1 

# TSCA N EW 

# C HEMICALS 

# R EVIEW 

# P ROGRAM



# S TANDARD 

# R EVIEW 

# R ISK 

# A SSESSMENT ON



# M EDIUM


# - C HAIN 

# C HLORINATED 

# P ARAFFINS 

# (PMN P

- 12

- 0282, P

- 12

- 0283)

AND



# L ONG


# - C HAIN 

# C HLORINATED 

# P ARAFFINS 

# (PMN P

- 12

- 0284)

This assessment was conducted under EPA’s TSCA Section 5 New Chemicals Program. EPA is assessing Medium

- chain Chlorinated Paraffin (MCCP) and Long

- Chain Chlorinated Paraffin (LCCP) chemicals as part of its New Chemicals Review program. As with all Premanufacture Notice (PMN) submissions, EPA followed the approaches, methods and statutory provisions of TSCA section 5 for the chlorinated paraffin PMNs assessments. United States Environmental Protection Agency Premanufacture Notification Number: P

- 12

- 0282

- 0284 Office of Chemical Safety and Pollution Prevention2 CONCLUSIONS

Based on its assessment of the available hazard and exposure information on P

- 12

- 0282, P

- 12

- 0283, and P

- 12

- 0284, EPA/OPPT concludes the follo wing pertaining to the manufacturing, processing and use 1 of these PMN substances :

1. Occupational Exposures : given the assumptions, data and scenarios evaluated in this assessment, there w ere low risks found for workers from either dermal or inhalation
exp osure s.

2. General Population Exposures (from environmental releases) : given the assumptions, data and scenarios eval uated in this assessment, there were low risks found to humans from environmental releases via exposure to drinking water or fish ingestion.

3. Environmental Assessment :

a. Using estimated environ mental concentrations, the P MN substance s may present an unreasonable risk following acute and chronic exposures to aquatic organisms

. i. The two exceptions are the low

- end estimates for aquatic concentrations for water releases from plastic converting use and commercial use of solve nts, paints and adhesives ( fo r the MCCP PMN P

- 12

- 0283)

b. Using available measured concentrations of MCCP and LCCP congener groups in the

e nvironment as supporting information, the PMN substances : i. A re expected to partition to sediment and may partition to soil through land application of biosolids, and ii. M ay be release d to the environment at levels at or above estimated concentrations of MC CP and LCCP congener groups that may present an unreasonable risk following acute and chronic exposures to aquatic organisms

.

4. PBT Assessment : The PMN substance s may be very persistent and very bioaccumulative

.

1 The two MCCP PMNs (P

- 12

- 0282 and 0283) have six different uses, but the three major uses are: as an additive to plastics as a flame retardant and as a secondary plasticizer [43% of total production volume] and as an a dditive to rubber as a flame retardant and plasticizer [24%], and as an extreme pressure additive to lubricants [30%]. The LCCP PMN (P

- 12

- 0284) is used exclusively as an extreme pressure additive to metalworking fluids.3 

## TABLE OF CONTENTS

TABLE OF CONTENTS

................................ ................................ ................................ ............. 3

1 INTRODUCTION

................................ ................................ ................................ .............. 6

1.1 PMN S RECEIVED

................................ ................................ ................................ ................................ ............ 6 1.2 CHEMISTRY

................................ ................................ ................................ ................................ ................... 6 1.3 USES

................................ ................................ ................................ ................................ ............................. 9

2 ENVIRONMENTAL FATE

................................ ................................ ................................ 10

2.1 ENVIRONMENTAL PERSISTENCE

................................ ................................ ................................ ................. 10 2.2 BIOCONCENTRATION AND BIOACCUMULATION

................................ ................................ ........................ 12

3 ECOLOGICAL HAZARD OV ERVIEW

................................ ................................ ................. 13

4 HUMAN HEALTH HAZA RD OVERVIEW

................................ ................................ ........... 16

4.1 MCCP HEALTH DATA REVIEW

................................ ................................ ................................ ..................... 16 4.2 LCCP HEALTH DATA REVIEW

................................ ................................ ................................ ....................... 18

5 EXPOSURE INFORMATION

................................ ................................ ............................ 19

5.1 ENVIRONMENTAL MONITORING

................................ ................................ ................................ ................ 19 5.2 MODELED ENVIRONMENTAL RELEASES

................................ ................................ ................................ ..... 22 5.3 EXPOSURE ESTIMATES

................................ ................................ ................................ ................................ 25 5.3.1 OCCUPATIONAL EXPOSURE ESTIMATES

................................ ................................ ................................ . 25 5.3.2 CONSUMER EXPOSURE

................................ ................................ ................................ ........................... 28

6 RISK ASSESSMENT

................................ ................................ ................................ ........ 29

6.1 ENVIRONMENTAL ASSESSMENT

................................ ................................ ................................ ................. 29 6.1.1 Risk Estimates Using Environmental Monitoring Data

................................ ................................ ........... 29 6.1.2 Risk Estimates Using Modeled Exposures

................................ ................................ ............................... 31 6.2 HUMAN HEALTH

................................ ................................ ................................ ................................ ......... 33 6.2.1 Workers

................................ ................................ ................................ ................................ .................. 34 6.2.2 General Population (from Environmental Releases)

................................ ................................ .............. 36 6.2.3 Consumer

................................ ................................ ................................ ................................ ............... 39

7 CONCLUSI ONS

................................ ................................ ................................ .............. 40

8 REFERENCES

................................ ................................ ................................ ................. 41

9 APPENDICES

................................ ................................ ................................ ................. 55 Appendix A ENVIRONMENTAL FATE AND BIOACCUMULATION STUDY SUMMARIES

................................ ................................ ................................ .................... 56

A

- 1 ENVIRONMENTAL PERSISTENCE

................................ ................................ ................................ ..... 56

A

- 1

- 1 Abiotic Degradation

................................ ................................ ................................ .............................. 56

A

- 1

- 1

- 1 Fate in Air

................................ ................................ ................................ ................................ ....................... 56

A

- 1

- 2 Biodegradation

................................ ................................ ................................ ................................ ...... 57

A

- 1

- 2

- 1 Fate in Wastewater Treatment

................................ ................................ ................................ ........................ 57

A

- 1

- 2

- 2 Fate in Surface and Groundwater

................................ ................................ ................................ ................... 58

A

- 1

- 2

- 3 Fate in Soil

................................ ................................ ................................ ................................ ...................... 584 A

- 2 BIOCONCENTRATION AND BIOACCUMULATION

................................ ................................ ........ 63 Appendix B ECOTOXICITY STUDY SUMMARIES

................................ ..................... 69

B

- 1 MCCP ECOTOXICITY DATA

................................ ................................ ................................ ................ 69

B

- 1

- 1 Acute Fish Toxicity

................................ ................................ ................................ ................................ .. 69

B

- 1

- 2 Acute Aquatic Invertebrate Toxicity

................................ ................................ ................................ ....... 70

B

- 1

- 3 Algae Toxicity

................................ ................................ ................................ ................................ ......... 73

B

- 1

- 4 Chronic Fish Toxicity

................................ ................................ ................................ ............................... 74

B

- 1

- 5 Chronic Aquatic Invertebrate Toxicity

................................ ................................ ................................ .... 76

B

- 1

- 6 Chronic Aquatic Sediment Invertebrate Toxicity

................................ ................................ .................... 79

B

- 1

- 7 Avian Toxicity

................................ ................................ ................................ ................................ ......... 81

B

- 1

- 8 Terrestrial Invertebrate Toxicity

................................ ................................ ................................ ............. 82

B

- 1

- 9 Terrestrial Plant Toxicity

................................ ................................ ................................ ........................ 83

B

- 1

- 10 Conclusions

................................ ................................ ................................ ................................ .......... 84

B

- 2 LCCP E COTOXICITY D ATA

................................ ................................ ................................ ....................... 85

B

- 2

- 1 Acute Fi sh Toxicity

................................ ................................ ................................ ................................ .. 85

B

- 2

- 2 Acute Aquatic Invertebrate Toxicity

................................ ................................ ................................ ....... 86

B

- 2

- 3 Aquatic Plant Toxicity

................................ ................................ ................................ ............................ 88

B

- 2

- 4 Chronic Fish Toxicity

................................ ................................ ................................ ............................... 88

B

- 2

- 5 Chronic Aquatic Invertebrate Toxicity

................................ ................................ ................................ .... 88

B

- 2

- 6 Chronic Aquatic Sediment Invertebrate Toxicity

................................ ................................ .................... 92

B

- 2

- 7 Avian Toxicity

................................ ................................ ................................ ................................ ......... 92

B

- 2

- 8 Terrestrial Invertebrate Toxicity

................................ ................................ ................................ ............. 92

B

- 2

- 9 Terrestrial Plant Toxicity

................................ ................................ ................................ ........................ 92

B

- 2

- 10 Conclusions

................................ ................................ ................................ ................................ .......... 92 Appendix C HUMAN HEALTH HAZARD STUDY SUMMARIES

............................... 95

C

- 1 MCCP H EALTH D ATA R EVIEW

................................ ................................ ................................ ................. 95

C

- 1

- 1 Metabolism

................................ ................................ ................................ ................................ ............ 96

C

- 1

- 2 Acute Toxicity

................................ ................................ ................................ ................................ ......... 96

C

- 1

- 3 Irritation and Sensitization

................................ ................................ ................................ ..................... 96

C

- 1

- 4 Repeated

- dose Toxicity

................................ ................................ ................................ .......................... 96

C

- 1

- 5 Genotoxicity

................................ ................................ ................................ ................................ ........... 98

C

- 1

- 6 Carcinogenicity

................................ ................................ ................................ ................................ ....... 98

C

- 1

- 7 Developmental Reproductive Toxicity Review

................................ ................................ ..................... 102

C

- 2 LCCP H EALTH D ATA R EVIEW

................................ ................................ ................................ ................ 108

C

- 2

- 1 Metabolism

................................ ................................ ................................ ................................ .......... 109

C

- 2

- 2 Acute Toxicity

................................ ................................ ................................ ................................ ....... 109

C

- 2

- 3 Irritation and Sensitization

................................ ................................ ................................ ................... 1095 C

- 2

- 4 Repeated

- dose Toxicity

................................ ................................ ................................ ........................ 109

C

- 2

- 5 Genotoxicity

................................ ................................ ................................ ................................ ......... 110

C

- 2

- 6 Carcinogenicity

................................ ................................ ................................ ................................ ..... 110

C

- 2

- 7 Developmental Reproductive Toxicity Review

................................ ................................ ..................... 110 Appendix D ENVIRONMENTAL MONITORING DATA

................................ ........... 114

D

- 1 MCCP MONITORING DATA

................................ ................................ ................................ ..................... 114

D

- 1

- 1 Surface Water

................................ ................................ ................................ ................................ ...... 114

D

- 1

- 2 Sediment

................................ ................................ ................................ ................................ .............. 116

D

- 1

- 3 Biosolids and Soil

................................ ................................ ................................ ................................ . 125

D

- 1

- 4 Biota

................................ ................................ ................................ ................................ .................... 125

LCCP MONITORING DATA

................................ ................................ ................................ ............................... 128 Appendix E ENGINEERING (ChemSTEER) REPORTS ON P

- 12

- 0282, P

- 12

- 0283 and P

- 12

- 0284 129 Appendix F EXPOSURE SCENARIO ESTIMATES

................................ .................... 140 Appendix G SUPPLEMENTAL INFORMATION

................................ ........................ 1456 

## 1 INTRODUCTION

1.1 PMNs R ECEIVED Dover submitted three Premanufacture Notices (PMN s ) ident ified by EPA/ OPPT as either medium

- chain chlorinated paraffins (MCCPs ; P

- 12

- 0282 and P

- 12

- 0283 ) of varying chain lengths with the formula C x H (2x

- y+2) Cl y and x equaling 14 to 17 and y equaling 6 to > 24 or long

ch ain chlorinated paraffins (LCCPs; P

- 12

- 0284 ) of varying chain lengths with the formula of C x H (2x

- y+2) Cl y and x equaling 18 to 20 and y eq ualing 6 to > 24. Table 1 lists basic information Dover supplied on these three PMNs (two MCCPs and one LCCP).

Table 1: Identification, Production Volume and Use for P

- 12

- 0282 to 0284

PMN Chemical Name 1 st Year Production Volume (kg ) Uses Log K ow Water S olubility

P

- 12

- 0282 Alkanes, C 14

- 16 , chloro (MCCP) 11,749,910 See below for generic uses 4.70 (E) < 0.03 mg/L (E)

P

- 12

- 0283 Tetradecane, chloro (MCCP) 3,779,281 See below for generic uses 4.70 (E) < 0.03 mg/L (E)

P

- 12

- 0284 Octadecane, chloro (LCCP) 45,500 100 % as a lubricant in metal working fluid 8.00 ( E) < 0.005 mg/L (E) E = E stimated

T hough the specific PMNs in this application include MCCPs and LCCPs, this standard review presents data and information on short

- chain chlorinated paraffins (SCCPs) and on very long

chain chlorinated paraffins (vLCCPs ) analogs. The continuum of carbon chain length and degree/ weight percent chlorination (wt% Cl) in all of the chlorinated paraffins (CPs) and the

r elationship among them needs to be kept in mind.



### 1.2 CHEMISTRY
Sho wn below are the structures and chlorine content of each of the three PMNs. P

- 12

- 0282 contains chain lengths of between C 14

- 17 and both P

- 12

- 0283 and P

- 12

- 0284 represent C 14 and C 18 chain length congener groups, respectively.7 P

- 12

- 0282 : The average mo lecular formula would range from C 14 H 27 Cl 3 (at ~30% chlorination) to C 16 H 18 Cl 1 high wt (at ~ 70% chlorination).

P

- 12

- 0283: The average molecular formula would range from C 14 H 27 Cl 3 (at ~30% chlorination) to C 14 H 16 Cl 14 ( at ~ 75% chlorination )

.

P

- 12

- 0284 : The chlorinated products would range from C 18 H 35 Cl 3 (at ~30% chlorination) to C 18 H 20 Cl 18 (at ~70% chlorination).

CPs have an unknown or variable composition (classified as UVCB 2 compounds for TSCA Inventory purposes) of polychlorinated n

- alkanes. The carbon chain length usually varies between 10 and 30 carbon atoms and the degree of chlorination can vary between 30 and 75 wt %. EPA/OPPT subdivides CPs according to their carbon ch ain length into the following categories:

1. SCCPs, C 10

- 13 2. MCCPs, C 14

- 17

2 UVCB are chemical substances whos e composition is Unknown or Variable compositions, Complex reaction products and Biological materials.8 3. LCCPs, C 18

- 20 4. Very long chain CPs (vLCCPs, C >20 )

SCCPs and MCCPs exist as liquids at standard temperature and pressure. CPs with a carbon chain length > 18 are subdivided based on their physical state, which is a function of chain length and chlorine content. The LCCPs and vLCCPs up to 70 wt % Cl are typically liquids (40

– 55 wt % Cl) while above 70 wt % CL they are waxy solids.
CP products contain a v ariety of carbon chain lengths that have been chlorinated to different degrees ( i.e. , variation in the number and position of the chlorine atoms on the carbon chain). The individual isomer content of commercial CPs is rarely identified because the number o f possible individual congener group is extremely large. Consequently, the physicochemical properties of CPs vary by carbon chain length and chlorine content. Increased molecular weight correlates to higher melting and boiling points, lower vapor p ressures and water solubilities

, and greater Log K OW ( logarithm of octanol:water partition coefficient ). All three PMNs in this assessment are liquids with chlorination levels of between 30 to 75%. EPA/OPPT used the physicochemical properties listed in Table 2 for informing its evaluation of P

- 12

- 0282, 0283 and 0284.

Table 2: Summary of Physiochemical Information a ,b

wt% Cl Melting Point Boiling Point Vapor Pressure Water Solubility Log K OW

MCCPs > 40 < 25 °C (pour point) > 200 °C (dec) < 0.036 Pa at 20 °C 27 μg/L at 20 °C > 5.5 (measured) 8.30 (estimated) c,d

LCCPs > 40 < 25 °C (pour point) > 200 °C (dec) < 2.7 × 10

- 4 Pa at 20 °C 5 μg/L at 20 °C > 10

a Source: EURAR (ECB, 2008)

b Because most CP products are liquids and the CPs begin decomposing at 200 °C ( via loss of HCl), melting point and boiling points are considered less important in characterizing hazard and risk.

c Value calculated using the KOWWIN Program (v1.68) available in EPA’s Estimation Programs Interface (EPI) Suite TM. This estimate was generated using a representative MCCP ( i.e. , C 14 H 24 Cl 6 , 52 wt % Cl) with the following SMILES notation: CCC(Cl)CC(Cl)CCCl)CC(Cl)CC(Cl)CC(Cl)C.

d The EURAR (ECB, 2008) cited Renberg’s li quid chromatography to measure a log K OW between 5.5 and 8.2 and then chose to use log K OW = 7 as a representative log K OW for MCCP 45

– 52 wt % Cl.

Analytical challenges exist with evaluating CPs due to the sheer number of congener groups 3 that may be pr esent in CP products. The existence of multiple chain length s in the UVCBs such as P

- 12

- 0282 requires the use of analytical methods that separate congener groups based on retention time in a column and mass spectrum of the respective peaks. Several lines o f evidence support the use of a representative SCCP or MCCP product as a surrogate for congener groups present in MCCP or LCCP commercial products, respectively.

3 For this report, congener groups is used to recognize the existence of different chain lengths and degrees of chlorination that could be present in any given CP product.9 Hüttig (2006) and Hüttig and Oehme (2006) reported that most commercial MCCP products were > 60 wt % of the C 14 chain

- length congener groups and the C 15 chain

- length congener groups comprised the majority of the remaining
∼ 40 wt %. The authors found < 15 wt % C 16 chain

- length congener groups present in most commercial MCCP samples and little to no C 17 chain

- length congener groups. Additional studies have reported that the C 14 and C 15 chain

- length congener groups are the predominant MCCPs present in environmental m edia and in human breast milk (Bayen et al., 2006; Chen et al., 20 11; Reth et al., 2006; Wang et al., 2013) . Some variation is possible in commercial products where th e C 14 and C 15 chain

- length congener groups may not be the predominant congener groups in a specific MCCP product; however, even in these products, the C 14 and C 15 chain

- length congener groups may serve as reasonable worse case surrogates for the C 16

- 17 cha in

- length congener groups, due to their greater bioavailability and mobility in e nvironmental media (ECB, 2005).

These analyses, in conjunction with measured or estimated physicochemical and environmental fate properties, allow for the reasonable use of associating commercial products with CP levels found in environmental media and biota. For example, the experimental observation that MCCPs (C 14

– 17 ) are abundant in sediment has been explained using known water solubility and vapor pressure values in con junction with predicted degrada tion pathways (de Boer, 2010). EPA/OPPT determined that the toxic end points of interest were measured for only one CP commercial product ( i.e. , Cereclor S52 ® ). Therefore, this approach of relating one commercial product with other commercial products is critical for attributing the hazard characterization to CPs of different sources. Thus, available information (hazard and environmental monitoring data) on one commercial product ( i.e. , Cereclor S52 ® ) is used as a representati ve CP for this assessment

. Furthermore, experimental data on SCCPs show that these products are more toxic than the longer chain CPs ( e.g. , MCCPs). Therefore, when endpoint specific data wer e lacking for the PMNs

, EPA/OPPT used measured data from SCC Ps as a surrogate for potential haza rds and risks for the PMNs.



### 1.3 USES Dover reported up to six uses, including:

1) 43% as a flame retardant/plasticizer in polymers (no consumer use). Typical concentration in the plastic is 12

- 30%. 2) 30% as a lubricant in MWFs (no consumer use). The notification substance is blended into the MWFs at 5

- 10%. 3) 24% as a flame retardant/p lasticizer in rubber. Typical concentration in the rubber is 1

- 1 0%, with a median value of 5.5%. 4) 2% as lubricants in automotives. 20% co nsumer uses. The notification substance is blended into the lubr icant at 4%.

. 5) 0.5% as a plasticizer in solvent based paints, adhesives and sealants. The paints are specialty paints for applications where weather resistant coatings are needed, such as

s teel construction, industrial flooring, road marking, and swimming pools. The notification substance is ble nded into the paints at 5

- 15%. C ons umer use as sealants in windows is 0.1%

.10 6) 0.5% a flame reta rdant/waterproofer to textiles (no consumer use)

. The notification substance products are added at 5% of the textile coating finishing solution.

Table 3 p rovides the listed trade n ames , range of chlorinateion and percent PMN substance in the final use products.

Table 3: Trade Names, Percent Chlorination an d PMN Concentration in Final Use Product s

PMN Trade Names Percent Chlorination Percent PMN in final Product

P12

- 0282 (MCCP) Variations of PAROIL, CHLOROWAX, KLORO. Also

– P152, AO52, OA50

- 60, X1045, X1052 30

- 70 5

- 30

P

- 12

- 0283

(MCCP) Variations of PAROIL, DO, P358, Doverguard. Also

– P10NR, AO63NR, Entech Blend, Militec

- 1, US Pack Blend, HWDO8110 30

- 75 5

- 30

P

- 12

- 0284

(LCCP) DO 8707, DO 8110 30

- 70 5

- 10



## 2 ENVIRONMENTAL FATE

EPA/OPPT reviewed available information on the environmental fate of MCCPs and LCCPs in different environmental compartments and the properties that control transport (summarized in Appendix A). In addition, EPA/OPPT reviewed assessments performed by Canad a (EC, 2008a) and the EU (EA, 2009; ECB, 2005) to inform its assessment.



### 2.1 E NVIRONMENTAL PERSISTENCE
A biotic studies have shown that MCCPs and LCCPs are stable to hydrolysis and to direct photolysis in water and air. In laboratory studies using hydrocarbon solvents, CPs were shown to poorly absorb ultraviolet (UV) light and no direct photodegradation was o bserved. The atmospheric half

- life for MCCPs and LCCPs has been estimated at 1

- 2 days (EA, 2009; E CB, 2005) , based on estimated values for the second order rate constant for reaction with atmospheric hydroxyl radicals for MCCPs (40

- 56 wt% Cl) and LCCPs (42

- 54 wt% Cl) (EA, 2009; ECB, 2005). The persistence of MCCPs increases with carbon chain length and higher chlorine content (EA, 2009; ECB, 2005).

Existing biotic degradation data suggest there are a number of microbial species that are capable of degrading shorter chain, lower chlorinated MCCP congeners. Longer and higher chlorinated chemicals als o may be degraded, but at much slower rates (Allpress and Gowland, 1999; Muir, 2010; Omori et al., 1987) . The results from laboratory studies of microbial metabolism, using both isolated species and mixed cultures of acclimated microbes, show that MCCPs and LCCPs may be degraded by direct metabolism or co

- metabolism by some microbes and microbial consortia in soil, wastewater treatment systems, sediment, and other environmental media. Overall, the existing studies suggest that with microbial degradation, dechlorination and carbon11 chain cleavage may be p ossible in some media (see Table A

- 1); however the degree of degradation is generally low (Allpress and Gowland, 1999; Muir, 2010; Omori et al., 1987)

. In general, MCCP and LCCP congeners with longer chain lengths and higher degrees of chlorination are expected to be highly persistent in some environmental compartments. In contrast, shorter and less chl orinated congeners are likely to degrade rapidly, especially in aerobic environments. Because persistence increases with chain length, LCCPs are generally expected to be more persistent than MCCPs with comparable degrees of chlorination (EA, 2009; ECB, 200 5).

Based on the review of available literature and studies submitted by various manufacturers, including confidential business information (CBI) not publicly available , EPA/OPPT’s conclusions regarding environmental persistence of MCCPs and LCCPs are co nsistent with those provided by Canada and the EU.

Canad a’s assessment (EC, 2008a) states :

“Information on physical properties of MCCPs, and especially LCCPs, is limited. Values used in this assessment are based on extrapolations mainly from SCCPs or QSARs . The analysis of SCCPs and MCCPs in sediment cores and associated calculations provide strong evidence for the persistence of these substances in the environment. Even though there are no data for persistence of LCCPs in sediment, based on biodegradation data which indicate increasing stability with increasing carbon chain length, it is reasonable to conclude that LCCPs are persistent in sediment. ”

The EU assessment on MCCPs (ECB, 2005) states : 4

“No standard ready or inherent biodegradation tests results are available for medium

chain chlorinated paraffins. From the available information, medium

- chain chlorinated paraffins can be co nsidered to be not biodegradable in such test systems and so a biodegradation rate MCCPs of 0 day

- 1 is used in the risk assessment.

There is evidence that some microorganisms may be capable of degrading MCCPs in the environment in acclimated or co

- metabo lic systems. The potential for biodegradation appears to increase with decreasing chlorine content. However, it is not possible from the available data to derive rate constants for biodegradation in soil, surface water and sediment systems. As a worst case approach, no biodegradation will be assumed in these media in the PEC calculations.

Hydrolysis is not expected to be a significant degradation process for medium

- chain chlorinated paraffins in the environment. An atmospheric half

- life of 1

- 2 days is esti mated for reaction with hydroxyl radicals. A value for the rate constant for the reaction (k OH ) of 8 x 10

- 12 cm 3 molecule

- 1 s - 1 is used for the environmental modelling in the ris k assessment.”

4 Note, since the EU issued its assessment in 2005, standard inherent biodegradation studies were performed and are summarized in Appendix A.12 The UK assessment on LCCPs 5 (EA, 2009) concluded :

“Based on the laboratory studies and other data available, LCCPs are unlikely to be readily or inherently biodegradable. Although there is some evidence that they may biodegrade in the environment, it is thought likely that this process will be sufficiently slow that LCCPs meet the P or vP (very persistent) criteria.”

EPA/OPPT generally concurs w ith these characterizations. I n the absence of information on specific congener groups and data for MCCP or LCCP products, EPA/OPPT concludes that at least some congener groups present in both MCCP and LCCP products are persistent to very persistent; with estimated half

- lives in air exceeding two days and estimated half

- lives in water

, sediment, and soil exceeding two months (60 days) (ECB, 2005; EA, 2009).



### 2.2 B IOCONCENTRATION AND BIOACCUMULATION Recent reviews of the potential for MCCPs and LCCPs to bioaccumulate have shown that, while data are limited

, some congener groups are bioaccumulative or very bioaccumulative (EC, 2008a; ECB, 2005; Houde et al., 2008; Thompson and Vaughan, 2014) . A summary of studies reviewed by EPA/OPPT is provided in Appendix A.

Based o n EPA/OPPT’s review of existing studies (Bengtsson et al., 1979; CPC, 1980, 1983a, 1983b; Fisk et al., 1999; Fisk et al., 1998; Houde et al., 2008; Madeley and Maddock, 1983a, 1983b; Madeley and Thompson, 1983; Renberg et al., 1986; Thompson et al., 2000)

, EPA/OPPT concludes that the bioconcentration varies with the chain lengths and degree of chlorination within the CP mixture and species evaluated. Shorter and les s chlorinated chemicals are readily taken up by organisms but also may be excreted or degraded after absorption (Arnot, 2013)

. Longer and more highly chlorinated chemicals are typically not absorbed across cellular membranes and are not accumulated in tiss ues. Some MCCP chemicals with intermediate chain length and chlorination may be absorbed and retained. The available evidence for MCCP and LCCP congener groups with intermediate chain lengths and chlorination suggests that some may have bioconcentration fa ctors (BCFs) or bioaccumulation factors (BAFs) greater than 1000 or 5000 (EC, 2008a; ECB, 2005, 2008) . This suggests that some congener groups in MCCP and LCCP products may be bi oaccumulative or very bioaccumulative.

The Canadian assessment (EC, 2008a) on MCCPs and LCCPs states :

“On the basis of the available information, and in particular the field BAF estimates, it is concluded that MCCPs are bioaccumulative substances...”

“On the basis of the available information, and in particular the BAF model and empirical BMF estimates, it is concluded that C18

– 20 liquid LCCPs are bioaccumulative substances...” [page 27] and “...While there is a lack of empirical bioaccumulation data for L CCPs, the modelling results provided by the Modified Gobas BAF Model

- which

5 Note, the UK assessment evaluated the following CPs under the term LCCP: C 18

- 20 liquid products, C >20 liquid products, and C >20 solid products.13 suggest that of all the LCCPs congeners only liquid C18

- 20 LCCPs have significant bioaccumulation potential

-- are considered credible.”

The UK assessment (EA, 2009) on LCCPs states :

“The available data for LCCPs do show that uptake into fish from food occurs in the laboratory, and that this uptake can be significant in some case s. The degree of uptake appears to be highest for the C18

– 20 liquid chlorinated paraffins, but uptake of C>20 liquid chlorinated paraffins has also been demonstrated. The uptake of the highly chlorinated C>20 solid chlorinated paraffins from food appears t o be minimal.”

EPA/OPPT generally concurs with these characterizations . I n the absence of information on specific congener groups and data for MCCP or LCCP products, EPA/OPPT concludes that at least some congener groups present in both MCCP and LCC P prod ucts are bioaccumulative to very bioaccumulative base d on multiple lines of evidence including: Log K OW values , modeled BCF s , laboratory

- mea sured BCFs, field

- measured BAFs, field

- measured biomagnification factors (BMFs), laboratory

- measured biota

- sedime nt accumulation factors (BSAFs) and the presence of MCCPs and LCCPs in human and wildlife biota.



## 3 ECO LOGICAL HAZARD OVERVIEW The available ecotoxicity data on MCCPs and LCCPs are summarized in Appendix B, along with the criteria EPA/OPPT used for identifying the highest quality studies. Ecotoxicity studies for MCCPs have been conducted in fish, aquatic invertebrates and plants, sediment and soil invertebrates, and terrestrial plants and invertebrates. Though no avian reproduction studies were available on MCC Ps, a high quality study was available on an SCCP product (C 10

- 12 , 58 wt % Cl) with similar physicochemical properties to MCCPs and was used for informing EPA/OPPT ’s hazard evaluation (ECB, 2000)

. For LCCPs, ecotoxicity studie s were only identified for aquatic invertebrates and vertebrates. No data were available on sediment

- dwelling or terrestrial organisms. Overall, the available data on LCCPs were of low quality ; therefore, the EPA/OPPT used data on MCCPs to inform its hazar d evaluation of LCCPs. This decision was considered a reasonable worst

- case scenario

, because although P

- 12

- 0284 does not contain any C 17 congener groups, it is reasonable to assume that C 17 and C 18 congener groups have similar ecotoxicological effects when the wt. % Cl is comparable. EPA/OPPT concludes that the studies summarized in Table 4 were the highest quality for assessing potential h azards in the aquatic, sediment and terrestrial compartments

.14 Table 4: Summary of Aquatic, Sediment and Terrestrial Ecotoxicity Data for MCCPs and LCCPs

Test Substance Test Organism (Species) Test Guideline; Study type End

point Value a Reference

Aquatic Invertebrates

Cereclor S

- 52 (52

wt% Cl, C 14

- 17 ) Water flea ( Daphnia magna ) OECD 202, 1984; Acute immobilization test EC 50 0.0059 CPA (1996)

Cereclor S

- 52 (52

wt% Cl, C 14

- 17 ) Water flea ( Daphnia magna ) OECD 202

- Part II, 1984; Reproduction test NOEC LOEC MATC 0. 01 0.018 0.013 Thompson, Williams et al. (1997)
Sediment

- Dwelling Invertebrates

Cereclor S

- 52 ( 52

wt% Cl, C 14

- 17 ) Amphipod ( Hyalella azteca ) OECD 218

- Draft, 2001; 28

- day prolonged sediment toxicity study NOEC LOEC MATC 130 270 187 Thompson et al. (2002)

Terrestrial Invertebrates

Cereclor S

- 52 (52

w t% Cl, C 14

- 17 ) Earthworm ( Eisenia fetida ) OECD Guideline

- Draft, 2000; 28

- day reproductive toxicity test NOEC LOEC MATC 79 280 149 Thom pson et al. (2001d)

Terrestrial Vertebrates

Commercial CP ( 58 wt% Cl, C 10

- 12 ) Mallard duck ( Anas platyrhynchos ) EPA 560/6

- 82

- 002; 22

- week reproduction test NOEC LOEC 168 1000 ECB (2000)

a Units are mg/L for aquatic invertebrates, mg/kg dry weight (dw) sediment for sediment

- dwelling invertebrates; mg/kg dw soil for earthworm study; and mg/kg diet for the duck study.

Using the concentrations in the “value” column in Table 4 to represent hazard, EPA/OPPT derived concentrations of concern (COCs) by applying assessment factors of five or ten for acute or chronic exposures, respectively, which account for laboratory variability and represent s species sensitivity distributions (following USEPA, 2012). The COC s derived for aquatic

- , sediment

- , and terrestrial

- dwelling organisms are explained below and are summarized in Table 5

. The most reliable and acceptable toxicity studies to aquatic organisms for MCCPs and LCCPs, are from the CPA (1996) study for acute t oxicity and the Thompson et al. (1997) study for chronic toxicity:

• Acute COC: The 48

- hour EC 50 value 0.0059 mg/L is divided by an assessment factor of 5 to yield an acute concentration of concern (COC) of 0.00118 mg/L, or 0.001 mg/L, or 1 μg/L (1 ppb). Aquatic Acute COC = 1 ppb

.

