COMMENTS ON AN ANNEX XV DOSSIER FOR IDENTIFICATION OF A SUBSTANCE AS SVHC AND RESPONSES TO THESE COMMENTS

Substance name: Phenanthrene
CAS number: 85-01-8
EC number: 201-581-5

The substance is proposed to be identified as meeting the following SVHC criteria set out in Article 57 of the REACH Regulation: vPvB (Article 57e)

Disclaimer: Comments provided during public consultation are made available as submitted by the commenting parties. It was the commenting parties own responsibility to ensure that their comments do not contain confidential information. The Response to Comments table has been prepared by the competent authority of the Member State preparing the proposal for identification of a substance of very high concern. RCOM has not been agreed by the Member State Committee nor has the document been modified as result of the MSC discussions. The table does not contain any confidential information.

PART I: Comments and responses to comments on the SVHC proposal and its justification

General comments on the SVHC proposal
Number / Date	Submitted by (name, submitter type, country)	Comment	Responses	
5203
2018/10/17	Euroalliages,
Industry or trade association,
Belgium	See our position paper also submitted for Benzo[k]fluoranthene.	Thank you for your comments. 
We take note of this comment although it is not directly relevant for SVHC identification that is based on the intrinsic properties of the substance. In particular, the rationale provided to question the legal basis of the SVHC identification refers exclusively to authorisation and is therefore not relevant for SVHC identification. A distinct Annex XV dossier has been submitted to present the basis of the proposed SVHC identification for each of the PAHs in discussion. There is therefore no source of legal uncertainty.	
		Attachment: 5203_Position paper candidates listing of PAHs October 2018.pdf

5257
2018/10/19	Concawe,
Industry or trade association,
Belgium	-	Thank you for your comments.

Please see our response under comment 5257 in the next table (Specific comments).	
		Attachment: 5257_PHEN.zip

Specific comments on the justification
Number / Date	Submitted by (name, submitter type, country)	Comment	Responses	
5190
2018/10/11	Germany,
Member State	The DE MSCA supports the identification of Phenanthrene as a vPvB substance according to Article 57 (e) of REACH.	Thank you for your support.	
5197
2018/10/16	Belgium,
Member State	The Competent Authority of Belgium would like to thank the FR CA for submitting this Annex XV dossier. We recognize that the dossier submitter proposes to identify Phenanthrene as very persistent and very bioaccumulative (vPvB) according to Article 57 (e) of the REACH Regulation.
BE CA agrees with FR CA that an assessment of the vPvB properties of phenanthrene has already been carried out by the MSC in the context of the identification of CTPHT as SVHC, as documented in the MSC Support Document for identification of CTPHT as SVHC (ECHA, 2009), and that this information is still valid and allows a compact assessment of the substance properties with a focus on vPvB. Therefore, BE CA supports the identification of phenanthrene as very persistent and very bioaccumulative (vPvB) according to Article 57 (e) of the REACH Regulation.	Thank you for your support.	
5200
2018/10/16	Finland,
Member State	The Finnish CA supports the identification of Phenanthrene as a substance of very high concern, on the basis that it fulfils the criteria for very persistent and very bioaccumulative (vPvB) according to Article 57 (e) of Regulation (EC) No 1907/2006 (REACH).

As stated in the Annex XV report, Phenanthrene is identified as vPvB substance according to REACH Art. 57 (e) of REACH by comparing all relevant and available information according to Annex XIII of REACH with the criteria set out in the same Annex, in a weight-of-evidence determination.

We have the following specific comments on the Annex XV report:

3.1.2.1
Bahr et al. (2015) study.
Comment: In the Bahr et al. (2015) study, was there a control assay with non-pre-adapted inoculum? If yes, please add the result for the control. If not, then please include a sentence “There is no information on the effect of pre-adaptation on biodegradation in this study”. Please specify whether the four PAHs were studied as a mixture or as single substances.

“This study highlighted that phenanthrene degradation was slow in oxic aquifer and possibly persistent. Biodegradation was possible but needed the presence of an adapted bacterial community of sufficient density to achieve mineralisation over a long period of time (100 days of colonisation plus 62 days of exposure to reach 33% of mineralisation).”
Comment: Please replace the text above with “This study highlighted that phenanthrene degradation was slow in oxic aquifer, indicating possible persistence. Biodegradation was observed in the presence of an adapted bacterial community (100 days of colonisation plus 62 days of exposure to reach 33% of mineralisation).” In addition, please consider whether “colonisation” should be explained (probably a term related to this specific test method/technique?) or could it be replaced by “pre-exposure”?

3.1.2.2
“The latter condition (slurry) was not considered due to the adjunction of MBS who improved the bacterial growth, dispersion, accessibility and enhanced phenanthrene bioavailability.”
Comment: Please explain “was not considered”. Could it be replaced by “was not considered relevant for the present assessment”?

“The maximum extent of mineralisation of 14C-phenanthrene was 1.14% at 4°C for 35 days, indicating that, in natural temperature conditions of Livingstone Island, phenanthrene mineralisation is extremely slow and may even not occur.”
Comment: Please edit as: “The maximum extent of mineralisation of 14C-phenanthrene was 1.14% at 4°C for 35 days, indicating that, in natural temperature conditions of Livingstone Island, phenanthrene mineralisation is slow compared to higher temperatures.”

“At 12°C, maximum mineralisation extent was 35.15 % and 39.09% at 22°C indicating that phenanthrene degradation was limited, which was consistent with the plateau in degradation reached in soil 4 and 5”
Comment: Please modify as follows: “At 12°C, maximum mineralisation extent was 35.15 % and 39.09% at 22°C indicating that phenanthrene degradation was limited, which was consistent with the plateau in degradation rate reached in two of the studied soils (soil 4 and 5).”

“At 4 and 12 °C, degradation did not start after 35 days of incubation: it was reflected by lag phases (time to reach 5% mineralisation) that were not reached in the 5 soils at 4°C and 3 soils at 12°C and even 2 soils at 22°C.”
Comment: Please modify as follows: “At 4 and 12 °C, degradation did not start during 35 days of incubation which was reflected by lag phases (time to reach 5% mineralisation). 5% mineralization was not reached for the 5 soils at 4°C and 3 soils at 12°C and even 2 soils at 22°C.”. Please note that “after 35 days of incubation” seems to be incorrect because there does not seem to be any information available on degradation after 35 days.

“This study highlighted the fact that in soil with poor organic matter content, adsorption of phenanthrene will occur on mineral surfaces, reducing its bioavailability. When increasing temperature and exposure duration, the mineralisation of phenanthrene increases. Nevertheless, in natural conditions that can be found on Livingstone Island, no biodegradation of phenanthrene was observed, indicating that phenanthrene will remain over a long period of time in the environment. Moreover, when the contact time between phenanthrene and soils was increased, the maximum amounts of phenanthrene available for biodegradation was decreased. This was reflected by the plateau of 7% reached at 12°C after 150 days of pre-exposure indicating that soil could accumulate phenanthrene due to ageing (increase in contact time) (Okere et al., 2017)”
Comment: Please indicate whether this is the interpretation of the study authors or of the dossier submitter (also in 3.1.2.4).

“No biodegradation of phenanthrene measured as primary degradation was observed for 150 days, and this, even with the adjunction of fertilisers at day 86 of treatment, where the microcosms were fertilised with 7.45 g kg-1 dry soil of commercial fertiliser Nitrofoska® (BASF, Research Triangle Park, NC), taking the relation C/N/P to 100:5:2.”
Comment: Please clarify the sentence (can the words “and this” be deleted?).

Madueño, Alvarez, and Morelli study (2015).
Comment: Please indicate whether there was an abiotic control (considering that only primary degradation was measured) in the Madueño, Alvarez, and Morelli study (2015).

3.1.2.4
“indicating that phenanthrene was unlikely to biodegrade under arid conditions (Madueño, Alvarez, and Morelli, 2015)”
Comment: Please add the result obtained with increased water content or explain why it is not considered relevant? (Also relevant to 6.2.1.1)

”but never allowed reaching a sufficient mineralisation level.”
Comment: Please modify as follows: “but never allowed reaching a mineralisation level higher than x%”.

3.3.4
“Thus, the BCFs values were higher than 2000 and 5000.”
Comment: please modify as: “Thus, the BCFs values were higher than 2000 and in the majority of cases higher than 5000.”	Thank you for your support and specific comments. 

Specific comments are addressed below.

This information has been specified in the Support document.

It has been rephrased accordingly in the Support document.

It has been rephrased accordingly in the Support document.

It has been rephrased accordingly in the Support document.

It has been rephrased accordingly in the Support document.

It has been rephrased accordingly in the Support document.

The sentence has been clarified in the text of the Support document.

This information has been specified in the text of the Support document.

This information has been specified in the text of the Support document.

It has been rephrased accordingly in the Support document.

It has been rephrased accordingly in the Support document.

It has been rephrased accordingly in the Support document.	
		-		
5203
2018/10/17	Euroalliages,
Industry or trade association,
Belgium	See our position paper also submitted for Benzo[k]fluoranthene	No comment in this position paper relates to the justification of the proposed SVHC identification. Responses to general comments are provided in the corresponding table above.	
		See the corresponding embedded attachment in table 1 of Part I: 5203_Position paper candidates listing of PAHs October 2018.pdf		
5215
2018/10/19	Norway,
Member State	The Norwegian CA supports that phenanthrene should be identified as a substance of very high concern and should be included in the Candidate List.
Phenanthrene is a polyaromatic hydrocarbon (PAH), a group of chemicals where several similar structures are already identified as vPvB, as described in the MSC opinion on the SVHC properties on coal tar pitch, high temperature (CTPHT) in 2009. Phenanthrene meets the criteria for inclusion on the Candidate list as vPvB in 57 e) REACH.	Thank you for your support. 

	
5217
2018/10/19	cefic "Coal Chemicals Europe",
Industry or trade association,
Belgium	General observation
Industry notes that the conclusion on PBT/vPvB classification of PAHs in 2009 was part of the UVCB substance CTPHT evaluation. It was an agenda undisclosed to the public thus not following the classification procedures. No Annex XV dossier was presented for phenanthrene in 2009. Therefore, the vPvB classification of phenanthrene claimed in 2009 is null and void. The 2009 conclusion needs a thorough reconsideration and is not solely a formal confirmation process now restarted with the Annex XV dossier prepared by the French Competent Authority.

Bioaccumulation of phenanthrene
In the Foreword (p.5) and the Summary (p.6), the authors emphasise that the assessment of the vPvB properties in the present dossier and the conclusion that phenanthrene fulfils the criteria in Article 57 (e) were based mainly on the information in the MSC Support Document on CTPHT (ECHA, 2009) and supplemented with information from newer studies. All available information (such as the results of standard tests, modelling and (Q)SAR results) were considered together in a weight-of-evidence approach.
In this context, 3.3.1, 1st § (p.16) states: “When bioaccumulation data have been reassessed by the RIVM, these are used in preference to the CTPHT data.”

CCE notes that for BIOACCUMULATION…
1. on the contrary, new information was not adequately taken into account (information which in reality was not brand-new). This in particular concerns the outcome of the re-assessment by Bleeker and Verbruggen (RIVM 2009). Although Annex XV refers to this comprehensive analysis, it is obvious that this re-assessment was ignored, since it would have led to a different conclusion.
2. the dossier claims to follow a “Weight-of-Evidence” approach; however, the impression prevails that information deviating from the envisaged outcome is simply ignored. The rules laid down in the Guidance Chapter R.11, version 3.0 of June 2017 clearly state the primary role of fish data in BCF determination for PBT assessment (page 69 – 80), while results from invertebrates are taken as secondary sources and clearly play a subordinate role.
In this dossier results from invertebrates were given preference over valid fish data.
3. the well-founded classification proposal of Bleeker and Verbruggen (2009) is to classify phenanthrene as fulfilling the B- and not the vB criterion.

