depuration rate constant for the study.

It should also be noted that the BCF estimates assume that the uptake rate constant for TDM can be reliably predicted. A very low assimilation efficiency was found in the dietary study. The assimilation efficiency for TDM was determined to be around 7.1% in the original study; reanalysis of the data by the eMS indicates that the value could be lower than this, at around 0.3%. For comparison, the assimilation efficiency for HCB was estimated to be around 51.9% based on measurements carried out on day 1 of the uptake phase. There are two possible explanations for this. Firstly, TDM might not be effectively taken up through the gut wall, for example due to size limitations. Secondly, TDM might be rapidly removed from the gut (e.g. due to metabolism by gut microflora) before uptake through the gut wall can occur. It is important to distinguish between these two possibilities since, if the first explanation applies, it is possible that the uptake across the gills could be limited (resulting in a much smaller uptake rate constant than predicted, and hence much smaller BCF value than predicted). The second possibility would not necessarily affect uptake across the gills and so the predicted BCFs may be more relevant. Unfortunately it is not possible to distinguish between these two possibilities from the available data. However, given that the substance has only a moderate molecular weight (202.4 g/mole), it would appear unlikely that size limitations would apply to the uptake across the gut (although this does not rule out the possibility of other, as yet unidentified, limitations) and so, in the absence of other data, it has to be assumed that the methods used to predict the uptake rate constant would also be applicable to TDM

A paper by Comber and Thomas (2013) (not included in the registration dossier but provided to the eMS to accompany the dietary test) has also evaluated the available data in terms of estimation of a BCF. Comber and Thomas (2013) extrapolated the results of the dietary accumulation study to a BCF value using the following methods to estimate the uptake rate constant from water: Sijm et al. (1995), Omega/Hendriks (2001), Barber (2003) observed, Erickson and McKim (1990a), Spacie and Hamelink (1982) and Tolls and Sijm (1995). The k2g value was estimated based on the day 0 to 7 of depuration data (including the ‘not detected’ values as half the limit of quantification) and also the day 0 to 14 data (again including the ‘not detected’ values as half the limit of quantification). This yielded growth corrected depuration rate constants of 0.299 day-1 and 0.159 day-1, which are in reasonable agreement with the re-analysis of the data carried out by the eMS18. Using these along with the estimated uptake rate constants leads to lipid normalised BCFs of 1,541 to 3,645 l/kg (mean of the six methods was 2,578 l/kg) based on the day 0-14 estimate of the depuration rate constant and 819 to 1,937 l/kg (mean of the six methods 1,370 l/kg) based on the day 0-7 estimates of the depuration rate constant. These estimates are consistent with the estimates carried out by the eMS for the purposes of this evaluation. Comber and

18 The values obtained by Comber and Thomas (2013) appear to have be determined using the mean fish concentration at each time point.


51

Thomas (2013) indicated that there were some concerns with the depuration rate constant of 0.159 day-1 as the data were skewed by the variability in the two data points for which values were obtained and the poor fit with first order kinetics. When these estimates are removed, Comber and Thomas (2013) concluded that the weight of evidence was that the BCF for TDM was below 2,000 l/kg.

Using the structures that Industry included in their comments (see section 4.1.1) the eMS has obtained estimates for the metabolism rate constants from BCFBAF, and these are provided in Table 16 below. The BCFBAF predicts that all constituents will be metabolised. The rate constants for the C12 constituents are similar to, but a little lower than would be expected based on the experimental data (appreciating this relies on reading across from species to species). The metabolic rate constant is predicted to decrease with increasing carbon chain length. This means that all constituents with carbon chain lengths less than C12 would be expected to metabolise faster than the C12 constituents. It should be noted that the BCF predictions obtained using the BCFBAF model already include the metabolic rate constant predicted by the programme. However as shown in the previous discussion if the overall depuration rate constant is above around 0.2 day-1 it is unlikely that the BCF will be above 2,000 l/kg based on the estimated uptake rate constants from the various methods that are available and the same considerations would also apply to the C9, C10 and C11 constituents (as most of the methods for prediction of the uptake rate constant, e.g. Sijm et al 1995, are only dependent on fish size and not substance-related parameters).

The C13, C14 and C15 constituents are expected to metabolise at a slower rate than the C12 constituents. However, the BCFBAF program predicts that the BCF for these substances would be lower than the C12 constituents. The reason for this is most probably related to a reduced bioavailability as the BCFBAF program takes account of the adsorption of the substance onto suspended and dissolved organic matter and the BCFs calculated by the program are expressed in terms of total concentrations. If the BCFs for these constituents (and indeed all of the constituents) were expressed in terms of the freely dissolved concentrations the predicted BCFs would be higher than reported in the Industry paper, with the highest increase occurring for the substances with the higher log Kow values i.e. the C13, C14 and C15 constituents. However when considering this latter point it is important to note that in the feeding study the C12 constituents showed a low level of uptake (i.e. a low bioavailability from diet as evidenced by the low assimilation efficiency). Therefore although there is no data with which to read-across to the C13, C14 and C15 constituents, based on the decreasing predicted BCF values it would be expected that these constituents would show a similarly low (or even lower) assimilation from diet and so would also be expected to have very low BMFs.

There are other complications that can affect the depuration rate constant and predicted BCF e.g. fish lipids. The fish assumed in the BCFBAF model generally have lipid contents higher than the 5% value usually assumed in the REACH guidance (the lipid content is understood to increase with increasing trophic level within the model).

Overall, the predictions provide some support that the C9, C10 and C11 constituents will be metabolised at faster rates than the C12 constituents and that the accumulation of the higher chain length constituents is likely to be limited by reduced bioavailability. While it is appreciated that this does not categorically prove that the BCF for these other constituents is less than 2,000 l/kg, it is felt that there is no evidence to suggest that

conclusions made from the BMF value obtained for the C12 consitutuents cannot be applied to the C9-C11 and C13-C15 constituents.


