GUIDANCE

Guidance on the Drinking Water Directive

Volume II: Methodologies for accepting starting

substances, compositions and constituents for use in the
manufacture of materials or products in contact with water
intended for human consumption

Version 1.0
 Guidance on the DWD: Volume II
Version 1.0 – January 2025

Legal notice
This document aims to assist users in complying with their obligations under Directive (EU)
2020/2184 of the European Parliament and of the Council of 16 December 2020 on the quality
of water intended for human consumption (recast), hereafter referred to as the drinking water
directive (DWD). However, users are reminded that the text of the DWD (Article 11(2)(a)) and
Commission Implementing Decision (EU) 2024/365 are the only authentic legal references and
that the information in this document does not constitute legal advice. Usage of the information
remains the sole responsibility of the user. The European Chemicals Agency does not accept any
liability with regard to the use that may be made of the information contained in this document.

Guidance on the Drinking Water Directive – Volume II: Methodologies for accepting
starting substances, compositions and constituents for use in the manufacture of
materials or products in contact with water intended for human consumption

Reference: ECHA-24-G-12-EN
Cat. number: ED-01-24-023-EN-N
ISBN: 978-92-9468-446-2
DOI: 10.2823/6310589
Publ. date: January 2025
Language: EN

© European Chemicals Agency, 2025

Luxembourg: Publications Office of the European Union, 2025

If you have questions or comments in relation to this document, please send them (quoting
the reference and issue date) using the information request form. The information request form
can be accessed via the Contact ECHA page at http://echa.europa.eu/contact.

European Chemicals Agency

Mailing address: P.O. Box 400, FI-00121 Helsinki, FINLAND

Visiting address: Telakkakatu 6, FI-00150 Helsinki, FINLAND
Guidance on the DWD: Volume II
Version 1.0 – January 2025 3

Preface
Guidance on the Drinking Water Directive – Volume II (accepting methodology), is to be used in
preparing applications for starting substances, compositions and constituents for approval for
inclusion in the European positive lists.
This document describes the applicants’ obligations according to Article 11(2)(a) of Directive
(EU) 2020/2184 and how to fulfil them.
The guidance provides technical scientific advice on how to comply with the methodologies for
testing and how to perform the risk assessment and the exposure assessment for the evaluation
of human health set out in Commission Implementing Decision (EU) 2024/365.
Notes for the reader

This guidance must be read in conjunction with the relevant DWD implementing
legislation, particularly:

• Commission Implementing Decision (EU) 2024/365 of 23 January 2024 laying down

rules for the application of Directive (EU) 2020/2184 of the European Parliament and
of the Council as regards methodologies for testing and accepting starting substances,
compositions and constituents to be included in the European positive lists;
• Commission Implementing Decision (EU) 2024/367 of 23 January 2024 laying down
rules for the application of Directive (EU) 2020/2184 of the European Parliament and
of the Council by establishing the European positive lists of starting substances,
compositions and constituents authorised for use in the manufacture of materials or
products that come into contact with water intended for human consumption;
• Commission Delegated Regulation (EU) 2024/369 of 23 January 2024 supplementing
Directive (EU) 2020/2184 of the European Parliament and of the Council by laying
down the procedure regarding inclusion in or removal from the European positive lists
of starting substances, compositions and constituents.

In addition, this guidance document is part of wider guidance on DWD applications

comprising a total of five volumes. The remaining volumes are:

• DWD Guidance Volume I on methodologies for testing starting substances,

compositions and constituents for use in the manufacture of materials or products in
contact with water intended for human consumption;
• DWD Guidance Volume III on the scope of DWD applications;
• DWD Guidance Volume IV on the contents of a notification of intention;
• DWD Guidance Volume V on calibrants and representative samples of accepted
compositions in connection to applications submitted under Commission Delegated
Regulation (EU) 2024/369 (this guidance was prepared by the European Commission
Joint Research Centre).
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Contents
Legal notice .................................................................................................... 2
Preface ........................................................................................................... 3
Contents .......................................................................................................... 4
List of figures .................................................................................................... 8
List of tables ..................................................................................................... 9
Abbreviations .................................................................................................... 10
1. Introduction .................................................................................................. 12
1.1. Evaluation ............................................................................................ 12
1.2. Assessment .......................................................................................... 13
2. General principles .......................................................................................... 15
3. Hazard assessment: hazard identification ......................................................... 17
3.1. Collection and evaluation of all available data ............................................ 21
3.1.1. Collection of data (step 4 in Figure 1 and Figure 2) ............................ 21
3.1.2. Evaluation of the data (step 5 in Figure 1 and Figure 2) ...................... 21
3.1.2.1. Completeness of data ................................................................. 21
3.1.2.2. Adequacy of data ....................................................................... 21
3.1.2.2.1. Reliability of data ................................................................... 22
3.1.2.2.2. Relevance of data .................................................................. 23
3.1.2.2.3. Human data .......................................................................... 24
3.1.2.2.4. In vitro data .......................................................................... 26
3.1.2.2.5. (Quantitative) structure–activity relationships ........................... 26
3.1.3. Read-across approach .................................................................... 31
3.1.4. Weight of evidence (WoE) ............................................................... 32
3.2. Toxicokinetics ....................................................................................... 33
3.2.1. Definitions and uses of toxicokinetic studies ...................................... 33
3.2.2. Main principles of toxicokinetic studies ............................................. 34
3.2.2.1. Accumulation ............................................................................ 35
3.2.3. Prediction of toxicokinetics .............................................................. 35
3.2.3.1. Absorption ................................................................................ 35
3.2.3.2. Distribution ............................................................................... 38
3.2.4. Accumulation potential ................................................................... 39
3.2.4.1. Metabolism ............................................................................... 40
3.2.4.2. Excretion .................................................................................. 40
3.2.5. Generating and integrating toxicokinetic information .......................... 42
3.2.6. Other methods for generating absorption, distribution, metabolism and excretion
data ............................................................................................. 42
3.2.6.1. In vitro studies .......................................................................... 42
3.2.6.2. In silico methods: kinetic modelling ............................................. 43
3.2.7. Variability and uncertainty in toxicokinetics ....................................... 43
3.2.8. Metabolism studies as a basis for internal dose considerations............. 43
3.2.9. Extrapolation ................................................................................. 45
3.2.9.1. Interspecies extrapolation ........................................................... 46
3.2.9.2. Interroute extrapolation.............................................................. 47
3.3. Repeated dose toxicity ........................................................................... 48
3.3.1. Definition of repeated dose toxicity .................................................. 48
3.3.2. Data to be used in the hazard assessment ........................................ 49
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3.3.2.1. Non-human data for repeated dose toxicity ................................... 49

3.3.2.1.1. Non-testing data for repeated dose toxicity ............................... 49
3.3.2.1.2. Testing data for repeated dose toxicity ..................................... 49
3.3.3. Specific system/organ toxicity ......................................................... 59
3.3.3.1. General aspects ......................................................................... 59
3.3.3.2. Neurotoxicity ............................................................................. 59
3.3.3.2.1. Definition of neurotoxicity ....................................................... 59
3.3.3.2.2. Introduction .......................................................................... 59
3.3.3.2.3. Structure–activity considerations ............................................. 60
3.3.3.2.4. Assessment of available information or results from initial testing 60
3.3.3.2.5. Further neurotoxicity testing ................................................... 61
3.3.3.3. Immunotoxicity ......................................................................... 61
3.3.3.3.1. Definition of immunotoxicity .................................................... 61
3.3.3.3.2. Introduction .......................................................................... 61
3.3.3.3.3. Immunosuppression ............................................................... 62
3.3.3.4. Effects on the endocrine system .................................................. 64
3.3.3.5. Remaining uncertainty ................................................................ 64
3.3.3.5.1. Threshold of the critical effect.................................................. 65
3.3.3.5.2. Other considerations .............................................................. 65
3.3.4. Conclusions on repeated dose toxicity .............................................. 65
3.3.5. Concluding on suitability for risk assessment ..................................... 66
3.4. Mutagenicity ......................................................................................... 67
3.4.1. Definition ...................................................................................... 67
3.4.2. Data to be used in the assessment of genotoxicity/mutagenicity .......... 68
3.4.2.1. Non-human data for mutagenicity ................................................ 68
3.4.2.1.1. Non-testing data for mutagenicity ............................................ 68
3.4.2.1.2. Testing data for mutagenicity .................................................. 68
3.4.2.2. Human data .............................................................................. 73
3.4.3. Remaining uncertainty on mutagenicity ............................................ 73
3.4.4. Concluding on suitability for risk assessment ..................................... 74
3.4.4.1. Considerations on dose–response shapes and mode of action of mutagenic
substances in test systems ............................................................. 74
3.5. Carcinogenicity ...................................................................................... 75
3.5.1. Definition ...................................................................................... 75
3.5.2. Data to be used in the hazard assessment ........................................ 76
3.5.2.1. Non-human data for carcinogenicity ............................................. 76
3.5.2.1.1. Non-testing data for carcinogenicity ......................................... 76
3.5.2.1.2. Testing data for carcinogenicity ............................................... 77
3.5.2.2. Human data .............................................................................. 82
3.5.3. Remaining uncertainty.................................................................... 82
3.5.4. Concluding on suitability for risk assessment ..................................... 83
3.6. Reproductive toxicity.............................................................................. 84
3.6.1. Definition ...................................................................................... 84
3.6.2. Data to be used in the effects assessment ........................................ 85
3.6.2.1. Non-human data........................................................................ 85
3.6.2.1.1. Non-testing data .................................................................... 85
3.6.2.1.2. Testing data for reproductive toxicity ....................................... 85
3.6.2.1.3. Relation between maternal toxicity and developmental toxicity .... 91
3.6.2.1.4. Reproductive toxicity via lactation ............................................ 92
3.6.2.2. Human data .............................................................................. 93
3.6.3. Remaining uncertainty on reproductive toxicity.................................. 93
3.6.4. Conclusions on reproductive toxicity ................................................. 93
3.6.5. Concluding on suitability for risk assessment ..................................... 93
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4. Hazard assessment: hazard characterisation (dose–response / concentration-response

relationship) ..................................................................................................... 94
4.1. Introduction .......................................................................................... 94
4.2. Identification of critical effects ................................................................. 96
4.2.1. Hazard information underlying the derivation of a derived no-effect level ... 96
4.2.1.1. Toxicokinetics and absorption .............................................................. 96
4.2.1.2. Systemic (target organ) toxicity after repeated dose administration ......... 96
4.2.1.3. Toxicity to reproduction and development ............................................. 97
4.2.1.4. Other toxicity end points .................................................................... 97
4.2.2. Derived no-effect level derivation for threshold effects ............................. 97
4.2.2.1. Relevant time frames in DNEL derivation .............................................. 98
4.2.2.2. Application of appropriate assessment factors ....................................... 98
4.2.2.3. Intraspecies differences ...................................................................... 100
4.2.3. Dose–response assessment .................................................................. 102
4.3. Non-threshold effects ............................................................................. 102
5. Exposure assessment ..................................................................................... 103
5.1. Determination of concentration at the tap from physical testing .................. 103
5.1.1. Organic materials ................................................................................ 103
5.1.2. Metallic materials ................................................................................. 104
5.1.2.1. Determination of concentration at the tap values from physical testing in accordance
with EN 15664-1 ................................................................................. 104
5.1.3. Cementitious materials ......................................................................... 105
5.1.4. Enamel, ceramic and other inorganic materials ........................................ 105
5.1.5. Surface layers ..................................................................................... 106
5.2. Determination of concentration at the tap from modelling for organic materials 107
6. Risk acceptance............................................................................................. 108
6.1. Main principles ...................................................................................... 108
6.1.1. Approaches to maximum tolerable concentration at the tap adoption for relevant
chemical species .................................................................................. 108
6.1.2. Concentration at the tap and migration tiers ........................................... 108
6.2. Limited approach to the adoption of a maximum tolerable concentration at the tap value
109
6.2.1. Limited approach 1: adoption of a maximum tolerable concentration at the tap value
on the basis of a robust past assessment ................................................ 109
6.2.2. Limited approach 2: maximum tolerable concentration at the tap value for relevant
chemical species classified as carcinogenic, mutagenic or toxic to reproduction
category 1A or 1B / endocrine disruptor human health category 1/PBT/vPvB/PMT/vPvM
112
6.2.3. Limited approach 3: maximum tolerable concentration at the tap value for relevant
chemical species that are potential carcinogens, mutagens or endocrine disruptors with
a non-threshold mode of action ............................................................. 114
6.2.4. Limited approach 4: maximum tolerable concentration at the tap value for relevant
chemical species for which available toxicological information is insufficient 114
6.3. Comprehensive approach to the adoption of a maximum tolerable concentration at the
tap value on the basis of toxicological information ................................... 118
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6.3.1. Maximum tolerable concentration at the tap for a relevant chemical species of low
migration potential .............................................................................. 118
6.3.2. Maximum tolerable concentration at the tap for a relevant chemical species of medium
migration potential .............................................................................. 118
6.3.3. Maximum tolerable concentration at the tap for a relevant chemical species of high
migration potential .............................................................................. 119
6.4. Selection of allocation factors .................................................................. 120
6.5. Risk acceptance methodology.................................................................. 121
6.5.1. Standard methodology ......................................................................... 121
6.5.2. Metallic compositions ........................................................................... 121
6.5.2.1. Metallic compositions tested according to EN 15664-1 ............................ 122
6.5.2.2. Metallic compositions tested according to EN 16056 (passivating alloys) ... 123
6.5.3. Surface layers (coatings, platings and linings) ......................................... 124
7. References ................................................................................................... 125
8. Annex I – Schematic of the metallics acceptance process ................................... 133
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List of figures
Figure 1. Schematic representation of the stepwise approach for fulfilling the testing
methodology for human health hazard for starting substances and constituents, for the
purpose of the DWD .......................................................................................... 19
Figure 2. Schematic representation of the stepwise approach for fulfilling the testing
methodology for human health hazard for compositions, for the purpose of the DWD 20
Figure 3. Use of increasing knowledge on substance metabolism ............................. 44
Figure 4. Schematic representation of the steps for performing the hazard assessment 95
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List of tables
Table 1. How different types of data can be used to interpret oral/GI absorption ....... 37
Table 2. How different types of data can be used to interpret substance distribution .. 39
Table 3. How data from different sites can be used to interpret accumulation after oral
exposure .......................................................................................................... 40
Table 4. How physico-chemical data can affect excretion after oral exposure ............ 41
Table 5. Overview of in vivo oral repeated dose toxicity test guideline studies ........... 54
Table 6. Overview of other in vivo test guideline studies giving information on repeated dose
toxicity ............................................................................................................. 57
Table 7. In vitro test methods ............................................................................. 69
Table 8. Somatic cells – in vivo test methods ........................................................ 70
Table 9. Germ cells – in vivo test methods............................................................ 71
Table 10. Overview of in vivo OECD test guidelines for reproductive toxicity ............. 89
Table 11. Allometric scaling factors for different animal species when compared with humans
(assuming the human body weight is 70 kg) ......................................................... 99
Table 12. Migration tiers equivalents for FCM and DWCM assessments ..................... 109
Table 13. Sources of existing MTCtap values for different material types .................... 110
Table 14. Summary of limited risk acceptance approaches ...................................... 116
Table 15. MTCtap values from toxicological data according to different migration tiers under the
comprehensive approach .................................................................................... 120
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Abbreviations
ADME absorption, distribution, metabolism and excretion

AF assessment factor

ALF allocation factor

AUC area under the curve

BMD benchmark dose

CLP regulation Regulation (EC) No 1272/2008 on classification, labelling and packaging of

substances and mixtures

CMR carcinogenic, mutagenic or toxic to reproduction

CNS central nervous system

Ctap concentration at the tap

DNEL derived no-effect level

DTU Technical University of Denmark

DWCM drinking water contact material

DWD drinking water directive

ECETOC European Centre for Ecotoxicology and Toxicology of Chemicals

ECHA European Chemicals Agency

ECVAM European Centre for the Validation of Alternative Methods

ED endocrine disruptor

EEA European Union + Norway, Iceland and Liechtenstein

EFSA European Food Safety Authority

EUPL European positive list

FCM food contact material

GI gastrointestinal

GLP good laboratory practice

IPCS International Programme on Chemical Safety

JRC Joint Research Centre

LOAEC lowest observed adverse effect concentration

LOAEL lowest observed adverse effect level

Kow partition coefficient octanol–water

KP,W partition coefficient organic material-water

MEPa(T) average of the arithmetic mean of the equivalent pipe concentrations

(MEPn(T)) of the three test lines in one rig
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MTCtap maximum tolerable concentration at the tap

N(L)OAEL NOAEL and/or LOAEL

NOAEC no observed adverse effect concentration

NOAEL no observed adverse effect level

OECD Organisation for Economic Co-operation and Development

PBPK physiologically based pharmacokinetic

PBT persistent, bioaccumulative and toxic

PMT persistent, mobile and toxic

(Q)SAR (quantitative) structure–activity relationship

REACH regulation Regulation (EC) No 1907/2006 on the registration, evaluation, authorisation

and restriction of chemicals

SML specific migration limit

TG Test Guideline

TK toxicokinetic

TOC total organic carbon

US EPA United States Environmental Protection Agency

vPvB very persistent and very bioaccumulative

vPvM very persistent and very mobile

w/w weight per weight ratio

WoE weight of evidence

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1. Introduction
1.1. Evaluation
To ensure a high level of protection of human health, any risks arising from relevant chemical
species that migrate into water intended for human consumption as a result of the use of a
starting substance, composition or constituent must be identified.

This document provides technical and scientific advice on how to fulfil the common principles set
out in Annex VI to Commission Implementing Decision (EU) 2024/365 of the acceptance
methodology for starting substances, compositions and constituents, based on the assessment
of the risk raised by the relevant chemical species.

According to Section 2 of Annex VI to Commission Implementing Decision (EU) 2024/365, the

acceptance methodology should be based on risk assessment that shall result in:

• determining the maximum tolerable concentration at tap water (MTCtap) for each relevant chemical
species;

• ensuring that Ctap [the concentration at the tap] for each relevant chemical species is lower than
its MTCtap.

In addition to the information required under Annexes I, II and III, a risk assessment shall take into
account any other relevant technical or scientific information which is available addressing worst
foreseeable conditions of use. Where appropriate, conditions of use shall be implemented.

The information provided in the risk assessment shall allow the Committee for Risk Assessment of ECHA
to evaluate and reach an opinion on whether the starting substance, composition or constituent
complies with the criteria set out under Article 11(1) of Directive (EU) 2020/2184.

This guidance is primarily addressed to applicants seeking approval for a starting substance,
composition or constituent to be included in a European positive list (EUPL) and also those
seeking the removal or updating of an existing entry in an EUPL.

This guidance needs to be considered together with Annex V (‘Toxicological properties’) to

Commission Implementing Decision (EU) 2024/365 and Guidance on the Drinking Water
Directive – Volume I, hereafter DWD Guidance Volume I, on testing methodologies.

In this document, unless explicitly described otherwise, the term ‘applied-for entry’ is used to
refer to starting substances, compositions and constituents for which applicants may seek
approval for inclusion in or removal from an EUPL, or their renewal, from the European Chemicals
Agency (ECHA). This is how the starting substance, composition or organic cementitious
constituent would be presented if included in an EUPL. In contrast, for the purpose of
toxicological properties assessment and risk assessment and acceptance, the term ‘relevant
chemical species’, as identified in accordance with Section 3 of Annex IV to Commission
Implementing Decision (EU) 2024/365, is used.

A glossary of key terms, also relevant to this guidance document, is provided in DWD Guidance
Volume I.
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1.2. Assessment
The risk assessment process in relation to human health entails a sequence of actions outlined
below.

• Hazard assessment, comprising the following.

o Hazard identification. Identification of the adverse effects that relevant chemical

species have an inherent capacity to cause.

o Hazard characterisation. Dose (concentration)–response (effects) assessment, which

is the estimation of the relationship between dose, or level of exposure to relevant
chemical species, and the incidence and severity of an effect, where appropriate. Based
on this information, a threshold concentration called maximum tolerable concentration at
the tap (MTCtap) should be calculated for each relevant chemical species.

• Exposure assessment. Estimation of the concentrations/doses to which human populations

(i.e. consumers) are or may be exposed. In the context of the drinking water directive
(DWD), this is to establish the concentration at the tap (Ctap), that is, the value that
constitutes the amount of the relevant chemical species in drinking water to which humans
are potentially exposed.

• Risk acceptance. In the context of the DWD, the applied-for entry does not pose an
unacceptable risk if, for all relevant chemical species, the Ctap is lower than the corresponding
MTCtap.

A risk assessment covering all of the steps identified above must be carried out for all relevant
chemical species, unless they meet one of the criteria listed in Section 1 of Annex VI to
Commission Implementing Decision (EU) 2024/365. For such substances, the first step – hazard
assessment – is omitted (see also Section 6.1.1 of DWD Guidance Volume I and Sections 6.2.1–
6.2.4 of this document).

The risk assessment for human health must address the following potential toxic effects and the
exposure of the general human population by the oral route:

• systemic (target organ) toxicity after repeated dose administration;

• mutagenicity;

• toxicity for reproduction;

• carcinogenicity;

• neurotoxicity;

• immunotoxicity;

• endocrine disruption.

The risk assessment should be carried out based on all data available, applying the methods
described in the following sections of this guidance.

Applicants should note that applications under the DWD cannot be made for uses of chemical
substances that would result in the weakening of a restriction under Annex XVII to the
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registration, evaluation, authorisation and restriction of chemicals (REACH regulation) or the

authorisation regime of the REACH regulation.

Note for the reader:

The text of Commission Implementing Decision (EU) 2024/365 can be found at
http://data.europa.eu/eli/dec_impl/2024/365/oj.
The term ‘chapter’ is used to refer to the highest level parts of the document (Chapter 1,
2, 3, etc.). Lower level headings (e.g. 1.1) are referred to as ‘sections’.
Guidance on the DWD: Volume II
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2. General principles
In the context of the DWD, the assessment should be based on the risks arising from all relevant
chemical species.

Section 3 of Annex IV to Commission Implementing Decision (EU) 2024/365 defines relevant

chemical species as follows.

Relevant chemical species are those that are covered by the requirements set out in Annex V in order
to demonstrate that the starting substance, composition or organic cementitious constituent meets the
acceptance criteria set out in Annex VI. Relevant chemical species include the following:

(a) starting substances and organic cementitious constituents which function as a monomer or other
reactant of a main polymer in the material;

(b) starting substances, organic cementitious constituents, substance constituents and non-
intentionally added species originating from the starting substance or organic cementitious constituent
which show one of the human health hazards referred to in Section 1.1. of Annex VI irrespective of
their levels of migration;

(c) starting substances, organic cementitious constituents, substance constituents and non-
intentionally added species originating from a starting substance or organic cementitious constituent
which do not fall under points (a) or (b) and which have been tested in accordance with Table 1 [of the
legal act] and have been found to migrate into water intended for human consumption with a
concentration at the tap (Ctap) exceeding 0,1 µg/l;

(d) metallic composition constituents or impurities, which have been tested in accordance with Table 1
[of the legal act];

(e) enamel, ceramic or other inorganic composition constituents or impurities of an enamel, ceramic
or other inorganic composition which have been tested in accordance with Table 1 [of the legal act].

Based on the Ctap of a relevant chemical species in the water intended for human consumption,
the applicant must submit toxicological information, as required under Section 2 of Annex VI to
Commission Implementing Decision (EU) 2024/365. This information, together with any other
relevant technical or scientific information that is available, must be considered in order to
perform a robust risk assessment.

In the context of the DWD, the procedure for the human health risk assessment of relevant
chemical species consists of comparing the Ctap derived based on the migration studies
(‘exposure’) with the MTCtap, an estimate of the amount of a substance in drinking water that
can be ingested daily over a lifetime without appreciable health risk. For threshold effects, the
MTCtap is based on the derived no-effect level (DNEL), which is calculated on the basis of
threshold levels such as the benchmark dose (BMD), no observed adverse effect level (NOAEL),
lowest observed adverse effect level (LOAEL) and no observed adverse effect concentration
(NOAEC), with the use of assessment factors (AFs), and which constitute the outcome of the
hazard characterisation. The NOAEL and/or LOAEL (N(L)OAEL) values are determined based on
results from animal testing or on the basis of available human data. For non-threshold effects
(e.g. for mutagenicity or carcinogenicity), a DNEL cannot be derived. For relevant chemical
species exerting toxicity with a non-threshold mode of action, an MTCtap of 0.1 µg/l should be
applied (further explanation is provided in Section 6.3).

The derivation and use of dose–response relationships for each of the effects to be assessed are
discussed in detail in Chapter 3 of this guidance document.
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In their risk assessment, an applicant must identify any significant uncertainty (e.g. uncertainties
related to the selection of the most relevant critical toxicological effect or to an allocation factor
(ALF)) and adequately address them (e.g. by applying a reasonable worst-case approach) if
needed to ensure the protection of human health.

For both the hazard assessment and exposure assessment, data on physico-chemical properties,
including chemical reactivity, are needed. These data are required, for example, to accurately
design and assess both the migration tests and toxicity studies. Physico-chemical properties may
also provide indications of the absorption of the relevant chemical species after exposure via the
oral route (the relevant route for drinking water). Chemical reactivity may be of importance, for
example in the estimation of the exposure to the relevant chemical species, and also has an
impact on its toxicokinetics and its metabolism. Information on the migration of the applied-for
entry, the identification of the relevant chemical species and their migration concentrations into
the test water is also needed, because it is used to determine the toxicological requirements for
each relevant chemical species, as explained in DWD Guidance Volume I and according to the
principles set out in Section 2 of Annex V to Commission Implementing Decision (EU) 2024/365.

It is the responsibility of the applicant to secure any necessary rights to use any third-party
intellectual property in their application. As set out in the terms and conditions of the ECHA
submission tool ( 1), when submitting the application, the applicant is required to warrant that
they have obtained all the necessary permission, rights or licences to use the third-party
intellectual property for the purpose of their application. The fact that a study or any other third-
party material is available on ECHA’s website does not mean that this material is not subject to
the third-party’s intellectual property rights or that it may be freely used without their
permission.

(1) The tool will be available on ECHA’s website.

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3. Hazard assessment: hazard identification

The principles of hazard assessment are stated in Part 2 of Section 2 of Annex VI to Commission
Implementing Decision (EU) 2024/365.

For each relevant chemical species, the hazard assessment process in relation to human health
must entail assessment of effects, comprising the following.

• Hazard identification. The aim of hazard identification is to identify the effects of concern
that a relevant chemical species has an inherent capacity to cause.

• Hazard characterisation: dose (concentration)–response (effect) assessment. This

is the estimation of the relationship between dose, or level of exposure to a substance, and
the incidence and severity of an effect. In this chapter it is referred to as ‘dose–response’.
At this step the NOAEL, or, where it is considered more appropriate, the LOAEL or the BMD,
is determined based on the observed effects, and a DNEL is derived. If appropriate, the shape
of the dose–response curve should also be considered.

The conditions under which hazard assessment does not apply to a relevant chemical species
are stated under Section 1 of Annex VI to Commission Implementing Decision (EU) 2024/365.

During both steps of the hazard assessment, it is very important to evaluate the available data
for adequacy and completeness. The evaluation of adequacy must address the reliability and
relevance of the data.

For the effects of concern for which it is not possible to determine a N(L)OAEL, it is generally
sufficient to evaluate whether the relevant chemical species has an inherent capacity to cause
such an effect. Where for such an effect it is possible to draw a relationship between the dose
or concentration of the relevant chemical species and the severity of an adverse effect, this
relationship should be determined.

Human data are in principle the most relevant source of information on human toxicity; however,
since there may be limitations in terms of the reliability of human studies (see Section 3.1.2.2.3),
they are normally considered together with animal and other data. Studies conducted with
human volunteers are problematic from an ethical point of view. Results from such studies should
be used only in justified cases (e.g. tests that were conducted for the authorisation of a medicinal
product or when effects in already available human volunteer studies of existing substances have
been observed to be more severe than deduced from prior animal testing). The potential
differences in sensitivity of human studies and studies in animals should be taken into account
in the hazard assessment, on a case-by-case basis. In relation to hazard identification, the
relative lack of sensitivity of human data may cause particular difficulty: negative data from
studies in humans will not usually be used to override the classification of substances that have
been classified on the basis of data from studies in animals in accordance with the criteria given
in Regulation (EC) No 1272/2008 on classification, labelling and packaging of substances and
mixtures (the CLP regulation), unless the classification is based on an effect that clearly would
not be expected to occur in humans. For more information, see Guidance on the Application of
the CLP Criteria ( 2).

(2) Guidance on the Application of the CLP Criteria Overview.

 Guidance on the DWD: Volume II
18 Version 1.0 – January 2025

The structure of the section on hazard identification for each end point is as follows:

• definition of the effect of concern;

• data to be used in the hazard assessment;

• remaining uncertainty;

• conclusion on risk assessment.

For hazard identification, DWD Guidance Volume I needs to be considered together with this
guidance and that in Introductory Guidance on the CLP Regulation ( 3). The CLP regulation,
Regulation (EC) No 1272/2008 ( 4), is an umbrella for assessing hazards under EU chemical
legislation and provides general rules that are relevant for hazard assessment under the DWD
(e.g. a substance is classified as carcinogenic unless there is ‘strong evidence’ that the
mechanism is not relevant for humans (CLP regulation, Annex I, Section 3.6.1.1)). It must be
noted, however, that the hazard assessment under the DWD does not aim to classify substances
according to CLP regulation criteria.

For the purpose of hazard assessment, the stepwise approach summarised in Figure 1 (for
starting substances and organic cementitious constituents) and Figure 2 (for compositions) has
to be followed. Steps 1–3 are taken first to identify if hazard assessment is needed. If it is, then
steps 4 and 5 have to be followed. They include the collection of all available information and its
assessment before deciding if additional testing needs to be performed. Once new test results
become available, as part of step 6 using DWD Guidance Volume I, these results should be
evaluated in accordance with the guidance in this chapter.

(3) https://echa.europa.eu/documents/10162/2324906/clp_introductory_en.pdf/b65a97b4-8ef7-4599-b122-
7575f6956027?t=1547546145023.
(4) https://eur-lex.europa.eu/eli/reg/2008/1272/oj.
Guidance on the DWD: Volume II
Version 1.0 – January 2025 19

Figure 1. Schematic representation of the stepwise approach for fulfilling the testing
methodology for human health hazard for starting substances and constituents, for the purpose
of the DWD

Step 1
Does the relevant chemical species fall under any of the cases that justify the adoption of a limited acceptance methodology,
as described in Section 1 of Annex VI to Commission Implementing Decision (EU) 2024/365?

Yes
No
No or limited, existing toxicological
Step 2 information is required.
Estimate migration and identify the Refer to Section 6.2 of DWD Guidance
relevant chemical species and their Ctap < 0.1 µg/l
Volume II for deriving an appropriate
migration tier MTC tap value
(Chapter 5 of DWD Guidance Volume I)

Step 3
Is the relevant chemical
Yes species a monomer or other
Ctap ≥ 0.1 µg/l
reactant of a main
monomer?
No

Step 4
Application for listing on the
Collect all available information on
EUPL may not be necessary.
toxicological properties of the relevant
Refer to Section 2.1.4 of
species
DWD Guidance Volume III
(Chapter 3 of DWD Guidance Volume II)

Step 5
Evaluate all relevant available information
(Chapters 3–4 of DWD Guidance Volume II)

Step 6
Generate any missing toxicological
information to fulfil the information
requirements in Annex V to Commission
Implementing Decision (EU) 2024/365
(Chapter 6 of Guidance Volume I)

Step 7
Evaluate the new toxicological information
(Chapters 3–4 of DWD Guidance Volume II)

Use available toxicological data to derive an

appropriate MTCtap value for the relevant
chemical species
(Chapter 6 of DWD Guidance Volume II)

NB: Ctap, concentration at the tap; MTCtap, maximum tolerable concentration at the tap.
 Guidance on the DWD: Volume II
20 Version 1.0 – January 2025

Figure 2. Schematic representation of the stepwise approach for fulfilling the testing
methodology for human health hazard for compositions, for the purpose of the DWD

Step 1
Estimate migration according to Annex IV to Commission Implementing Decision (EU) 2024/365 to
establish Ctap for the constituent or impurity present in the composition
(Chapter 5. of DWD Guidance Volume I)

Step 3
Step 2
Review the available toxicological
Does the constituent/impurity have a MTCtap
YES information on the constituent/impurity to
value in Annex V to Commission Decision
verify if the existing MTCtap value is still valid.
Implementing Decision (EU) 2024/367?
Is there new toxicological information?

NO

Step 4
Collect all available information on
toxicological properties
(Chapter 3 of DWD Guidance Volume II) NO

Step 5
Evaluate all relevant available information
(Chapters 3–4 of DWD Guidance Volume II)
No toxicity data
required

Step 5 YES
Perform new testing if needed to fulfil the
information requirements of
Annex V to Commission Implementing
Decision (EU) 2024/365
(Chapter 6 of DWD Guidance Volume I)

Step 6
Evaluate new information
(Chapters 3–4 of DWD Guidance Volume II)

Use available toxicological data to calculate Use existing MTCtap value shown in Annex V
MTC tap to Commission Implementing Decision (EU)
(Chapter 6 of DWD Guidance Volume II) 2024/367

NB: Ctap, concentration at the tap; MTCtap, maximum tolerable concentration at the tap.
Guidance on the DWD: Volume II
Version 1.0 – January 2025 21

3.1. Collection and evaluation of all available data

3.1.1. Collection of data (step 4 in Figure 1 and Figure 2)

For step 4 of the process, various sources exist for gathering all available information on the
relevant chemical species. The eChemPortal ( 5) and the QSAR Toolbox ( 6) are recommended for
the collection of existing information on toxicological properties and for the determination of
potential application of non-test methods in the hazard assessment. Literature databases should
also be considered. An additional list of sources to be considered during step 4 is available in
Guidance on Information Requirements and Chemical Safety Assessment – Chapter R.3:
Information gathering ( 7).

3.1.2. Evaluation of the data (step 5 in Figure 1 and Figure 2)

During the hazard assessment, it is very important to evaluate the data with regard to their
adequacy and completeness. This is particularly important for well-studied existing substances,
for which there may be a number of test results available for each effect but some or all of these
tests may not have been carried out to current standards. This section puts forward general
guidelines on data evaluation. The term ‘adequacy’ is used here to cover the reliability of the
available data and the relevance of those data for human hazard and risk assessment. In addition
to the guidance provided in this section, Guidance on Information Requirements and Chemical
Safety Assessment – Chapter R.4: Evaluation of available information ( 8), provides further
guidance on assessing the relevance, reliability and adequacy of the information.

3.1.2.1. Completeness of data

Section 2 of Annex V to Commission Implementing Decision (EU) 2024/365 gives the legal basis
for toxicity data requirements for the relevant chemical species. Part 2 of Section 2 of Annex V
to that decision specifies the general rules for the adaptation of the data requirements.

3.1.2.2. Adequacy of data

The adequacy of data can be defined by two basic elements:

• reliability, covering the inherent quality of a test relating to test methodology and the way
that the performance and results of the test are described;

• relevance, covering the extent to which a test is appropriate for a particular hazard or risk
assessment.

Reliable, relevant data can be considered valid for use in the risk assessment. When there is
more than one set of data for each effect, the greatest weight is attached to the most reliable
and relevant.

(5) http://www.echemportal.org.
(6) http://www.qsartoolbox.org.
(7) https://echa.europa.eu/documents/10162/17235/information_requirements_r3_en.pdf/41895234-1125-4977-
b058-50a98e36fa48.
(8) https://echa.europa.eu/documents/10162/17235/information_requirements_r4_en.pdf/d6395ad2-1596-4708-
ba86-0136686d205e
 Guidance on the DWD: Volume II
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The evaluation of animal test data with respect to reliability is outlined below.
Sections 3.1.2.2.3–3.1.2.2.5 consider issues specific to the reliability of human and in vitro data,
relevance to humans and the (quantitative) structure–activity relationship ((Q)SAR).
3.1.2.2.1. Reliability of data

For relevant chemical species for which tests conducted in accordance with the EU test methods
regulation (Regulation (EC) No 440/2008) and in compliance with the principles of good
laboratory practice (GLP) are available, many of the issues addressed in this section will not be
relevant.

For some substances, the test data available were generated prior to the requirements of GLP
and the standardisation of testing methods. Those data may still be used for risk assessment,
but the data and the methodology used must be evaluated in order to determine their reliability
for assessment purposes. The evaluation needs expert judgement and must be transparent, so
that the use made of a particular dataset is clearly justified. The requirements of the appropriate
standardised test method and GLP principles should be regarded as a reference when evaluating
the available test data. That is, studies carried out in accordance with current methods (e.g. the
EU test methods regulation, Organisation for Economic Co-operation and Development (OECD)
test guidelines programme ( 9) or United States Environmental Protection Agency (U.S. EPA) Test
Guidelines ( 10), appropriately reported, should be considered the most reliable for risk
assessment. Klimisch et al. (1997) developed the following scoring system to assess the
reliability of data, particularly from toxicological studies.

1 = Reliable without restrictions, which refers to ‘studies or data … generated according to

generally valid and/or internationally accepted testing guidelines (preferably performed
according to GLP) or in which the test parameters documented are based on a specific (national)
testing guideline … or in which all parameters described are closely related/comparable to a
guideline method’.

2 = Reliable with restrictions, which refers to ‘studies or data … (mostly not performed
according to GLP), in which the test parameters documented do not totally comply with the
specific testing guideline, but are sufficient to accept the data or in which investigations are
described which cannot be subsumed under a testing guideline, but which are nevertheless well
documented and scientifically acceptable’.

3 = Not reliable, which refers to ‘studies or data … in which there are interferences between
the measuring system and the test substance or in which organisms / test systems were used
which are not relevant in relation to the exposure (e.g., unphysiological pathways of application)
or which were carried out or generated according to a method which is not acceptable, the
documentation of which is not sufficient for an assessment and which is not convincing for an
expert judgment’.