• Chronic COC: The chronic value 0.013 mg/L is divided by an assessment factor of 10 to yield 0.0013 mg/L or 1.3 μg/L or 1.3 ppb. Aquatic Chronic COC = 1 ppb

.15 The most reliable and acceptable value for the acute toxicity to aquatic sediment invertebrate organisms is based on the MCCP material from the Thompson et al. (2002) 28

- d study. The 28

- d sediment invertebrate GMATC value of 187 mg/kg dry wt sediment is us ed to assess hazard. Using methods in USEPA (2012):

• Acute COC: The chronic value 187 mg/kg dry wt. is multiplied by an acute to chronic ratio for invertebrates (10) to yield 1,870 mg/kg dry wt. This value is then divided by an assessment factor of 5 to y ield 374 mg/kg dry wt. Aquatic Sediment Acute COC = 374 mg/kg dry wt sediment.

• Chronic COC: The 28

- d sediment invertebrate GMATC of 187 mg/kg dry wt sediment is divided by an assessment factor of 10 to yield 18.7 mg/kg dry wt sediment. Aquatic Sediment C hronic COC = 18.7 mg/kg dry wt sediment.

The most reliable and acceptable value for acute toxicity to terrestrial invertebrates is based on the MCCP material from the Thompson et al. (2001a) study. The 28

- d terrestrial invertebrate GMATC value of 149 mg/k g dry wt soil from this study will be used. Using methods in USEPA (2012):

• Acute COC: To calculate an acute concern concentration from the chronic value the value 149 mg/kg dry wt, is multiplied by an acute to chronic ratio for invertebrates (10) to yiel d 1,490 mg/kg dry wt. This value is then divided by an assessment factor of 5 to yield 298 mg/kg dry wt. Terrestrial Invertebrate Acute COC = 298 mg/kg dry wt.

• Chronic COC: The 28

- d terrestrial invertebrate GMATC of 149 mg/kg dry wt is divided by an asses sment factor of 10 to yield 14.9 mg/kg dry wt. Terrestrial Invertebrate Chronic COC = 14.9 mg/kg dry wt.

The most reliable and acceptable value for acute toxicity to terrestrial vertebrates is based on the MCCP material from the ECB (2001) study. The 22

- week terrestrial vertebrate NOEC value of 168 mg/kg dry wt soil from this study will be used. Using methods in USEPA (2012):

• Acute COC: To calculate an acute concern concentration from the chronic value the value 168 mg/kg diet is multiplied by an acute to chronic ratio for invertebrates (10) to yield 1,680 mg/kg diet. This value is then divided by an assessment factor of 5 to yield 336 mg/kg diet. Terrestrial Vertebrate Acute COC = 336 mg/kg diet.

• Chronic COC: The 22

- week terrestrial vertebrate NOEC of 168 mg/kg diet is divided by an assessment factor of 10 to yield 16.8 mg/kg diet. Terrestrial Vertebrate Chronic COC = 16.8 mg/kg diet.16 Table 5: COCs for E nvironme ntal Toxicity of MCCPs and LCCPs

.

Compartment Test organism Endpoint Value Assessment factor COC

Surface water Water flea EC 50 0.0059 mg/L 5 0.001 mg/L

21

- day MATC 0.013 mg/L 10 0.001 mg/L

Sediment Amphipod MATC 187 mg/kg dw 10 18.7 mg/kg dry wt. sediment

Terrestrial Earthworm 28

- day MATC 149 mg/kg dw 10 14.9 mg/kg dry wt. soil

Mallard duck 22

- week NOEC 168 mg/kg diet 10 16.8 mg/kg diet



## 4 HUMAN HEALTH HAZARD OVERVIEW A summary of EPA/OPPT ’s evaluations on MCCPs and LCCPs is provided in sections 4.1 and 4.2, respectively; individual study reviews are provided in Appendix C.



### 4.1 MCCP H EALTH DATA REVIEW There is no information on inhalation absorption of MCCPs in humans or in animals. Based on their low vapor pressure and low water solubility, absorption following inhalation or dermal exposure is expected to be limited ; previous ev aluations concluded that absorption by the inhalation and dermal routes of exposure will not exceed 50 or one percent, respectively (ECB 2005; EA 2009). Some MCCPs demonstrated moderate absorption and metabolism following oral exposure in animals. In gener al, absorption and metabolism are related to their carbon chain length and degree of chlorination; the longer the carbon chain length and the higher the degree of chlorination, the less absorption and metabolism.

No information is available on the toxici ty of MCCPs in humans; however, the toxicology of these compounds has been evaluated in experimental animals. Studies in rats and rabbits have shown that MCCPs only caused slight skin irritation and have low eye irritation potential. No evidence of skin se nsitization was found when tested in guinea pigs. The liver, kidney and thyroid are the target organs of MCCPs in oral repeated dose studies in experimental animals (see Table C

- 1 in Appendix C). MCCPs induced increased liver weight, enzyme activity, and h istopathological changes at high dose levels. Some of these hepatic effects are likely related to an increase in metabolic demand as an adaptive response, as well as to peroxisome proliferation, which are considered of limited toxicological significance to humans. However, liver necrosis was observed in a 90

- day study in rats at 360 mg/kg

- bw/day; this effect is considered relevant to humans. The reported effects in the kidney may have been produced by the parent compound or from metabolites. Mechanistic dat a cannot totally rule out that some kidney effects are relevant to humans. From the data available, a LOAEL of 625 mg/kg

- bw/day based on histopathological changes in the kidneys of female rats is identified in a 90

- day toxicity study, and a NOAEL of 23 mg/kg

- bw/day based on increased kidney weight at 222 mg/kg

- bw/day is identified from another 90

- day study in rats (CXR, 2005) . Repeated dose studies in rats reported some changes in histopathology and hormone levels of the thyroid. However, it may be conclude d based on an17 evaluation of the mechanistic data that the thyroid effects observed in rats is of little relevance to chronic toxicity in humans.

There is no information on the carcinogenicity of MCCPs; however, carcinogenicity studies on a SCCP and a vLCC P are available. These studies, along with the genotoxicity data on MCCPs, were used to inform the carcinogenic potential of MCCPs. When administered by gavage, a SCCP (C 12 , 60 wt% Cl) caused increased incidences of liver tumors in male and female rats, ki dney tumors in male rats, and thyroid tumors in female rats. However, based on mechanistic considerations, these tumors are considered to be of little or no relevance to humans (details in ECB, 2008 and in Appendix C) . An increased incidence of malignant l ymphoma in male mice was reported at the highest dose of 5,000 mg/kg

- bw/day in carcinogenicity studies of a LCCP (C 23 , 43 wt% Cl) in male and female rats and mice. However, malignant lymphoma is one of the more variable tumors in mice and has a viral origi n in many cases. No increased incidence of malignant lymphoma was observed in the carcinogenicity study on an SCCP. Further, MCCPs

are non

- genotoxic. Therefore, it may be concluded that MCCPs are unlikely to pose a carcinogenic hazard to humans.

A series of range

- finding and definitive prenatal developmental and reproductive toxicity studies were conducted in rats and rabbits with MCCPs. These studies were conducted between 1981 and 1986. They appear to be valid toxicity studies, conducted according to the standard methodologies available at the time.

In several prenatal developmental toxicity studies with MCCPs conducted via gavage, no signs of maternal toxicity were seen at doses as high as 500 mg/kg

- bw/day in rats and 100 mg/kg

bw/day in rabbits. Likew ise, no signs of developmental toxicity were observed at doses as high as 5000 mg/kg

- bw/day in rats and 100 mg/kg

- bw/day in rabbits.

Two reproductive toxicity studies with MCCPs in rats have been conducted. A one

- generation reproductive toxicity range

- fi nding study showed that administration of approximately 100 and 400 mg/kg

- bw/day MCCPs via the diet had no effect on fertility or other reproductive parameters; however, internal hemorrhaging and deaths in pups were observed beginning from 74 mg/kg

- bw/day (1000 ppm) up to approximately 400 mg/kg

- bw/day (6250 ppm). These effects in the pups were not seen in a more recent definitive one

- generation reproductive toxicity study with exposure to MCCPs for 11

- 12 weeks to doses as high as 100 mg/kg

- bw/day (1200 ppm). Internal hemorrhaging was not seen in the adult animals in either of these studies at d oses as high as 400 mg/kg

- bw/day (6250 ppm), or in another study in non

- pregnant female rats repeatedly exposed to doses as high as 1000 mg/kg

- bw/day. However, when dams were exposed to approximately 500 mg/kg

- bw/day (6250 ppm) MCCPs during cohabitation, g estation, and lactation, signs of hemorrhaging were observed in dams that died at the time of parturition. Taken together, the results of these studies suggest that newborns during lactation and pregnant females at the time of parturition are a potentially sensitive subpopulation with a possible LOAEL for interna l hemorrhaging and deaths in pup s at an oral dose of 74/mg/kg

- bw/day.

Additional studies with MCCPs have been conducted in an effort to clarify the possible causes of the hemorrhaging in the pups. One (single

- dose; 6250 ppm or 538 mg/kg

- bw/day) study showed maternal death during parturition due to low levels of vitamin K and related hemorrhaging,18 suggesting that the act of parturition places dams at higher risk. It was concluded in data from this s tudy and a cross

- fostering study that the fetus relies on clotting factors via mother’s milk and severe deficiencies in vitamin K levels and related clotting factors in the pups results in hemorrhaging.

No guideline developmental neurotoxicity studies on MCCPs were located. It is not clear if any developmental neurotoxicity endpoints were measured in the available prenatal developmental/reproductive toxicity studies; none were explicitly stated. The only information available regarding behavior during dev elopment is from cage

- side observations in pups through lactation day 21. In these cases, no dose

- related differences were reported in F 1 post

- weaning
appearance or cage

- side behaviors. While thyroid hormone induced effects were observed in adults, no data exist for developmental studies. Current studies do not evaluate developmental neurotoxicity following perinatal exposures.

In this assessment, the lowest NOAEL (90

- day value of 23 mg/kg/d from the rat study described above; CXR, 2005) will be used to as sess occupational and non

- occupational ( i.e. , general population) risk of MCCPs

.


### 4.2 LCCP H EALTH DATA REVIEW There is no information on inhalation absorption of v/ LCCPs in humans or in animals. Based on their low vapor pressure and water solubility, absorption following inhalation or dermal exposure is expected to be limited. Some absorption and metabolism following oral exposure are possible for v/ LCCPs with shorte r carbon chain length and lower degree of chlorination.

No information is available on the toxicity of v/ LCCPs in humans. Acute oral toxicity data in animals show that v/ LCCPs are of very low acute toxicity. Studies in animals have shown that

some v/ LCCPs may have the potential to cause slight skin irritation and sensitization but no eye irritation potential. The liver is the main target organ of v/ LCCPs in repeated dose studies in experimental animals. Inflammatory and necrotic changes of the liver w ere observed in rats exposed to a C 20

- 30 v/ LCCP with 43 wt% Cl at dose levels of 100 mg/kg

- bw and above. For another v / LCCP with C 20

- 26 70 wt% Cl, effects in the liver occurred at a very high exposure level of 3,750 mg/kg

- bw/day; the NOAEL was 900 mg/kg

- bw /day.

An increased incidence of malignant lymphoma in male mice was reported at the highest dose of 5,000 mg/kg

- bw/day when tested using a C 23 v LCCP with 43 wt% Cl in carcinogenicity studies in male and female rats and mice. However, malignant lymphoma i s one of the more variable tumors in mice and has a viral origin in many cases. Data on the analogous SCCPs have shown no increase in the incidence of malignant lymphoma in a carcinogenicity study of SCCPs. v/ LCCPs are non

- genotoxic and they are not expect ed to pose a carcinogenic hazard to humans.

Based on the LOAEL (100 mg/kg

- bw /day ) of the liver effects in female rats of repeated dose studies, Health Canada calculated a tolerable daily intake (TDI) of 71 μ g/kg

- bw/day with LCCPs. Using upper bounding in take estimates ranging from 0.007 μ g/kg

- bw/day for 60+ age group to 0.024 μ g/kg

- bw/day for 0.5 years age group, Environment Canada determined that the exposure levels are 10,000 and 3,000 times lower, respectively, than the TDI.19 The National Research Co uncil (NRC, 2000) reviewed the toxicological risks of selected flame retardant s , including a v LCCP containing C 24 with 70 wt% Cl. Based on the NOAEL of 900 mg/kg

- bw/day (liver toxicity), the NRC derived an RfD of 0.3 mg/kg

- bw/day. Using this RfD and the wo rst case average daily exposure to be 0.16 mg/kg/day, NRC concluded: “ LCCP do not pose a noncancer risk when incorporated into residential furniture at the estimated application levels .” Further, it was concluded that: “ LCCP are not likely to be a human carcinogen and derivation of a cancer potency factor is unnecessary. ”

In this assessment, the LOAEL of 100 mg/kg

- bw /d ay from the 90

- day study described above) will be used to assess occupational and non

- occupational ( i.e. , general population) risk of LCCP s

.


## 5 EXPOSURE INFORMATION

EPA/OPPT used the information in this section and standard PMN approaches to estimate potential worker exposures from activities associated with manufacturin g, processing and use of P

- 12

- 0282, 0283 and 0284. Environmental releases from these activities were also estimated for use in assessing risk to both human health (general population) and the environment (aquatic organisms). In addition, EPA/OPPT reviewed the available information on measured environmental concentrations of MCC Ps and LCCPs, which are not normally available for PMNs.



### 5.1 ENVIRONMENTAL MONITORING For this assessment, environmental monitoring data consisting of me asured levels of MCCPs and LCCPs in surface water, sediment and soil were used to characterize potential environmental exposure to MCCPs and LCCPs. These data are not amenable to determining the ultimate release source ( i.e. , manufacturing, processing, or use) into the environment; however, they provide some insight on the geographical and temporal distributi on of MCCPs and LCCPs. Appendix D contains information on the data used in this risk assessment.

Studies published between 1980 and 2013 that reported environmental concentrations of MCCPs and/or LCCPs were reviewed for this assessment. Monitoring studies from the early 1980s could not distinguish between the different chain lengths of CPs. The introduction of modern techniques, such as electron capture negative ion mass spectrometry (ECNI

- MS) allowed for the detection of specific congeners, although diffi culties with these methods have persisted ( e.g. , detection of low chlorination congeners in samples). Tomy (2010) performed a round robin laboratory study of SCCPs that highlighted the inability of the ECNI

- MS method to consistently measure a reference sample, with concentrations varying up to a factor of six. Subsequent work showed that significant errors (up to a factor of t en) could be introduced by the improper selection of the calibration standards (Coelhan et al., 2000) . A more recent inter

- laboratory study of SCCPs found good agreement amongst the laboratories that used ECN I

- MS (Pellizzato et al., 2009) , but similar inter

- laboratory studies for MCCPs or LCCPs have not been completed (Tomy, 2010)

. The majority of the monitoring data were collected in Europe, and some more recent monitoring data we re collected in China. Over time and across countries, industrial practices and effluent20 pre

- treatment have varied. Some of the monitoring studies only published their final measured concentrations, and did not include the details of the analytical techniq ues and sampling locations. Generally, EPA/OPPT used studies sponsored by the environmental agencies, but full documentation is lacking for even these studies. The industrial sectors studied by other countries also are present in the US, suggesting that co nditions in the US may be similar.

The level of detail provided in the studies varied. Some studies provided detailed information regarding sampling locations ( e.g. , impacted sites), analytical methodology, and final sample results including detection li mits, quantitation limits, and estimated values. In contrast, other studies provided only a summary of the results combined from a number of studies. These summaries also did not provide details of the data analysis to obtain sample results. In addition,

c ertain studies reported concentrations within a given country but did not provide additional details about the exact sampling location. Given the disparate conditions ( i.e. , number of sites sampled, temporal period over which samples collected, differing a nalytical methods, etc .) across the data sets, EPA/OPPT was unable to determine a central tendency or distribution for the data sets, and a range was used instead. Studies using older analytical techniques that did not distinguish CP congeners were not use d in this assessment. Other nations’ assessments that used newer, more reliable, analytical techniques were considered.

EPA/OPPT used the following selection criteria to identify the studies included in this
assessment:

• Specific mention of MCCP/LCCP chain length;

• Use of modern analytical techniques to distinguish categories of CPs; and

• At a minimum, general information on sampling location

.

EPA/OPPT used the monitoring data summarized in Tables 6 and 7 for this assessment. When a limit of detection (LOD) value was reported for non

- detectable results 6

, EPA/OPPT used one half of the LOD value.

Even though the existing monitoring data were limited in quality and quantity, and it remains unclear how well the measured data describe the potential range of US MCCP and LCCP use scenarios, EPA/OPPT concluded that the data in Tables 6 and 7 represented the best available monitoring information for MCCPs and L CCPs, respectively. These data provide some evidence that MCCPs and LCCPs are released into the environment; however, these data reflect discrete locations and times, and the extent to which they are representative of the overall distribution of MCCPs and LCCPs is unknown.

6 Examples would be “not detected” (ND), negligible, or with a “less than” qualifier.21 Ta b le 6: Summary of Measured Concentrations of MCCPs in Environmental Media and Biota

Media Category n Min Unit Max Unit References

Surface water (non

- marine) 15 <2.50 × 10

- 10 mg/L 1.49 × 10

- 3 mg/L Coelhan (2010) ; EC (2008b) ; Houde et al. (2008) ; IPCS (1996) ; Muir et al. (2003) ; Petersen et al. (2006) a ; USEPA (1988)

Sediment (non

- marine) 78 2.00 × 10

- 3 mg/kg** 6.51 × 10 1 mg/kg dw Borgen et al. (2003) ; Chen et al. (2011) ; EC (2008a) ; Iozza et al. (2008) ; IPCS (1996) ; Nicholls et al. (2001) ; Petersen et al. (2006) ; Pribylova et al. ( 2006) ; Tomy et al. (1998); Tomy et al. (1999) ; USEPA (1988)

Sediment (marine) 54 5.00 × 10

- 3 mg/kg dw 1.64 × 10 1 mg/kg dw Hüttig et al. (2004) ; Hüttig and Oehme (2005, 2006) ; Kemmlein et al. (2002) ; Muir et al. (2000)
Sludge 9 5.00 × 10

- 5 mg/kg ** 9.70 × 10 3 mg/kg dw Stevens et al. (2003) ; Pribylova et al. (2006)

Soil 12 2.1 × 10

- 6 mg/kg dw 8.5 × 10

- 2 mg/kg dw Iozza (2010) ; Wang et al. (2013)

Biota (aquatic) 120 <2.00 × 10

- 7 mg/kg 2.63 mg/kg

ww Bennie et al. (2000) ; EC (1993, 2008a) ; Houde et al. (2008) ; IVL (2009) ; Ke mmlein et al. (2002) ; Muir (2010) ; Muir et al. (2003); Muir et al. (2000) ; USEPA (1988) ; Tomy et al., (1999a) Biota (terrestrial) 8 5.00 × 10

- 3 mg/kg ww 3.70 × 10

- 1 mg/kg

ww Reth et al. (2006)

a Peterson (2006) reported results for 2 water samples; EPA/OPPT assumes these are non

- marine surface water samples.

b The weight type was not reported ( i.e. , wet, dry, or lipid weight). Notes: 1. All values provided in the table above represent total MCCP and not individual MCCP isomers 2. The “n” value represents the number of media

- specific MCCP monitoring data values that were compiled from various articles in the raw data table (provided in Appendix D). 3. In some cases, the minimum values in the table are preceded by “<”. This indicate s that the value reported in article was reported as a non

- detect. In such cases, one half of the lowest reported detection limit was compiled as the ‘minimum’ reported monitoring data. 4. dw

– dry weight and ww

– wet weight

Table 7 below summarizes the available environmental monitoring data for LCCPs. Environmental data were available for marine sediment and aquatic invertebrates. Though no data were available for other media categories ( e.g. , surface water, non

- marine sediment, terrestrial invertebrates), limited high quality data (from Table 6 ) were available for MCCPs which could be used for informing concentrations of LCCPs in the environment. This decision is based on the following inform ation: 1) P

- 12

- 0284 does not contain any C 17 congener groups; however, it is reasonable to assume that C 17 and C 18 congener groups have similar ecotoxicological effects when the wt% Cl is comparable , 2) LCCPs are also expected to behave in a manner similar to the MCCPs when released to the environment, and 3) MCCPs and LCCPs22 have similar uses (see Table 1) and hence may have similar releases at facilities that process and use these chemicals.

Table 7 : Summary of Measured Concentrations of LCCPs in Environ mental Media and Biota

Media Category n Min Units Max Units References

Sediment (marine) 4 1.02×10

- 1 mg/kg dw 4.31×10

- 1 mg/kg dw Kemmlein et al . (2002)

Biota (aquatic) 2 2.80×10

- 6 mg/kg lw 6.90×10

- 6 mg/kg lw Kemmlein et al. (2002)

Notes: 1. All values provided in the table abo ve represent total LCCP C 18

- 20 and not individual LCCP isomers

. 2. The “n” value represents the number of media

- specific LCCP C 18

- 20 monitoring data values that were compiled from various articles in the raw data table (provided in Appendix D). 3. dw

– dry weight and lw

– lipid weight

.


### 5.2 MODELED ENVIRONMENTAL RELEASES
EPA/OPPT used screening

- level models to generate environmental r elease estimates for P

- 12

- 0282 to 0284, which were used to calculate exposure concentrations for estimating risks to humans and aquatic organisms. EPA/OPPT used the Chemical Screening Tool for Exposures and Environmental Releases (ChemSTEER ver.2) to estimate environmental releases from industrial processes; the results are provided in Appendix E . Inputs to the ChemSTEER ver. 2 release modeling were based on multiple sources, including information provided by Dover, published OECD Emission Scenario Documents, EPA/OPPT Generic Scenarios, and EPA/OPPT models for estimating environmental releases. Ta ble 8 provides a general summary of the release
assessment.

Table 8: Summary of Estimated Release to Water PMN Chemical Name Manufacturing Processing 1 Use 1

P

- 12

- 0282 Alkanes, C 14

- 16 , chloro (MCCP)

No Yes, from 4/5 processes Yes, from 5/6 uses P

- 12

- 0283 Tetradecane, chloro (MCCP) Yes, from 2/5 processes

P12

- 0284 Octadecane, chloro (LCCP) Yes, from one process Yes, from one use

1 For P

- 12

- 0282 and P

- 12

- 0283, water releases were estimated to be released for one day/year for the formulation of plasticizers in paints, sealants and adhesives. Similar release estimates were done for the formulation of metal working fluids and engine lu bricants for P

- 12

- 0283 (but not P

- 12

- 0282). No releases to water were expected from the commercial use of engine oil lubricant from P

- 12

- 0282 or P

- 12

- 0283.

Exposure pathways of interest for human health include drinking water, fish ingestion, and air
st ack emissions. For aquatic organisms, the exposure pathway of concern is from direct releases
to water. EPA/OPPT assessed each of these pathways by using the ChemSTEER ver. 2 release estimates as inputs to the Exposure and Fate Assessment Screening Tool (E

- FAST V2.0) for estimating industrial releases and concentrations in the foregoing exposure pathways. EPA/OPPT assumed that potential releases to water occurred from indirect discharges to publicly owned treatment works (POTW). The E

- FAST V2.0 modeling app lied an assumption of 90 %23 removal of MCCPs and LCCPs at the POTW. Water concentrations were estimated using E

FAST V2.0’s probabilistic dilution model (PDM), which predicts downstream chemical concentrations from industrial discharges. These values were r eported as the central tendency for a median flow site, a low flow site, and the lowest seven

- day average flow that occurs on average once every ten years ( i.e. , 7Q10). These estimated water concentrations were compared to the COC of 1 μg/L for chronic aqu atic invertebrates (see Table 5). Air stack emissions were estimated using generic scenarios, which assumed inhalation exposures occurring 100 meters downwind of a facility. EPA/OPPT used these esti mated values for calculating human health and environmenta l risk s of P

- 12

- 0282, 0283 and 0284.

The results of the E

- FAST V2.0 modeling are provided in Appendix F. Tables 9 (MCCPs) and 10 (LCCP) present the values used in the risk assessment. As explained in the footnotes to these Tables, the values represent rea sonable worst

- case scenarios based on the processing and use scenarios presented in Table 8. However, these estimates require consideration of two important caveats: (1) limited environmental monitoring data are available, and (2) MCCPs and LCCPs are
expec ted to partition to particulates and sediment; however, the E

- FAST V2.0 models do not account for this partitioning

.24 Table 9: E

- FAST M odeling Values Used for MCCP s 1

Scenario 2 Water Release

– Human 3 Air Release

– Human 3 Water Release

– Aquatic Organisms 4

Drinking Water (mg/kg

- bw /d ay ) Fish Ingestion (mg/kg

bw /d ay ) Stack Air LADD (mg/kg

bw /d ay ) Range of Concentrations ( μ g/L)

P

- 12

- 0282 Manufacturing No releases No releases 1.11 × 10

- 4 No releases Proc1

– Plastics Compounding 3.76 × 10

- 4 4.6 × 10

- 3 1.89 × 10

- 3 9

- 67 (344)

Proc2

– Rubber Compounding 1.32 × 10

- 4 1.62 × 10

- 3 6.66 × 10

- 4 10

- 72 (367)

Proc3

– Formulating MWF 4.93 × 10

- 5 6.03 × 10

- 4 4.24 × 10

- 4 8

- 59 (299)

Proc4

– Formulating engine oil 2.91 × 10

- 5 3.56 × 10

- 4 1.21 × 10

- 4 4

- 30 (155)

Proc5

– Formulating plasticizer in paints, sealants, adhesives Water release of only one day/year, so values not used in risk assessment.

Use1

– Plastics converting 4.30 × 10

- 4 5.26 × 10

- 3 1.32 × 10

- 3 2.7

- 80 (447)

Use2

– Rubber converting 4.80 × 10

- 4 5.88 × 10

- 3 6.94 × 10

- 4 7.7

- 90 (774)

Use3

– Commercial use of MWF 3.81 × 10

- 4 4.67 × 10

- 3 1.01 × 10

- 3 5.7

- 73 (652)

Use4

– Commercial use of engine oil No releases No releases 8.90 × 10

- 4 No releases

Use5

– Commercial use of paints, sealants, adhesives 3.51 × 10

- 6 3.07 × 10

- 5 3.54 × 10

- 6 0.045

- 0.51 (5.4)

Use6

– Commercial use of textiles containing PMN as flame retardant/plasticizer 5.96 × 10

- 5 7.29 × 10

- 4 1.24 × 10

- 5 1.3

- 9.1 (46)

P

- 12

- 0283 (“Proc” and “Use” parameters are the same as P

- 12

- 0282) Manufacturing No releases No releases 1.11 × 10

- 4 No releases Proc1 7.84 × 10

- 5 9.59 × 10

- 4 3.94 × 10

- 4 8

- 59 (300) Proc2 4.26 × 10

- 5 5.21 × 10

- 4 2.14 × 10

- 4 (8.4

- 61 (313) Proc3 Water release of only one day/year, so values not used in risk assessment. Proc4 Water release of only one day/year, so values not used in risk assessment. Proc5 Water release of only one day/year, so values not used in risk assessment. Use 1 1.36 × 10

- 4 1.67 × 10

- 3 4.15 × 10

- 4 0.86

- 25.2 (142) Use 2 1.54 × 10

- 4 1.89 × 10

- 3 2.25 × 10

- 4 2.5

- 28.8 (249) Use 3 3.81 × 10

- 4 4.67 × 10

- 3 1.01 × 10

- 3 5.7

- 73 (652) Use 4 No releases No releases 8.90 × 10

- 4 No releases Use 5 2.49 × 10

- 6 3.05 × 10

- 5 3.51 × 10

- 6 0.045

- 0.51 (5.4) Use 6 1.94 × 10

- 5 2.37 × 10

- 4 4.61 × 10

- 6 1.3

- 9.1 (46)25 1 Taken from Appendix F. Values represent the highest concentrations/estimated (lifetime average daily dose, or LADD) doses for chronic ( i.e. , repeated exposure scenarios) for human health.

2 Proc1

- 5 = Processing, or formulating for listed products. MWF = metalworking fluids (MWFs). Use 1

- 6 = corresponding commercial use of products.

3 Estimated exposure values were corrected for absorption by the oral (50 % absorption) and inhalation (50 % absorption) routes of exposure (ECB 2005; EA 2009).

4 For PMNs, environmental risk wa s evaluated by performing a PDM as described above and in the E

- FAST Manual (2007). The ranges encompass concentrations from the central tendency for a median flow site ( e.g. , 9 μg/L) up to the central tendency for a low flow site ( e.g. , 67 μg/L)

– this example taken from the second row. The central tendency was calculated using the harmonic mean flow. The value in parentheses represents the 7Q10 ( e.g. , 344 μg/L) value normally us ed to determine chronic risk to aquatic organism s

.

Table 10: E

- FAST Model ing Values Used for the LCCP (P

- 12

- 0284) 1

Scenario 2 Water Release

– Human 3 Air Release

– Human 3 Water Release

– Aquatic Organisms 4

Drinking Water (mg/kg

- bw /d ay ) Fish Ingestion

(mg/kg

bw /d ay ) Stack Air LADD (mg/kg

bw /d ay ) Range of Concentrations ( μ g/L) Manufacturing No releases to water greater than one day/year Proc1

– Formulation of MWF No releases to water greater than one day/year

Use1

– Commercial use of MWFs 2.03 × 10

- 4 8.71 × 10

- 4 1.01 × 10

- 3 1.3

- 38.7 (215)

1 Taken from Appendix F. Values represent the highest concentrations/estimated (lifetime average daily dose, or LADD) doses for chronic ( i.e. , repeated exposure scenarios) for human health.

2 MWF = metalworking fluids,

3 Estimated exposure values were corrected for absorption by the oral (50 % absorption) and inhalation (50 % absorption) routes of exposure (ECB 2005; EA 2009).

4 For PMNs, environmental risk wa s evaluated by performing a PDM as described above and in the E

- FAST Manual (2007). The ranges encompass concentrations from the central tendency for a median flow site ( e.g. , 1.3 μg/L) up to the central tendency for a low flow site ( e.g. , 38.7 μg/L). The central tendency was calculated us ing the harmonic mean flow. The value in parentheses represents the 7Q10 ( e.g. , 215 μg/L) value normally used to determine potential chronic risk to aquatic organism s.



### 5.3 EXPOSURE ESTIMATES

**5.3.1 OCCUPATIONAL EXPOSURE ESTIMATES**  EPA/OPPT calculated screening

- level workplace exposure estimates with ChemSTEER ver. 2. Tables 11 (MCCPs) and 13 (LCCP ) provide a summary of the exposure estimates used in this risk assessment for evaluating worker exposures to P

- 12

- 0282 to 0284; detailed informat ion is provided in Appendix E. Table 12 provides details on the number of sites, workers, and days working/year for each of the two MCCP PMNs.26 Table 11: Summary of Occupational Exposure Estimates Used for the MC CP s 1,2

Scenario Route of Exposure 3 Inhalation (mg/d ay ) Dermal (mg/d ay ) Manufacturing Negligible due to low vapor pressure 880 (sampling liquid) 1800 (loading liquid) Proc1

– Plastics Compounding 0.03

- 4.4 1800 (unloading liquid)

Proc2

– Rubber Compounding 0.1

- 0.7 1800 (unloading liquid)

Proc3

– Formulating MWF Negligible due to low vapor pressure 180 (loading liquid) 1800 (unloading liquid) Proc4

– Formulating engine oil Negligible due to low vapor pressure 71 (loading liquid) 1800 (unloading liquid)

Proc5

– Formulating plasticizer in paints, sealants, adhesives Negligible due to low vapor pressure 260 (loading liquid) 1800 (unloading liquid)

Use1

– Plastics converting 1.2

- 15 Non

- quantifiable 4

Use2

– Rubber converting 0.7

- 15 Non

- quantifiable 5

Use3

– Commercial use of MWF 0.98

- 3.6 (mist) 1800 (liquid) 490 (mist)

Use4

– Commercial use of engine oil Negligible due to low vapor pressure 71 (unloading liquid)

Use5

– Commercial use of paints, sealants, adhesives 1.1

- 3.6 260 (unloading liquid) 1300 (coating using hand

- held spray gun)

Use6

– Commercial use of textiles containing PMN as flame retardant/plasticizer Negligible due to low vapor pressure 430 (during finishing) 1800 (unloading)

1 All values apply to both P

- 12

- 0282 and P

- 12

- 00283, with exceptions noted below. If one value is listed, it represents a “high end” exposure estimates if there are two values they represent the range of low end to high end. Some of the inhalation exposure values represent actual monitoring data from Europe and are cited appropriately below, others ar e estimates. Most dermal values are estimates; for sampling scenarios the assumption is that one hand is used, for loading/unloading, it is assumed that two hands are used. Exceptions are noted below.