Table 16: Estimates of rate constant for metabolism estimated for constituents of TDM using BCFBAF v3.01

Carbon chain length

Branching levela

SMILES a log Kow

Metabolism rate constant - km (day-1)

10 g fish 100 g fish 1 kg fish 10 kg fish

C9 1 CCCCCCC(C)(C)S 5.813 0.257 0.145 0.081 0.046

C10 1 CCCCCCCC(C)(C)S 6.463 0.167 0.094 0.053 0.030

C11 1 CCCCCCCCC(C)(C)S 6.963 0.120 0.068 0.038 0.021

C12 1 CCCCCCCCCC(C)(C)S 7.563 0.081 0.046 0.026 0.014

C13 1 CCCCCCCCCCC(C)(C)S 8.173 0.054 0.030 0.017 0.0096

C14 1 CCCCCCCCCCCC(C)(C)S 8.723 0.038 0.021 0.012 0.0067

C15 1 CCCCCCCCCCCCC(C)(C)S 9.283 0.026 0.015 0.0082 0.0046

C9 2 SC(CC)(CCC)CCC 5.773 0.286 0.161 0.091 0.051

C10 2 SC(CCC)(CCC)CCC 6.343 0.197 0.111 0.062 0.035

C11 2 SC(CCC)(CCCC)CCC 6.897 0.126 0.071 0.040 0.022

C12 2 SC(CCC)(CCCC)CCCC 7.543 0.076 0.043 0.024 0.013

C13 2 SC(CCCC)(CCCC)CCCC 8.043 0.051 0.028 0.016 0.0090

C14 2 SC(CCCC)(CCCCC)CCCC 8.603 0.035 0.020 0.011 0.0062

C15 2 SC(CCCC)(CCCCC)CCCCC 9.253 0.023 0.013 0.0071 0.0040

C9 3 SC(C)(C)C(C)(C)C(C)C 5.723 0.223 0.125 0.070 0.040

C10 3 SC(C)(C)C(C)(C)C(C)(C)C 6.153 0.130 0.073 0.041 0.023

C11 3 SC(C)C(C)(C)C(C)(C)C(C)C 6.713 0.115 0.065 0.036 0.020

C12 3 SC(C)(C)C(C)(C)C(C)(C)C(C)C 7.313 0.060 0.034 0.019 0.011

C13 3 SC(C)(C)C(C)(C)C(C)(C)C(C)(C)C 7.733 0.035 0.020 0.011 0.0062


Carbon chain length

Branching levela

SMILES a log Kow

Metabolism rate constant - km (day-1)

10 g fish 100 g fish 1 kg fish 10 kg fish

C14 3 SC(C)C(C)(C)C(C)(C)C(C)(C)C(C)(C) 8.333 0.030 0.017 0.0095 0.0054

C15 3 SC(C)(C)C(C)(C)C(C)(C)C(C)(C)C(C)C 8.933 0.016 0.0088 0.0049 0.0028

C9 4 CC(C)(C)C(S)C(C)(C)C 5.453 0.270 0.152 0.085 0.048

C10 4 CC(C)(C)C(S)(C)C(C)(C)C 5.913 0.154 0.087 0.049 0.027

C11 4 CC(C)(C)CC(S)CC(C)(C)C 6.543 0.132 0.074 0.042 0.023

C12 4 CC(C)(C)CC(C)(S)CC(C)(C)C 7.003 0.075 0.042 0.024 0.013

C13 4 CC(C)(C)C(C)C(S)C(C)C(C)(C)C 7.463 0.071 0.040 0.022 0.013

C14 4 CC(C)(C)C(C)C(S)(C)C(C)C(C)(C)C 7.923 0.040 0.023 0.013 0.0072

C15 4 CC(C)(C)C(C)(C)C(S)C(C)(C)C(C)(C)C 8.633 0.019 0.011 0.0061 0.0034

Notes: The same approach as in the personal communication from industry has been assumed here. Four levels of branching in the carbon chain, with the degree of branching increasing in each level, has been assumed.


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4.1.3 Bioaccumulation in terrestrial organisms (soil dwelling organisms,

vertebrates)

No data are available on soil dwelling organisms.

No toxicokinetic, metabolism or distribution studies were identified for tert- dodecanethiol. The registration dossier noted that simple thiols are metabolised by several pathways in mammalian systems including S-methylation, reaction with glutathione to form mixed disulphides and, especially for low molecular weight thiols, oxidative desulphuration to yield CO and SO4. The relevance of these pathways to TDM is unclear.

4.1.4 Summary and discussion of bioaccumulation

A dietary study with zebrafish gave a BMF clearly below 0.01, indicating a very low potential for accumulation via dietary routes. The derivation of the overall depuration rate constant depends on the method used, possibly a result of the presence of undigested food in the gut of the fish at certain sampling points, and estimates of the k2g range between 0.104 day-1 and 0.6026 day-1. Despite the uncertainty over the exact kinetic parameters derived in this study, all interpretations of the data show a low level of uptake and rapid depuration from the fish.

No experimental BCF value is available for TDM (and such a test would be technically very difficult to perform given the very low water solubilty of TDM and the lack of a representative radiolabelled substance). Both modelling approaches and extrapolation of the data from the dietary study to a BCF have been used to estimate the likely BCF for the substance. The overall depuration rate constant, dominated by metabolism, is an important factor in estimating the BCF for TDM and there is some uncertainty in the precise value for this. The analysis carried out shows that the substance clearly has a BCF below 5,000 l/kg. The situation as to whether the BCF is above 2,000 l/kg is less clear and, although the majority of estimates show the BCF to be below 2,000 l/kg some estimates are above 2,000 l/kg.

The key to understanding the bioaccumulation potential of TDM (particularly whether it has a BCF exceeding 2,000 l/kg) is the overall growth corrected depuration rate constant (which includes a substantial contribution from metabolism). The overall depuration rate constant of 0.61 day-1 may be biased (i.e. too high) as a result of the method used to analyse the data (possibly influenced by undigested food in the gut). In contrast, the lower estimates of around 0.1-0.15 day-1 are probably too low, taking into account the apparent level of metabolism (from estimated metabolic rate constants based on in vitro measurements). This latter conclusion is important, because it means that the BCF of the substance will be below 5,000 l/kg.

Based on the re-analysis of the data by the eMS, the overall depuration rate constant is probably around 0.2 day-1 or possibly higher. The 95% confidence intervals around this value are estimated to be 0.089 to 0.3 day-1. This is consistent with the available metabolism rate constants extrapolated from in vitro data (estimated to be 0.112 day-1


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or 0.165 day-1 depending on the assumptions over blood flow) and from QSAR estimates (values in the range 0.056 day-1 to 0.22 day-1 normalised to a 10 g fish have been estimated). An overall depuration rate constant of 0.2 day-1 leads to lipid normalised and growth corrected BCF values that are close to 2,000 l/kg (some estimates are just above and some estimates are just below). The analysis by Comber and Thomas (2013) shows that when the overall depuration rate constant approaches 0.3 day-1 (as estimated in their re-analysis of the feeding data) the BCF value is below 2,000 l/kg.

Although there is uncertainty around the actual BCF value (resulting from uncertainties in the uptake rate constant, overall depuration rate constant and metabolism rate constant (the latter two are linked)), taking into account the fact that the substance is depurated rapidly from fish, and has a very low biomagnification potential and low dietary assimilation efficiency (based on an experimental fish feeding study), the eMS considers that it is highly likely that the actual BCF value will be below 2,000 l/kg.