CCE does not see any scientific justification for the vB classification in view of valid BCF-data in fish.

In the summary part on p.6, Bioaccumulation, part of the justification for vB is as follows:
“The bioaccumulation of phenanthrene was measured in three studies with fish (BCFs ranging from 2229 to 6760 L/kg),…” (see also 3.3.4 Summary and Discussion).

This contains misleading information. There are three relevant fish studies:
Carlson et al. 1979 (newly introduced here, but not included in the MSC Support document of CTPht of 2009), Jonsson et al. 2004, and De Maagd 1996. The latter, previously a key study with RL2 in the Support Doc. 2009 and providing the highest BCF values among them 6760, had been downgraded to reliability 3 in the re-assessment by Bleeker and Verbruggen (2009).
The BCF range that remains relates to the first two considered as valid key studies: after lipid normalisation the range is 1149 – 4751.

The Annex XV dossier continues justifying their vB-decision by referring to several invertebrate studies with BCFs, all of them significantly above 5000.
Hence, the apparently plausible conclusion:
“Thus, BCFs greater than 2000 and 5000 were obtained….
Thus, the B and the vB criteria of REACH Annex XIII are fulfilled by phenanthrene.”

This statement is not accepted: Namely, with the two valid fish BCFs lower than 5000 (see above), this biased vB conclusion could not have been justified. Therefore, the highest BCF-value of 6760 from the RL3-study was not discarded, although scientifically founded.

CCE claims that the WoE procedure was not carried out properly due to the aim of reaching the envisaged vB classification.

3.3.1, 2nd § (p.16):
The relevant test was from Jonsson et al. 2004, sheephead minnows (Please, insert the missing source in the text!). The BCF of 2229 was before lipid normalisation, thereafter it was distinctly below 2000, namely 1149 (not mentioned by Bleeker and Verbruggen (2009) but by Verbruggen and Herwijnen 2011, Tab. 4.).
The “new” study could be referenced in the text for clarity: The study newly taken into account was an early but evaluable flow-through study within a comprehensive US EPA-screening programme and considered valid by Bleeker and Verbruggen (2009). Please reference the authors “Carlson et al 1979” in the text. The result was BCFK = 3611 L/kg in fathead minnows (Pimephales promelas), which resulted in BCF = 4751 after lipid-normalisation. This value was adopted as the “highest reliable value” by Bleeker and Verbruggen (2009, Tab.2).
The reference “De Maagd and colleagues (De Maagd et al., 1996)” should read “de Maagd 1996”, since it was a doctoral thesis. As this work was downgraded to RL3 by Bleeker and Verbruggen (2009), it was no longer accepted as key study.
Reference:
Verbruggen, E.M.J.; van Herwijnen, R. (2011): Environmental risk limits for phenanthrene. RIVM Letter report 601357007

3.3.1, 3rd § (p.16):
The statement is misleading, as Bleeker and Verbruggen (2009) had not supported this value of 5100 due to the considerable fluctuation of values: Instead of using the reported BCF values for day 28, they calculated kinetic BCF values based on the reported fish and water concentrations to minimise variability. This should be mentioned.

Moreover, the high depuration rate observed in both studies (Jonsson et al. 2004 and Carlson et al 1979) was ignored in the WoE. Either case indicate low bioaccumulation: k2 > 0.8/day, i.e. elimination of >90% within 3 days (see publications, Tab. 5 and 16, respectively), clearly supporting no to low bioconcentration.

As a matter of fact, Bleeker and Verbruggen (2009) re-analysed in total 10 reports covering some 12 fish species exposed to phenanthrene. Only two were classified Reliability (RL) 1 and 2, a total of 7 citations obtained RL 3 and one RL4. This was obviously not contradicted by the Annex XV dossier. The great majority of all studies produced BCFs distinctly below 2000.

3.3.1, 4th and 5th §§ (p.16/17):
The high number of BCF values for invertebrates collected in these sections, the majority of which with BCF values greater than 5000, cannot be taken as convincing argument for disqualifying valid fish bioconcentration studies. First, it is well-known that invertebrate species lack capabilities to metabolise PAHs to a great extent, and thus “bioconcentration” can easily be proven by using this killing argument, while a scientific discussion would become futile on this simplistic basis, and secondly, the fish on the higher trophic level is acknowledged to be the representative species for estimating bioaccumulating/bioconcentrating potential in the aquatic environment and has to be given preference over invertebrate data in PBT assessment (again ECHA guidance Chapter R.11).
This clear principle is turned to the opposite in the conclusion giving a high preference to invertebrates (3.3.4 (p.18).

CCE fully supports the conclusion of Bleeker and Verbruggen (2009) that - based on weight of evidence - the BCF values derived from fish support the qualification “bioaccumulative(B)” rather than “very bioaccumulative(vB)” for the purpose of PBT-classification. Given the high depuration rates observed in fish species, it is clearly indicative of B-classification to be sufficiently conservative.	Thank you for providing your 
comments

Specific comments are addressed below.

Response to General statement on PBT assessment

The agreement of the MSC on the identification of CTPHT as SVHC (ECHA, 2009), adopted on the 2nd December 2009, concludes on page 4: “Based on the data available, it is concluded that seven of the 12 PAH-constituents present in CTPHT in concentrations equal to or above 0.1% are to be considered as both vPvB and PBT substances. These are: fluoranthene, pyrene, benz(a)anthracene, chrysene, benzo(a)pyrene, benzo(k)fluoranthene and benzo(ghi)perylene. 
Phenanthrene fulfils the vPvB criteria, but not the PBT criteria.” The SVHC dossier submitter correctly reflected this fact and added more recent data.

Response to Bioaccumulation

For paragraph 2 

The lipid normalisation for sheephead minnows has been added in the text.

For paragraph 3 

The study by Bleeker and Verbruggen provides the same reliability score to static BCF than the calculated kinetic one, which was not inconsistent with the data set analysed. Moreover, when lipid normalization was applied to this static BCF, it provides a BCF value of 6118 (PBT factsheet on three ring PAHs bioaccumulation, PBT expert group of October 2018, NL, 2018). This data has been incorporated in the support document, summary has been adapted with this new evidence and this is in line with the previous conclusion.
Thus, Bleeker and Verbruggen (2009) concluded in the page 27 of their report that phenanthrene do accumulate in fish to a lesser extent than in invertebrates, due to biotransformation, but is very bioaccumulative at lower trophic levels (especially in molluscs)…  “For phenanthrene and fluoranthene, however, re-evaluation of data deemed these compounds B instead of vB when based on fish. Based on invertebrates (crustaceans for phenanthrene; mussels and crustaceans for fluoranthene) both compounds are considered vB.” Moreover, the authors also stated that “The results from De Maagd (1996) […] are still important as circumstantial evidence in a weight of evidence approach to decide on the bioaccumulation potential of the studies PAHs.”
So, the conclusion of this SVHC dossier are in line with the CTPHT assessment and in line with the conclusion raised by Bleeker and Verbruggen (2009) after their reevaluation of data.

Paragraphs 4 and 5

Cefic disagrees with the approach to conclude on the B and vB properties of phenanthrene based on data for lower trophic levels. However, the most sensitive organisms towards PAHs are those lacking a cytochrome P450 enzyme system. It has been shown without a doubt that bioaccumulation occurs in those organisms likely because they cannot degrade PAHs. This has been clearly assessed in the SVHC proposal. 
Furthermore,  the Guidance Chapter R11 version 3.0 of June 2017 state the following
(page 69): “Also use of other taxonomic groups than fish (e.g. mussel bioconcentration test, ASTM, 2003) is possible for measuring bioconcentration in the aquatic environment and the valid BCFs determined in other taxonomic groups can be used in assessing whether or not the B/vB criteria are met.” Therefore it is completely valid to use the most reliable data with invertebrate to conclude in the B/vB properties of phenanthrene.

	
		-		
5221
2018/10/19	EEB,
International NGO,
Belgium	EEB supports the proposal to identify Phenanthrene as SVHC based on its vPvB properties. MSC agreed already before on the vPvB properties of Phenanthrene in the context of the discussion on Coal Tar Pitch High Temperature containing Phenanthrene. Results from newer studies confirm the vPvB properties.	Thank you for your support. 	
5228
2018/10/19	Netherlands,
Member State	1. General Comments
NL supports the identification of phenanthrene as vPvB, which is based on the results presented in the European Union Risk Assessment Report on CTPHT (European Commission, 2008) performed by NL. The dossier also considered the evaluation of the bioaccumulation of polycyclic aromatic hydrocarbons in aquatic organisms by Bleeker & Verbruggen (2009). However, an update of the bioaccumulation data, including new data and reliability scoring, was presented in in Verbruggen and van Herwijnen (2011). The additional toxicity data included in the SVHC dossier were however not published in Verbruggen and van Herwijnen (2011) but in the updated assessment from RIVM for PAHs (Verbruggen, 2012), including reliability scoring. NL-CA would like this reference to be amended (https://www.rivm.nl/bibliotheek/rapporten/607711007.html).	Thank you for your support and specific comments. 

The reference has been modified accordingly.
	
5243
2018/10/19	CHEM Trust,
National NGO,
United Kingdom	CHEM Trust supports inclusion of Phenanthrene under Article 57 d) and e)  in the REACH candidate list based on the presented data in the Annex XV dossier.	Thank you for your support. 	
5251
2018/10/19	ChemSec,
International NGO,
Sweden	ChemSec supports the identification of phenantrene as an SVHC based on the evidence laid out in the dossier, and on it being a constituent of the earlier identified SVHC CTPHT.	Thank you for your support. 	
5257
2018/10/19	Concawe,
Industry or trade association,
Belgium	Introduction
Concawe welcomes the opportunity to comment on the draft SVHC assessment report on Phenanthrene, as submitted by the French Competent Authority in September 2018.
As Phenanthrene has not been registered under REACH, the obligation of gathering and assessing the missing science lies with the Rapporteur, who should, as a minimum, apply the standards that any Competent Authority demands Industry to employ when delivering their assessment reports on any substance. Before this has been done, it is difficult for any report reviewer to come to a balanced conclusion.
A review by Concawe indicates that much of the reliable data has not been incorporated in the assessment. This is not consistent with a weight of evidence determination. According to REACH Annex XIII, for identification of PBT substances a weight-of-evidence determination should be applied considering all relevant and available information. Appropriate weighting of reliability and relevance should be applied to all the available information, and a conclusion drawn based on the total weight of evidence. If individual data have not been used in the assessment due to concerns of reliability or relevance then this should be made clear in the Annex XV report and the reasons documented. We would consider it unusual for a piece of evidence to be rejected outright unless it was determined to have serious methodological flaws affecting the reliability or relevance of the study such that it is of absolutely no value to the assessment.
The additional information presented in these comments should therefore be taken into account in a new determination that is based on weight-of-evidence. Once the document has been revised, Concawe requests that there is another opportunity for stakeholder review to ensure an objective weight of evidence evaluation of all available information is being presented to the Member State Committee.
PART I
1. IDENTITY OF THE SUBSTANCE AND PHYSICAL AND CHEMICAL PROPERTIES
1.2 Composition of the substance
In paragraph 1.2 PHE is described as “PHE is not produced intentionally but does occur as constituent in UVCB2 – often together with other PAHs - mainly derived from the coal and petroleum stream”.
The fact that PHE is not produced intentionally is correct. However, it would be better to use the phrase “PHE is not produced, as such. However, it may occur as a UVCB constituent in UVCB-substances that are derived from coal or in several petroleum streams.”
The original sentence hints that all coal and petroleum stream derived UVCBs contain PHE, which is not correct, especially for lighter petroleum substances.
3. ENVIRONMENTAL FATE PROPERTIES
Based on the available weight of evidence, and in line with Annex XIII and the integrated testing and assessment strategy (ITS) detailed in ECHA R.11 Guidance, Concawe does not agree with a conclusion of vP for phenanthrene. The reasoning for this, along with data from several additional studies not included in the Annex XV report, are presented in the following comments. A poster on the persistence assessment of Phenanthrene (Hughes et al., 2018) has also been prepared for the recent CEFIC-LRI/Concawe persistence workshop and is included as an attachment with these comments.
3.1.1.3 Phototransformation/photolysis
Photodegradation is an important fate process for PAHs. Photolysis rates measured in laboratory tests equate to half-lives in the order of minutes to hours for phenanthrene, anthracene, fluoranthene, pyrene and benzo[a]pyrene (see attached data summary). Including photodegradation in risk assessment models (including the limited photic zone) shows decreases of up to 98% in the simulated water concentrations (Parkerton and Redman, 2015). It is therefore inappropriate to exclude this important fate process from consideration in the persistence assessment.
Furthermore, if photodegradation is excluded as a relevant fate process, then it should follow that photo-enhanced toxicity also be excluded from the hazard assessment, as this process would also be considered to not have a significant impact on the overall toxicity of PAHs in the environment.