2 See Table 13 for details on the number of sites, work ers, and days/year for each scenario. The following represent the estimated number of sites and workers per scenario:

3 Proc1 = Inhalation values based on range of monitoring data from EU

– 0.003 to 0.44 mg/m 3 converted to mg/day by multiplying by 10 m 3

– the volume of air taken in by a worker over an eight

- hour work shift (EU cited as 2008 EU risk assessment; see engineering report for P

- 12

- 0282 on pdf page 8). Proc2 = Inhalation values based on range of monitoring data from EU

– 0.01 to 0.07 mg/m 3 conv erted to mg/day by multiplying by 10 m 3

– the volume of air taken in by a worker over an eight

- hour work shift (EU cited as 2008 EU risk assessment; see engineering report for P

- 12

- 0282 on pdf page 12).27 Use 1 = Inhalation values. The low end of the range i s based on the high end of the range of monitoring data from EU

– 0.01 to 0.12 mg/m 3 converted to mg/day by multiplying by 10 m 3

– the volume of air taken in by a worker over an eight

- hour work shift. The high end is based on the OSHA PEL for “paraffin oil mist”

, which is converted from 5 mg/m 3 to 1 .5 mg/m 3 to account for 30% of product containing PMN substance ( see: http://www.cdc.gov/niosh/npg/npgd0472.html ). Use 2 = Inhalation values. The low end of the range is based on the high end of the range of mon itoring data from EU

– 0.01 to 0.07 mg/m 3 converted to mg/day by multiplying by 10 m 3

– the volume of air taken in by a worker over an eight

- hour work shift. The high end is based on the OSHA PEL which is converted from 15 mg/m 3 to1.5 mg/m 3 to account for 10% of product containing PMN substance (EU cited from two sources, 2008 EU risk assessment and 2010 EU CSR; see engineering report for P

- 12

- 0282 at pdf page 16). Use 3 = Inhalation exposures. User

- defined Inhalation Model. The 2011 ESD for MWF Operations estimates typical mist concentrations ranging from 0.19 to 0.39 mg/m 3 and high end concentrations ranging from 0.87 to 1.42 mg/m 3 depending on the type of MWF used. These estimates are based o n the geometric mean (typical) and 90th percentile data (high end) of data collected by NIOSH from 942 machinists at 79 shops across the U.S. A range of inhalation exposure estimates are based the high end for the range of geometric means and 90 th percenti les. It is assumed that the PMN is 25% of the mist (generic scenario assumptions: water conc. is 60% and the PMN is then estimated to be 10/40 = 25% of the mist). Dermal exposures. User

- defined Dermal Model. The 2011 ESD for MWF Operations estimates an ave rage dermal surface loading rate for MWFs of 2.9 mg/cm 2

- hr during metalworking operations. These data are more conservative than the EPA/OPPT 2

- Hand Dermal Contact Model and are used to provide a conservative estimate of dermal exposure during MWF use due to uncertainty at downstream operations (see engineering report for P

- 12

- 0282 at pdf pages 20

- 21). Use 5 = Inhalation exposures. Coating using hand

- held spray gun; EPA/OPPT Automobile Refinish Spray coating Inhalation Model (non

- volatile non

- polyisocyanate s). A 2002 EU RA on MCCPs and a 2010 EU CSR on MCCPs presents air monitoring data from paint spray application with values ranging from 0.002 to 0.19 mg/m 3 with a median of 0.04 mg/m 3 and a 90 th percentile of 0.085 mg/m 3 . The EPA/OPPT Autorefinishing Spray Coating Model calculates an exposure of 9 mg/m 3 . The EPA/OPPT model is used to provide a conservative estimate due to uncertainty in downstream use sites. Taking the highest measured value (0.19) and the model value (9) and converted them to mg/day by mul tiplying by 10 m 3

– the volume of air taken in by a worker over an eight

- hour work shift; results in the range of 1.9

– 90 mg/day. (SEE TEXT IN SECITON 6.2.1 FOR FURTHER EVALUATION OF THE EU MONITORING DATA).

4 PMN will be encapsulated in rubber sheets/blocks during handling of rubber raw material. While some surface contact may occur, dermal exposure to solids in this form is non

- quantifiable.

Table 12: Occupational E xposure Details for the MCCPs 1

Scenario Number of Sites Number of Workers Numb er of Days/year Manufacturing 2 8 250, 93 Proc1

– Plastics Compounding 32 768 250, 60 Proc2

– Rubber Compounding 33 792 96, 31 Proc3

– Formulating MWF 110 440 38, 12 Proc4

– Formulating engine oil 15 225 34, 12 Proc5

– Formulating plasticizer in paints, sealants, adhesives 33 1287 5, 2

Use1

– Plastics converting 32 1536 250 Use2

– Rubber converting 33 1584 250 Use3

– Commercial use of MWF 776, 250 37,248; 12,000 247 Use4

– Commercial use of engine oil 48, 16 240, 80 250 Use5

– Commercial use of paints, sealants, adhesives 117, 38 351, 114 250, 249

Use6

– Commercial use of textiles containing PMN as flame retardant/plasticizer 11 33 250, 9528 Scenario Number of Sites Number of Workers Numb er of Days/year

1 If a single value is present, it is the same for both P

- 12

- 0282 and P

- 12

- 0283. If there are two value s, the first one is P

- 12

- 0282 and the second is P

- 12

- 0283

Table 13: Summary of Occupational Exposure Estimates Used for the LCCP s 1

Route of Exposure Manufacturing Processing Use 3
Inhalation (mg/day) Negligible due to low
vapor pressure Negligible due to low vapor pressure 0.98

- 3.6

Dermal 2 (mg/day) 880 (sampling) 1800 (loading) 220 (loading) 1800 (unloading) 180 (unloading liquid) 490 (mist)

1 The following represent the estimated number of sites and workers per scenario: Manufacturing (one site, four workers

– three days/year) Processing = formulation of metalworking fluids, MWFs (24 sites, 192 workers

– three days/year) Use = use of MWFs (10 sites, 480 workers, 247 days/year)

2 Most dermal values are estimates; for sampling scenarios the assumption is that one hand is used, for loading/unloading, it is assumed that two hands are used. Exceptions are noted below.

3 Inhalation exposures

. User

- defined Inhalation Model. The 2011 ESD for MWF Operations estimates typical mist concentrations ranging from 0.19 to 0.39 mg/m 3 and high end concentrations ranging from 0.87 to 1.42 mg/m 3 depending on the type of MWF used. These estimates are based on the geometric mean (typical) and 90 th percentile data (high end) of data collected by NIOSH from 942 machinists at 79 shops across the U.S. A range of inhalation exposure estimates are based the high end for the range of geometric means and 90 th percentiles. It is assumed that the PMN is 25% of the mist (generic scenario assumptions: water conc. is 60% and the PMN is the n estimated to be 10/40 = 25% of the mist). Dermal exposures. User

- defined Dermal Model. The 2011 ESD for MWF Operations estimates an average dermal surface loading rate for MWFs of 2.9 mg/cm 2

- hr during metalworking operations. These data are more conserva tive than the EPA/OPPT 2

- Hand Dermal Contact Model and are used to provide a conservative estimate of dermal exposure during MWF use due to uncertainty at downstream operations (see engineering report for P

- 12

- 0284 at pdf pages 8

- 9).

**5.3.2 CONSUMER EXPOSURE**  For the two MCCP PMNs, estimates were made for “do it yourself” consumers being exposed to P

- 12

- 0282 and P

- 12

- 0283 via a used motor oil exposure scenario. Table 14 provides the estimates and the assumptions used.

Table 14: Consumer Exposure Results Summary for Dermal Exposure to Used Motor Oil 1

Scenario: CEM Used Motor Oil Dermal Inhalation

ADR LADD ADR LADD
mg/kg

- bw /day mg/kg

- bw /day mg/kg

- bw /day mg/kg

- bw /day P

- 12

- 0282 and P

- 12

- 0283 24.6 0.205 No inhalation exposure expected

1 ADR = acute dose rate, LADD is lifetime average daily dose. Assumptions include: used oil is handled four times/year, amount retained is 0.0105 grams/cm 2 per handling event, and the averaging time for the LADD is 57 years.29 

## 6 RISK ASSESSMENT

6.1 E NVIRONMENTAL ASSESSMENT PMN risk assessments typically use modeled exposure values because new chemical substances are not in the stream of US commerce; however, for MCCPs and LCCPs, measured environmental data are available for some locations in the US and abroad. Though these data are not specific to P

- 12

- 0282 to 0284, the data contain MCCP and LCCP congener groups that may be present in the PMN substances. However, EPA/OPPT used modeled exposure values as an important source for its assessment of risk because the modeled expos ure values were generated using exposure scenarios that were specific to the types of uses and releases that may occur with P

- 12

- 0282 to 0284. In contrast, the measured environmental data are generally not amenable for identifying the types of uses or releases from which the measured congeners originated. Therefore, EPA/OPPT us ed these measured data as supporting information, along with modeled exposure values, to calculate potential environmental risks using the risk quotient (RQ) method.

The RQ method integrates the results of exposure and ecotoxicity data (USEPA, 1998 )

.
An RQ is defined as:

RQ = Environmental Concentration ÷ Effect Level

where, the environmental concentration represents measured (see Tables 6 and 7) or estimated (see Tables 9 and 10) values for each compartment ( i.e. , water, sediment, and soil), and the effect level represents the COC for aquatic, benthic, or terrestrial species (see Table 5).

An RQ greater than one serves as a benchmark for identifying whether aquatic concentrations of P

- 12

- 0282 to 0284 may present a risk to aquatic

- and sediment

- dwelling organisms.

**6.1.1 Risk**  **Estimates Using Environmental Monitoring Data**  The RQs shown in Table 15 suggest that measured concentrations of MCCPs and LCCPs in water and sediment may present a risk of acute and chronic injur y to aquatic organisms and may present a risk of chronic injury to sediment

- dwelling organisms. However, several limitations must be noted about the monitoring studies and the level of uncertainty that they contribute to the basis of these findings. First, the reported concentrations represent m inimum and maximum values that span, at a minimum, several orders of magnitude and translate to RQs of less than one ( i.e. , low risk finding) or greater than one ( i.e. , risk finding), respectively. Second, the temporal and geographical distributions of the se data, along with the different types of uses and releases that may have served as the originating sources, make it impossible to describe the central tendency of these data. Finally, the frequency and magnitude of locations with relevant use and release scenarios to the PMN substances, which may result in environmental releases of MCCPs or LCCPs that exceed the relevant COCs, is unknown. In addition to these general limitations, there are specific limitations and uncertainties that preclude using these v alues as the sole source from which to inform potential environmental concentrations and risks that may result from the specific uses and releases associated with P

- 12

- 0282 to 0284.30 Table 15: Risk Quotients Calculated from Mon itoring D ata for Surface Wat er, Sediment and the Terrestrial Environment Environmental Concentration Effect Level ( i.e. , COC) RQs 1

Acute Risk Aquatic Species < 2.50×10

- 10 to 1.49×10

- 3 mg/L 0.001 mg/L < 2.50×10

- 7 to 1.49

Chronic Risk Aquatic Species < 2.50×10

- 10 to 1.49×10

- 3 mg/L 0.001 mg/L < 2.50×10

- 7 to 1.49

Chronic Risk Sediment

- dwelling Species Non

- marine Environment 0.002 to 65 mg/kg dw 18.7 mg/kg dw 1.07×10

- 4 to 3.5

Chronic Risk Terrestrial Species Insufficient Data 14.9 mg/kg dw 2 Not calculated

1 Bolded values represent those that may present a risk of injury.

2 The COCs for terrestrial invertebrates and vertebrates were 14.9 mg/kg dw and 16.8 mg/kg diet, respectively. Since these values were comparable, EPA/OPPT used the lowest value for calcul ating RQs for this compartme nt

.
For surface water, EPA/OPPT based the aquatic risk findings for MCCPs and LCCPs on the highest concentration reported by Petersen et al. (2006). These authors collected two surface water samples from an undisclosed location(s) in Norway and measured the concentration of MCCP congener groups ( i.e. , C 14

- 17 ). The authors reported a concentration of 1.49 × 10

- 3 mg/L for MCCP congener groups in one sample; however, a numerical value was not provided for the second sample, rather the distribution of congen er groups was displayed in a bar graph. Based on the ordinate scale, the concentration of MCCP congener groups in the second sample was greater than zero, but less than 5.0 × 10

- 4 mg/L. Of the monitoring studies reviewed by EPA/OPPT (see Appendix D), the P etersen et al. (2006) value of 1.49 × 10

- 3 is the only surface water concentration that resulted in an RQ greater than one. All other surface water concentrations are at least one order of magnitude below 1.49 × 10

- 3 mg/L ( i.e. , RQs < 1 ).

For sediment concentrations, EPA/OPPT reviewed multiple studies, some of which reported values that exceeded the COC. Nicholls et al. (2001) reported the most relevant data for P

- 12

- 0282

- 0284. These authors measured concentrations of MCCPs at locations in the United Kingdom where specific industries were known to employ MCCPs in the use categories identified for the PMN substances ( e.g. , lubricant in MWFs, plasticizer in PVC resins, and lubricant in sealants). Eight locations were sampled at three distances downstream ( i.e. , 100 meters, 300 meters, and 1

- 2 kilometers) from the respective sewage treatment works. At four of the locations, at least one of the sampled downstream values exceeded the COC ( i.e. , RQs > 1, risk finding). Though it is not possible to parse out the contribution of specific uses to the measured values, these data support that releases occur at locations with relevant uses to the PMN substances, which contribute to the environmental load of MCCP congener groups and in some cases result in RQs greater than one.

For soil concentrations, EPA/OPPT was unable to calculate RQs for terrestrial organisms due to the absence of relevant measured data from biosolid

- amended soils . Though Iozza (2010) a nd Wang et al. (2013) reported measured levels of MCCPs in soil, the authors obtained these samples from sites in remote alpine locations or industrialized areas, respectively. These data are relevant for assessing airborne deposition of MCCPs/LCCPs ; howev er , the reported31 concentrations are of questionable relevance with informing concentrations of MCCPs/LCCPs that may occur in biosolid

- amended agricultural soils.

Due to the foregoing limitations and resulting uncertainties with the measured environmental data, EPA/OPPT used these data in a limited capacity for estimating potential risks associated with the use categories identified for P

- 12

- 0282 to 0284. Specifically, these data were used as supporting information to inform the relevant pathways for estima ting potential releases from relevant use categories for the PMN substances. A summary of the estimated release values and associated RQs that EPA/OPPT used as the primary basis for evaluating the potential risks of P

12

- 0282 to 0284 is presented in the fo llowing section.

**6.1.2 R**  **isk Estimates Using Modeled**  **Exposure s**  The RQs shown in Tables 16, 17 and 18 (for each PMN substance) suggest that the majority of the intended processes and uses for P

- 12

- 0282, 0283 and 0284 are expected to result in releases to surfac e water at concentrations that may present a risk of injury to aquatic organisms. There are three exceptions ( i.e. , those scenarios for which the RQ is less than 1):

• P

- 12

- 0282: The median stream flow scenario for Use5 (Commercial use of paints, sealants, adhesives

– RQ of 0.045)

• P

- 12

- 0283: The median stream flow scenario for Use1 (Plastics converting

– 0.86)

• P

- 12

- 0283: The median stream flow scenario for Use5 (Commercial use of paints, sealants, adhesives

– RQ of 0.045)

It is noteworthy that these estimat ed concentrations are within the range of measured surface water concentrations reported for MCCP congener groups (Table 6). Though there is uncertainty whether the form ( i.e. , dissolved or particle bound) of MCCP impacts the aquatic toxicity, the estimate d values suggest that either form may exist. The median stream flow estimates are all below the reported water solubility for the P

- 12

- 0282 and 0283 and slightly above or below the water solubility reported for P

- 12

- 0284 (Table 1). Since the available aqua tic toxicity data support that dissolved MCCP congener groups cause toxicity, the median stream flow values suggest that the risk finding for this scenario is plausible. The low stream flow and 7Q10 flow scenarios estimate water concentrations that far exc eed the estimated water solubility of all three PMN substances. Under these scenarios, the MCCP or LCCP congener groups would likely be bound to particulates and would eventually settle out in sediment. Nicholls et al. (2001) provided support for this path way and showed that sediment concentrations of MCCP congener groups generally increased with distance downstream from the source outfall. Based on the foregoing information, EPA/OPPT concludes that the median stream flow values were adequate for determinin g that environmental releases of P

- 12

- 0282 to 0284 may present a risk of injury to aquatic organisms.32 T able 16: Risk Assessment of Aquatic Organism s using Model ed Exposures to P

- 12

- 0282 1

Scenario Estimated Water Concentrations ( μ g/L) 2 RQs Median Stream Flow Scenario Low Stream Flow Scenario 7Q10 Flow Scenario Proc1

– Plastics Compounding 9 67 344 9

- 344 Proc2

– Rubber Compounding 10 72 367 10

- 367 Proc3

– Formulating MWF 8 59 299 8

- 299 Proc4

– Formulating engine oil 4 30 155 4

- 155 Use1

– Plastics converting 2.7 80 447 2.7

- 447 Use2

– Rubber converting 7.7 90 774 7.7

- 774 Use3

– Commercial use of MWF 5.7 73 652 5.7

- 652

Use5

– Commercial use of paints, sealants, adhesives 0.045 0.51 5.4 0.045

- 5.4

Use6

– Commercial use of textiles containing PMN as flame retardant/plasticizer 1.3 9.1 46 1.3

- 46

1 Taken from full model run of summary data presented in Appendix E

. 2 For PMNs, EPA/OPPT evaluated potential environmental risks by performing a PDM as described above and in the “Exposure a nd Fate Assessment Screening Tool (E

- FAST) Version 2.0 Documentation Manual (2007)”, available at: http://www.epa.gov/opptintr/exposure/pubs/efast2man.pdf

Table 17: Risk A ssessment of Aquatic Organisms using Modeled Exposures to P

- 12

- 0283 1

Scenario Estimated Water Concentrations ( μ g/L) 2 RQs Median Stream Flow Scenario Low Stream Flow Scenario 7Q10 Flow Scenario Proc1

– Plastics Compounding 8 59 300 8

- 300

Proc2

– Rubber Compounding 8.4 61 313 8.4

- 313

Use1

– Plastics converting 0.86 25.2 142 0.86

- 142 Use2

– Rubber converting 2.5 28.8 249 2.5

- 249 Use3

– Commercial use of MWF 5.7 73 652 5.7

- 652

Use5

– Commercial use of paints, sealants, adhesives 0.045 0.51 5.4 0.045

- 5.4

Use6

– Commercial use of textiles containing PMN as flame retardant/plasticizer 1.3 9.1 46 1.3

- 46

1 Taken from full model run of summary data presented in Appendix E

. 2 For PMNs, EPA/OPPT evaluated potential environmental risks by performing a PDM as described above and in the “Exposure and Fate Assessment Screening Tool (E

- FAST) Version 2.0 Documentation Manual (2007)”, available at: http://www.epa.gov/opptintr/exposure/pubs/efast2man.pdf33 T able 18: Risk A ssessment of Aquatic Organisms using Modeled Exposures to P

- 12

- 0284 1

Scenario Estimated Water Concentrations ( μ g/L) 2 RQs Median Stream Flow Scenario Low Stream Flow Scenario 7Q10 Flow Scenario Proc1

– Formulation of MWF No releases to water for more than one day/year

Use1

– Commercial use of MWFs 1.3 38.7 215 1.3

- 215

1 Taken from full model run of summary data presented in Appendix E

. 2 For PMNs, EPA/OPPT evaluated potential environmental risks by performing a PDM as described above and in the “Exposure and Fate Assessment Screening Tool (E

- FAST) Version 2.0 Documentation Manual (2007)”, available at: http://www.epa.gov/opptintr/exposure/pubs/efast2man.pdf



### 6.2 H UMAN HEALTH
EPA/OPPT assessed potential risks to workers and the general population by calculating margins of exposure (MOE). This approach is performed according to the following equation:

MOE = Point of Departure (POD) ÷ Estimated human exposure

For the PODs, EPA/OPPT identified effect levels from three oral repeated dose toxicity studies, which served as the basis for calculating human equivalent doses (HEDs) (CXR 2005; NTP 198 6). In the first study, CXR (2005) reported a NOAEL of 23 mg/kg

- bw/day based on increased kidney weight at 222 mg/kg

- bw/day in male rats exposed through diet for 90 days to an MCCP congener group (C 14

- 17 , 52 wt% Cl). In the second and third studies, NTP (1 986) reported LOAELs of 100 mg/kg

- bw/day based on granulomatous inflammation of the liver in female rats administered an LCCP congener (C 23 , 43 wt% Cl) by gavage for 5 days/week for 12 months or two years.

Using the effect levels of 23 mg/kg

- bw/day (for the MCCP PMNs: P

- 12

- 0282 and P

- 12

- 0283) or 100 mg/kg

- bw/day (for the LCCP PMN [P

- 12

- 0284]), EPA/OPPT performed route

- to

- route extrapolations to develop HEDs for inhalation and dermal exposures in workers and for inhalation and oral exposures in the general population. EPA/OPPT did not assess oral exposures for workers, due to the unlikely nature of exposures occurring by this route. The respective HEDs served as the PODs for calculating MOEs, along with the pr eviously reported estimated human exposure values for workers (Table s 11 and 13 ) and the general population (Table s 9 and 10 ).

EPA/OPPT compared the MOEs to a benchmark value that consisted of a multiplicative composite of three possible uncertainty fact ors (UFs): intraspecies variability (UF H ; default value = 10), interspecies variability (UF A ; default value = 10), and LOAEL

- to

- NOAEL extrapolation uncertainty (UF L ; default value =10). The UF H and UF A may each be subdivided to account for toxicokinetics ( TK; default value = 3.16) and toxicodynamics (TD; default value = 3.16). When effect levels from experimental animal studies are converted to HEDs, EPA/OPPT ’s default approach is to reduce the TK subfactor of UF A to 1 ( i.e. , UF A = TK × TD = 1 × 3.16 ≈ 3).34 EPA/OPPT interpretes MOEs that were equal to or below a benchmark value ( e.g. , MOE < 1000 [UF H × UF A × UF L = 1000]) as an indication that the scenario(s) may present a risk of injury to human health, whereas MOEs that were above the bench mark value as a low risk finding. In the following sections, more detailed descriptions are provided on: 1) converting effect levels to
route

- and exposure

- specific HEDs; 2) determining the appropriate UFs for the benchmark value, and 3) evaluating risk estimates for work ers and the general population.

**6.2.1 Workers**  EPA/OPPT performed route

- to

- route extrapolations to convert the oral NOAEL of 23 mg/kg

bw/day ( i.e. , MCCP congener groups) and the oral LOAEL of 100 mg/kg

- bw/day ( i.e. , LCCP congener) to HED values for inhalation exposures to workers ( i.e. , HED INHAL

- WORKER ) using the following equation:

HED INHAL

- WORKER = NOAEL ORAL × (1 ÷ sRV RAT ) × (ABS ORAL

- RAT ÷ ABS INHAL

- HUMAN ) × (sRV HUMAN ÷ wRV)

where,

NOAEL ORAL = 23 or 100 mg/kg

- bw/day sRV RAT = rat standard respiratory volume for 8

- hours = 0.38 m 3 /kg bw ABS ORAL

- RAT = percent absorption by the oral route in rats = 50 % ABS INHAL

- HUMAN = percent absorption by inhalation in humans = 50 % sRV HUMAN = human standard respiratory volume for 8

- hours = 6.7 m 3 wRV = worker respiratory volume for 8

- hours = 10 m 3

For the oral NOAEL of 23 mg/kg

- bw/day and the oral LOAEL of 100 mg/kg

- bw/day, the HED INHAL

- WORKER values equal 41 mg/m 3 and 176 mg/m 3 , respective ly.

EPA/OPPT calculated the HED values for dermal exposures to workers ( i.e. , HED DERM

- WORKER ) based on the following equation:

HED DERM

- WORKER = NOAEL ORAL × (ABS ORAL

- RAT ÷ ABS DERMAL

- HUMAN ) × (BW RAT ÷ BW HUMAN ) 1/4

where,

NOAEL ORAL = 23 or 100 mg/kg

- bw/day ABS ORAL

- RAT = percent absorption by the oral route in rats = 50 % ABS DERM

- HUMAN = percent absorption by the dermal route in humans = 1 % BW RAT = rat bodyweight = 0.250 kg BW HUMAN = human bodyweight = 71.8 kg

The resulting HED DERM

- WORKER values equal 4600 mg/kg

- bw/day for MCCP congener groups and 20000 mg/kg

- bw/day for the LCCP congener.

EPA/OPPT used the foregoing HED values to inform the appropriate application of UFs to derive benchmark values. For MCCP congener groups and the LCCP congener, a benchmark value of 30 or 300 was applied. These values consisted of the following individual UFs. A35 default UF H of 10 was applied to each benchmark value, due to the absence of experimental data to inform the TK and TD subfactors of t his UF. A reduced UF A of 3 was applied to each benchmark value, which accounted for a TK subfactor of 1 after converting the effect levels to HEDs. The UF A of 3 accounted for the remaining uncertainty associated with TD variability. A default UF L of 10 was only used for the benchmark value compared to the MOE derived from an LCCP congener because the underlying study reported a LOAEL, not a NOAEL.

EPA/OPPT used the HED INHAL

- WORKER and HED DERM

- WORKER values for calculating the respective MOEs using the esti mated exposure values presented in Table s 11 (P

- 12

- 0282 and P

12

- 0283) and 13 (P

- 12

- 0284) . As shown in Table 19 , the MOEs for P

- 12

- 0 282

, P

- 12

- 0 28 3 and P

- 12

- 0284 exceeded the respective benchmark values for all the dermal exposure scenarios; which indicate a finding of low risk to workers for the processes and uses evaluated in this assessment. For the inhalation exposures , the same low risk finding was observed for all scenarios

. Use1 and Use5 were evaluated using two separate approaches, both of which did not indicate potential risks of concern. For Use1, as noted in the footnote in Table 11, t he high end value is based on the OSHA PEL for “paraffin oil mist” . The low

- end value is the highest value reported in an EU workplace monitoring study and, although identified as “low

- end” here, represents a reasonable worst

- case from a CP workplace monitoring study. For Use5, again as noted in the footnote in Table 11, the high

- end value pr esented is the EPA/OPPT Auto refinishing Spray Coating Model calculation. This is used to provide a conservative estimate to account for uncertainty in downstream uses. The footnote also provides a range of values from an EU monitoring study

. Each of the three tasks in the monitoring study were performed in duplicate ( i.e. , apply a first coat, let it dry, and then apply a second coat). One individual performed all of the tasks ( e.g. , painting a fence). The time

- weighted average from all tasks w as 1.1 mg/m 3

, whi ch represents a reasonable worst case scenario for a painter undertaking multiple tasks

.36 Table 19 : Occupational MOEs for P

- 12

- 0282, P

- 12

- 0283 and P

- 12

- 0284

Exposure

Route Exposure Scenario Margins of Exposure

P

- 12

- 0282 and P

- 12

- 0283 1 Benchmark MOE= 30

Inhalation Proc1

– Plastics Compounding 186

- 27,333 Proc2

– Rubber Compounding 1171

- 8200 Use1

– Plastics converting 55

- 683 Use2

– Rubber converting 55

- 1171 Use3

– Commercial use of MWF 228

- 837 Use5

– Commercial use of paints, sealants, adhesives 228

- 745

Dermal Manufacturing 18,349

- 37,532 Proc1

– Plastics Compounding 18,349 Proc2

– Rubber Compounding 18,349 Proc3

– Formulating MWF 18,349

- 183,489 Proc4

– Formulating engine oil 18,349

- 465,183 Proc5

– Formulating plasticizer in paints, sealants, adhesives 18,349

- 127,031

Use3

– Commercial use of MWF 18,349

- 67,404 Use4

– Commercial use of engine oil 465,183 Use5

– Commercial use of paints, sealants, adhesives 25,046

- 127,031 Use6

– Commercial use of textiles containing PMN as flame retardant/plasticizer 18,349

- 76,809

P

- 12

- 0284 2 Benchmark MOE= 300 Inhalation Commercial use of MWF 978

- 3592

Dermal Manufacturing 79,778

– 161,382 Processing

- Formulating MWF 79,778

– 652,727 Commercial use of MWF 293,061

- 797,778

1 There were no inhalation exposures for the following scenarios for P

- 12

- 0282 and P

- 12

- 0283 : Manufacturing, Proc3, Proc4, Proc5, Use 4 and Use6. There were no dermal exposures for the following scenarios: Use1 and Use2. Proc3 = Formulating MWFs, Proc4 = formulating engine oil; Proc5 = formulating plasticizer in paints, sealants and adhesives.

2 Manufacture process and use is only for metalworking fluids. There was no inhalation exposure for the manufacturing and processing scenarios.

6.2.2 General Population (from Environmental Releases) EPA/OPPT converted the oral NOAEL of 23 mg/kg

- bw/day ( i.e. , MCCP congener groups) and the oral LOAEL of 100 mg/kg

- bw/day ( i.e. , LCCP congener) to HED values for oral exposures to the general population ( i.e. , HED ORAL

- GENPOP ) by the environment using the following equation:

HED ORAL

- GENPOP = NOAEL ORAL × (ABS ORAL

- RAT ÷ ABS ORAL

- HUMAN ) × (BW RAT ÷ BW HUMAN ) 1/4 × (5 days ÷ 7 days) a

where,

NOAEL ORAL = 23 or 100 mg/kg

- bw/day37 ABS ORAL

- RAT = percent absorption by the oral route in rats = 50 % ABS ORAL

- HUMAN = percent absorption by the oral route in humans = 50 % BW RAT = rat bodyweight = 0.250 kg BW HUMAN = human bodyweight = 71.8 kg

a A duration

- specific adjustment was only applied to the oral LOAEL of 100 mg/kg

bw/day because the animals were gavaged five days per week.

For assessing inhalation exposures to the genera l population, EPA/OPPT performed route

- to

route extrapolations to convert the oral NOAEL of 23 mg/kg

- bw/day ( i.e. , MCCP congener groups) and the oral LOAEL of 100 mg/kg

- bw/day ( i.e. , LCCP congener) to HED values for inhalation exposures to the general popu lation ( i.e. , HED INHAL

- GENPOP ) using the following equation:

HED INHAL

- HUMAN = NOAEL ORAL × (1 ÷ sRV RAT ) × (ABS ORAL

- RAT ÷ ABS INHAL

- HUMAN ) × (5 days ÷ 7 days) a

where,

NOAEL ORAL = 23 or 100 mg/kg

- bw/day sRV RAT = rat standard respiratory volume for 8

- hours = 1.15 m 3 /kg bw ABS ORAL

- RAT = percent absorption by the oral route in rats = 50 % ABS INHAL

- HUMAN = percent absorption by inhalation in humans = 50 %

a A duration

- specific adjustment was only applied to the oral LOAEL of 100 mg/kg

bw/day because the animals were gavaged five days per week.

For the oral NOAEL of 23 mg/kg

- bw/day and the oral LOAEL of 100 mg/kg

- bw/day, the HED INHAL

- GENPOP values equal 20 mg/m 3 and 62 mg/m 3 , respectively.

The same benchmark values of 30 or 300 were used for evaluating the general population MOEs. These benchmark values consisted of the same individual UFs and rationale discussed previously for workers.

EPA/OPPT used the HED ORAL

- GENPOP and HED INHAL

- GENPOP values for calculating the respective MOEs using the estimated exposure values presented in Table s 9 (MCCPs) and 10 (LCCP) . As shown in Table 20 , the MOEs for P

- 12

- 0 282,

- 0283, and

- 0284 all exceed the respective benchmark values , which indicate a finding of low risk to the general population for environmental exposures that may occur due to the processes and uses evaluated in this
assessment.38 Table 20 : E

- FAST Modeling Values Used for MCCP PMNs and the Risk Calculations 1 ,2 Scenario 2 Water Release

- Human Air Release

- Human Drinking Water MOE Fish Ingestion MOE Stack Air MOE

P

- 12

- 0282 (Benchmark MOE=30) Manufacturing No releases to water 1.8 × 10 5 Proc1

– Plastics Compounding 2.4 × 10 5 2.0 × 10 4 1.1 × 10 4

Proc2

– Rubber Compounding 7.0 × 10 5 5.7 × 10 4 3.0 × 10 4

Proc3

– Formulating MWF 1.9 × 10 6 1.5 × 10 5 4.7 × 10 4

Proc4

– Formulating engine oil 3.2 × 10 6 2.6 × 10 5 1.7 × 10 5

Proc5

– Formulating plasticizer in paints, sealants, adhesives Water release of only one day/year, so values not used in risk assessment.

Use1

– Plastics converting 2.1 × 10 5 1.7 × 10 4 1.7 × 10 4

Use2

– Rubber converting 1.9 × 10 5 1.6 × 10 4 2.9 × 10 4

Use3

– Commercial use of MWF 2.4 × 10 5 2.0 × 10 4 2.0 × 10 4

Use4

– Commercial use of engine oil No releases to water 2.2 × 10 4

Use5

– Commercial use of paints, sealants, adhesives 2.6 × 10 7 3.0 × 10 6 5.6 × 10 6

Use6

– Commercial use of textiles containing PMN as flame retardant/plasticizer 1.5 × 10 6 1.3 × 10 5 1.6 × 10 6

P

- 12

- 0283 3 (Benchmark MOE=30) Manufacturing No releases to water 9.0 × 10 6 Proc1 1.2 × 10 6 9.6 × 10 4 2.5 × 10 6 Proc2 2.2 × 10 6 9.0 × 10 5 4.7 × 10 6 Proc3 Water release of only one day/year, so values not used in risk assessment. Proc4 Water release of only one day/year, so values not used in risk assessment. Proc5 Water release of only one day/year, so values not used in risk assessment. Use 1 6.8 × 10 5 5.5 × 10 4 2.4 × 10 6 Use 2 6.0 × 10 5 4.9 × 10 4 4.4 × 10 6 Use 3 2.4 × 10 5 2.0 × 10 4 9.9 × 10 5 Use 4 No releases to water 1.1 × 10 6 Use 5 3.7 × 10 7 3.0 × 10 6 2.8 × 10 8 Use 6 4.7 × 10 6 3.9 × 10 5 2.2 × 10 8
P

- 12

- 028 4 (Benchmark MOE=300) Manufacturing Water release of only one day/year, so values not used in risk assessment. PROC1: Formulation of MWFs Water release of only one day/year, so values not used in risk assessment.