5 HUMAN HEALTH HAZARD ASSESSMENT RELEVANT FOR THE

PBT ASSESSMENT

The self-classification proposed by the registrant(s) concerns skin and eye irritation and skin sensitisation. The key study in the registration dossier for repeat dose studies is a 28-day inhalation study with rats. The LOAEC was 220 mg/m3. No repeat dose oral studies are available for TDM but a NOAEL for reproduction/developmental toxicity of 50 mg/kg bw/day was reported in the registration dossier for an analogue substance, n-octyl mercaptan.

No further evaluation by the eMS has taken place for the purposes of this evaluation.


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6 ENVIRONMENTAL HAZARD ASSESSMENT

The substance is registered as an intermediate and so limited data are presented by the registrant. The following sections consider the information available in the registration dossier and EA (2005).

6.1 Aquatic compartment (including sediment)

6.1.1 Toxicity data

It should be noted that only limited information was provided on the toxicity of TDM to aquatic organisms in the original registration dossier. However the updated registration dossier (September 2013) contained robust study summaries of several studies and an assessment of the available data (Thomas and Comber, 2013). It has been noted where robust study summaries are available in the updated dossier but as the focus of this evaluation is on the bioaccumulation test, these have not been evaluated in-depth as part of the current evaluation.

The Thomas and Comber (2013) paper reviewed a total of 14 ecotoxicity studies (9 acute toxicity test, one chronic study with Daphnia magna, one chronic sediment toxicity study on chironomids, one earthworm reproduction test and two activated sludge respiration inhibition tests). Owing to methodological problems only five of these studies were considered valid by Thomas and Comber (2013). These were the two activated sludge respiration inhibition test, one acute toxicity study with Daphnia magna, one chronic toxicity study with Daphnia magna and the chronic sediment study. Several studies were given Klimisch code 4 and were considered by Thomas and Comber (2013) to be suitable for use in a weight of evidence, notably an acute toxicity study with rainbow trout (Oncorhynchus mykiss) and the chronic earthworm study. Overall, Thomas and Comber (2013) concluded that the acute toxicity studies for TDM demonstrate no effects at concentrations at or close to the solubility limit of TDM (taken to be 0.00393 mg/l). It was noted that in some cases effects were seen at intermediate concentrations (i.e. a poor dose-response), but not at higher or lower concentrations but Thomas and Comber (2013) believed that these effects were likely to be physical effects (resulting from the presence of undissolved test substance) rather than true toxic effects.

6.1.1.1 Fish

6.1.1.1.1 Short-term toxicity to fish

No data are reported in the original registration dossier. The results from four acute fish toxicity studies with TDM were reviewed in EA (2005) and these data are summarized in Table 15.


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Table 15: Overview of short-term toxicity to fish (taken from EA, 2005)

Species Value Remarks

Brachydanio rerio 96-h LC0 ≥ 10,000 mg/l Evidence that undissolved substance remained on the surface of the test medium as an oil layer. Result based on nominal concentrations.

Leuciscus idus 48-h LC0 = 50 mg/l

48-h LC100 = 100 mg/l Static test, result based on nominal concentrations.

Salmo salar LC50 = 0.9 mg/l Static test, result based on nominal concentrations.

Oryzias latipes 96-h LC50 = 0.377 mg/l

Although toxic effects were evident in the tests with Leuciscus idus, Salmo salar and Oryzias latipes, the nominal LC50 for these species is above the water solubility of the substance (0.0039 mg/l). It is probable that the actual dissolved concentrations were much lower (as also suggested by studies with invertebrates – see Section 6.1.1.2.1) and therefore there is significant uncertainty over the precise value of the LC50 for these species. The updated registration dossier considers two studies, one study with Danio rerio and one study with Oncorhychus mykiss. Both studies were considered in the updated registration dossier to be technically flawed (nominal concentrations above solubility and insufficient analysis of the actual concentrations). The test with Danio rerio used at water accommodated fraction prepared from a loading of 100 mg/l however as the tanks were opened regularly volatilistaion of the test substance could not be ruled out. The test with Oncorhychus mykiss used a rage of concentrations (0.01, 0.06, 0.35, 1.92, 10.6, 58.2 and 320 mg/l (nominal); all above the water solubility of 0.00393 mg/l) and effects were seen only in the mid-range concentrations and not the lower and higher concentrations. This was considered by Thomas and Comber (2013) to be consistent with physical effects from undissolved test material rather than a true toxic effect. A study using Oryzias latipes was considered in the registration dossier. However as this was carried out in open conditions using an emulsifier it was not considered in the weight of evidence in the registration dossier/Thomas and Comber (2013). Further studies were considered in Thomas and Comber (2013) but little information beyond the effect concentration was available for these studies and no effects were reported at, or close to, the solubility limit in these studies. As toxicity is not the main focus of the current PBT evaluation, the interpretation of the data in the updated registration and the Thomas and Comber (2013) paper has not been investigated in detail by the evaluating Member State.

6.1.1.1.2 Long-term toxicity to fish

No data are reported in the registration dossier or EA (2005).


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6.1.1.2 Aquatic invertebrates

6.1.1.2.1 Short-term toxicity to aquatic invertebrates

The original registration dossier contains the results of four studies. These are summarized in Table 16. The same data are given in EA (2005), along with additional studies which are also briefly summarized below.

Table 16: Overview of short-term toxicity to aquatic invertebrates

Species Value Remarks

Daphnia magna 48-h EC50 = 0.16 mg/l

48-h EC50 = 0.068 mg/l (reinterpreted)

Key study from registration dossier. Substance tested had a stated purity of 98.58%.

Data reinterpreted by EA (2005) in terms of the likely geometric mean concentration (see text).

48-h EC50 = 0.29 mg/l Supporting study from

registration dossier. No data on the purity of the substance tested.

Nominal concentrations according to EA (2005).

48-h EC50 = 0.598 mg/l Supporting study from registration dossier. Substance tested had a stated purity of 100%.

48-h EC50 = 0.0178 mg/l From EA (2005). Based on mean measured concentrations. Significant concentration losses were reported to occur in this test.

48-h EC50 ≥0.25 mg/l From EA (2005). Only one concentration tested.

24-h EC50 = 3.9 mg/l From EA (2005). Nominal concentrations. Oily droplets were evident in the test medium.

24-h EC50 = 1.4-24 mg/l From EA (2005). Nominal concentrations.

Ceriodaphnia dubia 48-h EC50 <1 mg/l (survival)

48-h EC50 <0.5 mg/l (total progeny)

Supporting study from registration dossier. No data on the purity of the substance tested.

Nominal concentrations according to EA (2005).