4 = Not assignable, which refers to studies or data ‘ … which do not give sufficient
experimental details and which are only listed in short abstracts or secondary literature (books,
reviews, etc.)’.

The use of such scoring tools (e.g. the abovementioned Klimisch codes) allows the information
to be ranked and organised for further review. This implies focusing on the most relevant tools,
taking into account the hazard being measured or estimated. The evaluation of reliability is
performed based on certain formal criteria, using international standards as references. The

(9) http://www.oecd.org/env/ehs/.
( ) https://www.epa.gov/test-guidelines-pesticides-and-toxic-substances/final-test-guidelines-pesticides-and-toxic.
10
Guidance on the DWD: Volume II
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scoring of information (e.g. using the Klimisch codes) should not exclude all unreliable data from
further consideration by expert judgement because of the possible pertinence of these data in
relation to the evaluated hazard. In general, some types of data that are not reliable (i.e. those
for which documentation is not sufficient to make an assessment) and data for which it is not
possible to assign reliability may be used only as supporting data.

When looking at a test report, the assessor should consider the following.

• Whether the purity/impurities and the origin of the test material are reported.

• Whether a complete test report is available or the test has been described in sufficient detail,
and whether the test procedure described is in accordance with the relevant guidelines and/or
generally accepted scientific standards. If the information is considered to be reliable, it can
be used for the hazard and risk assessment.

• Whether the reliability of the data cannot be fully established or the test procedure described
differs in some respects from the test guidelines and/or generally accepted scientific
standards. The assessor must decide in those cases whether the data will be taken into
consideration in the hazard and risk assessment and how they will be used (e.g. as supporting
information where a reliable study has already been identified) or whether they should be
regarded as invalid.

• Whether the following factors, among others, can be used to support the view that these
data may be acceptable for use in a hazard and risk assessment:

o there are other studies or calculations available on the relevant chemical species, and the
data under consideration are consistent with them;

o other studies, for example on isomers with similar structure activity profiles, homologues,
relevant precursors, breakdown products or other chemical analogues, are available, and
the data under consideration are consistent with them;

o an approximate value is sufficient for taking a decision on the result of the risk
characterisation.

In principle, the same criteria apply to test data reported in the published literature. The
amount of information presented will provide the basis for deciding on the reliability of the
data reported. In general, publications in peer-reviewed journals are preferable. High-quality
reviews may be used as supporting information. Summaries or abstract publications may also
provide supporting material.

3.1.2.2.2. Relevance of data

In order to evaluate the relevance of the available data, it is necessary to judge, among other
things, if an appropriate species has been studied, if the route of exposure is relevant for the
population and if the substance tested is representative of the substance as supplied. For
relevant chemical species identified in the water intended for human consumption, the oral route
is the only relevant route of exposure.

Relevant human data of an adequate quality can sometimes be the best available data; however,
more frequently, the available human, animal and non-animal data are considered together in
order to reach a conclusion about the relevance to humans of effects observed in studies in
animals.
 Guidance on the DWD: Volume II
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The evaluation of the relevance for humans of data from studies in animals is aided by use of
data on the toxicokinetics, including metabolism of a substance in both humans and the animal
species used in the toxicity tests, even when these data are relatively limited. Clear, well-
documented evidence of a species-specific effect/response (e.g. light hydrocarbon-induced
nephropathy in the kidneys of male rats) should be used as justification for the conclusion that
a particular effect is not expected to occur in humans exposed to the relevant chemical species.

In the absence of such information (on the relevant chemical species itself or, if it can be
scientifically justified, on a close structural analogue), ‘threshold’ adverse effects observed in
studies in animals will normally be assumed to be likely to also occur in humans exposed to the
relevant chemical species above a certain level of exposure.

In any case, the dose–response relationships in the animal studies (or the severity of the effect,
when only a single dose was tested) are also assessed as a part of the hazard and risk
assessment process. These assessments are taken into account at the risk characterisation stage
when a judgement is made on the likelihood of occurrence of an adverse effect in humans at a
particular level of exposure.

Interpretation of the relevance of data derived from tests conducted in vitro should be taken into
account whether the results seen have been observed or could be expected to occur (e.g. from
knowledge of the toxicokinetics of the substance) in vivo. According to the validation procedures
established by the European Centre for the Validation of Alternative Methods (ECVAM), the
relevance of an alternative (non-animal) test, such as an in vitro test, is assessed according to
the scientific basis of the test system (scientific relevance) and the predictive capacity (predictive
relevance) of the prediction model, which is an algorithm for extrapolating from in vitro data to
an in vivo end point (Worth and Balls, 2001).

The results of in vitro tests (in addition to the standard test guideline protocols for the
assessment of specific end points, such as mutagenicity, that are mentioned in DWD Guidance
Volume I) may provide helpful information, which, for instance, may be used to facilitate the
interpretation of the relevance for humans of data from studies in animals or to gain a better
understanding of the mechanism of action of a substance.

The guidance given above on the evaluation of the adequacy (relevance and reliability) of
information in general covers mainly the evaluation of data from animal studies. Some additional
specific guidance is given below for human data, in vitro data and (Q)SARs.

3.1.2.2.3. Human data

The evaluation of human data usually requires a thorough critical assessment of the reliability
of the data, considering some unique aspects distinct from the evaluation of animal data.
Human data in the form of epidemiological studies or case reports can contribute to the hazard
identification process as well as to the risk assessment process itself.
Criteria for assessing the adequacy of epidemiology studies include an adequate research design,
the proper selection and characterisation of the exposed and control groups, adequate
characterisation of exposure, sufficient length of follow-up for the disease as an effect of the
exposure to develop, valid ascertainment of effect, proper consideration of bias and confounding
factors, proper statistical analysis and a reasonable statistical power to detect an effect. These types
of criteria have been described in more detail and can be derived from epidemiology textbooks
(Hernberg, 1991; Checkoway et al., 2004; Rothman and Greenland, 2008).
The results from human experimental studies, which are often not designed for the purpose of
hazard identification and hence may lack sensitivity, may often be limited by several factors,
Guidance on the DWD: Volume II
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such as a relatively small number of subjects, short duration of exposure and low dose levels,
resulting in poor sensitivity in detecting effects.
Therefore, epidemiological studies with negative results cannot prove the absence of an intrinsic
hazardous property of a substance, but well-documented ‘negative’ studies of good quality may
be useful in the hazard assessment. Four major types of human data may be submitted: (1)
analytical epidemiology studies on exposed populations, (2) descriptive or correlation
epidemiology studies, (3) case reports and (4) in very rare, justified cases, controlled studies
involving human volunteers.
(1) Analytical epidemiology studies are useful for identifying a relationship between human
exposure and effects such as biological effect markers, early signs of chronic effects, disease
occurrence and increased mortality rates, and may provide the best data for hazard and risk
assessment. Study designs include:

• case–control (case–referent) studies, where groups of individuals with (cases) and without
(controls/referents) a particular effect are identified and compared to determine differences
in exposure;

• cohort studies, where groups of ‘exposed’ and ‘non-exposed’ individuals are identified and
differences in effect occurrence are studied;

• cross-sectional studies, where a population (e.g. a workforce) is studied so that morbidity at

a given point in time can be assessed in relation to concurrent exposure.

The strength of the epidemiological evidence for specific health effects depends on, among other
things, the type of analyses and the magnitude and specificity of the response. Confidence in
the findings is increased when comparable results are obtained in several independent studies
on populations exposed to the same agent under different conditions and using different study
designs.

Criteria for assessing the adequacy of epidemiology studies include the proper selection and
characterisation of the exposed and control groups, adequate characterisation of exposure,
sufficient length of follow-up for disease occurrence, valid ascertainment of effect, proper
consideration of bias and confounding factors and a reasonable statistical power to detect an
effect.

(2) Descriptive epidemiology studies examine differences in disease rates among human
populations in relation to age, gender, race and differences in temporal or environmental
conditions. These studies are useful for identifying areas for further research but are not very
useful for hazard and risk assessment. Typically, these studies can identify patterns or trends in
disease occurrence over time or in different geographical locations but cannot ascertain the
causal agent or degree of human exposure.

(3) Case reports describe a particular effect in an individual or a group of individuals who were
exposed to a substance. They may be particularly relevant when they demonstrate effects that
cannot be observed in experimental animal studies.

(4) When they are already available, well-conducted controlled human exposure studies in
volunteers, including low-exposure toxicokinetic (TK) studies, can also be used in hazard and
risk assessment in some rare cases. Those studies have to be conducted in line with the World
Medical Association Declaration of Helsinki, which describes the general ethical principles for
medical research involving human subjects.
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It should be emphasised that testing with human volunteers is strongly discouraged; however,
when there are good quality data already available, they should be used as appropriate, in well-
justified cases.

Experimental human toxicity studies must not be conducted specifically for the purpose of
inclusion in an EUPL.

3.1.2.2.4. In vitro data

It can be expected that some of the available data will have been derived from studies conducted
in vitro – the basic (and perhaps additional) studies on genotoxicity, for example. There may
also be data from in vitro studies on, for instance, metabolism and/or mechanisms of action
(including studies in cell cultures from different species), and various aspects of toxicity (e.g.
tests for cytotoxicity in different types of cells, macromolecule binding studies, tests using
embryo culture systems, sperm motility tests). For any of these studies, their usefulness will be
influenced by their adequacy in the light of some of the general criteria already discussed, for
example how well the study is reported, how well the test material is characterised and to what
extent the requirements of the method described in the EU test methods regulation (Regulation
(EC) No 440/2008 ( 11)) and the principles of good in vitro method practice ( 12) have been met
for the end point under consideration.

However, there are also some criteria that need particular attention when assessing the
adequacy of in vitro studies, for example:

• the range of exposure levels used, taking into account the toxicity of the substance towards
the bacteria/cells, its solubility and, as appropriate, its effect on the pH and osmolality of the
culture medium;

• the maintenance of effective concentrations of the volatile substances in the test system;

• the use of an appropriate exogenous metabolism mix (e.g. S9 from induced rat liver or
hamster liver) when necessary;

• the use of appropriate negative and positive controls as integral parts of the tests;

• the use of an adequate number of replicates (within the tests and of the tests);

• the use of the appropriate test system (e.g. appropriate cell lines).

In studies conducted in accordance with, for instance, the EU test methods regulation, OECD
test guidelines programme ( 13) or US EPA test guidelines ( 14), the above criteria are evaluated
against the specifications provided in the test guidelines.

3.1.2.2.5. (Quantitative) structure–activity relationships

When data do not exist for a given end point, or when data are limited, the use of structure–
activity relationships (SARs) may be considered. It should be noted that (Q)SAR models are

(11) https://eur-lex.europa.eu/legal-content/EN/ALL/?uri=celex%3A32008R0440.
(12) https://www.oecd.org/env/guidance-document-on-good-in-vitro-method-practices-givimp-9789264304796-
en.htm.
(13) http://www.oecd.org/env/ehs/.
(14) http://www.epa.gov/ocspp/pubs/frs/home/guidelin.htm.
Guidance on the DWD: Volume II
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currently not well developed for application in hazard and risk assessment, especially in relation
to long-term mammalian toxicity. The (Q)SARs that are used for risk assessment purposes are
usually of a more qualitative nature and do not address quantitative aspects. (Q)SARs may be
of value in indicating a potential hazard, TK properties or the need for further testing. Additional
guidance is provided in Guidance on Information Requirements and Chemical Safety
Assessment – Chapter R.6: QSARs and grouping of chemicals ( 15).

(Q)SAR predictions can be used together with other information in the weight of evidence (WoE)
determination. When using (Q)SARs to predict a substance property, an assessment of both the
model and the prediction is needed. A (Q)SAR model must be scientifically valid (using OECD
principles (OECD, 2004, 2007)), and adequate and reliable documentation must be provided. A
valid (Q)SAR model does not necessarily produce an acceptable prediction. For an acceptable
(Q)SAR prediction, the input is correct, the substance falls within the applicability domain of the
model, the prediction is reliable and the outcome is fit for the regulatory purpose. The validity
of models and predictions can be assessed by using the OECD (Q)SAR assessment framework
(OECD, 2023).

Transparent documentation of the validity of the models must be provided as well as information
relevant for judging the reliability of predictions for individual compounds or other comparable
documentation. A (Q)SAR model reporting format displays a description of the (Q)SAR model
relative to the five OECD (Q)SAR validation principles in a systematic and summarised way
(OECD, 2004, 2007; minor update: OECD, 2023). The information about the (Q)SAR prediction
is reported in the (Q)SAR prediction reporting format. An updated (Q)SAR prediction reporting
format template was published in 2023, and it reflects the newly established OECD (Q)SAR
prediction principles (OECD, 2023) ( 16).

More information can be found in OECD (Q)SAR assessment framework documents (e.g. OECD,
2023), in the ECHA guidance on (Q)SARs and grouping of chemicals (Chapter R.6 ( 17)) and in
the ECHA practical guide How to Use and Report (Q)SARs ( 18).

(Q)SAR models for predicting genotoxicity. In the context of the DWD, results from (Q)SAR
models that met all conditions described under Section 1.3 of Annex XI to Regulation (EC)
No 1907/2006 on the registration, evaluation, authorisation and restriction of chemicals (the
REACH regulation) may be used as an alternative to experimental testing for genotoxicity, only
for relevant chemical species constituents or a non-intentionally added species, if it can be
justified that the testing is technically not possible (e.g. it cannot be isolated and tested as such).
In such cases, the recommendations described below have to be considered.
There are hundreds of (Q)SAR models available in the literature for predicting genotoxicity test
results (Honma, 2020). There are local (Q)SAR models, for relatively small sets of congeneric
substances, that is, substances with the same basic structure and same mechanism of action,
and global (Q)SAR models for a wide variety of non-congeneric substances. Global models may
provide valuable predictive tools for estimating a number of mutagenic end points, if essential
features of the information domain are clearly represented. However, the quality of reporting
varies from model to model, and predictivity must be assessed case by case on the basis of clear
documentation.

(15) http://echa.europa.eu/guidance-documents/guidance-on-information-requirements-and-chemical-safety-
assessment.
(16) https://one.oecd.org/document/ENV/CBC/MONO(2023)32/ANN2/en/pdf.
(17) https://echa.europa.eu/documents/10162/17224/information_requirements_r6_en.pdf.
(18) https://echa.europa.eu/documents/10162/17250/pg_report_qsars_en.pdf/407dff11-aa4a-4eef-a1ce-
9300f8460099.
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A list of the available (free and commercial) predictive software for ecotoxicological, toxicological
and environmental end points, including mutagenicity models, has been compiled within the
frame of an EU project ( 19). This website contains information on about 450 models, sorted by
end point. The Joint Research Centre (JRC) website hosts the JRC (Q)SAR model inventory,
which is an inventory of information on (Q)SAR models that have been submitted to the JRC ( 20).
This inventory currently contains a list of 154 models, of which 21 cover mutagenicity end points.

The bacterial reverse mutation (Ames test) is the most commonly predicted genotoxicity end
point for global models. The Ames test is reported to be well predicted, while the reliability of
the (Q)SAR models for other genotoxicity assays / end points is still quite far from optimal
(Benigni et al., 2020). However, for certain classes of chemicals, the predictivity of other
genotoxicity assays / end points can be quite good, averaging the performance for predicting
the reverse mutation test (Ames test), while for other chemical classes the predictive statistics
are poor, possibly also due to the poor availability of for these chemical classes. Recently, there
has been significant progress in improving the predictive ability of (Q)SAR models, which have
achieved high predictive power, including for predicting clastogenicity (Morita, 2019; van
Bossuyt, 2020).

As a general trend, the combination of (Q)SAR models increases sensitivity, but at the expense
of specificity.

The Danish Environmental Protection Agency and the (Q)SAR group at the National Food
Institute at the Technical University of Denmark (DTU) have developed the Danish (Q)SAR
Database, which contains predictions from a number of mutagenicity models. In addition to
assorted reverse mutation test (Ames test) models, the database provides predictions of the
following in vitro end points: chromosomal aberrations (CHO and CHL cells), mouse
lymphoma/tk, CHO/Hprt gene mutation assays and unscheduled DNA synthesis (rat
hepatocytes). It also provides predictions of the following in vivo end points: Drosophila SLRL,
mouse micronucleus, rodent dominant lethal, mouse sister chromatid exchanges in bone marrow
and mouse comet assay. The Danish (Q)SAR Database is a repository of model predictions from
more than 200 (Q)SAR models and is considered a good screening tool. All organic single-
constituent substances that were preregistered or registered under the REACH regulation
(around 93 000) are included in the structure set. In addition, chemical structures from other
relevant databases are included, leading to a structure set of more than 650 000 total / 600 000
unique chemical structures. When possible, the end points have been modelled in three software
systems, Leadscope, CASE Ultra and SciQSAR. Most DTU-developed models and a number of
commercial models from MultiCASE have been modelled in two or three systems. For the set
structure, predictions are provided in the different systems separately and as an overall battery
prediction. A user manual with information on the individual models, including training set
information and validation results, is available at the Danish (Q)SAR Database website. (Q)SAR
model reporting format documentation of all DTU-developed and commercial models is also
available in the database. Predictions from a number of OECD QSAR Toolbox profilers have been
included as supporting information to the (Q)SAR predictions. The database also includes
predictions from other software (e.g. VEGA). The Danish (Q)SAR Database ( 21) is freely
accessible and is also integrated into the OECD QSAR Toolbox.

The OECD QSAR Toolbox ( 22) is freely available software developed by the Laboratory of
Mathematical Chemistry, Burgas University, Bulgaria, driven by and in collaboration with ECHA,

(19) https://www.life-concertreach.eu/results/results-gateway/.
(20) https://jeodpp.jrc.ec.europa.eu/ftp/jrc-opendata/EURL-ECVAM/datasets/QSARDB/LATEST/qsardb.html.
(21) https://qsar.food.dtu.dk/.
(22) https://qsartoolbox.org/.
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OECD and the OECD (Q)SAR management group. It is an expert system that supports
reproducible and transparent chemical hazard assessment. It offers functionalities for retrieving
experimental data, simulating metabolism and profiling properties of chemicals. These pieces of
information and tools can also be used to identify structurally and mechanistically defined
analogues and chemical categories, which can serve as sources for read-across and trend
analysis.

Concerning mutagenicity, the OECD QSAR Toolbox covers information relevant for the bacterial
reverse mutation test (Ames test), in vitro chromosomal aberration, in vivo chromosomal
aberration (micronucleus test) and genotoxic carcinogenicity end points. The OECD QSAR
Toolbox includes a number of databases with relevant experimental data: REACH (ECHA), Food
TOX Hazard (European Food Safety Authority (EFSA)), bacterial mutagenicity (In vitro
Salmonella typhimurium mutagenicity (Isssty)), genotoxicity and carcinogenicity (ECVAM),
genotoxicity (Outcome and Assessment Information Set (OASIS)), genotoxicity pesticides
(EFSA), micronucleus (In vitro micronucleus test (Issmic)), micronucleus (OASIS), Toxicity
Japan (Ministry of Health. Labour and Welfare (MHLW)) and the Transgenic Rodent Database.
The OECD QSAR Toolbox allows the user to mine these databases for the purpose of (computer-
assisted) read-across analysis, looking for relevant structural analogues in these databases. It
also offers profilers for predicting modes of actions that are relevant for mutagenicity, such as
profilers for DNA binding by OASIS and the OECD (Q)SAR Toolbox, DNA alerts for Ames,
chromosomal aberration and micronucleus test by OASIS, and in vitro and in vivo mutagenicity
alerts by Intercom Station Software (ISS). Data and profilers can be used in combination to
identify mechanistically and structurally relevant analogues for read-across predictions. Profilers
identify structures associated with a hazard and are not predictive of the hazard or lack of hazard
per se. There can be additional factors, such as metabolism, clearance and bioavailability, that
may contribute to the genotoxicity potential of the substance. In addition, the profilers are often
not associated with defined applicability domains. Unbiased subcategorisation is extremely
important when using the profilers in the OECD QSAR Toolbox.

In addition to the data and profilers, the OECD QSAR Toolbox includes predictions of
mutagenicity from a number of (Q)SAR models from the Danish (Q)SAR Database.

The Danish (Q)SAR Database and the OECD QSAR Toolbox profilers can support predictions for
multiple genotoxicity end points and may assist those making WoE decisions regarding the
mutagenic potential of untested substances.

For mutagenicity predictions, the potential of the substances to generate metabolites of concern
should also be considered. Some models for predicting genetic toxicity include a metabolic
simulator and prediction of metabolites for genotoxicity potential. For models not including a
metabolic simulator, separate in silico approaches to predict the likely metabolites based on
molecular structure are available. Reliability of simulated metabolism is discussed in Dermen et
al. (2022). In some cases, models may implicitly assume metabolism, such as in the use of the
aromatic amine alert (OECD QSAR Toolbox, Derek Nexus, VEGA, MultiCASE) as an indication of
mutagenic potential. Metabolism can also explain some differences between in vitro and in vivo
genotoxicity (Petkov et al., 2022). In addition, a new (Q)SAR modelling concept was introduced
in which the potency is related to the amount of DNA adducts as an element of metabolism
prediction.
When using (Q)SARs to predict a relevant chemical species property, an assessment of both the
model and the prediction is needed. Further guidance on (Q)SARs can be found in Section R.6.2
of Guidance on Information Requirements and Chemical Safety Assessment – Chapter R.6:
QSARs and grouping of chemicals and in the OECD (Q)SAR assessment framework, which
provides guidance and practical advice on how to assess the validity of models and their
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predictions (OECD, 2023) ( 23). The assessment framework states that the validity of a model
should be assessed according to the OECD validation principles for (Q)SARs (OECD, 2004, 2007),
while the validity of predictions can be assessed against the newly established principles for the
assessment of (Q)SAR predictions and results presented in the assessment framework
guidance ( 24).
For prediction of gene mutation in bacteria, it is important that all relevant strains of the Ames
test are addressed. Metabolic activation should be taken into account for adequacy and
equivalence to tests (in order to make the predictions suitable for meeting the information
requirements of the DWD). For statistical models, verification that the substance falls within the
applicability domain and information on analogues supporting the predictions are important.
Negative predictions from an alert-based system such as Derek Nexus, for example, can be
considered only in the vicinity of very similar compounds that tested negative in the respective
tests, assuming that all strains and metabolic activation are covered. These can be searched
outside Derek Nexus (e.g. using Sarah Nexus, Vitic or another database source, such as the
OECD QSAR Toolbox ( 25)). The documentation that Derek Nexus provides in the results window,
however, is currently not sufficient to assess a prediction.

The Danish (Q)SAR Database includes statistical models from SciQSAR, Leadscope and CASE
Ultra models for the Ames test. However, the documentation for the Danish (Q)SAR Database
does not always allow users to verify that the substance falls within the applicability domain and
verify information from analogue substances to support the prediction. Information on the
models is available from the website ( 26).

If well-documented and applicable (Q)SAR data are available, they should be used to help reach
the decision points described below in this section. In many cases, the accuracy of such methods
will be sufficient to enable either a test or a specific regulatory decision to be made. In other
cases, the uncertainty may be considered too high, and further data will be needed.
The accuracy of available methods may be assessed using substances that were not originally
included in the training set of the models (so-called external validation). There was an
international challenge project for predicting the bacterial reverse mutation test (Ames test), the
results of which are described in Honma et al. (2019). Evaluating new data on mutagenicity can
lead to expansion of the training sets (Amberg et al., 2019; Petkov et al., 2019a).
Documentation can include reference to a related and relevant substance or group of substances
that leads to the conclusion of concern or lack of concern. This can be presented in accordance
with a scientific logic (read-across justification) or sometimes as a mathematical relationship of
chemical similarity. It should be noted that when an in silico tool such as the OECD QSAR Toolbox
is used as a first step, to find analogues for performing read-across of data but not to make
predictions based on (Q)SAR, this is considered a read-across approach and, as such, the
justification and predictions should comply with the conditions of a read-across adaptation (see
Section 3.1.6).

The lack of mechanistic justification may, depending on the application context, limit the use of
(Q)SAR predictions, particularly for regulatory decisions. Workflows based on a combination of
mechanistic (Q)SAR, read-across analysis and expert knowledge may be derived to allow users
to make a transparent decision on the final prediction based on WoE (Petkov et al., 2019b). In
the case of consistent predictions, expert input may not be needed to make a final decision.

(23) https://one.oecd.org/document/ENV/CBC/MONO(2023)32/en/pdf.
(24) https://ec.europa.eu/eurostat/documents/64157/4392716/ESS-QAF-V2.0-final.pdf.
(25) https://qsartoolbox.org/.
(26) https://qsar.food.dtu.dk/.
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Nonetheless, expert input may be useful to expand the set of read-across analogues from
literature sources and/or to provide a rationale for the end-point-specific similarity between
source analogue(s) and the target substance and, not least, to assess how the differences in the
chemical structure between these may affect the end point under assessment. Advice on how to
interpret some freely available models is provided by Mombelli et al. (2016). The WoE approach
for Ames test predictions is also discussed in Mombelli et al. (2022).

Substances for which no test data exist or for which testing is technically not possible represent
a special case in which reliance on non-testing data may be absolute. Many factors will dictate
the acceptability of non-testing methods in reaching a conclusion based on no tests at all. It may
be discussed whether WoE decisions based on multiple genotoxicity and carcinogenicity
estimates can equal or exceed those obtained by one or two in vitro tests, and whether general
rules for adaptation of the standard testing regime as described in Part 2 of Annex V to
Commission Implementing Decision (EU) 2024/365 may be invoked based on such estimates.
This must be considered on a case-by-case basis.

(Q)SAR models are continuously updated to improve predictions, with new versions typically
released on a yearly basis. It is important to understand the impact of model updates on
mutagenicity predictions over time. Such analysis has been done, for instance, by Hasselgren et
al. (2020) on computational methods used for the prediction of the mutagenic properties of drug
impurities.

(Q)SAR models for predicting repeated dose toxicity (including target organ toxicity,
toxicity to reproduction and development). Overall, (Q)SAR approaches are currently not
well validated for repeated dose toxicity and, consequently, no firm recommendations can be
made concerning their routine use in a testing strategy in this area. One of the reasons is that
there are a large number of potential targets/mechanisms associated with repeated dose toxicity
that currently cannot be adequately covered by a battery of (Q)SAR models. Therefore, a
negative result from current (Q)SAR models without other supporting evidence cannot be
interpreted as demonstrating a lack of a toxicological hazard or a need for hazard classification.
Another limitation of (Q)SAR modelling is that dose–response information, including the
N(L)OAEL, is not provided. Similarly, a valid (Q)SAR model may identify a potential toxicological
hazard, but, because of limited confidence in this approach, such a result would not be adequate
to support hazard classification.
It has to be acknowledged, however, that there are emerging models, some of which are
currently under development and have not undergone the OECD validation process yet.

In some cases, (Q)SAR models could be used as part of a WoE approach, when considered
alongside other data, provided the input is correct, the substance falls within the applicability
domain of the model, the prediction is reliable and the outcome is fit for the regulatory purpose.
The validity of models and predictions can be assessed by using the OECD (Q)SAR assessment
framework (OECD, 2023). In addition, (Q)SARs can be used as supporting evidence when
assessing the toxicological properties by read-across within a substance grouping approach,
providing the applicability domain is appropriate. Positive and negative (Q)SAR modelling results
can be of value in a read-across assessment and for classification purposes.

3.1.3. Read-across approach

In the context of the DWD, read-across is provided as an alternative to testing to predict the
toxicological properties of relevant chemical species.

If the read-across approach is used, the conditions stated in Section 1.5 (on the read-across
approach) of Annex XI to Regulation (EC) No 1907/2006 should be met.
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The read-across approach needs to be adequately and appropriately documented to support the
read-across hypothesis and predictions.

The read-across assessment framework (RAAF) documents of March 2017 ( 27) ( 28) describe a
general framework and principles for the scientific assessment of the suitability of a read-across
approach based on different scenarios. These scenarios are selected according to the type of
read-across approach used (analogue or category approach), the basis for the read-across
hypothesis ((bio)transformation of the analogues into common compound(s) or different
compounds having qualitatively similar properties) and whether quantitative variations in the
predicted properties are expected between the different analogues.

To justify the validity of a read-across approach, one important element of the read-across
assessment framework is the need to specify why the commonalities between two or more
analogue structures suggest a similar biological action. A justification also needs to be provided
as to why structural dissimilarities are not expected to result in dissimilar biological actions.

To assess the suitability of selected analogues as source substances, the following general
questions also need to be addressed.

• Is the same end point considered?

• Are there any additional functional groups or additional substituents that may influence the
reactivity and mutagenicity potential (applicability domain considerations)?

• Are the physico-chemical parameters similar (applicability domain considerations)?

• Are there impurities that influence the toxicity profile?

• Is the likely chemical mechanism the same?

A read-across approach can also support a conclusion for a property within a WoE approach.

3.1.4. Weight of evidence (WoE)

When applying the WoE approach, the conditions stated in Section 1.2 of Annex XI to Regulation
(EC) No 1907/2006 (the REACH regulation) should be met. In addition, the provisions of
Section 1.1.1 of Annex I to Regulation (EC) No 1272/2008 (the CLP regulation) should be
considered.

The REACH regulation, Annex XI, Section 1.2, states that there may be sufficient WoE from
several independent sources of information to enable, through a reasoned justification, a
conclusion on the information requirement, while the information from each single source alone
is insufficient to fulfil the information requirement.

According to Section 1.1.1 of Annex I to the CLP regulation, ‘weight of evidence determination
means that all available information bearing on the determination of hazard is considered
together, such as the results of suitable in vitro tests, relevant animal data, information from
the application of the category approach (grouping, read-across), (Q)SAR results, human

(27) https://echa.europa.eu/support/registration/how-to-avoid-unnecessary-testing-on-animals/grouping-of-
substances-and-read-across.
(28) https://echa.europa.eu/documents/10162/13630/raaf_uvcb_report_en.pdf/3f79684d-07a5-e439-16c3-
d2c8da96a316.
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experience such as occupational data and data from accident databases, epidemiological and
clinical studies and well-documented case reports and observations. The quality and consistency
of the data shall be given appropriate weight. Information on substances or mixtures related to
the substance or mixture being classified shall be considered as appropriate, as well as site of
action and mechanism or mode of action study results. Both positive and negative results shall
be assembled together in a single weight of evidence determination.’

Relevant chemical species for which no test data exist, for example relevant chemical species
that cannot be tested, present a special case in which reliance on non-testing data may be
absolute. Many factors will dictate the acceptability of non-testing methods in reaching a
conclusion based on no tests at all. It has yet to be established whether WoE decisions based on
multiple genotoxicity and carcinogenicity estimates can equal or exceed those obtained by one
or two in vitro tests. This must be considered on a case-by-case basis.

WoE is based on expert judgement. This expert judgement has to be made transparent and
understandable by documenting all information used, all steps carried out in the evaluation
process and all conclusions drawn. Furthermore, the recommendations on how to perform WoE
given in EFSA’s guidance on the use of the WoE approach ( 29) and in the ECHA guidance
(Chapter R.4) ( 30) can be followed.

3.2. Toxicokinetics

The aim of this section is to provide a general overview of the main principles of toxicokinetics
and guidance on the use of TK information in the human health risk assessment of relevant
chemical species. For the generation of new TK data, this guidance document should be used
together with DWD Guidance Volume I.

3.2.1. Definitions and uses of toxicokinetic studies

TK studies evaluate the fate of a substance in an organism, that is, its absorption, distribution,
metabolism and excretion (ADME) (see ECETOC, 2006; Regulation (EC) No 440/2008
(Section B.36); and OECD Test Guideline (TG) 417 (2010)).

Toxicokinetics may help in understanding the results from both in vivo and in vitro toxicology
studies by elucidating the ADME of the test substance. Basic TK parameters determined from
these studies will also provide information on the potential for accumulation of the test substance
in tissues and/or organs and the potential for induction or inhibition of biotransformation by
measuring the circulating moieties (parent substance and its metabolites).

Adequate TK data can further be used for in silico predictions and physiologically based
TK/pharmacokinetic modelling or for grouping of substances and read-across approaches.

The TK data should, however, not be used for dose selection or to disregard positive results for
hazard evaluation in the absence of remarkable and unspecific general toxicity.

(29) https://efsa.onlinelibrary.wiley.com/doi/epdf/10.2903/j.efsa.2017.4971.
(30) https://echa.europa.eu/documents/10162/17235/information_requirements_r4_en.pdf/d6395ad2-1596-4708-
ba86-0136686d205e?t=1323782558175.
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3.2.2. Main principles of toxicokinetic studies

TK studies are designed to obtain species- and route-dependent parameter data on the
concentration–time course of a parent substance and its metabolites in, for instance, blood,
urine, bile, faeces, exhaled air and organs.

According to OECD TG 417 on TK studies, data on the following parameters should be assessed:

• half-life and potential for accumulation under repeated or continuous exposure;

• rate and extent of absorption and bioavailability;

• distribution of the relevant chemical species in the body;

• biotransformation;

• rate and extent of pre-systemic (first-pass) and systemic metabolism after oral or inhalation
(if available) exposure;

• rate and extent of excretion in the urine or faeces, via exhalation, and in other biological
fluids (e.g. milk, bile, sweat, etc.);

• information on enterohepatic circulation.

Enterohepatic circulation may pose problems for route-to-route extrapolation since systemic
availability after oral administration may be greater than after non-oral administration. This will
result in an area under the curve (AUC) (which reflects both the absorption / systemic availability
of the test substance and the extent of recirculation). As the relative extent of absorption
following different routes of exposure is often calculated from the ratio of AUCs by different
routes, the absorption after oral exposure may be overestimated when enterohepatic
recirculation takes place. For this reason, TK studies are preferably conducted using the oral
route to administer the substance. Integrating TK data into toxicity studies can improve the
quality of risk assessment in many ways, including aiding understanding of the differences in
responses or sensitivity between individual animals, species or life stages.

In the absence of in vivo data, standardised in vitro / ex vivo assays may, to a limited extent,
provide data on some of the TK parameters. These include parameters of metabolic steps, such
as Vmax ( 31), Km ( 32) and intrinsic metabolic clearance, permeation rates (e.g. intestinal,
placental, blood–brain barrier ( 33)) and the distribution coefficient. Physiologically based TK
modelling techniques may be used to simulate the concentration–time profile in blood and at the
target site.

However, these assays are non-stand-alone methods that can be used prior to an in vivo TK
study. These include:

(31) Maximum velocity; reflects how fast the enzyme can catalyse the reaction.
(32) Michaelis constant; describes the substrate concentration at which half the enzyme’s active sites are occupied by
the substrate.
(33) Currently, there is no standardised and regulatory-process-compliant in vitro methods for permeation rate via
placental and blood–brain barrier.
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• protein binding and binding to erythrocytes, if relevant (in vitro / ex vivo studies);

• hepatic metabolic clearance with primary human hepatocytes or liver cell lines (in vitro
studies);

• intestinal permeability (the in vitro Caco-2 permeability assay).

3.2.2.1. Accumulation

Everything in a biological system has a biological half-life, that is, a measure of how long it will
stay in that system until it is lost by mainly excretion, degradation or metabolism. To put it
another way, the amount of a substance eliminated from the blood in a unit of time is the product
of clearance (the volume of blood cleared per unit of time) and concentration (the amount of a
compound per unit of volume). For the first-order reactions, clearance is a constant value that
is a characteristic of a substance. If the absorption rate of a substance into an organism is
greater than the rate at which the substance is eliminated, the organism is said to be
accumulating the substance. When the internal concentration has increased so that the amount
eliminated equals the amount of substance absorbed, there will be a constant concentration, a
steady state. The extent of accumulation reflects the relationship between the body burden and
the steady state. Species differences in clearance will determine the difference in the steady
state and body burden between experimental animals and humans (Kroes et al., 2004).

3.2.3. Prediction of toxicokinetics

3.2.3.1. Absorption

Absorption is a function of the potential for a substance to cross biological membranes. These
membranes consist of lipidic layers and aqueous phases, which require the substance to be
soluble both in lipid and in water in order to cross through the membrane by passive diffusion.
Therefore, apart from molecular weight, the partition coefficient (log P) – or the n-octanol–water
partition coefficient (log Kow) – value and the water solubility of a substance are useful
parameters in understanding the biological mechanisms behind absorption. The log Kow value
provides information on the relative solubility of the substance in water and in the hydrophobic
solvent octanol (used as a surrogate for lipids). A log Kow value of > 0 indicates that the
substance is lipophilic and, therefore, more soluble in octanol than in water. A log Kow value of
< 0 indicates that the substance is hydrophilic and, hence, more soluble in water than in octanol.
In general, log Kow values between – 1 and 4 are favourable for absorption. Nevertheless, a
substance with such a log Kow value can be poorly soluble in lipids and, hence, not readily
absorbed when its water solubility is very low. It is therefore important to consider both the
water solubility of a substance and its log Kow value when assessing the absorption potential of
a substance. Substances that do not absorb by passive diffusion can cross the membrane by
facilitated diffusion, active transport or pinocytosis.

Once absorbed, the substance enters the blood system for transportation to the liver, where the
substance may undergo first-pass metabolism. The parent compound and/or its metabolites may
then continue through the circulatory systemic and further excreted via urine or into the bile
(known as the first-pass effect or pre-systemic elimination). Large molecules or lipophilic
substances, on the other hand, may instead enter the lymphatic system. Substances in the
lymphatic fluid are then emptied into the subclavian veins and, thereby, will bypass the liver and
its first-pass metabolism.

Oral/gastrointestinal (GI) absorption. The GI tract is an organ with a large absorptive

surface area that stretches from the mouth to the rectum. Absorption of a substance can occur
at different sites along the GI tract and with different transport mechanisms. The small intestine
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has a very large surface area, and the transit time through this section is the longest, making
this the predominant site of absorption within the GI tract. Compared with that of the small
intestine, the rate and extent of absorption within the large intestine is low. Table 1 provides an
overview of different types of data that can be considered for estimating GI absorption, including
physico-chemical characteristics of a substance, type of dosing vehicle used in the in vivo study
and the possibility of the hydrolysis of a substance. Regarding the hydrolysis possibility, it should
also be noted that substances could undergo chemical changes in the GI fluids as a result of
metabolism by intestinal cells, GI flora and/or enzymes released into the GI tract, or by
spontaneous hydrolysis. These changes will alter the physico-chemical characteristics of the
substance; hence, predictions based on the properties of the parent substance may no longer
apply (see Appendix 1-1 to Guidance on the Biocidal Products Regulation – Volume III human
health – Assessment & evaluation (Parts B+C) ( 34) for a detailed list of physiological factors,
data on stomach and intestine pH, and data on transit time in the intestine, among other things).
In addition, the presence of food in the GI tract can affect (generally decrease) the GI absorption
considerably. In that sense, absorption can differ considerably between routes of oral
administration, for example via gavage or diet.