USE1: Use of MWF 4.5 × 10 5 1.1 × 10 5 9.1 × 10 439 1 Taken from Appendix F. Values represent the highest concentrations/estimated (lifetime average daily dose, or LADD) doses for chronic ( i.e. , repeated exposure scenarios) for human health.

2 For PMNs, EPA/OPPT evaluated potential risks to humans from environmental exposures by performing a PDM as described above and in the “Exposure and Fate Assessment Screening Tool (E

- FAST) Version 2.0 Documentation Manual (2007)”, available at: http://www.epa.gov/opptintr/exposure/pubs/efast2man.pdf

. 3 Proc1

- 5 = Processing, or formulating for listed products. MWF = metalworking fluids (MWFs). Use 1

- 6 = corresponding commercial use of products.

6.2.3 Consumer The consumer use assessed in this assessment is for dermal exposure to used motor oil for the MCCP PMNs (P

- 12

- 0282 and P

- 12

- 0283). Table 14 presents the acute and chronic exposure estimates.

A n ac ute risk assessment was not performed for this consumer exposure scenario. The estimated
acute exposure value was approximately 25 mg/kg

- bw. The only available acute hazard information on MCCPs is an oral LD 50 value of over 4000 mg/kg

- bw (see Appendix C for details). Given the available information on dermal absorption of MCCP (i.e., up to one percent), a quick calculation shows an equivalent dermal LD 50 value would be 400,000 mg/kg

- bw (or 400 grams). Compared to 25 mg/kg

- bw (or 0.025 gra ms/kg

- bw), it likely that risks would be low to

consumers from this use.

For the chronic value, u sing the same approach described above for assessing dermal exposures to workers, EPA/OPPT calculated the HED values for dermal exposures to consumers ( i.e.
, HED DERM

- CONSUMER ) based on the following equation:

HED DERM

- CONSUMER = NOAEL ORAL × (ABS ORAL

- RAT ÷ ABS DERMAL

- HUMAN ) × (BW RAT ÷ BW HUMAN ) 1/4

where,

NOAEL ORAL = 23 mg/kg

- bw/day ABS ORAL

- RAT = percent absorption by the oral route in rats = 50 % ABS DERM

- HUMAN = percent absorption by the dermal route in humans = 1 % BW RAT = rat bodyweight = 0.250 kg BW HUMAN = human bod yweight = 71.8 kg

The resulting HED DERM

- CONSUMER value equal s 4600 mg/kg

- bw/day for MCCP congener.

Using an absorption factor of 1 % for the dermal exposure value results in an exposure estimate of 0.01 x 25 = 0.25 mg/kg

- bw/day; which is then used to determine the MOE of 4600/0.25 = 18,400. This value is well above the benchmark MOE of 30 , showing low risk to consumers from this use.40 

## 7 CONCLUSIO NS Based on its assessment of the available hazard and exposure information on P

- 12

- 0282, P

- 12

- 0283, and P

- 12

- 0284, EPA/OPPT concludes the follo wing pertaining to the manufacturing, processing and use 7 of these PMN substances :

5. Occupational Exposures : giv en the assumptions, data and scenarios evaluated in this assessment, there w ere low risks found for workers from either dermal or inhalation
exp osure s.

6. General Population Exposures (from environmental releases) : given the assumptions, data and scenarios e valuated in this assessment, there were low risks found to humans from environmental releases via exposure to drinking water or fish ingestion.

7. Environmental Assessment :

a. Using estimated environ mental concentrations, the PMN substances may present an unreasonable risk following acute and chronic exposures to aquatic organisms

. i. The two exceptions are the low

- end estimates for aquatic concentrations for water releases from plastic converting use and commercial use of solvents, paints and adhesives (fo r the MCCP PMN P

- 12

- 0283)

b. Using available measured concentrations of MCCP and LCCP congener groups in the environment as supporting information, the PMN substances: i. A re expected to partition to sediment and may partition to soil through land application of biosolids, and ii. May be released to the environment at levels at or above estimated concentrations of MC CP and LCCP congener groups that may present an unreasonable risk following acute and chronic exposures to aquatic organisms

.

8. PBT Assessment : The PMN substances may be very persistent and very bioaccumulative

.

7 The two MCCP PMNs (P

- 12

- 0282 and 0283) have six different uses, but the three major uses are: as an add itive to plastics as a flame retardant and as a secondary plasticizer [43% of total production volume] and as an additive to rubber as a flame retardant and plasticizer [24%], and as an extreme pressure additive to lubricants [30%]. The LCCP PMN (P

- 12

- 028 4) is used exclusively as an extreme pressure additive to metalworking fluids.41 

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- 740.55 

## 9 APPENDICES56 

## Appendix A ENVIRONMENTAL FATE AND BIOACCUMULATION STUDY SUMMARIES

A

- 1 ENVIRONMENTAL PERSISTENCE

A

- 1

- 1 Abiotic Degradation Generally, CPs are stable to hydrolysis and to direct photolysis in air and water, though very limited data exist on hydrolysis and direct and indirect photolysis in soil, water, or air. In studies using aliphatic hydrocarbon solvents, CPs were shown to be poor absorbers of UV light and no direct photodegradation was observed (Friedman and Lombardo, 1975; Lombardo et al., 1975)

. Koh and Thiemann (2001) studied photolysis of aqueous solutions for CPs products with chain lengths ranging from C 10 to C 24 including an MCCP product, Hoechst CP52, with chain lengths from C 12 to C 18 and an average of 52 wt % Cl. A mercury vapor lamp with main radiation wavelengths of 254, 302, 313, 366,405/408, and 436 was used in batch experiments. Following a 5 hour radia tion time, estimated atmospheric degradation rates showed photolysis half

- lives of less than 20 hours based on measurement of free chloride and analysis of degradation products. The MCCP product had a T 1/2 of 12.8 hour in aqueous solution. The addition of peroxide or acetone increased the photolysis rate suggesting that indirect photolysis may be significant. The authors also reported that longer chain CPs were formed during this study and speculated that recombination of smaller alkyl radicals could occur under some conditions.

Thermal degradation data for MCCPs and LCCPs are limited, but studies of SCCPs and Polyvinyl chlorides suggest MCCPs are degraded rapidly at 250

- 350 °C (Bergman et al., 1984)

. Dehydrohalogenation may lead to the formation of a large number of aliphatic and aromatic compounds. Chlorine radical formation can lead to production of highly chlorinated aromatics including polychlorinated biphenyls. Higher Cl content results in production of greater numb ers and amounts of chlorinated aromatics (Bergman et al., 1984)

.
A

- 1

- 1

- 1 Fate in Air As noted above, CPs lack structural components that absorb light in the UV or visible spectrum, so direct photolysis is not expected to occur. Th e atmospheric half

- life has been estimated at 1

- 2 days (EA, 2009; ECB, 2005) , based on estimate d values for the second order rate constant for reaction with atmospheric hydroxyl radicals for MCCPs with lower chlorine contents between 40 and 56 wt %. EPA/OPPT also estimated atmospheric half

- lives for MCCPs (40 and 70 Cl wt %) calculated using EPI Sui teTM/AOPWINTM (v. 1.92a) that range from about 1 to > 4 days (see Table_Apx A

- 1 ). MCCPs with the shorter chain lengths and higher chlorine contents were calculated to be more persistent.

MCCPs have low estimated vapor pressures (4.5 × 10

- 8 to 2.27 × 10

- 3 Pa at 20

- 25° C) and a Henry’s law constant (HLC) (0.014

- 51.3 Pa × m 3 /mol for C 14

- 17 congener groups) and are not expected to partition to air. They may be transported associated with particulate matter, and have been reported in indoor and outdoor air and house dust (Barber et al., 2005; Fridén et al., 2011; Hi lger et al., 2013) . Wide spread soil contamination and occurrence in artic samp les suggest that MCCPs behave similarly to other chlorinated persistent organic pollutants (POPs) with high57 production volumes and releases, and are subject to long range transport (Dick et al., 2010; Medeiros et al., 2011; Tomy et al., 2000)

.
Table_Apx A

- 1 : Estimated Atmospheric Half Lives Using EPI SuiteTM/AOPWINTM (v. 1.92a) for Varying MCCP Chain Length and Chlorination Percents Based on Wt.

Chain Length 40 Cl Wt % 70 Cl Wt %

C 14 1.0 4.4 C 15 0.8 3.0 C 16 0.8 3.0 C 17 0.8 2.9



### A

- 1

- 2 Biodegradation EPA/OPPT reviewed studies from the open litera ture and submitted to EPA/OPPT including those described in the Canada and EU assessments and referenced in Table_Apx A

- 2 (EC, 2008a; ECB, 2005) to determine biodegradation under a variety of environmental conditions. Some of these studies used modified test conditions to enhance or maximize biodegradation. EPA/OPPT concurs with the EU’s conclusions that under these modi fied test conditions, C 14 41.3 % by wt. Cl and a C 14 45.5% by wt. Cl substances are readily biodegradable. C 15 51% by wt. Cl were found to be inherently degradable and possibly readily degradable in modified OECD 301 and 301D tests. This suggests that CPs with these chain lengths and shorter, and this degree of chlorination and lower, are inherently degradable. More highly chlorinat ed and longer carbon chain CPs (C 14

- 17 51.7 % by wt. Cl, C 14 55 % by wt. Cl, C 14 60.2% by wt. Cl, and C 14 - 17 63.2% by wt. Cl) biodegraded over a range of 2

– 54% in 28 days to 4

– 57% at up to 60 da ys. The most highly chlorinated , (C14

- 17, 63.2 wt% Cl) biodegraded 5% in 28 days and 10% at 60 days in the enhanced biodegradation studies, suggesting that longer chain and higher chl orination can contribute to greater persistence under most environmental conditions. (Van Ginkel, 2014 a and b; Van Ginkel 2010 a

- d; Van Ginkel and Louwerse 2010 a and b)

A

- 1

- 2

- 1 Fate in Wastewater Treatment In its review of the available measured data on MCCPs i n wastewater treatment from data in from other countries, EPA/OPPT determined that CPs are present in the majority of municipal waste water treatment plant (WWTP) influent (Coelh an, 2010; Nicholls et al., 2001; Stevens et al., 2003; Zeng et al., 2013)

. Low water solubility and relatively high partitioning coefficients suggest that most of the MCCPs and LCCPs entering WWTP systems will associate with solids. Some biodegradation of shorter chain, lower chlorinated MCCP congener groups may occur, while long er chain length, more chlorinated congener groups will be resistant to aerobic and anaerobic degradation. Shorter and lower chlorinated congener groups have higher vapor pressure and may be lost to the vapor phase during aeration. WWTP effluent also contai ns some particulate

- associated MCCPs. Because of their low water solubility, little MCCP or LCCP will be in the dissolved phase, and the majority will be removed along with settled sludge. Once associated with the sludge, the CPs will generally be stable i n sludge treatment and remain in the58 residual biosolids. Land application of biosolids will transfer the MCCPs and LCCPs to agricultural and other soils (Nicholls et al., 2001 ; Stevens et al., 2003) . Because 50

- 60% of biosolids in the US are land applied, the majority of MCCPs and LCCPs entering WWTPs may be released to the environment via application to soil, and may be transpor ted from contaminated soil to other locations and media by soil erosion, runoff, and wind borne particulates, and volatilization.

A

- 1

- 2

- 2 Fate in Surface and Groundwater Because they generally have low water solubility, high sorption coefficients, and tend to part ition to solids, MCCPs and LCCPs released to surface water will partition to surficial sediment where they may be buried and removed from potential degradation processes. This explains what is found in the limited monitoring data that exist

- MCCP concentr ations in surface water are generally in the low pg/L range, while sediment concentrations are several orders of magnitude higher (EC, 2008a)

. MCCPs may leach from soil and be transported to groundwater, but low solubility and high sorption will act to kee p dissolved concentrations very low. Facilitated transport with colloids and particulates may occur so that MCCPs can be transported in groundwater, but in general, concentrations in this compartment are expected to be very low. MCCPs that are introduced t o groundwater will tend to partition to the solid phase and not be mobile.

A

- 1

- 2

- 3 Fate in Soil Existing monitoring data suggest that MCCPs are present in soil, probably as a result of atmospheric transport and deposition. Areas near sources, such as land receivi ng wastewater biosolids, manufacturing and processing facilities, and electronic waste processing and recycling facilities are shown to have higher levels than the background from atmospheric deposition (Wang et al., 2013) . MCCPs are expected to be stabl e in soil, and once deposited, could remain/persist in the soil for years or decades. Burial and advective transport away from the site of deposition are the major dissipation processes. No data are available on soil photolysis, although aqueous photolysis data suggest that indirect photolysis may result in degradation to shorter and less chlorinated CP congener groups. No soil biodegradation data exists, but some strains of bacteria that can co

- metabolize MCCPs have been identified (Allpress and Gowland, 1999) . If degradation does occur, it is expected to be slow with T 1/2 of at least months to years.59 Table_Apx A

- 2 : Review of MCCP and LCCP Biodegradation Studies.

Biodegradation Studies on MCCPs (C 14 - 17 ) and LCCPs (C >18 )

Study Authors Publication Date MCCP/LCCP Chemicals Evaluated ( i.e. , C

- length, wt% Cl) Method Study Duration Noteworthy Results and Implications

MCCPs

Van Ginkel 2010 d C 14 45 wt% Cl Closed bottle 28 days Approximately 64% degraded in 28 days Based on oxygen demand

Van Ginkel 2010 b C 14

- 17 45.6 wt% Cl Closed bottle 28 days Approximately 51% in 28 days and 63% in 42 days degradation Based on oxygen demand

Van Ginkel 2010 c C 14

- 17 51.7 wt% Cl Closed bottle 28 days Approximately 27% degradation in 28 days and 57% after 60 days Based on oxygen demand

Van Ginkel 2010 a C 14

- 17 63.2 wt% Cl Closed bottle 28 days Approximately 5 % degradation after 28 days and 10% after 60 days.

Van Ginkel and Louwerse 2010 a C 14 41.3

- 60.2 wt% Cl % Closed bottle with river water and sludge inoculum 28 days Approximately 66% (41.3 wt% Cl) to 11 (60 .2 wt% Cl) degradation in 28 days respectively

Van Ginkel and Louwerse 2010 b C 14 41.3

- 50 wt% Cl Batch
reactor 21 and 105 days 41.3 wt% Cl: 79 % degradation in 21 days and 94% at 105 days

50 wt% Cl: 14% degradation by 21 days 5 wt% Cl in 80 days based on quantitation of released chloride
Conclusions: Quantification of degradation was by oxygen uptake or chloride release. No information on the chemical distribution in the test material or degradates was provided.

These studie s used modified test conditions to enhance or maximize biodegradation. Under these modified test conditions, C 14 41.3 wt% Cl and a C 14 45.5 wt% Cl substances are readily biodegradable. More highly chlorinated and longer carbon chain CPs (C 14

- 17 51.7 wt% Cl , C 14 55 wt% Cl, C 14 60.2 wt% Cl )

, biodegraded over a range of 2

– 54% in 28 days to 4

– 57% at up to 60 da ys. The most highly chlorinated , ( C 14

- 17 63.2 wt% Cl) biodegraded 5% in 28 days and 10% at 60 days in the enhanced biodegradation studies, suggesting that longer chain and higher chlorination can contribute to greater persistence under most environmental conditions.

Van Ginkel 2014 b C 15 51 wt% Cl Closed bottle (301D) 60 day 43 % and 63 % degradation at 28 and 60 days60 Biodegradation Studies on MCCPs (C 14 - 17 ) and LCCPs (C >18 )

Study Authors Publication Date MCCP/LCCP Chemicals Evaluated ( i.e. , C

- length, wt% Cl) Method Study Duration Noteworthy Results and Implications

Van Ginkel 2014 a C 15 51wt% Cl Closed bottle (301) 60 day 37 % and 57 % degradation at 28 and 60 days
Conclusions: Unlike the 2010 series of studies, these most recent biodegradation studies did not have significant protocol modifications and the C 15 51 wt% Cl were found to be inherently degradable and possibly readily degradable in OECD 301 and 301D tests.

Madeley and Birtley 1980 C 14

- 17 mixed product, 40 wt% Cl BOD test 25 days Approximately 15.5 % degradation as measured by theoretical BOD in non

acclimated samples and 22.5 % degradation in acclimated samples. 1

Madeley and Birtley 1980 C 14

- 17 mixed product, 45 wt% Cl BOD test 25 days Approximately 10% degradation as measured by theoretical BOD in non

- acclimated samples and 30% degradation with acclimated soil microbes added. 1

Madeley and Birtley 1980 C 14

- 17 mixed product, 52 wt% Cl BOD test 25 days Approximately 4% degradation as measured by theoretical BOD in non

- acclimated samples and 6% degradation with accli mated soil microbes added. 1

Madeley and Birtley 1980 C 14

- 17 mixed product, 58 wt% Cl BOD test 25 days No significant degradation

Conclusions: The data from Madeley and Birtley suggests the potential for biodegradation but has significant limitations. The BOD studies were done on mixed products. No attempt was made to determine which specific congeners were degraded or the reaction products. No identification of the congeners present was provided. The degradation was estimated from the BOD but other compounds may have contributed to the ThBOD in the bottles. BOD measurements are highly variable as evidenced by the decrease in the C 20

- 30 42 % of > 50 % between day 20 and 25.

1 The ThOD (theoretical oxygen demand) was estimated (ThOD (g O 2 /g substance) = 16[2 × c+0.5 × (h

cl)]/mw; where c=number of carbon atoms, h=number of hydrogen atoms, cl=number of chlorine atoms and MW = molecular weight). This is questionable for a product containing mixture of congeners as was used in all studies.

LCCPs61 Biodegradation Studies on MCCPs (C 14 - 17 ) and LCCPs (C >18 )

Study Authors Publication Date MCCP/LCCP Chemicals Evaluated ( i.e. , C

- length, wt% Cl) Method Study Duration Noteworthy Results and Implications

Madeley and Birtley 1980 C 20

- 30 mixed product, 42 wt% Cl BOD test 25 days Approximately 7.5 % degradation as measured by theoretical BOD in non

acclimated samples and 23 % degradation with acclimated soil microbes added. 1

Madeley and Birtley 1980 C 25 “chlorinated pentacosane” 14 C on central carbon 8 weeks (mean) 11% of 14 C

- released as CO 2. Non

- acclimated microbes.

Conclusions: The data from Madeley and Birtley suggests the potential for biodegradation but has significant limitations. The BOD studies were done on mixed products. No attempt was made to determine which specific congeners were degraded or the reaction products. No identification of the congeners present was provided. The degradation was estimated from the BOD but other compounds may have contributed to the ThBOD in the bottles. BOD measurements are highly variable as evidenced by the decrease in the C 20

- 30 42 % of > 50 % between day 20 and 25.

1 The ThOD (theoretical oxygen demand) was estimated (ThOD (g O 2 /g substance) = 16[2 × c+0.5 × (h

cl)]/mw; where c=number of carbon atoms, h=number of hydrogen atoms, cl=number of chlorine atoms and MW = molecular weight). This is questionable for a product containing mixture of congeners as was used in all studies.

Hildebrecht 1972 C 20

- 30 mixed product, 42 wt% Cl BOD test 5 days 25% degradation. Degradation was estimated by the authors as the % of the theoretical BOD based on the total carbon content of the test solution. Substances other than the chlorinated paraffin contributed to this total carbon content.

Hildebrecht 1972 > C 20

- 30 mixed product, 70 wt% Cl BOD test 5 days 2% degradation. Degradation was estimated by the aut hors as the % of the theoretical BOD based on the total carbon content of the test solution. Substances other than the chlorinated paraffin contributed to this total carbon
content.

Hildebrecht 1972 > C 20

- 30 mixed product, 70 wt% Cl BOD test 5 days 65% degradation. Degradation was estimated by the authors as the % of the theoretical BOD based on the total carbon content of the test solution. Substances other than the chlorinated62 Biodegradation Studies on MCCPs (C 14 - 17 ) and LCCPs (C >18 )

Study Authors Publication Date MCCP/LCCP Chemicals Evaluated ( i.e. , C

- length, wt% Cl) Method Study Duration Noteworthy Results and Implications

paraffin contributed to this total carbon content.
Conclusions: As desc ribed by the (EA, 2009) , Hildebrecht’s results are questionable (Hildebrecht, 1972) . This report is not available so it cannot be reviewed directly, but others have reported that it provided limited details. A su rfactant, other carbon sources, and nutrients were added that may have contributed BOD. The extent of degradation was determined by the comparing the oxygen consumption in the test with the theoretical oxygen demand (ThOD) based on oxidation to CO 2 of the total organic carbon present in the solution from all sources. This estimation of ThOD does not take into account oxygen consumption by other compounds or the unknown composition. Of the CPs in the mixture, as the UK report concludes, “It is not possible t o draw definite conclusions as to the degradability of the chlorinated paraffins in these tests” (EA, 2009)

.
Hoechst AG 1976 and 1977 C 18

- 20 , 35 wt% Cl BOD test 5 days 0.7% degradation

Hoechst AG 1976 and 1977 C 18

- 20 , 44 wt% Cl BOD test 5 days < 1.2% degradation

Hoechst AG 1976 and 1977 C 18

- 20 , 49 wt% Cl BOD test 5 days < 2.3% degradation

Hoechst AG 1976 and 1977 C 18

- 20 , 52 wt% Cl BOD test 5 days < 0.6% degradation

Conclusions: The Hoechst reports from early industry studies are not available so it is not possible to directly review the data (Hoechst, 1976, 1977) . Others (EA, 2009) have reported the limitations of the studies. Limited details of the studies were apparently reported by Hoechst. These tests were done on mixtures of congeners with unknown composition. They reported tha t the majority of the CPs were removed by sorption on to the solids so no degradation may have occurred that would have been detected as BOD. The tests were run for 5 days using non

- acclimated sludge microbes so degradation may have been possible but had n ot yet occurred.

Omori et al. 1987 C 24.5 H 44.5 Cl 6.5 , 40.5 wt% Cl Chloride release 48 hours 9.9% degradation using bacterial strain HK

- 3; 13% H15

- 4; 2.2% HK

- 6; 3.5% HK

- 8; 33% using mixed bacterial culture (HK

- 3, HK

- 6, HK

- 8 and HK

- 10)

Omori et al. 1987 C 24.5 H 41 Cl 10, 50 wt% Cl Chloride release 48 hours 3% degradation using bacterial strain HK

- 3; 9% H15

- 4; 1.8% HK

- 6; 2.6% HK

- 8

Omori et al. 1987 C 24.5 H 30 Cl 21, 70 wt% Cl Chloride release 48 hours 2.6% degradation using bacterial strain HK

- 3;63 Biodegradation Studies on MCCPs (C 14 - 17 ) and LCCPs (C >18 )

Study Authors Publication Date MCCP/LCCP Chemicals Evaluated ( i.e. , C

- length, wt% Cl) Method Study Duration Noteworthy Results and Implications

12% H15

- 4; 1.4% HK

- 6; 1.7% HK

- 8; 15% using mixed bacterial culture (HK

- 3, HK

- 6, HK

- 8 and HK

- 10)
Conclusions: Omori e t al. (1987) showed the potential for biodegradation using pure and mixed cultures in short (48 hour) incubations. No information on the starting mixtures were provided except average compositions. No data on the products were reported. Loss of Cl sugge sts dechlorination can occur and that lower Cl content or shorter chain lengths may be produced. Allpress and

Gowland 1999 C 18

- 20 , 48 wt% Cl Chloride release 71 days 11% degradation using Rhodococcus sp. bacteria Allpress and

Gowland 1999 C > 20 , 42 wt% Cl Chloride release 71 days 14% degradation using Rhodococcus sp. bacteria
Conclusions: Allpress and Gowland (1999) also showed that CPs have the potential to biodegrade using pure culture. They used mixed congener products and did not provide any information on the composition of the starting material or the degradation products. They found that the Rhodococcus sp. was able to use CPs as carbon source as well as an energy source.

1 The ThOD (theoretical oxygen demand) was estimated (ThOD (g O 2 /g substance) = 16[2 × c+0.5 × (h

- cl)]/mw; where c=number of carbon atoms, h=number of hydrogen atoms, cl=number of chlorine atoms and MW = molecular weight). This is questionable for a product containing mixture of congeners as was used in all studies.



### A

- 2 BIOCONCENTRATION AND BIOACCUMULATION
EPA/OPPT’s review of measured d ata on bioaccumulation of MCCPs and LCCPs are somewhat limited and conclusions vary with type of CP product and species evaluated (Bengtsson et al., 1979; CPC, 1980, 1983a, 1983b; Fisk et al., 1999; Fisk et al., 1998; Houde et al., 2008; Madeley and Maddock, 1983a, 1983b; Madeley and Thompson, 1983; Renberg et al., 198 6; Thompson et al., 2000)

. The limited measured data on MCCPs and LCCPs, informed by data on SCCPs, suggests that bioaccumulation is a function of chain length and degree of chlorination (see Table_Apx A

- 3 ). Some MCCP chemicals with intermediate chain length and chlorination may be absorbed and retained. The available evidence for MCCP congener groups with intermediate chain lengths and chlorination suggests that some may have BCFs or BAFs great er than 1000 or 5000 (EC, 2008b; ECB, 2008) . This suggests that some congener groups in MCCP products may be bioaccumulative or very bioaccumulative. In conclusion, some MCCP congener groups present in the products are both very persistent and very bioaccumulative.64 Additional evidence for bioaccumulation of MCCPs is provided by Houde et al. (2008) . Field

derived log BAFs for MCCPs (C 14

- 15 ), ranging from 6.5 to 7.3, were reported for several Lake Ontario aquatic species from multiple trophic levels. Canada’s assessment of MCCPs also indicates that modeled BAFs for a number of MCCPs (using the Modified Gobas BAF Model

w ith assumption of no metabolism), were all above 5000, suggesting high to very high bioaccumulation (EC, 2008a) . Evidence of bioaccumulation in sediment

- dwelling organisms is also provided in a study by Fisk et al. (1998)

. Biota

- sediment accumulation factors (BASFs) ranging from 0.6 to 4.4 were reported for oligochaetes, which indicate bioaccumulation of MCCPs from sediment to biota (USEPA, 2009).

The Houde et al. (2008) study also provides evidence of biomagnification of MCCPs. BMFs derived for food chains in Lake Ontario and Lake Michigan ranged from 1 to 15. More specifically, large BMFs were observed for all MCCP chain lengths in Lake Ontario, and for C 14 MCCPs in Lake Michigan, indicating biomagnification. BMFs (2.4

– 7.7) were also above 1 for smelt and lake trout in Lake Michigan.

In laboratory studies with rainbow trout and oligochaetes, lipid

- normalized equilibrium BMFs estimated from a first

- or der bioaccumulation model for constant dietary exposure ranged from 0.4

- 5.0 (Fisk et al., 1996; Fisk et al., 2000; Fisk et al., 1998)

. Most of the laboratory

- based BCF studies (Bengtsson et al., 1979; CPC, 1980, 1983a, 1983b; Fisk et al., 1999; Fisk et al. , 1998; Houde et al., 2008; Madeley and Maddock, 1983a, 1983b; Madeley and Thompson, 1983; Renberg et al., 1986; Thompson et al., 2000) , were reported to have been conducted at MCCPs concentrations above the water solubility limit and hence likely underestimate the true BCF (summarized in Appendix C). Furthermore, acetone as a solvent in these tests, so they do not adhere to OECD guidelines. Nonetheless, some BCF values estimated from these studies indicate MCCPs and LCCPs are bioaccumulative ( e.g. , bleak and rainbow trout (32

- 2856) and BCF of 6920 for common mussel).65 Table_Apx A

- 3 : Review of MCCP and LCCP Bioaccumulation Studies

Bioaccumulation studies on MCCPs (C 14

- 17 ) and LCCPs (> C 18 ) MCCPs

Study Authors Publication Date MCCP/LCCP Chemicals Evaluated ( i.e. , C

- length, % Cl) Method Study Duration Noteworthy Results and Implications

Houde et al. 2008 C 14

- 15 Only BAF = ([predator]/[water (filtered)]);

BMF) = [predator]/[prey]) where the concentrations in predator and prey are on a lipid basis Three sampling periods: October 2000, June 2002 , and July 2004. Issues related to the temporal variability of water concentrations over the period of biota sampling (1999

– 2004) in this study have been raised (ECB, 2005; EC, 2008) contributing to uncertainties associated with the reported BAF values.

Log BAF = Plankton: C14=6.2; C15=6.6; ∑=6.5 Alewife: C14=7.0; C15=6.8; ∑=6.9 Sculpin: C14=7.4; C15=7.2; ∑=7.3 Rainbow smelt: C14=7.4; C15=7.1; ∑=7.2 Lake trout: C14=6.8; C15=6.5; ∑=6.6

BMF (Lake Ontario) = 0.25 (lake trout

– alewife); 0.14 (lake trout

– smelt); 8.7 (sculpin

– Diporeia)

BMF (Lake Michigan) = 0.22 (lake trout

– alewife); 0.94 (lake trout

– sculpin); 0.88 (sculpin

– Diporeia)66 Thompson et al.; as summarized in ECB, 2005 2000 n

- pentadecane

- 8

- 14 C, 51% Cl mixed with a

non

- radio

labelled C 14

- 17 , 51% Cl chlorinated paraffin Freshwater; flow

through; acetone solvent used; 35 days Steady

- state may not have been achieved, so kinetic BCF data considered more reliable.

BCF = 860 L/kg at 35 days when exposed at 0.9
μ g/L BCF = 265 L/kg at 35 days when exposed at 4.9
μ g/L kinetic BCF = 1,087 L/kg at 35 days when exposed at 0.9 μ g/L kinetic BCF = 349 L/kg at 35 days when exposed at 4.9 μ g/L CPC (Madeley et al.); as summarized in ECB, 2005 1983 commercial product mixed with a n

pentadecane

- 8

- 14 C chlorinated to a similar degree freshwater; flow

through; rainbow trout; acetone solvent used 60 days Concentrations above water solubility; hence, water concentrations may be overestimated and BCF underestimated. Uncertainty as to w hether steady

- state was reached.

BCF = 32

- 45 L /kg on a wet weight basis when exposed at 1.05 mg/ L ; BCF = 42

- 67 l/kg on a wet weight basis when exposed at the 4.5 mg/ L CPC (Madeley and Pearson); as summarized in ECB, 2005 1980 C 14

- 17 , 45% Cl freshwater; flow

through; rainbow trout; 28 days Measured water concentrations questionable; water concentrations may be overestimated and BCF underestimated.

BCF = 50

- 60 L /kg based on nominal exposure concentrations BCF = 280

- 600 L /kg based on measured water concentrations Madeley and Maddock; as summarized 1983 Total MCCPs Bioconcentration factors. MCCPs concentrations were above the water No Information BCF = 32

- 2856 for common mussel, bleak and rainbow trout. May not have reached steady

- state.67 in ECB, 2005 solubility limit, using acetone as the

co

- solvent in the test solutions, and hence are not in compli ance with OECD guideline requirements Fisk et al. 1999 Average formula: C 14 H 23.3 Cl 6.7 , 55% Cl freshwater; medaka
eggs 20

- days Uncertainty as to whether steady

- state was reached; hence BCFs probably represent lower limit of true value

BCF = 32

- 680 L/kg Fisk et al. 1998 14 C 16 35 % Cl and 14 C 16 69 % Cl Lake sediments were spiked and worms added after 18 and 32 No Information Kinetic BAF probably represents the upper limit of the true bioaccumulation factor

14 C 16 35 % Cl 14

- day BSAFss = 0.7 Kinetic BSAF = 4.4

14 C 16 69 % Cl 14

- day BSAFss = 0.2 Kinetic BSAF = 0.6

Bengtsson et al. 1979 C 14

- 17 , 50% Cl seawater; semi

static; bleak; acetone solvent used 14 days Measured water concentrations questionable; water concentrations may be overestimated and BCF underestimated.

BCF
∼ 40 L/kg Madeley and Thompson;

as summarized 1983 commercial C 14

- 17 , 52 % Cl seawater; flow

through; mussel acetone solvent used 60 days BCFs = 2,182 L/kg (parent compound analysis) or 2,856 L/kg ( 14 C

- measurements) when exposed to 0.22 mg/ L68 in ECB, 2005 BCF = 339 L/kg (parent compound analysis) or 429 L /kg ( 14 C measurements) when exposed to 3.8 mg/ L

. Renberg et al. 1985 C 16 H 30.7 Cl 3.3 (34% Cl) and C 12 H 16 Cl 9.8 (68.5% Cl) mixture synthesized with
14 C radiolabel Flow through exposure to mussel ( Mytilus edulis )

C 16 - 0.13 and 5.0 μg/L C 16 - 28 day uptake C 12 - 21 day uptake followed by 28 day depuration Steady state BCF about 7 000 for C 16 and 140,000 for C 12 based on 14 C quantification.

No chemical specific analysis for CPs. Metabolism and accumulation of degradation products may have accounted for high values. LCCPs Bengtsson et al. ; as summarized in ECB, 2005 1979 C 18

– 26 concentrations were above the water solubility limit and hence are not in compliance with OECD guideline requirements No Information Concentrations above water solubility; hence, water concentrations may be overestimated and BCF underestimated. Uncertainty as to whether steady

- state was reached.