3.1.2.1 Biodegradation in water and sediments
In a ready biodegradability test according to the OECD 301C (MITI) guideline method, phenanthrene achieved 67.2% mineralisation (BOD/ThOD) over 28 days (Junker et al., 2016). The 10-day window and all guideline validity criteria were met. The substance therefore meets the criteria for ready biodegradability. A robust study summary has been included as an attachment with these comments.
According to OECD guidelines, ready biodegradability tests are designed such that positive results are unequivocal and that, given a positive result, it may be assumed that a chemical will undergo rapid and ultimate biodegradation in the environment. Due to the stringent nature of ready biodegradability tests, a positive test result should generally supersede negative test results (OECD, 2006).
With reference to the ready biodegradability data presented in the Annex XV dossier, which is missing a citation (we believe the source of these data is the Japan NITE database (http://www.safe.nite.go.jp/jcheck/detail.action?cno=85-01-8&mno=4-0635&request_locale=en - accessed 25/09/2018), here the degradation was also measured by GC (78.9%) and by UV-vis (71.9%), and a conclusion of readily biodegradable is published in the database.
Therefore, an overall conclusion that phenanthrene is readily biodegradable can be drawn based on the available data and the OECD guidelines. According to the ITS detailed in ECHA guidance document R.11 (and in accordance with REACH Annex XIII), if a substance is readily biodegradable then it can be concluded “not P”. Whilst no further consideration of data relevant to the persistence of phenanthrene is necessary, additional available evidence is presented, which also supports the conclusions of the screening-level assessment.
In total 17 measurements of phenanthrene degradation in environmental surface waters are available. Measured half-lives cover a range 0.1 – 16 days, with a geometric mean of 2.9 days. These results are clearly well below the cut-offs for persistence of 40 days for freshwater and 60 days for marine water. No correlation of degradation half-life with temperature is evident from these data (see Figure 2B in phenanthrene poster).
Degradation of phenanthrene in seawater and a rural lake water at low, environmentally relevant, concentrations has been measured (Birch et al, 2018). Test solutions were prepared by passive dosing of phenanthrene as part of composed mixtures. Half-lives based on first-order kinetics were 16 and 11 days for the rural lake and seawater, respectively. In previous SVHC public consultations, a concern has been raised as to the reliability of substance degradation data derived from testing of composed mixtures due to possible co-substrate effects. However, it should be noted that phenanthrene exists and is released to the environment in the presence of other hydrocarbons. These data should therefore be considered environmentally relevant and suitable for use in the persistence assessment. Furthermore, a more recent study investigating the influence of degradation rates of substances tested in isolation and in 3-, 8- and 16-part mixtures did not find a significant influence of number of components on measured degradation rates (Hammershoj et al, 2018 - see attached poster). Concerns have also previously been raised regarding the sampling site of the seawater inoculum and potential for pre-exposure of the inoculum. However, we would like to point out that, according to the study, sampling took place in the open sea and the distances from the nearest port and the main shipping lane were 8-10 km and 1-2 km, respectively. We would not consider this as resulting in excessive pre-exposure of the inoculum, and would like to stress that hydrocarbons have many natural sources of environmental release and have been in wide-dispersive commercial use for more than a century. Hence, some degree of pre-exposure of the natural environment is considered inevitable and unavoidable.
Degradation of phenanthrene has been measured in seawater dispersions in rotating carousel systems (Brakstad et al., 2015, 2018a, 2018b, Loftus et al., 2018, Ribicic et al, 2018). Crude oil dispersions were prepared using droplet generators, with and without dispersants, and in one case phenanthrene was dosed as part of a produced water (PW) from an offshore oil platform (Loftus et al., 2018). In one case seawater was pre-adapted to the test substance (Brakstad et al., 2015) but in other cases the inoculum was non-adapted and produced comparable results. A study comparing results derived with and without dispersant demonstrated no significant influence on the measured degradation rates (Brakstad et al., 2018a). Measured half-lives from these studies varied between 0.1 and 10 days.
Degradation of phenanthrene has been measured in a biodiesel product in freshwater and a physically dispersed crude oil in seawater (Prince et al., 2008 and 2013). Degradation half-lives were 2.6 days in both experiments. A further study of the degradation of phenanthrene in passive dosed composed mixtures in seawater resulted in a half-life of 5 days (Concawe, 2012).
Specific bacterial strains have also been demonstrated to rapidly utilise and mineralise PAHs. In one instance, dynamic passive dosing has been used to demonstrate mineralisation of freely-dissolved phenanthrene and fluoranthene at comparable rates by S. paucimobilis (Smith et al., 2012). In another case a bacterial co-culture of P. citronellolis and S. maltophilia isolated from contaminated soil was shown to degrade 50% of 100 ppm phenanthrene within 7 days at an initial concentration of 105 CFU/mL and temperature of 37-40°C (Nzila et al., 2018). This was faster than the degradation of anthracene (a supposed P, but not vP, substance) by a similarly prepared culture. Data from various similar studies showing comparable results are also presented in Nzila et al., 2018.
In addition the BioHCWin v1.01 QSAR, which is an EPA EPISuite model developed specifically for the biodedegradation half-life prediction of petroleum hydrocarbons, predicts a degradation half-life of 15 days for phenanthrene (Howard et al, 2005). Recent comparisons of BioHCWin predictions to experimental degradation half-lives in aquatic tests have generally found predictions to be conservative (Prosser et al., 2016, Birch et al., 2018).
A water-sediment screening tool (WSST) test system has been developed based on the OECD 301C (MITI) test method (Junker et al., 2016). In this system, phenanthrene achieved 51.3% mineralisation over 28 days, with a corresponding ultimate degradation half-life of 25.3 days.
A water-sediment simulation test according to OECD 308 has been conducted where DegT50, system values of 116 days were determined in two sediments (Meisterjahn et al. 2018). A study summary is attached with these comments. This test employed a static biometer (rather than flow-through) design due to volatile losses of phenanthrene in the flow-through setup. This was likely to reduce oxygen flux due to a lack of air flow in the system, which could have led to greater oxygen depletion and reduced biodegradation being observed. Nevertheless, the results are below the persistence cut-off of 120 days.
The standard OECD 308 test design has previously been criticised as unrepresentative of environmental conditions due to its high sediment: water ratio and lacking relevance to flowing water bodies typical of receiving environments (Shrestha et al., 2016). Arguably, the dimensions and oxic conditions of the WSST system are more representative of typical receiving water environments and this should be considered in the weight of evidence.
A biodegradation in activated sludge simulation test according to OECD 314B has been conducted where DegT50 of 0.5 days was determined for phenanthrene (Meisterjahn et al. 2018). This was accompanied by >99% removal of parent compound by day 7. A study summary is attached with these comments. Although there are no persistence criteria directly applicable to the OECD 314B test, these data provide further weight of evidence of the high biodegradation potential of phenanthrene. A similar half-life of 0.82 days was observed in another test where activated sludge filtrate was used as the inoculum (Birch et al, 2018).
Regarding the Bahr et al (2015) study presented in the Annex XV report, we would like to point out that the study authors observe that mineralisation of phenanthrene stabilises after 29 days (not 62 days) and that they expect that amount of PAH biodegraded to be 22-85% higher than the amount of PAH mineralised. We therefore consider this study to be reasonably in line with the overall weight of evidence for the aquatic compartment, considering the non-standard method and inoculum used, and disagree with the statement “Biodegradation was possible but needed the presence of an adapted bacterial community of sufficient density to achieve mineralisation over a long period of time (100 days of colonisation plus 62 days of exposure to reach 33% of mineralisation)”.

3.1.2.2 Biodegradation in soil
A soil screening tool (SST) test system has been developed based on the OECD 301C (MITI) test method, replacing the inoculated mineral medium with a sandy loam soil (Junker et al., 2016). In this system, phenanthrene achieved 39.8% mineralisation over 28 days. It was not possible to calculate a DegT50,ultimate for soil according to the method and is reported as > 28 days.
A soil simulation test according to OECD 307 has been conducted (Meisterjahn et al. 2018), which resulted in degradation half-lives of 6.8 – 17.3 days in four soils (geometric mean: 11.1 days). A study summary is attached with these comments. These results are comparable to data for spiked soils reported in the literature (Wild and Jones, 1993; Sigmund et al, 2018). However, they contrast with experimental data presented from the two studies in the Annex XV report (Wild et al, 1991; Wild and Jones, 1993). An important difference in these studies was that phenanthrene was introduced to the soil through application of sludge from a sewage treatment plant (STP). The OECD 314B activated sludge test described above indicates that phenanthrene is subject to extremely rapid and complete transformation in a STP environment, hence any material remaining in STP sludge would be expected to be of very limited bioavailability. A substantial fraction of the PAHs in these studies may have been from pyrogenic origin (i.e. incomplete combustion), which tend to be associated with soot carbon and of very limited bioavailability, whereas petrogenic PAHs are considered more susceptible to microbial degradation (Burgess et al., 2003).
The influence of bioavailability was further highlighted by a landfarming study investigating the degradation of PAHs from dredged sediments over a period of several years (Harmsen and Rietra, 2018). Here it was possible to identify fast, slowly and very slowly desorbing fractions of PAHs and to measure their individual degradation rates, which varied significantly.
Clearly, the bioavailability of a substance is critical to the rate at which it will biodegrade in any system, and is of similar importance to other environmental factors influencing biodegradation (e.g. availability of nutrients, oxygen and competent degraders). However, bioavailability is not explicitly addressed in the framework of a persistence assessment. It is important, as far as possible, to separate extrinsic environmental factors from substance properties in persistence assessment in order to arrive at conclusions which reflect the intrinsic biodegradability of the substance and allow for consistent comparison and prioritisation between substances. Simulation test data generated according to the standard information requirements, and according to which persistence assessment criteria have been developed, allow for a consistent comparison between substances and should therefore be given a high priority in assessments based on weight-of-evidence.
It should also be noted here that substances which are not bioavailable are also expected to not pose a risk to either human health or the environment, and that the available evidence indicates that if phenanthrene were to become bioavailable it would be very rapidly degraded by indigenous microbial populations. The greatly reduced bioavailability of PAHs in soils and sediments, particularly in the presence of black carbon, has previously been recognized by RIVM (Verbruggen, 2012). Here it was advocated to use bioavailability-based methods (e.g. SPME) for monitoring PAH levels and comparing to established environmental risk limits in order to account for this reduced bioavailability in risk assessments.
In addition to the concerns raised around the bioavailability of PAHs in the Wild et al (1991) field study, we would like to note some additional concerns.
• In a later publication the authors themselves highlight the inconsistency of this dataset with other field datasets and cite possible artefacts due to use of archived soils (Northcott and Jones, 2001).
• The PAH concentrations in control soils actually increase over the duration of the experiment indicating a continuous influx of PAHs not from the initial application, most likely as a result of aerial deposition.
• Data from only one of two soils is reported. The other soil appeared to show faster degradation.
• Initial concentrations in the experiments used to derive half-lives are estimates rather than measured values.
• Treatment groups appear to not have been replicated and measured concentrations appear based on single analyses.
• The results for different treatment groups and calculation methods produce very wide variation in half-lives for the same soil.
• PAHs dosed via STP sludge would be present as a complex mixture with other substances, including the reported high levels of metals, which may have influenced the indigenous microbial populations and degradation of the PAHs.
• The experiments are lacking true controls for the measurement of PAH degradation, as the controls were not amended with STP sludge and are therefore different in other respects than simply the amounts of PAHs present.
• The SVHC Support Document for coal tar pitch, high temperature (EC number: 266-028-2) states that the Wild et al (1991) field data are given priority as they represent realistic conditions. However, according to the ECHA R.11 Guidance, field studies do not represent high-tier studies and should not override other results as they are dependent on local conditions and not directly comparable to laboratory tests or P/vP criteria.
Regarding the latter three studies presented in the Annex XV report, we would question the relevance of data for Antarctic soils and arid conditions, and encourage that they be considered in a weight of evidence context along with the other available data, particularly those data that are more in line with standard testing methods.