For ionic substances (i.e. acids and bases), one consideration that could influence the absorption
is the varying pH of the GI tract. It is generally thought that ionised substances do not readily
diffuse across biological membranes. Therefore, when assessing the absorption potential of an
acid or a base, knowledge of its pKa (the pH at which 50 % of the substance is in ionised form
and 50 % in non-ionised form) is advantageous. Absorption of acids is favoured at pH < pKa,
whereas absorption of bases is favoured at pH > pKa.

Other mechanisms by which substances can be absorbed in the GI tract include the passage of
small water-soluble molecules with a molecular weight up to around 200 Da. These small
molecules can pass through the aqueous pores or be carried through with the bulk passage of
water.

The absorption of highly lipophilic substances (log Kow ≥ 4), on the other hand, may be limited
by the inability of such substances to dissolve into the GI fluids and make contact with the
mucosal surface. However, the bile salts in the GI tract aid in decomposing highly lipophilic
substances into simpler molecules, which are then formed into micelles that enhance the
absorption of such substances (Watkins and Klaassen, 2018). Substances absorbed as micelles
enter the circulation via the lymphatic system, thereby bypassing the liver. Although particles
and large molecules with a molecular weight of around 1000 g/mol (or Da) would normally be
considered too large to cross biological membranes, small amounts of such substances may be
transported into epithelial cells by pinocytosis or persorption.

(34)
https://echa.europa.eu/documents/10162/2324906/biocides_guidance_human_health_ra_iii_part_bc_en.pdf/
30d53d7d-9723-7db4-357a-ca68739f5094.
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Table 1. How different types of data can be used to interpret oral/GI absorption

Type of data What they tell us

Structure It may be possible to identify ionisable groups within the structure of the
molecule. Groups containing oxygen, sulphur or nitrogen atoms are potentially
ionisable (e.g. thiol (SH), sulphonate (SO3H), hydroxyl (OH–), carboxyl (COOH)
or amine (NH2)).

Molecular weight Generally, the smaller the molecule the more easily it may be taken up.
Molecular weights of < 500 g/mol are favourable for absorption;
molecular weights of > 1 000 g/mol do not favour absorption.

Particle size Generally, solids have to dissolve before they can be absorbed. It may be
possible for particles in the nanometre size range to be taken up through
pinocytosis or passive transport (Amornwachirabodee et al., 2018; Manzanares
and Ceña, 2020). The absorption of a substance that was administered as a dry
powder (e.g. as large particles, several hundreds of micrometres in diameter,
mixed in the diet) or in a suspension may be reduced because of the time taken
for the particles to dissolve. This would be particularly relevant for substances
with low solubility in water.

Water solubility Water-soluble substances will readily dissolve in the GI fluids. Absorption of very
hydrophilic substances via passive diffusion may be limited by the rate at which
the substance partitions out of the GI fluids. However, if the molecular weight is
low (< 200 g/mol), the substance may pass through aqueous pores or be carried
through the epithelial barrier by the bulk passage of water.

Log Kow Moderate log Kow values (between – 1 and 4) are favourable for absorption by
passive diffusion. Any lipophilic compound may be taken up by micellular
solubilisation, but this mechanism may be of particular importance for highly
lipophilic compounds (log Kow > 4), particularly those that have low solubility in
water (≤ 1 mg/l) and would otherwise be poorly absorbed.

Dosing vehicle If the substance has been dosed using a vehicle, the water solubility of the
vehicle and the vehicle/water partition coefficient of the substance may affect
the rate of uptake. Compounds delivered in aqueous media are likely to be
absorbed more rapidly than those delivered in oils. Compounds delivered in oils
that can be emulsified and digested, such as corn oil or arachis oil, are likely to
be absorbed to a greater degree than those delivered in non-digestible mineral
oil (liquid petrolatum).

Oral toxicity data If signs of systemic toxicity are present, then absorption has occurred (a).
Coloured urine and/or internal organs can also provide evidence that a coloured
substance has been absorbed. In addition, some clinical signs such as a hunched
posture could be due to discomfort caused by irritation or simply the presence
of a large volume of test substance in the stomach, and reduced feed intake
could be due to an unpalatable test substance. It must therefore be clear that
the effects that are being cited as evidence of systemic absorption are genuinely
due to absorbed test substance and not to local effects at the site of contact.

Hydrolysis test The hydrolysis test (EU testing method C.7; OECD TG 111) provides information
on the half-life of the substance in water at 50 °C and pH values of 4.0, 7.0 and
9.0. The test is conducted using a low concentration, 0.01 M or half the
concentration of a saturated aqueous solution (whichever is lower). Since the
temperature at which this test is conducted is much higher than that in the GI
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Type of data What they tell us

tract, this test will not provide an estimate of the actual hydrolysis half-life of
the substance in the GI tract. However, it may indicate that the parent compound
is present in the GI tract for only a limited period of time. Hence, TK predictions
based on the characteristics of an easily hydrolysable parent compound may be
of limited relevance.
(a) Ensure that systemic effects do not occur secondary to local effects.

Respiratory and dermal absorption. Inhalation and dermal exposure are not the most
relevant routes of exposure when it comes to relevant chemical species identified in the water
intended for human consumption. Therefore, if an inhalation or dermal study is provided to meet
the information requirements under the DWD, the applicant needs to provide a scientifically
robust justification for choosing a route that is different from the oral route of administration.
Nevertheless, as it cannot be excluded that humans will be exposed dermally and/or by
inhalation to drinking water, positive results seen after exposure via these routes cannot be
discarded. Furthermore, if such a study is used for hazard characterisation, extrapolation from
respiratory/dermal to oral absorption has to be done. If this is the case, the applicants are
advised to consult Guidance on the Biocidal Products Regulation – Volume III human health –
Assessment & evaluation (Parts B+C), Tables 2 and 3 ( 35).

3.2.3.2. Distribution

The concentration of a substance in blood or plasma (blood level) is dependent on the dose, on
the rate of absorption, distribution and elimination, and on the affinity to a tissue. Tissue affinity
is usually described using a parameter known as the volume of distribution, which is a
proportionality factor between the amount of substance present in the body and the measured
plasma or blood concentration. The larger the volume of distribution, the lower the blood level
will be for a given amount of substance in the body. A particularly useful term is the volume of
distribution at a steady state (Vdss). At a steady state, all distribution phenomena are completed,
the various compartments of the body are in equilibrium, and the rate of elimination is exactly
compensated by the rate of absorption. In non-steady-state situations, the distribution volume
varies with time except in the simplest case of a single-compartment model (this model assumes
that the entire body acts as a single, uniform compartment and the substance can enter or leave
the body, that is, the model is ‘open’). In theory, a steady state can be physically reached only
in the case of a constant zero-order input rate and at stable first-order distribution and
elimination rates. However, many real situations are reasonably close to a steady state, and
reasoning at a steady state is a useful method in kinetics.

The rate at which highly water-soluble molecules distribute may be limited by the rate at which
they cross cell membranes, and the access of such substances to the central nervous system
(CNS) or testes is likely to be restricted by the blood–brain and blood–testes barriers (Slitt,
2018). It is not clear what barrier properties the placenta may have. However, species
differences in transplacental transfer may occur due to varying placental structure and metabolic
capacity of the placenta and placental transporters.

Although protein binding can limit the amount of a substance available for distribution, it will
generally not be possible to determine from the available data which substances will bind to
proteins and how strongly they will bind. Furthermore, if a substance undergoes extensive first-

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pass metabolism, predictions made on the basis of the physico-chemical characteristics of the
parent substance may not be applicable.

Table 2 provides an overview of data that can be considered for estimating substance
distribution.

Table 2. How different types of data can be used to interpret substance distribution

Type of data What they tell us

Molecular weight In general, the smaller the molecule, the wider the distribution.

Water solubility Small water-soluble molecules and ions will diffuse through aqueous channels
and pores. The rate at which very hydrophilic molecules diffuse across
membranes could limit their distribution.

Log Kow If the molecule is lipophilic (log Kow > 0), it is likely to distribute into cells, and
the intracellular concentration may be higher than extracellular concentration,
particularly in fatty tissues.

Target organs If the parent compound is the toxicologically active species, it may be possible
to draw some conclusions about the distribution of the substance from its target
tissues. If the substance is a dye, colouration of internal organs can give evidence
of distribution.

Signs of toxicity Clear signs of any toxicity. For example, CNS effects indicate that the substance
(and/or its metabolites) has distributed to the CNS. However, not all behavioural
changes indicate that the substance has reached the CNS. The behavioural
change may be due to discomfort caused by some other effect of the substance.

Trace elements If the substance is a cationic trace element, absorption is likely to be very low
(< 1 %). Stable isotopes or radioisotopes should be used and background levels
determined to prevent analytical problems and inaccurate recoveries.

3.2.4. Accumulation potential

It is important to consider the potential of a substance to accumulate or to be retained within

the body because, with gradual build-up with successive exposures, the body burden can be
maintained for long periods of time.

Lipophilic substances have the potential to accumulate within the body if the dosing interval is
shorter than four times the whole-body elimination half-life. Although there is no direct
correlation between the lipophilicity of a substance and its biological half-life, substances with
high log Kow values tend to have longer half-lives unless high clearance counterbalances their
large volume of distribution. On this basis, there is the potential for highly lipophilic substances
(log Kow > 4) to accumulate in individuals who are frequently exposed to the substance. Once
the exposure stops, the concentration within the body will decline at a rate determined by the
half-life of the substance. Other substances that can accumulate within the body include
substances that bind to endogenous proteins, and certain metals and ions that interact with the
matrix of the bone (Slitt, 2018).

Table 3 provides an overview of data that can be considered for estimating the accumulation of
substances that migrate into the drinking water (oral exposure).
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Table 3. How data from different sites can be used to interpret accumulation after oral
exposure

Site Characteristics of substance of concern

Adipose tissue Lipophilic substances will tend to concentrate in adipose tissue and may
accumulate. If the interval between exposures is less than four times the whole-
body half-life of the substance, then there is the potential for the substance to
accumulate. Generally, substances with high log Kow values have long biological
half-lives. On this basis, daily exposure to a substance with a log Kow of ≥ 4 could
result in a build-up of the substance within the body. A substance with a log Kow
of ≤ 2 would be unlikely to accumulate with repeated intermittent exposure
patterns but may accumulate if exposures are continuous or there is limited
excretion. Once exposure to the substance stops, the substance will be gradually
eliminated at a rate dependent on the half-life of the substance. If fat reserves
are mobilised more rapidly than normal, for example if an individual or an animal
is under stress or lactating, there is the potential for higher quantities of the
parent compound to be released into the blood.

Bone Certain metals, such as lead, and small ions, such as fluoride, can interact with
ions in the matrix of bone. This interaction can displace the normal constituents
of bone, leading to retention of the metal or ion.

3.2.4.1. Metabolism

The main reason animal species toxicity and route-specific toxicity vary is the difference in the
metabolism of substances among species and tissues. The liver has the greatest capacity for
metabolism and commonly causes route-specific pre-systemic effects (first pass), particularly
following oral intake. However, route-specific (oral, for substances that migrate into drinking
water) toxicity may result from several phenomena, such as hydrolysis within the GI tract or
metabolism by GI flora or within the GI tract epithelia (mainly in the small intestine) (for a
review, see Rourke and Sinal, 2014).

Predicting the changes that a substance may undergo is difficult based on physico-chemical
information alone. Although it is possible to look at the structure of a molecule and identify
potential metabolites, it is not certain that these reactions will occur in vivo (e.g. the molecule
may not reach the necessary site for a particular reaction to take place). It is even more difficult
to predict the extent to which it will be metabolised along different pathways and what animal
species differences may exist. Consequently, experimental data can help in the assessment of
potential metabolic pathways. Data on TK parameters can be derived from in vivo studies and,
to a limited extent, from standardised in vitro / ex vivo studies (e.g. the hepatic clearance
assay).

3.2.4.2. Excretion

The major routes of substance excretion from the systemic circulation are via urine and/or
faeces, via bile and directly from the GI mucosa (see Slitt, 2018).

The excretion processes in the kidney are passive glomerular filtration through membrane pores
and active tubular secretion via carrier processes. Substances that are excreted in the urine tend
to be water soluble and of low molecular weight (< 300 g/mol or Da in rats, mostly anionic and
cationic compounds) and are generally conjugated metabolites (e.g. glucuronides, sulphates,
glycine conjugates) from phase II biotransformation. The kidneys will filter most of them out of
the blood, although a small amount may enter the urine directly by passive diffusion, and there
is the potential for reabsorption into the systemic circulation across the tubular epithelium.
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Biliary excretion involves active secretion rather than passive diffusion. Substances that are
excreted into the bile tend to have a higher molecular weight. It has been found that, in rats,
substances with a molecular weight of < ca 300 g/mol (or Da) do not tend to be excreted into
the bile (Renwick, 1994). Threshold values for molecular weights of organic anions have been
set to 400 Da in rats and 475 Da in humans, whereas no threshold values can be set for
molecular weights of cations or cation/neutral compounds (Yang et al., 2009). Hepatic function
highly influences the excretion of compounds via bile, as the metabolites formed in the liver may
be excreted directly into the bile without entering the bloodstream. Blood flow is therefore an
additional determining factor.

Substances in the bile pass through the intestines before they are excreted in the faeces. As a
result, the substances may undergo enterohepatic recycling (i.e. circulation of bile from the liver,
where it is produced, to the small intestine, where it aids the digestion of fats and other
substances, and back to the liver), which will prolong their biological half-life. This is a particular
problem, although uncommon, for conjugated molecules that are hydrolysed by GI bacteria to
form smaller, more lipid-soluble molecules that can be reabsorbed from the GI tract. Substances
with strong polarity and high molecular weight are less likely to recirculate. Other substances
excreted in the faeces are those that have diffused out of the systemic circulation into the GI
tract directly, substances that have been removed from the GI mucosa by efflux mechanisms,
and non-absorbed substances that have been ingested or inhaled and subsequently swallowed.
However, depending on the possible metabolic changes, the compound that is finally excreted
may have few or none of the physico-chemical characteristics of the parent compound.

Table 4 provides an overview of the data that can be used for estimating the excretion of
substances that migrate into drinking water (oral exposure).

Table 4. How physico-chemical data can affect excretion after oral exposure

Route Physico-chemical characteristics favourable for excretion

Urine Characteristics favourable for urinary excretion in rats are a low molecular weight of
< 300 g/mol (or Da), high water solubility and ionisation of the molecule at the pH of
urine.

Bile In rats, molecules that are excreted in the bile are amphipathic (containing both polar
and non-polar regions) and hydrophobic/non-polar and have a high molecular weight.
In general, it is unlikely that in rats more than 5–10 % of organic cations with a
molecular weight of < 300 g/mol (or Da) will be excreted in the bile, however, for
some organic cations (e.g. quaternary ammonium ions), this cut-off may be even
lower. Substances excreted in bile may undergo enterohepatic circulation. Substances
with strong polarity and high molecular weight are less likely to recirculate. Little is
known about the determinants of biliary excretion in humans.

Breast milk Substances present in plasma generally may also be found in breast milk. Lipid-soluble
substances may be present at higher concentrations in milk than in blood/plasma.
Although lactation is a minor route of excretion, exposure of neonates via nursing may
have toxicological significance for some substances.

Saliva/sweat Non-ionised and small lipid-soluble molecules may be excreted in saliva or sweat. In
saliva, the molecules may be swallowed.

Hair/nails Metal ions may be incorporated into hair and nails.

Exfoliation Highly lipophilic substances that penetrated the stratum corneum but did not penetrate
the viable epidermis may be sloughed off with skin cells.
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3.2.5. Generating and integrating toxicokinetic information

In vivo studies provide an integrated perspective on the relative importance of different

processes in an intact biological system, which can be used for comparison with the results of
toxicity studies. To ensure a valid set of TK data, an in vivo TK study has to consist of several
experiments that include blood/plasma kinetics, mass balances and excretion experiments, as
well as tissue distribution experiments. Depending on the problem to be solved, certain
experiments (e.g. plasma kinetics) may be sufficient to provide the data needed for further
assessments (e.g. bioavailability).

The high dose level administered in an ADME study should be linked to the dose levels that cause
adverse effects in toxicity studies. Ideally, there should also be a dose without toxic effect, which
should be in the range of expected human exposure. A comparison of kinetics between toxic
dose levels and dose levels that are likely to represent human exposure values may provide
valuable information for the interpretation of adverse effects and is thus essential for
extrapolation.

In an in vivo study, the systemic bioavailability is usually estimated by comparing either dose-
corrected amounts excreted or dose-corrected AUCs of plasma/blood/serum kinetic profiles,
after extravascular (e.g. oral) and intravascular (e.g. intravenous) administration. The systemic
bioavailability is the dose-corrected amount excreted or AUC determined after an extravascular
substance administration, divided by the dose-corrected amount excreted or AUC determined
after an intravascular substance application, which corresponds, by definition, to a bioavailability
of 100 %. This is valid only if the kinetics of the compound is linear (i.e. dose proportional) and
relies on the assumption that the clearance is constant between experiments. If the kinetics is
not linear, the experimental strategy has to be revised on a case-by-case basis, depending on
the type of non-linearity involved (e.g. saturated protein binding, saturated metabolism).

3.2.6. Other methods for generating absorption, distribution, metabolism

and excretion data

In addition to the predictive approaches described earlier and the test methods described in
Section 6.5 of DWD Guidance Volume I, in vitro studies and physiologically based
TK/pharmacokinetic modelling should also be considered for generating ADME data.

3.2.6.1. In vitro studies

Generally, in vitro studies provide data on specific aspects of toxicokinetics, such as metabolism
or dermal absorption. A major advantage of in vitro studies is that it is possible to carry out
parallel tests on different species, including humans, thus facilitating interspecies comparisons
(e.g. metabolite profile, metabolic rate constants). In recent years, methods using the
appropriate physiologically based TK/pharmacokinetic models to integrate a number of in vitro
results into a prediction of ADME in vivo have been developed. Such methods allow both the
prediction of in vivo kinetics in humans and the progressive integration of all available data into
a predictive model of ADME. The uncertainty associated with the prediction depends largely on
the amount of available data.

In particular, generating TK data should aim to provide essential information for building
physiologically based kinetic models, to enable more accurate estimation of internal exposure,
where relevant.
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3.2.6.2. In silico methods: kinetic modelling

An overview of in silico methods and guidance for use in TK assessment is provided in Guidance
on the Biocidal Products Regulation – Volume III human health – Assessment & evaluation
(Parts B+C) ( 36). Additional guidance has been developed by the OECD on the characterisation,
validation and reporting of physiologically based kinetic models for regulatory purposes (Series
on Testing and Assessment, No 331). This guidance document builds on previous documents on
best practice in physiologically based kinetic model development and application, including the
US EPA document Approaches for the application of physiologically based pharmacokinetic
(PBPK) models and supporting data in risk assessment (2006), the International Programme on
Chemical Safety (IPCS) guidance document Characterization and application of physiologically
based pharmacokinetic models in risk assessment (2010) and EFSA’s Scientific opinion on good
modelling practice in the context of mechanistic effect models for risk assessment of plant
protection products (2014). However, such modelling is still very uncommon.

3.2.7. Variability and uncertainty in toxicokinetics

Uncertainty and variability are inherent to TK studies and potentially affect the conclusions of
these studies. It is necessary to minimise uncertainty in order to assess the variability that may
exist between individuals so that there is confidence in the TK results such that they can be
useful for risk analysts and decision-makers.

Variability typically refers to differences in the physiological characteristics among individuals

(interindividual variability) or across time within a given individual (intraindividual variability). It
may stem from genetic differences, activity level, lifestyle, physiological status, age, sex, etc.
Variability is characteristic of animal and human populations. It can be observed and registered
as information about the population, but it cannot be reduced. An important feature of variability
is that it does not tend to decrease when larger samples of a population are examined.

Variability in the population should, then, be considered in TK studies. The application of

probability distributions to the parameters representing the distribution of physiological
characteristics in the population may introduce variability into physiologically based kinetic
models. The propagation of the variability to model predictions may be evaluated using Monte
Carlo simulation methods.

Uncertainty can be defined as the inability to make precise and unbiased statements. It is
essentially the result of a lack of knowledge. Uncertainty in terms of information may decrease
with the size of the sample studied. Further optimised experiments and better understanding of
the process under study can theoretically eliminate or at least reduce uncertainty.

Uncertainty may be related to the experimental nature of the data, the modelling procedure or
the high inherent variability of biological systems.

3.2.8. Metabolism studies as a basis for internal dose considerations

Biotransformation of a substance produces metabolites that may have different toxicological

properties than the substrate from which they are formed. Although metabolism is generally
considered to have a detoxification purpose, there are also many examples where metabolites
have higher intrinsic toxicity than the parent compound itself (bioactivation). Therefore, it is
necessary to know if the test substance is metabolised and to which metabolites. This enables

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the assessment of results from toxicity studies in respect of grouping approaches and defines
the internal dose (see Figure 3).

Figure 3. Use of increasing knowledge on substance metabolism

Is the test substance metabolised ?

no yes

Parent compound
is relevant internal Metabolites known?
dose metric

no yes

Tox. information on
Identification via
metabolites
in vitro / in vivo meta-
available ? (with respect
bolism studies
to species specificities)

no yes

Perform studies
of metabolites Define relevant
(include grouping internal dose metric
approach)

Apply in RA

If the test substance is not metabolised, the parent compound is the relevant marker for
measuring and defining the internal dose. If the test substance is metabolised, the knowledge
of which metabolites are formed is essential for any further steps in an assessment. When this
information is not available, it can be investigated by appropriate in vitro and/or in vivo
metabolism studies. In special cases, metabolites may show a high degree of isomeric specificity,
and this should be kept in mind when designing and interpreting mixtures of isomers, including
racemates. If the metabolites are known and the toxicity studies are available for these
metabolites, the risk assessment may be carried out based on these data, and an assessment
based on the definition of the internal dose can be made. If the toxicity profile of the metabolites
is unknown, studies that address the metabolites’ toxicity may be performed, with special
consideration given to the potential of group approaches, particularly if a substance is the
metabolite of different compounds, for example carboxylic acid as a metabolite of different
esters.
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TK information can be very helpful in bridging various gaps encountered across the whole risk
assessment, from toxicity study design and biomonitoring set-up ( 37) to the derivation of the
threshold levels and various extrapolations, as are usually needed (cross-dose, cross-species
(including human), cross-exposure regimen, cross-route and cross-substance extrapolations).
The internal dose is the central output parameter of TK studies and therefore the external
exposure–internal dose concept is broadly applicable in the various extrapolations mentioned.
If, for that purpose, route-to-route extrapolation is necessary, and if assessment of combined
exposure (via different routes) is needed, for systemic effects, internal exposure may have to
be estimated.

Exposure should normally be understood as an external exposure, which, in the context of the
DWD, can be defined as the amount of substance ingested.

Determination of the level of systemic exposure is considered synonymous with determination

of the substance bioavailability to the general circulation. It should be emphasised that it is
usually not possible to show that after oral ingestion the amount of a substance bioavailable is
zero; however, in some cases, it may be predicted to be very low. The degree of certainty of the
prediction will depend on the case. Important factors include the accuracy and reliability of the
in vivo, in vitro or in silico model used, the performance of the methods used to assay the
substance or its metabolites, and the estimated variability in the target population. It should be
noted that when a very low bioavailability is predicted, the degree of certainty of this prediction
should be very high.

The compound’s tissue distribution characteristics can be an important determinant of its

potential to cause toxicity in specific tissues. In addition, tissue distribution may be an important
determinant of the ability of a compound to accumulate upon repeated exposure. However, the
accumulation is substantially modified by the rate at which the compound is cleared. Correlation
of tissue distribution with target tissues in toxicity studies should be accomplished while
substantial amounts of the substance remain present in the body (e.g. around one or more times
the peak blood concentration following oral absorption). Such data should quantify the parent
compound and the metabolites to the extent feasible. If the metabolites are unknown or difficult
to quantify, subtracting the parent compound from total radioactivity will provide an estimate of
the behaviour of the total metabolites formed.

3.2.9. Extrapolation

For ethical reasons, if data allowing model parameters to be estimated are of poor quality and
sparse, and do not often concern human populations, recourse to extrapolation is needed. In
most cases TK data are gathered for a limited number of different concentrations (usually < 5
different concentrations) and exposure times. However, risk evaluation should assess the
different doses (concentrations and exposure times). Interdose/interexposure time extrapolation
is a common way to satisfy this demand (Clewell and Andersen, 1996).

In the rare cases where data on human volunteers are available, they concern only a very limited
number of subjects. Extrapolation to other populations can be done (interindividual
extrapolation). The problem of sensitive populations also arises, and hence a TK study should
assess different genders, ages, ethnic groups, etc. As it is nearly impossible to control the
internal dose in humans in practice, an alternative animal study is often proposed. Since risk

(37) Biological monitoring information should be seen as equivalent to (i.e. as having neither greater nor lesser
importance than) other forms of exposure data. It should also be remembered that biological monitoring results
reflect an individual’s total exposure to a substance from any relevant route, that is, from consumer products
and/or the environment, and not just occupational exposure. Data from controlled human exposure studies are
even less likely to be available. This is due to the practical and ethical considerations involved in deliberate exposure
of individuals.
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assessment aims to protect human populations, interspecies extrapolation (Davidson et al.,

1986; Watanabe and Bois, 1996) should be done. For practical reasons, the administration route
in an experimental study can be different from the most likely exposure route. In this case, risk
assessment suggests deciding on a route other than the one experimentally studied. Interroute
extrapolation should be performed.

Default values have been derived to match the extrapolation idea in a general way. The
incorporation of quantitative data on interspecies differences or human variability in
toxicokinetics and toxicodynamics into the dose–response / concentration–response assessment
may be implemented through the development of substance-specific assessment factors (AF).
The use of substance-specific AFs may improve risk assessment of individual substances.
Currently, data relevant for consideration are often restricted to the component of uncertainty
related to interspecies differences in toxicokinetics. At the present time, there are commonly
fewer data to address interspecies differences in toxicodynamics and interindividual variability
in toxicokinetics and toxicodynamics. It is anticipated that the availability of such information
will increase with a better common understanding of its appropriate nature (IPCS, 2001). The
types of TK information that could be used include the rate and extent of absorption, the extent
of systemic availability, the rate and extent of pre-systemic (first-pass) and systemic
metabolism, the extent of enterohepatic recirculation, information on the reactive metabolites
formation and possible species differences, and knowledge of the half-life and potential for
accumulation under repeated exposure.

The need for these extrapolations can lead to more frequent use of physiological TK models than
empirical models (Davidson et al., 1986; Watanabe and Bois, 1996). Physiologically based
kinetic models facilitate the required extrapolations (interspecies, intersubjects, etc.). For
example, by changing anatomical parameters such as organ volumes or blood flows, a
physiologically based kinetic model can be transposed from rat to human.

3.2.9.1. Interspecies extrapolation

The use of animal data for toxicological risk assessment raises the question of how to extrapolate
experimentally observed kinetics to human subjects or populations. The ability to compare data
from animals and humans enables the definition of substance-specific interspecies extrapolation
factors to replace the default values. One option is extrapolation based on different body sizes,
which calculates the allometric factors.

Allometric scaling is a commonly employed extrapolation approach. It is based on the principle

that biological diversity is largely explained by body size (Schneider et al., 2004). Allometric
scaling captures the correlations of physiological parameters or toxicokinetics with body size.
More precisely, allometric equations relate the quantity of interest (e.g. a tissue dose) to a power
function of body mass fitted across species, as follows:

Y = a BWb

where Y is the quantity of interest, a is the species-independent scaling coefficient ( 38),BW is

body weight and b is the allometric exponent.

The value of b depends on whether the quantity of interest scales approximately with body mass
(b = 1), metabolic rate ( 39) (b = 0.75) or body surface area (b = 0.67 ( 40)) (Fiserova-Bergerova

(38) Fits single data points together to form an appropriate curve.

(39) In this context, it is not metabolism of compounds. The factor adapts different levels of oxygen consumption.
(40) This scaling factor is generally justified on the basis of the studies by Freireich et al. (1966), who examined the
interspecies differences in the toxicity of a variety of antineoplastic drugs.
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and Hugues, 1983; Davidson et al., 1986; West et al., 1997). As it is easy to apply the allometric
scaling, it is probably the most convenient approach to interspecies extrapolation. However, it
is very approximate and may not hold for the substance of interest. It can therefore be
considered a default approach to be used only in the absence of specific data on the relevant
chemical species of interest.

For a substance that demonstrates significant interspecies variation in animal toxicity

experiments, the most susceptible animal species are generally used as the reference point for
extrapolation. The need for additional uncertainty factors should be evaluated case by case.

To better estimate tissue exposure across species, physiologically based kinetic models may be
used for the toxicant of interest (Watanabe and Bois, 1996). These models account for transport
mechanisms and metabolism within the body. The same equation set then models the processes
for all species considered. Differences between species are assumed to be due to different
(physiological, chemical and metabolic) parameter values. Extrapolation of physiologically based
kinetic models then relies on replacing the model parameter values of one species with the
parameter values of the species of interest. For physiological parameters, numerous studies
(Arms and Travis, 1988; Brown et al., 1997; ICRP, 2002) give standard parameter values for
many species. Chemical (log P) and metabolic parameter values are usually less easily found.
When the parameter values of the physiologically based kinetic model are not known for the
species of interest, the option of in vitro data, (Q)SAR predictions or allometric scaling of those
parameters is still possible. To consider population variability in the extrapolation process,
probability distributions of parameters may be used rather than single parameter values.
Physiologically based kinetic models can be particularly useful where data are being extrapolated
to population subgroups for which only little information is available, for example pregnant
women or infants (Luecke et al., 1994; Young et al., 2001).

3.2.9.2. Interroute extrapolation

As explained in Chapter 5, the oral route is the most relevant one for substances migrated in
drinking water. If studies with a route of exposure other than the oral route are used, route-to-
route extrapolation should be applied.

Route-to-route extrapolation is defined as the prediction of the total amount of a substance

administered by one route that would produce the same systemic toxic response as that obtained
for a given amount of a substance administered by another route.

In general, route-to-route extrapolation is considered to be a poor substitute for toxicity data

obtained using the appropriate route of exposure. Uncertainties in extrapolation increase when
performing risk assessment with toxicity data obtained by an administration route that does not
correspond to the human route of exposure.

Because of the high level of uncertainty in extrapolation, it is recommended that inhalation and,
in particular, dermal studies not be used for hazard assessment and risk acceptance for the
purpose of the DWD. However, if a study performed via a different route of administration (e.g.
inhalation) is used, route-to-route extrapolation has to be considered, taking into account all the
uncertainties mentioned below.

When route-to-route extrapolation is used, the following aspects should be carefully considered.

• Nature of the effect. Route-to-route extrapolation is applicable only for the evaluation of
systemic effects. For the evaluation of local effects after repeated exposure, only results from
toxicity studies performed with the route under consideration can be used.

• TK data (ADME). The major factors responsible for differences in toxicity due to route of
exposure include:
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o differences in bioavailability or absorption;

o differences in metabolism (first-pass effects);

o differences in internal exposure pattern (i.e. internal dose).

In the absence of relevant kinetic data, route-to-route extrapolation is possible only if the
following assumptions are reasonably valid:

• absorption can be quantified;

• toxicity is a systemic effect, not a local one (the compound is relatively soluble in body fluids,
therefore systemically bioavailable), and internal dose can be estimated;

• first-pass effects are minimal.

For further guidance on how to perform route-to-route extrapolation, please follow the
recommendations provided in ECHA’s Guidance on Information Requirements and Chemical
Safety Assessment – Chapter R.8: Characterisation of dose [concentration]–response for human
health ( 41).

3.3. Repeated dose toxicity

The sections on repeated dose toxicity, neurotoxicity, endocrine disruption and immunotoxicity
in DWD Guidance Volume I (Sections 6.6 and 6.10) should be considered together with the
elements described in this section for the assessment of repeated dose toxicity. Information
from experimental and non-test approaches with regard to other end points (e.g. toxicokinetics
and genotoxicity) should be assessed using a WoE approach in the assessment of toxicological
findings following repeated dose administration. The ultimate goal is to identify the potential
mode of action and underlying key events.

3.3.1. Definition of repeated dose toxicity

Repeated dose toxicity comprises the adverse general (i.e. excluding reproductive, genotoxic or
carcinogenic effects) toxicological effects occurring as a result of repeated daily dosing with or
exposure to a substance for a part of the expected lifespan (subacute or subchronic exposure)
or for a major part of the lifespan, in the case of chronic exposure.

The term general toxicological effects (hereafter referred to as general toxicity) includes effects
on body weight and/or body weight gain, absolute and/or relative organ and tissue weights,
alterations in clinical chemistry, urinalysis and/or haematological parameters, functional
disturbances in the nervous system as well as in organs and tissues in general, and pathological
alterations in organs and tissues as examined macroscopically and microscopically. Repeated
dose toxicity studies may also examine parameters that have the potential to identify specific
manifestations of toxicity, such as neurotoxicity, immunotoxicity, endocrine-mediated effects,
reproductive toxicity and carcinogenicity.

An adverse effect is a change in the morphology, physiology, growth, development,

reproduction or lifespan of an organism, system or (sub)population that results in an impairment
of functional capacity, an impairment of the capacity to compensate for additional stress or an
increase in susceptibility to other influences.

(41) https://echa.europa.eu/documents/10162/17224/information_requirements_r8_en.pdf/e153243a-03f0-44c5-
8808-88af66223258?t=1353935239897.
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A chemical substance may induce systemic and/or local effects, defined as follows:

• a local effect is an effect that is observed at the site of first contact, caused irrespective of
whether a substance is systemically available;

• a systemic effect is an effect that is normally observed distant from the site of first contact,
that is, after having passed through a physiological barrier (mucous membrane of the GI
tract or the respiratory tract, or the skin) and becomes systemically available.

It should be noted, however, that toxic effects on surface epithelia may reflect indirect effects
as a consequence of systemic toxicity or secondary effects as a consequence to systemic
distribution of the substance or its active metabolite(s).

Repeated dose toxicity tests provide information on possible adverse effects likely to arise from
repeated exposure of target organs and information on dose–response relationships.

The determination of the dose–response relationship should lead to the identification of the
NOAEL/LOAEL/BMD. As part of the risk assessment process for relevant chemical species, data
on the adverse effects a substance may cause and the dose levels at which the effects occur are
evaluated in the light of the likely extent of human exposure to the relevant chemical species
via water intended for human consumption, so that the potential risk(s) to health may be
ascertained.

The objectives of assessing repeated dose toxicity are to evaluate:

• whether exposure of humans to a substance has been associated with adverse toxicological
effects occurring as a result of repeated daily exposure for a part of the expected lifespan
or for a major part of the lifespan;

• whether administration of a substance to experimental animals causes adverse

toxicological effects as a result of repeated daily exposure for a part of the expected
lifespan or for a major part of the lifespan – effects that may be predictive of possible
adverse human health effects;

• the target organs, potential cumulative effects and the reversibility of the adverse
toxicological effects;

• the dose–response relationship and threshold for any of the adverse toxicological effects
observed in the repeated dose toxicity studies;

• the basis for risk characterisation and classification and labelling of substances for systemic
target organ toxicity.

3.3.2. Data to be used in the hazard assessment

3.3.2.1. Non-human data for repeated dose toxicity

3.3.2.1.1. Non-testing data for repeated dose toxicity

See Section 3.1.2.2.5 (on (Q)SARs) and Section 3.1.3 (on the read-cross approach).

3.3.2.1.2. Testing data for repeated dose toxicity

In vitro data. While several assays measuring repeated in vitro exposure of, for example, liver,
kidney and brain cells are under development, currently the in vitro data cannot be used on their
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own in regulatory decisions such as risk assessment. However, such data may be helpful in the
assessment of repeated dose toxicity, for instance to detect local target organ effects and/or to
clarify the mechanisms of action. Since, at present, there are no validated in vitro methods that
have received regulatory acceptance, the quality of each of these studies and the adequacy of
the data provided should be carefully evaluated.

Generic guidance is given in ECHA’s Guidance on Information Requirements and Chemical Safety
Assessment – Chapter R.4: Evaluation of available information ( 42) and Guidance on Information
Requirements and Chemical Safety Assessment – Chapter R.5: Adaptation of information
requirements ( 43), for judging the applicability and validity of the outcome of various study
methods and assessing the quality of the conduct of a study and the reproducibility of data and
aspects such as vehicle, number of replicates, exposure/incubation time, GLP compliance and
comparable quality description.

Animal data. The most appropriate data on repeated dose toxicity for use in hazard
characterisation and risk assessment of relevant chemical species identified in the migration
water are primarily obtained from oral studies in experimental animals conforming to
internationally agreed test guidelines. In some circumstances, repeated dose toxicity studies not
conforming to conventional test guidelines may also provide relevant information for this end
point.

The information that can be obtained from the available EU/OECD test guideline studies for
repeated dose toxicity is briefly summarised below. Table 5 summarises the parameters
examined in these OECD test guideline studies in more detail to facilitate overview of the
similarities and differences between the various studies.

Repeated dose 28-day toxicity oral study

The relevant guideline for this end point is EU B.7 / OECD TG 407. The 28-day studies provide
information on the toxicological effects arising from exposure to the substance during a relatively
limited period of the animal’s lifespan.