BCF reported = 8

- 16 L/kg69 

## Appendix B ECO TOXICITY STUDY SUMMARIES

B

- 1 MCCP E COTOXICITY DATA
**B**

**- 1** 

**- 1**  **Acute Fish Toxicity** (1) A series of 96

- hour acute fish toxicity studies were conducted by Mayer an Ellersieck (1986) with Paroil 1048 (50

- 52% Cl, C 15 H 26 C l6 ) similarly to ASTM (1980). Bluegill sunfish ( Lepomis macrochirus ) and yellow perch ( Perca flavescens ) were exposed to the test substance in a flow

through test system and channel catfish ( Ictalurus punctatus ) and rainbow trout ( Oncorhynchus mykiss ) were exposed to the test substance in a static test system. Solvent use was not specified for this compound. The average pH level was between 7.4 and 7.5 for all tests. Test temperature was 12 oC for bluegill sunfish, rainbow tr out, and yellow perch and 20 oC for channel catfish. Dilution water hardness was 44 mg CaCO 3 /L in the rainbow trout and channel catfish test system and 314 mg CaCO 3 /L for the bluegill sunfish and yellow perch test system. Reported effect levels are conside red to be nominal with LC 50 values of >10 mg/L for bluegill sunfish, channel catfish, and yellow perch and >0.011 mg/L for rainbow trout; all values are greatly above the limit of solubility.

EPA/OPPT Conclusion Using a weight

- of

- evidence approach, these studies were considered acceptable to characterize the acute fish toxicity endpoint. 96

- hr LC 50 = NES

(2) A 96

- hour acute fish toxicity study was published by Linden et al. (1979). Groups of 10 Bleak ( Alburnus alburnus ) were exposed to six nominal unspec ified concentrations of Cereclor S52 ® (C 14

- 17 , 52% Cl), Chloroparaffin huls 40G (C 15.5 , 40% Cl), and Witaclor 50 (C 14

- 17 , 50% Cl) under static test conditions. Salinity was 7 ppt, pH was 7.8, temperature was 10 oC, and dissolve oxygen was considered by stu dy authors to be satisfactory. EPA/OPPT requires reporting of dissolved oxygen concentrations to determine study adequacy. EPA/OPPT also does not consider the test species, the bleak, a standard test species. The 96

- hour fish LC 50 values

were > 10,000 mg/L, >5,000, and >5,000 for Cereclor S52 ® , Chloroparaffin huls 40G, and Witaclor 50.

EPA/OPPT Conclusion Given effect levels observed in Mayer and Ellersieck (1986) and the reported water solubility of medium chain paraffins, these studies were considered acceptable using a weight

- of

- evidence approach to characterize the acute saltwater fish toxicity endpoint. 96

- hr LC 50 = NES

(3) Bengtsson et al. (1979) also studied the toxicity of a medium

- chain chlorinated paraffin to Bleak ( Alburnus alburnus ) as part of a bioaccumulation study. The chlorinated paraffin tested was a C14

- 17, 50% wt. Cl substance. The tests were performed at 10 °C using a semi

- static procedure in which the test solutions containing 125 μg/L of the substance were renewed every
tw o to three days over the 14

- day exposure period. The water used in the experiment was Baltic Sea water with a salinity of 7‰, and acetone was present in all aquaria, including controls at a concentration of 0.1 ml/l. The fish used in the experiment had an average weight of 4.5 g and were not fed during the exposure period. Six groups of 15 fish were used for both the exposure70 and control solutions. No mortality or effect on behavior was seen in fish exposed to the medium

- chain chlorinated paraffin during th e test.

EPA/OPPT Conclusion This data review was part of a BAF study and as such will be used as a weight of evidence to support other data for this category of organisms.



### B

**- 1** 

**- 2**  **Acute Aquatic Invertebrate Toxicity** (1) A 48

- hour acute Daphnia magna toxicity st udy was conducted by Thompson et al. (1996) according to OECD TG 202 (1984) with GLP compliance using a static test system. The test substance was identified as Cereclor S52 ® , a C 14

- 17 chlorinated paraffin with 52% chlorination that contained 0.3% epoxy so ya bean oil stabilizer as well as a small amount of radiolabelled n

pentadecane

- 8

- 14 C (51% chlorinated). Four replicates of 5 Daphnia magna Straus (<24 hours old) were exposed to nominal concentrations of 0 (dilution water control), 0 (solvent control), 0. 0032, 0.0056, 0.01, 0.018, 0.032, 0.056, and 0.1 mg/L test substance in acetone (0.1 mL/L). Test solutions were prepared by adding the appropriate stock solution to dilution water while continuously and vigorously stirring with a magnetic follower. Appeara nce of test solutions was not provided. Corresponding mean measured concentrations determined by radiochemical methods were 0.0025, 0.0041, 0.0094, 0.015, 0.024, 0.047, and 0.095 mg/L. Daphnid loading was 25 daphnids/L. Over the course of the study dissolv ed oxygen concentrations remained between 9 and 9.2 mg/L, pH remained within 8 and 8.1, and temperatures were 20 ±1 oC. Dilution water had a total water hardness of 248 mg CaCO 3 /L. At 48 hours, 0%, 45%, 90%, 75%, 85%, 100%, and 100% immobilization was obse rved at the mean measured concentrations of 0.0025, 0.0041, 0.0094, 0.015, 0.024, 0.047, and 0.095 mg/L, respectively. Red coloration on parts of the exoskeleton was observed in animals exposed to each of the test substance treatments, which the laboratory notes as being of an uncertain significance.

EPA/OPPT Conclusion The study is acceptable. 48

- hr EC 50 = 0.0059 mg/L

(2) A 48

- hour acute Daphnia magna toxicity study was conducted the University of Bremen, Department of Physical and Environmental Chemist ry with CP 52 (C 12

- 18 , 52% chlorination) according to DIN 38412 by Koh and Thiemann (2001). Study methods were not fully characterized. Additional communications with the study author Wolfram Thiemann clarified that a static test system was used with nominal test concentrations. Based on communications with the study author, local (Bremen, Germany) tap water was used without adjustments. Presumed pH was between 5 and 6 and water hardness was between 35.7 and 53.5 mg CaCO 3 /L. Ambient laboratory air temp erature was around 21 oC. The solvent acetone was used to maintain test substance in solution. Floating effects at the surface of the water were observed in individual cases due to undissolved oil slicks, but communications with the study author noted that there was no significant loss of daphnids due to mechanical trapping since most daphnid swam away from these occasional slicks observed at the higher test concentrations.

EPA/OPPT Conclusion This study is considered acceptable. 48

- hr EC 50 = 0.052 mg/L71 (3) A 48

- hour acute Daphnia magna toxicity study was conducted by Thompson et al. (1994) according to OECD TG 202 (1984) with GLP compliance using a static test system. The test substance was identified as Cereclor S52 ® , a C 14

- 17 chlorinated paraffin with 52% chlorination that was mixed with an equal weight of radiolabeled n

- pentadecane

- 8

- 14 C (51% chlorinated). Four replicates of 5 Daphnia magna (<24 hours old) were exposed to 0% (dilution water control), 6.3%, 12.5%, 25%, 50%, and 100% of stock solution c ontaining test substance. The test substance was prepared in solution by (1) combining 0.75 g test substance and 25 mL acetone to a borosilicate glass conical flask, (2) evaporation of the acetone using a stream of nitrogen, (3) addition of 1.5 L dilution water, (4) stirring for three days, and (5) filtration of the aqueous phase. Radiochemical methods were used to determine the concentration of test substance in solution. Nominal concentrations of 0 (dilution water control), 0.14, 0.28, 0.55, 1.1, and 2.2 mg/L were within 86

- 100% of measured concentrations. Concerns regarding test solution preparation methods and analytical technique were identified by the submitter that included increasing the level of more soluble impurities (i.e., short chain chlorinated paraffins), questionable analytical monitoring results due to the presence of radio

- labeled impurities, and abnormally low recovery of the chlorinated paraffin into hexane. Over the course of the study dissolved oxygen concentrations remained between 8 an d 9 mg/L, pH remained within 8 and 8.1, and temperatures were 20 ±1 oC. Dilution water had a total water hardness of 237 mg CaCO 3 /L. Observed immobilization was limited to the highest test concentration (100% solution) with 55% immobilization.

EPA/OPPT C onclusion The study is unacceptable since EPA/OPPT agrees that methods used to prepare the test solution and analyze the test concentrations were questionable.

4) The following study summary (Frank, 1993; Frank and Steinhäuser, 1994) provided in the 2005 European Chemical Bureau Risk Assessment of MCCP was considered supportive of the aquatic invertebrate hazard determination. The chlorinated paraffin used in these studies was a commercial C 14

- 17 product with a 52% by weight chlorine content. Daphnia magna were exposed to nominal concentrations of either 100 mg/L or 10,000 mg/L. The 100 mg/L solution was sonicated for 1 hour and then left to stand in the dark for 48 hours before use. The 10,000 mg/L solution also stood for 48 hours in the dark before use, b ut this time without sonication. After this period, both solutions were filtered firstly with glass filters and then with membrane filters to remove undissolved test substance. The concentrations of medium

- chain chlorinated paraffin in the water soluble fr actions were then determined by AOX (adsorbable organic halogen) analysis (detection limit of 10 μg/L Cl was equivalent to around 20 μg/L of the chlorinated paraffin). This analysis showed that the concentration of chlorinated paraffin present in the water soluble fraction was around 0.404

- 0.500 mg/L for the 10,000 mg/L nominal solution and 0.071

- 0.142 mg/L L for the 100 mg/L stock solution. The acute (48

- hour) toxicity tests were carried out using dilutions of the two prepared water soluble fractions. The method used was DIN 38 412, Teil 11, which is equivalent to OECD 202.

In the tests using the water soluble fraction from the 100 mg/L nominal solutions no toxicity was seen at concentrations up to the undiluted stock solution (i.e. no effects up to aroun d 0.071

- 0.142 mg/L). In experiments using the water soluble fraction from the 10,000 mg/L stock solution, an EC 0 of 0.140 mg/L (also reported as 0.100

- 0.110 mg/L in the paper) and an EC 25 of 0.423 mg/L (also reported as 0.420

- 0.470 mg/L in the paper) was d etermined (maximum mortality seen was 25%) (Frank, 1993). The latter results for the 10,000 mg/L stock solution were reported by Frank and Steinhäuser (1994) as EC 0 = 0.140 mg/L and EC 25 = 0.339 mg/L, and it was noted that some72 of the Daphnia were floating on the surface of the test solution. In the later study (Frank and Steinhäuser, 1994), the results of further acute toxicity studies were reported using the same test method. An EC 50 of 0.037 mg/L and an EC 0 of 0.009 mg/L were determined using the water

s oluble fraction from the 100 mg/L stock solution and no toxic effects were seen in tests with the water soluble fraction from the 10,000 mg/L stock solution (approximately EC 0 ≥0.525 mg/L). The authors noted that the effects seen in the acute tests showed poor reproducibility, probably because effects were seen only around the water solubility limit of the substance. However, the authors thought that the possibility of undissolved droplets affecting the results could be ruled out, as floating Daphnia were o nly sporadically observed in the test.

EPA/OPPT Conclusion Tthe results of these studies should be treated with caution, as the effects were mainly seen in the saturated solutions only. 48

- hour EC 50 = 0.037 mg/L; 100 mg/L stock (Frank and Steinhäuser, 199 4)

5) The following study summary (Thompson and Gore, 1999) provided in the 2005 European Chemical Bureau Risk Assessment of MCCP was considered support ive of the aquatic invertebrate hazard determination. The acute toxicity of C14

- 17, 52% wt. Cl substanc e was tested using the freshwater crustacean Gammarus pulex and the freshwater daphnid , Daphnia
magna . The medium

- chain chlorinated paraffin used was dissolved in acetone and then added to beakers in two separate studies containing either Gammarus or D. ma gna to give nominal concentrations of 0.1, 0.32, and 1.0 mg/L. A control and solvent control (containing 0.1 mL/L acetone) were also run. The tests were carried out for 96 hours at 15 oC, with the solutions being renewed after 48 hours. The water used in t he study had a hardness of 220 mg/L as CaCO 3 and had a pH of 8.0

- 9.2. No mortalities of the Gammarus were seen in any of the test substance solutions or control. One animal died in the solvent control. Therefore, no significant toxic effects were seen with the medium

- chain chlorinated paraffin over the concentration range tested. This contrasted markedly to the situation when Daphnia magna were exposed usin g the same test system at 20 oC over 48 hours, where complete immobilization was seen at the lowest
te st concentration (0.1 mg/L).

EPA/OPPT Conclusion The high immobilization rate observed in Daphnia magna in this study appears consistent with the other studies and Gammarus pulex appear to be a less sensitive to medium chained chlorinated paraffins then Daphnia
magna

. EPA/OPPT reserves judgment on the acceptability of this study until further details become available. 96

- hr EC 50 ( Gammarus pulex ) > 1 mg/L 48

- hr EC 50 ( Daphnia magna ) < 0.1 mg/L

6) The following study summary (Tarkpea et al., 1981; as quoted in WHO, 1996) provided in the 2005 European Chemical Bureau Risk Assessment of MCCP was considered support ive of the aquatic invertebrate hazard determination. The results of tes ts with the brackish water harpacticoid Nitocra spinipes have been reported (Tarkpea et al., 1981). No other details of the test were reported but the test method was probably the same as reported by Tarkpea et al.73 (1986), where a static method was employe d using water of salinity 7‰ at a temperature of 20

- 22 oC without aeration, probably using acetone as cosolvent.

EPA/OPPT Conclusion The results are considered supportive to address aquatic invertebrate acute toxicity. 96

- hour LC 50 = 9 mg/L (C14

- 17, 45% wt Cl) 96

- hour LC 50 >10,000 mg/L (C14

- 17, 52% wt. Cl)



### B

**- 1** 

**- 3**  **Algae Toxicity** (1) A 72

- hour algae toxicity study was conducted by the University of Bremen, Department of Physical and Environmental Chemistry with CP 52 (C 12

- 18 , 52% chlorination) according to DIN 38412 by Koh and Thiemann (2001). Study methods were not fully characterized. Additional communications with the study author Wolfram Thiemann clarified that a static test system was used with nominal test concentration s. Scenedesmus subspicatus were exposed to the test substance and cell density was determined using a particle counter. Based on communications with the study author, local (Bremen, Germany) tap water was used without adjustments. Presumed pH was between 5 and 6 and water hardness was between 35.7 and 53.5 mg CaCO 3 /L. Ambient laboratory air temperature was around 21 oC. The solvent acetone was used to maintain test substance in solution. Effects were calculated based on growth rate. No effects were observed up to 0.1 mg/L.

EPA/OPPT Conclusion Due to deficiencies/missing details in the study methods, the study alone was not acceptable to characterize aquatic toxicity to plants. 72

- hr NOEC = 0.1 mg/L (Highest Test Concentration)

(2) A 96

- hour algae toxicity study was conducted according to OECD TG 201 (2006). The test substance was a commercial product of a C 14

- 17 chlorinated paraffin with 52% chlorination that contained 0.3% epoxy soya bean oil stabilizer as well as a small amount of radiolabeled n

pentadec ane

- 8

- 14 C (51% chlorinated). Selenastrum capricornutum were exposed to nominal concentrations of 0 (dilution water control), 0 (solvent control), 0.1, 0.18, 0.32, 0.56, 1, 1.8, and 3.2 mg/L test substance in the solvent acetone. Six replicates were tested for each control and three replicates were tested for each treatment. A mean measured concentration of 0.49, 0.77, and 1.2 mg/L was determined using radiochemical analysis for the nominal concentrations of 1, 1.8, and 3.2 mg/L, respectively, but effects we re reported based on nominal test concentrations. At the start of the test, the pH was 7.4

- 7.5, but had reached 10.0

- 10.3 by the end of the test. The shift in pH was thought to be a function of the high control growth rates observed in the test according t o the study summary. The section

- by

- section coefficient of variation for the solvent control remained below 35% indicating acceptable control growth rates throughout the duration of the study.

EPA/OPPT Conclusion The maximum inhibition in the growth rate and biomass seen was 3% and 18%, respectively, but a dose response relationship was not seen. The nominal NOEC was 0.1 mg/L and the nominal LOEC based on 18% biomass inhibition was 0.18 mg/L. A GMATC of 0.134 mg/L was calculated. The study was considered acceptable. 72

- hour EC 50 = >3.2 mg/L (nominal); 1.2 mg/L (mean measured).74 96

- hour EC 50 = >3.2 mg/L (nominal); 1.2 mg/L (mean measured). 72

- hr NOEC b = 0.1 mg/L 72

- hr LOEC b = 0.18 mg/L 72

- hr GMATC = 0.134 mg/L



### B

**- 1** 

**- 4**  **Chronic Fish Toxicity** (1) A 60

- day fish toxicity study was conducted by Brixham Laboratories in 1983 with radio

labeled chlorinated (52%) n

- pentadecane (Trade Name: Cereclor S52 ® ) under flow through testing conditions. A full non

- CBI study report was submitted under TSCA in 1983 as DCN 40

- 83321 84 (OTS Fiche 0507258). Two replicates of 3 immature rainbow trout ( Salmo gairdneri ) per concentration were exposed to nominal concentrations of 0 (dilution water control), 0 (acetone control, 500 ppm), 1, or 5.6 mg/L in 500 ppm acetone. Corresponding mean measured concentrations were 0, 0, 1.05, and 4.5 mg/L. Test concentrations were determined by radio activity measurements. Flow rate of the test system was 0.25 mL/minute for exposure concentrations. No mortality or adverse sub

- lethal behavioral effects w ere observed for the duration of the 60

- day exposure period. Effects observed were limited to the highest test concentration and involved sluggish movements. The measured NOEC was identified as 4.5 mg/L. In addition to the hazard assessment, the submitter provided an assessment of bioconcentration which indicated that analytically determined exposure concentrations of 1.05 and 4.5 mg/L resulted in fish tissue concentrations of 34 and 190 μg/g wet weight, respectively.

EPA/OPPT Conclusion This study appears to have been previously reviewed by EPA/OPPT in 1985. The previous conclusion that “a fish full life cycle toxicity test or modification thereof is needed to address the effects of CPs present in fish eggs duri ng embryonic development” (US EPA Memorandum, 1991) is still relevant for MCCPs. Thus, this study is considered unacceptable to characterize chronic population

- level effects in fish. 60

- day NOEC = 4.5 mg/L

(2) A 28

- day fish toxicity study was conducted by Brixham Laboratories in 1978 with a C 14

- 17
c hlorinated paraffin having 45% chlorination under unspecified testing conditions. The study report was submitted under TSCA in 1992 as DCN 88920006972 (OTS Fiche 0545375). Rainbow trout ( Salmo gairdneri ) per concentration were exposed to nominal concentrat ions of 0 (dilution water control), 0.1, or 1 mg/L in acetone. The age and size of the rainbow trout used in the study were not specified in the study. The specific environmental conditions of the test, such as pH, temperature and water quality were not sp ecified in the report. Concentrations of the chlorinated paraffin in the test water were measured using TLC analytical procedures resulting in mean measured concentrations of 0.01 and 0.18 mg/L. Mortality and behavior (response to food, general behavior, s wimming behavior and pigmentation) were assessed during the 28

- day study. Survival was 96.6 and 100% for the mean measured exposures of 0.01 and 0.18. No behavioral effects were seen over the course of the study.

EPA/OPPT Conclusion The study was conside red unacceptable to characterize the chronic fish toxicity endpoint since insufficient study details were provided including the age and/or the life

- stages of the exposed organisms.75 (3) A 20

- day Japanese medaka ( Oryzias latipes ) embryo toxicity study was conducted with the formulation C 14 H 24.9 C l5.1 , 48%Cl (composition: 10.5% 1, 2, 13, 14

- tetrachlorotetradecane (42.3% Cl); 74.3% x, 1, 2, 13, 14

- pentachlorotetradecane (47.7% Cl); 14.2% x, y, 1, 2, 13, 13

- hexachlorotetradecane (52.6% Cl); 1.0% x, y, z, 1, 2, 13, 14

- heptachlorotetradecane (56.4% Cl)) by Fisk et al. (1999) under static testing conditions. Five sets of 10 vials containing 1 egg each were exposed to nominal concentrations of 0.001, 0.010, 0.100, 1, or 10 mg/L test substance starting after fertiliz ation and terminating approximately 3 days post

- hatch. No adverse effects were reported in exposed embryos.

EPA/OPPT Conclusion The study was considered unacceptable primarily due to insufficient exposure duration and insufficient number of eggs per exposure concentration.

(4) A 20

- day Japanese medaka ( Oryzias latipes ) embryo toxicity study was conducted with the formulation 14 C

- C 14 H 23.3 Cl 6.7, 55% Cl (composition: 0.2% C 14 H 26 Cl 4 (42.3% Cl), 4.4% C 14 H 25 Cl 5 (47.7% Cl), 34% C 14 H 24 Cl 6 (52.6% Cl), 45% C 14 H 23 Cl 7 (56.4% Cl), 14% C 14 H 22 Cl 8 (59.9% Cl), and 1.9% C 14 H 21 Cl 9 (62.8% Cl)) by Fisk et al. (1999) under static testing conditions. Five sets of 10 vials containing 1 egg each were exposed to measured concentrations of 0.0014, 0.012, 0.120, 0.420, or 1.6 mg /L test substance starting after fertilization and terminating approximately 3 days post

- hatch. No adverse effects were reported in exposed embryos. Concentrations of the test substance were found in larvae and eggs in a dose

- dependent manner (with excepti on of the highest concentration) suggesting that the substance can diffuse through the egg. Corresponding measured concentrations in eggs were 0.04, 8.4, 63, 110, and 72 mg/kg and corresponding measured concentrations in larvae were 0.24, 8.2, 45, 84, and 51 mg/L.

EPA/OPPT Conclusion The study was considered unacceptable primarily due to insufficient exposure duration and insufficient number of eggs per exposure concentration.

(5) The following summary provided in the 2005 European Chemical Bureau Risk A ssessment of MCCP was considered supportive, but did not characterize all fish life

- cycle stages. Cooley et al. (2001) studied the toxicity of C 14 H 24.9 C l5.1 , 48%Cl (as described in Fisk et al., 1999) to juvenile rainbow trout ( Oncorhynchus mykiss ) through dietary exposure. Treatment groups of 10 fish were exposed to 0.78 and 2.9 mg/kg for 21 days and 0.082 mg/kg for 85 days. Three control groups were also run. Histological examination and analysis of the chlorinated paraffin concentration was performed in f ive fish per treatment after 21 days in the two higher test concentrations and in three fish per treatment after 85 days in the lowest test concentration. Three fish were also sacrificed from each low exposure group and the remaining control group (but wer e not analyzed) after 21 days of exposure. Quantitative histomorphological measurements were also carried out on livers and thyroid of the exposed fish in the middle exposure group after 21 days, and also the low exposure group after 85 days. The parameter s investigated included hepatocyte nuclear diameter, hepatocyte volume index, nucleus:cytoplasm area ratio and thyroid epithelium cell height. Livers displaying mild hepatocyte necrosis and moderate to severe depletion of glycogen/lipids were reported for the 0.78 mg/kg exposure. At 2.9 mg/L abnormal behavior was observed from day 3 onwards. Quantitative effects following 21 days of exposure were limited to a significantly (p=0.05) reduced mean hepatocyte volume in 2.9 mg/L exposure group.76 EPA/OPPT Conclu sion This study is considered unacceptable to characterize chronic mortality in fish because it did not characterize life stages but instead characterized physiological effects.

(6) Cooley et al. (2001) studied the toxicity of another medium chain chlori nated paraffin with a slightly different chemical composition and at slightly different concentration levels. The chemical formula was 14 C

- C 14 H 23.3 Cl 6.7, 55% Cl (as described in Fisk et al., 1999) and was juvenile rainbow trout ( Oncorhynchus mykiss ) were e xposed through the diet. Treatment groups of 10 fish were exposed to 29 and 78 mg/kg for 21 days and 5.7 mg/kg for 85 days. Three control groups were also run. Histological examination and analysis of the chlorinated paraffin concentration was performed in five fish per treatment after 21 days in the two higher test concentrations and in three fish per treatment after 85 days in the lowest test concentration. Three fish were also sacrificed from each low exposure group and the remaining control group (but w ere not analyzed) after 21 days of exposure. Quantitative histomorphological measurements were also carried out on livers and thyroid of the exposed fish in the middle exposure group after 21 days, and also the low exposure group after 85 days. The paramet ers investigated included hepatocyte nuclear diameter, hepatocyte volume index, nucleus:cytoplasm area ratio and thyroid epithelium cell height. At 29 mg/kg abnormal behavior was observed from day 2 onwards and livers exhibited mild to moderate hepatocyte necrosis and moderate to severe depletion of glycogen lipids. Abnormal behavior from day 3 onward was also observed at 78 mg/kg.

EPA/OPPT Conclusion This study is considered unacceptable to characterize chronic mortality in fish because it did not charac terize life stages but instead characterized physiological effects.



### B

**- 1** 

**- 5**  **Chronic Aquatic Invertebrate Toxicity** (1) A 21

- day chronic Daphnia magna reproduction toxicity study was conducted by Thompson et al. (1997b) according to OECD 202, Part II (1984) using a static

- renewal test system with renewal 3 times/week. The test substance was identified as Cereclor S52 ® , a C 14

- 17 chlorinated paraffin with 52% chlorination that contained 0.3% epoxy soya bean oil stabilizer as well as a small amount of radiolabelled n

- pentadecane

- 8

- 14 C (51% chlorinated). Ten replicates of 1 Daphnia magna Straus (<24 hours old) were tested per exposure concentration, which did not comply with OECD 202, Part II requirements that at least 40 daphnid be tested per concentration. Nominal co ncentrations were 0 (dilution water control), 0 (solvent control), 0.0056, 0.01, 0.018, 0.032, 0.056, and 0.1 mg/L test substance in acetone (0.025 mL/L). Results from the acute daphnid study by the same author do not appear to have been considered when

se lecting concentrations for this study. Test solutions were prepared by adding the appropriate stock solution to dilution water while continuously and vigorously stirring with a magnetic follower. The submitter does not indicate whether renewal of the stati c

- renewal test systems was carried out at regular intervals (e.g, Monday

- Wednesday

- Friday). Corresponding mean measured concentrations determined by radiochemical methods were 0.0037, 0.005, 0.01, 0.018, 0.032, and 0.065 mg/L and were 78

- 94% of nominal con centrations at the start of the renewal period and 7.3

- 61% of nominal concentrations at the end of the renewal period indicating a notable loss of test substance. Also, analysis of test concentrations appears to be at irregular intervals. In the dilution w ater control, 20% mortality was observed. Overall, dilution water control and solvent77 control results were significantly different for reproductive parameters. The test was carried out at temperatures of 19.5

- 20.3 °C, at pH levels of 7.41

- 8.13, and at diss olved oxygen concentrations of 6.2

- 9.2 mg/L. A significant decrease in the number of live offspring was reported at the mean measured concentration of 0.018 mg/L and delayed release of first offspring was observed at higher concentrations. Percentage dead offspring reported was 0%, 0%, 5.9%, 20.4%, and 18.5% for the 0.0037, 0.005, 0.01, 0.018, 0.032, and 0.065 mg/L mean measured
exposures.

EPA/OPPT Conclusion The inability to maintain test concentrations, unspecified renewal periods, and a smaller population size may have affected subsequent reproductive results. Given the uncertainties of the test, reported effect levels may not represent a worst case scenario but do exhibit a clear dose response relationship with a clearly defined statistically si gnificant effect level. Thus, using a weight of evidence approach, the study is considered acceptable for this endpoint. 21

- day LC 50 value of 0.025 mg/L (parent mortality) 21

- d NOEC = 0.01 mg/L 21

- d LOEC = 0.018 mg/L 21

- d GMATC = 0.013 mg/L

(2) A 60

- day mussel toxicity study was conducted by Brixham Laboratories in 1983 with radio

labeled chlorinated (52%) n

- pentadecane (Trade Name: Cereclor S52 ® ) under flow through testing conditions. A full non

- CBI study report was submitted under TSCA in 1983 as DCN 4 0

- 8332184 (OTS Fiche 0507258). Blue mussels ( Mytilus edulis ) were exposed to nominal concentrations of 0 (dilution sea water control), 0 (acetone control, 500 ppm), 0.56, or 5.6 mg/L in 500 ppm acetone. Two replicates of 50 mussels were tested for the dilu tion water and solvent controls and a single replicate of 50 mussels was exposed for each treatment concentration. Corresponding mean measured concentrations were 0, 0, 0.22, and 3.8 mg/L. Test solutions were cloudy at the higher test concentration. Test c oncentrations were determined by radio activity measurements. Flow rate of the test system was 0.25 mL/minute for exposure concentrations. Over the course of the study, water temperature ranged from 14.6

– 15.6 °C, pH ranged from 8.0

– 8.3, and the dissolv ed oxygen concentrations ranged from 6.1

– 8.25 mg/L. Dilution water salinity was 34

- 35.5 ppb, which is high by OCSPP standards.

EPA/OPPT Conclusion One mussel exposed to 0.56 mg/L died, and two mussels exposed to the controls died; this was not consider ed to be a test substance related effect. Decreases in filter feeding were observed at 5.6 mg/L. In addition, the submitter provided an assessment of bioconcentration, but this assessment does not appear to include a depuration phase. Overall, the 60 day N OEC and LOEC were 0.22 and 3.8 mg/L based on reduced filtration. The study was acceptable to characterize mussel toxicity, but mussels are not considered a standard species to fulfill the chronic aquatic invertebrate toxicity endpoint. 60

- d NOEC = 0.22 mg/ L 60

- d LOEC = 3.8 mg/L (reduced filtration)

(3) The following study summary (Frank, 1993; Frank and Steinhäuser, 1994) provided in the 2005 European Chemical Bureau Risk Assessment of MCCP was considered support ive of the78 aquatic invertebrate hazard deter mination. A 21

- day chronic Daphnia magna reproduction toxicity study was conducted using a static

- renewal test system (renewal 3 times/week). The test substance was identified as C 14

- 17 chlorinated paraffin with 52% chlorination, which was tested
as a wate r soluble fraction of two stock solutions (dilutions used 1:2 to 1:32). Nominal test concentrations prepared from the 100 mg/L stock solution were 3.125, 6.25, 12.5, 25, and 50 mg/L. Nominal test concentrations prepared from the 10,000 mg/L stock solution were 312.5, 625, 1250, 2500, and 5000 mg/L. Analytical monitoring of test concentrations was conducted, but only the final effect levels were presented as measured concentrations. Methods for test solution preparation were not provided in the summary. The tests we re carried out at 20 °C and
at pH 7.79

- 8.44.

In the experiments using the 100 mg/L stock solution the mortality seen in the exposed populations was 0% at 3.125 mg/L, 0% at 6.25 mg/L, 20% at 12.5 mg/L, 90% at 25 mg/L, and 100% at 50 mg/L. In the e xperiments using the 10,000 mg/L stock solution the mortality seen in the exposed populations was 0% at 312.5 mg/L, 30% at 625 mg/L, 70% at 1250 mg/L, and 100% at lower dilutions (>1250 mg/L). In the experiments using the 100 mg/L stock solution the
averag e number of young/adult was 82 at 3.125 mg/L, 89 at 6.25 mg/L, 80 at 12.5 mg/L, 15 at 25 mg/L and 0 at 50 mg/L (all parents died). Similarly in the experiments using the 10,000 mg/L stock solution the average number of young/adult was 74 at 312.5 mg/L, 64 at 625 mg/L, 43 at 1250 mg/L, and 0 at 2,500 and 5,000 mg/L (all parents died). Based on these effects, survivability/mortality appears to be the more sensitive endpoint. Based on the known measured concentrations in the stock solutions and the dilution ra tes used the NOEC for mortality was around 0.0044

- 0.0089 mg/L for the 100 mg/L nominal stock solution experiments and 0.0126

- 0.0156 mg/L for the 10000 mg/L nominal stock solution experiments. The corresponding LOECs were 0.0089

- 0.0178 mg/L (100 mg/L nomin al stock) and 0.0253

- 0.0313 μg/L (10 g/L nominal stock). The GMATC of 0.006 mg/L was calculated using the geometric mean from the most conservative NOEC (0.0044 mg/L) and LOEC (0.0089 mg/).