3.2.3 Distribution modelling
Contrary to the text, Fugacity level III does not release in equal rates to air, water, soil and sediment – only air, water and soil.
The calculation has been repeated using a water half-life of 15 days, based on both BioHCWin prediction and EPA guidance for substances passing ready biodegradability tests (https://www.epa.gov/tsca-screening-tools/interim-guidance-using-ready-and-inherent-biodegradability-tests-derive-input#eqc), the latter approach being supported by several publications from US and EU (Boethling et al., 1997; Struijs and Van den Berg, 1995; Federle et al., 1997). Default EPA conversion factors of 2 and 9 were applied for soil and sediment, respectively. These inputs can be considered conservative, based on the experimental data presented in the above sections. This resulted in an overall persistence time of 20.9 days, which is not consistent with a vP substance. The calculated compartmental mass amounts were 19.7, 69.2 and 9.62 % for water, soil and sediment, respectively.

3.3 Bioaccumulation
General approach to bioaccumulation assessments
Bioaccumulation assessments are a logical step in characterizing the potential significance of indirect exposure pathways of substances in the environment.  REACH PBT assessments utilize a laboratory Bioconcentration Factor (BCF) of 2000 L/kg WW and 5000 L/kg WW for determining bioaccumulative (B) and very bioaccumulative (vB) substances.  These are useful screening criteria and were formalized in the 2001 Stockholm Convention on persistent organic pollutants.  The basis of these metrics was deduced from legacy persistent organic chemicals that exhibited no or limited degradation in the environment and were observed to biomagnify (i.e. tissue concentrations progressively increase with trophic level) through aquatic foodchains and terrestrial foodchains (Matthies, 2016; McLachlan, 2018).  Thus, the intent of the B and vB designation is to identify substances that potentially undergo biomagnifcation and pose higher exposures and subsequent risks than might be anticipated from direct air or water exposure routes.  However, use of BCF metrics alone reflects science that is more than two decades old.  Thus, recent advances in bioaccumulation science must be considered in order to compliment BCF data and provide a credible, objective, weight of evidence (WoE) evaluation to logically inform B decision-making.
SETAC Pellston Workshops bring together experts from several stakeholder groups, including scientists from regulatory, industry and academic backgrounds and work to establish the state of the science on a particular topic.  In 2008 SETAC organized a Pellston workshop to evaluate the state of PBT assessment science (Klečka et al., 2009). This resulted in several key publications that outlined consensus guidance to improve B assessments.  First, Gobas et al (2009), which was co-authored by RIVM, advocated for a tiered approach to B assessments, (as outlined in Figure 1 of this paper – see attachment ‘’graphs’’).  The highest tier includes evaluation of substance bioaccumulation in a field food web as expressed in terms of trophic magnification factors (TMF).  Biomagnification factors (BMF) obtained using laboratory based dietary exposures with fish were identified as the next best tier to reliably characterize substance biomagnification potential.  This recommendation was a key motiviation for the subsequent development of an OECD guideline for conducting dietary bioaccumulation tests (OECD 2015).  The rationale for considering TMF and BMF as higher tier endpoints than the BCF test is that the underlying mechanistic basis for biomagnification is driven by the dietary route of exposure that is explicitly addressed using these B metrics. Use of bioconcentration factors (BCF) are used as lower tier assessments.  The BCFs can vary widely by species due to their metabolic capabilities.  The lowest tier is to screen based on logKOW. Biota-sediment accumulation factors (BSAF) can also be considered in weight of evidence acknowledging the sediment as a source to certain foodchains (Burkhard et al., 2011).

It is recognised that logKOW screening is very simple so most evaluations rely upon BCF data, which are relatively common.  However, the choice of test organisms are important due to the role of metabolism.  The ultimate goal of B assessments isto identify additional risks through successive trophic levels.  Therefore, BCF data derived from fish that typically reside at an intermediate position in the foodweb are considered most relevant for B assessment.  This is outlined in the ECHA 2017 Guidance on endpoints R11 “For the start, it should be noted that, in normal cases where experimental information on bioaccumulation is needed, a flow-through bioaccumulation test with fish according to OECD TG 305-I or OECD TG 305-II is preferred due to the best possibilities of reliably comparing the results from such test with the B/vB criteria.”  This approach provides a degree of conservatism since higher trophic organisms (birds, mammals) can exhibit enhanced enzymatic capacity to metabolize substances.

Evaluation of BCFs on invertebrates with low metabolic capability, such as mussels, essentially equate to evaluating substances based on log KOW.  This is true because logKOW simple reflects the extent of equilibrium partitioning to lipids but is insensitive to metabolism, which is a less important process in mussels.  Screening on logKOW carries the lowest weight in a B assessment, and by extension invertebrate BCF data should be given low weight in a B assessment especially if reliable fish BCF data are available.

While TMF data are valuable in characterizing substance transfer through ecosystems, experience has proven that such field studies are prone to artefacts due to seasonality, or specific foodweb dynamics such as specific predator-prey relationships, which can complicate the interpretation of TMF data (Kim et al., 2016; Burkhard et al., 2013).  Therefore, BMF data are viewed as particularly influential in current B evaluations.   BMF tests are performed with spiked diets that usually deliver better defined exposures than BCF tests since aqueous concentrations of insoluble chemicals are more variable and translated to greater uncertainty in BCF endpoints (Lo et al., 2016). Therefore, BMF becomes a critical objective line of evidence in evaluating vB potential of organic chemicals.

Measurements of BCF have a number of important considerations including careful maintenance and measurements of the exposure concentrations (aqueous or diet), as well as characterizing the uptake and depuration kinetics.  At times experimental evaluations rely on radiolabelled compounds to track the bioaccumulation and metabolism kinetics in a test organisms.  While this is potentially useful data, the use of total radioactivity includes the accumulation of metabolic break down products that can be routed into the catabolic processes of the organism normal metabolism.  Potential incorporation of the radiolabelled material into subsequent biomass has the potential to overestimate the BCF for a given chemical.  For these reasons, consensus guidance concludes radiolabelled BCF data are generally considered not fit for purpose in B assessments (Parkerton et al., 2008). Additional consideration for judging the reliability of BCF studies for B decision-making include: (1) exposure concentrations that are below the solubility limit of the substances and (2) use of test concentrations at levels that are below acute and chronic effect levels to avoid confounding effects due to potentially altered physiology.

The outlined approach above represents the state of science related to bioaccumulation assessments.  To support this review, B data referenced in the draft PBT dossiers, the Concawe PBT report from 2009, other reviews (Bleeker, 2010), as well as other data identified from a recent literature search in Google/Scholar performed during October 2018 were compiled and critically assessed for reliability.  Tables of the available data are presented attached with comments on how to consider these data.

Specific results for Phenanthrene
As discussed above, radiolabelled BCF data are not considered fit for purpose of B assessments.  Also, mussel and invertebrate BCF data are not sufficient to characterize additional risks due to biomagnification.  Therefore, fish BCF data are considered the most reliable.  Fish BCF data are given in Table 1 (see attachment ‘’graphs’’), and invertebrate BCF data are given in Table 2 (see attachment ‘’graphs’’) for reference, but are not used to evaluate B potential.  There are data on 10 fish species showing a range of BCF from 75 to 2900 L/kg.  With the exception of one study on fathead minnow all data are below the B criteria of 2000 L/kg.  Several of these are new data and were not in the SVHC proposal and should be included in the evaluation.  This suggests that PHE is a borderline B substance and needs additional lines of evidence.

Comments on specific studies
The Carlson et al 1979 study represents the highest measured BCFs.  This is a challenging study to evaluate since the exposure concentrations vary notably during the exposure.  They vary between 2 and 3.5 ug/L, which is low absolute variability, but the relative change can have (almost 2-fold) a substantial impact on the calculated BCF.  Therefore, this dataset can be used with caution.  Prior evaluations of this datasets have limited the analysis to periods of the exposure with more stable concentrations.  Therefore, the Carlson et al 1979 data are further evaluated in Lampi and Parkerton et al 2009.
Another key study is Newsted and Giesy (1987), in which the BCF for D. magna is measured. Concawe strongly disagrees that the BCF for D. magna, an organism which has limited enzymatic ability for biotransformation, serves as the rationale for concluding this substance meets the B criteria. This is particularly egregious given that fish aqueous BCF data for PHE are available. In both the REACH regulation, and in the Guidance document, it is clearly stated that bioaccumulation data from fish is the preferred test metric to support B decision-making (see general comments above). Fish represent a higher trophic level than crustaceans, and hence through both fish tests involving aqueous and dietary test exposures, a better judgement can be made on whether the substance is B or vB. The study by Newsted and Giesy (1987) is based on a single 24 hour exposure and possibly high enough exposure concentration to cause toxicity and is thus not an ideal study for BCF assessments.
Given that PHE is a borderline B substance, and that the highest point was based on a non-ideal study, additional lines of evidence are needed to conclude on bioaccumulation potential.