Repeated dose 90-day oral toxicity studies

Separate guidelines are available for studies using oral administration (OECD TG 408 / EU B.26
and OECD TG 409 / EU B.27 in rodent and non-rodent species, respectively). The principle of
these study protocols is identical, although the revised OECD TG 408 (2018) protocol includes
additional parameters enabling the identification of a neurotoxic potential, immunological effects
or reproductive organ toxicity.

The 90-day study provides information on the general toxicological effects on target organs
arising from subchronic exposure (a prolonged period of the animal’s lifespan) covering post-
weaning maturation and growth well into adulthood.

Chronic toxicity studies

The chronic toxicity studies (OECD TG 452 / EU B.30) provide information on the toxicological
effects arising from repeated exposure over a prolonged period of time covering a major part of
the animal’s lifespan. The duration of the chronic toxicity studies should be at least 12 months.

(42) https://echa.europa.eu/documents/10162/17235/information_requirements_r4_en.pdf/d6395ad2-1596-4708-
ba86-0136686d205e?t=1323782558175.
(43) https://echa.europa.eu/documents/10162/17235/information_requirements_r5_en.pdf/51ffb7a7-baef-43ef-bac7-
501e95b5a1d5?t=1323783044916.
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The combined chronic toxicity / carcinogenicity studies (OECD TG 453 / EU B.33) include an
additional high-dose satellite group for evaluation of pathology other than neoplasia. The satellite
group should be exposed for at least 12 months and the animals in the carcinogenicity part of
the study should be retained in the study for the majority of the normal lifespan of the animals.

Ideally, the chronic studies should allow for the detection of general toxicity effects
(physiological, biochemical and haematological effects, etc.) but could also provide information
on neurotoxic, immunotoxic, reproductive and carcinogenic effects of the relevant chemical
species. However, in 12-month studies, it is possible that non-specific life-shortening effects,
which require a long latent period or are cumulative, may not be detected in this study type. In
addition, the combined study will allow for the detection of neoplastic effects and determination
of carcinogenic potential and life-shortening effects.

Repeated dose toxicity study combined with reproduction/developmental toxicity screening test

The combined repeated dose toxicity / reproductive screening study (OECD TG 422 / EU B.64)
provides information on the toxicological effects arising from repeated exposure (generally oral
exposure) over a period of about 6 weeks for males and approximately 54 days for females (a
relatively limited period of the animal’s lifespan) and information on reproductive toxicity. For
the repeated dose toxicity part, OECD TG 422 / EU B.64 is in concordance with OECD TG 407 /
EU B.7 except for use of pregnant females and the longer exposure duration in OECD TG 422 /
EU B.64 than in OECD TG 407 / EU B.7.

Neurotoxicity studies

The neurotoxicity study in rodents (OECD TG 424 / EU B.43) is designed to further characterise
potential neurotoxicity observed in repeated dose systemic toxicity studies. The neurotoxicity
study in rodents will provide detailed information on major neurobehavioural and
neuropathological effects in adult rodents.

Delayed neurotoxicity studies of organophosphorus substances

The delayed neurotoxicity study (OECD TG 419 / EU B.38) is specifically designed to be used in
the assessment and evaluation of the neurotoxic effects of organophosphorus substances. This
study provides information on the delayed neurotoxicity arising from repeated exposure over a
relatively limited period of the animal’s lifespan.

Other studies providing information on repeated dose toxicity

Although they are not aimed at investigating repeated dose toxicity per se, other available
OECD/EU test guideline studies involving repeated exposure of experimental animals may
provide useful information on repeated dose toxicity. These studies are summarised in Table 6.

It should be noted that the repeated dose toxicity studies, if carefully evaluated, may provide
information on potential reproductive toxicity and on carcinogenicity (e.g. pre-neoplastic
lesions).

The one-generation, two-generation and extended one-generation reproductive toxicity studies

(OECD TG 415 / EU B.34; TG 416 / EU B.35; TG 443 / EU B.56) may provide information on the
general toxicological effects arising from repeated exposure over a prolonged period of time
(about 90 days for parental animals), as clinical signs of toxicity, body weight, selected organ
weights, and gross and microscopic changes of selected organs are recorded.

The prenatal developmental toxicity study (OECD TG 414 / EU B.31), reproduction/developmental

toxicity screening study (OECD TG 421 / EU B.63) and developmental neurotoxicity study
(OECD TG 426) may give some indications of general toxicological effects arising from repeated
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exposure over a relatively limited period of the animal’s lifespan, as clinical signs of toxicity and
body weight are recorded.

The carcinogenicity study (OECD TG 451 / EU B.32) will, in addition to providing information on
neoplastic lesions, provide information on the general toxicological effects arising from repeated
exposure over a major portion of the animal’s lifespan, as clinical signs of toxicity, body weight,
and gross and microscopic changes of organs and tissues are recorded.

The basic concept of repeated dose toxicity studies to generate data on target organ toxicity
following subacute to chronic exposure is to treat experimental animals for 4 weeks, 13 weeks
or longer. In addition, other studies performed in experimental animals may provide useful
information on repeated dose toxicity. While the majority of non-animal methods are currently
in the research and development phase and not yet poised as substitutes for subchronic/chronic
animal studies, there is the potential to enhance data collection for risk assessment. This
presents opportunities for increased efficiency, reduced animal usage and more effective
resource utilisation.

Consideration of in vitro data as well as TK data can be useful in the evaluation of the repeated
dose toxicity information, as these data can assist in the correct derivation of internal exposure
values, the assessment of human relevance and the correct application of AFs in deriving
threshold levels and in the design and development of new and more biologically relevant tests.

The following general guidance is provided for the evaluation of repeated dose toxicity data and
the development of the WoE; in this respect, all other information, including non-test methods,
must be taken into account in the WoE building.

• Studies on the most sensitive animal species should be selected as the significant ones,
unless TK and toxicodynamic data show that this species is less relevant for human risk
assessment.

• Studies using an appropriate route, duration and frequency of exposure in relation to the
expected route(s), frequency and duration of human exposure have greater weight. For
relevant chemical species identified in the migration water, those are oral subchronic and/or
chronic repeated dose toxicity studies.

• Studies enabling the identification of the relevant point of departure (NOAEL/LOAEL/BMD)

and robust hazard identification have a greater weight.

• Studies of a longer duration should be given greater weight than a repeated dose toxicity
study of a shorter duration in the determination of the most relevant NOAEL (e.g. chronic
study is preferred over subchronic).

If sufficient evidence is available to identify the critical effect(s) (with regard to the dose–
response relationship(s) and the relevance for humans) and the target organ(s) and/or tissue(s),
greater weight should be given to specific studies investigating this effect in the identification of
the NOAEL. The critical effect can be a local as well as a systemic effect.

While data available from repeated dose toxicity studies not performed in accordance with
conventional guidelines and/or GLP may still provide information of relevance for risk assessment
and classification and labelling, such data require extra careful evaluation.

Data from non-guideline studies must be considered to be equivalent to data generated by

corresponding test methods if the following conditions are met:
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• they are adequate for the purpose of risk assessment;

• there is adequate coverage of the key parameters (e.g. observations, body weight,
food/water consumption, haematology and clinical biochemistry, gross pathology and
histopathology) that the corresponding test guideline methods anticipate will be investigated;

• exposure duration is comparable to or longer than the corresponding test guideline method
if exposure duration is a relevant parameter;

• adequate and reliable documentation of the study is provided.

In all other situations, non-guideline studies may contribute to the overall weight of the evidence
but cannot stand alone as a hazard and risk assessment of a relevant chemical species and,
thus, cannot serve as the sole basis for an assessment of repeated dose toxicity or for exemption
from the standard information requirements for repeated dose toxicity. That is, they cannot be
used to identify a relevant chemical species as being adequately controlled in relation to repeated
dose toxicity.

The existing information is considered sufficient when, based on a WoE analysis, the critical
effect(s) and target organ(s) and tissue(s) can be identified, the dose–response relationship(s)
and NOAEL(s) and/or LOAEL(s) and/or BMD(s) for the critical effect(s) can be established, and
the relevance for human beings can be assessed.

It should be noted that potential effects in certain target organs (e.g. thyroid) following repeated
exposure may not be observed within the span of shorter studies, for example the 28-day study.
Attention is also drawn to the fact that the protocols for the 28-day and 90-day oral studies
include additional parameters compared with those for the 28-day and 90-day dermal and
inhalation protocols.

Where it is considered that the existing data are as a whole inadequate to provide a clear
assessment of this end point, the need for further testing should be considered in view of all
available relevant information on the relevant chemical species, including use pattern, the
potential for human exposure, physico-chemical properties and structural alerts.

Specific investigations such as studies of neurotoxicity, endocrine disruption or immunotoxicity

are also elements in the testing strategy that should be taken into account.

Regarding neurotoxicity and immunotoxicity, standard 28-day and 90-day oral toxicity studies
include end points capable of detecting such effects. Indicators of neurotoxicity include clinical
observations, a functional observational battery, motor activity assessment and
histopathological examination of the spinal cord and sciatic nerve. Indicators of immunotoxicity
include changes in haematological parameters and serum globulin levels, alterations in immune
system organ weights such as the spleen and thymus, and histopathological changes in immune
organs such as the spleen, thymus, lymph nodes and bone marrow. Where data from standard
28-day and 90-day oral studies identify evidence of direct test substance-related adverse effects
for neurotoxicity or immunotoxicity, it is advised that those effects be considered in the WoE
approach to decide on the need for further testing. It should be noted that very limited end
points indicative of neurotoxicity and immunotoxicity are examined in the standard 28-day and
90-day dermal or inhalation repeated dose toxicity studies.

Regarding endocrine disruptors (EDs), ECHA/EFSA guidance on the identification of EDs ( 44) is
available and describes how to gather, evaluate and consider all relevant information for the

(44) https://doi.org/10.2903/j.efsa.2018.5311.
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assessment, conduct a mode of action analysis and apply a WoE approach, in order to establish
whether the ED criteria are fulfilled.

Furthermore, guidance to facilitate the interpretation of hazard data derived from screens and
tests in the OECD conceptual framework ( 45) was published in 2012 and updated in 2018 to
reflect new and updated OECD test guidelines.

In general, results from toxicological studies requiring repeated administration of a substance,

such as reproduction and developmental toxicity studies and carcinogenicity studies, can
contribute to the assessment of repeated dose toxicity. However, such toxicological studies
rarely provide the information obtained from a standard repeated dose toxicity study and,
therefore, cannot stand alone as the sole basis for the assessment of repeated dose toxicity or
for exemption from the standard information requirements for repeated dose toxicity.

Studies such as acute toxicity and in vivo genotoxicity studies contribute limited information to
the overall assessment of repeated dose toxicity. However, such studies may be useful in
deciding on the dose levels for use in repeated dose toxicity studies.

Guidance on the dose selection for repeated dose toxicity testing is provided in detail in the EU
and OECD test guidelines. Unless limited by the physico-chemical nature or biological effects of
the test substance, the highest dose level possible should be chosen, with the aim of inducing
toxicity but not death or severe suffering.

TK studies may be helpful in the evaluation and interpretation of repeated dose toxicity data, for
example in relation to the accumulation of a substance or its metabolites in certain tissues or
organs as well as the mechanistic aspects of repeated dose toxicity and species differences.

Table 5. Overview of in vivo oral repeated dose toxicity test guideline studies

Test Design End points

OECD TG 407 (EU B.7): Exposure for 28 days Clinical observations
repeated dose 28-day oral At least three dose levels Functional observations (fourth exposure
toxicity study in rodents plus control week – sensory reactivity to stimuli of
At least five males and different types, grip strength, motor
females per group activity)
Preferred rodent species: Body weight and food/water consumption
rat Haematology (haematocrit, haemoglobin,
erythrocyte count, total and differential
leucocyte count, platelet count, blood
clotting time/potential)
Clinical biochemistry
Urinalysis (optional)
Gross necropsy (full, detailed, all animals)
Organ weights (all animals – liver, kidneys,
adrenals, testes, epididymides, thymus,
spleen, brain, heart)
Histopathology (full, at least control and
high-dose groups – all gross lesions, brain,
spinal cord, stomach, small and large
intestines, liver, kidneys, adrenals, spleen,
heart, thymus, thyroid, trachea and lungs,
gonads, accessory sex organs, urinary
bladder, lymph nodes, peripheral nerve, a

(45) http://www.oecd.org/env/ehs/testing/oecdworkrelatedtoendocrinedisrupters.htm#GD_Standardized_TG.
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Test Design End points

section of bone marrow)
Endocrine end points: T3, T4 and TSH
(optional)
OECD TG 408 (EU B.26): Exposure for 90 days Clinical observations
repeated dose 90-day oral At least three dose levels Ophthalmological examination
toxicity study in rodents plus control Functional observations (towards end of
At least 10 males and exposure period – sensory reactivity to
females per group stimuli of different types, grip strength,
Preferred rodent species: motor activity)
rat Body weight and food/water consumption
Haematology (haematocrit, haemoglobin,
erythrocyte count, total and differential
leucocyte count, platelet count, blood
clotting time/potential)
Clinical biochemistry
Urinalysis
Gross necropsy (full, detailed, all animals)
Organ weights (all animals – liver, kidneys,
adrenals, testes, epididymides, uterus,
ovaries, thymus, spleen, brain, heart)
Histopathology (full, at least control and
high-dose groups – all gross lesions, brain,
spinal cord, pituitary, thyroid, parathyroid,
thymus, oesophagus, salivary glands,
stomach, small and large intestines, liver,
pancreas, kidneys, adrenals, spleen, heart,
trachea and lungs, aorta, gonads, uterus,
accessory sex organs, female mammary
gland, prostate, urinary bladder, gall
bladder (mouse), lymph nodes, peripheral
nerve, a section of bone marrow, and
skin/eyes on indication)
Endocrine end points: T4 and thyroid
weight (mandatory); T3, TSH and other
hormone measurements (optional)
Observations of the spermatogenesis and
oestrous cycle
OECD TG 409 (EU B.27): Exposure for 90 days Clinical observations
repeated dose 90-day oral At least three dose levels Ophthalmological examination
toxicity study in non- plus control Body weight and food/water consumption
rodents
At least four males and Haematology (as in OECD TG 408 /
females per group EU B.26)
Preferred species: dog Clinical biochemistry
Urinalysis
Gross necropsy (full, detailed, all animals)
Organ weights (as in OECD TG 408 /
EU B.26 – additional: gall bladder, thyroid,
parathyroid)
Histopathology (as in OECD TG 408 /
EU B.26 – additional: gall bladder, eyes)
OECD TG 452 (EU B.30): Exposure for at least Clinical observations, including neurological
chronic toxicity studies 12 months changes
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Test Design End points

At least three dose levels Ophthalmological examination
plus control Body weight and food/water consumption
Rodents: at least 20 Haematology (haematocrit, haemoglobin,
males and females per erythrocyte count, total leucocyte count,
group platelet count, clotting potential)
Non-rodents: at least Clinical biochemistry
four males and females
Urinalysis
per group
Gross necropsy (full, detailed, all animals)
Preferred rodent species:
rat Organ weights (all animals – brain, liver,
kidneys, adrenals, gonads,
Preferred non-rodent
thyroid/parathyroid (non-rodents only))
species: dog
Histopathology (full, at least control and
high-dose groups – all grossly visible
tumours and other lesions, as well as
essentially the same organs and tissues as
in the 90-day studies (OECD TG 408/409 /
EU B.29/B.27))
OECD TG 453 (EU B.33): Exposure for at least Essentially as in OECD TG 452 / EU B.30
combined chronic toxicity / 12 months (satellite
carcinogenicity studies groups) or majority of
normal lifespan
(carcinogenicity part)
At least three dose levels
plus control
At least 50 males and
females per group
Satellite group: at least
20 males and females
per group
Preferred species: rat
OECD TG 422 (EU B.64): Exposure for a minimum Clinical observations as in OECD TG 407 /
combined repeated dose of 4 weeks (males) or EU B.7
toxicity study with the from 2 weeks prior to Functional observations as in OECD
reproduction/developmental mating until at least TG 407 / EU B.7
toxicity screening test postnatal day 4
Body weight and food/water consumption
(females – at least
6 weeks of exposure) Haematology as in OECD TG 407 / EU B.7
At least three dose levels Clinical biochemistry
plus control Urinalysis (optional)
At least 10 males and Gross necropsy (full, detailed, all adult
females per group animals)
Organ weights (testes and epididymides –
all males; liver, kidneys, adrenals, thymus,
spleen, brain, heart – in five animals of
each sex per group, that is, as in OECD
TG 407 / EU B.7)
Histopathology (ovaries, testes,
epididymides, accessory sex organs, all
gross lesions – all animals in at least
control and high-dose groups; brain, spinal
cord, stomach, small and large intestines,
liver, kidneys, adrenals, spleen, heart,
thymus, thyroid, trachea and lungs, urinary
bladder, lymph nodes, peripheral nerve, a
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Test Design End points

section of bone marrow – in five animals of
each sex in at least control and high-dose
groups, that is, as in OECD TG 407 /
EU B.7)
Endocrine end points: anogenital
difference; nipple retention, T4
OECD TG 424 (EU B.43): Exposure for at least Detailed clinical observations
neurotoxicity study in 28 days Functional observations (sensory reactivity
rodents Dose levels: not to stimuli of different types, grip strength,
specified motor activity, more specialised tests on
At least 10 males and indication)
females per group Ophthalmological examination
Preferred rodent species: Body weight and food/water consumption
rat Haematology (haematocrit, haemoglobin,
Generally oral route of erythrocyte count, total and differential
administration leucocyte count, platelet count, blood
clotting time/potential)
Clinical biochemistry
Histopathology: at least five animals of
each sex per group) for neuropathological
examinations (brain, spinal cord and
peripheral nerves); remaining animals to be
used either for specific neurobehavioural,
neuropathological, neurochemical or
electrophysiological procedures that may
supplement the histopathology or,
alternatively, for routine pathological
evaluations in accordance with the
guidelines for standard repeated dose
toxicity studies
OECD TG 419 (EU B.38): Exposure for 28 days Detailed clinical observations
delayed neurotoxicity of At least three dose levels Body weight and food/water consumption
organophosphorus plus control Clinical biochemistry (neuropathy target
substances – 28-day
At least 12 birds per esterase activity, acetylcholinesterase
repeated dose study
group activity)
Species: domestic laying Gross necropsy (all animals)
hen Histopathology (neural tissue)
NB: T3, thyroid hormone triiodothyronine; T4, thyroid hormone thyroxine; TSH, thyroid stimulating hormone.
Repeated dose dermal and inhalation toxicity studies are not listed in the table.
Table 6. Overview of other in vivo test guideline studies giving information on repeated dose
toxicity

Test Design End points (general toxicity)

OECD TG 416 (EU B.35): Exposure before Clinical observations
two-generation reproduction mating for at least Body weight and food/water consumption
toxicity study one spermatogenic Gross necropsy (all parental animals)
cycle until weaning of Organ weights (reproductive organs, brain,
second generation liver, kidneys, spleen, pituitary, thyroid, adrenal
At least three dose glands and known target organs)
levels plus control Histopathology (reproductive organs, previously
At least 20 parental identified target organ(s) – at least control and
males and females high-dose groups)
per group
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Test Design End points (general toxicity)

OECD TG 415 (EU B.34): Exposure before As in OECD TG 416 / EU B.35
one-generation reproduction mating for at least
toxicity study one spermatogenic
cycle until weaning of
first generation
At least three dose
levels plus control
At least 20 parental
males and females
per group
OECD TG 443 (EU B.56): As described in OECD As described in OECD TG 443 / EU B.56
extended one-generation TG 443 / EU B.56
reproductive toxicity study
OECD TG 414 (EU B.31): Exposure at least Clinical observations
prenatal developmental from implantation to Body weight and food/water consumption
toxicity study 1 or 2 days before Macroscopical examination of all dams for any
expected birth structural abnormalities or pathological changes
At least three dose that may have influenced the pregnancy
levels plus control
At least 20 pregnant
females per group
OECD TG 421 (EU B.63): Exposure from Clinical observations
reproduction/developmental 2 weeks prior to Body weight and food/water consumption
toxicity screening test mating until at least Gross necropsy (adult animals, special attention
postnatal day 4 to reproductive organs)
At least three dose Organ weights (all adult males: testes,
levels plus control epididymides)
At least 8–10 parental Histopathology (reproductive organs in at least
males and females control and high-dose groups)
per group
OECD TG 426 (EU B.53): Exposure at least Clinical observations
developmental neurotoxicity from implantation and Body weight and food/water consumption
study throughout lactation
(postnatal day 20)
At least three dose
levels plus control
At least 20 pregnant
females per group
OECD TG 451 (EU B.32): Exposure for majority Clinical observations (special attention to
carcinogenicity studies of normal lifespan tumour development)
At least three dose Body weight and food consumption
levels plus control Gross necropsy
At least 50 males and Histopathology (all groups – all grossly visible
females per group tumours or lesions suspected of being tumours;
at least control and high-dose groups – brain,
pituitary, thyroid, parathyroid, thymus, lungs,
heart, salivary glands, liver, spleen, kidneys,
adrenals, oesophagus, stomach, duodenum,
jejunum, ileum, caecum, colon, rectum, uterus,
urinary bladder, lymph nodes, pancreas,
gonads, accessory sex organs, female
mammary gland, skin, musculature, peripheral
nerve, spinal cord, sternum with bone marrow
and femur, eyes)
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3.3.3. Specific system/organ toxicity

3.3.3.1. General aspects

For some specific system/organ effects, the EU testing methods and the OECD guidelines may
not provide for adequate characterisation of the toxicity. There may be indications of such effects
in the standard studies for systemic toxicity or from SARs. For adequate characterisation of the
toxicity and, hence, the risk to human health, it may be necessary to conduct studies using other
established methods.

Some specific investigation of organ/systemic toxicity (e.g. hepatotoxicity and nephrotoxicity) is

undertaken as part of the repeated dose toxicity tests. Reproductive toxicity is specifically
examined using appropriate methods (see Section 3.6 of this guidance document). Specific
investigation (or further investigation) of any organ/system toxicity (e.g. kidney, cardiac,
adrenal, thyroid) may sometimes be considered necessary and should be addressed on a case-
by-case basis.

3.3.3.2. Neurotoxicity

3.3.3.2.1. Definition of neurotoxicity

Neurotoxicity is the induction by a chemical of adverse effects in the central or peripheral nervous
system or in the sense organs. It is useful for the purpose of hazard and risk assessment to
differentiate sense-organ-specific effects from other effects that lie within the nervous system.
A substance is considered neurotoxic if it induces a reproducible lesion in the nervous system or
a reproducible pattern of neural dysfunction.

3.3.3.2.2. Introduction

It is recommended that a hierarchical approach is taken in the investigation of the potential

neurotoxicity of substances. The starting point for the testing strategy should be exposure
considerations, in vitro data and SARs, and should proceed using data already available from
base set tests to more specific testing. Thus, any indications of specific or non-specific
neurotoxicity in the acute and repeated dose toxicity tests should be carefully noted. In addition,
if there are already alerts from SARs or available information on the relevant chemical species
or similar substances, and repeated dose toxicity is planned, it would be of benefit to investigate
neurotoxicity within the repeated dose test. The same would apply for consideration of
developmental neurotoxicity investigations within the reproductive toxicity generation tests.

The present EU and OECD 28-day and 90-day oral tests (EU B.7 and OECD TG 407, 1995;
EU B.26 and OECD TG 408, 1998) examine a number of simple nervous system end points (e.g.
clinical observations of motor and autonomous nervous system activity, histopathology of nerve
tissue), which should be regarded as the starting point for evaluation of a substance’s potential
to cause neurotoxicity. It should be recognised that the standard 28- and 90-day tests measure
only some aspects of nervous system structure and function, while other aspects, for example
learning and memory and sensory function, are not or only superficially tested. SAR
considerations may prompt the introduction of additional parameters to be tested in standard
toxicity tests or the investigation of, for example, delayed neurotoxicity (EU B.37 and OECD TG
418; EU B.38 and OECD TG 419). Any indication of the potential neurotoxicity of substances can
also be a trigger for testing for developmental neurotoxicity (see also DWD Guidance Volume I).

If there is no indication of neurotoxicity in humans, no indication in adequately performed acute

and repeated dose toxicity tests and none from SARs, it is not necessary to conduct any
additional tests for neurotoxicity.
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3.3.3.2.3. Structure–activity considerations

Structural alerts are used only as a positive indication of neurotoxic potential. Substance classes
with an alert for neurotoxicity may include organic solvents (for chronic toxic encephalopathy),
organophosphorus compounds (for delayed neurotoxicity) and carbamates (for cholinergic
effects). Several estimation techniques are available, one of which is the rule-based Derek Nexus
system. The rule base comprises the following hazards and structural alerts: organophosphate
(for direct and indirect anticholinesterase activity), N-methyl or N,N-dimethyl carbamate (for
direct anticholinesterase activity) and gamma-diketones (for neurotoxicity). Another repository
of models is the JRC’s (Q)SAR Model Database ((Q)SAR model reporting framework
inventory) ( 46), which is intended to provide information on (Q)SAR models submitted to the JRC
for peer review. It should be noted that no formal adoption process for (Q)SAR models exists or
is provided for under the DWD. The validity, applicability and adequacy of (Q)SAR models are
assessed individually, with the prediction generated for the target chemical ( 47).

3.3.3.2.4. Assessment of available information or results from initial

testing

Signs of neurotoxicity in standard acute or repeated dose toxicity tests may be secondary to
other systemic toxicity or to discomfort from physical effects such as a distended or blocked GI
tract. Nervous system effects seen at dose levels near or above those causing lethality should
not be considered, in isolation, to be evidence of neurotoxicity. In acute toxicity studies where
high doses are administered, clinical signs are often observed that are suggestive of effects on
the nervous system (e.g. observations of lethargy, postural or behavioural changes), and a
distinction should be made between specific and non-specific signs of neurotoxicity.

Neurotoxicity may be indicated by the following signs: morphological (structural) changes in the
central or peripheral nervous system or in special sense organs, neurophysiological changes
(e.g. electroencephalographic changes), behavioural (functional) changes or neurochemical
changes (e.g. neurotransmitter levels).

The type, severity, number and reversibility of the effect should be considered. Generally, a
pattern of related effects is more persuasive evidence of neurotoxicity than one or a few
unrelated effects. Available data should be assessed in a WoE to conclude whether an effect may
be considered a concern for neurotoxicity.

It is important to ascertain whether the nervous system is the primary target organ. The
reversibility of neurotoxic effects should also be considered. Reversible effects may be of high
concern depending on the severity and nature of the effects. In this context, it should be kept
in mind that effects observed in experimental animals that appear harmless might be of high
concern in humans depending on the setting in which they occur (e.g. sleepiness in itself may
not be harmful, but in relation to operating machinery it is an effect of high concern).
Furthermore, the possibility that a permanent lesion has occurred cannot be excluded, even if
the overt effect is transient. The nervous system possesses reserve capacity, which may
compensate for the damage, but the resulting reduction in the reserve capacity should be
regarded as an adverse effect. Compensation may be suspected if a neurotoxic effect slowly
resolves during the lifespan. This could be the case for developmental neurotoxicants (see
Section 3.6). Irreversible neurotoxic effects are of high concern and usually involve structural
changes, although, at least in humans, lasting functional effects (e.g. depression, involuntary

(46) https://jeodpp.jrc.ec.europa.eu/ftp/jrc-opendata/EURL-ECVAM/datasets/QSARDB/LATEST/qsardb.html.
(47) https://echa.europa.eu/documents/10162/13655/pg_report_qsars_en.pdf/407dff11-aa4a-4eef-a1ce-
9300f8460099.
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motor tremor) are suspected to occur as a result of neurotoxicant exposure, apparently without
morphological abnormalities.

3.3.3.2.5. Further neurotoxicity testing

If the data acquired from the standard systemic toxicity tests are inadequate or provide
indications of neurotoxicity that are not sufficiently investigated for risk characterisation, the
nature of further investigation will need to be considered. Additional guidance is provided in
DWD Guidance Volume I and in ECHA’s Guidance on the Biocidal Products Regulation –
Volume III: Human health – Part A: Information requirements ( 48).

3.3.3.3. Immunotoxicity

3.3.3.3.1. Definition of immunotoxicity

Immunotoxicity is defined as any adverse effect on the immune system that can result from
exposure to a range of environmental agents, including chemicals (IPCS, 2012).

An immunotoxic response may occur when the immune system is the target of a chemical insult;
this in turn can result in either immunosuppression and a subsequent decreased resistance to
infection and certain forms of neoplasia, or immune dysregulation, which exacerbates allergy or
autoimmunity reaction. Alternatively, toxicity may arise when the immune system responds to
an antigenic specificity of the chemical as part of a specific immune response (i.e. allergy or
autoimmunity). Changes of immunological parameters may also be a secondary response to
stress resulting from effects on other organ systems. Therefore, it must be recognised that in
principle all chemical substances may be able to influence parameters of the immune system if
administered at a sufficiently high dosage; however, these may not constitute substance-specific
effects. An immunotoxic effect should not be disregarded until an adequate investigation has
been performed.

The IPCS published Guidance for Immunotoxicity Risk Assessment for Chemicals (IPCS
Harmonization Project Document No 10), focused on immunosuppression, inadvertent
immunostimulation and autoimmunity caused by chemical exposure ( 49). Therefore, the IPCS
Guidance for Immunotoxicity Risk Assessment for Chemicals (IPCS, 2012) must be consulted
together with this volume of the DWD guidance when performing the assessment of this end
point.

3.3.3.3.2. Introduction

The toxicological significance of immune responses is currently under discussion by several

scientific groups (e.g. European Centre for Ecotoxicology and Toxicology of Chemicals (ECETOC),
IPCS). Immunotoxicity is of particular concern for test substances that induce toxicity on the
immune system at dose levels below those that induce toxicity at other target sites. If the
immunotoxicity is the critical effect, it is recommended to assess immune effects in the risk
assessment process as for any other toxic effect (IPCS, 1996). As the revised test methods
(OECD TG 407 and 408) begin to be applied routinely, it is expected that the database on the
immunotoxic potential of substances will grow and experience on the evaluation of immune

(48) https://echa.europa.eu/documents/10162/2324906/bpr_guidance_vol_iii_part_a_en.pdf/05e4944d-106e-9305-
21ba-f9a3a9845f93?t=1648525287369.
(49) https://iris.who.int/bitstream/handle/10665/330098/9789241503303-eng.pdf?sequence=1.
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effects will increase. The test guidelines are intended primarily as a screening for
immunotoxicity, and, depending on the results, immediate further testing may be needed.

3.3.3.3.3. Immunosuppression

In vivo animal studies. The basis of the currently recommended approach to assessment of
the potential immunotoxicity of a new substance is that many immunotoxic substances can be
identified in the context of more general toxicity testing, such as the 28-day repeated dose
toxicity study (OECD TG 407), the 90-day repeated dose toxicity study (OECD TG 408) and the
extended one-generation reproduction toxicity study (OECD TG 443). Special studies to
characterise effects of concern for immunotoxicity are used only when necessary for adequate
risk characterisation. The nature of special studies and when they should be conducted need to be
decided on a case-by-case basis. In particular, the use of in vivo tests should not be undertaken
without detailed consideration of the need for such studies (IPCS, 2012).

The protocols of both the EU and the OECD TG 407 (28-day) and OECD TG 408 (90-day) studies
include the measurement of thymus and spleen weights and histopathological examination of
certain lymphoid tissues (i.e. thymus, draining and distant lymph nodes, Peyer’s patches, bone
marrow section) in addition to the total and differential white blood cell counts and spleen
histopathology required by the previous EU method. These tissues all have immunological
functions, and changes to them can be indicative of adverse effects on the immune system.

Indications of immunotoxicity from standard repeated dose toxicity studies may include one or
more of the following signs:

• morphological changes of lymphoid organs and tissues, including bone marrow (e.g. altered
cellularity/size of major compartments), in the absence of other relevant substance-specific
effects that cause this change;

• weight changes of lymphoid organs in the absence of other relevant substance-specific

effects that cause this change;

• changes in haematology parameters (e.g. white blood cell number, differential cell counts of
lymphocytic, monocytic and granulocytic cells);

• changes in clinical chemistry parameters (e.g. serum protein levels, immunoglobulin

concentrations if determined).

It has to be noted that the effects listed above may not be sufficient by themselves to come to
a conclusion on immunotoxicity but should rather be considered in a WoE approach, taking into
account all the available information.

If there are no indications of immunotoxicity in the 28-day (or 90-day) toxicity test and none
from SARs, no further specific investigation for immunotoxicity is normally required.

However, the abovementioned effects may also raise concern for mechanisms/modes of action
associated with developmental immunotoxicity. For regulatory purposes, such concern is further
investigated in cohort 3 of the extended one-generation reproductive toxicity study via the T-
cell-dependent antibody response assay, which is a so-called functional assay. However, it has
to be considered that, as the concern for justifying the request for developmental immunotoxicity
study is based on data generated in adult animals, it is possible to miss substances causing
immunomodulation in developing animals.

Further testing to investigate immune function (e.g. a T-cell function test for substances that
cause histopathological changes in the thymus, host resistance models) should be conducted
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only if the results of such studies can be interpreted in relation to the risk assessment for the
relevant chemical species. In many cases, the observation of the morphological changes or
changes in haematology and clinical chemistry parameters, together with a NOAEL/LOAEL/BMD
for those changes, will be sufficient for screening. Functional assays may provide valuable
information for identifying immunotoxic effects and, in some cases, can be more sensitive than
non-functional assays. However, it should be noted that the observation of the immunological
changes discussed above may not necessarily reflect a direct immunotoxic effect but may be
secondary to other effects.

For immunotoxicity testing, a list of reviews of principles and methods from the IPCS is provided
in its guidance on information requirements (IPCS, 2012), along with a list of available test
methods, including assays for the assessment of autoimmunity.

Non-animal methods

Depending on the mode of action suspected and the effects observed, in vitro studies may be
able to provide useful mechanistic insight for the assessment of immunotoxicity. To assess the
immunotoxicity of chemicals, a huge number of non-animal testing models have been published;
however, only a few cell-based test methods have reached the stage of validation and
acceptance by international regulatory bodies, such as OECD TG 444A (IL-2 Luc assay).
Therefore, the majority of reported tests still fall into the category of non-guideline methods;
however, the results obtained from these tests may nevertheless be used to support regulatory
decisions. Following the developments in the field of immunotoxicity is recommended, as new
methods and approaches to assess immunotoxicity may become available.

A detailed review paper on non-animal approaches that could be used to test chemicals for their
potential immunotoxic effects has been published under the Inter-Organization Programme for
the Sound Management of Chemicals 50. The detailed review paper aims to present and discuss
the application and interpretation of in vitro immunotoxicity assays, covering mainly
immunosuppression, and to define an in vitro tiered approach to testing and assessment.

Currently, there are some initiatives by the Johns Hopkins Center for Alternatives to Animal
Testing to establish a framework to foster the refinement and development of new alternative
test methods suitable for screening of developmental immunotoxic compounds ( 51).

Despite the need for an in vitro tiered system to evaluate immunotoxicity, there is currently no
combination of OECD test guidelines that can be used to conclude on immunosuppression in
vitro. It is clear that one assay alone will not be able to cover all of the potential adverse effects
of chemicals on the immune system and that a larger set of assays that will cover the spectrum
of immunotoxicity is needed.

A tiered testing strategy is proposed to assess immunotoxicity in vitro (Gennari et al., 2005). In
the proposed tiered approach, pre-screening for direct immunotoxicity in vitro begins by
evaluating myelotoxicity (tier 1). Compounds capable of damaging or destroying bone marrow
cells will most likely have immunotoxic effects, as the majority of immune cells are derived from
a common precursor located in adult bone marrow. If compounds are not potentially myelotoxic,
they should be tested for direct leukotoxicity, defined as toxicity to any cell of the lymphoid or
myeloid lineages (tier 2). Compounds should then be tested for immunotoxicity using various
approaches, such as T-cell-dependent antibody response, lymphocyte proliferation assay, mixed
leucocyte reaction, natural killer cell assay, dendritic cell maturation assay, human whole-blood
cytokine release assay (HWBCRA) and fluorescent cell chip assay (tier 3).

(50) ENV/CBC/MONO(2022)16 (https://one.oecd.org/document/ENV/CBC/MONO(2022)16/En/pdf).

(51) https://caat.jhsph.edu/developmental-immunotoxicity/.
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Among these assays, the HWBCRA has the advantage of comprising multiple cell types in their
natural proportion and environment, allowing the evaluation of both monocyte and lymphocyte
functions by using selective stimuli (Langezaal et al., 2001, 2002), while omics techniques can
provide additional mechanistic understanding and hold promise for the characterisation of
classes of compounds and prediction of specific toxic effects (Hochstenbach et al., 2010; Shao
et al., 2014; Schmeits et al., 2015). The IL-2 Luc assay also allows high-throughput analysis
(Kimura et al., 2018), which will greatly expand the opportunities for in vitro testing. The colony
forming unit – granulocyte/macrophage assay, the HWBCRA as a pyrogen test and the IL-2 Luc
assay have undergone validation for reproducibility and predictive capacity.

It is likely that multiple assays will be required to define immunotoxicants because of the
complexity and varied components of the immune system (e.g. innate or adaptive immune
responses). For example, for a combination of in vitro assays that may be used to predict the
immunotoxicity of chemicals, please refer to the detailed review paper ( 52).

3.3.3.4. Effects on the endocrine system

The endocrine system consists of a set of glands that includes the thyroid, gonads and the
adrenal glands, and the hormones they produce, such as thyroxine, oestrogen, testosterone and
adrenaline, which help guide the development, growth, reproduction and behaviour of animals,
including human beings.

According to the CLP regulation (Regulation (EC) No 1272/2008 ( 53)), Annex 1,

Section 3.11.1.1(a), ‘“endocrine disruptor” means a substance or a mixture that alters one or
more functions of the endocrine system and consequently causes adverse effects in an intact
organism, its progeny, populations or subpopulations’.

EDs are believed to interfere with the endocrine system, leading to adverse effects by one or
more modes of action, depending on the individual substance. There are three possible general
modes of action, as listed below:

• by mimicking the action of a naturally produced hormone such as oestrogen or testosterone,

thereby setting off similar chemical reactions in the body;

• by blocking the receptors in cells receiving the hormones (hormone receptors), thereby
preventing the action of normal hormones;

• by affecting the synthesis, transport, metabolism and excretion of hormones, thus altering
the concentration of normal hormones.