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The EC50 from the key study with Daphnia magna (48-h EC50 = 0.16 mg/l) from the original registration dossier is close to, but below, the water solubility of TDM in the registration dossier (estimated to be ≤ 0.3 mg/l) but well above the recent experimentally value of 0.0039 mg/l. EA (2005) also reviewed this study and noted that the 48-h EC50 of 0.16 mg/l was based on the initial measured concentration of TDM in the test medium. Instead, EA (2005) recommended that the data should be reinterpreted in terms of the likely geometric mean concentrations throughout the test period as this would result in a more representative 48-h EC50 value (0.068 mg/l). EA (2005) considered it likely that this value would still to be an overestimate of the true EC50 value because the concentrations in the lowest treatments were below the limit of quantitation and in some cases below the limit of detection of the analytical method.

Several of the values reported in the registration dossier and EA (2005) exceed the water solubility of the test substance, and in some cases undissolved test material was

also evident. The lack of measured concentration data means that the precise EC50

from these tests is uncertain. The lowest 48-h EC50 for aquatic invertebrates reported in EA (2005) was 0.0178 mg/l. This is supported by the value determined in the key study from the registration dossier discussed above. It should be noted however that this EC50 value is still above the water solubility of TDM that has recently been determined (0.0039 mg/l) and so the actual EC50 is uncertain.

The available acute invertebrate data have been considered further in the updated (September 2013) registration dossier and Thomas and Comber (2013). This considered the key study to be a study carried out by CERI (2000). The test was considered to show no effects at 0.056 mg/l (nominal). Effects were seen at only the highest concentration of 0.1 mg/l (nominal). The time weighted average concentrations measured were 0.027 mg/l at the nominal 0.1 mg/l treatment and 0.017 mg/l at the nominal 0.056 mg/l treatment.

Four further studies were considered in the updated registration dossier as supporting studies. These studies all suffered from technical limitations and although effects were seen in some of the studies the results were difficult to interpret (e.g. nominal concentrations and/or effects occurring only at concentrations in excess of the water solubility of TDM).

Three further studies were considered in the review by Thomas and Comber (2013). In these studies only 24-h EC50’s were available and these were all >0.25 mg/l but it was noted that insufficient information was available on the studies in order to prepare a robust study summary.

As toxicity is not the main focus of the current PBT evaluation, the interpretation of the data in the updated registration and Thomas and Comber (2013) paper has not been investigated in detail by the evaluating Member State.

6.1.1.2.2 Long-term toxicity to aquatic invertebrates

No long-term toxicity data with aquatic invertebrates are reported in the original registration dossier.

EA (2005) gives the result of a 21-day Daphnia magna reproduction study performed in accordance with the OECD 211 test guideline and conducted to GLP. A 21-d


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NOEC of 0.0108 mg/l was reported based on time-weighted mean measured concentrations (significant loss of test substance was observed during the test). It should be noted that this value is above the actual water solubility of TDM (recently determined to be 0.0039 mg/l). The test was carried out using a cosolvent (DMSO at a concentration of 100 µl/l) and so the solubility of TDM in this test system may be enhanced over that in pure water. This complicates the interpretation of the results but could mean that the true NOEC is actually lower than the reported value.

The updated registration dossier (September 2013) contains two robust study summaries. As toxicity is not the main focus of the current PBT evaluation, the interpretation of the data in the updated registration has not been investigated in detail by the evaluating Member State. The key study in the updated registration dossier is the 21-d study with Daphnia magna, and the NOEC is considered to be 0.0108 mg/l as above. The supporting study in the updated registration dossier is a 7-d NOEC of 0.35 mg/l obtained with Ceriodaphnia dubia. However there are reported to be some limitations with the study (e.g. nominal concentrations and differences between the control and solvent control).

6.1.1.3 Algae and aquatic plants

The original registration dossier contains the results of two algal toxicity studies. The key study is an OECD 201 test guideline study with Pseudokirchnerella subcapitata using TDM (purity 99.1%). The test was carried out using water accommodated fractions (WAFs) prepared from TDM loadings of 100, 32, 10, 3.2, 1 and 0.3 mg/l. The WAFs were prepared by direct weight addition of TDM following by 24 hours stirring and a one hour settling period before separation of the WAF. The 72-h EC50 was determined as the 100 mg/l loading rate and the NOEC was determined as the 3.2 mg/l loading rate. Unfortunately it was not possible to analyse the concentration of TDM actually in solution in this test and therefore the results are difficult to interpret in terms of the actual NOEC for the substance. However, the water solubility of TDM is around 0.0039 mg/l and this implies that effects could have been seen at concentrations around this level.

The second study in the original registration dossier is a supporting study for the read across substance dodecane-1-thiol (CAS No. 112-55-0) with Pseudokirchnerella subcapitata. The method used was the OECD 201 and one concentration of the test substance appears to have been tested. This was prepared from a loading rate of 5 mg/l of the test substance. Analytical monitoring of the test solution found the actual concentration of dodecane-1-thiol present was below 14.5 µg/l (this was the approximate water solubility of the test substance and also the limit of quantification of the analytical method used). At this concentration a 54% inhibition of biomass and a 14% inhibition of growth rate over 72 hours was evident. It was concluded that the 72-h NOEC and EC10 were below 14.5 µg/l for this substance based on both growth rate and biomass and that the 72-h EC50 was below 14.5 µg/l based on biomass. The concentrations measured in this test system are above the water solubility of TDM of 0.0039 mg/l determined in pure water.

EA (2005) included the results of further algal toxicity tests with TDM, but concluded that none of the data were valid.


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The available studies have been considered further in the updated registration dossier (September 2013) and Thomas and Comber (2013). As toxicity is not the main focus of the current PBT evaluation, the interpretation of the data in the updated registration and the Thomas and Comber (2013) paper has not been investigated in detail by the evaluating Member State. The updated registration dossier considers the key study to be the read-across study using dodecane-1-thiol (as above). Two studies with TDM are considered, one using

Pseudokirchnerella subcapitata exposed to water accommodated fractions between 0.3 and 100 mg/l (same study as above) but as only nominal loading rates are available and the flasks were opened at regular intervals this study is considered to be technically flawed in the updated registration dossier and Thomas and Comber (2013). The second study considered in the updated registration dossier and Thomas and Comber (2013) is a study reporting the EC50 for TDM to be >100 mg/l but too little information was available on this study to prepare a robust study summary.