EPA/OPPT Conclusion EPA/OPPT reserves judgment on the acceptabili ty of this study until further details become available. 21

- d NOEC = 0.0044 mg/L 21

- d LOEC = 0.0089 mg/L 21

- d GMATC = 0.006 mg/L

3) The following study summary (TNO, 1993) provided in the 2005 European Chemical Bureau Risk Assessment of MCCP was conside red support ive of the aquatic invertebrate hazard determination. A 21

- day chronic Daphnia magna reproduction toxicity study was conducted using a static

- renewal test system (renewal 3 times/week). The test substance was identified as C 14

- 17 chlorinated par affin with 52% chlorination. The test solutions were prepared by stirring 20 g of the test substance in 2 litres of heated water (60 °C) with stirring and then filtration through a 0.8 μm and 0.2 μm filter. This resulting stock solution was referred to as a water soluble fraction, but given that each concentration was not independently prepared, the test solutions is considered by EPA/OPPT to be merely a mixed and filtered solution that was subsequently diluted. Following dilution of the stock solution, exposure concentrations were analytically determined using the extractable organic halogen method, but not provided in the study summary. T he test was carried out at 20 ± 1 oC and solutions were gently aerated from day 9 onwards. The pH of the test water varied between 7.7 and 8.3, the dissolved oxygen concentration was > 7 mg/L, and the hardness was 214 mg CaCO3/L. Test solutions were clear.79 EPA/OPPT Conclusion Although analytical results obtained were considered to be too erratic to allow precise determination of concentrations (according to the study summary in ECB, 2005), the NOEC was reported based on survivability and/or reproductive effects. A LOEC in mg/L was not reported in the summary, nor could EPA/OPPT extrapolate one. EPA/OPPT reserves judgment on the acceptability of this study until further details become available. 21

- d NOEC = 0.004

- 0.008 mg/L



### B

**- 1** 

**- 6**  **Chronic Aquatic Sediment Invertebrate Toxicity** 1) Thom pson et al. (2001c) conducted a 28

- day prolonged sediment invertebrate toxicity study with spiked sediment was conducted according to OECD 218 draft guideline (February 2000 version) using a static test system. The substance used in the test was commercial C 14

- 17 , 52% wt. Cl substance containing no stabilizers (the substance was reported to have a C 14

- 17 content of 99.06% with 0.67% of C10

- 13 chain length substances) that was mixed with a small amount of a radio labeled n

- pentadecane

- 8

- 14 C, 51% wt. Cl subst ance (radiochemical purity >96.6%). The sediment used in the test was an artificial sediment that did not fully adhere to the final OECD TG recommendations, but the composition of 10% sphagnum moss peat, 70% quartz sand, 20% kaolinite clay, and <0.1% calci um carbonate are not considered to be significantly different. The sediment had a mean organic carbon content of 4.9% and a pH of 6.0. The sediment was spiked with the test substance by firstly mixing a solution of the test substance in acetone with the dr y sand component of the sediment and allowing the acetone to evaporate overnight under an air stream. Measured concentrations were determined using radiochemical analysis. Over the course of the study, temperature was maintained at 20 oC, pH levels were 6. 2

- 7.6, and dissolved oxygen in overlying water was maintained at 7.3

- 8.6 mg/L. Three replicates of 15 midge ( Chironomus riparius ) larvae (<48 hours post hatch) were exposed to mean measured concentrations of 0 (sediment control), 0 (solvent sediment contro l), 36, 110, 370, 1200, 3800, or 13000 mg/kg dry wt. sediment. Time to first emergence, mean emergence time, mean number emerged per replicate, and sex ratio was assessed for each exposure group. Statistically significant effects (p = 0.05) were limited to a decrease in mean number emerged per replicated in the 13,000 mg/kg dry wt. sediment exposed midges.

EPA/OPPT Conclusion The overall NOEC of 3,800 mg/kg dry wt. sediment corresponded to 1,460 mg/kg on a wet weight basis. The study is acceptable. 28

- d N OEC = 3,800 mg/kg dry wt sediment 28

- d LOEC = 13,000 mg/kg dry wt sediment 28

- d GMATC = 7,029 mg/kg dry wt sediment

2) Thompson et al. (2001d) conducted a 28

- day prolonged sediment invertebrate toxicity study with spiked sediment according to methods des cribed in Phipps et al. (1993) using a static test system. The substance used in the test was commercial C 14

- 17 , 52% wt. Cl substance containing no stabilizers (the substance was reported to have a C 14

- 17 content of 99.06% with 0.67% of C10

- 13 chain length substances) that was mixed with a small amount of a radio labelled n

pentadecane

- 8

- 14C, 51% wt. Cl substance (radiochemical purity >96.6%). The sediment used in the test was an artificial sediment consisting of 10% sphagnum moss peat, 70% quartz sand, 20%80 kaolinite clay, and <0.1% calcium carbonate. The sediment had a mean organic carbon content of 4.9% and a pH of 6.0. The test sediments were made up by adding the test substance to the sand phase as a solution in acetone, evaporating the acetone overnight and mixing the spiked sand with the rest of the sediment for 16 hours. Six replicates of 10 oligochaete ( Lumbriculus variegatus ) adults were exposed to mean measured concentrations of 0 (sediment control), 0 (solvent sediment control), 39, 130, 410, 1300, 4000, or 13000 mg/kg dry wt. sediment. Throughout the duration of the study, water temperature was maintained at 20 ± 1 Co, and pH remained between 6.3 and 7.9. Mortality and reproductive success were determined by total number of worms at study terminati on since differentiation of adult and young worms is difficult. Mean number of worms per replicate and mean total dry weight of worms per replicate was significantly different from controls (p = 0.01) at mean measured concentrations of 410 mg/kg dry weight sediment and greater. Statistical methods used were not reported, and a p

value of 0.01 was used.

EPA/OPPT Conclusion Given that a clear decline in the mean number of worms per replicate and mean total dry weight of worms per replicate was observed at th e lowest test concentration (39 mg/kg dry weight sediment) a conservative LOEC of 39 mg/kg dry weight sediment will be used based on noticeable differences. EPA/OPPT reserves judgment on the acceptability of this study until further details become availabl e regarding the analytical measurements of the chlorinated paraffin mixture. 28

- d NOEC = 130 mg/kg dry wt sediment 28

- d LOEC = 410 mg/kg dry wt sediment GMATC = 230.9 mg/kg dry wt sediment

3) Thompson et al. (2002) conducted a 28

- day prolonged sediment toxicity study with amphipod Hyalella azteca in spiked sediment using a static

- renewal test system with weekly renewals. The substance used in the test was a mixture of a commercial medium

- cha in chlorinated paraffin product (C14

- 17, 52.5% wt. Cl) mixed with a small amount of a radiolabelled chlorinated n

pentadecane

- 8

- 14C (51% wt. Cl). The sediment used in the test was an artificial sediment consisting of 10% sphagnum moss peat, 70% quartz sand , 20% kaolinite clay, and <0.1% calcium carbonate. The sediment had a mean organic carbon content of 4.9% and a pH of 6.0. The test sediments were made up by adding the test substance to the sand phase as a solution in acetone, evaporating the acetone over night and mixing the spiked sand with the rest of the sediment and water. Six replicates per concentration of ten juvenile Hyalella azteca (~7

- day

- old) were exposed to 0 (sediment control), 0 (acetone sediment control), 38, 75, 150, 300, or 600 mg/kg dry w eight sediment. The concentration of the test substance was measured in the sediment phase by radiochemical analysis with concentrations at the start of the exposure period of 85

- 97% of the nominal values and concentrations at the end of the 29

- day exposur e period of 78

- 90% of nominal. Results of the test were expressed as the arithmetic mean concentration. Over the course of the study, dissolved oxygen ranged from 7.7 to 8.4 mg/L, pH ranged from 7.0 to 7.6, water hardness ranged from 41 to 42 mg CaCO 3 /L, a nd temperature ranged from 22.4

- 23.2oC. The endpoints investigated in the study included survival, growth (dry weight) and sexual development of females (proportion of gravid females).

EPA/OPPT Conclusion Controls responded adequately. For the survival e ndpoint, a statistically significant (p=0.05) reduction in survival was seen at 470 mg/kg dry weight. A statistically significant reduction in mean weight was seen in81 females only at exposures of 470 g/kg dry weight sediment and a statistically significant (p=0.05) reduction in mean weight was seen at 270 mg/kg dry weight. For the sexual development endpoint, there was a statistically significant (p=0.05) reduction in the proportion of gravid females in the 470 mg/kg dry weight treatment. This study was

acc eptable. 28

- d NOEC = 130 mg/kg dry wt sediment 28

- d LOEC = 270 mg/kg dry wt sediment 28

- d GMATC = 187 mg/kg dry wt sediment



### B

**- 1** 

**- 7**  **Avian Toxicity** ( 1) An acute avian toxicity study conducted according to OPPTS guidelines was published by Madeley & Birtley (1980). Following a range

- finding study, groups of 5 male and 5 female ring

necked pheasants ( Phasianus colchicus ) were exposed by gavage to 0 (control) or 24,606 mg/kg Cereclor S52 ® (C 14

- 17 , 52% Cl) and then observed for 14 days. Based on reported tissue concentrations, the test substance is believed to have been absorbed by the ring

- necked pheasant. Doses up to 24,606 mg/kg failed to produce any abnormal clinical signs or mortality.

EPA/OPPT Conclusion Acute LD50 > 24,606 ppm

(2) An acute avian toxicity study conducted according to OCSPP guidelines was published by Madeley & Birtley (1980). Following a range

- finding study, groups of 5 male and 5 female mallard ducks ( Anas platyrynchos ) were exposed by gavage to 0 (control) or 10,280 mg/kg Cereclor S52 ® ( C14

- 17, 52% Cl) and then observed for 14 days. Based on reported tissue concentrations, the test substance is believed to have been absorbed by the mallard ducks. Doses up to 10,280 mg/kg failed to produce any abnormal clinical signs or mortality.

EPA/OPP T Conclusion Acute LD 50 > 10,280 ppm

(3) A sub

- acute dietary avian toxicity study conducted according to OPPTS guidelines was published by Madeley & Birtley (1980). Following a range

- finding study, groups of 5 male and 5 female ring

- necked pheasants ( Phasianus colchicus ) were exposed to diets containing 0 (control), 1,000, or 24,063 ppm Cereclor S52 ® (C14

- 17, 52% Cl) for 5 days. Three groups were exposed to the negative control and two groups were exposed to each of the treatment concentrations. Based on reported tissue concentrations, the test substance is believed to have been absorbed by the ring

- necked pheasant. Good health was noted in all control and treatment groups. No abnormal effects were noted at necropsy.

EPA/OPPT Conclusion 5

- day LD 50 > 24 ,063 ppm

(3) A sub

- acute dietary avian toxicity study conducted according to OPPTS guidelines was published by Madeley & Birtley (1980). Following a range

- finding study, groups of 5 male and 5 female mallard ducks ( Anas platyrynchos ) were exposed to diets containing 0 (control), 1,000, or 24,063 ppm Cereclor S52 ® (C14

- 17, 52% Cl) for 5 days. Three groups were exposed to the negative control and two groups were exposed to each of the treatment concentrations. Inferior food intake was noted for ducks, but we ight gain was comparable to controls. Based on reported82 tissue concentrations, the test substance is believed to have been absorbed by the mallard ducks. Good health was noted in all control and treatment groups. No abnormal effects were noted at
necropsy.

EPA/OPPT Conclusion 5

- day LD 50 > 24,063 ppm



### B

**- 1** 

**- 8**  **Terrestrial Invertebrate Toxicity** 1) A 28

- day earthworm reproductive toxicity test was conducted by Thompson et al., 2001a according to OECD guideline (2000 draft version). The substance tested was commercial C14

- 17, 52% wt. Cl substance containing no stabilizers (the substance was reported to have a C14

- 17 content of 99.06% with 0.67% of C10

- 13 chain length substances) and a small amount of 14C

labelled n

- pentadecane, 51% wt. Cl substance. Four replicates per concentration of 10 adult earthworm ( Eisenia fetida ) were exposed to nominal concentrations of 0 (soil control), 0 (solvent soil control), 100, 320, 1000, 3200, or 10,000 mg/kg dry wt. soil. Corresponding mean measured concentrations of 0 (soil control), 0 (solvent soil control), 79, ~280, 900, ~2,800, or 9,300 mg/kg dry wt.was determined using radiochemical analysis; concentrations identified as approximate (~) were approximated using the mean % of nominal (87%) determined in other treatments. Measured tis sue concentrations in adults on day 28 were 169, 802, and 732 mg/kg wet weight for the 79, 900, and 9,300 mg/kg dry weight exposure groups. Measured tissue concentrations in juveniles on day 56 were 140 and 1,011 mg/kg wet weight for the 79 and 900 mg/kg d ry weight exposure groups. The soil used in the test was an artificial soil consisting of 10% sphagnum moss peat, 70% quartz sand, 20% kaolinite clay, and 0.25% calcium carbonate. The soil had an organic carbon content of 4.7% and a pH of 6.66

- 7.09. Nomina l test temperatures remained at 20 ± 1 oC. The soils were prepared up by firstly adding the test substance in solution with acetone to a small portion of soil, evaporating out the acetone overnight under a stream of compressed air, and then mixing with the remainder of the soil. Before use, distilled water was added to the dry soil to provide a soil wet:dry ratio of 1.35. Following the 28

- day parental exposure period, adult earthworms were removed, and vessels were incubated for an additional 28 days to all ow hatching of any egg cocoons produced by parent. Effects assessed were parental survival, growth as determined by change in weight of parents, and reproduction as determined by number of live offspring. A statistically significant (p = 0.05) reduction in parental survival (85%) was observed at 9,300 mg/kg dry wt. soil. A statistically significant (recalculated with Dunnett’s Procedure, P = 0.05) reduction in parental weight was reported at 2800 mg/kg dry wt. soil. A statistically significant (recalculated with Dunnett’s Procedure, P = 0.05) reduction in number of live offspring was reported at 280 mg/kg dry wt. soil. In addition, the submitter assesses corresponding tissue concentrations in earthworms and determines that at nominal concentrations of 100 mg /kg dry wt. soil the concentration in parental earthworm tissue after 28 days is 850 mg/kg dry wt.and in juvenile worms after 56 days was 703 mg/kg dry wt..

EPA/OPPT Conclusion The study was acceptable. 28

- d NOEC = 79 mg/kg dry wt soil 28

- d LOEC = 280 mg /kg dry wt soil 28

- d ChV = 149 mg/kg dry wt soil

2) The following study summary (Thompson et al., 2001a) provided in the 2005 European Chemical Bureau Risk Assessment of MCCP was used to characterize terrestrial invertebrate hazard in a screening level ri sk assessment. A 28

- day earthworm reproductive toxicity test was83 conducted according to OECD guidelines. The substance tested was commercial C 14

- 17 , 52% wt. Cl substance containing no stabilizers (the substance was reported to have a C 14

- 17 content of 99.06% with 0.67% of C10

- 13 chain length substances). The test substance contained a small amount of 14C

- labelled n

- pentadecane, 51% wt. Cl substance so that radiochemical analysis could be used to analytically determine test concentrations in soil. The so il used in the test was an artificial soil consisting of 10% sphagnum moss peat, 70% quartz sand, 20% kaolinite clay, and 0.25% calcium carbonate. The soil had an organic carbon content of 4.7% and a pH of 6.66

- 7.09. The soils were prepared up by firstly a dding the test substance in solution with acetone to a small portion of soil, evaporating out the acetone overnight under a stream of compressed air, and then mixing with the remainder of the soil. Before use, distilled water was added to the dry soil to p rovide a soil wet:dry ratio of 1.35. Four replicates per concentration of 10 adult earthworm ( Eisenia fetida ) were exposed to mean measured concentrations of 0 (soil control), 0 (solvent soil control), 79, ~280, 900, ~2800, or 9300 mg/kg dry wt. for 28 day s; concentrations identified as approximate (~) were approximated using the mean % of nominal (87%) determined in other treatments. Measured tissue concentrations in adults on day 28 were 169, 802, and 732 mg/kg wet weight for the 79, 900, and 9300 mg/kg d ry weight exposure groups. Measured tissue concentrations in juveniles on day 56 were 140 and 1011 mg/kg wet weight for the 79 and 900 mg/kg dry weight exposure groups. Following the 28

- day parental exposure period, adult earthworm were removed, and vessel s were incubated for an additional 28 days to allow hatching of any egg cocoons produced by parent. Effects assessed were parental survival, growth as determined by change in weight of parents, and reproduction as determined by number of live offspring. St atistical methods used to calculate significance were not provided. A statistically significant (p = 0.05) reduction in parental survival (85%) was observed at 9300 mg/kg dry wt. soil. A statistically significant (p = 0.01) reduction in parental weight was reported at 2800 mg/kg dry wt. soil and a noticeable reduction in parental weight was reported at 280 mg/kg dry wt. soil. A statistically significant (p = 0.01) number of live offspring was reported at 1000 mg/kg dry wt. soil and a noticeable reduction in number of live offspring was reported at 320 mg/kg dry wt. soil. A clear decline in parental weight and number of live offspring was observed at 280 mg/kg dry wt. soil.

EPA/OPPT Conclusion This study is acceptable. 28

- d NOEC = 79 mg/kg dry wt soil 28

- d LOEC = 280 mg/kg dry wt soil 28

- d GMATC = 149 mg/kg dry wt so il



### B

**- 1** 

**- 9**  **Terrestrial Plant Toxicity** 1) Thompson et al., 2001ab conducted a 28

- day seed germintation and vegetative vigor study. The toxicity of a C 14

- 17 , 52% wt. Cl substance (99.06% purity) to wheat ( Triticum aestivum ; monocotyledon), oilseed rape ( Brassica napus ; ditcotyledon), and mung bean ( Phaseolus

aureus ; dicotyledonous legume) has been studied using OECD guideline 208 (July, 2000 Revision). The test substance contained a small amount of 14C

- lab elled n

- pentadecane, 51% wt. Cl substance so that radiochemical analysis could be used to analytically determine test concentrations in soil. The soils were prepared up by firstly adding the test substance in solution with acetone to dry silver sand, evapo rating the acetone overnight, and mixing the spiked sand with the soil. Four replicate pots per exposure concentration each containing 9 seeds were exposed for 28 days to nominal exposure concentrations of 0 (soil control), 0 (solvent soil control), 50, 15 8, 500, 1,580, or 5,000 mg/kg dry wt.. According to the E.U. Risk Assessment, use of nominal test concentrations to determine effect levels is based on test substance stability84 determined at the 50, 500, and 5,000 mg/kg dry wt. concentrations (correspondin g mean measured concentrations: 49, 520, and 5,800 mg/kg dry wt.). Effects assessed were seed germination, emergence (% emerged plants on Day 14), vegetative growth (mean shoot dry weight per plant), and visual appearance of seedling. No statistically sign ificant differences (p = 0.05) were observed in wheat, oilseed rape. A statistically significant reduction in growth was seen at 1,580 and 5,000 mg/kg dry wt. for mungbean when compared to soil control results.

EPA/OPPT Conclusion Since soil control and s olvent control means were equal (two tailed T

Test) indicating no solvent interference, comparison of treatments was made to the soil control. Thus, the NOEC and LOEC for terrestrial plants was 500 and 1,580 mg/kg dry wt. soil and the GMATC (geometric mean of the NOEC and LOEC) was 888.8 mg/kg dry wt. soil. The study was acceptable to characterize both monocot (wheat) and dicot (mung bean) seed germination and vegetative vigor; reproductive effects remain uncharacterized. 28

- d NOEC = 500 mg/kg dry wt soil 2 8

- d LOEC = 1,580 mg/kg dry wt soil 28

- d GMATC = 888.8 mg/kg dry wt soil



### B

**- 1** 

**- 10**  **Conclusions** Sufficient data were available to characterize the acute fish, the acute aquatic invertebrate, the chronic aquatic invertebrate, the chronic aquatic sediment invertebrate, avian, and terrestrial plant toxicity endpoints for MCCPs. Data for other toxicity endpoints (i.e., chronic fish, aquatic plant, etc.) were inconclusive due to lack of study details, uncertainties in analytical methods, or test material prepa ration methods; thus, these data are included in order to characterize risk in a qualitative manner, but are used as supportive for the categories under which they are provided. Supporting data were included in order to provide a weight

- of

- evidence approac h used to characterize some endpoints.

Most of the data provided in this review indicated several difficulties were encountered when testing in an aquatic environment. These included: (1) getting the material into solution, (2) measuring the material in solution, and (3) characterizing the effects for each study listed. Often there were many details of a given study omitted, prohibiting a full and robust review of the data. The (estimated) physical

- chemical properties of MCCPs (water solubility values of approximately 30 μg/L and Log K ow values between 4

- 8) suggest these materials may not partition to the aquatic media or elicit toxicity to aquatic organisms within the water column.

The most reliable and acceptable studies indicate that for MCCPs, the to xicity to aquatic organisms for acute endpoints are from the Thompson et al. 1996 study for aquatic invertebrates. Where the 48

- hour EC 50 value = 0.0059 mg/L. Using the methods described in the Su stainable Futures/P2 Manual (US EPA, 2012), the acute and chronic concentrations of concern (CoC) are derived as follows:

• Acute CoC: The 48

- hour EC 50 value = 0.0059 mg/L is divided by an assessment factor of 5 to yield an acute concentration of concern (CoC) of 0.00118 mg/L, or 0.001 mg/L, or 1 μg/L (1 ppb). Aqua tic Acute Concern Concentration= 1 ppb

• Chronic CoC: The aquatic invertebrate chronic value of 0.013 mg/L from the 1997 Thompson et al. study based on a MCCP material is divided by an assessment factor of85 10 to yield 0.0013 mg/L or 1.3 μg/L or 1.3 ppb. Aqua tic Chronic Concern Concentration = 1 ppb

The most reliable and acceptable value for the toxicity to aquatic sediment invertebrate organisms acute endpoint is based on the MCCP material from the Thompson et al. 2002 28

- d study. The 28

- d sediment invertebr ate GMATC value of 187 mg/kg dry wt sediment is used to assess haz ard. Again, using methods in US EPA (2012):

• Acute CoC: Calculating an acute concern concentration from the chronic value of 187 mg/kg dry wt. The 187 value is multiplied by an acute to chron ic ratio for invertebrates (10) to yield 1,870 mg/kg dry wt. This value is then divided by an assessment factor of 5 to yield 374 mg/kg dry wt. Aquatic Sediment Acute Concern Concentration = 374 mg/kg dry wt sediment.

• Chronic CoC: The 28

- d sediment inverte brate GMATC of 187 mg/kg dry wt sediment is divided by an assessment factor of 10 to yield 18.7 mg/kg dry wt sediment. Aquatic Sediment Chronic Concern Concentration = 19 mg/kg dry wt sediment.

The most reliable and acceptable value for the toxicity to te rrestrial invertebrates acute endpoint is based on the MCCP material from the Thompson et al. 2001a study. The 28

- d terrestrial invertebrate GMATC value of 149 mg/kg dry wt sediment from this study will be used. Again, using methods in US EPA (2012):

• A cut e CoC: Calculating an acute concern concentration from the chronic value of149 mg/kg dry wt, this value is multiplied by an acute to chronic ratio for invertebrates (10) to yield 1,490 mg/kg dry wt. This value is then divided by an assessment factor of 5 t o yield 298 mg/kg dry wt. Terrestrial Invertebrate Acute Concern Concentration = 298 mg/kg dry wt sediment.

• Chronic CoC: The 28

- d terrestrial invertebrate GMATC of 149 mg/kg dry wt is divided by an assessment factor of 10 to yield 14.9 mg/kg dry wt sedimen t. Terrestrial Invertebrate Chronic Concern Concentration = 15 mg/kg dry wt sediment.

The most reliable and acceptable value for the toxicity to terrestrial plants is based on the MCCP material from the Thompson et al. 2001ab study. For LCCPs, the analog approach using the values from this study may be used. However, there is no OPPT guidance regarding assessing concern concentrations for terrestrial plants.



### B

- 2 LCCP Ecotoxicity Data Where data are absent for long chain chlorinated paraffins for ecotoxicity endpoints, data from sources using medium chain chlorinated paraffins will be used to inform the hazard for these endpoints.



### B

**- 2** 

**- 1**  **Acute Fish Toxicity** (1) Johnson and Finely 1980 A series of 96

- hour acute fish toxicity studies were conducted by the United States Geological Survey’s Columbia National Fisheries Research Laboratory, over the years 1965

– 1978, with several long

- chain chlorinated paraffins (Chlorowax LV, C >17 , 39 % Cl; Chlorowax 40, C >20 , 40

- 42 % Cl; Chlorowax 50, C >20 , 48

- 54 % Cl; Chlorowa x 70, C >20 , 70 % Cl) published by Johnson and Finely (1980). Bluegill sunfish ( Lepomis macrochirus ) and rainbow trout ( Oncorhynchus mykiss ) were exposed to the test substance (100 % commercial formulation) in a static test system. Stock solutions were prepared immediately before each test with acetone used86 as a carrier solvent. The average pH level was between 7.2 and 7.5 for all tests and the test temperature was 20 ± 1°C for Bluegill sunfish and 10 ± 1 oC for Rainbow trout. Diluti on water hardness ranged from 40 to 50 mg/L as CaCO 3

. Reported effect levels are considered to be nominal with LC 50 values of >300 mg/L for bluegill sunfish and rainbow trout for all LCCPs tested; all values are above the limit of solubility.

EPA/OPPT Con clusions Using a weight

- of

- evidence approach, these studies were considered acceptable to characterize the acute fish toxicity endpoint. 96

- hr LC 50 = NES (> 300 mg/L)

(2) Bengtsson et al. 1979 A 96

- hour acute fish toxicity study as part of a bioaccumulati on study was published by Bengtsson et al. (1979). Bleak ( Alburnus alburnus ) were exposed to a long

- chain chlorinated paraffin, Witaclor 549 (C 18

- 26 , 49 % Cl). The test was performed at 10 °C under semi

- static testing conditions. A nominal concentration o f 0.125 mg/L was prepared by first dissolving Witaclor in acetone and then added to the dilution water. The treatment vessels and acetone control vessels consisted of six groups of 15 fish (average 4.5 g per fish). The acetone in all treatments did not exceed 0.1 ml/L. The test solutions were renewed every two to three days over the 14

- day exposure period. Specific details of the test conditions were not provided other than the tests were performed at 10°C in seawater with a salinity of 7 ppt. Even thoug h the test duration was 14

- days, no mortality occurred within 96

- hours or 14

- days, thus the 96

- hour LC 50 was > 0.125 mg/L.

EPA/OPPT Conclusions This study was considered supplemental to characterize the acute saltwater fish toxicity endpoint using a weig ht

- of

- evidence approach. 96

- hr LC 50 = NES (> 0.125 mg/L)



### B

**- 2** 

**- 2**  **Acute Aquatic Invertebrate Toxicity** (1) Frank (1993) and Frank and Steinhäuser (1994) The following study summary (Frank, 1993; Frank and Steinhauser, 1994) , provided in the ECB (2005) MCCP risk assessment, was considered supportive of the aquatic invertebrate hazard determination. The chlorinated paraffin used in these studies was a commercial C 14

- 17 product with a 52 wt% Cl. Daphnia magna were exposed to nominal concentrations of eith er 100 mg/L or 10,000 mg/L. The 100 mg/L solution was sonicated for 1 hour and then left to stand in the dark for 48 hours before use. The 10,000 mg/L solution also stood for 48 hours in the dark before use, but this time without sonication. After this per iod, both solutions were filtered firstly with glass filters and then with membrane filters to remove undissolved test substance. The concentrations of medium

- chain chlorinated paraffin in the water soluble fractions were then determined by AOX (adsorbable organic halogen) analysis (detection limit of 10 μg/L Cl was equivalent to around 20 μg/L of the chlorinated paraffin). This analysis showed that the concentration of chlorinated paraffin present in the water soluble fraction was around 0.404

- 0.500 mg/L for the 10,000 mg/L nominal solution and 0.071

- 0.142 mg/L for the 100 mg/L stock solution. The acute (48

- hour) toxicity tests were carried out using dilutions of the two prepared water soluble fractions. The method used was DIN 38 412, Teil 11, which is equivalent to OECD 202.87 In the tests using the water soluble fraction from the 100 mg/L nominal solutions no toxicity was seen at concentrations up to the undiluted stock solution (i.e., no effects up to around 0.071

- 0.142 mg/L). In experiments using th e water soluble fraction from the 10,000 mg/L stock solution, an EC 0 of 0.140 mg/L (also reported as 0.100

- 0.110 mg/L in the paper) and an EC25 of 0.423 mg/L (also reported as 0.420

- 0.470 mg/L in the paper) was determined (maximum mortality seen was 25%) (Frank, 1993). The latter results for the 10,000 mg/L stock solution were reported by Frank and Steinhauser (1994) as EC 0 = 0.140 mg/L and EC 25 = 0.339 mg/L, and it was noted that s ome of the Daphnia were floating on the surface of the test solution. In the later study (Frank and Steinhauser, 1994) , the results of further acute toxicity studies were reported using the same test method. An EC 50 of 0.037 mg/L and an EC 0 of 0.009 mg/L were determined using the water soluble fraction from the 100 mg/L stock solution and no toxic effects were seen in tests with the water soluble fraction from the 10,000 mg/L stock solution (appro ximately EC 0 ≥ 0.525 mg/L). The authors noted that the effects seen in the acute tests showed poor reproducibility, probably because effects were seen only around the water solubility limit of the substance. However, the authors thought that the possibilit y of undissolved droplets affecting the results could be ruled out, as floating Daphnia were only sporadically observed in the test.

EPA/OPPT Conclusions This study could not be adequately assessed due to inconsistencies in the hazard data for the different dilutions of the test material. In addition, the authors noted that the effects seen showed poor reproducibility, most likely due to effects observed only around the solubility limit in the test system used. The results of this test should theref ore be treated with caution, as the effects were mainly seen in the saturated solutions only.

(2) Tarkpea et al. 1981 The Tarkpea et al. (1981) ; as quoted in IPCS (1996) summary, provided in the ECB (2005) MCCP risk assessment, was considered supportive of the aquatic invertebrates hazard determination.

The results of tests with the brackish water harpacticoid Nitocra spinipes have been reported (Tarkpea et al., 1981) . No other details of the test were reported but the test method was probably the same as reported by Tarkpea et al. (1981) , where a static method was employed using water of salinity 7‰ at a temperature of 20

- 22°C without aeration, probably using

acetone as co

- so lvent.

EPA/OPPT Conclusions This study could not be adequately assessed due to the lack of details provided regarding specific conditions of the test and the preparation of the test solutions. The lack of detail prohibits a review and adequacy of the stud y. However, the harpaticoid, Nitocra spinipes , is not a standard test species and the test concentrations greatly exceed the limit of solubility. The 96

- hour LC 50 values for both chlorinated paraffins were identified, but the reliability of the results can not be determined. Overall, EPA/OPPT reserves judgment on the acceptability of this study. Therefore, data from sources using medium chain chlorinated paraffins will be used to inform the hazard for these endpoints.88 

### B

**- 2** 

**- 3**  **Aquatic Plant Toxicity** (1) Koh and Thiemann 2001 A 72

- hour algae toxicity study using Scenedesmus subspicatus , was conducted at the University of Bremen, Department of Physical and Environmental Chemistry with CP 30 (C 17

- 24 , 35% chlorination), CP 40 (C 17

- 20 , 44% chlorination) and Hordaflex LC50 (C 17

- 20 , 52% chlorination) according to DIN 38412 by Koh and Thiemann (2001). Additional communications with the study author Wolfram Thiemann clarified that nominal test concentrations and local tap water (Bremen, Germany) was used without adjustment s however, the study methods were not fully characterized. The stock solutions were prepared to a concentration 6 mg/mL of test substance in
acetone. The individual stock solutions were diluted with distilled water to prepare individual standard solutions of 0.250 mg/L for CP 30 and CP 40 and 0.125 mg/L for Hordaflex LC 50. pH was between 5 and 6 and water hardness was between 35.7 and 53.5 mg CaCO 3 /L. Ambient laboratory air temperature was ~ 21 oC. The solvent acetone was used to maintain test substance in solution. Effects were calculated based on growth rate; no effects were observed up to 0.250 and 0.125 mg/L.

EPA/OPPT Conclusions All three of the chlorinated paraffins tested contain C 17 and C 18 constituents which are considered to have LCCP

- like proper ties. More information concerning composition would be needed to accept these results for long

- chain chlorinated paraffins. Overall, EPA/OPPT reserves judgment on the acceptability of this study. 72

- hr EC 50 (growth rate) = NES (> 0.250 mg/L; CP 30 and CP 4 0) 72

- hr EC 50 (growth rate) = NES (> 0.125 mg/L; Hordaflex LC 50) 72

- hr NOEC ≥ 0.250 mg/L (CP 30 and CP 40) 72

- hr NOEC ≥ 0.125 mg/L (Hordaflex LC 50)



### B

**- 2** 

**- 4**  **Chronic Fish Toxicity**
No data are available.



### B

**- 2** 

**- 5**  **Chronic Aquatic Invertebrate Toxicity** (1) Frank 1993 and Frank Steinhäuser 1994 The following data review of Frank (1993) and Frank and Steinhäuser (1994) were directly excerpted from the U.K. Environmental Risk Assessment of long

- chain chlorinated paraffins:

The test substance was identified a s C 18

- 20 chlorinated paraffin with 52% chlorination. Frank (1993) carried out a series of acute and longer

- term studies with Daphnia magna using a commercial C 18

- 20 , 52% wt. Cl product. The tests were carried out using dilutions of the
water

- soluble fracti on of the chlorinated paraffin. Stock solutions of the chlorinated paraffin were made up in water to give nominal concentrations of either 100 mg/L or 10 g/L. The 100 mg/L solution was sonicated for one hour and then left to stand in the dark for 48 hours before use. The 10 g/L solution also stood for 48 hours in the dark before use, but this time without sonication. After this period, both solutions were filtered firstly with glass filters and then with membrane filters to remove undissolved test material (microscopic and spectroscopic investigation of the filtered solutions gave no indication of the presence of droplets) to give the respective water

- soluble fractions. The concentration of the chlorinated paraffin in the water

- soluble fractions was determin ed by AOX (adsorbable organic halogen) analysis. The detection limit of the method used was around 10 μg Cl/L, which is equivalent89 to around 20 μg/L of the chlorinated paraffin. This analysis showed that the concentration of chlorinated paraffin present in the water

- soluble fraction was around 462

– 519 μg/L for the 10 g/L nominal solution but was not detectable in the 100 mg/L solution (i.e. <20 μg/L). Experiments were carried out to show that in the test vessels, although the concentration of chlorinated pa raffin present fell over time, it remained within 80 per cent of the initial concentration over 2

– 3 days. This time period was used in the long term tests as the renewal period for the solution (semi

- static method).