In relation to hazard identification, the procedures described in the ECHA/EFSA guidance on the
identification of EDs in the context of Regulations (EU) No 528/2012 and (EC) No 1107/2009 ( 54)
should be considered.

3.3.3.5. Remaining uncertainty

The following elements contribute to the uncertainty in the determination of a threshold for the
critical effects and the selection of the AF (see also Section 4.2.2.2).

(52) https://one.oecd.org/document/ENV/CBC/MONO(2022)16/En/pdf.
(53) https://eur-lex.europa.eu/eli/reg/2008/1272/oj.
(54) https://echa.europa.eu/fi/-/guidance-on-identifying-endocrine-disruptors-published.
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3.3.3.5.1. Threshold of the critical effect

In the determination of the overall threshold for repeated dose toxicity, all relevant information
is evaluated to determine the lowest dose that induces an adverse effect (i.e. LOAEL or lowest
observed adverse effect concentration (LOAEC)) and the highest level with no biologically or
statically significant adverse effects (i.e. NOAEL or NOAEC). In this assessment, all toxicological
responses are taken into account and the critical effect is identified. The uncertainty in the
threshold depends on the strength of the data and is largely determined by the design of the
underlying experimental data. Parameters such as group size, study type/duration and the
methodology need to be taken into account in the assessment of the uncertainty in the threshold
of the critical effect(s).

The NOAEL is typically used as the starting point for the derivation of the threshold level (e.g.
DNEL). In cases where a NOAEL has not been achieved, a LOAEL may be used, provided the
available information is sufficient for a robust hazard assessment. The BMD may also be used as
the starting point.

The selection of the NOAEL or LOAEL is usually based on the dose levels used in the most relevant
toxicity study, without considering the shape of the dose–response curve. Therefore, the
NOAEL/LOAEL may not reflect the true threshold for the adverse effect. In contrast, the BMD is
a statistical approach for the determination of the threshold and relies on the dose–response
curve. Alternatively, mathematical curve-fitting techniques or statistical approaches exist for
determining the threshold for an adverse effect. The use of such approaches (e.g. BMD) to
estimate the threshold should be considered on a case-by-case basis and are usually used for
higher-tier hazard characterisation refinement.

3.3.3.5.2. Other considerations

Another situation may arise when testing is not technically possible, a waiving option indicated
in Part 2, Section 2 of Annex V to Commission Implementing Decision (EU) 2024/365. In such
cases, approaches such as (Q)SAR, category formation and read-across may be helpful in the
hazard characterisation; they should also be considered for information that might be suitable
as a surrogate for a dose descriptor. Alternatively, generic threshold approaches, for example
the threshold of toxicological concern, might be considered for the starting point of a risk
characterisation as risk management tools to estimate negligible exposure potential (for more
information, see Appendix 1-4 of Guidance on the Biocidal Products Regulation – Volume III
human health – Assessment & evaluation (Parts B+C) ( 55)).

3.3.4. Conclusions on repeated dose toxicity

Potentially relevant studies should be judged for quality, and studies of high quality should be
given more weight than those of lower quality. When both epidemiological and experimental
data are available, similarity of effects between humans and animals is given more weight. If
the mechanism or mode of action is well characterised, this information is used in the
interpretation of observed effects in either human or animal studies. WoE is not to be interpreted
as simply tallying the number of positive and negative studies, nor does it imply an averaging
of the doses or exposures identified in individual studies that may be suitable as starting points
for risk assessment. The study or studies used for the starting point are identified by an informed
and expert evaluation of all the available evidence. Recommendations on how to perform a WoE

(55)
https://echa.europa.eu/documents/10162/2324906/biocides_guidance_human_health_ra_iii_part_bc_en.pdf/
30d53d7d-9723-7db4-357a-ca68739f5094?t=1512979002065.
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are provided in Section 3.1.4 of this guidance, as well as in the EFSA guidance on the use of the
WoE approach ( 56) and in Chapter R.4 of the ECHA guidance ( 57).

The available repeated dose toxicity data should be evaluated in detail for a characterisation of
the health hazards upon repeated exposure. In this process, an assessment of all toxicological
effect(s), their dose–response relationships and possible thresholds are taken into account. The
evaluation should include an assessment of the severity of the effect, its adversity, or if it is a
precursor to a more significant effect or secondary to general toxicity. Correlations between
changes in several parameters, for example between clinical or biochemical measurements,
organ weights and (histo)pathological effects, will be helpful in the evaluation of the nature of
the effects. Further guidance on this issue can be found in the publications of ECETOC (2021) ( 58)
and Inchem (2020) ( 59).

The effects data are also analysed for indications of potential serious toxicity of target organs or
specific organ systems (e.g. neurotoxicity or immunotoxicity), delayed effects or cumulative
toxicity. Furthermore, the evaluation should take into account the study details and determine
if the exposure conditions and duration and the parameters studied are appropriate for an
adequate characterisation of the toxicological effect(s).

If an evaluation allows the conclusion that the information of the repeated dose toxicity is
adequate for a robust characterisation of the toxicological hazards, including an estimate of a
dose descriptor (BMD/NOAEL/LOAEL), and the data are adequate for risk assessment under the
DWD, no further testing will be necessary unless there are indications of further risk.

Another consideration is whether the study duration was appropriate for an adequate expression
of the toxicological effects. If the critical effect involves serious specific system or target organ
toxicity (e.g. haemolytic anaemia, neurotoxicity or immunotoxicity), delayed effects or
cumulative toxicity, and a threshold has not been established, dose extrapolation may not be
appropriate and further studies are required. In this case a specialised study is likely to be more
appropriate for an improved hazard characterisation and should be considered instead of a
standard short-term rodent or subchronic toxicity test at this stage.

3.3.5. Concluding on suitability for risk assessment

It is necessary to identify the so-called dose descriptor, that is, an appropriate threshold dose
for the critical effect as the starting point for DNEL derivation, namely a BMD or NOAEL. If a
NOAEL cannot be identified, the LOAEL may be used instead, provided the data are adequate for
a robust hazard assessment.

It must be noted that the dose descriptor should be route specific. As explained in Section 5, the
oral route is the relevant one for relevant species migrated in the drinking water. Thus, in cases
when only animal data with dermal or inhalation exposure are available, and humans are
exposed via ingestions, as is the case for drinking water, a threshold level for the oral route is
needed, that is, route-to-route extrapolation is needed. Guidance on this route-to-route
extrapolation is provided in Chapter 4 ‘Hazard assessment’.

(56) https://efsa.onlinelibrary.wiley.com/doi/epdf/10.2903/j.efsa.2017.4971.
(57) https://echa.europa.eu/documents/10162/17235/information_requirements_r4_en.pdf/d6395ad2-1596-4708-
ba86-0136686d205e?t=1323782558175.
(58) https://www.ecetoc.org/wp-content/uploads/2021/10/ECETOC-TR-138-Guidance-on-Dose-Selection_Final.pdf.
(59) https://cdn.who.int/media/docs/default-source/food-safety/publications/chapter5-dose-response.pdf?sf.
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3.4. Mutagenicity

The section on mutagenicity in DWD Guidance Volume I should be considered together with the
elements described in this section for the assessment of mutagenicity.

Section 1 of Annex VI to Commission Implementing Decision (EU) 2024/365 lists the criteria
when no hazard assessment needs to be performed.

3.4.1. Definition

Mutagenicity refers to the induction of permanent transmissible changes in the amount or

structure of the genetic material of cells or organisms. These changes may involve a single gene
or gene segment, a block of genes or chromosomes. Several mutagenicity end points (i.e. gene
mutation, clastogenicity and aneugenicity) can be distinguished. The term gene mutation refers
to permanent changes in the base sequences of a certain gene. The term clastogenicity is used
for agents giving rise to structural chromosome aberrations. A clastogen can cause breaks in
chromosomes that result in the loss or rearrangement of chromosome segments. Aneugenicity
(aneuploidy induction) refers to the effects of agents that give rise to a change (gain or loss) in
chromosome number in cells. An aneugen can cause loss or gain of chromosomes, resulting in
cells that do not have an exact multiple of the haploid number. For example, three chromosomes
of the same number, or trisomy (e.g. trisomy 21, characteristic of Down’s syndrome), is a form
of aneuploidy.
Genotoxicity is a broader term that refers to processes that alter the structure, information
content or segregation of DNA and are not necessarily associated with mutagenicity. Thus, tests
for genotoxicity include tests that provide an indication of induced damage to DNA (but not direct
evidence of mutation) via effects such as DNA strand breaks, unscheduled DNA synthesis, sister
chromatid exchange and DNA adduct formation or mitotic recombination, as well as tests for
mutagenicity.
The chemical and structural complexity of the chromosomal DNA and associated proteins of
mammalian cells, and the multiplicity of ways in which changes to the genetic material can be
affected make it difficult to give more precise, discrete definitions.
In the risk assessment of the relevant chemical species, it is necessary to address the potential
effect of mutagenicity. It can be expected that some of the available data will have been derived
from tests conducted to investigate potentially harmful effects on genetic material
(genotoxicity). Hence, both the terms mutagenicity and genotoxicity are used in this document.

The aim of testing for genotoxicity is to assess the potential of relevant chemical species to
induce genotoxic effects that may lead to cancer or cause mutagenicity in humans. Genotoxicity
data are also used in risk characterisation of substances.

Alterations to the genetic material of cells may occur spontaneously or be induced as a result of
exposure to ionising or ultraviolet radiation, or to genotoxic substances. In principle, human
exposure to substances that are mutagens may result in increased frequencies of mutations
above the background level.

Mutations in somatic cells may be lethal or may be transferred to daughter cells with deleterious
consequences for the affected organism (e.g. cancer may result when they occur in proto-
oncogenes, tumour suppressor genes and/or DNA repair genes), ranging from trivial to
detrimental or lethal.

There is considerable evidence of a positive correlation between the mutagenicity of substances

in vivo and their carcinogenicity in long-term studies with animals. Genotoxic carcinogens are
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chemicals for which the most plausible mechanism of carcinogenic action involves genotoxicity
(Kang et al., 2013).

Heritable damage to the offspring, and possibly to subsequent generations, may follow after
parents are exposed to substances that are mutagens if mutations are induced in parental germ
cells. To date, all known germ cell mutagens are also mutagenic in somatic cells in vivo.
Substances that are mutagenic in somatic cells may produce heritable effects if they, or their
active metabolites, reach the genetic material of germ cells. Conversely, substances that do not
induce mutations in somatic cells in vivo would not be expected to be germ cell mutagens.

3.4.2. Data to be used in the assessment of genotoxicity/mutagenicity

Genotoxicity is a complex end point and requires evaluation by expert judgement. For both steps
of the effects assessment, namely hazard identification and dose (concentration)–response
(effect) assessment, it is very important to evaluate the data with regard to their adequacy and
completeness. The evaluation of adequacy must address the reliability and relevance of the data
in a way outlined in the introductory section. The completeness of the data refers to the
conclusion drawn on whether the available information is adequate and fulfils the requirements
under Section 2 of Annex V to Commission Implementing Decision (EU) 2024/365. Such a
conclusion relies on WoE approaches, which categorise available information based on the
methods used: guideline tests, non-guideline tests and other types of information that may
justify adaptation of the standard testing regime. Such a WoE approach also includes an
evaluation of the available data as a whole, that is, both over and across toxicological end points.

This approach provides a basis for deciding whether further information is needed on end points
for which specific data appear inadequate or not available, or whether the requirements are
fulfilled.

3.4.2.1. Non-human data for mutagenicity

3.4.2.1.1. Non-testing data for mutagenicity

See Section 3.1.2.2.5 (on (Q)SARs) and Section 3.1.3 (on the read-across approach).

3.4.2.1.2. Testing data for mutagenicity

Test methods preferred for use are listed in Tables 7–9. Some of these have officially adopted
EU/OECD guidelines for genotoxicity testing.
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In vitro data

Table 7. In vitro test methods

Test Genotoxic end points Principle of the test method EU/OECD guideline
measured
Bacterial reverse Gene mutations The test uses amino acid requiring EU B.13/14 / OECD
mutation test strains of bacteria to detect TG 471
(reverse) gene mutations (point
mutations and frameshifts).
In vitro Gene mutations The test identifies chemicals that EU B.17 / OECD
mammalian cell induce gene mutations in the Hprt TG 476
gene mutation and Xprt genes of established cell
tests using the lines.
Hprt and Xprt
genes
In vitro Gene mutations and The test identifies chemicals that EU B.67 / OECD
mammalian cell structural chromosome induce gene mutations in the TK TG 490
gene mutation aberrations gene of the L5178Y mouse
tests using the lymphoma or TK6 human
thymidine kinase lymphoblastoid cell line. If colonies
(TK) gene in a TK mutation test are scored
using the criteria of normal growth
(large / early appearing) and slow
growth (small/late-appearing)
colonies, gross structural
chromosome aberrations may be
measured, since mutant cells that
have suffered the most extensive
genetic damage have prolonged
doubling times and are more likely
to form small/late-appearing
colonies.
In vitro Structural (and The test identifies chemicals that EU B.10 / OECD
mammalian numerical) chromosome induce chromosome aberrations in TG 473
chromosome aberrations cultured mammalian established
aberration test cell lines, cell strains or primary
cell cultures. An increase in
polyploidy may indicate that a
chemical has the potential to
induce numerical chromosome
aberrations. However, this test is
not designed to measure
aneuploidy.
In vitro Structural and numerical The test identifies chemicals that EU B.49 / OECD
mammalian chromosome aberrations induce micronuclei in the TG 487
micronucleus test cytoplasm of interphase cells.
These micronuclei may originate
from acentric fragments or whole
chromosomes, and the test thus
has the potential to detect both
clastogenic and aneugenic
chemicals.
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Animal data

Table 8. Somatic cells – in vivo test methods

Test Genotoxic end Principle of the test method EU/OECD

points measured guideline
In vivo mammalian Structural (and The test identifies chemicals that EU B.11 / OECD
bone marrow numerical) induce structural chromosome TG 475
chromosome chromosome aberrations in the bone marrow cells of
aberration test aberrations animals, usually rodents. An increase
in polyploidy may indicate that a
chemical has the potential to induce
numerical chromosome aberrations.
In vivo mammalian Structural and The test identifies chemicals that cause EU B.12 / OECD
erythrocyte numerical micronuclei in erythroblasts sampled TG 474
micronucleus test chromosome from bone marrow and/or peripheral
aberrations blood cells of animals, usually rodents.
These micronuclei may originate from
acentric fragments or whole
chromosomes, and the test thus has
the potential to detect both clastogenic
and aneugenic chemicals.
Unscheduled DNA DNA repair The test identifies chemicals that EU B.39 / OECD
synthesis test with induce DNA repair (measured as TG 486
mammalian liver unscheduled ‘DNA’ synthesis) in liver
cells in vivo (a) cells of animals, commonly rats. The
test is usually based on the
incorporation of tritium labelled
thymidine into the DNA by repair
synthesis after excision and removal of
a stretch of DNA containing a region of
damage.
Transgenic rodent Gene mutations / The test can measure gene mutations EU B.58 (b) /
somatic and germ in any tissue of an animal and may, OECD TG 488
cell gene mutation therefore, also be used in specific site
assay of contact tissues.
In vivo mammalian DNA strand breaks in The test can measure DNA strand EU B.62 / OECD
alkaline comet assay eukaryotic cells breaks in any tissue of an animal and TG 489
may, therefore, also be used in specific
site of contact tissues.
Mammalian Gene mutations The erythrocyte Pig-a assay uses an OECD TG 470
erythrocyte Pig-a endogenous mammalian gene, the
gene mutation assay phosphatidylinositol glycan class A gene
(Pig-a), as a reporter of somatic cell
gene mutations in erythroid precursor
cells, primarily found in the bone
marrow. The test can identify
substances that cause gene mutations
in these precursor cells, which are
reflected in erythrocytes sampled from
peripheral blood cells of animals,
usually rodents. The test can be
conducted without euthanising the
animals, which facilitates integration of
the Pig-a assay into many in vivo
rodent testing protocols.
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(a) This test is no longer considered appropriate; therefore, data from existing tests may be used in the risk
assessment, but no new tests should be carried out for the purpose of Commission Implementing Decision (EU)
2024/365 (see Commission Regulation 2023/464 of 3.3.2023 amending, for the purpose of its adaptation to technical
progress, the Annex to Regulation (EC) No 440/2008 laying down test methods pursuant to Regulation (EC)
No 1907/2006 of the European Parliament and of the Council on the registration, evaluation, authorisation and restriction
of chemicals (C(2023)1099 final), https://environment.ec.europa.eu/publications/commission-regulation-amending-
purpose-its-adaptation-technical-progress-annex-regulation-ec-no_en).
(b) Does not include the latest update to OECD TG 488.

Table 9. Germ cells – in vivo test methods

Test Genotoxic end points Principle of the test method EU/OECD

measured guideline
Mammalian Structural (and numerical) The test measures structural EU B.23 /
spermatogonial chromosome aberrations chromosome aberrations in OECD TG 483
chromosome mammalian, usually rodent,
aberration test spermatogonial cells and is, therefore,
expected to be predictive of induction
of heritable mutations in germ cells. An
increase in polyploidy may indicate
that a chemical has the potential to
induce numerical chromosome
aberrations. This test is not designed
to measure numerical abnormalities.
Rodent Structural and numerical The test measures dominant lethal EU B.22 /
dominant chromosome aberration effects causing embryonic or fetal OECD TG 478
lethal test (a) death resulting from inherited
dominant lethal mutations induced in
germ cells of an exposed parent,
usually the male. It is generally
accepted that dominant lethals are due
to structural and numerical
chromosome aberrations. Rats or mice
are recommended as the test species.
Transgenic Gene mutation The test measures gene mutations in EU B.58 (b) /
rodent somatic spermatocytes of an animal and may, OECD TG 488
and germ cell therefore, be used to obtain
gene mutation information about the mutagenic
assay activity of a chemical in germ cells.
(a) Data from existing tests may be used in the risk assessment; however, due to the known limitations of the test
(e.g. high number of animals), no new tests should be carried out for the purpose of Commission Implementing Decision
(EU) 2024/365.

(b) Does not include the latest update to OECD TG 488.

It should be noted that, although at present the in vivo mammalian alkaline comet assay as
described in OECD TG 489 is not considered appropriate for measuring DNA strand breaks in
mature germ cells, in some cases available in vivo comet data may include analysis of male
gonadal cells from the seminiferous tubules. If such results are available, they may be relevant
for the overall assessment of germ cell mutagenicity.

Evaluation of genotoxicity test data should be carried out with care.

Each test guideline contains criteria for the acceptability of the study based on important
parameters related to the study design and test conditions (e.g. acceptable cell type or animal
species, number of cells used and scored or animals tested per group, dose/concentration levels
and the number of test doses/concentrations, recommended negative and positive controls,
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treatment schedule, exposure and sampling time(s), acceptable levels of (cyto)toxicity, evidence
of target tissue exposure, laboratory proficiency demonstration) and criteria for the evaluation
and interpretation of results (definition of clearly positive and clearly negative responses based
on, for example, statistical analysis or threshold values, or comparison with historical control
ranges for the negative and positive controls).

In addition, further aspects described below need to be considered to determine the validity of
study results.

Regarding positive findings, responses generated only at highly toxic/cytotoxic concentrations

should be interpreted with caution, and the presence or absence of a dose–response relationship
should be considered.

Particular points to take into account when evaluating negative test results include the following.

• The doses or concentrations of the test substance used (were they high enough?).

• Was the test system used sensitive to the nature of the genotoxic changes that might have
been expected? For example, some in vitro test systems will be sensitive to point mutations
and small deletions but not to mutagenic events that create large deletions.

• The volatility of the test substance (were concentrations maintained in tests conducted in
vitro?).

• For studies in vitro, the possibility of metabolism not being active in the system, including
those in extrahepatic organs.

• Was the test substance taken up by the test system used for in vitro studies?

• Was a sufficient number of cells scored/sampled for studies in vitro? Has the appropriate
number of samples / technical replicates been scored to support statistical significance of the
putative negative result?

• For studies in vivo, was/were the most appropriate tissue(s) sampled? Did the substance
reach the target organ? Or was the substance expected to act only at the site of contact due
to its high reactivity or insufficient systemic availability (also taking TK data into
consideration; for example, rate of hydrolysis and electrophilicity may be factors that need
to be considered)?

• For studies in vivo, was sampling appropriate? (Was a sufficient number of animals used?
Were sufficient sampling times used? Was a sufficient number of cells scored/sampled?)

Different results between different test systems should be evaluated with respect to their
individual significance. Examples of points to be considered are as follows.

• Different results obtained in non-mammalian systems and in mammalian cell tests may be
addressed by considering possible differences in substance uptake, metabolism or the
organisation of genetic material. Although the results of mammalian tests may be considered
of higher significance, additional data may be needed to resolve contradictions.

• If the results of indicator tests detecting putative DNA lesions (e.g. DNA binding, DNA
damage, DNA repair, sister chromatid exchange) are not in agreement with results obtained
in tests for mutagenicity, the results of mutagenicity tests are generally of higher
significance, provided that appropriate mutagenicity tests have been conducted. This is
subject to expert judgement.
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• If different findings are obtained in vitro and in vivo, in general, the results of in vivo tests
indicate a higher degree of reliability. However, for evaluation of negative results in vivo, it
should be considered whether the most appropriate tissues were sampled and whether there
is adequate evidence of target tissue exposure. The assessment of contradicting data of
comparable quality assessing similar end points should be done on a case-by-case basis,
while considering the limitations of the respective source of the data.

• The sensitivity and specificity of different test systems may vary for different classes of
substances. If available testing data for other related substances permits assessment of the
performance of difference assays for the class of substance under evaluation, the result from
the test system known to produce more accurate responses would be given higher priority.

Different results may also be also available from the same test performed by different
laboratories or on different occasions. In this case, expert judgement should be used to evaluate
the data and to reach an overall conclusion. In particular, the quality of each of the studies and
of the data provided should be evaluated, with special consideration of the study design,
reproducibility of data, dose–effect relationships and biological relevance of the findings. The
purity of the test substance may also be a factor to take into account. In cases where an
EU/OECD test guideline is available for a test method, the quality of a study using the method
is regarded as being higher if it was conducted in compliance with the requirements stated in
the test guideline. Furthermore, compared to non GLP-studies, studies compliant with GLP for
the same assay generally provide more documentation and details of the study, which are
important factors to consider when assessing study reliability/quality.

When assessing the potential mutagenicity of a substance or considering the need for further
testing, data from various tests and genotoxic end points may be found. Both the strength and
the weight of the evidence should be taken into account. The strongest evidence will be provided
by modern, well-conducted studies with internationally established test protocols and well-
characterised test substance. For each test type and each genotoxic end point, there should be
a separate WoE analysis. It is not unusual for positive evidence of mutagenicity to be found in
just one test type or for only one end point. In such cases the positive and negative results for
different end points are not necessarily conflicting but may illustrate the advantage of using test
methods for a variety of genetic alterations to increase the probability of identifying substances
with mutagenic potential. Hence, results from methods testing different genotoxic end points
should not be combined in an overall WoE analysis but should be subjected to such analysis
separately for each end point. Based on the whole dataset, one has to consider whether an
appropriate conclusion/assessment can be made or whether there are data gaps. If there are
data gaps, further testing should be considered.

3.4.2.2. Human data

See Section 3.1.2.2.3.

3.4.3. Remaining uncertainty on mutagenicity

Reliable data can be generated from well-designed and well-conducted studies in vitro and in
vivo. However, due to the lack of human data available and the degree of uncertainty, which is
inherent in testing, a certain level of uncertainty remains when extrapolating these testing data
to the effect in humans.
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3.4.4. Concluding on suitability for risk assessment

3.4.4.1. Considerations on dose–response shapes and mode of action of

mutagenic substances in test systems

Consideration of the dose (concentration)–response effect relationship and possible mechanisms

of action is an important component of a risk assessment. The default assumption for genotoxic
substances has for a long time been that they have a linear dose (concentration)–response
relationship. However, this assumption has been challenged by experimental evidence showing
that both direct and indirect acting genotoxins can possess non-linear or threshold dose
(concentration)–response curves.

Examples of mechanisms of genotoxicity that may be demonstrated to lead to non-linear or

threshold dose (concentration)–response (effect) relationships include extremes of pH, ionic
strength and osmolarity, inhibition of DNA synthesis, alterations in DNA repair, overloading of
defence mechanisms (antioxidants or metal homeostatic controls), interaction with microtubule
assembly leading to aneuploidy, topoisomerase inhibition, high cytotoxicity, metabolic overload
and physiological perturbations (e.g. induction of erythropoiesis) (Dearfield et al., 2011). The
mechanisms underlying non-linear or threshold dose (concentration)–response relationships for
some DNA reactive genotoxic substances such as alkylating agents seem to be linked to DNA
repair capacity.

Assessment of the significance to be assigned to genotoxic responses mediated by such

mechanisms would include assessing whether the underlying mechanism can be induced at
substance concentrations that can be expected to occur under relevant in vivo conditions.

In general, several concentrations/doses are tested in genotoxicity assays. At least three

experimental concentrations/doses have to be tested, as recommended in the OECD test
guidelines for genotoxicity. Determination of experimental dose (concentration)-dependent
effects is one of several pieces of experimental information that are important for assessing the
genotoxic potential of a substance and may be used as indicated below. It should be recognised
that not all of these considerations may be applicable to in vivo data.

• The OECD overview document of the genotoxicity test guidelines (OECD, 2017) lists the
relevant criteria to be fulfilled for a result to be considered a clear positive one: (1) the
increase in genotoxic response is concentration- or dose-related, (2) at least one of the data
points exhibits a statistically significant increase compared with the concurrent negative
control and (3) the statistically significant result is outside the distribution of the historical
negative control data (e.g. 95 % confidence interval). In practice, the criterion for a dose
(concentration)-related increase in genotoxicity will be most helpful for in vitro tests, but
care is needed to check for cytotoxicity or cell cycle delay, which may cause deviations from
a dose (concentration)–response-related effect in some experimental systems.

• Genotoxicity tests are not designed to support derivation of no-effect levels. However, the
lowest dose with an observed effect (i.e. the lowest observed effect dose) may, on certain
occasions, be a helpful tool in risk assessment. This is true specifically for genotoxic effects
caused by threshold mechanisms, for example aneugenicity. Furthermore, it can give an
indication of the mutagenic potency of the substance in the test at issue. Modified studies
with additional dose or concentration points and improved statistical power may be useful in
this regard. The BMD approach presents several advantages over the no observed effect
dose / lowest observed effect dose approach and can be used as an alternative strategy for
dose (concentration)–response assessment (see ECHA’s Guidance on Information
Requirements and Chemical Safety Assessment – Chapter R.8: Characterisation of dose
[concentration]–response for human health).
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• Unusual shapes of dose (concentration)–response curves may contribute to the identification

of specific mechanisms of genotoxicity. For example, extremely steep increases suggest an
indirect mode of action or a metabolic switching, which could be confirmed by further
investigation.

3.5. Carcinogenicity

The section on carcinogenicity in DWD Guidance Volume I should be considered together with
the elements described in this section for the assessment of carcinogenicity.

Section 1 of Annex VI to Commission Implementing Decision (EU) 2024/365 lists the criteria
when no hazard assessment needs to be performed.

3.5.1. Definition

A substance is defined as carcinogenic if it induces cancer or increases its incidence. Substances

that have induced benign and malignant tumours in well-performed experimental studies on
animals are also considered to be presumed or suspected human carcinogens unless there is
strong evidence that the mechanism of tumour formation is not relevant for humans (Guidance
on the Application of the CLP Criteria ( 60)). Carcinogenic substances can increase the tumour
incidence and/or malignancy and shorten the time to tumour occurrence. Benign tumours that
are considered to have the potential to progress to malignant tumours are generally considered
along with malignant tumours. Substances may induce cancer by any route of exposure, but
carcinogenic potential and potency may depend on the conditions of exposure (e.g. route, level,
pattern and duration of exposure). In the context of the DWD, the oral route is considered the
most appropriate route of exposure. Carcinogens may be identified from epidemiological studies,
animal experiments and/or other appropriate means that may include (Q)SAR analyses and/or
extrapolation from structurally similar substances (read-across approach). Each strategy for
identifying potential carcinogens is discussed in detail below. The determination of the
carcinogenic potential of a substance is based on a WoE approach. Classification criteria are
given in the CLP regulation.

The process of carcinogenesis involves the transition of normal cells into cancer cells via a
sequence of stages that entail both genetic alterations (i.e. mutations) and non-genetic events.
Non-genetic events are defined as those alterations/processes that are mediated by mechanisms
that do not affect the primary sequence of DNA and yet increase the incidence of tumours or
decrease the latency time for the appearance of tumours. For example, altered growth and death
rates, (de)differentiation of the altered or target cells and modulation of the expression of specific
genes associated with the expression of neoplastic potential (e.g. tumour suppressor genes or
angiogenesis factors) are recognised to play an important role in the process of carcinogenesis
and can be modulated by a chemical agent in the absence of genetic change to increase the
incidence of cancer.

Carcinogenic chemicals have conventionally been divided into two categories according to the
presumed mode of action: genotoxic or non-genotoxic. Genotoxic modes of action involve
genetic alterations caused by the chemical interacting directly with DNA to result in a change in
the primary sequence of DNA. A chemical can also cause genetic alterations indirectly following
interaction with other cellular processes (e.g. secondary to the induction of oxidative stress).
Non-genotoxic modes of action include epigenetic changes, that is, effects that do not involve
alterations in DNA but that may influence gene expression, altered cell–cell communication, or
affect other factors involved in the carcinogenic process. For example, chronic cytotoxicity with
subsequent regenerative cell proliferation is considered a mode of action by which tumour

(60) http://echa.europa.eu/web/guest/guidance-documents/guidance-on-clp.
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development can be enhanced: the induction of urinary bladder tumours in rats may, in certain
cases, be due to persistent irritation/inflammation, tissue erosion and regenerative hyperplasia
of the urothelium following the formation of bladder stones. Other modes of non-genotoxic action
can involve specific receptors (e.g. peroxisome proliferator-activated receptor alpha (PPARα),
which is associated with liver tumours in rodents, or tumours induced by various hormonal
mechanisms). As with other non-genotoxic modes of action, these can all be presumed to have
a threshold.

The objective of investigating the carcinogenicity of chemicals is to identify potential human

carcinogens, their mode(s) of action and their potency.

Carcinogenic potential and potency information from human epidemiology studies, if available,
requires expert consideration because of uncertainties in the exposure assessment and/or
limited sensitivity and statistical power. Reliable human epidemiological studies (e.g. cohort
case–control studies) can be considered using a WoE approach. In the absence of human data,
animal carcinogenicity tests may be used to differentiate carcinogens from non-carcinogens.
However, the results of these studies subsequently have to be extrapolated to humans, in both
qualitative and quantitative terms. This introduces uncertainty with regard to potency and
relevance to humans due to species-specific factors such as differences in chemical metabolism
and toxicokinetics and difficulties inherent in extrapolating from the high doses used in animal
bioassays to those normally experienced by humans.

Once a chemical has been identified as a carcinogen, there is a need to elucidate the underlying
mode of action, that is, whether the chemical is directly genotoxic or not. In risk assessment a
distinction is made between the different types of carcinogens - genotoxic or non-genotoxic.

For genotoxic carcinogens exhibiting direct interaction with DNA, it is not generally possible to
infer the position of the threshold from the no observed effect level on a dose–response curve,
even though a biological threshold below which cancer is not induced may exist.

For non-genotoxic carcinogens, no-effect thresholds are assumed to exist and to be discernible
(e.g. if appropriately designed studies of the dose–response for critical non-genotoxic effects are
conducted). No-effect thresholds may also be present for certain carcinogens that cause genetic
alterations via indirect effects on DNA following interaction with other cellular processes (e.g.
carcinogenic risk would manifest only after chemically induced alterations of cellular processes
had exceeded the compensatory capacity of physiological or homeostatic controls). However, in
this situation the scientific evidence needed to convincingly underpin this indirect mode of
genotoxic action may be more difficult to achieve. Human studies are generally not available for
making a distinction between the abovementioned modes of action; a conclusion on this, in fact,
depends on the outcome of mutagenicity/genotoxicity testing and other mechanistic studies. In
addition to this, animal studies (e.g. the carcinogenicity study, repeated dose studies and
experimental studies with initiation–promotion protocols) may also inform on the underlying
mode of carcinogenic action.

3.5.2. Data to be used in the hazard assessment

3.5.2.1. Non-human data for carcinogenicity

3.5.2.1.1. Non-testing data for carcinogenicity

See Section 3.1.2.2.5 (on (Q)SARs) and Section 3.1.3 (on the read-across approach).
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3.5.2.1.2. Testing data for carcinogenicity

In vitro data. The following in vitro data, which provide direct or indirect information useful in
assessing the carcinogenic potential of a substance and (potentially) the underlying mode(s) of
action, may be available. No single end point or effect in and of itself possesses unusual
significance for assessing carcinogenic potential but must be evaluated within the context of the
overall toxicological effects of the substance under evaluation. Standardised protocols do not
exist for most of the in vitro end points noted; rather, studies are conducted in accordance with
expert judgement using protocols tailored to the specific substance, target tissue and cell type
or animal species under evaluation.

• Genotoxicity studies. The ability of substances to induce mutations or genotoxicity can be

indicative of carcinogenic potential. However, correlations between
mutagenicity/genotoxicity and carcinogenesis are stronger when effects are observed in
appropriately designed in vivo as opposed to in vitro studies.

• In vitro cell transformation assay results. Such assays assess the ability of chemicals to
induce changes in the morphological and growth properties of cultured mammalian cells that
are presumed to be similar to phenotypic changes that accompany the development of
neoplastic or pre-neoplastic lesions in vivo (OECD, 2006). Similarly to in vitro assays, cell
transformation assays are restricted to the detection of effects of chemicals at the cellular
level and will not be sensitive to carcinogenic activity mediated by effects exerted at the level
of intact tissues or organisms.

• Mechanistic studies. Data from these studies include the following.

o Cell proliferation. Sustained cell proliferation can facilitate the growth of neoplastic/pre-
neoplastic cells and/or create conditions conducive to spontaneous changes that promote
neoplastic development.

o Altered intercellular gap junction communication. Exchange of growth suppressive

or other small regulatory molecules between normal and neoplastic/pre-neoplastic cells
through gap junctions is suspected to suppress phenotypic expression of neoplastic
potential. Disruption of gap junction function, as assessed by a diverse array of assays
for fluorescent dye transfer or the exchange of small molecules between cells, may
attenuate the suppression of neoplastic potential by normal cells.

o Hormone or other receptor binding. A number of agents may act through binding to
hormone receptors or sites for regulatory substances that modulate the growth of cells
and/or control the expression of genes that facilitate the growth of neoplastic cells.
Interactions of this nature are diverse and generally very compound specific.

Other targeted mechanisms of action are as follows.

• Immunosuppressive activity. Neoplastic cells frequently have antigenic properties that

permit their detection and elimination by normal immune system function. Suppression of
normal immune function can reduce the effectiveness of this immune surveillance function
and permit the growth of neoplastic cells induced by exogenous factors or spontaneous
changes.

• Ability to inhibit or induce apoptosis. Apoptosis, or programmed cell death, is a sequence

of molecular events that results in the death of cells, most often by the release of specific
enzymes that result in the degradation of DNA in the cell nucleus. Apoptosis is integral to the
control of cell growth and differentiation in many tissues. Induction of apoptosis can eliminate
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cells that might otherwise suppress the growth of neoplastic cells; inhibition of apoptosis can
permit pre-neoplastic/neoplastic cells to escape regulatory controls that might otherwise
result in their elimination.

• Ability to stimulate angiogenesis or the secretion of angiogenesis factors. The

growth of pre-neoplastic/neoplastic cells in solid tumours will be constrained in the absence
of vascularisation to support the nutritional requirements of tumour growth. Secretion of
angiogenesis factors stimulates the vascularisation of solid tumour tissue and enables
continued tumour growth.

In vitro data can give only preliminary information about the carcinogenic potential of a
substance and possible underlying mode(s) of action. For example, in vitro genotoxicity studies
may provide information about whether or not the substance is likely to be genotoxic in vivo,
and thus a potential genotoxic carcinogen, and about the potential mode of action underlying
carcinogenicity – with or without a threshold.

Besides genotoxicity data, other in vitro data, such as in vitro cell transformation, can help to
decide, in a WoE evaluation, whether a substance possesses carcinogenic potential. Cell
transformation results in and of themselves do not inform about the actual underlying mode(s)
of action, since they are restricted to the detection of effects exerted at the level of the single
cell and may be produced by mechanistically distinct processes.

Studies can also be conducted to evaluate the ability of substances to influence processes
thought to facilitate carcinogenesis. Many of these end points are assessed by experimental
systems that have yet to be formally validated and/or are the products of continually evolving
basic research. Formalised and validated protocols are thus lacking for the conduct of these tests
and their interpretation. Although it is difficult to give general guidance on each test due to the
variety and evolving nature of tests available, it is important to consider them on a case-by-case
basis and carefully consider the context in which the test was conducted.

A number of the test end points evaluate mechanisms that may contribute to neoplastic
development, but the relative importance of each end point will vary as a function of the overall
toxicological profile of the substance being evaluated. It should also be noted that there are
uncertainties associated with extrapolating in vitro data to an in vivo situation, which requires
careful consideration of the in vitro to in vivo extrapolation model used. In many cases, in vitro
studies can offer valuable insights, which may be used in a WoE approach.

Animal data. Carcinogenicity in experimental animals can be evaluated using conventional

bioassays, bioassays that employ genetically modified animals and other in vivo bioassays that
focus on one or more of the critical stages of carcinogenesis. In the absence of data from
conventional long-term bioassays or assays with neoplasia as the end point, consistently positive
results in several models that address several stages in the multistage process of carcinogenesis
should be considered in evaluating the degree of evidence of carcinogenicity in experimental
animals.