6.1.1.4 Sediment organisms

No data are reported in the original registration dossier or EA (2005). The updated registration dossier (September 2013) includes a robust study summary for a study using Chironomus riparius. As toxicity is not the main focus of the current PBT evaluation, the interpretation of the data in the updated registration and Thomas and Comber (2013) paper has not been investigated in detail by the evaluating Member State. The 28-d NOEC was reported to be 488 mg/kg dry weight (based on nominal concentrations) or 300 mg/kg dry weight (based on mean measured concentrations). The sediment used had an organic carbon content of 2.9%. Thomas and Comber (2013) used equilibrium partitioning to back calculate from this NOEC based on the sediment concentration to an equivalent NOEC of 0.75 mg/l based on the pore water concentration (which exceeds the measured water solubility value by a factor of about 200).

6.1.1.5 Other aquatic organisms


6.2 Terrestrial compartment

6.2.1 Toxicity data

No data are reported in the original registration dossier or EA (2005).

The updated registration dossier (September 2013) and Thomas and Comber (2013) includes the results for a 56-day study using earthworms (Eisenia fetida). As toxicity is not the main focus of the current PBT evaluation, the interpretation of the data in the updated registration and the Thomas and Comber (2013) paper has not been investigated in detail by the evaluating Member State.


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The 56-d NOEC was reported to be ≥100 mg/kg dry weight (no effects were seen at nominal concentrations up to 100 mg/kg dry weight). However, Thomas and Comber (2013) considered that the study has some technical limitations as no analytical verification of the exposure concentration was undertaken and volatile loss was possible.

6.3 Atmospheric compartment


6.4 Microbiological activity in sewage treatment systems

6.4.1 Toxicity to aquatic micro-organisms

The original registration dossier gives the results of two OECD 209 activated sludge respiration inhibition tests. The key study used TDM with a purity of 99.1% and resulted in a 3-h NOEC of 8.6 mg/l (no effects at the highest concentration tested). The supporting study gave a 3-h EC50 of >10 g/l. These two studies are also included in the updated registration dossier (September 2003), although the key study is now considered to be the supporting study from the original dossier (i.e. 3-h EC50 of >10 g/l).

EA (2005) also report the results of the supporting study along with two further studies using single species cultures of bacteria (Pseudomonas putida and Pseudomonas florescens). The 16-h EC50 values determined for P. putida was >10 g/l and the 24-h EC0 value determined for P. florescens was >10 g/l.

6.5 Non compartment specific effects relevant for the food chain

(secondary poisoning)

6.5.1 Toxicity to birds


6.5.2 Toxicity to mammals

The key study in the registration dossier for repeat dose studies is from a 28-day inhalation study with rats. The LOAEC was 220 mg/m3. No repeat dose oral studies are available for TDM in the registration dossier but a NOAEL for reproduction/developmental toxicity of 50 mg/kg bw/day was reported in the registration dossier for an analogue substance n-octyl mercaptan. The registration dossier also gives a NOAEC of 733 mg/m3 for TDM for developmental toxicity for exposure via inhalation.


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7 CONCLUSIONS ON THE PBT/VPVB ASSESSMENT

7.1 Assessment of PBT/vPvB properties – comparison with the

criteria of Annex XIII

7.2 Persistence

TDM is predicted to degrade rapidly in the atmosphere by reaction with hydroxyl radicals and the half-life estimated for this reaction is 1.6 days.

Oxidation of TDM to the corresponding disulphide could theoretically occur in the environment but the available screening experimental data suggests that this is not likely to be a major removal process in aquatic systems.

The substance has been shown to be not readily biodegradable in standard test systems. However, the concentrations tested in some of the studies were well above the water solubility of TDM which may mean that the bioavailability of the substance could have been limited in the test systems used.

A recent study has been carried out to investigate the biodegradation of TDM using a modified OECD 310 test system. The results of this test showed that TDM undergoes substantial primary biodegradation in the presence of sewage sludge (or substances derived from sewage sludge). As can be seen from Table 7 in Section 3.1.2.1 the half- lives obtained in the systems using 4 mg/l MLSS, when corrected for temperature, are longer than the Annex XIII criterion for persistent (half-life of 40 days in freshwater) and very persistent (half-life of 60 days in freshwater). The situation with the experiments carried out using the 30 mg/l MLSS concentration is less clear as it depends, at least in part, on how the degradation rate constant is derived. Some interpretations of the data would result in a half-life >40 days (but less than 60 days) in the abiotic control systems but not the viable system.

It should be noted that the half-lives determined in this study should be used with caution as the test system is designed to establish whether or not a substance is readily biodegradable rather than to obtain a half-life for degradation that is applicable to the

general environment. The MLSS concentrations used in the study (4 or 30 mg/l) are in fact similar to the suspended matter content that is assumed in freshwater environments under REACH (e.g. a suspended matter concentration of 15 mg/l is usually assumed) but the similarity (in terms of physical properties, microbial populations, organic carbon etc.) to the suspended matter found in natural waters is unknown. Therefore, these half-lives should be considered as indicative values rather than precise values for the likely environmental half-life of TDM.

It is also important to note that the study assessed primary degradation only. There is no information available on the identity of the products from primary degradation. In addition covalent binding to solids, if it occurred, would not be detectable by the analytical method used.

The available test results appear to show that extensive primary degradation of TDM may occur but there are still some uncertainties remaining over the interpretation and


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relevance of the results. In particular, a possibility remains that at least some of the loss of the substance seen in the tests with MLSS could have resulted from covalent bonding to the MLSS (which would have not been detectable with the analytical method used). It could be argued that if covalent bonding did occur this could be considered a removal process, as the covalently bound substance would have different properties to TDM itself and such binding may or may not be reversible.

Overall, based on the available data, it cannot be concluded that the substance definitively meets the Annex XIII criteria for persistent or very persistent. However equally there are no results available to demonstrate that the half-life for TDM in freshwater is <40 days or <60 days, i.e. that the substance is not persistent or very persistent.

7.3 Bioaccumulation

A very low biomagnification factor in fish (BMF ~ 0.001) was determined in a dietary study with zebrafish. Although there is some uncertainty over the exact kinetic parameters derived in this study, all interpretations of the data show a low level of uptake and rapid depuration from the fish. This is an important finding as it shows that TDM has a low potential to accumulate from food in the environment.

No experimental BCF value is available for TDM. Both modelling approaches and extrapolation of the data from the dietary study to a BCF have been used to estimate the likely BCF for the substance. The overall depuration rate constant is an important factor in estimating the BCF for TDM and there is some uncertainty in the precise value for this.