Long

- term (21

- day) reproduction studies were also performed using dilutions of the water

soluble fractions of the two stock solutions. The dilutions used were 1:2, 1:4, 1:8 and 1:16 for the 100 mg/L loading and 1:4, 1:8, 1:16, 1:32 and 1:64 for the 10 g/L loading. In these experiments, the test medium was changed three times per week and 10 animals were used per concentration. The tests were carried out at 20°C and a pH of 7.79

– 8.44. Two endpoints were determined in the study: effects on parent mortality and effects on reproduction (number of of fspring per adult). Parent mortality in the controls was 0 per cent in the test carried out with the 10 g/L nominal stock solution and 10 per cent in the test carried out with the 100 mg/L nominal stock solution.

Elevated mortality was seen in the exposed populations. For the 10 g/L stock solution the LOEC was determined as the 1:8 dilution (approximately 58

– 65 μg/L) and the NOEC was determined as the 1:16 dilution (approximately 29

– 32 μg/L). For the 100 mg/L nominal stock solution the LOEC was determined as the 1:4 dilution and the NOEC was determined as the 1:8 dilution. These dilutions are based on the detection limit for the analysis of the 100 mg/L stock solution, and equate to LOEC and NOEC [values] of <5 and <2.5 μg/L respectively). From the dose res ponse curves it appears that 100 % parent mortality occurred at a concentrations of around <10 μg/L in the 100 mg/L nominal stock solution experiments and around 125 μg/L in the 10 g/L nominal stock solution experiments.

For the reproduction endpoint, the average number of young per adult in controls was 72.3 in the 100 mg/L nominal stock solution series of experiments and 73.5 in the 10 g/L nominal stock solution experiments. A significant reduction in the number of young per adult was seen in some of the exposed organisms. For the 100 mg/L nominal stock solution, this effect on reproduction was significantly different from the control groups at the lowest concentration tested (a 1:16 dilution which is equivalent to a chlorinated paraffin concentration of <1.2 μg/L, based on the detection limit of the analytical method used). Thus the NOEC/LOEC for this series of experiments was <1.2 μg/L. Similarly, for the 10 g/L nominal stock solution effects were again seen at the lowest concentration tested (a 1:64 dil ution, which is equivalent to a chlorinated paraffin concentration of 7.3

– 8.1 μg/L). This value is treated as the LOEC for this series of experiments. The report also indicates that the NOEC is very close to this value, since using a different statistical method (Dunnett’s Test rather than Williams’ Test), the effects seen at this concentration were not statistically significantly different from controls.

EPA/OPPT Conclusions The interpretation of the results is complicated by the difficulties interpreting the effects from the different loading and dilution concentrations used. The actual exposure concentration in the 100 mg/L nominal stock solution is unknown and the measured concentration in the 10 g/L nominal stock solution (500 μg/L) is above the report ed (estimated) water solubility of LCCPs. It was also noted in the U.K.90 environmental risk assessment, that the data for the 10 g/L loading were reanalyzed by Thompson (2001). The reanalysis suggests that the statistical significance of parent mortality is questionable and that the 1:8 dilution (20 % mortality) considered as the LOEC is a marginal effect at best. The NOEC for parent mortality could be considered to be the 1:8 dilution (0.058

– 0.065 mg/L). In addition, there was a serious error found by Tho mpson (2001) in the statistical method. They determined that the statistical software misinterpreted increasing dilutions as increasing concentrations based on how the data were entered. Re

- analysis of the data showed that effects were statistically signif icant compared with the controls only at the 1:8 dilution, leading to a NOEC at the 1:16 dilution (0.029

– 0.032 mg/L). This problem with the statistical analysis may also explain why the NOEC and LOEC from the 100 mg/L stock were observed below the lowest concentration tested; thus, the NOEC/LOEC of < 0.0012 mg/L is questionable. The data from these studies should be approached with caution due to deficiencies and uncertainties with the statistical analysis. The NOEC and LOEC values that follow are questio nable. 21

- day NOEC = 0.029

– 0.032 mg/L (10 g/L solution; parent mortality) 21

- day LOEC = 0.058

– 0.065 mg/L (10 g/L solution; parent mortality) 21

- day NOEC < 0.0025 mg/L (below detection limit; 100 mg/L solution; parent mortality) 21

- day LOEC ≤ 0.0050 mg/L (below detection limit; 100 mg/L solution; parent mortality) 21

- day NOEC < 0.0073 mg/L (10 mg/L solution; reproduction) 21

- day LOEC ≤ 0.0073 mg/L (10 mg/L solution; reproduction) 21

- day NOEC < 0.0012 mg/L (100 mg/L solution; reproduction) 21

- day LOEC ≤ 0.0012 mg/L (100 mg/L solution; reproduction )

(2) TNO 1993 The test was carried out according to the OECD [211] methodology using a semi

- static test procedure (test solution renewal was carried out every 48

– 72 hours). The test substance was identified as C 18

- 20 chlorinated paraffin with 52% chlorination (Chloroparaffin Hoechst 56 Flüssig, Chloroparaffin Hoechst 52 Flüssig, and Hordaflex LC50. The dilution water used was a synthetic medium (DSWL) prepared by the addition of various salts to ground water. The hardness of the medium was 214 mg/L as CaCO 3 . The test was carried out using saturated solutions of the chlorinated paraffin using a column technique. The column was prepared by firstly dissolving/suspending 0.1 g of the test substance in 25 mL of ace tone. This solution was then added to 10 g of the packing material for the column (chromosorb 60/80 mesh) and the acetone removed by rotation evaporation. The coated packing material was stored at room temperature in the dark until needed. The columns were stainless steel (25 cm long with an internal diameter of 4.3 mm) filled with 1 g of the coated packing material. The column was conditioned by pumping dilution water through at a flow rate of 6.2 mL every three minutes; the first 500 mL was collected and discarded. Around 18 litres of dilution water was then collected in a bottle and continually re

- circulated through the column at a flow rate of 3.4 mL per minute throughout the test. The required amount of the saturated solution needed for the start of the test, and at each renewal period, was then taken from the bottle.91 Four replicates (10 daphnids in 400 mL of test solution) were carried out for each treatment. The tests were performed in 600 mL beakers and these were conditioned to the test solutions f or two days prior to the start of the test. The solutions were renewed every Monday, Wednesday and Friday during the test.

The concentration of test substance was determined at each renewal time in both the “fresh” solution and the “spent” solution. The analytical method used was based on extractable organic halogen (EOX; similar in principle to AOX) analysis. The mean EOX measured in the test solution over the course of the test was around 1 μg/L (the range found in the “fresh” solutions was 1.0

– 1.5 μg/L and the range in the “spent” solutions was 0.5

– 1.5 μg/L. The EOX concentrations in the control solutions were generally <0.5 μg/L in the “fresh” solution but the range found in the spent solutions was <0.5

– 1.0 μg/L for the blank control and 0.5 2.0 μg/L f or the column control.

The temperature, DO, and pH during the test were 19.2

- 20.3 oC, ≥7.1 mg/L and 7.6

- 8.5, respectively. The parent survival in both of the control groups was 97.5 %. In the C 18

– 20 , 52% wt. Cl treatment group the parent survival was 92.5 %. This survival was not significantly different (at the p=0.05 level) from the control group. Therefore it was concluded that no treatment related effects on parent morality were seen in the study. The reproduction rate (expressed as the cumulative number of young per living female) in the study was 113.6 in the blank control and 127.3 in the column control. The response of the two controls was not significantly different (at the p=0.05 level). The reproduction rate in the C18

– 20, 52% wt. Cl treatmen t group was 100.8. This was 88.7 per cent of the blank control response and 79.1 per cent of the column control response. These responses were analyzed statistically in TNO (1993) using the two

- tailed Dunnett

- test. No statistically significant differences were found between the C 18

– 20 , 52% wt. Cl treatment group and the blank control, or the column control.

EPA/OPPT Conclusions Thompson (2007) conducted a further analysis of the results, due to the author’s perception of discrepancies in the UK’s (2009) E nvironmental Risk Assessment. The discrepancy was based on a seemingly contradictory statement from pg. 7 of the TNO report where the reproduction rate of the Hodaflex LC50 and Hoechst 52 Flüssig was stated as being significantly different from the control s (Pg.7 of the report) but not significantly different from the controls (Pg. 16 of the report). The data were re

- analyzed by Thompson (2007) and no differences in the reproduction rate were observed between the C 18

- 20 , 52 wt % chlorination treatment group s Hodaflex LC50 and Hoechst 52 Flüssig, the column control, blank control or the pooled control group (combined blank control and column control). An additional statistical analysis was conducted by OPPT and no differences were determined for reproduction rate in the treatment groups versus controls using SAS (v. 9.3). Regardless of the discrepancy in the TNO report, statistically significant differences were reported by Thompson (2007) for adult mortality, reproduction rate, and mortality in general for Da phnia exposed to Hoechst 52 Flüssig compared to both controls. However, it is still unclear as to the levels of saturation of each solution the organisms were exposed to based on the extraction technique and nominal loading rates. Therefore, EPA/OPPT will reserve judgment on the acceptability of this study based on a weight of evidence approach.92 21

- day NOEC = 0.002 mg/L (reproduction Hoechst 56 Flüssig)



### B

**- 2** 

**- 6**  **Chronic Aquatic Sediment Invertebrate Toxicity** No data are available. Data from secondary sources using medium

- chain chlorinated paraffins for chronic aquatic sediment invertebrate toxicity can be used to fill data gaps for the long

- chain chlorinated paraffin; and these data were described above in the MCCP section.



### B

**- 2** 

**- 7**  **Avian Toxicity**
No data are available. Data from secondary sources using medium

- chain chlorinated paraffins for avian toxicity can be used to fill data gaps for the long

- chain chlorinated paraffins; and these data were described above in the MCCP section.



### B

**- 2** 

**- 8**  **Terrestrial Invertebrate Toxicity**
No data are available. Data from secondary sources using medium

- chain chlorinated paraffins for terrestrial invertebrate toxicity can be used to fill data gaps for the long

- chain chlorinated paraffins; and these data were described above in the MCCP section.



### B

**- 2** 

**- 9**  **Terrestrial Plant Toxicity** No data are available. Data from secondary sources using medium

- chain chlorinated paraffins for terrestrial plant toxicity can be used to fill data gaps for the long

- chain chlorinated paraffins; and these data were described abo ve in the MCCP section.



### B

**- 2** 

**- 10**  **Conclusions** Sufficient data were available to characterize the acute fish, the acute aquatic invertebrate, the chronic aquatic invertebrate, the chronic aquatic sediment invertebrate, avian, and terrestrial plant toxicity endpoints for MCCPs and LCCPs by read across in some instances. Data for other toxicity endpoints (i.e., chronic fish, aquatic plant, etc.) were inconclusive due to lack of study details, uncertainties in analytical methods, or test material preparation me thods; thus, these data are included in order to characterize risk in a qualitative manner, but are used as supportive for the categories under which they are provided. Supporting data were included in order to provide a weight

- of

- evidence approach used to characterize some endpoints.

Most of the data provided in this review indicated several difficulties were encountered when testing in an aquatic environment. These included: (1) getting the material into solution, (2) measuring the material in solution, and (3) characterizing the effects for each study listed. Often there were many details of a given study omitted, prohibiting a full and robust review of the data. The (estimated) p hysical

- chemical properties of LCCPs and MCCPs (water solubility values be tween 2 and 30 μg/L and Log K ow values between 4

- 8) suggest these materials may not partition to the aquatic media or elicit toxicity to aquatic organisms within the water column.

Specifically for the chronic and acute aquatic invertebrate, aquatic sedim ent, avian, and terrestrial plant endpoints for LCCPs, other analog data provided was acceptable using compounds with chlorination percentage of 52 wt % and carbon chain lengths of C 14

- 17 which is defined as a MCCP material. These data are used in this ass essment to fill data gaps for the C 18

- 20 LCCPs as this would be a conservative approach to charactering hazard in the absence of data. Concern concentrations based on these data are again a very conservative approach in the93 absence of data for the LCCP mat erials themselves and therefore may not inherently characterize toxicity to LCCPs directly.

The most reliable and acceptable studies indicate that for MCCPs, the toxicity to aquatic organisms for acute endpoints are from the Thompson et al. 1996 study for aquatic invertebrates. Where the 48

- hour EC 50 value = 0.0059 mg/L. Using the methods described in the Su stainable Futures/P2 Manual (US EPA, 2012), the acute and chronic concentrations of concern (CoC) are derived as follows:

• Acute CoC: The 48

- hour EC 50 v alue = 0.0059 mg/L is divided by an assessment factor of 5 to yield an acute concentration of concern (CoC) of 0.00118 mg/L, or 0.001 mg/L, or 1 μg/L (1 ppb). Aquatic Acute Concern Concentration= 1 ppb

• Chronic CoC: The aquatic invertebrate chronic value of 0.013 mg/L from the 1997 Thompson et al. study based on a MCCP material is divided by an assessment factor of 10 to yield 0.0013 mg/L or 1.3 μg/L or 1.3 ppb. Aquatic Chronic Concern Concentration = 1 ppb

For LCCPs, the acute concern concentration may be derived from the Johnson and Finely (1980) studies. For the chronic concern concentration, the results from the Thompson et al. 1997 study based on a MCCP material will be used as a conservative qualitative assessment due to the lack of overly reliable da ta for this endpoint for LCCPs.

• Acute CoC: Aquatic Acute Concern Concentration= NES?????

• Chronic CoC: The aquatic invertebrate chronic value of 0.013 mg/L from the 1997 Thompson et al. study based on a MCCP material is divided by an assessment factor of 1 0 to yield 0.0013 mg/L or 1.3 μg/L or 1.3 ppb. Aquatic Chronic Concern Concentration = 1 ppb (MCCP and LCCP)

The most reliable and acceptable value for the toxicity to aquatic sediment invertebrate organisms acute endpoint is based on the MCCP material fr om the Thompson et al. 2002 28

- d study. For both MCCPs and LCCPs, the 28

- d sediment invertebrate GMATC value of 187 mg/kg dry wt sediment is used to assess hazard. The 28

- d sediment invertebrate GMATC value of 187 mg/kg dry wt sediment is used to assess ha z ard. Again, using methods in US EPA (2012):

• Acute CoC: Calculating an acute concern concentration from the chronic value of 187 mg/kg dry wt. The 187 value is multiplied by an acute to chronic ratio for invertebrates (10) to yield 1,870 mg/kg dry wt. Thi s value is then divided by an assessment factor of 5 to yield 374 mg/kg dry wt. Aquatic Sediment Acute Concern Concentration = 374 mg/kg dry wt sediment. (MCCP and LCCP)

• Chronic CoC: The 28

- d sediment invertebrate GMATC of 187 mg/kg dry wt sediment is divi ded by an assessment factor of 10 to yield 18.7 mg/kg dry wt sediment. Aquatic Sediment Chronic Concern Concentration = 19 mg/kg dry wt sediment. (MCCP and LCCP)

The most reliable and acceptable value for the toxicity to terrestrial invertebrates acute en dpoint is based on the MCCP material from the Thompson et al. 2001a study. For LCCPs, the 28

- d terrestrial invertebrate GMATC value of 149 mg/kg dry wt sediment will be used as an analog94 approach to assess hazard. The 28

- d terrestrial invertebrate GMATC va lue of 149 mg/kg dry wt sediment from this study will be u sed. Again, using methods in US EPA (2012):

• Acute CoC: Calculating an acute concern concentration from the chronic value of149 mg/kg dry wt, this value is multiplied by an acute to chronic ratio fo r invertebrates (10) to yield 1,490 mg/kg dry wt. This value is then divided by an assessment factor of 5 to yield 298 mg/kg dry wt. Terrestrial Invertebrate Acute Concern Concentration = 298 mg/kg dry wt sediment. (MCCP and LCCP)

• Chronic CoC: The 28

- d terrestrial invertebrate GMATC of 149 mg/kg dry wt is divided by an assessment factor of 10 to yield 14.9 mg/kg dry wt sediment. Terrestrial Invertebrate Chronic Concern Concentration = 15 mg/kg dry wt sediment. (MCCP and LCCP)

The most reliable and acceptable value for the toxicity to terrestrial plants is based on the MCCP material from the Thompson et al. 2001ab study. For LCCPs, the analog approach using the values from this study may be used. However, there is no OPPT guidance regarding assessing concern concentrations for terrestrial plants.95 

## Appendix C HUMAN HEALTH HAZARD STUDY SUMMARIES

C

- 1 MCCP Health Data Review There is no information on inhalation absorption of MCCPs in humans or in animals. Based on their low vapor pressure and low water solubility, a bsorption following inhalation or dermal exposure is expected to be limited. Some MCCPs demonstrated moderate absorption and metabolism following oral exposure in animals. In general, absorption and metabolism are related to their carbon chain length and d egree of chlorination; the longer the carbon chain length and the higher the degree of chlorination, the less absorption and metabolism.

No information is available on the toxicity of MCCPs in humans; however, the toxicology of these compounds has been evaluated in experimental animals. Studies in rats and rabbits have shown that MCCPs only caused slight skin irritation and have low eye irritation potential. No evidence of skin sensitization was found when tested in guinea pigs. The liver, kidney and thy roid are the target organs of MCCPs in oral repeated dose studies in experimental animals. MCCPs induced increased liver weight, enzyme activity, and histopathological changes at high dose levels. Some of these hepatic effects are likely related to an incr ease in metabolic demand as an adaptive response, as well as to peroxisome proliferation, which are considered of limited toxicological significant to humans. However, liver necrosis was observed in a 90

- day study in rats at 360 mg/kg

- bw/day; this effect i s considered relevant to humans. The reported effects in the kidney may have been produced by the parent compound or from metabolites. Mechanistic data cannot totally rule out that some kidney effects are relevant to humans. From the data available, a LOAE L of 625 mg/kg

- bw/day based on histopathological changes in the kidneys of female rats is identified in a 90

- day toxicity study, and a NOAEL of 23 mg/kg

- bw/day based on increased kidney weight at 222 mg/kg

- bw/day is identified from another 90

- day study in rats. Repeated dose studies in rats reported some changes in histopathology and hormone levels of the thyroid. However, it may be concluded based on an evaluation of the mechanistic data that the thyroid effects observed in rats is of little relevance to c hronic toxicity in humans.

There is no information on the carcinogenicity of MCCPs; however, carcinogenicity studies on a short

- chain chlorinated paraffin (SCCP) and a long

- chain chlorinated paraffin (LCCP) are available. These studies, along with the gen otoxicity data on MCCPs, may be used to inform the carcinogenic potential of MCCPs. When administered by gavage, a SCCP (C 12 , 60 wt% Cl) caused increased incidences of liver tumors in male and female rats, kidney tumors in male rats, and thyroid tumors in female rats. However, based on mechanistic considerations, these tumors are considered to be of little or no relevance to humans. An increased incidence of malignant lymphoma in male mice was reported at the highest dose of 5,000 mg/kg

- bw/day in carcinogen icity studies of a LCCP (C 23 , 43 wt% Cl) in male and female rats and mice. However, malignant lymphoma is one of the more variable tumors in mice and has a viral origin in many

cases. No increased incidence of malignant lymphoma was observed in the carcino genicity study on an SCCP. Further, MCCPs are non

- genotoxic. Therefore, it may be concluded that MCCPs are unlikely to pose a carcinogenic hazard to humans.

When evaluating the risks of workers from exposure to MCCPs based on the available repeated

dose t oxicology studies, the EU’s draft Risk Assessment Report (RAR) on MCCPs concluded that except metal working fluids (MWF) use, “ There is at present no need for further96 information and/or testing or for risk reduction measures beyond those which are being ap plied already ” (EURAR, 2008).

Using the NOAEL (23 mg/kg

- bw/day) of kidney toxicity identified in the 90

- day oral study in rats (CXR, 2005), a MOS of 70,000 was estimated for dermal exposure of consumers resulting from wearing leather clothes treated with MCCP. For inhalation exposure of consumers using metal fluids containing MCCP, a MOS of 2,875 was obtained. Therefore, it was also concluded that: “ There is present no need for further information and/or testing or for risk reduction measures beyond those which are being applied already ” (EURAR, 2008).



### C

**- 1** 

**- 1**  **Metabolism** There is no information on inhalation absorption of MCCPs in humans or in animals. Based on their low vapor pressure and low water solubility, absorption following inhalation or dermal exposure i s expected to be limited. An in vitro study using human skin showed that after 24 hours, approximately 0.7 % of a C 15 chlorinated paraffin was absorbed (Scott, 1984; cited in: EURAR, 2008). Oral studies (IRDC, 1984, CXR, 2005; cited in: EURAR, 2008) showed that approximately 50 % of a single dose of [8

- 14 C]

- labeled C 15 chlorinated paraffin (52 wt% Cl) was absorbed from the GI tract in rats. Excretion via feces was the major route of elimination of radiolabeled material. Elimination of radioactivity from bod y tissues occurred with an elimination half

- life of approximately 2

- 5 days (liver and kidney) or approximately 2 weeks (adipose tissue).



### C

**- 1** 

**- 2**  **Acute Toxicity** There is no information on the effects of a single exposure to MCCPs in humans. No deaths and only lim ited, non

- specific clinical signs of toxicity resulting from exposure of rats to very high doses were observed in an acute oral toxicity study of MCCPs (C 14

- 17, 51

- 60 wt% Cl); the LD 50 was reported to be > 4,000 mg/kg bw (Birtley et al ., 1980; cited in : IP CS, 1996). Though no acute toxicity data are available for MCCP by the inhalation or dermal route of exposure, the low acute toxicity data for SCCPs by these routes suggest that MCCPs are likely to have low acute inhalation and dermal toxicity.



### C

**- 1** 

**- 3**  **Irritation and Sensitization**  No signs of skin irritation were seen with MCCPs (C 14

- 17 , 45 wt% Cl), and only slight erythema on the shaved skin was reported in one rabbit at 24 hours exposed to MCCPs (C 14

- 17 , 40 wt% Cl) (Chater, 1978; cited in: EURAR, 2008). A mild s kin irritancy response was reported in one of nine unpublished skin irritation studies of MCCPs (C 14

- 17 , 51

- 60% Cl) in rats (Birtley et al

., 1980; cited in: EURAR, 2008). The material caused slight, transient eye irritation in rabbits (Birtley et al ., 1980; Kuhnert et al ., 1986; cited in: EURAR, 2008).

No skin sensitization reactions were produced in guinea pig maximization tests conducted on MCCPs (C 14

- 17 , 40

- 45 wt% Cl) (Murmann, 1988; Chater, 1978; cited in: EURAR, 2008).



### C

**- 1** 

**- 4**  **Repeated**

**- dose Toxicity**  There are a number of repeated dose toxicity studies (up to 90

- days duration) of MCCPs (C 14

- 17 40 wt% Cl or 52 wt% Cl) in rats by oral exposure (CXR, 2005; Poon et al ., 1995; IRDC, 1984; Birtley et al ., 1980; and Wyatt et al ., 1997; cited in: EURAR, 2008). Though the quality and reliability of these studies differs, the liver, kidney, and thyroid were consistently established as the target organs. A summary of the results from these studies is provided in Table 1.97 MCCPs caused an increase in liver weigh t in male rats at exposure levels of > 100 mg/kg

bw/day) and in female rats at exposure levels of > 32 mg/kg

- bw/day. Liver enzyme induction was reported in male and female rats starting from 222 and 100 mg/kg

- bw/day, respectively. Liver hypertrophy of trac e to minimal severity was reported in male rats at dose levels of > 100 mg/kg

- bw/day and higher. Collectively, these changes are likely to be related to an increase in metabolic demand as an adaptive response and to peroxisome proliferation, both of which are considered of no or limited toxicological significance to humans. Though Poon et al . (1995) reported various histopathological effects in the liver of male and female rats at dose levels > 36 mg/kg

- bw/day, there are a number of deficiencies with this s tudy, including the scoring and classification of histopathological findings and limited reporting of data, which preclude its utility in hazard evaluation. This conclusion is consistent with previous evaluations of this study (EURAR, 2008). Further, despi te the consistency of findings reported in the review article by Birtley et al . (1980) with other 90

- day studies, these findings should be viewed cautiously because the original full study report is not available. Based on the available data, the studies b y IRDC (1984) and CXR (2005) provide the most reliable data for identifying effect levels of MCCPs on the liver. For the purposes of this assessment, a NOAEL of 100 mg/kg

- bw/day was chosen based on increases in absolute liver weight ( i.e. , 22

- 26 %), liver hypertrophy of trace severity, and enzyme induction ( i.e. , 30 % increase).

Kidney effects have been reported in a number of studies, with effect levels typically being observed at the limit doses. MCCPs (C 14

- 17 52 wt% Cl) caused significant increases (9

- 13 %) in kidney weight at 222 mg/kg

- bw/day (CXR, 2005; cited in: EURAR, 2008), as well as “chronic nephritis” and tubular pigmentation in the kidney of female rats at 625 mg/kg

- bw/day (IRDC, 1984, cited in: EURAR, 2008). One study reported a dose

- related increase in congestion starting at 32 mg/kg

- bw/day; however, no information was provided on the incidence or severity of this effect (Birtley et al ., 1980; cited in: EURAR, 2008). An additional study reported minimal to mild hyaline

- droplet like cytoplasmic inclusions, starting at > 0.4 mg/kg

- bw/day in male rats. This effect is considered of limited relevance to humans. The authors also reported minimal dose

- related increases with inner medullary tubular dilat ion at an incidence of 1/10, 4/10, and 8/10 female rats at 4, 42, and 420 mg/kg

- bw/day, respectively (Poon et al ., 1995; cited in: EURAR, 2008). Though this effect is considered relevant to humans, the study suffers from a number of limitations, which prec lude utilizing it for hazard evaluation. However, based on the incidence reported by the authors, the NOAEL of 42 mg/kg

- bw/day for kidney effects is consistent with the NOAEL of 23 mg/kg

- bw/day reported in the CXR (2005) study. Therefore, a NOAEL of 23 mg/ kg

- bw/day was chosen for the kidney, based on increases in organ weight at the next highest dose level.

MCCPs (C 14

- 17 52 wt% Cl) have been reported to cause minimal to mild adaptive histopathological changes in the thyroid ( i.e. , follicular cell hypertro phy and hyperplasia) in two studies in rats starting at 50 ppm (4 mg/kg

- bw/day) and above (Poon et al ., 1995; IRDC, 1985). Decreased T 4 levels and increased TSH levels in the plasma were also seen at similar dose levels. As noted previously, these results have been drawn into question based on the scoring and classification for histopathology, the limited reporting of data, and the inconsistent findings from other more robust studies (EURAR, 2008). Therefore, these studies will not be considered further for hazard identification. IRDC (1985) reported mild to moderate hypertrophy and hyperplasia in male rats at dose levels of > 10 mg/kg

- bw/day and higher, whereas changes in absolute organ weights of male and female rats were not observed except at the limit d ose of 625 mg/kg

- bw/day (IRDC, 1985; cited in: EURAR, 2008). The remaining studies that evaluated thyroid hormone levels identified a decrease in plasma free T 3 in male rats, but not total T 3 or98 free/total T 4 , and an increase in TSH in female rats at dose levels of 24.6 or 242 mg/kg

- bw/day, respectively (CXR, 2005; cited in: EURAR, 2008), or fluctuations in thyroid hormones in male or female rats at doses of > 312 mg/kg

- bw/day or higher (Wyatt et al ., 1997; cited in: EURAR, 2008). There is evidence that the thyroid effects observed are attributable to stimulation of this organ arising from a negative feedback effect arising from plasma T 4 depletion following increased excretion of this hormone. This depletion of plasma T 4 results from the induction of hepatic UDPG

- transferase, increased glucuronidation, and ultimately excretion of T 4 following exposure to MCCPs. The pituitary responds to the decreased levels of T 4 by releasing more TSH, which in turn leads to increased pro duction of T 4 by the thyroid. The continuous stimulation of the thyroid in response to the increased excretion of plasma T 4 is predicted to ultimately give rise to hypertrophy and hyperplasia in this organ. Humans, unlike rodents, possess T 4 - globulin bindi ng protein and are therefore less susceptible to plasma T 4 depletion and hence any resultant thyroid stimulation. The thyroid effects observed in rats are not considered to be of relevance to chronic human health at relevant levels of exposure, although th ese changes may be relevant for assessing potential adverse outcomes during reproduction and development, as discussed under section.



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**- 1** 

**- 5**  **Genotoxicity** MCCPs (C 14

- 17 40

- 52 wt% Cl) are not mutagenic to bacteria. Three in vivo bone marrow studies also show that MCCPs are not clastogenic ( cited in: EURAR, 2008). Therefore, it may be concluded that MCCPs possess a low potential to cause genotoxic effects.



### C

**- 1** 

**- 6**  **Carcinogenicity** There is no information on the carcinogenicity of MCCPs. When administered by gavage, a SCCP (C 12 , 60 wt% Cl) caused increased incidences liver tumors in male and female rats, kidney tumors in male rats, and thyroid tumors in female rats (NTP, 1986). However, on mechanistic considerations, these tumors are considered to be of little or no relevanc e to humans. This conclusion is consistent with previous carcinogenicity evaluations (EURAR, 2008). An increased incidence of malignant lymphoma in male mice was reported at the highest dose of 5,000 mg/kg

- bw/day in carcinogenicity studies of a LCCP (C 23 , 43 wt% Cl) in male and female rats and mice. However, malignant lymphoma is one of the more variable tumors in mice and has a viral origin in many cases. No increased incidence of malignant lymphoma was observed in the carcinogenicity study on a SCCP (C 12 , 60 wt% Cl). Based on structure

- activity relationships and the absence of positive genotoxicity data on MCCPs, the available carcinogenicity studies on a SCCP and a LCCP suggest that MCCPs are not expected to pose a carcinogenic hazard to humans.99 Table Apx_C

- 1: Summary of Results from 90

- day S tudies in Rats A dministered MCCPs. Strain (sample size) Test substance and dose levels Target
organ Effect levels

F

- 344 (10 rats/sex/group) 1 C 14

- 17 , 52 wt% Cl

Dietary intake for ♂: 0, 2.38, 9.34, 23.0, or 222 mg/kg

- bw/day.

Dietary intake for ♀: 0, 2.51, 9.70, 24.6, or 242 mg/kg

- bw/day. Liver ♂ at 222 and ♀ at 242 mg/kg

- bw/day, 13

- 31 % ↑ in organ weight
♂ at 222 mg/kg

- bw/day, minimal centrilobular hypertrophy in 9/10 a nimals
♂ at 222 mg/kg

- bw/day, 82 % ↑ in microsomal T 4 - UDPGA

- glucuronyl transferase activity

♀ at 100, 300, and 300 mg/kg

- bw/day, 30, 30, and 252 % ↑ in microsomal T 4 - UDPGA

- glucuronyl transferase activity, respectively

Kidney ♂ at 222 and ♀ at 242 mg/kg

- bw/day, 9

- 13 % ↓ in organ weight
♂ at > 222 and ♀ at 242 mg/kg

- bw/day, no treatment

- related histopathology

Thyroid ♂ at 222 mg/kg

- bw/day, 17 % ↑ in plasma TSH
♂ at 23.0 and 222 mg/kg

- bw/day, 26 % or 22 % ↓ in plasma free T 3 , respectively, but no effects on total T 3 or on free/total T 4 at any dose

♀ at > 242 mg/kg

- bw/day, no effects on free/total T 3 or T 4
♀at 24.6 and 242 mg/kg

- bw/day, 20 and 39 % ↑ in plasma TSH
♂at > 222 and ♀ at 242 mg/kg

- bw/day, no treatment

- related histopathology

Sprague

- Dawley (10 rats/sex/group) 2 C 14

- 17 , 52 wt% Cl
Dietary intake for ♂: 0, 0.4, 4, 36, or 360 mg/kg

- bw/day.