Sources of information include the following.

• Carcinogenicity studies (conventional long-term or lifetime studies in experimental

animals). Carcinogenicity testing is typically conducted using rats and mice, less commonly
using animals such as the guinea pig and Syrian hamster and occasionally using minipigs,
dogs and primates. The standard rodent carcinogenicity bioassay would be conducted using
rats or mice randomly assigned to treatment groups. Exposures to test substances may be
via oral, inhalation or dermal exposure routes. The selection of exposure route is often
dictated by a priori assumptions regarding the routes of exposure relevant to humans and/or
other data sources (e.g. epidemiology studies or repeated dose toxicity studies in animals)
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that may indicate relevance of a given exposure route. In the context of the DWD, the
appropriate route is oral. Substances that have induced benign and malignant tumours in
well-performed experimental studies on animals are also considered to be presumed or
suspected human carcinogens unless there is strong evidence that the mechanism of tumour
formation is not relevant for humans (CLP regulation, Annex I, Section 3.6.1.1).
Standardised protocols for such studies have been developed and validated (e.g. OECD
TGs 451 and 453 and US EPA 870.4200).

• Short- and medium-term bioassay data (e.g. mouse skin tumour, rat liver foci
model, neonatal mouse model). Multiple assays have been developed that permit the
detection and quantitation of putative pre-neoplastic changes in specific tissues. The
induction of such pre-neoplastic foci may be indicative of carcinogenic potential. Such studies
are generally regarded as adjuncts to conventional cancer bioassays and, while less validated
and standardised, may be applicable on a case-by-case basis for obtaining supplemental
mechanistic and dose–response information that may be useful for risk assessment
(Enzmann et al., 1998).

• Genetically engineered (transgenic) rodent models (e.g. Xpa–/–, p53+/–, rasH2 or

Tg.AC). Animals can be genetically engineered such that one or more of the molecular
changes required for the multistep process of carcinogenesis has been accomplished. This
can increase the sensitivity of the animals to carcinogens and/or decrease the latency with
which spontaneous or induced tumours are observed. The genetic changes in a given strain
of engineered animal can increase sensitivity to carcinogenesis in a broad range of tissues or
can be specific to the change’s requisite for neoplastic development in one or a limited
number of tissues (ILSI-HESI, 2001; Pritchard et al., 2003; Jacobson-Kram, 2004). There
are, thus, limitations in the interpretation of data from such studies, although data from
these models may be used in a WoE analysis of a chemical’s carcinogenicity.

• Genotoxicity studies in vivo. The ability of substances to induce mutations or genotoxicity

can be indicative of carcinogenic potential. There is, in general, a good correlation between
positive genotoxicity findings in vivo and animal carcinogenicity bioassay results (Ashby and
Tennant, 1991; Ashby and Paton, 1993). A decision on carcinogenicity should not be made
from genotoxicity studies alone, but these studies may be used in a WoE analysis of a
chemical’s toxicological profile.

• Repeated dose toxicity tests. Such tests can identify tissues that may be specific targets
for toxicity and subsequent carcinogenic effects. Particular significance can be attached to
the observation of pre-neoplastic changes (e.g. hyperplasia or metaplasia) suspected to be
conducive to tumour development, and the observation of these pre-neoplastic changes
assist in the development of dose–effect relationships (Elcombe et al., 2002). However, it
should be noted that histopathological lesions in subchronic toxicity studies do not show a
perfect correlation with the carcinogenic potential of a substance (Elcombe et al., 2002).

• Studies on the induction of sustained cell proliferation. Substances can induce

sustained cell proliferation via compensatory processes that continuously regenerate tissues
damaged by toxicity. Some substances can also be tissue-specific mitogens, stimulating cell
proliferation in the absence of overt toxic effects. Mitogenic effects are often associated with
the action of tumour promoters. Both regenerative cell proliferation and mitogenic effects
can be necessary, but not sufficient, for tumour development, but have a sufficiently different
mechanistic basis that care should be exercised in assessing which is occurring (Cohen and
Ellwein, 1991; Cohen et al., 1991).
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• Studies on immunosuppressive activity. As noted earlier, suppression of normal immune

surveillance functions can interfere with normal immune system functions that serve to
identify and eliminate neoplastic cells.

• Studies on toxicokinetics. Such studies can identify tissues or treatment routes that may
be the targets for toxicity and can deliver data on exposure and metabolism in specific
organs. Linkages to subsequent carcinogenic impacts may or may not exist, but such data
can serve to focus carcinogenesis studies upon specific tissue types or animal species.

• Other studies on mechanisms/modes of action (e.g. OMICs studies:

toxicogenomics, proteomics, metabonomics and metabolomics). Carcinogenesis is
associated with multiple changes in gene expression, transcriptional regulation, protein
synthesis and other metabolic changes. Specific changes diagnostic of carcinogenic potential
have yet to be validated, but these rapidly advancing fields of study may one day permit
assessment of a broad array of molecular changes that might be useful in the identification
of potential carcinogens.

It is noted that the above tests inform differently on hazard identification, mode of action and
carcinogenic potency. For example, conventional bioassays are used for hazard identification and
potency estimation (i.e. derivation of a dose descriptor), whereas studies using genetically
engineered animals are informative on potential hazard and possibly mode of action, but less so
on carcinogenic potency, as they are considered to be highly sensitive to tumour induction.

In vivo data can give direct information about the carcinogenic potential of a substance, possible
underlying mode(s) of action and its potency.

Carcinogenicity testing should be addressed by an in vivo test in accordance with OECD TG 451
or 453, unless the substance is classified as mutagen category 1A or 1B, and a conclusion is
based on a comparison of the incidence, nature and time of occurrence of neoplasms in treated
animals and controls. Other tests may contribute to a WoE evaluation, for example by providing
supporting information or mechanistic data.

Knowledge of the historical tumour incidence for the strain of animal used is important
(laboratory-specific data are preferable). Attention to the study design is also essential because
of the requirement for statistical analyses. The quality, integrity and thoroughness of the
reported data from carcinogenicity studies are essential to the subsequent analysis and
evaluation of studies. A qualitative assessment of the acceptability of study reports is therefore
an important part of the process of independent evaluation. Sources of guidance in this respect
are IEH (2002), CCCF (2004) and OECD (2002). If the available study report does not include
all the information required by the standard test guideline, judgement is required to decide if
the experimental procedure is acceptable or not and if essential information is lacking.

The final design of a carcinogenicity bioassay may deviate from OECD guidelines if expert
judgement and experience in the testing of analogous substances supports the modification of
protocols. Such modifications to standard protocols can be considered a function of the specific
properties of the material under evaluation.

Carcinogenicity data may sometimes be available in animal species other than those specified in
standard test guidelines (e.g. guinea pig, Syrian hamster and occasionally minipigs, dogs and
primates). Such studies may be used in addition to, or instead of, studies in rats and mice, and
they should be considered in any evaluation.

Data from non-conventional carcinogenicity studies, such as short- and medium-term

carcinogenicity assays with neonatal or genetically engineered (transgenic) animals, may also
be available (OECD, 2002; CCCF, 2004). Genetically engineered animals possess mutations in
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genes that are believed to be altered in the multistep process of carcinogenesis, thereby
enhancing animal sensitivity to chemically induced tumours. A variety of transgenic animal
models exist, and new models are continually being developed. The genetic alteration(s) in a
specific animal model can be those suspected of facilitating neoplastic development in a wide
range of tissue types, or the alterations can be in genes suspected of being involved in tissue-
specific aspects of carcinogenesis. The genetic alterations must be applied while recognising both
their experimental nature and the specific mechanistic pathways they are designed to evaluate.
For example, a transgenic animal model sensitive to mesothelioma induction would be of limited
value in the study of a suspected liver carcinogen. While such animal model systems hold
promise for the detection of carcinogens in a shorter period of time and using fewer animals,
their sensitivity and specificity remains to be determined. Due to a relative lack of validation,
such assays have not yet been accepted as alternatives to the conventional lifetime
carcinogenicity studies, but existing data from these assays may be useful for screening
purposes or to determine the need for a rodent 2-year bioassay. Several evaluations of these
types of study have been published (e.g. ILSI-HESI, 2001; Pritchard et al., 2003; Jacobson-
Kram, 2004).

When data are available from more than one study of acceptable quality, consistency of the
findings should be established. When they are consistent, arriving at a conclusion is usually
straightforward, particularly if the studies were in more than one species or if there is a clear
treatment-related incidence of malignant tumours in a single study. If a single study only is
available and the test substance is not carcinogenic, scientific judgement is needed to decide on
whether (1) this study is relevant and (2) additional information is required to provide confidence
that it should not be considered as carcinogenic.

In addition, study findings may not clearly demonstrate a carcinogenic potential, even when
approved study guidelines have been followed. For example, there may be an increase in the
incidence of benign tumours or tumours that have a high background incidence only in control
animals. Although less convincing than an increase in malignant and rare tumours, and
recognising the potential oversensitivity of this model (Haseman, 1983; Ames and Gold, 1990),
a detailed and substantiated rationale should be given before such positive findings can be
dismissed as not relevant.

Repeated dose toxicity studies may provide helpful additional information to the WoE gathered
to determine whether a substance has the potential to induce cancer and the potential underlying
modes of action (Elcombe et al., 2002). For example, the induction of hyperplasia (through
cytotoxicity and regenerative cell proliferation, mitogenicity or interference with cellular control
mechanisms) and/or the induction of pre-neoplastic lesions may contribute to the WoE for
carcinogenic potential. However, it should be noted that histopathological lesions in subchronic
toxicity studies do not show a perfect correlation with the carcinogenic potential of a substance
(Elcombe et al., 2002). Toxicity studies may also provide evidence for immunosuppressive
activity, a condition favouring tumour development under conditions of chronic exposure.

Finally, TK data may reveal the generation of metabolites with relevant structural alerts. They
may also provide important information about the potency and relevance of carcinogenicity and
related data collected in one species and its extrapolation to another, based upon differences in
ADME of the substance. Species-specific differences mediated by such factors may be
demonstrated through experimental studies or by the application of TK modelling.

Positive carcinogenic findings on animals require careful evaluation, and this should be done with
reference to other toxicological data (e.g. in vitro and/or in vivo genotoxicity studies, TK data,
mechanistic studies, (Q)SAR evaluations) and the exposure conditions (e.g. route). Such
comparisons may provide evidence for a specific mechanism(s) of action, a significant factor to
consider whenever possible, which may then be evaluated with respect to relevance for humans.
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A conceptual framework that provides a structured and transparent approach to the WoE
assessment of the mode of action of carcinogens has been developed (see Sonich-Mullin et al.,
2001; Boobis et al., 2006). This framework should be followed when the mechanism of action is
key to the risk assessment being developed for a carcinogenic substance; it can be critical, in
particular, for determining whether a substance induces cancer via genotoxic or non-genotoxic
mechanisms.

For example, a substance may exhibit limited genotoxicity in vivo, but the relevance of this
property to carcinogenicity is uncertain if genotoxicity is not observed in tissues that are the
targets of carcinogenesis, or if genotoxicity is observed via routes not relevant to exposure
conditions (e.g. intravenous injection) but not when the substance is administered via routes of
administration known to induce cancer. In such instances, the apparent genotoxic properties of
the substance may not be related to the mechanism(s) believed to underlie tumour induction.
For example, oral administration of some inorganic metal compounds will induce renal tumours
via a mechanism believed to involve organ-specific toxicity and forced cell proliferation. Although
genotoxic responses can be induced in non-target tissues for carcinogenesis via intravenous
injection, there is limited evidence to suggest that this renal carcinogenesis entails a genotoxic
mechanism (IARC, 2006). The burden of proof in drawing such mechanistic inferences can be
high but can have a significant impact upon underlying assumptions made in risk assessment.

In general, tumours induced by a genotoxic mechanism (known or presumed) are, in the absence
of further information, considered to be of relevance to humans even when observed in tissues
with no direct human equivalent. Tumours shown to be induced by a non-genotoxic mechanism
are, in principle, also considered relevant to humans, but there is a recognition that some non-
genotoxic modes of action do not occur in humans (see OECD, 2002). This includes some specific
types of rodent kidney, thyroid, urinary bladder, forestomach and glandular stomach tumours
induced by rodent-specific modes of action, that is, by mechanisms/modes of action not
operating in humans or operative in humans under extreme and unrealistic conditions. Reviews
are available for some of these tumour types, providing a detailed characterisation that includes
the key biochemical and histopathological events that are needed to establish these rodent-
specific mechanisms that are not relevant for human health (IARC, 1999). The IPCS has
developed a framework and provided some examples of how to evaluate the relevance to
humans of a postulated mode of action in animals (Meek et al., 2003, 2014; Boobis et al.,
2006;).

The information available for substances identified as carcinogenic based on testing and/or non-
testing data should be further evaluated to identify underlying mode(s) of action and potency
and to subsequently enable a proper quantitative risk assessment. As already pointed out, the
use of existing data from non-standard animal models (e.g. transgenic or neonatal animals)
needs careful evaluation, using expert judgement, of how to apply the results obtained for hazard
and risk assessment purposes; it is not possible to provide guidance for such evaluations.

3.5.2.2. Human data

See Section 3.1.4.

3.5.3. Remaining uncertainty

As indicated in the previous sections, adequate human data for evaluating the carcinogenic
properties of a chemical are most often not available, and alternative approaches have to be
used.

In addition, test systems for identifying genotoxic carcinogens are reasonably well developed
and adequately cover this property. There is also agreement that animal carcinogens that act by
a genotoxic mode of action may be regarded as human carcinogens unless there is convincing
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evidence that the mechanisms by which mutagenicity and carcinogenicity are induced in animals
are not relevant to humans. The relationship between carcinogenic potency in animals and in
humans is unclear, however, and therefore introduces some uncertainty.

There is, on the other hand, a shortage of sensitive and selective test systems to identify non-
genotoxic carcinogens, apart from the carcinogenicity bioassay. In the absence of non-testing
information on the carcinogenicity of structurally related chemicals, indications of possible
carcinogenic properties may come from existing repeated dose toxicity data or in vitro cell
transformation assays. However, whereas existing repeated dose toxicity data will have a low
sensitivity (e.g. in the case of a 28-day study), for the in vitro cell transformation assay it is not
possible to quantify the reliability and uncertainty associated with these assays to predict
carcinogenic potential.

Non-genotoxic carcinogens may be difficult to identify in the absence of animal carcinogenicity

test data. However, risk assessment methodology should cover the risk for carcinogenic effects
via this mode of action as well; current risk assessments for many non-genotoxic carcinogens
are based on NOAELs for precursor effects or target organ toxicity, with the application of
conservative AFs to address uncertainty. Once identified as a non-genotoxic carcinogen (from
testing or non-testing data) there may be uncertainty about the relevance to humans of this
observation, that is, about the relevance to humans of the underlying mode of action. In the
absence of specific data on this, observations in the animal are taken as relevant to humans.
However, additional uncertainty will exist about the relationship between carcinogenic potency
in animals and in humans; this uncertainty, though, will be addressed in the procedure for
deriving human standards (Meek et al., 2003).

Finally, conventional assays of carcinogenicity in animals have been found to be insensitive for
some well-established human carcinogenic substances (e.g. asbestos and arsenic compounds).
These substances can be shown to be carcinogenic when the test conditions are modified, thus
illustrating that there will always be the possibility that a chemical poses a carcinogenic hazard
in humans that was not detected in conventional animal studies. This is also true for other
toxicological end points and should be considered in the risk assessment whenever applicable.

3.5.4. Concluding on suitability for risk assessment

Besides the identification of a chemical as a carcinogenic agent from either animal data or
epidemiological data, or both, dose–response assessment is an essential further step in
characterising carcinogenic risks for certain exposure conditions or scenarios. A critical element
of this assessment is the identification of the mode of action underlying the observed tumour
formation, whether this induction of a tumour is thought to be via a genotoxic mechanism or
not.

In regulatory work, it is generally assumed that, in the absence of data to the contrary, an effect
threshold cannot be identified for genotoxic carcinogens exhibiting direct interaction with DNA,
that is, it is not possible to define a no-effect level for carcinogenicity induced by such agents.
However, in certain cases, even for these compounds, a threshold for carcinogenicity may be
identified in the low-dose region. For example, it has in certain cases been clearly demonstrated
that an increase in tumours did not occur at exposures below those associated with local chronic
cytotoxicity and regenerative hyperplasia (Waddell, 2003). It is also recognised that for certain
genotoxic carcinogens causing genetic alterations, a practical threshold may exist for the
underlying genotoxic effect. For example, this has been shown to be the case for aneugens
(agents that induce aneuploidy – the gain or loss of entire chromosomes, resulting in changes
in chromosome number) and for chemicals that cause indirect effects on DNA that are secondary
to another effect (e.g. through oxidative stress that overwhelms natural antioxidant defence
mechanisms) (Elhajouji et al., 2011; Greim and Albertini, 2012). However, substance-specific
evidence needs to be provided in order to demonstrate that there is a threshold for genotoxicity.
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Non-genotoxic carcinogens exert their effects through mechanisms that do not involve direct
DNA reactivity. It is generally assumed that these modes of actions are associated with threshold
doses, and it may be possible to define no-effect levels for the underlying toxic effects of concern.
There are many different modes of action thought to be involved in non-genotoxic
carcinogenicity. Some appear to involve direct interaction with specific receptors (e.g. oestrogen
receptors), whereas others appear to be non-receptor mediated. Chronic cytotoxicity with
subsequent regenerative cell proliferation is considered a mode of action by which tumour
development can be induced. For example, the induction of urinary bladder tumours in rats may,
in certain cases, be due to persistent irritation/inflammation/erosion and regenerative
hyperplasia of the urothelium following the formation of bladder stones, which eventually results
in tumour formation. Specific cellular effects, such as inhibition of intercellular communication,
have also been proposed as facilitating the clonal growth of neoplastic/pre-neoplastic cells.

The identification of the mode of action of a carcinogen is based on a combination of results from
genotoxicity tests (both in vitro and in vivo) and observations in animal experiments, for example
site and type of tumour and parallel observations from pathological and microscopic analysis.
Reliable human epidemiological studies (e.g. cohort case–control studies) can be considered in
a WoE approach.

Once the mode of action of tumour formation is identified as having a threshold or not, the
following has to be considered:

• if the mode of action of tumour formation is defined as having a threshold, a dose descriptor
(DNEL) has to be derived to conclude the risk assessment;

• if the mode of action of tumour formation is identified as non-threshold, a limit value (MTCtap)
of 0.1 µg/l should be applied (see Sections 4.3 and 6.3.1).

3.6. Reproductive toxicity

The section on reproductive toxicity in DWD Guidance Volume I should be considered together
with the elements described in this section for the assessment of reproductive toxicity.

Section 1 of Annex VI to Commission Implementing Decision (EU) 2024/365 lists the criteria
when hazard assessment needs to be performed.

3.6.1. Definition

The term reproductive toxicity is used to describe the adverse effects induced (by a substance)
on sexual function and fertility in adult males and females, developmental toxicity in the offspring
and effects on or mediated via lactation, as defined in part 3 of the globally harmonised system
of classification and labelling of chemicals. In practical terms, reproductive toxicity is
characterised by multiple diverse end points that relate to impairment of male and female
reproductive functions or capacity (fertility) and the induction of non-heritable harmful effects
on the progeny (developmental toxicity). Effects on male or female fertility include adverse
effects on libido, sexual behaviour and any aspect of spermatogenesis or hormonal or
physiological response that interferes with the capacity to fertilise, fertilisation itself or the
development of the fertilised ovum up to and including implantation. Developmental toxicity
includes any effect interfering with normal development, both before and after birth. It includes
effects induced or manifested either pre- or postnatally. This includes embryotoxic/fetotoxic
effects such as reduced body weight, growth and developmental retardation, organ toxicity,
death, abortion, structural defects (teratogenic effects), functional effects, peri- and postnatal
defects, and impaired postnatal mental or physical development up to and including normal
pubertal development.
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The general objectives of the testing are to establish:

• whether exposure of humans to the relevant chemical species can be associated with adverse
effects on reproductive function or capacity;

• whether, in studies in animals, administration of the relevant chemical species to males

and/or females prior to conception and during pregnancy and lactation causes adverse effects
on reproductive function or capacity;

• whether exposure of the relevant chemical species during the period of pre- or postnatal
development induces non-heritable adverse effects in the progeny.

The dose–response relationship for any substance-related adverse effects on reproduction are
always of potential concern. It is important, however, to distinguish, where possible, between a
primary effect on the reproductive system (due to the intrinsic property of the substance) and a
secondary reproductive effect, usually a consequence of general toxicity (e.g. reduced food or
water intake, maternal stress). Hence, reproductive toxicity should be assessed alongside
parental toxicity in the same study. Further guidance on the assessment of developmental
toxicity in relation to maternal toxicity is presented below.

3.6.2. Data to be used in the effects assessment

3.6.2.1. Non-human data

3.6.2.1.1. Non-testing data

For reproductive toxicity, a grouping and category approach and WoE adaptation are currently the
best fit-for-purpose tools for non-animal approaches to adapt the (standard) information
requirements for reproductive toxicity. However, appropriate justification and documentation must
be provided. For more information, see Section 3.1.3 (on the read-across approach) and
Section 3.1.4 (on WoE).

There are a large number of potential targets/mechanisms associated with reproductive toxicity,
which, on the basis of current knowledge, cannot normally be adequately covered by a battery
of (Q)SAR models. For further details, see Section 3.1.2.2.5 (on (Q)SARs).

3.6.2.1.2. Testing data for reproductive toxicity

In vitro data. The design of alternatives to in vivo testing for reproductive toxicity is particularly
challenging in view of the complexity of the reproductive process and large number of potential
targets/mechanisms associated with this broad area of toxicity. In addition, many in vitro
approaches do not include elements of biotransformation, which may differ depending on the
organ (Coecke et al., 2006).

Information on the current developments in in vitro tests and methodology can be found on the
ECVAM website ( 61) and other international centres for validation of alternative methods. ECHA’s
website is also updated with new internationally accepted non-animal approaches ( 62). However,
regulatory acceptance of these studies and approaches to replace animal testing for reproductive
toxicity has not been achieved, as they currently do not provide equivalent information and thus
cannot be used alone for risk assessment. In spite of this, they may support category and read

(61) https://joint-research-centre.ec.europa.eu/index_en.
(62) https://www.echa.europa.eu/support/oecd-eu-test-guidelines.
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across justification and WoE adaptation. They may also provide important information on
mechanisms and modes of action, or preliminary screening information that can be used in
planning further testing.

Currently, there are only three officially adopted EU test methods or OECD test guidelines for in
vitro tests of relevance to modes of action for reproductive toxicity: two measuring
oestrogenicity (OECD TGs 455 and 457) and the other measuring steroidogenesis (EU B.57 /
OECD TG 456). Most assays under development and in the process of international validation
focus on agonist/antagonistic properties measured by binding and activating or blocking a steroid
(or a thyroid) hormone receptor.

Three in vitro embryotoxicity tests to predict developmental toxicity have been validated but
have not been accepted for regulatory use (Genschow et al., 2004; Piersma et al., 2004;
Spielmann et al., 2004, 2006). These three tests, the embryonic stem cell test, the limb bud
micromass culture test and the whole embryo culture test, showed high predictivity for certain
strongly embryotoxic chemicals. However, due to the nature of the methods and limitations in
their predictivity, they may be used only as supporting information along with other more reliable
data to predict developmental toxicity. The value of these validated methods could be increased
by incorporating molecular-based markers through the application of proteomic and
toxicogenomic approaches (Piersma, 2006; van Dartel et al., 2010). The embryonic stem cell
method may be combined with physiologically based biokinetic modelling to derive quantitative
points of departure in vitro, which are then extrapolated to in vivo points of departure for use in
risk assessment (Worth et al., 2014).

The combination of assays in a tiered and/or battery approach may improve predictivity, but the
in vivo situation remains more than the sum of the areas modelled by a series of in vitro assays
(see Piersma (2006) for a review). Therefore, a negative result predicting absence of a particular
property for a substance with no supporting information cannot be interpreted as demonstrating
the absence of a reproductive hazard with the same confidence as an animal study. Another
limitation of in vitro tests is that a N(L)OAEL and other dose–response information required for
a risk assessment is not provided.

However, a positive result predicting a particular reproductive hazard in a validated in vitro test
could provide justification for the need for further testing beyond the standard information
requirement, depending on the effective concentration and taking account of what is known
about the TK profile of the substance. However, because of limited confidence in this approach
at this time, such a result in isolation would not be adequate to support hazard classification.

In addition, validated and non-validated in vitro tests, provided the applicability domain is
appropriate, could be used with other data in a WoE approach to gather information on hazardous
properties. In vitro techniques can be used in mechanistic investigations, which can also provide
support for regulatory decisions.

In vitro tests can also be used as supporting evidence when assessing the toxicological properties
by read-across within a substance grouping approach, providing the applicability domain is
appropriate. Positive and negative in vitro test results may be of value in a read-across
assessment and in a category approach as supporting information.

Current developments in adverse outcome pathways to build a combination of studies and

investigations to cover key events, from an initiating molecular event to an adverse outcome,
may provide information on certain pathways, particularly developmental toxicity for certain
malformations. Approaches may combine different methods (e.g. in vitro tests, (Q)SARs, in
chemico assays). As these pathways do not cover all potential mechanisms/modes of action,
negative results predicting absence of a particular property from those approaches do not
provide enough confidence for regulatory decision-making to demonstrate absence of a
reproductive hazard. In addition, currently they do not provide a N(L)OAEL value or other dose–
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response information for risk assessment. However, they may provide necessary support for
read-across justification and categories and contribute to a WoE approach.

Animal data. Data may be available from a wide variety of animal studies that give different
amounts of direct or indirect information on the potential reproductive toxicity of a substance,
for example:

• screening studies (e.g. EU B.63 / OECD TG 421 or EU B.64 / OECD TG 422);

• one-, two- or multigeneration studies (e.g. EU B.56 / OECD TG 443, EU B.34 / OECD TG 415
or EU B.35 / OECD TG 416);

• prenatal developmental toxicity tests (e.g. EU B.31 / OECD TG 414);

• developmental neurotoxicity studies (e.g. EU B.53 / OECD TG 426);

• repeated dose toxicity studies (e.g. EU B.7 / OECD TG 407 or EU B.26 / OECD TG 408) if
relevant parameters are included, for example semen analysis, oestrous cyclicity and/or
reproductive organ histopathology;

• male or female fertility studies of non-standard design;

• mechanistic and TK studies.

Repeated dose toxicity studies

Although not aimed directly at investigating reproductive toxicity, repeated dose toxicity studies
(e.g. EU B.7 / OECD TG 407 or EU B.26 / OECD TG 408) may reveal clear effects on reproductive
organs in adult animals. However, if these findings occur in the presence of marked systemic
toxicity (up to the highest dose level tested in a repeated dose study), it may lower concerns for
effects on fertility and can contribute to decisions on further testing requirements. However, this
does not rule out the possibility that the substance may have the capacity to affect fertility.

The observation of effects on reproductive organs in repeated dose toxicity studies may also be
sufficient for identifying a N(L)OAEL for use in the risk assessment. It should, however, be noted
that the sensitivity of repeated dose toxicity studies for detecting effects on reproductive organs
may be less than that of reproductive toxicity studies (e.g. extended one-generation
reproductive toxicity studies) because of the lower number of animals per group. Some effects
seen in repeated dose toxicity studies may be difficult to interpret, for example changes in sex
hormone level, and should be investigated further as part of other studies. If there are indications
of adverse effects on the reproductive organs of adult animals in repeated dose toxicity studies,
further studies may be triggered (for details on the triggers, consult Section 6.9 of DWD
Guidance Volume I). Repeated dose toxicity studies may also provide indications on potential
developmental neurotoxicity and/or developmental immunotoxicity, which may need to be
further investigated.

In vivo reproductive toxicity tests

The available OECD test guidelines specifically designed to investigate reproductive toxicity are
shown in Table 10.

The purpose of the reproduction/developmental toxicity screening test (EU B.63 / OECD TG 421
and EU B.64 / OECD TG 422) is to provide information on the effects on male and female
reproductive performance in rodents, such as gonadal function, mating behaviour, conception,
development of conceptus and parturition. The observation in these tests of clear evidence of
adverse effects on reproduction or reproductive organs may be sufficient to meet the information
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needs for a classification and risk assessment (using an appropriate AF) and provide a N(L)OAEL
from which a DNEL can be identified. The results should be interpreted with caution, because
OECD TGs 421 and 422 are screening assays that were not designed as an alternative to or a
replacement for the definitive reproductive toxicity studies (EU B.34 / OECD TG 415, EU B.56 /
OECD TG 443 and EU B.35 / OECD TG 416). These screening tests are not meant to provide
complete information on all aspects of reproduction and development. In particular, the
postnatal effects associated with prenatal exposure (e.g. undetected malformations affecting
viability or functional effects) or effects resulting from postnatal or lactational exposure are not
covered in these studies. Furthermore, the exposure duration in these studies may not be
sufficient to detect all effects on the spermatogenic cycle, although it is likely that in practice the
2-week exposure period will be sufficient to detect the majority of testicular toxicants (Ulbrich
and Palmer, 1995). However, the number of animals per dose group is limited, which may affect
the statistical power of the study to detect an effect. These screening tests may in some cases
give indications of reproductive effects (e.g. fertility and postnatal effects) that cannot be
investigated in a prenatal developmental toxicity study (EU B.31 / OECD TG 414). A negative
result in a screening study may lower concerns for reproductive toxicity, but this will not provide
reassurance about the absence of this hazardous property. However, a negative result can
provide the basis for DNEL derivation in relation to reproductive toxicity derived from the highest
dose level used in the study and using an AF that takes account of the limitations of this study.
An evaluation of EU B.63 / OECD TG 421 and EU B.64 / OECD TG 422 has confirmed that these
tests are useful for initial hazard assessment and can contribute to decisions on further test
requirements (Reuter et al., 2003; Gelbke et al., 2004).

The two-generation study (EU B.35 / OECD TG 416) is a general test that enables evaluation of
the effects of the test substance on the complete reproductive cycle, including libido, fertility,
development of the conceptus, parturition, postnatal effects in both dams (lactation) and
offspring, and the reproductive capacity of the offspring.

The extended one-generation reproductive toxicity study (EU B.56 / OECD TG 443) addresses
the main limitation of EU B.34 / OECD TG 415 by incorporating additional postnatal evaluations,
which include clinical pathology, a functional observation battery, immunotoxicity end points,
oestrous cyclicity and semen analysis, and, using an extended F1 generation dosing period (to
postnatal day 70), end points addressing developmental neurotoxicity. The study has a
shortened F0 male premating dosing period, justified by the observation of no differences in the
detection rates for adverse effects on fertility between 4- and 9-week premating dosing periods
in a number of studies (reviewed by Ulbrich and Palmer, 1995).

The prenatal developmental toxicity study (EU B.31 / OECD TG 414) provides a focused
evaluation of potential effects on prenatal development, although only effects that are
manifested before birth can be detected.

Positive results in these studies will be relevant to hazard classification and the human health
risk assessment, unless there is information to show that effects seen in these studies are not
relevant for humans. N(L)OAELs can be identified from all studies discussed above.

Developmental neurotoxicity studies (EU B.53 / OECD TG 426 and EU B.56 / OECD TG 443) are
designed to provide information on the potential functional and morphological hazards to the
nervous system arising in the offspring from exposure of the mother during pregnancy and
lactation. These studies investigate changes in behaviour due to effects on the CNS and the
peripheral nervous system. As behaviour may also be affected by the function of other organs
such as the liver, kidneys and the endocrine system, toxic effects on these organs in offspring
may also be reflected in general changes in behaviour. No single test is able to reflect the entire
complex and intricate function of behaviour. For testing behaviour, therefore, a range of
parameters in a functional test battery, is used to identify changes in individual functions.
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The severity and nature of the effect should be considered. Generally, a pattern of effects (e.g.
impaired learning during several consecutive trials) is more persuasive evidence of
developmental neurotoxicity than one or a few unrelated changes. The reversibility of effects
should be considered, too. Irreversible effects are clearly serious, while reversible effects may
be of less concern. However, it is often not possible to determine whether an effect is truly
reversible. The nervous system possesses reserve capacity, which may compensate for damage,
but the resulting reduction in reserve capacity should be regarded as an adverse effect. If
developmental neurotoxicity is observed only during some of the lifespan, then compensation
should be suspected. In addition, effects observed, for example, during the beginning of a
learning task but not at the end should not be interpreted as reversible effects. Rather, the
results may indicate that the speed of learning is decreased.

The experience of offspring, particularly during infancy, may affect their later behaviour. For
example, frequent handling of rats during infancy may alter the physiological response to stress
and the behaviour in tests for emotionality and learning. In order to control for environmental
experiences, the conditions under which the offspring are reared should be standardised within
experiments with respect to variables such as noise level, handling and cage cleaning. The
performance of the animals during the behavioural testing may be influenced by, for example,
the time of day and the stress level of the animals. Therefore, the most reliable data are obtained
in studies where control and treated animals are tested alternatively and environmental
conditions are standardised.

Equivocal results may need to be followed up by further investigation. The most appropriate
method for further investigation should be determined on a case-by-case basis. Additional
guidance is provided in DWD Guidance Volume I and ECHA’s Guidance on Information
Requirements and Chemical Safety Assessment, Chapter R.8, and Guidance on the Biocidal
Products Regulation – Volume III: Human health – Part A: Information requirements ( 63).

Positive results in a developmental neurotoxicity study will be relevant to hazard classification

and the human health risk assessment, providing a N(L)OAEL, unless there is information to
show that effects seen in these studies are not relevant for humans.

For more detailed reviews of how to interpret the test guidelines mentioned in this section,
including discussions of their strengths and limitations, see the reports from ECETOC (2002) and
WHO (2001).

Table 10. Overview of in vivo OECD test guidelines for reproductive toxicity

Test Design End points

EU B.56 / OECD TG 443: Exposure of 10 weeks (unless specific Parental (P) generation:
extended one- reasons to shorten) prior to mating (P) and Growth, survival and fertility
generation study until postnatal day 90–120 (cohorts 1A and
Oestrus cyclicity and sperm
1B). If the extension of cohort 1B is
quality
triggered, then until postnatal day 4 or 21
(F2) Pregnancy length and litter
size
Three dose levels plus control; highest dose
level must be chosen with the aim of Histopathology and weight
inducing some systemic toxicity but not of reproductive and non-
death or severe suffering reproductive organs
Preferred species: rat Haematology and clinical
chemistry
Preferred route: oral
Offspring (F1):
N = sufficient mating pairs to produce 20

(63) ttps://echa.europa.eu/documents/10162/2324906/bpr_guidance_vol_iii_part_a_en.pdf/05e4944d-106e-9305-
21ba-f9a3a9845f93?t=1648525287369.
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Test Design End points

pregnant animals per dose group (P Growth, survival and sexual
generation) maturation
N = 20 mating pairs (extension of Histopathology and weight
cohort 1B, if triggered) of reproductive and non-
N = 10 males and 10 females per dose reproductive organs
group (cohorts 2A, 2B and 3, if triggered) (cohort 1A)
Weight of reproductive
organs and optional
histopathology (cohort 1B)
Haematology and clinical
chemistry
Fertility of F1 animals to
produce F2 generation
(extension of cohort 1B)
under certain conditions
Developmental neurotoxicity
(cohorts 2A and 2B or a
separate study) in case of a
particular concern
Developmental
immunotoxicity (cohort 3 or
a separate study) in case of
a particular concern
EU B.35 / OECD TG 416: Exposure before mating for at least Fertility
two-generation study 10 weeks until the end of the mating period Oestrus cyclicity and sperm
(males) until weaning of second generation quality
(females)
Pregnancy outcome (e.g.
At least three dose levels (unless limit test) dystocia)
plus control
Growth, development and
Sufficient number of animals to yield viability
preferably not less than 20 pregnant
Anogenital distance if
females per dose group
triggered
Preferred species: rat
Sexual maturation
Histopathology and weight
of reproductive organs,
brain and target organs
Recommended: motor
activity, sensory function
and reflex ontology in F1
generation
EU B.34 / OECD TG 415: Males: exposure before mating for at least Fertility
one-generation study one spermatogenic cycle until end of mating Growth, development and
period viability
Females: exposure before mating for at Histopathology and weight
least 2 weeks until weaning of first of reproductive organs,
generation brain and target organs
At least three dose levels (unless limit test)
plus control
Sufficient number of animals to yield about
20 pregnant females per dose group
Species: rat or mouse
EU B.31 / OECD TG 414: Exposure at least from implantation to 1– Clinical observations
prenatal developmental 2 days before expected birth Body weight and food/water
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Test Design End points

toxicity study At least three dose levels (unless limit test) consumption
(teratology study) plus control Implantation and
Sufficient number of females to result in resorptions
approximately 20 female animals with Fetal growth
implantation sites
Morphological variations and
Preferred rodent species: rat malformations
Preferred non-rodent species: rabbit
EU B.53 / OECD TG 426: Exposure at least from implantation and Birth and pregnancy length
developmental throughout lactation (postnatal day 20) Growth, development and
neurotoxicity study Three dose levels plus control viability
N = 20 pregnant females Physical and functional
maturation
Behavioural changes due to
central and peripheral
nervous system effects
Brain weights and
neuropathology
EU B.63 / OECD TG 421 Males: exposure before mating for at least Clinical observations
and EU B.64 / OECD 2 weeks until end of mating period Body weight and food/water
TG 422: reproduction/ Females: exposure before mating for at consumption
developmental toxicity least 2 weeks until at least postnatal Fertility
screening test day 13.
Pregnancy length and birth
At least three dose levels (unless limit test)
Thyroid hormones (T4)
plus control
Clinical chemistry
At least 10 males and 12–13 females per
group Gross necropsy
Organ weights
Histopathology
Fetal and pup growth and
survival until day 4
NB: T4, thyroid hormone thyroxine.