TMF data
TMF data can be useful since they characterize the trophic transfer of contaminants in field datasets.  Due to variability notes above, multiple studies should be critically evaluated as part of weight of evidence for bioaccumulation assessments.  For Phenanthrene there are five studies reporting TMFs: Bohai Bay 0.43 (Wan et al., 2007), Baltic Sea 0.82 (Nfon, 2008), Tokyo Bay 0.75 (Takeuchi et al., 2009), Passaic River estuary 0.34 (Khairy, 2014), and Taihu Lake 1.27 (Wang et al., 2012).
The study by Wang et al (2012) is the sole study to cite a TMF > 1, which is inconsistent with available BMF data and other existing TMF data.  Further review of the data underlying this study (Table 1, Table S2) indicates that data for a single species (catfish) are driving this observed TMF greater than 1.  This data point is >30-fold higher than the other data points in table S2 and do not support a trend analysis. The other data collected are all of similar magnitude suggesting a general lack of biomagnification.  Further regression analysis of this data set using Excel indicates confidence intervals of 0.81-2.07, which is not significantly different than the benchmark of 1.0 and reflects the high uncertainty of this dataset.  Therefore, this TMF is not considered reliable and the weight of evidence from the other studies confirms the lack of trophic magnification behaviour.
Lastly, another study (Moermond et al., 2007) measured PAH concentrations for different categories of organisms at different trophic levels, but did not include measurements of trophic position.  However comparing fish concentrations to concentrations in zooplankton suggests (Table S4) TMF < 1.
All 6 of these studies show that PHE does not biomagnify in the food chain (TMF values < 1 [ECETOC 2012]).  Therefore, Concawe is of the opinion that PHE is not bio-accumulative.BSAF
Data from a US EPA BSAF database (US Environmental Protection Agency, 2008) provides information on field-derived fish BSAF factors for 17 PAHs, PHE included. The median and 90th percentile for PHE, are less than 0.05, and 0.5, respectively. While no threshold values are available for used BSAF values, these data provide further evidence that PHE does not biomagnify in the field, and is therefore not bioaccumulative.
BSAF
Data from a US EPA BSAF database (US Environmental Protection Agency, 2008) provides information on field-derived fish BSAF factors for 17 PAHs, PHE included. The median and 90th percentile for PHE, are less than 0.05, and 0.5, respectively. While no threshold values are available for used BSAF values, these data provide further evidence that PHE does not biomagnify in the field, and is therefore not bioaccumulative.

BMF
As discussed above BMF have the largest weight of evidence for bioaccumulation assessments.  There are 5 studies that report BMF data for PHE for lab and field studies (Table 1 - see attachment ‘’graphs’’).  The reported BMFs range from 0.08 to 0.01, which indicate a lack of transfer from diet items to fish.  This confirms the observations for TMF and confirms low B potential for PHE.  Therefore, Concawe is of the opinion that PHE is not bio-accumulative.

5. ENVIRONMENTAL HAZARD ASSESSMENT
The draft SVHC assessment report on phenanthrene, as submitted by the French competent authorities, does not include discussions on the T criterion of Article 57 (a), (b), (c) or (d) of REACH. CONCAWE supports the conclusion that phenanthrene does not meet the T criteria for classification as an environmental hazard or CMR.

6. CONCLUSIONS ON THE SVHC PROPERTIES
6.2.1 Assessment of PBT/vPvB properties
In conclusion, in line with Annex XIII and the ECHA ITS, the available weight of evidence does not support a conclusion of vP for phenanthrene. The data currently presented in the Annex XV dossier Support Document do not represent a complete or representative compilation of the available weight of evidence. Phenanthrene is readily biodegradable according to OECD 301C, which is a screening trigger for a conclusion of “not P”. Furthermore, data from OECD simulation testing and other comparable literature sources indicate that the half-lives of phenanthrene are well below the respective P and vP cut-offs for water, soil and sediment.
In Concawe’s opinion, there is new data available on the bioaccumulation potential for PHE under field and laboratory conditions that do not support the vB classification of PHE. A refined assessment according to the state-of-the-art of the bioaccumulation potential of PHE also indicates that this substance does not biomagnify in the food chain, which in Concawe’s view is the main protective goal of a B-assessment. Therefore, it should be considered neither a B or vB substance.
In view of this recent scientific data, and in line with the correct interpretation of these data, Concawe is of the opinion that PHE should not be classified to be a vPvB substance.
The Coal Tar Pitch, High Temperature (CTPHT) Decision comes to the PBT/vPvB conclusion for PHE by the application of a hazard assessment that does not take into account higher tier observations in the environment (e.g. TMF values), which should be included in this current PBT assessment using a Weight of Evidence approach.
This is in line with Concawe’s view that any assessment should be made upon the state-of-the-art of the science, especially, when these latest insights give rise to the revision of decisions taken in the past. In Concawe’s view this is a key point, because otherwise the main driver for performing dedicated research to address lacunas in knowledge becomes superfluous. When any new result that overrules hypotheses or decisions from the past are not used to correct the results of historic research, which no longer can be regarded to be state-of-the-art, this will prevent any future research into the fate and effects of chemical substances, which is one of the main objectives of the REACH regulation.

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Nfon, E., I.T. Cousins, and D. Broman, Biomagnification of organic pollutants in benthic and pelagic marine food chains from the Baltic Sea. Science of the Total Environment, 2008. 397(1-3): p. 190-204.
Niimi, A. and V. Palazzo, Biological half-lives of eight polycyclic aromatic hydrocarbons (PAHs) in rainbow trout (Salmo gairdneri). Water Research, 1986. 20(4): p. 503-507.
Northcott, G.L. and Jones, K.C., 2001. Partitioning, Extractability, and Formation of Nonextractable PAH Residues in Soil. 1. Compound Differences in Aging and Sequestration. Environmental Science and Technology 35, 1103-1110.
Nzila et al., 2018. Isolation and characterisation of bacteria degrading polycyclic aromatic hydrocarbons: phenanthrene and anthracene. Archives of Environmental Protection 44, 43-54.
OECD, 2006. Revised Introduction to the OECD Guidelines for Testing of Chemicals, Section 3, OECD Guidelines for the Testing of Chemicals, Section 3, OECD Publishing, Paris, https://doi.org/10.1787/9789264030213-en.
Parkerton, T. and Redman, A., 2015. Assessment of Photochemical Processes in Environmental Risk Assessment of PAHs. Report to Concawe.
Parkerton, T.F., et al., Guidance for evaluating in vivo fish bioaccumulation data. Integrated environmental assessment and management, 2008. 4(2): p. 139-155.
Prince et al., 2008. The primary aerobic biodegradation of biodiesel B20. Chemosphere 71, 1446–1451.
Prince et al., 2013. The primary biodegradation of dispersed crude oil in the sea. Chemosphere 90, 521–526.
Prosser, C.M., Redman, A.D., Prince, R.C., Paumen, M.L., Letinski, D.J., Butler, J.D., 2016. Evaluating persistence of petroleum hydrocarbons in aerobic aqueous media. Chemosphere 155, 542−549.
Ribicic et al., 2018. Oil type and temperature dependent biodegradation dynamics - Combining chemical and microbial community data through multivariate analysis. BMC Microbiology 18:83.
Shrestha et al., 2016. Simulation Studies to Explore Biodegradation in Water–Sediment Systems: From OECD 308 to OECD 309. Environmental Science and Technology 50, 6856-6864.
Sigmund et al., 2018. Influence of compost and biochar on microbial communities and the sorption/degradation of PAHs and NSO-substituted PAHs in contaminated soils. Journal of Hazardous Materials 345, 107–113.
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Soehl, A., Wu, C. Y., 2012. Assessment of Benzo[a]pyrene Emissions in the Great Lakes Region, 2012 International Emission Inventory Conference, August 16, 2012.  (http://www.epa.gov/ttnchie1/conference/ei20/session10/asoehl_pres.pdf).
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Verbruggen, E., 2012. Environmental risk limits for polycyclic aromatic hydrocarbons (PAHs) - For direct aquatic, benthic, and terrestrial toxicity. National Institute for Public Health and the Environment (RIVM). Report number: 607711007/2012.
Wan, Y., et al., Trophic dilution of polycyclic aromatic hydrocarbons (PAHs) in a marine food web from Bohai Bay, North China. Environmental science & technology, 2007. 41(9): p. 3109-3114.
Wang, D.-Q., et al., Polycyclic aromatic hydrocarbons and organochlorine pesticides in fish from Taihu Lake: Their levels, sources, and biomagnification. Ecotoxicology and Environmental Safety, 2012. 82: p. 63-70.
Wang, H., et al., Relationship between metabolic enzyme activities and bioaccumulation kinetics of PAHs in zebrafish (Danio rerio). Journal of Environmental Sciences, 2018. 65: p. 43-52.
Wild et al., 1991. The persistence of polynuclear aromatic hydrocarbons (PAHs) in sewage sludge amended agricultural soils. Environmental pollution 72, 141–157.
Wild, S.R. & Jones, K.C., 1993. Biological and abiotic losses of polynuclear aromatic hydrocarbons (PAHs) from soils freshly amended with sewage sludge. Environmental Toxicology and Chemistry 12, 5–12.
Xia, X., et al., How does predation affect the bioaccumulation of hydrophobic organic compounds in aquatic organisms? Environmental science & technology, 2015. 49(8): p. 4911-4920.
Yokley, R.A., Garrison, A.A., Wehry, E.L., Mamantov, G., 1986. Photochemical transformation of pyrene and benzo[a]pyrene vapor-deposited on eight coal stack ashes. Environ. Sci. Technol. 20, 86–90. https://doi.org/10.1021/es00143a011	Thank you for your detailed comment on SVHC assessment report on Phenanthrene. 

General Response:

CONCAWE remarks critically that only comparably old references have been used in the SVHC proposal. This is due to the fact that mainly information provided in the RAR for CTPHT has been used for the dossier. Based on this valid information also provided in the SVHC dossier for CTPHT, the Member State Committee has already concluded on the vPvB properties of phenanthrene. More recent data were also added when relevant and reliable.

Many of the studies identified as missing by CEFIC in this comment were therefore not initially included because of poor reliability or lack of relevance for the discussion as detailed in the specific responses below. 

For transparency and when relevant, the additional information that you mentioned in your comment has been added to the support document. However, ANSES concluded that the information provided does not change the original conclusions on the identification of phenanthrene as SVHC for its PBT and vPvB properties.

FR welcomes any efforts undertaken by Concawe to characterize the fate and behaviour of phenanthrene.

Specific comments are addressed below.

Response to 1.2. Composition of the substance: 
We agree and rephrased opening sentence as suggested for more clarity.

Response to 3. Environmental fate properties:

As stated in ECHA R.11 Guidance, “Appropriate data need to be available to conclude the P/vP-assessment with a conclusion “not P/vP” on all three compartments (or five, with marine compartments): water (marine water), sediment (marine sediment) and soil. […]if a conclusion “P” or “vP” is reached for one compartment, no further testing or assessment of persistence of other environmental compartments is normally necessary.” As stated in the SVHC dossier, phenanthrene is considered very persistent in sediment, thus allowing its classification as a vP substance.

3.1.1.3 Phototransformation/photolysis

The R11 ECHA guidelines clearly states in page 43, that data derived from abiotic studies such as photodegradation cannot be used on their own for the assessment of persistence. 

3.1.2.1 Biodegradation in water and sediments

The data from the study of Junker et al., 2016 has been added to the SVHC annex XV document. As explained in the dossier, based on the available information from the ready biodegradation tests showing not ready biodegradation for one result (Junker et al., 2016), and ready biodegradation from the other (INERIS, 2010), biodegradation of phenanthrene may occur in water. 

Nevertheless, a classification as vP is possible based on the vP properties of phenanthrene in other compartments, as stated in the ECHA R.11 guidance document. 
A water-sediment screening tool (WSST) test system performed in this study also stated that phenanthrene achieved 51.3% mineralisation over 28 days, which indicates that phenanthrene is not readily degradable. Moreover, this study also states that according to their SST screening test, there is evidence that phenanthrene is persistent in soil.
Mineralisation of phenanthrene at 5°C in seawater was 63.2 days, above the criteria for vP in marine waters. (Brackstad et al., 2018, Biodegradation-mediated alterations in acute toxicity of water accommodated fraction and single crude oil components in cold seawater, chemosphere (204), 87-91.)