Based on the re-analysis of the data by the eMS, the overall depuration rate constant is probably around 0.2 day-1 or possibly higher. The 95% confidence intervals around this value are estimated to be 0.089 to 0.3 day-1. This is consistent with the available metabolism rate constants extrapolated from in vitro data for a second species (estimated to be 0.112 day-1 or 0.165 day-1 depending on the assumptions over blood flow) and from QSAR estimates (values in the range 0.056 day-1 to 0.22 day-1

normalised to a 10 g fish have been estimated). An overall depuration rate constant of 0.2 day-1 leads to lipid normalised and growth corrected BCF values that are close to 2,000 l/kg (some estimates are just above and some estimates are just below). The analysis by Comber and Thomas (2013) shows that when the overall depuration rate constant approaches 0.3 day-1 (as estimated in their re-analysis of the feeding data) the BCF value is below 2,000 l/kg.

Although there is considerable variation in the calculated BCF value due to uncertainties in the uptake rate constant, overall depuration rate constant and metabolism rate constant, taking into account the fact that the substance is depurated rapidly from fish, and has a very low biomagnification potential and low dietary assimilation efficiency (based on an experimental fish feeding study), the eMS considers that it is highly likely that the actual BCF value will be below 2,000 l/kg, and so the substance is not considered to meet the Annex XIII criteria for B or vB.

This is the conclusion of the eMS. A few Member States have expressed concern that the low assimilation factor may be a result of metabolism of the substance in the fish digestion tract. Therefore low uptake may have been due to the route of exposure, and a higher assimilation factor may have been achieved via for example gill exposure. In the


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opinion of the eMS, this uncertainty can be addressed sufficiently using the current data. In particular:

• The S9 study, which uses liver cells demonstrates metabolism would be expected outside of the fish stomach/gut.

• Rapid depuration is observed in the feeding study for the uptake that does occur in fish

• The fish feeding study is specifically designed for poorly soluble chemicals, where uptake in food is expected to be more significant than via aqueous exposure

• Thiols are known to be metabolised in other vertebrate animals, such as rats, by a number of mechanisms. While fish may have a lower metabolic capacity than in rats, the results of the fish feeding study do not contradict the known biochemistry of thiols

7.4 Toxicity

7.4.1 Fulfilment of the T criterion based on human health classification

The substance does not meet the Annex XIII criteria for T based on the human health classification proposed by the registrants. However, insufficient data appear to be available to draw a definitive conclusion.

7.4.2 Fulfilment of the T criterion based on ecotoxicity data

The available ecotoxicity dataset is not sufficient to conclude whether the substance meets the Annex XIII criteria for toxicity (T). The majority of acute toxicity tests with aquatic organisms indicate effects, but were performed above the water solubility limit without analytical verification of test concentrations. The lowest acute toxicity result is for invertebrates, with a 48-h EC50 of 0.0178 mg/l for Daphnia magna. This triggers the screening criterion for T, although it should also be noted that the result is an estimate because of significant concentration losses in the test. In contrast, a long- term study with Daphnia magna resulted in a 21-day NOEC of 0.0108 mg/l, which is close to, but above, the Annex XIII criteria for T. This value is also an estimate due to significant concentration losses. There are no long-term toxicity data for fish.

Effects have also been seen with algae but the actual concentration of TDM causing the effects is unclear. A structurally related substance (dodecane-1-thiol) has an algal 72-h NOErC below 0.014 mg/l, but it is not possible to use this information to confirm the likely NOEC for TDM itself.

It should be noted that the water solubility of TDM included in the registration dossier is ≤ 0.3 mg/l. A recent paper by Comber and Thomas (2013) has suggested that the water solubility of TDM could be lower than assumed (the paper quotes a slow-stir water solubility study giving a water solubility of 0.0039 mg/l). This study has been evaluated by the eMS and is considered to be valid. This is important in interpretation of the available ecotoxicity data as it implies that the effects that were being seen in the various test systems could have been at actual dissolved concentrations below

0.01 mg/l and hence that TDM can be considered to meet the T criterion. It is therefore recommended that the registrant updates the registration dossier to take account of this information.

It should also be noted that only limited information was provided on the toxicity of TDM to aquatic organisms in the original registration dossier. However an updated


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registration dossier (September 2013) contains a larger data set (Thomas and Comber, 2013b)with a total of 14 ecotoxicity studies (9 acute toxicity test, one chronic study with Daphnia magna, one chronic sediment toxicity study on chironomids, one earthworm reproduction test and two activated sludge respiration inhibition tests). Owing to methodological problems only five of these studies were considered valid by Thomas and Comber (2013). These were the two activated sludge respiration inhibition test, one acute toxicity study with Daphnia magna, one chronic toxicity study with Daphnia magna and the chronic sediment study. Several studies were given Klimisch code 4 and were considered by Thomas and Comber (2013) to be suitable for use in a weight of evidence, notably an acute toxicity study with rainbow trout (Oncorhynchus mykiss) and the chronic earthworm study. Overall, Thomas and Comber (2013) concluded that the acute toxicity studies for TDM demonstrate no effects at concentrations at or close to the solubility limit of TDM (taken to be 0.0039 mg/l). It was noted that in some cases effects were seen at intermediate concentrations (i.e. a poor dose-response), but not at higher or lower concentrations but Thomas and Comber (2013) believed that these effects were likely to be physical effects (resulting from the presence of undissolved test substance) rather than true toxic effects.

Overall, based on the low reported water solubility for TDM (0.0039 mg/l) there is a high probability that the effects noted in many of the available acute ecotoxicity tests occurred at concentrations well above the water solubility limit of the test substance. As concluded by Thomas and Comber (2013), these were likely to be physical effects rather than direct toxicity of the substance itself. Although the available data set (acute and chronic) suggests that the substance does not cause toxic effects at concentrations below the 0.01 mg/l cut-off for the T-criterion, in the opinion of the evaluating Member State, it is not possible to conclude unambiguously that this is the case at present. This is because although the water solubility of TDM is clearly low, the value of 0.0039 mg/l was obtained in pure water and the actual solubility in the test medium used in the ecotoxicity tests is not known precisely. So, where no effects were seen at concentrations substantially higher than 0.0039 mg/l, it can be concluded that the substance is not toxic at its solubility limit. However, where effects were seen at concentrations within about an order of magnitude of the water solubility value, it is possible that they were due to exposure to the dissolved test substance.

A conclusion about whether the T-criterion is met on the basis of mammalian effects cannot be drawn based on the available data.

In summary, a definitive conclusion about whether or not the T-criterion is met cannot be made on the basis of the available data.

7.5 Summary and overall conclusions on the PBT, vPvB properties

The substance screens as P/vP, but there are no definitive data to refine this assessment.

Based on a fish dietary bioaccumulation study, it is considered that the substance does not meet the Annex XIII criteria for B or vB.

The available acute and chronic ecotoxicity studies suggest that the substance does not meet the T criterion but it is not currently possible to draw a definitive conclusion about this.

Overall, the substance does not meet the Annex XIII criteria for vPvB or PBT, because of the B/vB conclusion.