Developmental effects should be considered in relation to adverse effects occurring in the

parents. Since adverse effects in pregnancy or the postnatal period may result as a secondary
consequence of maternal toxicity, reduced food or water intake, maternal stress, lack of maternal
care, specific dietary deficiencies, poor animal husbandry, intercurrent infections, etc., it is
important that the effects observed should be interpreted in conjunction with possible
concomitant maternal toxicity (ECB, 2004; Fleeman et al., 2005; Cappon et al., 2005). The
nature, severity and dose–response of all effects observed in progeny and parental animals
should be considered and compared to achieve a balanced integrated assessment of available
data on all end points relevant for reproductive toxicity.

3.6.2.1.3. Relation between maternal toxicity and developmental toxicity

Developmental toxicity may be mediated by maternal toxicity, as experience has shown.

However, in most studies, evidence for a causal relationship is lacking. Developmental toxicity
occurring in the presence of maternal effects does not itself imply a causal relationship between
the two; therefore, it is not appropriate to discount developmental toxicity that occurs only in
the presence of maternal toxicity. Developmental effects that occur even in the presence of
maternal toxicity are considered to be evidence of developmental toxicity, unless it can be
unequivocally demonstrated on a case-by-case basis that the developmental effects are
secondary to maternal toxicity.
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For the assessment of the interrelationship between developmental toxicity and maternal
toxicity, please refer to Section 3.7.2.4 of Annex I to Regulation (EC) No 1272/2008 (the CLP
regulation) ( 64) and Guidance on the Application of the CLP Criteria ( 65).

3.6.2.1.4. Reproductive toxicity via lactation

Reproductive toxicity may occur through lactation in several ways. Substances may reach the
milk and result in exposure of the newborn. Alternatively, the quality and quantity of the milk
may be affected by maternal exposure to the substance, resulting in nutritional effects on the
newborn. Three aspects are crucial in the risk assessment of lactational effects, as indicated
below.

• The concentration of the substance transferred via the milk. TK aspects should be considered,
including the chemico-physical properties of the compound, the timing and duration of
exposure, the bioavailability and the persistence of the substance. Fat-soluble chemicals that
may be mobilised during lactation are of particular concern.

• The sensitivity of the newborn compared with the adult. A wide spectrum of toxic effects may
occur in the newborn, ranging from general toxic effects, which may present as reduced
weight gain or delayed general development, to specific effects on the maturation of organs
or physiological systems. The newborn may be more sensitive than the adult, not only
because of specific developmental end points, but also in view of a possibly higher intake of
the substance per kilogram of body weight and the immaturity of detoxification pathways
and physiological barriers. Moreover, some effects may become apparent only later in life.

• Effects on milk quality and/or quantity. These effects will usually be detected only through
effects on the growth and development of the newborn. Any effect on the quantity or quality
of the breast milk is likely to be due to systemic effects in the mother. However, overt
maternal toxicity may not be seen (e.g. the substance may affect only the transfer of a
nutrient into the milk, with no consequence for the mother). If a substance causes marked
overt systemic toxicity in the mother at the same dose level, then it is possible that this may
indirectly impair milk production or impair maternal care as a non-specific secondary effect.
The type and magnitude of the maternal effects and their potential influence on lactation /
milk production needs to be considered on a case-by-case basis using expert judgement. In
general, the two-generation study (EU Annex V B.35 or OECD TG 416) or the extended one-
generation study (OECD TG 443) are the best guideline-based studies available for
identifying effects on or via lactation. In the case of specific questions regarding lactation,
the protocol may have to be amended in view of any existing information on the substance
under study, including physico-chemical, TK and general toxic properties. Cross-
fostering ( 66) may establish whether toxicity to the offspring is the result of lactational effects
or uterine exposure.

(64) https://eur-lex.europa.eu/eli/reg/2008/1272/oj.
(65) https://echa.europa.eu/documents/10162/2324906/clp_overview_en.pdf/515227da-8a2e-c554-1320-
ee20e7cccb7c?t=1730719032697.
(66) Cross-fostering is a technique used in animal husbandry, animal science, genetic and nature versus nurture studies,
and conservation, whereby offspring between two predefined groups are usually exchanged after birth, thus being
removed from their biological parents and raised by surrogates.
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3.6.2.2. Human data

See Section 3.1.2.2.3.

3.6.3. Remaining uncertainty on reproductive toxicity

Unless the effect is a very specific one of low ‘normal’ incidence, there may be a high level of
uncertainty in human studies of effects on reproduction (see Section 3.1.2.2.3).

It is obvious that there are limitations in many of the types of non-human studies relating to
reproductive toxicity. Well-conducted tests following EU B.35/B.31 or OECD TG 416/414
standards can be used with confidence to identify whether substances are toxic to reproduction
or not in relation to the end points addressed in the test. However, other studies, including tests
conducted in accordance with EU B.63/B.64 or OECD TG 421/422, may provide clear (in the case
of the OECD methods) or indicative evidence of reproductive toxicity, but will not provide
sufficient evidence for confidence about the absence of reproductive toxicity. The WoE from other
studies (including human, TK and/or mechanistic data), when available, can help in reducing this
uncertainty.

3.6.4. Conclusions on reproductive toxicity

Reproductive toxicity end points should be considered collectively, using a WoE approach to
establish the most relevant end point and its NOAEL or critical effect dose to be used in risk
assessment.

A WoE assessment involves the consideration of all data that are available and may be relevant
to reproductive toxicity. There can be no firm rules for conducting a WoE assessment, as this
process involves expert judgement and because the mix and reliability of information available
for a particular substance will probably be unique. In addition, the WoE assessment should
consider all toxicity end points together, and not look at reproductive toxicity in isolation. Further
recommendations on how to perform WoE are given in the EFSA guidance on the use of the WoE
approach ( 67), and Chapter R.4 of the ECHA guidance ( 68) can be followed.

One example of a WoE assessment is the pooling of information from several available in vivo
reproductive toxicity studies. Individually, these studies may have deficiencies, such as brief
reporting, small group size, limited range of end points evaluated, the dose levels or the dosing
schedule not being appropriate for a comprehensive evaluation of potential effects on the
reproductive cycle, and the study not complying with GLP. However, taking account of their
reliability and relevance and consistency of findings, collectively these studies could provide a
level of information similar to that of the EU or OECD test guideline studies, and therefore meet
the information requirements needed for the classification decision and risk assessment.

3.6.5. Concluding on suitability for risk assessment

In order to be suitable for risk assessment, appropriate threshold levels (NO(A)EL/LO(A)EL,

BMD) have to be established. Reproductive effects due to endocrine disrupting mode of action
are regarded as non-threshold effects, and a limit value (MTCtap) of 0.1 µg/l should be applied
instead (see Sections 4.3 and 6.3.1).

(67) https://efsa.onlinelibrary.wiley.com/doi/epdf/10.2903/j.efsa.2017.4971.
(68) https://echa.europa.eu/documents/10162/17235/information_requirements_r4_en.pdf/d6395ad2-1596-4708-
ba86-0136686d205e?t=1323782558175.
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4. Hazard assessment: hazard characterisation (dose–

response / concentration-response relationship)
4.1. Introduction

For the derivation of a DNEL, all available hazard information regarding systemic toxicity and
local effects at the port of entry (e.g. GI tract) needs to be evaluated (see Chapter 3 of this
guidance document) and, where possible, dose descriptors (BMD, N(L)OAEL, etc.) need to be
established. The following is noted:

• a systemic effect is defined as an effect that is normally observed distant from the site of
first contact, that is, after having passed through a physiological barrier (mucous membrane
of the GI tract or of the respiratory tract, or the skin), and becomes systemically available;

• a local effect is an effect that is observed at the site of first contact (GI mucosa – in the
context of the DWD), caused irrespective of whether a substance is systemically available.

It should be noted, however, that toxic effects on surface epithelia may reflect indirect effects
as a consequence of systemic toxicity or secondary to systemic distribution of the substance or
its active metabolite(s). Figure 4 provides a schematic representation of the steps for performing
hazard characterisation.

Hazard characterisation involves the following steps.

• Identification of critical effects (outcome of the hazard identification).

• For effects where a threshold is identified:

o dose–response assessment –identification of most relevant dose descriptor (e.g.

N(L)OAEL/BMD);

o modification of relevant dose descriptor – application of appropriate AFs to the dose

descriptor to address uncertainties (see Section 4.2.2);

o derivation of systemic DNEL.

• For effects where a threshold is not identified (see also Chapter 6):

o if the relevant chemical species falls into the lowest migration tier (only genotoxicity data
provided) and it is non-genotoxic, then MTCtap = 2.5 µg/l;

o if the assessment identifies carcinogenicity and/or mutagenicity and/or ED hazard,

irrespective of the migration tier, and there is no robust scientific evidence of an existing
threshold, then an MTCtap of 0.1 µg/l must apply (see Section 6.2.3).
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Figure 4. Schematic representation of the steps for performing the hazard assessment

Identification of critical effects

• Consider both systemic and route-specific (oral) local effects (GI irritation)
• Establish most relevant animal species and study
• Consider dose–response relationships

Is the relevant
chemical species
genotoxic?

No

Is information Yes or data insufficient

available on
genotoxicity only
(due to low
Yes migration tier)?

No

Does the relevant

chemical species
show any other Yes
non-threshold
effect (Carc or ED)?

No

Dose response assessment

Relevant chemical
Identification of dose Identification of dose Carc or Muta or ED
species is not
descriptor for descriptor for local hazard concern
genotoxic at low
systemic effects effects (GI irritation) identified
migration band

See Section 6.3.1 of Select effect with Local effects at the See Sections 6.2.2
DWD Guidance relevant LOAEL; use site of entry (GI and 6.2.3 of DWD
Volume II respective NOAEL as tract). Guidance Volume II
starting point for Give as relevant
(MTCtap = 2.5 µg/l) DNEL external concentration (MTCtap = 0.1 µg/l)

Modification of dose descriptor

Application of appropriate AF

Derivation of Systemic DNELs

Establish MTCtap value

(Sections 6.3 and 6.4 of DWD Guidance Volume II)
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4.2. Identification of critical effects

In the first step of hazard assessment, the whole data package should be evaluated for
assessment of the most relevant critical effects, considering the biological plausibility of the
dose–effect relationship, its consistency over the whole data package, the severity and
reversibility of the effect, and the mode of action (if known) and its relevance for humans. For
the latter, the IPCS developed a framework for analysing the relevance of a non-cancer (Boobis
et al., 2008) or cancer mode of action for humans (Boobis et al., 2006). The IPCS framework on
mode of action / species concordance analysis was updated to take into account new
developments in the field of risk assessment (Meek et al., 2003). This framework gives the
opportunity to present in a transparent manner the evidence for the key events leading to an
adverse effect and to identify a causal linkage (through dose–response and time concordance).

Furthermore, the data package should be evaluated with respect to local effects at the port of
entry. If, instead of an oral study, an inhalation study is used to cover the repeated dose
information requirement for substances under the DWD, local effects, such as lesions in the
airways in inhalation studies or on the skin in dermal studies, should not be considered relevant
for derivation of a threshold for local toxicity. Only local effects in the GI tract need to be
considered.

Before deriving a DNEL on the basis of the dose descriptors, it is important to determine whether
the relevant chemical species exerts its effects by a non-threshold mode of action (e.g. non-
threshold mutagens, carcinogens, EDs) or whether a threshold is possible to derive (e.g.
systemic toxicity, reproductive toxicity).

If the substance exerts its effects by a threshold mode of action, a DNEL value must be derived
for the most critical effect(s) (see Section 4.2.2).

When the decision on a threshold and a non-threshold mode of action cannot be unambiguously
made, the non-threshold mode of action should be assumed.
4.2.1. Hazard information underlying the derivation of a derived no-effect
level

4.2.1.1. Toxicokinetics and absorption

Data on TK will provide information on the possible fate of the substance in the human body.
Sufficient information on absorption should be available to support route-to-route extrapolation
in the risk characterisation where it is needed or to address species-specific mechanisms if
relevant. Furthermore, absorption differences between species may be taken into consideration
for modification of the dose descriptor during the risk assessment (see Section 4.2.2.2).

4.2.1.2. Systemic (target organ) toxicity after repeated dose

administration

Information on systemic (target organ) toxicity after repeated dose administration can be
derived from short-term (28-day), subchronic (90-day) and chronic/carcinogenicity studies.

Most systemic (target organ) effects can be assessed using quantitative risk characterisation
and therefore depend upon the difference in dose levels at which adverse effects are seen in
animals (or humans) and the estimated exposure for the substance. The key factors are the
most sensitive, relevant NOAEL, the effects it is based upon and the dose response that occurs
at higher doses.
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4.2.1.3. Toxicity to reproduction and development

Effects on the reproductive system are often threshold based, allowing a quantitative risk
characterisation to be carried out. There could be, however, some exceptions, in particular when
non-threshold endocrine disrupting mode of action or genotoxic mechanisms underly the
reproductive effects or the effects on the development of offspring. In such cases, an MTCtap of
0.1 µg/l should be applied.

If the DNEL is based on severe reproductive effects, the need for an additional AF should be
considered. The AF will depend upon the severity of effects, their relationship to toxicity observed
in the mothers and the exposure level at which they occurred compared with effects seen in
other animals.

Fertility and developmental effects are relevant end points for exposure scenarios involving
repeated exposure to contaminants in drinking water.

4.2.1.4. Other toxicity end points

In addition to the abovementioned effects, other effects, such as neurotoxicity and

immunotoxicity, must be considered. They may be reversible or irreversible. In either case, the
acceptability of the effects will be reflected by the relevant DNEL accounting for oral, long-term
exposure.

4.2.2. Derived no-effect level derivation for threshold effects

A dose below which no adverse effect will occur exists for most types of toxicity, according to
current knowledge. For relevant chemical species that have such dose limit, a DNEL value should
be established. The most sensitive end point in the most relevant oral study should be used in
the derivation of a DNEL, as follows:

𝑁𝑁𝑁𝑁𝑁𝑁𝑁𝑁𝑁𝑁 𝑜𝑜𝑜𝑜 𝐿𝐿𝐿𝐿𝐿𝐿𝐿𝐿𝐿𝐿 𝑜𝑜𝑜𝑜 𝐵𝐵𝐵𝐵𝐵𝐵𝐵𝐵

𝐷𝐷𝐷𝐷𝐷𝐷𝐷𝐷 =
𝑜𝑜𝑜𝑜𝑜𝑜𝑜𝑜𝑜𝑜𝑜𝑜𝑜𝑜 𝐴𝐴𝐴𝐴
where:

• NOAEL = no observed adverse effect level;

• LOAEL = lowest observed adverse effect level;

• BMDL = lower confidence limit on the benchmark dose;

• AF = assessment factor.

The DNEL value is used further to calculate the maximum concentration at tap (MTCtap) (see
Section 4.1). Usually, the study in the most sensitive and relevant animal species resulting in
the most relevant lowest dose descriptor (e.g. BMDs, NOAELs, NOAECs, LOAELs, LOAECs) will
be selected for DNEL derivation. Often, several studies addressing a certain end point are
available for one relevant chemical species. Different dose spacing in these studies results in
different dose descriptors (e.g. BMDs, NOAELs, LOAELs). If study design and end points
addressed are comparable, it might be appropriate to consider these studies together. Special
attention should be paid to the relevance of effects, animal species and exposure duration when
choosing the dose descriptor, and the choice must be justified.
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As a general rule, if several relevant NOAELs (or other dose descriptors) are available, the one
that would result in the lowest DNEL should be chosen. However, the lowest dose descriptor may
not always provide the lowest DNEL value, as it depends on the AFs that will be used for its
derivation. Therefore, the choice of the critical dose descriptor should be made bearing in mind
that the resulting DNEL should also be the critical one for use in risk characterisation. The
duration of the study should also be considered when deciding on the critical dose descriptor. In
general, the experimental NOAEL will decrease with increasing exposure times and other, and
more serious, adverse effects may appear with increasing exposure times. Consequently, to end
up with the most conservative DNEL for repeated dose toxicity, chronic exposure is the ‘worst
case’. Therefore, if an appropriate chronic toxicity study is available, this is the preferred starting
point, and no AF for duration extrapolation is needed. If only a subacute or subchronic toxicity
study is available, appropriate AFs must be applied (see Section 4.2.2.3).

Alternatively, the assessor should derive all possible DNELs from the available relevant dose
descriptors for each end point and then choose the most critical DNEL to apply in the risk
assessment.

4.2.2.1. Relevant time frames in DNEL derivation

A DNEL for the general population’s long-term exposure to substances used in materials coming
into contact with drinking water can be derived.

Toxicity studies that give information on these possible long-term effects of a substance are
repeated dose toxicity studies, reproductive toxicity studies (including developmental toxicity
studies) and carcinogenicity studies. Here, ‘long-term’ is used as a more general term, including
subchronic (usually 90 days) as well as chronic (usually 1.5–2 years) studies.

When valid developmental toxicity studies are available, all relevant critical effects should be
evaluated together with other observations from other studies. If the dose descriptor (e.g.
NOAEL) derived from relevant effects in a valid developmental toxicity study is lower than that
from a short-term repeated dose toxicity study and this cannot be explained by dose spacing,
the dose descriptor (e.g. NOAEL) from the developmental toxicity study should be used for the
derivation of the DNEL value. This will apply to the general population (thus protecting both
pregnant and non-pregnant women). Developmental studies are often the only studies that use
gavage dosing with the aim of determining a dose descriptor (e.g. NOAEL). Maternal effects can
be regarded as critical effects for deriving a long-term DNEL if they are deemed relevant in
comparison with other critical effects observed in other valid repeated dose toxicity studies.

4.2.2.2. Application of appropriate assessment factors

Derivation of reference values such as a DNEL requires the selection of AFs. AFs are numerical
values. They are used to address the differences between the experimental data and the human
situation, considering the uncertainties in the extrapolation procedure and in the available
dataset. In principle, all data on a relevant chemical species need to be reviewed thoroughly in
order to use, as far as possible, substance-specific information for the establishment of
appropriate values for the various AFs. When substance-specific information is not available,
data on analogues, which act with the same mode of action as the relevant chemical species
under consideration, should be considered. However, when the available data do not allow the
derivation of substance-specific or analogue-specific AFs, default AFs should be applied. Although
it is very often necessary to rely upon them, default AFs represent a fallback position rather than
a starting point.
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The setting of the overall AF is a critical step, which considers, but is not limited to:

• interspecies differences in toxicokinetics;

• interspecies differences in toxicodynamics;

• human variability in toxicokinetics;

• human variability in toxicodynamics.

Interspecies differences. Data from animal studies are a typical starting point for risk
characterisations; thus, differences in sensitivity between experimental animals and humans
need to be addressed, with the default assumption that humans are more sensitive than
experimental animals. Where human data are used as the starting point for the risk
characterisation, no extrapolation and no AF is necessary for interspecies differences in
sensitivity.

Interspecies differences result from variation in the sensitivity of species due to differences in
toxicokinetics and toxicodynamics. Some of the TK differences can be explained by differences
in body size (and related differences in basal metabolic rate). Information on interspecies
differences may be gathered from the toxicological database of a substance or by using
physiologically based pharmacokinetic (PBPK) modelling. If no substance-specific data are
available, the standard procedure for threshold effects, used as a default, is to correct for
differences in metabolic rate (allometric scaling) and apply an additional factor of 2.5 for other
interspecies differences, that is, TK differences not related to metabolic rate (small part) and
toxicodynamic differences (larger part). In cases where substance-specific information shows
specific susceptibility differences between species that are not related to differences in basal
metabolic rate, the additional factor of 2.5 for ‘remaining differences’ should be modified
accordingly.

Allometric scaling extrapolates doses according to an overall assumption that equitoxic doses
(when expressed in mg/kg body weight per day) scale with body weight to the power of 0.75.
This results in different default allometric scaling factors for the different animal species when
compared with humans (see Table 11).

Table 11. Allometric scaling factors for different animal species when compared with humans
(assuming the human body weight is 70 kg)

Species Body weight (kg) Allometric scaling factor

Rat 0.250 4
Mouse 0.03 7
Hamster 0.11 5
Guinea pig 0.8 3
Rabbit 2 2.4
Dog 18 1.4

Allometric scaling is based on the assumption, originally predicted mathematically and

subsequently substantiated by empirical investigations (Schneider et al., 2004), that the effects
of toxicological relevance are driven by the basal metabolic rate, as it affects physiological
processes such as cardiac output, blood flow and perfusion of liver and kidneys, which, in turn,
affect the elimination/clearance of most chemicals.
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Allometric scaling is an empirical approach for interspecies extrapolation of a significant number

of kinetic processes related to toxicity. It is generally applicable to substances that are essentially
renally excreted, but not to compounds that are highly extracted by the liver and excreted in
the bile. It appears that species differences in biliary excretion and glucuronidation are
independent of caloric demand (Walton et al., 2001).

Allometric scaling according to caloric demand applies most appropriately to those substances
for which the not-metabolised parent or a stable metabolite is the relevant toxic species and
clearance is according to first-order processes. Conversely, the applicability of allometric scaling
is less well supported when toxicity is a consequence of exposure to a very reactive parent
compound (or metabolite) that is not removed from the site of formation.

Allometric scaling should not be applied if the effects are not dependent on metabolic rate or
systemic absorption, for example in the case of local effects. Generally, as long as route-to-route
extrapolation is not needed, allometric scaling should also not be applied in cases where doses
in experimental animal studies are expressed as concentrations (e.g. in mg/m3 in air, ppm in
diet, or mg/l in drinking water). The concentrations are assumed to be already scaled according
to the allometric principle, since ventilation rate and food intake directly depend on the basal
metabolic rate. However, once the concentration (e.g. ppm in diet) has been converted into a
dose (e.g. mg/kg/day), an allometric scaling factor has to be used. Thus, it is the dose unit
(original or transformed), and not the (experimental) route of application, that triggers the need
for a species-specific factor for allometric scaling.

For systemic effects in cases where allometric scaling is not applicable, AFs established on the
basis of substance-specific information should be well justified and used in a case-by case
manner.

4.2.2.3. Intraspecies differences

Humans differ in sensitivity to toxic insult due to a multitude of biological factors, such as genetic
polymorphism affecting, for example, toxicokinetics/metabolism, age, gender, health status and
nutritional status. These differences can be the result of genetic and/or environmental influences.
This intraspecies variation is greater in humans than in the more inbred experimental animal
population.

It is recognised that in order to always account for the most sensitive person exposed to any
chemical, a very large default AF would be required. That is of course not workable, and it is
usually assumed that a default AF of 10 is sufficient to protect the larger part of the population,
including children and the elderly. For threshold effects, this factor of 10 is the standard
procedure, used as a default, when assessing exposure to the general population. It is recognised
that there are differences between children and adults in toxicokinetics (especially babies in their
first months) and toxicodynamics (especially at different stages of development). These
differences may render children susceptible to the toxic effects of a substance. A higher
intraspecies AF for children (US EPA ( 2009) recommends from 10 up to 100 when assessing
pesticides in relation to food safety) should be considered when the following two criteria are
fulfilled:

• there are indications, obtained from, for example, experiments in adult animals,
epidemiological studies, in vitro experiments and/or SARs, of effects on organ systems and
functions that are particularly vulnerable under development and maturation in early life (in
particular the nervous, reproductive, endocrine and immune systems, and also the metabolic
pathways);

• there are deficiencies in the database on such effects in young animals.

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This line of reasoning and these criteria of course also apply to the unborn child, that is, to the
pregnant woman.

For additional information on allometric scaling and its use, please refer to ECHA’s Guidance on
Information Requirements and Chemical Safety Assessment – Chapter R.8: Characterisation of
dose [concentration]–response for human health, Section R.8.4.3.1 ( 69).

In addition to uncertainties in interspecies differences and intraspecies variability, additional AFs

for the following elements should be considered.

• The nature and severity of the effect.

• Deviations between the exposure in the study providing the NOAEL and the estimated human
exposure (e.g. lifetime for exposure to drinking water).

• Duration extrapolation. AFs for duration extrapolation should be handled on a case-by-case

basis, to use the best available data in derivation of reference values. It is specifically noted
that the possibility of duration extrapolation cannot be used in justifying a study waiver.

o subchronic to chronic: AF of 2.

o Subacute to chronic: an AF of 6.

• Dose–response relationship:

o extrapolation from LOAEL to NOAEL;

o the slope of the dose–response curve.

• The overall quality of the toxicity data package.

If the severity of the critical effect at the LOAEL (even if a NOAEL has been identified) is judged
to be of particular significance, an additional AF might be considered necessary. Based on expert
judgement, this AF has ranged from 2 to 10. Quantification should be determined on a case-by-
case basis taking into account the dose–response data.

If the derivation of the DNEL was based on a LOAEL and not a NOAEL, an additional AF has to
be considered. This factor will vary depending on the slope of the dose–response curve and the
magnitude of the effect at the LOAEL. This extrapolation step should be based on expert
judgement. The BMD concept can also be used when data allows and if it is deemed appropriate.
Guidance for using the BMD approach can be found in Guidance on the Biocidal Products
Regulation – Volume III human health – Assessment & evaluation (Parts B+C),
Section 2.3.2.1 ( 70). The use of LOAELs to set DNELs should be a last resort; however, where
the effects at the LOAEL are of moderate magnitude and not severe, the use of a LOAEL and an
appropriate AF reduces the need for additional animal studies.

To account for uncertainties related to data used to identify the accumulation potential,
additional AFs may need to be applied (see Section 6.5.1 of DWD Guidance Volume I). In

(69) https://echa.europa.eu/documents/10162/17224/information_requirements_r8_en.pdf/e153243a-03f0-44c5-
8808-88af66223258?t=1353935239897.
(70)
https://echa.europa.eu/documents/10162/2324906/biocides_guidance_human_health_ra_iii_part_bc_en.pdf/
30d53d7d-9723-7db4-357a-ca68739f5094?t=1512979002065.
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addition, when available, data from PBPK modelling can be used for the purpose of refining the
AFs. PBPK models will not remove all the uncertainty from the risk assessment process. The
rationale for using PBPK models in risk assessment is that they provide a documentable,
scientifically defensible means of bridging the gap between animal bioassays and human risk
estimates. Guidance on the use of PBPK modelling is available from the IPCS project on the
harmonization of approaches to the assessment of risk from exposure to chemicals, and should
be followed (IPCS, 2010).

4.2.3. Dose–response assessment

The quantitative extrapolation of hazard from animal experiments to exposed humans is based
on the most relevant end points. In most cases, these end points should correspond to relevant
dose descriptors (e.g. BMD, NOAEL, NOAEC, LOAEL, LOAEC). For more details on the
identification and modification of dose descriptors for systemic toxicity, please refer to ECHA’s
Guidance on Information Requirements and Chemical Safety Assessment – Chapter R.8:
Characterisation of dose [concentration]–response for human health, Section R.8.4.3.1 ( 71), and
EFSA’s 2022 guidance on the use of the benchmark dose approach in risk assessment ( 72).

4.3. Non-threshold effects

For substances migrating into the drinking water exerting genotoxic, carcinogenic or endocrine
disrupting effects, non-threshold mode of action is assumed, unless a threshold mode of action
is demonstrated. For non-threshold effects, a conservative MTCtap of 0.1 µg/l should be applied
(see Section 6.2.3).

(71) https://echa.europa.eu/documents/10162/17224/information_requirements_r8_en.pdf/e153243a-03f0-44c5-
8808-88af66223258?t=1353935239897.
(72) https://efsa.onlinelibrary.wiley.com/doi/epdf/10.2903/j.efsa.2022.7584.
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5. Exposure assessment
Exposure is a quantitative estimate of the dose/concentration of a substance that humans are
exposed to. Routes of exposure to humans are oral, inhalation and dermal. In the context of
water intended for human consumption, the route of (chronic) exposure, likely to cause the
highest exposure to humans, is the oral one (i.e. by direct ingestion of drinking water from the
tap or indirectly via beverages or food, etc.). In fact, compared with the oral route, other routes
of exposure via domestic water (e.g. dermal, when taking a shower or bath) are more acute in
nature and considered less important here. To estimate exposure via drinking water to relevant
chemical species originating from drinking water contact materials (DWCMs), accurate
quantification of the concentrations of the relevant chemical species migrating from and released
by DWCMs into drinking water is essential.

This section develops the concept of assessing risks related to substances migrated from
DWCMs. In practice, this means how to assess data from physical testing introduced in DWD
Guidance Volume I with the aim of establishing the Ctap. This value, which is strictly linked to the
exposure (dose) to which humans are potentially exposed, is then used to assess whether a
starting substance, composition or constituent fulfils the criteria of risk acceptance (see
Chapter 6).

5.1. Determination of concentration at the tap from physical testing

The value of Ctap compared with the limit concentration MTCtap must obtained for relevant
chemical species in each material type by physical testing in the worst foreseeable conditions of
use, as instructed in Chapter 5 of DWD Guidance Volume I. The steps are as follows.

• Step 1. Determine the material type. Determine whether the starting substance,
composition or constituent is used in (and thus belongs to) organic, metallic, cementitious
or enamel, ceramic and other inorganic material. Once the material type is determined, the
instructions in step 2 must be followed to determine the Ctap.

• Step 2. If the starting substance, composition or constituent is used in a bulk material, the
instructions below for that particular material type must be followed. If the starting
substance, composition or constituent is used in a surface layer (e.g. coating, plating or
lining), the instructions below for surface layers must be followed.

5.1.1. Organic materials

All testing, including preparation of the test sample and test procedures, must be conducted in
the worst foreseeable conditions of use, as instructed in DWD Guidance Volume I, Section 5.1.2.
All analyses must be conducted in accordance with the analytical quality criteria described in
DWD Guidance Volume I, Section 5.1.7.

The general procedure to determine the Ctap from the physical migration measurements is as
follows:

• calculate the migration rate, M, of a substance from its measured concentration(s), ci, in the
migration water of a specified migration period (3rd or 7th), as instructed in European
standard EN 12873-1;

• calculate the Ctap using the migration rate as determined above and a conversion factor using
(1) the default worst-case conversion factor of 20 (e.g. no limitation on use for a specific
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product type) or (2) a product-specific conversion factor ( 73) (< 20) if the applicant would
like to limit the approval to necessary use of certain product(s).

The migration periods chosen for the calculation of migration rate should be as follows.

• Cold water. The Ctap value is calculated from the 3rd migration period obtained from a
physical migration test in accordance with EN 12873-1 in cold chlorinated and non-
chlorinated water. Testing in two waters is needed to demonstrate the worst foreseeable
conditions because these may vary (see also Table 2, Section 5.1.2 in DWD Guidance
Volume I).

• Warm or hot water. The Ctap value is calculated from the 7th migration period obtained
from a physical migration test in accordance with EN 12873-1 in warm or hot non-chlorinated
water. Additionally, if the test piece containing the applied-for entry is a multilayer material
and the applied-for entry is not contained in the surface layer, an extended test in high
temperature is necessary, and the Ctap value is calculated from the 22nd migration period.

If testing in accordance with EN 12873-1 is not possible and this is demonstrated by a written
scientific justification, use of modelling can be considered. Ctap using modelling is calculated
similarly to that from physical testing by using the calculated results instead of physical test
results.

5.1.2. Metallic materials

All testing, including preparation of the test piece and testing procedures, must be conducted in
the worst foreseeable conditions of use, as instructed in DWD Guidance Volume I, Section 5.1.2.
All analyses must be conducted in accordance with the analytical quality criteria described in
DWD Guidance Volume I, Section 5.1.7.

In the case of metallic compositions, exposure assessment is necessary for compositions that
are tested in accordance with EN 15664-1 and are intended for use in product groups A, B or C.
Compositions that are tested and accepted in accordance with EN 16056 according to the criteria
for passive compositions or that are intended for use in product group D have negligible releases,
and thus insignificant exposure is assumed.

5.1.2.1. Determination of concentration at the tap values from physical

testing in accordance with EN 15664-1

The concentration value(s) at the tap for metallic compositions used to compare concentrations
in test water against the corresponding MTCtap value are determined by the following.

• The equivalent pipe concentration, c*EP,n(T,t), as defined in EN 15664-1. For pipe samples,
c*EP,n(T,t) equals the concentration of metals in water. For pipe pieces (i.e. test pieces in the
standard), it is formed by multiplying the concentration of metals in water by VS/VTP, which
are sample volume from the rig and volume of water in the test pieces. It corrects for the
decreased surface area from full pipe.

(73) The conversion factors typical for different product groups are reported in Table 5 of Annex I to Commission
Implementing Decision (EU) 2024/368.
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• The mean of the equivalent pipe concentrations, MEPn(T), as defined in EN 15664-1. MEPn(T)
is the arithmetic mean of the equivalent pipe concentrations of stagnation samples from test
line n according to the sampling plan in Annex B to EN 15664-1.

• MEPa(T) that is the average of the MEPn(T) of the three test lines.

• A multiplication factor ‘a’ according to the relevant product group (see Table 8, Annex II to
DWD Guidance Volume I).

• Ctap = MEPa(T) × a.

In the expression of MEPn(T) and MEPa(T) , T is the operation period in weeks.

If the metallic composition does not belong to an existing metallic composition category (see
Table 14, Annex IV to DWD Guidance Volume I and the ECHA website), the applicant must follow
the testing instructions in Section 5.1.4.3 of DWD Guidance Volume I. Otherwise, the applicant
must follow the instructions in Section 5.1.4.4 of DWD Guidance Volume I.

5.1.3. Cementitious materials

All testing, including preparation of the test piece and testing procedures, must be conducted in
the worst foreseeable conditions of use, as described in DWD Guidance Volume I, Section 5.1.2.
All analyses must be conducted in accordance with the analytical quality criteria described in
DWD Guidance Volume I, Section 5.1.7.

The general procedure to determine Ctap from the physical migration measurements is equivalent
to that for organic materials (see above). The migration periods chosen for the calculation of
migration should be as follows.

• Cold water. The Ctap value is calculated from the 3rd migration period obtained from a
physical migration test in accordance with EN 14944-3 in cold chlorinated or non-chlorinated
water.

• Warm or hot water. The Ctap value is calculated from the 7th migration period obtained
from a physical migration test in accordance with EN 14944-3 in warm or hot chlorinated or
non-chlorinated water.

5.1.4. Enamel, ceramic and other inorganic materials

All testing, including preparation of test piece and testing procedures, must be conducted in the
worst-known conditions of use, as described in DWD Guidance Volume I, Section 5.1.2. All
analyses must be conducted in accordance with the analytical quality criteria described in DWD
Guidance Volume I, Section 5.1.7.

The general procedure to determine the Ctap from the physical migration measurements is
equivalent to that for organic materials (see above). The migration periods chosen for the
calculation of migration rate should be as follows.

• Cold water. The Ctap value is calculated from the 3rd migration period obtained from a
physical migration test in accordance with EN 12873-1 in cold demineralised water.

• Warm or hot water. The Ctap value calculated from the 7th migration period obtained from
a physical migration test in accordance with EN 12873-1 in warm or hot demineralised water.
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5.1.5. Surface layers

Surface layers, used here to refer to coatings and platings, generally refer to any layer of material
applied on a bulk substrate that differs from the bulk in its properties. Coatings and platings
include organic, metallic, cementitious, and enamel, ceramic and other inorganic materials
applied by any existing method, for example spraying, brushing, hot-dipping and vacuum
deposition.

All testing, including preparation of the test sample and test procedures, must be conducted in
the worst foreseeable conditions of use, as described in DWD Guidance Volume I, Section 5.1.2,
depending on the coating or plating material in question. All analyses must be conducted in
accordance with the analytical quality criteria described in DWD Guidance Volume I,
Section 5.1.7.

The value of Ctap must be obtained for relevant chemical species in coatings and linings by
physical testing in the worst foreseeable conditions of use, as described earlier in this chapter.
The testing should follow the criteria for the group of materials to which the type of the coating
belongs. For metallic platings on metallic substrate materials, a long-term migration test in
accordance with EN 16058 or EN 15664-1 can be used where appropriate, taking into
consideration the shape of the test piece. If EN 16058 is used, three products should be tested.
Generally, the presentation of results must be as for EN 15664-1.

Deviations from bulk material testing include the following.

• The details of the substrate material must be reported.

• Migration from the substrate material, for example permeability, must be assessed. This may
be accomplished by, for example (1) a screening analysis and/or targeted analysis,
potentially followed by a targeted analysis of the migration water if the substrate material
does not contain substances similar to the surface layer or (2) a migration test with the
intended substrate and an additional migration test with a substrate that does not contain
any substances similar to the surface layer, followed by an analysis as in (1), and comparison
of the results to estimate potential migration through the surface layer from the substrate.
Other methods may be used and their suitability should be scientifically justified.

• Surface layer (coating, plating, lining) thickness, permeability and processing details must
be reported.

• For cementitious coatings (linings), a porosity test must be conducted in accordance with
EN 14944-3.

• For platings formed using a galvanic bath, there may be non-intended residual substances
from the plating bath incorporated into the plating. A migration test in accordance with
EN 12873-1 should be performed to demonstrate the absence of such substances. The
normal test procedures, including flushing and prewashing, apply, as the plated products
would normally undergo similar operations after plating. Certain platings may contain
residual substances or compounds incorporated during plating process that may be gradually
removed, for example via evaporation or other natural processes; thus, plating may take
some time after processing before reaching its final state. The intention of this test is to
ensure the absence of more stable residues that may represent a risk to the quality of water
intended for human consumption.
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5.2. Determination of concentration at the tap from modelling for

organic materials

Determination of the Ctap from modelling in accordance with CEN/TR 16364 follows, in principle,
the instructions laid out in Section 5.3.1 of DWD Guidance Volume I for starting substances, for
example determination of Ctap from simulated 3rd or 7th migration periods.

When calculating a simulated Ctap, reasonable worst-case values for parameters representing
the interaction between substance, material and water must be used, for example:

• diffusion coefficient;

• partition coefficient of the substance between organic material and water (KP,W);

• initial substance concentration in the final material.

The boundaries of the applicability domain of the modelling (e.g. applies only to organic material
where diffusion and partitioning can be estimated for starting substances), policy requirements
for its use (e.g. physical testing is not reliable due to analytical limit of quantification challenges)
and estimation of reasonable worst-case values of the parameters needed for the modelling are
reported in CEN/TR 16364 and summarised in Section 5.3.1 of DWD Guidance Volume I.
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6. Risk acceptance
6.1. Main principles

6.1.1. Approaches to maximum tolerable concentration at the tap adoption

for relevant chemical species

The starting substance, composition or constituent can be considered to pose an acceptable risk
if the Ctap is lower than the corresponding MTCtap for each of the relevant chemical species
associated with the specified use.