See our previous answer above this point (ITS) (Response to point 3. Environmental fate properties).

In the study of Birch et al., 2018, the authors used as inoculum activated sludge filtrate, seawater and lake water. Activated sludge filtrate could be considered as pre-adapted to hydrocarbons (some of the hydrocarbons were present at or above 10% of the test concentration) and seawater could be considered also as pre-adapted to hydrocarbons (obtained in the vicinity of a trafficked shipping port). In addition, the tests were conducted in relative small systems (15mL), which implies a rather limited population of microorganisms. Considering the remarks above FR considers that the study suitability for the P assessment is limited, as it presents methodological limitations.

Brakstad et al. (2015) used a pre-adapted seawater to determine biodegradation of dispersed oil droplets at low temperature. The method used is not suitable to evaluate P in a PBT/vPvB assessment for pristine environments.

In the study of Brakstad et al., 2018 a, b, the authors studied the biodegradation of oil with and without dispersion in natural seawater at 5°C. The experiments conducted with dispersant are not considered as an environmentally relevant condition. Furthermore, no abiotic control was used and it is not possible to determinate the loss of the substance by abiotic process. In addition, no information about the dissolved phenanthrene is available in the study.  The authors calculated the half-life from rate coefficient using the equation t1/2=0.693/k1. Therefore, FR considers that the study is of limited suitabilityfor the P assessment.

The study of Loftus et al., 2018 is also questionable for P-assessment, due to the use of “produce water” from offshore oil production installation containing dispersed oil compounds. As the sample contained dispersants, this condition is not considered as an environmentally relevant condition. In addition, for the determination of half-lifes, the authors expressed the values relative to internal marker (30 ab Hopane) and no justification is provided in the paper about the use of this marker, especially about the similarity with phenanthrene.

In the study of Ribicic et al, 2018. The authors investigated a comparative multivariable approach for studying the biodegradation of dispersed oil in seawater. The use of dispersant in the experiments is not considered as an environmentally relevant condition. In the figure S1 the authors show biotransformation of PAHs relative to an internal marker 30ab Hopane. No justification is provided in the paper about the use of this marker, especially about the similarity with phenanthrene.  No information is available regarding the determination of DT50 for every PAHs. Overall, considering all these remarks, FR concluded that the study is of limited suitability for the P assessment. 

Regarding the study of Prince et al 2008 and 2013, some losses observed during the assessment might be attributable to evaporation as stated by the authors. No abiotic control was included to indicate and correct for effect of dissipation. It was unclear whether all individual constituents were fully dissolved and half-lives were expressed relative to internal markers. These internal markers were not selected based on their similarity in physical and chemical properties (and thus cannot be expected to behave similarly: evaporation, solubility, etc.).
For the study of 2008, it is unclear how the inoculum from rain water depot is sufficiently representative of environmental conditions. It is also the case for the addition of mineral media in the experiment.

Regarding the study of Concawe, 2012, indeed it is noted that for phenanthrene a half-life of 5 days was calculated, however the reliability of the study is questionable, due to the addition of non-sterilised mineral media and the relatively small test systems (15 mL) used in the experiments, which implies a rather limited population of microorganisms.

Specific bacterial strain cited cannot be used in regulatory assessment. They are part of bioaugmentation process, and are considered as adapted bacteria.

No information on BIOWIN predictions have been included in the dossier because the available experimental information was considered enough and more relevant to conclude on persistence in soils.
Moreover, in a recent PBT factsheet on 3 ring aromatic PAHs (September 2018), it was highlighted that BioHCWin provided very different half-lives than BioWin3 for phenanthrene, leading to inconclusive results and cannot be used without experimental data. BioWin 3, which is largely used in regulatory assessments, provide an estimation of vP for phenanthrene, at the opposite of BioHCWin.

Regarding the water-sediment simulation test OECD 308, the half-life value is very close to the threshold for the classification as P when looking at the results. Nevertheless, the test was performed at 20°C and, for a regulatory purpose, the recommended temperature is 12°C. No recalculation was performed by the data owners by using the Arrhenius equation as recommended by ECHA R.7. When the recalculation is performed, the DT50 value are all > cut-off value of 180 days (between 216.2 to 284.5 days with the simplified method recommended by ECHA, or between 242.1 to 318.5 days with the standard methodology). According to the extended summary provided by Concawe, the study presented the appropriate control and stick to the recommendation of OECD (positive and negative control, water-sediment ratio …).
Thus, mineralisation is quite low, the NER fraction reachs 30% in one sediment, DT50 in sediment is >304 days for coarse texture sediment and > 40 days for water. These data indicate that phenanthrene is vP as stated in the SVHC support document and confirm the conclusion.

Regarding the study of Wild and Jones, 1993, this study was conducted with different types of soils, some of them could be considered as pre-adapted. However, the experiments conducted with the rural agricultural soil with no previous sludge application exhibited a half-life for phenanthrene of 193 days (up to 643 days after correction of the temperature), the soil from coniferous forest stand: half-life of 112days (up to 373 after temperature correction). The study is considered valid and provides a value above the cut-off value for vP criteria. This was stated during the assessment of the CTPHT and agreed by the MSC.

The OECD 314B is not suitable for P assessment according to guidance document R7 and R11.

Regarding the Bahr et al., 2015 study, your remark is noted. Nevertheless, FR considers that this study is in line with the weight-of-evidence presented.

Response to 3.1.2.2. Comments on Biodegradation in soil

Regarding the soil simulation test according to OECD 307, the test was performed at 20°C and, for a regulatory purpose, the recommended temperature is 12°C. No recalculation was performed by the data owners by using the Arrhenius equation as recommended by ECHA R.7. Moreover, it seems to have a deviation regarding the recovery of radioactivity, exceeding quality criteria. However, this study indicates that degradation of phenanthrene can occur in soil.

It is mentioned by Concawe that the source of PAHs can have an influence on their bioavailability. The differences between pyrogenic and petrogenic origin is highlighted in the article by Burges et al (2003). However, FR believes that this is not related to the ad hoc circumstances of specific tests/studies. Furthermore, Concawe underlines that bioavailability is not explicitly addressed in the persistence assessment, which FR agrees upon because we consider this is not directly relevant for SVHC identification that shall be based on the intrinsic properties of phenanthrene.  
Regarding the study of Wild et al (1991), the half-life calculated under field conditions supports the very low degradation of phenanthrene in soils, observed in the study under laboratory conditions of Wild and Jones (1993). FR highlights that the MSC has agreed on the relevance of this study on the assessment of persistence properties of phenanthrene in the context of CTPHT assessment.

Regarding the Harmsen and Rietra, 2018, study, it is suggested that the degradation rate of phenanthrene may depend on the initial concentration and that a certain fraction can remain for a long period of time in the environment. 

In conclusion, the current information does not allow for a final conclusion on persistency of phenanthrene in soil.

Response to 3.2.3. Comments on Distribution modelling

It is observed that the results from the modelisation presented in the dossier and the one done by Concawe are close, indicating that soil is the main receptacle of phenanthrene. Moreover, in the study by Brackstad et al., 2018, the authors, as cited by Concawe, also indicate that EPI suite modelisation provides a vP ranking for phenanthrene and that no ready biodegradation is expected (Brackstad et al., 2018, Biodegradation-mediated alterations in acute toxicity of water accommodated fraction and single crude oil components in cold seawater, chemosphere (204), 87-91.).

Response to 3.3. Comments on Bioaccumulation

The agreement of the MSC on the identification of CTPHT as SVHC (ECHA, 2009), adopted on 2 December 2009 concludes on page 4: “Based on the data available, it is concluded that seven of the 12 PAH-constituents present in CTPHT in concentrations equal to or above 0.1% are to be considered as both vPvB and PBT substances. These are: fluoranthene, pyrene, benz(a)anthracene, chrysene, benzo(a)pyrene, benzo(k)fluoranthene and benzo(ghi)perylene. 
Phenanthrene fulfils the vPvB criteria, but not the PBT criteria.” The SVHC dossier submitter correctly reflected this fact and added more recent data.

Concawe disagrees with the approach followed to conclude on the B and vB properties of phenanthrene based on data for lower trophic levels. However, the most sensitive organisms towards PAHs are those lacking a cytochrome P450 enzyme system. It has been shown without a doubt that bioaccumulation occurs in those organisms likely because they cannot degrade PAHs. This has been clearly assessed in the SVHC proposal. 
Furthermore, the Guidance Chapter R11 version 3.0 of June 2017 states the following:
(page 69) “Also use of other taxonomic groups than fish (e.g. mussel bioconcentration test, ASTM, 2003) is possible for measuring bioconcentration in the aquatic environment and the valid BCFs determined in other taxonomic groups can be used in assessing whether or not the B/vB criteria are met.” Therefore it is completely valid to use the most reliable data with invertebrate to conclude in the B/vB properties of phenanthrene.

We note that the PBT/vPvB assessment addresses the differences between the various BCF values as well as differences between the various reported B-metrics. This encompasses a discussion on the difference of the bioaccumulation potential in invertebrates compared to, e.g., fish. It should be noted, that the purpose of the PBT/vPvB assessment is not only to protect from biomagnification in the food chain, but also from high bioaccumulation at any individual food chain steps.   

Regarding the table that Concawe presented as supporting material for fish BCF data for phenanthrene, some of the studies were already included in the SVHC proposal (Johnson et al., 2004, Carlson, 1974). The other data were added in the support document for transparency. However, the low BCF values reported for these fish species does not change the overall conclusion of the B/vB properties of phenanthrene based not solely on fish but also mainly on invertebrate data. Moreover, some of these data were already assessed by the RIVM and were rated as R3 and 4. 

Most of the information presented in the dossier was already assessed by Bleeker and Verbruggen, 2009. The authors assigned analysis Klimisch scores to each BCF value to indicate the reliability. In addition, the authors clearly indicate that a validity of 4 is allocated to BCF values being based on total radioactivity (which deems it impossible to distinguish between the concentration of the parent compound and that of its transformation products), studies with uncertainties about the exposure concentration (e.g. exposure concentration reported in graphs only, nominal concentration reported only, exposure via sediment or diet, exposure to oil) or studies with various (pre-)exposure durations (e.g. for field collected animals). Assessment of reliability scores by Bleeker and Verbruggen, 2009 has been used in the current SVHC proposal for phenanthrene and these studies are therefore not further considered for the conclusion.

According to recent work presented during the PBT expert group in October 2018, a PBT factsheet on three aromatic ring PAHs, the lipid normalized BCF in fish can reach 6118 L/kg and is considered of R2 (Carlson et al., 1979). Again, this data allow the classification of phenanthrene as vB. This is reflected in the support document.

This study (Newsted and Giesy (1987) is not used in the support document for B assessment.

According to the results from field studies and biomagnification, Concawe is of the opinion that phenanthrene is not bio-accumulative. However, the R.11 ECHA guidance document on page 77, indicates clearly that “[…] as dietary and trophic biomagnification represent different process than bioconcentration in aquatic organisms, BMF and/or TMF values<1 cannot be directly used to disragard a valid assessment based on reliable BCF data indicating that a substance meetis the numerical B/vB criteria in Annex XIII to the Reach regulation”. Thus, the fact that no trophic magnification is observed in the studies reported does not invalid the previous conclusion that phenanthrene fulfills the B/vB criteria.