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8 REFERENCES

ATOCHEM (1990). Cited in industry IUCLID, MTC, 2003 (taken from EA (2005)).

Baltussen E (2013). Determination of the water solubility of tertiododecylmercaptan. Report No. 499720, WIL Research Europe B.V.

Barber M C (2001). Bioaccumulation and aquatic system simulator (BASS) User’s Manual, Beta Test Version 2.1. EPA/600/R-01/035. United States Environmental Protection Agency, Office of Research and Development, Athens, GA.

Barber M C (2003). A review and comparison of models for predicting dynamic chemical bioconcentration in fish. Environ. Toxicol. Chem., 22, 1963-1992.

Barber M C, Suárez L A and Lassiter R R (1991). Modelling bioaccumulation of organic pollutants in fish with an application to PCBs in Lake Ontario salmonids. Can. J. Fish. Aquat. Sci., 48, 318-337.

Bayer AG internal study (1973). Cited in industry IUCLID, MTC, 2003 (taken from EA (2005)).

Brooke D N, Crookes M J and Merckel D A S (2012). Methods for predicting the rate constant for uptake of organic chemicals from water by fish. Environ. Toxicol. Chem., in press. CERI (2000). Acute swimming inhibition test of tert-dodecanethiol on Daphnia magna. No. 2000-07-30.Chemicals Evaluation and Research Institute, Japan. Report (as reported in the updated registration dossier (September 2013).

Comber M H I and Thomas P C (2013). Weight of evidence report addressing the bioaccumulation potential of tert-dodecanethiol. Report dated 11th April 2013.

Cowan-Ellsberry C E, Dyer S D, Erhardt S, Bernhard M J, Roe A L, Dowty M E and Weisbrod A V (2008). Approach for extrapolating in vitro metabolism data to refine bioconcentration factor estimates. Chemoshere, 70, 1804-1817.

Crookes M and Brooke D (2011). Estimation of fish bioconcentration factor (BCF) from depuration data. Science Report SCHO0811BUCE-E-E, Environment Agency for England and Wales (available from http://publications.environment- agency.gov.uk/PDF/SCHO0811BUCE-E-E.pdf).

Davis J W and Erhardt S (2009). Tert-Dodecanethiol: Estimation of bioaccumulation using the in vitro trout liver S9 metabolism assay. Study Report for Mercaptan Thiol Council (MTC), Study ID: 081106, Toxicology & Environmental Research and Consulting, The DOW Chemical Company, Midland, Michigan.

Davis J W, and Hancock G (2009). Bioaccumulation assessment of tert-dodecanethiol (CAS# 25103-58-6) using predictive approaches. Unpublished, Dow Chemical Company internal memo.

Davis J W, Gonsior S J, Perala A W and Hales C A (2009). The aerobic biodegradation of tert-dodecanethiol in a modified ready biodegradability test. Study Report for Mercaptan Thiol Council (MTC), Study ID: 081090, Toxicology & Environmental Research and Consulting, The DOW Chemical Company, Midland, Michigan.


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Dimitrov S D, Dimitrova N C, Pakerton T F, Comber M, Bonnell M and Mekenyan O (2005). Base-line model for identifying the bioaccumulation potential of chemicals. SAR and QSAR in Environmental Research, 16(6), 531-55.

Egeler P, Gilberg D and Baltussen E (2013). Tertiododecylmercaptan: A study on the dietary accumulation in zebrafish. Study Report for Mercaptan Thiol Council (MTC), Study Number (ECT): 11CN1FX, NOTOX Phase ID: 496666, ECT Oekotoxikologie GmbH, Germany and NOTOX B.V., the Netherlands.

Environment Agency [EA] (2005). Environmental risk assessment report: Tert- dodecanethiol. Science Report, January 2005. Authors: Fisk P R, Girling A E, McLaughlin L and Wildey R J. Environment Agency, Bristol. Environment Agency (2012). Depuration rate constant: growth correction and use as an indicator of bioaccumulation potential. Environment Agency Science Report, Product code: LIT 7371, October 2012. Authors: Brooke D and Crookes M. Environment Agency, Bristol. (http://a0768b4a8a31e106d8b0-50dc802554eb38a24458b98ff72d550b.r19.cf3.rackcdn.com/LIT_7371_44228a.pdf).

Erickson R J and McKim J M (1990a). A simple flow-limited model for exchange of organic chemicals at fish gills. Environ. Toxicol. Chem., 9, 159-165.

Erickson R J and McKim J M (1990b). A model for exchange of organic chemicals at fish gills: Flow and diffusion limitations. Aquatic Toxicol., 18, 175-198. Fisk AT, Norstrom R J, Cymbalisty and Muir D C G (1998). Dietary accumulation and depuration of hydrophobic organochlorines: Bioaccumulation parameters and their relationship with the octanol/water partition coefficient. Environ. Toxicol. Chem., 17, 951–961. Gobas F A P C, McCorquodale J R and Haffner G D (1993). Intestinal absorption and biomagnification of organochlorines. Environ. Toxicol. Chem., 12, 567–576.

Hayton W L and Baron M G (1990). Rate-limiting barriers to xenobiotic uptake by the gill. Environ. Toxicol. Chem., 9, 151-157.

Hendriks A J , Van Der Linde A, Cornelissen G and Sijm D T H M (2001). The power of size. 1. Rate constants and equilibrium ratios for accumulation of organic substances related to octanol-water partition ratio and species weight. Environ. Toxicol. Chem., 20, 1399-1420. Inoue Y, Hashizume N, Yoshida T, Murakami H, Suzuki Y, Koga Y, Takeshige R, Kikushima E, Yakata N and Otsuka M (2012). Comparison of bioconcentration and biomagnification factors for poorly water-soluble chemicals using common carp (Cyprinus carpio L.). Arch. Environ. Contam. Toxicol., 63, 241-8.

ISO 10634 (1995). Water quality – Guidance for the preparation and treatment of poorly water-soluble organic compounds for the subsequent evaluation of their biodegradability in an aqueous medium. International Organization for Standardization.

Jenkins (1990). Tertiary dodecyl mercaptan: assessment of its ready biodegradability. Closed bottle test. Study No 90/PSVO40/0649, Life Science Research Limited, 15/10/1990.


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Niimi A J and Oliver B G (1988). Influence of molecular weight and molecular volume on dietary adsorption and assimilation efficiency of chemicals by fishes. Can. J. Fish. Aquat. Sci., 45, 222–227. OECD (2012). Validation report of a ring test for the OECD 305 dietary exposure bioaccumulation fish test (Part 1) with additional report including comparative analysis of trout and carp results (Part II). Series on Testing and Assessment No. 175, ENV/JM/MONO(2012)20, Organisation for Economic Co-operation and Development (http://search.oecd.org/officialdocuments/displaydocumentpdf/?cote=env/jm/mono(2012)20&doclanguage=en).