The estimation of MTCtap values for relevant chemical species may follow one of the following
approaches, depending on the situation.

• Limited approach. A limited acceptance approach based on Section 1 of Annex VI to

Commission Implementing Decision (EU) 2024/365 may be followed.

o Limited approach 1. A predetermined MTCtap value can be adopted on the basis of a

robust past assessment of the toxicological properties of the relevant chemical species.
No toxicological information should be generated for the purposes of the DWD application,
unless new or updated information has emerged since that past assessment. This
approach does not depend on the migration tier of the relevant chemical species
(Section 6.2.1).

o Limited approach 2. The identification of a hazard with a non-threshold mode of action

results in a default stringent MTCtap value of 0.1 μg/l, in some cases together with
additional conditions on the use of the applied-for entry (see Sections 6.2.2 and 6.2.3).

o Limited approach 3. The available information is insufficient to exclude genotoxicity,

thus resulting in a default stringent MTCtap value of 0.1 μg/l (see Section 6.2.4).

o Limited approach 4. The available information for the relevant chemical species is
sufficient to exclude genotoxicity but does not allow a conclusion on toxic threshold
effects, thus resulting in a default stringent MTCtap value of 2.5 μg/l (see Section 6.2.4).

• Comprehensive approach. A comprehensive acceptance methodology that starts from

toxicological information requirements linked to the migration tier of the relevant chemical
species. It considers all toxicological information available to adopt an MTCtap value, including
by calculation via a systemic DNEL value (see Section 6.3).

6.1.2. Concentration at the tap and migration tiers

As set down in Chapter 6 of DWD Guidance Volume I, there are three migration tiers with
associated toxicological data information requirements, which expand with the increased
migration of the substance from the DWCM:

• low migration tier – Ctap < 2.5 µg/l;

• medium migration tier – 2.5 µg/l ≤ Ctap < 250 µg/l;

• high migration tier – Ctap ≥ 250 µg/l.

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These migration tiers originate from the migration tiers used by EFSA in its guidance on preparing
an application for the safety assessment of a substance to be used in plastic food contact
materials (FCMs) ( 74), taking into account the requirements of Annex V to the DWD. Annex V to
the DWD prescribes that the MTCtap be based on a specific migration limit (SML) set in
Commission Regulation (EU) No 10/2011 by considering a 10 % ALF and water consumption of
2 l per day. Numerically, the division of the boundaries of the EFSA migration tiers by 20 has
been used to derive the DWD migration tier thresholds. The equivalent migration tiers for FCMs
and DWCMs is shown in Table 12.

Table 12. Migration tiers equivalents for FCM and DWCM assessments

Migration tier FCM migration tier DWCM migration tier

Low < 0.05 mg/kg food Ctap < 2.5 µg/l

Medium 0.05–5 mg/kg food 2.5 µg/l ≤ Ctap < 250 µg/l

High 5–60 mg/kg food Ctap ≥ 250 µg/l

It should be noted that the Ctap is determined as instructed, for example the 3rd migration period
for organic materials in cold water. For metals, the migration tier for toxicological analysis should
be taken as the mean of MEPa(T) for T = 16, 21 and 26 weeks for the relevant chemical species
in a rig test in accordance with EN 15664-1, resulting from the worst test water where possible.

6.2. Limited approach to the adoption of a maximum tolerable

concentration at the tap value

6.2.1. Limited approach 1: adoption of a maximum tolerable concentration

at the tap value on the basis of a robust past assessment

Section 1.3 of Annex VI to Commission Implementing Decision (EU) 2024/365 outlines cases
where an MTCtap value for a relevant chemical species can be generated from available data and
the applicant has no obligation to supply toxicological information, unless there are new or
updated data that need to be taken into consideration in setting an updated MTCtap value. Such
cases of limited acceptance include relevant chemical species for which the following is
applicable.

• A parametric value is set under Annex I to the DWD, in which case the MTCtap value is
calculated by application of an appropriate ALF to consider potential multiple routes of
exposure to the relevant chemical species, besides migration from products in contact with
drinking water. That MTCtap value is to be used for the purposes of risk acceptance. For
metals, those values will be available in Annex V to Commission Implementing Decision (EU)
2024/367; for certain substances listed in Annex I to the DWD, their MTCtap value will be
found in the corresponding entry of the relevant EUPL in the annexes to Commission
Implementing Decision (EU) 2024/367.

• An MTCtap value for the relevant material type is set under the corresponding annex to
Commission Implementing Decision (EU) 2024/367 following a decision by the European
Commission on a DWD application for a starting substance, composition or organic
cementitious constituent that has been previously submitted to ECHA. That MTCtap value may

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be used for the purposes of risk acceptance, unless information not included in the application
underlying the earlier European Commission decision materially impacts upon that MTCtap
value, thus requiring an updated, more appropriate value to be generated.

• An authorisation as an active substance has been granted under Regulation (EU)

No 528/2012 (the biocidal products regulation) on the basis of an agency opinion setting a
defined safety threshold for the oral route and used as such for material in contact with water
under product type 6. In such cases, that safety threshold is to be used for the purpose of
generating an MTCtap value in accordance with Section 4.2.2. Biocidal active substances may
be used with the intention to exert such activity in organic and cementitious materials.
However, only biocidal active substances of product type 6 (preservatives for products during
storage), in accordance with Regulation (EU) No 528/2012, may be used as starting
substances or organic cementitious constituents.

• An SML for the relevant chemical species that was set in Regulation (EU) No 10/2011 less
than 15 years in advance of the date of submission of the DWD application exists. In that
case, that SML divided by 20 and expressed in μg/l is to be used as the MTCtap for the
purposes of risk acceptance. Applicants are reminded that an SML is tied to a FCM migration
tier and the migration tiers for the same relevant chemical species migrating from FCM and
DWCM may differ. If the SML is based on an FCM migration tier that is lower than the
applicable tier for DWCM, the MTCtap value calculated from the existing SML value may be
lower than the established Ctap value in drinking water. Acceptable risk from the use of the
applied-for entry can thus not be demonstrated. In such cases, only through the provision of
toxicological information for the corresponding DWCM migration tier that allows the
derivation of a higher, more appropriate MTCtap value for DWD purposes can risks be shown
to be acceptable. In addition, any additional toxicological information that exists (e.g. a new
study that could not be considered when establishing an earlier SML) and can lead to a more
protective MTCtap value than what can be calculated from an existing SML must also be
considered by the applicant.

Table 13 summarises sources of existing MTCtap values that applicants may rely upon in their
applications. It is important to note that it is up to the applicants to ensure that any existing
MTCtap value (or parametric value) is based on the latest available toxicological information for
the relevant chemical species in question.

Table 13. Sources of existing MTCtap values for different material types

Material Table in Commission Description

type Implementing Decision
(EU) 2024/367
Organic Table 1 of Annex I For certain starting substances (e.g. acrylamide, benzene,
vinyl chloride), the table includes MTCtap values that have
been based on a parametric value shown in Annex I to the
DWD or an SML value, and for which no new or updated
data have become available since their last review.
The table will ultimately provide MTCtap values for starting
substances and other relevant chemical species that have
already been reviewed by RAC, and for which no new or
updated data have become available since their last
review.
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Material Table in Commission Description

type Implementing Decision
(EU) 2024/367
Table 1 of Annex V For certain metals relevant to metallic starting
substances, the table includes MTCtap values that are
presented in Annex V to the decision and based on a
parametric value in Annex I to the DWD or an SML value
from Annexes I and II to Regulation (EU) No 10/2011),
and for which no new or updated data have become
available since their last review.
Table 3 of Annex I The table provides total MTCtap values for relevant
chemical species that may arise from multiple starting
substances.
Table 4 of Annex I The table provides MTCtap values for degradation products
of starting substances with alkylphenol structural
moieties.
Metallic Table 1 of Annex V The table includes metallic elements for which an MTCtap
is available and for which no new or updated data have
become available since their last review.
Cementitious Table 3 of Annex III For certain organic cementitious constituents (e.g.
acrylamide, benzene, vinyl chloride), the table already
includes MTCtap values that have been based on a
parametric value in Annex I to the DWD, and for which no
new or updated data have become available since their
last review.
The table will ultimately provide MTCtap values for organic
cementitious constituents and other relevant chemical
species that have already been reviewed by ECHA’s
Committee for Risk Assessment, and for which no new or
updated data have become available since their last
review.
Table 1 of Annex V For certain metals relevant to organic cementitious
constituents, an MTCtap is available on the basis of a
parametric value in Annex I to the DWD or an SML value
from Annexes I and II of Regulation (EU) No 10/2011.
Enamel, etc. Table 1 of Annex V For certain elements, the table already includes MTCtap
values that have been based on the corresponding
existing parametric value shown in Annex I to the DWD,
and for which no new or updated data have become
available since their last review.
Table 1 of Annex V or For certain organic impurities (e.g. polycyclic aromatic
Table 1 of Annex I hydrocarbons), the table includes MTCtap values that are
based on a parametric value shown in Annex I to the
DWD, and for which no new or updated data have become
available since their last review.
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6.2.2. Limited approach 2: maximum tolerable concentration at the tap

value for relevant chemical species classified as carcinogenic,
mutagenic or toxic to reproduction category 1A or 1B / endocrine
disruptor human health category 1/PBT/vPvB/PMT/vPvM

Consumers should not be exposed to substances of very high concern, particularly those of very
high concern for human health, as a result of the use of starting substances, compositions and
constituents in the manufacture of DWCMs. In the context of DWCM, a relevant chemical species
that is a starting substance, an organic cementitious constituent, a substance constituent or a
non-intentionally added species is considered to be of very high concern if either of the following
apply:

• it has the hazard classification, whether harmonised or self-classification, of (1) carcinogenic,

mutagenic or toxic to reproduction (CMR) category 1A or 1B, (2) ED human health category
1, (3) persistent, bioaccumulative and toxic (PBT), (4) very persistent and very
bioaccumulative (vPvB), (5) persistent, mobile and toxic (PMT) or (6) very persistent and
very mobile (vPvM) established under Regulation (EU) No 1272/2008;

• it is identified as a substance of very high concern under the candidate list established under
Article 59 of Regulation (EC) No 1907/2006, except those identified on the basis of
Article 57(f) of Regulation (EC) No 1907/2006 only for the environment.

Starting substances, compositions or constituents that themselves fall under one of the above
classes or the use of which gives rise to migration of relevant chemical species that fall under
one of the above classes should not be used as far as possible in the manufacture of DWCM.

Nevertheless, if an application involves a relevant chemical species that falls under one or more
of the above classes, the approval of the applied-for entry into the EUPL is only possible if a very
stringent migration limit applies to the relevant chemical species. Moreover, the concentration
of the relevant chemical species in the final material should be restricted.

The rationale behind this approach is that:

• all consumers, including vulnerable groups, are regularly and continuously exposed to
substances migrating from DWCMs;

• products in contact with drinking water have a long-lasting article service life, with difficulty
in guaranteeing low migration from them after several years (due to ageing effects);

• there is analytical uncertainty around the very low migration limit potentially applicable to
the CMR/ED/PBT/vPvB substances;

• the approach is in line with chemicals strategy for sustainability, which shapes the future
landscape in chemical risk management in the EU; the strategy discusses a generic restriction
on CMR substances in several article types, including FCM;

• the use of CMR, ED and PBT/vPvB substances can also be largely avoided in all life-cycle
stages, including the production and waste stage (which are not covered by the DWD);

• PBT/vPvB/PMT/vPvM substances can always be considered relevant for posing risks to human
health due to the lifetime direct exposure of humans to such substances as a result of their
migration from DWCM, but also for the following additional reasons.
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o They may cause exposure via the environment, as, if present in drinking water, they will
eventually be released into the environment where they will persist, accumulate and
contaminate the soil, ground water, etc., leading to human exposure.

o Due to their (very) bioaccumulative, (very) persistent and/or (very) mobile properties,
accumulation of such substances in the environment is difficult to reverse, and the effects
of this accumulation are often difficult to predict in the long term. Therefore, it is
considered necessary to minimise emissions and exposure, since potential risks to human
health cannot be reliably assessed and excluded.

The reader may note that Section 6.5 of Annex I to Regulation (EC) No 1907/2006 (the REACH
regulation) requires that:

For substances satisfying the PBT and vPvB criteria the manufacturer or importer shall use the
information as obtained in Section 5, Step 2 when implementing on its site, and recommending for
downstream users, RMM [risk management measures] which minimise exposures and emissions to
humans and the environment, throughout the life-cycle of the substance that results from manufacture
or identified uses.

The applied-for entry may be accepted in the corresponding EUPL only if any of the relevant
chemical species that is of very high concern meets all the following criteria.

• It does not migrate at a level resulting in a Ctap value higher than 0.1 µg/l or higher than the
corresponding MTCtap calculated from a parametric value set under Annex I to the DWD by
application of an appropriate ALF.

• It is present in the final material at a concentration lower than 0.1 % weight per weight ratio
(w/w). If the available analytical method does not demonstrate unequivocally that the MTCtap
value has been met, the content of the substance in the final material must be lower than
0.02 % w/w; this lower concentration limit aims to ensure that migration is as low as
possible ( 75) ( 76).

• The relevant chemical species is only one of the following:

o a non-intentionally added species,

o a substance constituent,

o a monomer of a main polymer of the contact material.

For such relevant chemical species, applicants generally do not need to provide any toxicological
data. However, if information (e.g. a parametric value or toxicological data) exists and this
supports the derivation of an MTCtap value that is lower than 0.1 μg/l, such information must be
considered by the applicant. Migration data and information on their concentration in the final
material will always be required, however.

(75) ECHA has previously investigated the presence of CMR category 1A/1B substances in childcare articles.
Appendix A6.1 of the report provides details of analytical methods for CMR substances that may be of relevance
for measuring such substances in polymeric matrices. The investigation was published on 31 October 2023
(https://echa.europa.eu/completed-activities-on-restriction).
(76) There is evidence that a content lower than 0.1 % in the material does not guarantee a migration lower than
0.1 µg/l. See ‘Model calculations and specifications based on an acceptability of 0.1 µg/l of substances in tap water’,
unpublished meeting document of the subcommittee (‘W4’) to the Dutch committee of experts, document No W4
03-033, 2003, https://wetten.overheid.nl/BWBR0030279/2024-01-01.
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Meeting the requirement detailed in the second point above requires the estimation of a chemical
concentration in a solid matrix. For non-reactive relevant chemical species, a calculation-based
estimation may be sufficient. For example, for a non-reacting additive to an organic material, its
maximum concentration should be sufficient. For a non-reactive impurity of a starting substance
or of an organic cementitious constituent, the estimation can be based on a simple multiplication
of the maximum concentration of the impurity in the starting substance or organic cementitious
constituent by the maximum dosage of the starting substance or the cementitious constituent
in the formulation to produce the final material.

For relevant reactive chemical species (e.g. a monomer or other reactant or a reactive additive),
a measurement of its residual concentration in the solid matrix using an appropriate analytical
method may be required.

Important note. These requirements for classified relevant chemical species do not apply to
elements released into the migration water from constituents or impurities of metallic
compositions or from constituents of enamels, ceramic or other inorganic compositions. Those
only need to respect the corresponding MTCtap value shown in Annex V to Commission
Implementing Decision (EU) 2024/367. Where the MTCtap value is missing from that annex or is
not based on a parametric value in Annex I to the DWD or an SML established in Regulation (EU)
No 10/2011, the applicant must provide the toxicological information required under the
migration tier for the element established in physical migration tests and use that information
to generate and propose in their application a new MTCtap value.

6.2.3. Limited approach 3: maximum tolerable concentration at the tap

value for relevant chemical species that are potential carcinogens,
mutagens or endocrine disruptors with a non-threshold mode of
action

A relevant chemical species that is a starting substance, an organic cementitious constituent, a

substance constituent or a non-intentionally added species and is not classified as a carcinogen,
mutagen or ED human health category 1 according to Regulation (EC) No 1272/2008 (the CLP
regulation) but which, from the available toxicological data, is suspected to potentially exert
mutagenic, carcinogenic or endocrine disruptive effects for human heath via a non-threshold
mode of action, needs to comply with an MTCtap of 0.1 µg/l or an MTCtap value calculated from
an existing parametric value. This also applies to relevant chemical species that are classified
according to Regulation (EC) No 1272/2008 as mutagenic, carcinogenic or EDs for human health
category 2.

The use of substances with genotoxic properties should be avoided as much as possible.

6.2.4. Limited approach 4: maximum tolerable concentration at the tap

value for relevant chemical species for which available toxicological
information is insufficient

Section 1.4 of Annex VI to Commission Implementing Decision (EU) 2024/365 outlines additional
cases where a default MTCtap value for a relevant chemical species can be adopted in the light
of missing information, as follows.

• If the available information for the relevant chemical species is insufficient to exclude
genotoxicity, an MTCtap of 0.1 µg/l applies.

• If the available information for the relevant chemical species is sufficient to exclude
genotoxicity but does not allow a conclusion on toxic effects listed under Part 2.1.2 of
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Section 2 of Annex VI to Commission Implementing Decision (EU) 2024/365, an MTCtap of

2.5 µg/l applies. However, such a generic limit cannot be applied for genotoxicity,
carcinogenicity or endocrine disruption affecting human health with a non-threshold mode of
action.

More generally, for every relevant chemical species for which toxicological information is
required, provision of data on its genotoxicity should be considered the absolute minimum.

A summary of the scenarios described above for the different types of relevant chemical species
is provided in Table 14.
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Table 14. Summary of limited risk acceptance approaches

Relevant chemical species

cementitious

intentionally

Composition

Composition
constituent

constituent

constituent
Substance
substance
Scenario Risk acceptance criterion

impurity
Starting

Organic

added
Non-
Ctap < 0.1 µg/l (or < MTCtap from a parametric value)
Classified CMR category 1, ED human    
and
health category 1, PBT/vPvB, PMT/vPvM
Concentration in final material < 0.1 % w/w or
< 0.02 % w/w if analytical method is uncertain and
The relevant chemical species is any of the following:
• a non-intentionally added substance,
• a substance constituent,
• a monomer or other reactant of a main polymer in
the material.

Potential carcinogen, mutagen or ED Ctap < 0.1 µg/l (or < MTCtap from a parametric value)    
human health with identified non-
threshold mode of action based on
available toxicological data

Parametric value is available under Ctap < MTCtap from a parametric value      
Annex I to the DWD (and in Annex V to
Commission Implementing Decision
(EU) 2024/367)

MTCtap adopted by the European Ctap < existing MTCtap (unless new or updated      
Commission is available, following a toxicological information is available)
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Relevant chemical species

cementitious

intentionally

Composition

Composition
constituent

constituent

constituent
Substance
substance
Scenario Risk acceptance criterion

impurity
Starting

Organic

added
Non-
previous DWD application to ECHA

Safety threshold (oral) is available from Ctap < MTCtap generated from the existing safety value  
a biocidal active substance approval
Only biocidal active substances of product type 6 can be
under Regulation (EU) No 528/2012
used in DWCM

SML value is available in Regulation (EU) Ctap < SML ÷ 20 (unless new toxicological information      
No 10/2011 and is less than 15 years old requires a more appropriate MTCtap value or in certain
at the time of the DWD application to the cases where the FCM and DWCM migration tiers differ)
agency

The available information is insufficient Ctap < 0.1 µg/l      

to exclude genotoxicity

The available information is sufficient to Ctap < 2.5 µg/l      

exclude genotoxicity but does not allow
a conclusion on other toxic effects with a
threshold mode of action

The available information is sufficient to Ctap < 0.1 µg/l (or < MTCtap from a parametric value)      
exclude genotoxicity but raises concern
over toxic effects with a non-threshold
mode of action
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6.3. Comprehensive approach to the adoption of a maximum tolerable

concentration at the tap value on the basis of toxicological
information

6.3.1. Maximum tolerable concentration at the tap for a relevant chemical

species of low migration potential

If the relevant chemical species has a low migration potential (Ctap < 2.5 µg/l from migration
studies) and does not have a parametric value, the MTCtap will be calculated according to the
following rules.

• If genotoxicity studies or other available toxicological information indicate potential toxicity

exerted by non-threshold mode of action (e.g. endocrine disruption), an MTCtap value of
0.1 µg/l will apply. This limit also applies when toxicological data are not available, for
example for a non-intentionally added species for which it is not possible to set an appropriate
MTCtap (see Section 6.5).

• If the genotoxicity/mutagenicity studies show clearly negative results (see Section 3.4) and
no other toxicity data that show hazard that leads to an MTCtap lower than 2.5 µg/l is
available, the MTCtap is set to 2.5 µg/l.

In general, when the toxicological information provided by the applicant covers only the
information requirements for the low migration tier, the MTCtap cannot be higher than the upper
boundary of the tier, that is, it cannot be higher than 2.5 µg/l. However, there may be cases
where the applicant considers that the Ctap value is relatively close to the upper boundary of the
tier or cases where the Ctap is uncertain due to challenges with the analytical method used. Such
cases may raise concern that the products that are associated with the relevant chemical species
may fail their certification if the Ctap in the migration tests conducted for product certification
exceeds 2.5 µg/l. In such cases, the applicant may wish to consider assessing the standard
toxicological information of the next migration tier up in order to establish (by using
equation (1)) a DNEL leading to a more appropriate MTCtap value that is higher than 2.5 µg/l.
The applicant must justify the provision of toxicological information that corresponds to the
information requirements of the medium migration tier. Again, if this approach is taken, the
MTCtap proposed (and accepted into the EUPL) cannot be higher than the upper boundary of that
tier, that is, it cannot be higher than 250 µg/l.

The MTCtap can be lower than the upper boundary of the tier, even in the absence of a positive
genotoxicity study, if there is additional toxicological information that shows a toxicological effect
with a non-threshold mode of action (e.g. endocrine disruption) or if the additional information
supplies a DNEL that, according to equation (1), results in an MTCtap value lower than 2.5 µg/l.
The reader is reminded that under Part 1.5 of Section 2 of Annex V to Commission Implementing
Decision (EU) 2024/365, ‘any other relevant toxicological information that is available [beyond
the standard information for the relevant migration tier] shall be identified and considered’.

6.3.2. Maximum tolerable concentration at the tap for a relevant chemical

species of medium migration potential

If the relevant chemical species shows a medium migration potential (Ctap is in the range of 2.5–
250 µg/l), an MTCtap can be established from the toxicological data required under this migration
tier; in particular, it is possible to derive a dose at which no effect will occur, the DNEL, from
repeated dose toxicity data (see Chapter 4). The MTCtap is therefore calculated according to the
following equation:
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DNEL (mg/kg/d) × 60 (kg)

(1) MTCtap (mg/l) =
2 (l/d)
× ALF (77)

Essentially, the MTCtap is calculated assuming a dose below which no effect will occur when
ingested by an adult (with a body weight of 60 kg) who consumes 2 l of drinking water per day,
and by applying the so-called ALF (see Section 6.4). The supplementary ALF is introduced to
take into account that only a fraction of the dose of a substance may originate from the product
in contact with drinking water for which the MTCtap is calculated, as exposure to the same
substance may occur via other products or originate from different sources into the drinking
water.

In general, the MTCtap cannot be higher than the upper boundary of the tier; therefore, in this
case, it cannot be higher than 250 µg/l.

The MTCtap is to be set at 0.1 µg/l if there is toxicological information that shows a toxicological
effect with a non-threshold mode of action (e.g. genotoxicity or endocrine disruption).

6.3.3. Maximum tolerable concentration at the tap for a relevant chemical

species of high migration potential

If the relevant chemical species shows a high migration potential (Ctap > 250 µg/l), the MTCtap
will be calculated using equation (1). Again, the MTCtap is to be set at 0.1 µg/l if there is
toxicological information that shows a toxicological effect with a non-threshold mode of action
(e.g. genotoxicity or endocrine disruption).

For organic starting substances, there is another parameter that may play a role in setting the
MTCtap for the high migration tier: a limit based on total organic carbon (TOC). TOC is used for
product acceptance for organic materials in contact with drinking water; for organic materials,
the TOC as measured in the product migration tests cannot be higher than 0.5 mg/l. Therefore,
if the MTCtap of a starting substance is higher than the TOC limit (0.5 mg/l) ‘normalised’ to the
carbon content in the starting substance, then the applicant needs to indicate this in the
application. Consequently, the MTCtap value for the starting substance will be highlighted in the
EUPL to enable the certification process to take this into consideration.

In practice, this means that, for starting substances with a Ctap of > 250 µg/l, the applicant
should:

• calculate the MTCtap according to equation (1);

• compare the MTCtap with the TOC limit normalised to organic carbon (= 0.5 mg/l × molecular
weight substance / molecular weight of carbon in substance) to decide whether the MTCtap
value should be highlighted in the EUPL.

Note that the TOC is determined as non-purgeable organic carbon in water (in accordance with
EN 1484). The process of TOC determination typically involves removal of inorganic carbon and
conversion of organic carbon into CO2, which can be measured, and the amount of carbon then
determined. According to EN 1484, suitable methods for CO2 determination include infrared
spectrometry, titration, thermal conductivity, conductometry and coulometry. In addition to
organic carbon, water typically contains inorganic carbon such as CO2 or ions of carbonic acid,
which needs to be removed prior to TOC determination by purging with a suitable gas. If the
substance is purgeable (e.g. benzene, toluene), it will not contribute to the TOC, and therefore

(77) Note that the MTCtap in the equation is expressed in mg/l, while the migration tier thresholds are expressed in μg/l.
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the TOC criterion does not apply. Nonetheless, it is considered that the majority of TOC (~ 90 %)
is non-purgeable and methods such as EN 1484 are valid. In cases where purgeable organic
carbon analysis is necessary, methods exist but may have high uncertainties.

The MTCtap calculation according to different migration tiers is summarised in Table 15.

Table 15. MTCtap values from toxicological data according to different migration tiers under the
comprehensive approach

Relevant migration tier Equation for Conditions and MTCtap value

MTCtap

Low migration tier Not applicable Substance is not genotoxic  MTCtap = 2.5 µg/l
(Ctap < 2.5 µg/l)
Medium migration tier Equation (1) No concern for toxicity with a non-threshold mode of
(2.5 µg/l ≤ Ctap < 250 µg/l) action  MTCtap is calculated; MTCtap cannot be higher
than 250 μg/l

High migration tier Equation (1) No concern for toxicity with a non-threshold mode of
(Ctap ≥ 250 µg/l) action  MTCtap is calculated

If MTCtap from equation (1) exceeds 0.5 mg/l × MW /

MW_C  MTCtap is calculated and highlighted in
EUPL (a)

All migration tiers Not applicable Concern for toxicity with non-threshold mode of action
 MTCtap = 0.1 µg/l

(a) Criterion valid for starting substances only where TOC is determined to be non-purgeable organic carbon.
NB: MW, molecular weight of a substance; MW_C, molecular weight of carbon in a substance.

6.4. Selection of allocation factors

Annex V to Commission Implementing Decision (EU) 2024/367 presents MTCtap values for
several elements and ions for which a parametric value exists in Annex I to the DWD or an SML
appears in Regulation (EU) No 10/2011. Where a parametric value is used as the basis for the
MTCtap, it has been multiplied by an ALF (expressed as a percentage) to exclude other possible
sources of the substance in the drinking water from being taken into account. The principles
behind the selection of the ALF are presented below and can be considered where applicants
wish to generate MTCtap values that are missing from Annex V to Commission Implementing
Decision (EU) 2024/367.

• The standard ALF is equal to 10 % for starting substances and organic cementitious
constituents. It accepts a maximum contribution from organic and cementitious DWCMs
equal to 10 % of the maximum concentration at consumers’ taps (e.g. the parametric value
of a substance, where that exists in Annex I to the DWD). In a few cases, where it can be
demonstrated that the DWCM is the only source of the substance at the tap, the coefficient
can be set to 100 %. This applies to, for example, vinyl chloride, which is used as a monomer
for PVC polymer production, and no other source for it into drinking water is known).

• For elements released from metallic compositions, the standard ALF is equal to 50 % on the
assumption that other sources of contamination are possible. An exception to this assumption
is lead, where the only source of contamination is metallic contact materials (WHO, 2017).
In a few cases, an ALF of 90 % can be used where metallic products in contact with drinking
water constitute the only major source of contamination or where the toxicological
(parametric) value is substantially high (e.g. the case of copper or zinc), so that contributions
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from other sources into the drinking water would reliably not result in the exceedance of the
toxicological (parametric) value.

• For elements released from enamel, ceramic and other inorganic materials, the standard ALF
is also set to 10 %. When setting up the ALF, lower (e.g. currently 5 % for cadmium and
lead) or higher (e.g. currently 50 % for aluminium and cobalt) values may be considered.
The applicant should consider the toxicological concern and all sources from which the
element may end up in drinking water when proposing a value. Higher values could
potentially be more acceptable for low toxicity elements that do not have other sources. In
contrast, high toxicity elements that have multiple sources from which to enter drinking
water may be considered for lower values. Evaluation to determine a new ALF may be done
on a case-by-case basis.

Upon its adoption into legal text, Annex V to Commission Implementing Decision (EU) 2024/367
does not include MTCtap values that are based on toxicological values that are not reported in EU
legislation; as such, values originating from the work of the World Health Organization, national
competent authorities, etc., have not been considered. Applicants may review any such available
value and decide on and justify in their applications its use as a basis for generating a new
MTCtap. More importantly, applicants should justify their choice of ALF.

6.5. Risk acceptance methodology

6.5.1. Standard methodology

The aim of the risk assessment is to ensure that the DWCM will pose an acceptable risk to human
health or, in other words, that migration of relevant chemical species from the DWCM to water
will be lower than their corresponding maximum tolerable concentrations.

This means that for each relevant chemical species the Ctap as defined in accordance with
Chapter 5 needs to be lower than the MTCtap as derived in accordance with Sections 6.2 and 6.3.
The applicant is responsible for identifying specific conditions of use if required to lower the Ctap
below the MTCtap.

Special care should be taken by the applicant in those cases where the Ctap as calculated from
migration studies is very close to the MTCtap as derived according to the rules set in Sections 6.2
and 6.3., especially when the MTCtap is set to be equal to the upper boundary of the migration
tier (e.g. MTCtap = 2.5 µg/l for a relevant chemical species at the low migration tier). In such
cases, the applicant is advised to evaluate whether providing toxicity data relevant to the next
tier up (medium migration, 2.5–250 µg/l) would allow the setting up of a more appropriate, but
still protective, higher MTCtap.

6.5.2. Metallic compositions

Acceptance of a metallic composition is determined according to the following steps (see also
Annex I).

Step 1. Determination of whether (1) the composition falls within an existing metallic material
category or (2) the creation of a new category is necessary. In the case of (2), follow the
instructions for creating a new category in Section 5.1.4.3 of DWD Guidance Volume I.

Step 2. Determination of the product group. For intended use in product group A, B or C,
proceed to step 3. For intended use in product group D, no assessment is required, as the
potential exposure and thus risk to human health is considered negligible.
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Step 3. Determination of whether the metallic composition exhibits passivity in water. If yes,
then the instructions for compositions tested in accordance with EN 16056, that is, passive
compositions, in Section 6.5.2.2 should be followed. Otherwise, the instructions for metallic
compositions tested in accordance with EN 15664-1 in Section 6.5.2.1 should be followed. See
also step 4.

Step 4. If the metallic composition is used as a bulk material, follow the instructions for testing
and exposure assessment in Sections 5.1.4 and 5.2.2 of DWD Guidance Volume I and
Section 5.1.2 of this document, respectively. If the metallic composition is used in plating, that
is, in a surface layer, follow the instructions for testing and exposure assessment in
Sections 5.1.4 and 5.2.2 of DWD Guidance Volume I and for surface layers in Section 5.1.5 of
this document.

6.5.2.1. Metallic compositions tested according to EN 15664-1

Determine the relevant concentration values according to Section 5.1.2. Metallic materials tested
in accordance with EN 15664-1 may be accepted by testing.

In the acceptance of metallic compositions, the criteria considered are the absolute
concentrations of metals (criterion A) and the trends in concentration change (criterion B) as a
function of time.

For the assessment of the test rig results in accordance with EN 15664-1, the Ctap derived from
the arithmetic mean of the equivalent pipe concentrations MEPa(T), the ‘a’ factor (see also
Section 5.1.2), and the arithmetic mean (c*EP(T,4 hour)) of the 4-hour stagnation values
(c*EP,n(T,4h)) are to be considered.

The metallic composition can be accepted for a product group with the assumed contact surface
a (see Table 8, Annex II to DWD Guidance Volume I), if the following criteria are met for all
required test waters.

Criterion A

The applicable MTCtap values (also refer to Annex V to Commission Implementing Decision (EU)
2024/367) must be met for all analysed elements beginning from week 16. It is considered that
a pipe rig test has three periods, (1) conditioning, (2) transition and (3) stability, all of which
are associated with different metal migration behaviours (Eisnor and Gagnon, 2003). Risk
acceptance should be evaluated during the stability period, as during this time potential corrosion
product layers on the surface, etc., have formed and metal migration rates have stabilised. In
EN 15664-1, week 16 is considered as sufficient time for this to take place.

Criterion A is considered fulfilled if:

• Ctap (MEPa(T) × a) ≤ MTCtap for T = 16, 21 and 26 weeks or T = 16, 21, 39 and 52 weeks if
the test is extended to 52 weeks due to non-compliance in criterion B.

Criterion B

Metal concentrations (parameters) should not increase so that there is a risk of exceeding the
MTCtap beyond the duration of the test.
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Criterion B is considered fulfilled if one of the three criteria below is fulfilled:

• MEPa(Tb) ≥ MEPa(T) for {Tb, T} = {12, 16}, {16, 21} and {21, 26} {weeks};

• a negative slope of a linear fit of the c*EP(T,4-hour) for T > 12 weeks is obtained;

• c*EP(T,4 hour) × a ≤ 0.5 × MTCtap for T = {16–26 weeks}.

The test may be extended up to 1 year if criterion B is not met after 26 weeks.

If the test is extended to 1 year, criterion B is considered fulfilled if one of the three criteria
below is fulfilled:

• MEPa(Tb) ≥ MEPa(T) for {Tb, T} = {26, 39} and {39, 52} weeks;

• a negative slope of a linear fit of the c*EP(T,4 hour) for T > 26 weeks is obtained;

• c*EP(T,4 hour) × a ≤ 0.5 × MTCtap for T = {26–52} weeks.

For cases where all testing and water analyses were conducted properly, the assessment of
whether criteria A and B have been fulfilled may be quite straightforward. This could happen
when, for example, the Ctap values are well below the corresponding MTCtap values for all relevant
chemical species and the visual representations of the results as curves of concentration versus
time reveal a clearly decreasing trend. Nonetheless, when assessing acceptance, both criteria A
and B must be numerically met according to the conditions set above.

In other cases, compliance with these criteria can be difficult to establish clearly. This could be
due to closeness of measured Ctap values to MTCtap values or because of deviations (outliers) in
the test results due to uncertainties in analysis or variations in the test water composition. In
such cases, the complete set of available data must be considered. For the test rig, according to
EN 15664-1 these are:

• results of individual test lines;

• 4-hour stagnation results;

• parameters of water composition;

• temperature of the test rig;

• stagnation samples that were in addition to the requirements in EN 15664-1.

In addition, for these unclear cases it is advisable to use expert assessment in interpreting the
results.

6.5.2.2. Metallic compositions tested according to EN 16056 (passivating

alloys)

In cases where an alloy is claimed to be passive, forming a stable corrosion layer to prevent
general corrosion, this must be proven by testing the stability of the passive layer via an
electrochemical test in accordance with EN 16056 or equivalent.

Sufficient stability is displayed if the measured pitting potential, Epit, is higher than the free
corrosion potential + 500 mV.
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6.5.3. Surface layers (coatings, platings and linings)

Evaluation of surface layers (coatings, platings and linings) of organic, metallic, cementitious
materials or materials of inorganic origin are to be carried out as for bulk compositions.

Due to the nature of surface layers, they have additional requirements to bulk material testing,
which are described in Section 3.3 of DWD Guidance Volume I. In practice, a surface layer and
its properties should be described in a sufficient manner so that an evaluation of the migration
of substances contained in the layer and in the substrate can be made. For this purpose, the
properties that must be described may include layer thickness, adhesion, ductility, hardness,
porosity, hydrogen embrittlement, fatigue and processing details.
Guidance on the DWD: Volume II
Version 1.0 – January 2025 125

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1.
Guidance on the DWD: Volume II
Version 1.0 – January 2025 133

8. Annex I – Schematic of the metallics acceptance

process
Identification of the composition and uses: standard number,
constituents, impurities, chemical composition and processing history
(see Chapters 2 and 3 in DWD Guidance Volume I)

Does the alloy exhibit

passivation due to surface
Yes No
treatment or chemical
composition?
Describe whether passivity is Will the composition
due to surface treatment or be used only in
chemical composition of bulk. product group D (i.E.
Electrochemical test in negligible water
accordance with EN 16056 contact surface area)? Yes

No testing
No necessary.
Composition
accepted
Is Epit > free corrosion
potential +500 mV Is the composition
Yes
used as a plating?

Follow the additional

Yes No
No conditions for
platings; see Section
6.5.3
Does the composition
Composition belong to an existing
Composition
accepted into passive composition
rejected
category category?

Yes No

Addition of a composition to an Creation of a new category by

existing category by testing In testing as a reference In
accordance with EN 15664-1; see accordance with EN 15664-1; see
Section 5.1.4.3 in DWD Guidance Section 5.1.4.4 in DWD Guidance
Volume I Volume I

Do all the elements have

No an accepted existing
MTCtap?
Yes

Provide toxicological data

Does the composition
according to migration tier
satisfy all relevant
and physico-chemical
acceptance criteria in
property data, as for organics
(see Decision 2024/365)
Section 6.5?

Composition Composition
Yes No
accepted rejected
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