The data from the US EPA BSAF (biota sediment accumulation factor) database was not considered in the assessment, because any information about the methodology followed in the experiments, the method of calculation of BSAF and the source of the data are reported in the database. In addition, as it is stated in the page 75 of R11 guideline, no threshold values for using sediment BSAF values in PBT assessment are available. Thus, for all these reasons, FR does not considered the information from the database.

Response to 6.2.1. Comments on Concawe’s assessment of PBT/vPvB properties

Concawe does not agree with the identification of phenanthrene as SVHC for its vPvB properties. Furthermore, Concawe addressed studies presented as new to take into consideration in the assessment especially in the persistence of the substance. FR assessed the data submitted and concluded that this information does not undermine and change the conclusion that phenanthrene is persistent and very persistent in sediment

Moreover, it is important also to remind that ECHA R.11 Guidance states the following: “Appropriate data need to be available to conclude the P/vP-assessment with a conclusion “not P/vP” on all three compartments (or five, with marine compartments): water (marine water), sediment (marine sediment) and soil. […] if a conclusion “P” or “vP” is reached for one compartment, no further testing or assessment of persistence of other environmental compartments is normally necessary.” It therefore acknowledges that persistence demonstrated in one compartment is sufficient to conclude that a substance is P/vP.
As concluded in the Annex XV dossier, phenanthrene is considered very persistent in sediment and therefore fulfills criteria for its identification as P and vP. 

Regarding bioaccumulation, the data presented by Concawe regarding fish and the observations about field studies, don’t change the fact that phenanthrene is bioaccumulable and very bioaccumulable.

Overall we come to the conclusion that the new information provided by Concawe does not disqualify the studies already included in the SVHC dossier and evaluated during MSC decision on coal tar pitch high temperature and that can thus not change the conclusion that phenanthrene fulfils the criteria of REACH Art. 57 e).

With consideration of all relevant data and in accordance to the methodology recommended in the appropriate guidance, phenanthrene therefore fulfils criteria of article 57 (e) as a vPvB substance.

	
		See the corresponding embedded attachment in table 1 of Part I: 5257_PHEN.zip		
5267
2018/10/30	Finland,
Member State	As an addition to the FI CA comments submitted before, please find below summaries of benzo[k]fluoranthene, fluoranthene, phenanthrene, and pyrene concentrations in biota as determined in field studies.

Benzo[k]fluoranthene:
Benzo[k]fluoranthene has been determined in 22 pike samples (white muscle) from Finnish lakes and rivers, and coastal areas of Baltic sea, in 2003-2004 (Finnish Environment Institute 2018). Benzo[k]fluoranthene concentration was above limit of determination in 13 samples (0.01-0.03 µg/kg). For the other 11 samples, Benzo[k]fluoranthene concentration was below limit of determination of 10 µg/kg.
Benzo[k]fluoranthene has been determined in 50 mussel samples (blue mussel, duck mussel, painter’s mussel, and unspecified freshwater mussels (Anodonta)) from Finnish lakes and rivers and coastal areas of Baltic sea, in 2016-2017 (Finnish Environment Institute 2018). Benzo[k]fluoranthene concentration was above the limit of determination in four blue mussel samples (1.1-4.4 µg/kg). For the other 46 samples (blue mussel, duck mussel, painter’s mussel, and unspecified freshwater mussels (Anodonta)), Benzo[k]fluoranthene concentration was below the limit of determination (limits of determination were 1 or 2µg/kg for different sets of samples).
Benzo [b + k + j] fluoranthenes have been detected in caged blue mussels (Mytilus trossolus) 2.9-10.3 ng/g dw outside Finnish (Porvoo and Hanko) and Swedish (Gävle and Sundsvall) towns in the Baltic Sea (Turja et al 2014; Turja et al 2013). In addition, a concentration of 14.3 ng/g dw for  benzo [b + k + j] fluoranthenes was observed in native mussels used for the experiment before caging (Turja et al. 2014). Hanko was the reference site of these studies whereas the other sites were affected by anthropogenic pollution.

There is no data on concentrations of benzo[k]fluoranthene in the water for these studies.
Bioaccumulation factors were not quantified in these studies. However, the results indicate the occurrence of benzo[k]fluoranthene in pike and in mussels. It is noted that high BCFs in mussels have been reported for other PAHs, such as fluoranthene (BCF 4120-5920), pyrene (4430-77000) and benzo[a]pyrene (41000-142000) (ECHA, 2009). Therefore, mussels may be relevant target species for bioaccumulation of Benzo[k]fluoranthene.
References:
ECHA (2009): Support Document for identification of Coal Tar Pitch, High Temperature as
a SVHC because of its PBT and CMR properties.
http://echa.europa.eu/documents/10162/73d246d4-8c2a-4150-b656-
c15948bf0e77
Finnish Environment Institute 2018. Open information. Ympäristötiedon hallintajärjestelmä Hertta. Kertymärekisteri (KERTY). Available at https://wwwp2.ymparisto.fi/scripts/kirjaudu.asp (registration required) (in Finnish)

Turja, R., Soirinsuo, A., Budzinski, H., Devier, M. H., Lehtonen, K. K. 2013. Biomarker responses and accumulation of hazardous substances in mussels (Mytilus trossulus) transplanted along a pollution gradient close to an oil terminal in the Gulf of Finland (Baltic Sea). Comparative Biochemistry and Physiology, Part C 157 (2013) 80–92. http://dx.doi.org/10.1016/j.cbpc.2012.09.006
Turja, R., Höher, N., Snoeijs, P., Baršienė, J., Butrimavičienė, L., Kuznetsova, T., Kholodkevich, S. V., Devier, M.-H., Budzinski, H., Lehtonen, K.K. 2014. A multibiomarker approach to the assessment of pollution impacts in two Baltic Sea coastal areas in Sweden using caged mussels (Mytilus trossulus). Science of the Total Environment 473-474, 398-409. https://doi.org/10.1016/j.scitotenv.2013.12.038

Fluoranthene:
Fluoranthene has been determined in 51 mussel samples (blue mussel, duck mussel, painter’s mussel, and unspecified freshwater mussels (Anodonta)) from Finnish lakes and rivers and coastal areas of Baltic sea, in 2016-2017 (Finnish Environment Institute 2018). Fluoranthene concentration was above the limit of determination in 30 mussel samples (1.1-10 µg/kg) (blue mussel, duck mussel, painter’s mussel, and unspecified freshwater mussels (Anodonta)). For the other 21 samples (blue mussel, duck mussel, freshwater mussels (Anodonta)), fluoranthene concentration was below the limit of determination of 1 µg/kg.
Fluoranthene has also been detected in caged blue mussels (Mytilus trossolus) 3.6-23.8 ng/g dw outside Finnish (Porvoo and Hanko) and Swedish (Gävle and Sundsvall) towns in the Baltic Sea (Turja et al 2014; Turja et al 2013). In addition, a fluoranthene concentration of 67.6 ng/g dw was observed in native mussels used for the experiment before caging (Turja et al. 2014). Hanko was the reference site of these studies whereas the other sites were affected by anthropogenic pollution.

There is no data on concentrations of fluoranthene in the water for these studies.
Bioaccumulation factors were not quantified in these studies. However, the results indicate the occurrence of phenanthrene in mussels.
References:
Finnish Environment Institute 2018. Open information. Ympäristötiedon hallintajärjestelmä Hertta. Kertymärekisteri (KERTY). Available at https://wwwp2.ymparisto.fi/scripts/kirjaudu.asp (registration required) (in Finnish)

Turja, R., Soirinsuo, A., Budzinski, H., Devier, M. H., Lehtonen, K. K. 2013. Biomarker responses and accumulation of hazardous substances in mussels (Mytilus trossulus) transplanted along a pollution gradient close to an oil terminal in the Gulf of Finland (Baltic Sea). Comparative Biochemistry and Physiology, Part C 157 (2013) 80–92. http://dx.doi.org/10.1016/j.cbpc.2012.09.006
Turja, R., Höher, N., Snoeijs, P., Baršienė, J., Butrimavičienė, L., Kuznetsova, T., Kholodkevich, S. V., Devier, M.-H., Budzinski, H., Lehtonen, K.K. 2014. A multibiomarker approach to the assessment of pollution impacts in two Baltic Sea coastal areas in Sweden using caged mussels (Mytilus trossulus). Science of the Total Environment 473-474, 398-409. https://doi.org/10.1016/j.scitotenv.2013.12.038

Phenanthrene:
Phenanthrene has been determined in 22 pike samples (white muscle) from Finnish lakes and rivers, and coastal areas of Baltic sea, in 2003-2004 (Finnish Environment Institute 2018). Benzo[k]fluoranthene concentration was above the limit of determination in 9 samples (0.18-0.82 µg/kg). For the other 13 samples, Benzo[k]fluoranthene concentration was below the limit of determination (limits of deteramination were 0.01 or 20 µg/kg for different sets of samples).
Phenanthrene has been determined in 51 mussel samples (blue mussel, duck mussel, painter’s mussel, and unspecified freshwater mussels (Anodonta)) from Finnish lakes and rivers and coastal areas of Baltic sea, in 2016-2017 (Finnish Environment Institute 2018). Phenanthrene concentration was above the limit of determination in 28 mussel samples (blue mussel, duck mussel, painter’s mussel, and unspecified freshwater mussels (Anodonta)) (1-3.7 µg/kg). For the other 23 samples (blue mussel, duck mussel, and unspecified freshwater mussels (Anodonta)), phenanthrene concentration was below the limit of determination of 0.01 µg/kg.
Phenanthrene has also been detected in caged blue mussels (Mytilus trossolus) 12.8-72.3 ng/g dw outside Finnish (Porvoo and Hanko) and Swedish (Gävle and Sundsvall) towns in the Baltic Sea (Turja et al 2014; Turja et al 2013). In addition, a phenanthrene concentration of 334.0 ng/g dw) was observed in native mussels used for the experiment before caging (Turja et al. 2014). Hanko was the reference site of these studies whereas the other sites were affected by anthropogenic pollution.

There is no data on concentrations of phenanthrene in the water for these studies.
Bioaccumulation factors were not quantified in these studies. However, the results indicate the occurrence of phenanthrene in pike and in mussels.

References:

Finnish Environment Institute 2018. Open information. Ympäristötiedon hallintajärjestelmä Hertta. Kertymärekisteri (KERTY). Available at https://wwwp2.ymparisto.fi/scripts/kirjaudu.asp (registration required) (in Finnish)

Turja, R., Soirinsuo, A., Budzinski, H., Devier, M. H., Lehtonen, K. K. 2013. Biomarker responses and accumulation of hazardous substances in mussels (Mytilus trossulus) transplanted along a pollution gradient close to an oil terminal in the Gulf of Finland (Baltic Sea). Comparative Biochemistry and Physiology, Part C 157 (2013) 80–92. http://dx.doi.org/10.1016/j.cbpc.2012.09.006
Turja, R., Höher, N., Snoeijs, P., Baršienė, J., Butrimavičienė, L., Kuznetsova, T., Kholodkevich, S. V., Devier, M.-H., Budzinski, H., Lehtonen, K.K. 2014. A multibiomarker approach to the assessment of pollution impacts in two Baltic Sea coastal areas in Sweden using caged mussels (Mytilus trossulus). Science of the Total Environment 473-474, 398-409. https://doi.org/10.1016/j.scitotenv.2013.12.038

Pyrene:
Phenanthrene has been determined in 51 mussel samples (blue mussel, duck mussel, painter’s mussel, and unspecified freshwater mussels (Anodonta)) from Finnish lakes and rivers and coastal areas of Baltic sea, in 2016-2017 (Finnish Environment Institute 2018). Phenanthrene concentration was above the limit of determination in 24 mussel samples (blue mussel, duck mussel, painter’s mussel) (1.1-11 µg/kg). For the other 27 samples (blue mussel, duck mussel, freshwater mussels (Anodonta), phenanthrene concentration was below the limit of determination of 0.01 µg/kg.
Pyrene has also been detected in caged blue mussels (Mytilus trossolus) 3.6-44.4 ng/g dw outside Finnish (Porvoo and Hanko) and Swedish (Gävle and Sundsvall) towns in the Baltic Sea (Turja et al 2014; Turja et al 2013). In addition, a pyrene concentration of 129.9 ng/g dw was observed in native mussels used for the experiment before caging (Turja et al. 2014). Hanko was the reference site of these studies whereas the other sites were affected by anthropogenic pollution.