OECD (2013). Validation report of a ring test for the OECD 305 dietary exposure bioaccumulation fish test. Additional report including results using a lower feeding rate. Series on Testing and Assessment No. 191, ENV/JM/MONO(2013)15, Organisation for Economic Co-operation and Development (http://search.oecd.org/officialdocuments/displaydocumentpdf/?cote=ENV/JM/MONO(2013)15&doclanguage=en).

PBT Working Group (2008). PBT Working Group – PBT list no. 98, UK_PBT_98_25103-58-6_Rev17, February 2008. PBT Working Group, European Chemicals Bureau.

Perala A W and Markham D A (no year). Method for the analysis of tert-dodecyl mercaptan in aqueous matrices. Study Report for Mercaptans/Thiols Council, Study ID 081090, the Dow Chemical Company. Personal communication to the evaluating Member State, 6th December 2013. Replies to the questions raised by the PBT Expert Group on the PBT assessment conducted for tert-dodecanethiol also known as tert-dodecyl mercaptan (TDM). Including ‘Annex I: iSafeRat® Regression-Fragment Approach (RFA) for log Kow prediction (version 1.1)’ and ‘Annex II: Information on the iSafeRat® GSE QSAR validity.

Sijm D T H M, Verberne M E, De Jonge W J, Pärt P and Opperhuizen A (1995). Allometry in the uptake of hydrophobic chemicals determined in vivo and in isolated perfused gills. Toxicol. Appl. Pharmacol., 131, 130-135.

Spacie A and Hamelink J L (1982). Alternative models for describing the bioconcentration of organics in fish. Environ. Toxicol. Chem., 1, 309-323.

Streit B and Siré E O (1993). On the role of blood proteins for uptake, distribution, and clearance of waterborne lipophilic xenobiotics by fish – a linear system analysis. Chemosphere, 26, 1031-1039.

Thiebaud H (1994). Biodegradabilite du n-octyl mercapan et du t-dodecyl mercaptan. Report No.53062/HT, Elf Atochem S.A., 26/8/1994.

Thomann R V (1989). Bioaccumulation model of organic chemical distribution in aquatic food chains. Environ. Sei. Technol., 23, 699-707. Thomas P C and Comber M H I (2013). Review and validity of available ecotoxicity data on tert-dodecanethiol. Report dated 19th April 2013

Tolls J and Sijm D T H M (1995). A preliminary evaluation of the relationship between bioconcentration and hydrophobicity for surfactants. Environ. Toxicol. Chem., 14, 1675-1685.


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INFORMATION ON USES AND EXPOSURE

1 MANUFACTURE AND USE(S)

1.1 Quantities

The registered quantity is confidential. The ECHA dissemination database indicates that the substance has been registered in the 1,000-10,000 tonnes/year band (joint submission).

1.2 Identified uses

The following identified uses of TDM are given in the ECHA dissemination database.

Manufacture

Manufacture of tert- dodecanethiol

Process categories PROC 1: Use in closed process, no likelihood of exposure

PROC 2: Use in closed, continuous process with occasional controlled exposure

PROC 8b: Transfer of substance or preparation (charging/discharging) from/to vessels/large containers at dedicated facilities

PROC 9: Transfer of substance or preparation into small containers (dedicated filling line, including weighing)

PROC 15: Use as laboratory reagent

Environmental release category

ERC 1: Manufacture of substances

Uses at Industrial Sites

Use as intermediate under strictly controlled conditions

Process category PROC 1: Use in closed process, no likelihood of exposure

PROC 3: Use in closed batch process (synthesis or formulation)



Chemical product category


Substance supplied to that use in form of

PC 19: Intermediate

ERC 6a: Industrial use resulting in manufacture of another substance (use of intermediates)

As such

Sector of end use SU 9: Manufacture of fine chemicals

SU 12: Manufacture of plastics products, including compounding and conversion

Subsequent service life No relevant for that use?


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Use as intermediate




PROC 4: Use in batch and other process (synthesis) where opportunity for exposure arises

PROC 8a: Transfer of substance or preparation (charging/discharging) from/to vessels/large containers at non-dedicated facilities


PROC 9: Transfer of substance or preparation into small containers (dedicated filling line, including weighing)





As such

Sector of end use SU 8: Manufacture of bulk, large scale chemicals (including petroleum products)

SU 9: Manufacture of fine chemicals

Subsequent service life No relevant for that use?

01b - Use as an Intermediate









As such

Sector of end use SU 0: Other: Industrial manufacturing (all)

SU 8: Manufacture of bulk, large scale chemicals (including petroleum products)


Subsequent service life relevant for that use?

Yes

01b - Use as an Intermediate





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PROC 4: Use in batch and other process (synthesis) where opportunity for exposure arises

PROC 8a: Transfer of substance or preparation (charging/discharging) from/to vessels/large containers at non-dedicated facilities




As such

Sector of end use SU 8: Manufacture of bulk, large scale chemicals (including petroleum products)


SU 0: Other: Industrial Manufacturing (All)

Subsequent service life relevant for that use

Yes

It should be noted that the information provided in the above Table was taken from the ECHA Dissemination database in August 2013. At the time the substance was registered as an intermediate. Since that time an updated registration dossier (September 2013) has been submitted for the substance covering uses other than intermediate uses (and including exposure scenarios and quantitative risk assessments for those uses). As the availability of this updated registration dossier was not clear to the evaluating Member State until October 2013, after this PBT evaluation was drafted, and as this evaluation is concentrating on the bioaccumulation aspect of TDM, the new identified uses have not been considered as part of this evaluation.

2 POTENTIAL EXPOSURE AND RISK(S)

2.1 Workers

Not relevant for the current factsheet.

2.2 Consumers

Not relevant for the current factsheet.

2.3 Environment

The originally registered uses of the substance are as an intermediate (see above). The release to the environment from this use is expected to be well controlled at industrial sites.

EA (2005) reported that the main use of TDM was as a chain transfer agent in the production of emulsion polymers, particularly styrene butadiene rubber (SBR or E- SBR) and acrylonitrile butadiene, or nitrile, rubber (NBR). In addition EA (2005)


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indicated other uses of the substance to include production of emulsion polymer dispersions and latices which can be supplied and further processed in the form of aqueous dispersions (for example for carpet backing and underlay, textiles, paper coatings, adhesives, dipped rubber goods and products for the construction industry).

2.3.1 Monitoring data

No data

2.4 Man exposed via the enviromnent

No information


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Confidential annex(es).

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