There is no data on concentrations of pyrene in the water for these studies.
Bioaccumulation factors were not quantified in these studies. However, the results indicate the occurrence of pyrene in mussels.
References:
Finnish Environment Institute 2018. Open information. Ympäristötiedon hallintajärjestelmä Hertta. Kertymärekisteri (KERTY). Available at https://wwwp2.ymparisto.fi/scripts/kirjaudu.asp (registration required) (in Finnish)

Turja, R., Soirinsuo, A., Budzinski, H., Devier, M. H., Lehtonen, K. K. 2013. Biomarker responses and accumulation of hazardous substances in mussels (Mytilus trossulus) transplanted along a pollution gradient close to an oil terminal in the Gulf of Finland (Baltic Sea). Comparative Biochemistry and Physiology, Part C 157 (2013) 80–92. http://dx.doi.org/10.1016/j.cbpc.2012.09.006
Turja, R., Höher, N., Snoeijs, P., Baršienė, J., Butrimavičienė, L., Kuznetsova, T., Kholodkevich, S. V., Devier, M.-H., Budzinski, H., Lehtonen, K.K. 2014. A multibiomarker approach to the assessment of pollution impacts in two Baltic Sea coastal areas in Sweden using caged mussels (Mytilus trossulus). Science of the Total Environment 473-474, 398-409. https://doi.org/10.1016/j.scitotenv.2013.12.038	Thank you for the information provided. After careful consideration, it has been decided not to be included to the support document, as the existing weight-of-evidence argumentation on the conclusion of the substance as vB was deemed enough.	

PART II: Comments and responses to comments on uses, exposures, alternatives and risks

Specific comments on use, exposure, alternatives and risks
Number / Date	Submitted by (name, submitter type, country)	Comment	Responses	
5200
2018/10/16	Finland,
Member State	11.1., Page 22: “Phenanthrene is not produced intentionally but occurs together with other PAHs as constituent of coal and petroleum stream UVCB substances. Among the PAHs benzo[a]pyrene (B[a]P, EC number: 200-028-5), chrysene (EC number 205-923-4), benz[a]anthracene (B[a]A, EC number 200-280-6) and anthracene (EC number: 204-371- 1) are included in the Candidate List due to PBT and vPvB properties already”.

Comment: Please note that in addition to the listed PAHs, also Benzo[ghi]perylene (B[ghi]P, EC Number: 205-883-8) has been included in the Candidate List due to PBT and vPvB properties.	Noted. 	
5203
2018/10/17	Euroalliages,
Industry or trade association,
Belgium	See our position paper also submitted for Benzo[k]fluoranthene	See response provided in the table for general comments.	
		See the corresponding embedded attachment in table 1 of Part I: 5203_Position paper candidates listing of PAHs October 2018.pdf		
5257
2018/10/19	Concawe,
Industry or trade association,
Belgium	7. REGISTRATION AND C&L NOTIFICATION STATUS
As there is no registration dossier for Phenanthrene, Concawe queries the relevance of this part of the dossier, because there is no REACH data source.  In that respect, it has to be clearly stated that Phenanthrene is never added to substances but is present due to its presence in starting materials, which often are of natural origin.
In view of this Concawe queries when declaring Phenanthrene as SVHC how Risk Management Options on the substance can be implemented.
Furthermore, as it will never lead to an authorisation process for the substance Phenanthrene, Concawe considers part II obsolete.  Moreover, in the analytical database in Concawe comprising of 295 samples, 244 samples exhibit Phenanthrene levels above 1 ppm and in a total of 51 samples Phenanthrene was not detected or quantitated.  Of these analytical results only 11 samples have levels > 1000 ppm and these are only found in the HFO (7), CGO (3) and VHGO (1) categories that are fuels and exempted from Authorisation under REACH through article 56 sub 4 c & d.
Notwithstanding this opinion, Concawe will comment on some of the content of this part.
As Phenanthrene is not produced as such the Rapporteur has to alter the text in Chapter 8, as Phenanthrene is only present is minor amounts in petroleum substances of which there are only 198 registered.  Therefore, claiming that there are “numerous petroleum UVCBs” is exaggerated and the correct number should have been used.  In addition, Concawe, after consulting its Membership, no longer supports the use of petroleum substances with high PAH-contents for road construction.  During the latest dossier update, this professional use is no longer supported.  Therefore, this use should be removed from this report.
Regarding the sources of Phenanthrene, Concawe observes that the focus of the Rapporteur does not mention any other anthropogenic sources that may introduce Phenanthrene into the environment or lead to human exposures other than the petroleum and coal based substance.  Concawe considers this to be a major omission, as without highlighting that residential wood burning being the major source for Phenanthrene observed in the atmosphere (Soehl & Wu, 2012; Jaarrapport, 2018), this provides a totally incorrect impression on the potential Phenanthrene sources.  Also smoking, barbequing and the overcooking of meat are considerable human exposure sources that are not mentioned.  Concawe would have expected that the Rapporteur should have referenced their own national findings on this, as well as, asked the other Member States on similar data and source reports.
As discussed in Part I, The source will heavily influence the environmental fate, behaviour and bioavailability of a PAH: PAHs are generally categorized as either of pyrogenic or petrogenic origin. Pyrogenic PAHs are formed rapidly during incomplete combustion of organic material: forest or grass fires, volcanic eruptions, combustion engine exhausts, residential wood burning, coal-fired power plant emissions, coke-oven emissions, and coal-tar from legacy manufactured gas plants and pavement sealants (Baldwin et al., 2017). Pyrogenic PAHs associated with soot carbon formed during the incomplete combustion process are far less available for bio- or photodegradation and are thus much more prevalent in the environment than petrogenic PAHs (Burgess et al., 2003, Butler and Crossley, 1981; Korfmacher et al., 1980; Yokley et al., 1986).
Petrogenic PAHs are formed naturally from biomass at low temperatures (100-150°C) over long geologic time scales and are released from natural oil seeps or unintentionally from unprocessed coal, crude oil or refined petroleum products (petroleum fuels, lubricants, asphalt). Petrogenic PAHs are released into the environment in complex mixtures of aliphatic and aromatic hydrocarbons and are associated with a local or point source, such as refineries and petroleum industries, offshore and onshore oil and gas drilling, and adjacent to roads and navigational routes. In contrast, pyrogenic PAHs occur on broader geographical scales (Burgess et al. 2003). In contrast to soot associated pyrogenic PAHs, released petrogenic PAHs are generally available for abiotic and biotic degradation and partitioning processes, as supported by the data presented in Part I. Atmospheric vapor phase PAHs are expected to be rapidly removed in the atmosphere by photochemical oxidation. Petrogenic PAHs reaching the aquatic environment (either by wet deposition or direct influx) will be exposed to photochemical oxidation, biodegradation, (re)volatization and sedimentation. In the sediment, dissolution and further biodegradation can occur in especially the oxic upper layers. In contrast, particle associated pyrogenic PAHs are far less available to degradation or partitioning process and tend to end up in the sediment, where their potential for dissolution or biodegradation is limited. This, along with the relative contributions from these sources, explains why pyrogenic PAHs tend to dominate in the aquatic environment and that petrogenic PAHs do not occur in the sedimentary record to the same extent (Burgess et al., 2003).
Therefore, Concawe questions, if measures on substances that contain Phenanthrene in minor quantities will be an effective way of mitigating the potential risks associated with this.
On page 21 the Rapporteur mentions that “Phenanthrene is not produced, as such and is not registered within the scope of the REACH Regulation. However, it may occur as a constituent in UVCB-substances that are derived from coal or in several petroleum streams. There is a broad range of different uses, e.g. production of other chemicals and rubber, uses in road construction and as lubricants”. Concawe is of the opinion that this statement is not entirely correct and should, as a minimum, be accompanied by a reference that allows the reader to verify and comment on the screening of the Rapporteur.
Moreover, the presence of Phenanthrene in any petroleum stream is due to its presence in crude oil, the processes that deliver these UVCB substances and the physical-chemical properties of Phenanthrene.  Therefore, Phenanthrene is only observed in a limited number of petroleum UVCBs.  The Concawe analytical database, comprising of 295 individual PS-samples, only has analytical results for 11 tested substances with a Phenanthrene content above the Classification threshold of 1000 ppm.
With respect to the information on the use of substances that might contain Phenanthrene, Concawe is surprised to read that “numerous UVCB substances that contain Phenanthrene”.  As far as Concawe is aware, from the 198 registered petroleum substances, there are only 92 substances that have demonstrable levels of Phenanthrene, as determined by analysis.  Of these, 22, contain less than 1 ppm Phenanthrene.

11. ADDITIONAL INFORMATION
11.2 Alternatives
Concawe has a severe objection to the statement in chapter 11.2 on page 22, namely: “articles derived from coal or mineral oil (see for example use of low aromatic oils in tyres, (KEMI, 2003)).”  Concawe is not aware of any article that are derived from petroleum, directly.  The example provide concerns the use of a petroleum substance during the manufacture of rubber tyres, as the tyres themselves are made from rubber that is either a natural product or synthesised from isoprene, butadiene or a combination of these.
This is incorrect for Petroleum Substances as the amount of PHE or any other PAH cannot be reduced without altering the substance’s properties and use potential. The technology available for that in the refining sector would be hydro-treatment that besides reacting with PHE will completely alter the composition of the substance completely, leading to a new substance that is no longer fit for the original purpose. Moreover, it would become another substance under REACH that, giving the fact that these upgrading processes already take place at many refineries, has been registered under another EINECS number.
The example quoted in section 11.2, the extender oil in tire manufacture, is a good example of this, as the PAH reduction is not achieved by treating the original extender oils but by shifting to other feedstocks of Petroleum Substances with similar physical-chemical properties that never contained PAHs, above the limits stipulated in the REACH Restriction #50.  However, as this restriction is not applicable to Pyrene this whole section should be removed.
Concawe remains of the opinion that the report would gain considerable value when the current conclusions are compared and placed into the correct context with the total Phenanthrene emissions and the major sources of Phenanthrene intake by humans like smoking, char-coal barbeques, eating of burned or barbequed meat, residential wood burning, ill-functioning stoves and open fire places, etc., etc.
Finally, Concawe would like to point out that the refining sector is consistently monitoring and attempting to reduce its PAH emissions.  However, as Phenanthrene is not subject to any regulatory obligation to be monitored, Concawe has no data on this substance.  However, as the trend in PAH-discharges still demonstrates a decrease in PAH concentrations and loads in effluents, it can be assumed that this also applies to Phenanthrene (Concawe Report, 2013).	Thank you for your comment and the information provided. 

Comments regarding use, exposure, alternatives and risks may, when relevant, be considered at later stages of the risk management process of the SVHC dossier.

		See the corresponding embedded attachment in table 1 of Part I: 5257_PHEN.zip