Dietary intake for ♀: 0, 0.4, 4, 42, or 420 mg/kg

- bw/day. Liver ♂ at 360 and ♀ at 420 mg/kg

- bw/day, 28 and 48 % ↑ in absolute and relative we ights, respectively
♂ and ♀ at < 4 mg/kg

- bw/day, no treatment

- related histopathology
♂at 36 and ♀ at 42 mg/kg

- bw/day, minimal increase in anisokaryosis and vesiculation of the nuclei
♂ at 360 and ♀ at 420 mg/kg

- bw/day, mild increase in anisokaryosis and vesiculation of the nuclei (7

- 10 animals)

♂ at 360 mg/kg

- bw/day, ↑ in perivenous homogeneity

♀ at 42 and 420 mg/kg

- bw/day, ↑ in perivenous homogeneity

♂ at 360 and ♀ at 420 mg/kg

- bw/day, ↑ in single cell necrosis (incidence not reported)

Kidney ♂ at 360 and ♀ at 420 mg/kg

- bw/day, 11 % ↑ in absolute and relative weights

♂ at > 0.4 mg/kg

- bw/day, minimal to mild hyaline

- droplet like cytoplasmic inclusions, with significant accumulation at100 Table Apx_C

- 1: Summary of Results from 90

- day S tudies in Rats A dministered MCCPs. Strain (sample size) Test substance and dose levels Target
organ Effect levels

the limit dose

♀ at > 4 mg/kg

- bw/day, minimal dose

- related inner medullary tubular dilation seen in 0/10, 0/10, 1/10, 4/10, and 8/10 animals

Thyroid ♂ at 36 and 360 mg/kg

- bw/day, minimal to mild morphological changes affecting the architecture ( i.e. , reduced follicle sizes and collapsed angularity) and the epithelium ( i.e. , increased height, cytoplasmic vacuolation, and nuclear vesiculation)

♀ at > 4 mg/kg

- bw/day, minimal to mild morphological changes affecting the architecture ( i.e. , reduced follicle sizes and collapsed an gularity) and the epithelium ( i.e. , increased height, cytoplasmic vacuolation, and nuclear vesiculation)

F

- 344 (15 rats/sex/group) 3 C 14

- 17 , 52 wt% Cl
Dietary intake for ♂ and ♀: 0, 10, 100, or 625 mg/kg

- bw/day. Liver ♂ and ♀ at 100 and 625 mg/kg

- bw/day, 22

- 26 % and 64

- 92 % ↑ in absolute weight values, respectively
♂ at 100 and 625 mg/kg

- bw/day, hypertrophy of trace severity seen in 1/15 and 13/15 animals, respectively

♀ at 625 mg/kg

- bw/day, hypertrophy of trace severity seen in 13/15 anima ls

Kidney ♂ and ♀ at 625 mg/kg

- bw/day, 18 % ↑ in absolute weight values

♂ at > 10 mg/kg

- bw/day, trace to mild nephritis seen in 1/15, 3/15, 4/15, and 10/15 animals

♀ at 625 mg/kg

- bw/day, tubular pigmentation (9/14 animals)

Thyroid ♂ at 625 mg/kg

- bw/day, 50 % ↑ in absolute weight values

♂ at > 10 mg/kg

- bw/day, mild to moderate hypertrophy observed in controls with a dose

- dependent trend towards ↑ severity in treated animals

♂ at > 10 mg/kg

- bw/day, trace to mild hyperplasia with a dose

- depen dent trend towards ↑ severity Adrenal ♂ and ♀ at 625 mg/kg

- bw/day, 25 % ↑ in absolute weight values

Wistar

- derived (24 rats/sex/group) 4 C 14

- 17 , 52 wt% Cl, containing epoxidized vegetable oil as a stabilizer
Dietary intake for ♂: 0, 33, 167, or 333 mg/kg

bw/day. Liver ♂ at 167 and 333 mg/kg

- bw/day, 15 and 22 % ↑ in relative weight values, respectively

♀ at 32, 160, and 320 mg/kg

- bw/day, 11, 21, and 48 % ↑ in relative weight values, respectivel y ♂ at 333 and ♀ at 320 mg/kg

- bw/day, no histopathological abnormalities

♂ at > 33 and ♀ at > 32 mg/kg

- bw/day, dose

- related ↑ in proliferation of smooth endoplasmic reticulum (electron microscopy)101 Table Apx_C

- 1: Summary of Results from 90

- day S tudies in Rats A dministered MCCPs. Strain (sample size) Test substance and dose levels Target
organ Effect levels

Dietary intake for ♀: 0, 32, 160, or 320 mg/kg

bw/day. Kidney ♂ and ♀ at limit dose, 15 % ↑ in relative weight
♂ and ♀, dose

- related ↑ in congestion (incidence and severity not reported)
♂ and ♀ at > limit dose, no histopathological abnormalities

F

- 344 (10 rats/sex/group) 5 C 14

- 17 , 40 wt% Cl
Oral gavage for ♂ and ♀: 0, 312, or 625 mg/kg

- bw/day Liver ♂ and ♀ at 312 and 625 mg/kg

- bw/day, 37 and 72 % ↑ in relative weight, respectively, (absolute weight and bodyweight not presented)
♂ and ♀, dose

- related ↑ in centrilobular hypertrophy (incidence and severity not reported)
♂ and ♀ at 312 and 625 mg/kg

- bw/ day, dose

- related ↑ in β

- oxidation from day 29 onwards (~2.7

- and 3.3

- fold ↑, respectively, at study termination)
♂ and ♀ at 312 and 625 mg/kg

- bw/day, dose

- related ↑ in UDPG

- transferase activity from day 15 onwards (up to 100 % ↑, respectively)

Thyroid ♂ and ♀ at 312 and 625 mg/kg

- bw/day, ↓ in levels of free and plasma T 3 , which reached statistical significance on days 15 and 57
♂ at 312 and 625 mg/kg

- bw/day, ↑ TSH up to 2

- fold on day 8 only

♀ at 312 and 625 mg/kg

- bw/day, T 3 significantly ↑ by day 91

♀ at 312 and 625 mg/kg

- bw/day, total plasma T 4 significantly ↓ by up to 25 % on day 57
♂ and ♀ at 312 and 625 mg/kg

- bw/day, ↑ follicular cell hypertrophy throughout the study, and accompanied by follicular cell hyperplasia on days 55 and 91 (incidence an d severity not reported)
♂ and ♀ at 312 and 625 mg/kg

- bw/day, significantly ↑ replicative DNA synthesis on day 29, but not on day 91

1 CXR (2005), cited in: EURAR (2008).

2 Poon et al . (1995), cited in: EURAR (2008).

3 IRDC (1984), cited in: EURAR (2008).

4 Birtley et al . (1980), cited in: EURAR (2008); note, this study was only summarized in the review by Birtley et al . (1980). The underlying original study report was not available.

5 Wyatt et al . (1997), cited in: EURAR (2008).102 

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**- 7**  **Developmental Reproductive Toxicity Review** A series of range

- finding and definitive prenatal developmental and reproductive toxicity studies were conducted in rats and rabbits with medium

- chain chlorinated paraffins (MCCPs). These studies were conducted between 1981 and 1986. They appear to be valid toxicity studies, conducted according to the standard methodologies available at the time. More recently, additional studies with MCCPs have been conducted in an attempt to determine th e cause of hemorrhaging in the pups observed in a one

- generation reproductive toxicity range

- finding study.

In several prenatal developmental toxicity studies with MCCPs conducted via gavage, no signs of maternal toxicity were seen at doses as high as 500 mg/kg

- bw/day in rats and 100 mg/kg

bw/day in rabbits. Likewise, no signs of developmental toxicity were observed at doses as high as 5000 mg/kg

- bw/day in rats and 100 mg/kg

- bw/day in rabbits.

Two reproductive toxicity studies with MCCPs in rats have been conducted. A one

- generation reproductive toxicity range

- finding study showed that administration of approximately 100 and 400 mg/kg

- bw/day MCCPs via the diet had no effect on fertili ty or other reproductive parameters; however, internal hemorrhaging and deaths in pups were observed beginning from 74 mg/kg

- bw/day (1000 ppm) up to approximately 400 mg/kg

- bw/day (6250 ppm). These effects in the pups were not seen in a more recent definit ive one

- generation reproductive toxicity study with exposure to MCCPs for 11

- 12 weeks to doses as high as 100 mg/kg

- bw/day (1200 ppm). Internal hemorrhaging was not seen in the adult animals in either of these studies at doses as high as 400 mg/kg

- bw/day ( 6250 ppm), or in another study in non

- pregnant female rats repeatedly exposed to doses as high as 1000 mg/kg

- bw/day. However, when dams were exposed to approximately 500 mg/kg

- bw/day (6250 ppm) MCCPs during cohabitation, gestation, and lactation, signs of hemorrhaging were observed in dams that died at the time of parturition. Taken together, the results of these studies suggest that newborns during lactation and pregnant females at the time of parturition are a potentially sensitive subpopulation.

The UK Risk Assessment (February, 2008) did not use the LOAEL of 74 mg/kg

- bw/day (1000 ppm) from the one

- generation reproductive toxicity range

- finder study as a point of departure because the pup deaths at that dose were not statistically significant. The study itself used a limited number of animals and was intended for dose range

- finding purposes only and, more importantly, the pup deaths were not repeated in a more recently conducted definitive study. With respect to developmental/reproductive toxicity, the U K Risk Assessment identified two subpopulations at risk: offspring during lactation and pregnant dams at parturition. The NOAELs from the definitive one

- generation reproductive toxicity study (a maternal NOAEL ~ 47 mg/kg

bw/day (600 ppm) for effects on the offspring mediated via lactation; and a maternal NOAEL ~ 100 mg/kg

- bw/day (1200 ppm) for effects on the dam during the time of parturition) were used to calculate risk. Assuming a conservative value of 50 % oral absorption, the margin of safety (MOS) for effects on the offspring mediated via lactation and effects on the dam during the time of parturition were calculated for workers, consumers, and other scenarios. In all but one scenario (oil

- based metal working fluids), the margins of safety were above 10 0 and in many cases, several fold above. In addition, margins of exposure were calculated for infants exposed via breast milk and via cow’s milk, and in both instances, large MOEs ( i.e. , > 100) were calculated.103 Additional studies with MCCPs have be en conducted in an effort to clarify the possible causes of the hemorrhaging in the pups. One (single

- dose; 6250 ppm or 538 mg/kg

- bw/day) study showed maternal death during parturition due to low levels of vitamin K and related hemorrhaging, suggesting tha t the act of parturition places dams at higher risk. It was concluded in data from this study and a cross

- fostering study that the fetus relies on clotting factors via mother’s milk and severe deficiencies in vitamin K levels and related clotting factors i n the pups results in hemorrhaging.

No definitive developmental neurotoxicity studies on MCCPs were located. It is not clear if any developmental neurotoxicity endpoints were actually measured in the available prenatal developmental/reproductive toxicit y studies; none were explicitly stated. The only information available regarding behavior during development is from cage

- side observations in pups through LD 21. In these cases, no dose

- related differences were reported in F 1 post

- weaning appearance

or ca ge

- side behaviors.

In the prenatal developmental toxicity study in rats, the LOAEL for maternal toxicity was 2000 mg/kg

- bw/day based on clinical signs. The NOAEL for maternal toxicity was 500 mg/kg/day. The NOAEL for developmental toxicity was 5000 mg/kg

- bw/day, the highest dose tested.

In the prenatal developmental toxicity study in rabbits, no adverse, treatment

- related effects were reported in the dams or the offspring. The NOAEL for both maternal and developmental toxicity was 100 mg/kg

- bw/day, th e highest dose tested.

In the reproduction range

- finding study in rats, the LOAEL for maternal toxicity was 6250 ppm (463 mg/kg

- bw/day) based on reductions in body weight gains. The NOAEL for maternal toxicity was 1000 ppm (74 mg/kg

- bw/day). The LOAEL for developmental toxicity was 1000 ppm (62/74 mg/kg/day) based on pup mortality associated with internal hemorrhages. The NOAEL for developmental toxicity was 100 ppm (6/8 mg/kg

- bw/day). No effects on any reproductive parameters were reported. The NOAEL f or reproductive toxicity was 6250 ppm (384/463 mg/kg

- bw/day).

In the one

- generation reproduction toxicity study in rats , the LOAEL for maternal toxicity was 1200 ppm (~100 mg/kg

- bw/day) based on increases in liver weight; the NOAEL for maternal toxicity was 600 ppm (~ 47 mg/kg

- bw/day). The NOAEL for developmental and reproductive toxicity was 1200 ppm (~ 84/99 mg/kg

- bw/day), the highest dose tested.

Basis for Conclusions In a range

- finding prenatal developmental toxicity study in pregnant Charles Riv er COBS CD rats administered MCCPs (C 14

- 17 , 52 wt% Cl) via gavage at dose levels of 0, 1000, 1500, and 2500 mg/kg

- bw/day on gestation days (GD) 6

- 20, no effects were observed in the dams at doses up to 2500 mg/kg

- bw/day (IRDC, 1983, 1984; cited in: EURAR, 2008). As a result, doses greater than 2500 mg/kg

- bw/day were selected for the definitive study.

In the definitive study, four groups of 25 pregnant Charles River COBS CD rats were administered MCCPs (C 14

- 17 , 52 wt% Cl) via gavage at doses of 0, 500, 2000, and 5000 mg/kg

bw/day on GD 6

- 19 (IRDC, 1984; cited in: EURAR, 2008). Unmated males and females were individually housed and acclimated for 21

- days in an environmentally controlled room. At the104 end of the acclimation perio d, all animals were weighed and subjected to a detailed physical examination. One female and one male rat were placed together for mating. Confirmation of mating was based on evidence of a copulatory plug or by vaginal smear for sperm. The day mating was c onfirmed was designated as day 0 of gestation. Test article was administered to pregnant females orally by gavage as a single daily dose on GD 6

- 19. During treatment, pregnant females were observed daily for mortality and clinical signs of toxicity. Any fe males not surviving to scheduled sacrifice were necropsied. Body weights were recorded on GD 0, 6, 9, 12, 16, and 20. All females were sacrificed on GD 20 and the uterus and ovaries excised for examination. The number and location of viable and nonviable fetuses, early and late resorptions, and the number of total implantations and corpora lutea were recorded. The uterus was weighed. The abdominal and thoracic cavities underwent gross examination. Maternal tissues were preserved for future histopathologica l analysis. Fetuses were weighed, sexed, tagged, and examined for external malformations and variations, including the palate and the
eyes. The fetuses underwent visceral and skeletal examinations for malformations and developmental variations.

The onl y effects reported in dams consisted of an increased incidence in wet matted and yellow stained haircoat in the anogenital area at 5000 mg/kg

- bw/day, and soft stool at > 2000 mg/kg

bw/day. No treatment

- related adverse effects were reported in offspring at doses up to 5000 mg/kg

- bw/day. The LOAEL for maternal toxicity was 2000 mg/kg

- bw/day based on clinical signs; the NOAEL for maternal toxicity was 500 mg/kg

- bw/day. The NOAEL for developmental toxicity was 5000 mg/kg

- bw/day, the highest dose tested.

In a range

- finding prenatal developmental toxicity study in pregnant Dutch Belted rabbits administered MCCPs (C 14

- 17 , 52 wt% Cl) via gavage at dose levels of 0, 100, 300, 1000, 2000, and 3000 mg/kg/day on GD 6

- 27, an increase in the number of abortions was obs erved at > 1000 mg/kg/day (IRDC, 1982a; cited in: EURAR, 2008). Body weight reductions in the dams were reported at 100 and 300 mg/kg/day. As a result, another range

- finding prenatal developmental toxicity study in rabbits was initiated. This second range

- finding study showed decreases in maternal weight gain at 80 and 160 mg/kg

- bw/day (IRDC, 1982b; cited in: EURAR, 2008).

Based on the results of these range

- finding studies, dose levels of 10, 30, and 100 mg/kg

- bw/day were selected for the definitive pren atal oral gavage developmental toxicity study (IRDC, 1983; cited in: EURAR, 2008). In the definitive study, four groups of 16 pregnant Dutch Belted rabbits were administered 0, 10, 30, and 100 mg/kg

- bw/day MCCPs (C 14

- 17 , 52 wt% Cl) via gavage on GD 6

- 27. U nmated males and females were individually housed and acclimated for 50

- days in an environmentally controlled room. As a result of a positive finding for parasites in stool samples collected during acclimation, all rabbits received sodium sulfamethazine in their drinking water for 16 days during the acclimation period. This treatment was terminated 4 weeks prior to study initiation and only rabbits testing negative for parasites were placed on study. At the end of the acclimation period, all animals were we ighed and subjected to a detailed examination. Females were impregnated via artificial insemination. Three weeks prior to artificial insemination, females were given chorionic gonadotropin via an injection in a marginal ear vein in order to induce superovu lation. Semen was collected from males of proven fertility and evaluated for motility. The day of artificial insemination was designated as day 0 of gestation. During treatment, pregnant females were observed for mortality and clinical signs of toxicity. Body weights were recorded on GD 0, 6, 12, 18, 24, and 28. Any females not surviving to scheduled sacrifice were necropsied. On GD 28, all surviving females were sacrificed and the105 uterus and ovaries excised for examination. The location and number of viab le and nonviable fetuses, early and late resorptions, and the number of total implantations and corpora lutea were recorded. The uterus was weighed. The thoracic and abdominal cavities underwent gross examination. Pooled samples of abdominal adipose tissue from 3 dams were frozen for future analysis. Each fetus was sexed, weighed, and examined for external malformations and variations, including the palate and the eyes, as well as visceral and skeletal examinations for malformations and developmental variat ions, including examination of the brain and the heart.

No adverse, treatment

- related effects were reported in the dams or the offspring at doses up to 100 mg/kg

- bw/day. The NOAEL for both maternal and developmental toxicity was 100 mg/kg

- bw/day, the highest dose tested.

In a one

- generation reproductive toxicity range

- finding study, four groups of 5 male and 10 female Charles River COBS SC rats were administered MCCP (C 14

- 17 , 52 wt% Cl) via the diet at 0, 100, 1000, and 6250 ppm (~ 0, 6, 62, and 384 mg/kg

- bw/day, respectively, in males; and 0, 8, 74, or 463 mg/kg

- bw/day, respectively, in females) (IRDC, 1985; cited in: EURAR, 2008). F 0 animals were exposed to test substance from 28 days prior to mating until sacrifice; F 1 animals
were t reated from weaning until sacrifice, with additional potential exposures occurring in utero and during lactation. All F 0 males were sacrificed after the mating period. Following the premating period, each male was cohabited with two females for 10 days. Fe males were examined for evidence of copulation by means of vaginal smears and/or the appearance of a vaginal plug. The day evidence of copulation was determined was designated as day 0 of gestation. Direct dosing began at 83 days of age for the F 0 parents and at 21 days of age for the F 1 weanlings. The F 0 and F 1 animals were observed for clinical signs of toxicity, changes in general appearance and behavior, and mortality. In the F 0 adults, body weights and food consumption were measured weekly; in addition , body weights were measured in F 0 females on GD 0, 7, 14, and 20; and on lactation days (LD) 0, 7, 14, and 21. Estrous cyclicity was determined in F 0 females prior to mating, during mating, and prior to dosing. All F 0 females were allowed to deliver. The day the entire litter was found and delivery was judged to be complete was designated as LD 0. Gestation duration was calculated. Following delivery, all pups were examined for external malformations and the numbers of live births and stillbirths (litter s ize) was recorded for each dam. Pups were weighed, sexed, and examined externally on LD 0, 7, 14, and 21. Litter size was determined on LD 0, 4, 10, and 21. The number of male and female pups was recorded on LD 4. Litters were examined daily for survival. F 0 females were examined for behaviors in nesting and nursing. On LD 21, all dams were sacrificed and a gross necropsy performed, including examination of the uterine contents for implantation sites; and ten F 1 weanlings/sex/dose were sacrificed and necrop sied. Five F 1 males and ten F 1 females/dose group were retained after LD 21 and sacrificed at 70 days (10 weeks) of age and necropsied. Due to high mortality in high

- dose F 1 pups, the surviving F 1 pups in the high

- dose group and an equal number of control pups were sacrificed on LD 6 and 7 and necropsied. Blood was collected via heart puncture and complete blood counts performed. Bone marrow smears were collected from the femur, and the abdominal contents of the pups with milk in the stomach were collected and frozen for future analyses.

Effects in the adults consisted of isolated reductions food consumption and body weight in the dams at 6250 ppm. Effects in the offspring consisted of significant reductions in pup survival at the high dose (none of the F 1 pups in the high

- dose group survived until lactation day 21); and slight (11 %, not statistically significant) decreases in pup survival, and labored breathing,106 subcutaneous hematoma, pale discoloration, blood around the orifices, pale liver, kidney, and spleen, and blood in the cranial cavity and brain beginning at the mid

- dose. No dose

- response, treatment

- related adverse effects were reported in the offspring in the low dose group. Reductions in body weight in F 1 male and female pups occurred during LD 7 , 14, and 21, but these reductions were not statistically significantly different from controls, and were seen only in the low

- and mid

- dose groups but not the high

- dose group. There were no dose

- related differences in F 1 post

- weaning appearance, behavior, food consumption, or clinical or anatomical pathology in the low

- and mid

- dose groups. Based on the results of this study, it was recommended that dosage levels in a two

- generation reproduction toxicity study not exceed 1000
ppm. The LOAEL for maternal to xicity was 6250 ppm (~463 mg/kg

- bw/day) based on reductions in body weight gains. The NOAEL for maternal toxicity was 1000 ppm (~74 mg/kg

bw/day). The LOAEL for developmental toxicity was 1000 ppm (~62/74 mg/kg

- bw/day) based on pup mortality due to hemorrh aging. The NOAEL for developmental toxicity was 100 ppm (~6/8 mg/kg

- bw/day). No effects on any of the reproductive parameters were reported. The NOAEL for reproductive toxicity was 6250 ppm (~384/463 mg/kg

- bw/day).

In an effort to determine the cause of hemorrhaging in the pups at the high dose from the reproductive toxicity range

- finding study, a screening level cross

- fostering developmental toxicity study was conducted in Charles River COBS Wistar rats fed diets containing either 0 or 6250 ppm (~ 3125 m g/kg

- bw/day) MCCPs (C 14

- 17 , 52 wt% Cl) for 4 weeks prior to mating and throughout pregnancy in a series of groups (Hart et al ., 1985; cited in: EURAR, 2008). Offspring from two of these groups (pups from control females reared from treated females, and pup s reared from their treated mothers) showed high

- pup mortality associated with internal hemorrhages. Hematological assays in the pups from these two groups showed decreases in factor X, resulting in a disruption of a vitamin K

- dependent clotting system (lo wer plasma vitamin K levels). It was concluded that the pup mortalities were due to internal hemorrhages caused by a decrease in the vitamin K

- dependent hemostatic mechanism (not examined in this study), induced during lactational exposures via the milk fr om mothers receiving MCCPs.

Additional studies have been conducted to investigate two hypotheses in an effort to clarify the possible causes of the hemorrhaging in the pups.

The first hypothesis proposes that MCCPs induce a catabolism of vitamin K in l actating rats leading to decreased plasma concentrations and ultimately low levels of vitamin K in the milk pups receive (vitamin K controls the formation of several clotting factors in the liver). In order to test this hypothesis, a preliminary study (CXR Biosciences Ltd., 2003; cited in: EURAR, 2008) was conducted in which three groups of 6 female adult Sprague

- Dawley were administered MCCPs (C 14

- 17 , 52 wt% Cl) via gavage at doses of 0, 500, or 1000 mg/kg

- bw/day for 21 days while being fed a normal diet or a vitamin K

- deficient diet. Following exposures to MCCPs, significant decreases in plasma concentrations of a clotting factor were seen in rats fed a normal diet; however, these decreases did not affect prothrombin clotting times. Reductions of a clotting factor in both treated and control groups were also seen in animals fed a vitamin

- K deficient diet. Plasma vitamin K levels were not affected by treatment in the normal diets, but they were lower in high

- dose animals fed vitamin K

- deficient diets. The results from this study suggested that MCCPs did not adversely affect the blood clotting system in adult female rats treated for 3 weeks up to a dose of 1000 mg/kg

- bw /day; and the hemorrhaging effects in pups are unlikely to be mediated by reduced vitamin K levels in breast milk.107 The second hypothesis proposes that MCCPs transferred to the pups through breast milk causes disruption of the pup clotting system. In order to test this hypothesis, a study (CXR Biosciences Ltd., 2004; cited in: EURAR, 2008) was conducted in two groups of 16 male and 32 female Sprague

- Dawley rats administered 0 or 6250 ppm (~ 0 and 513 and 538 mg/kg

- bw/day in males and females, respe ctively) MCCPs (C 14

- 17 , 52 wt% Cl) for 4 weeks prior to mating, during cohabitation, gestation, and lactation until study termination (at about 2 weeks after the first litters were born, due to high rate of pup mortality). Milk, blood, and liver samples fr om lactating dams, and blood and liver samples from lactating pups were assessed for plasma vitamin K levels. Five dams died or were killed at the time of parturition (16 % mortality). These deaths were considered to be treatment

- related as there was no in dication of obstruction or hindrance to delivery. The clinical necropsy of these dams showed effects suggestive of hemorrhaging in 3 out of the 5 dams and one male who died. Slight reductions in food consumption and body weight gains were observed during g estation and lactation. There were no effects on mating performance or duration of gestation. Concentrations of plasma vitamin K levels in adult females having gone through lactation and pregnancy was markedly decreased by treatment with MCCPs, which in tu rn produced a decrease in activity of the plasma clotting factors in treated dams. Prothrombin clotting times were not affected in the dams, suggesting that the functional reserve in these adult animals was sufficient. Pup plasma volumes were reportedly in sufficient to measure vitamin K directly, but clotting factor activities were possible to analyze. No effects on litter size at birth or on pup mortality from birth to LD 4 were reported; however, after pup mortality increased significantly after LD 4. The majority of these pups showed internal hemorrhages at necropsy. It was concluded that data from this study and the cross

fostering study performed by Hart et al . (1985) suggest that the fetus receives sufficient vitamin K via the placenta, but after birth becomes severely deficient in vitamin K and related clotting factors and relies on these factors via mother’s milk. In addition, the pups also receive considerable levels of MCCPs via lactation (through mother’s milk) which may also contribute to further reducing the vitamin K levels. These severe deficiencies in vitamin K levels and related clotting factors in the pups results in hemorrhaging. It was also concluded that the act of parturition places dams at higher risk.

More recently, a definitive one

- generation reproductive toxicity study was conducted to refine the NOAEL for effects in the offspring and to further explore the mechanisms of hemorrhaging (CXR, 2006; cited in: EURAR, 2008). This study was reportedly conduc ted in compliance with OECD TG 421 and Good Laboratory Practice standards. Four groups of 12

- 17 male and female Sprague

- Dawley rats were administered 0, 300, 600, and 1200 ppm (~ 0 and 21, 44, and 84 mg/kg

- bw/day in males; and 0, 23, 47, and 99 mg/kg

- bw/da y in females) MCCPs (C 14

- 17 , 52 wt% Cl) for 4 weeks prior to mating, during cohabitation, gestation, and lactation until study termination (for a total treatment of 11

- 12 weeks). Males were terminated on LD 4 (9 weeks of treatment) and females were allowed to litter and rear their offspring until PND 21. Females were sacrificed on LD 21. Adult males were assessed for signs of clinical toxicity, body weight, food consumption, and macropathology. Adult females were assessed for signs of clinical toxicity, body weight, food consumption, gestation length, parturition, liver weights, and macropathology. Mating performance and fertility were also evaluated. Offspring evaluations included clinical signs of toxicity, litter size, survival, sex ratio, body weight, and pathological examinations at necropsy. Milk, blood, and liver samples were obtained from selected offspring at specific time points between birth of litters and PND 21. In addition, blood, liver, and milk samples from a satellite group of five females and their litters from the control and high

- dose108 group (1200 ppm) were collected for future analysis. Analysis of these samples was still pending at the time of the UK assessment.

No adverse effects were reported in the adult animals for clinical condit ion, body weight, body weight gain, food consumption, estrous cycling, mating performance, pre

- coital interval, fertility, number of implantations, gestation lengths, or parturition. The only effect reported was for higher absolute and relative liver weigh ts in high

- dose females (1200 ppm; 99 mg/kg

- bw/day). Likewise, no adverse effects were in the offspring at any dose level for litter size, sex ratio, offspring survival, body weights, body weight gains, macropathology and liver weights. No adverse effects were reported on pre

- and post

- natal survival and growth up to sacrifice (weaning). Though no histopathology was performed, the body cavity and cranial cavity were opened and examined for any signs of hemorrhaging. None was reported. Based on the results o f this study, the LOAEL for maternal toxicity was 1200 ppm (~100 mg/kg

- bw/day) based on increases in liver weight; the NOAEL for maternal toxicity was 600 ppm (~ 47 mg/kg

- bw/day). The NOAEL for developmental and reproductive toxicity was 1200 ppm (~ 84/99 mg/kg

bw/day), the highest dose tested.



### C

- 2 LCCP Health Data Review There is no information on inhalation absorption of LCCPs in humans or in animals. Based on their low vapor pressure and water solubility, absorption following inhalation or dermal
exposure is expected to be limited. Some absorption and metabolism following oral exposure are possible for LCCPs with shorter carbon chain length and lower degree of chlorination.

No information is available on the toxicity of LCCPs in humans. Acute oral toxicit y data in animals show that LCCPs are of very low acute toxicity. Studies in animals have shown that some LCCPs may have the potential to cause slight skin irritation and sensitization but no eye irritation potential. The liver is the main target organ of LCCPs in repeated dose studies in experimental animals. Inflammatory and necrotic changes of the liver were observed in rats exposed to a C 20

- 30 LCCP with 43 wt% Cl at dose levels of 100 mg/kg

- bw and above. For another LCCP with C 20

- 26 70 wt% Cl, effects i n the liver occurred at a very high exposure level of 3,750 mg/kg

- bw/day; the NOAEL was 900 mg/kg

- bw/day.

An increased incidence of malignant lymphoma in male mice was reported at the highest dose of 5,000 mg/kg

- bw/day when tested using a C 23 LCCP with 4 3 wt% Cl in carcinogenicity studies in male and female rats and mice. However, malignant lymphoma is one of the more variable tumors in mice and has a viral origin in many cases. Data on the analogous short

- chain chlorinated paraffins (SCCPs) have shown no increase in the incidence of malignant lymphoma in a carcinogenicity study of SCCPs. LCCPs are non

- genotoxic and they are not expected to pose a carcinogenic hazard to humans.

No testing is needed for the PMN substances. Based on the LOAEL (100 mg/kg

- bw ) of the liver effects in female rats of repeated dose studies, Health Canada calculated a tole rable daily intake (TDI) of 71 μ g/kg

- bw/day. Using upper bounding intak e estimates ranging from 0.007 μ g/kg

bw/ day for 60+ age group to 0.024 μ g/kg

- bw/day for 0. 5 years age group, Environment Canada determined that the exposure levels are 10,000 and 3,000 times lower, respectively, than the TDI. Based on these evaluations, Health Canada concluded: “LCCP are not harmful to human health as defined in the Canadian en vironmental Protection Act. ”109 The National Research Council (NRC, 2000) reviewed the toxicological risks of selected flame retardant, including a LCCP containing C 24 with 70 wt% Cl. Based on the NOAEL of 900 mg/kg

- bw/day (liver toxicity), the NRC derived an RfD of 0.3 mg/kg

- bw/day. Using this RfD and the worst case average daily exposure to be 0.16 mg/kg/day, NRC concluded: “ LCCP do not pose a noncancer risk when incorporated into residential furniture at the estimated application levels .” Further, it was concluded that: “ LCCP are not likely to be a human carcinogen and derivation of a cancer potency factor is unnecessary. ”



### C

**- 2** 

**- 1**  **Metabolism** There is no information on inhalation absorption of LCCPs in humans or in animals. Based on their low vapor pressure and low water solubility, absorption following inhalation or dermal exposure is expected to be limited. Oral (gavage) studies (IRDC, 1981; cited in: UK, 2009) showed that approximately 82

- 95 % of a single dose of [8

- 14 C]

- labeled C 22

- 26 chlorinated paraffin (43 % Cl) was recovered in the feces in rats during the seven

- day collection period. Only 0.1

- 0.8 % of the radiolabel was excreted in the urine. For C 22

- 26 chlorinated paraffin (70 wt% Cl), it was found that 61

- 88 % of the administrated radioactivity was reco vered in the feces in rats during the seven

- day collection period. Less than 0.1

- 1 % was excreted in the urine.



### C

**- 2** 

**- 2**  **Acute Toxicity** Acute toxicity studies have been conducted in rats, mice or dogs on five LCCPs: C 20

- 30 , 41

- 50 wt% Cl; C 22

- 26 , 42 wt% Cl, C 23, 4 3 wt% Cl; C 20

- 30, 61

- 70 wt% Cl; C 24, 70 % Cl. The maximum dose levels used in these studies ranged from 4

- 50 g/kg

- bw. No deaths were reported in any of the studies (IUCLID, 2003; cited in: UK, 2009).



### C

**- 2** 

**- 3**  **Irritation and Sensitization** Skin irritation testing has been conducted on four LCCPs: C 19 , 44 wt% Cl; C 22

- 26 , 42 wt% Cl; C 20

- 30, 41

- 50 wt% Cl; C 20

- 30, 70 wt% Cl. No evidence of irritation was seen in three of the four LCCPs. For the C 22

- 26 , 42 wt% Cl product, erythema was observed in two of six animals test ed; the severity threshold was below the classification of the EU system (IUCLID, 2003; cited in: UK, 2009).

Evidence of slight eye irritation was seen in a test of a C 22

- 26 , 42 wt% Cl product. However, the criteria for classification as an eye irritant w ere not met (IUCLID, 2003; cited in: UK, 2009).

In a maximization test and a Buehler test using guinea pig, a C 22

- 26 , 42 wt% Cl product was negative (IUCLID, 2003; Bailey and Sheldon, 1998; cited in: UK, 2009). A C 18

- 27 , 40 wt% Cl product was reported to elicit a positive response in the guinea pig maximization test (IUCLID, 2000; cited in: UK, 2009). However, no information is available on the quality of this study.



### C

**- 2** 

**- 4**  **Repeated**

**- dose Toxicity** 
LCCPs with C 20

- 30, 43 wt% Cl were dissolved in corn oil and given by gavage at 100, 900 or 3,750 mg/kg

- bw/day to groups of 15 male and female Fisher 344 rats in a 14

- and a 90

- day studies (IRDC1981, 1984; cited in: Serront et al ., 1987). There was a treatment

- related effect on the liver of female rats at all dose levels, but no liver effects were seen in the males. Female liver weights were increased and a multifocal granulomatous hepatitis characterized by inflammatory changes a nd necrosis. Nephrosis was observed in the kidney of male rats and mineralization in110 the kidneys of female rats at 3,750 mg/kg

- bw/day. Similar liver effects were observed in the high

- dose (3,750 mg/kg

- bw/day) rats of both sexes in a 90

- day study on LCCPs w ith C 22

- 26, 70 wt% Cl (IRDC1981; cited in: Serront et al ., 1987) and in a 90

- day study as well as a 2

- year bioassay on a LCCP with an average of C 23, 43 wt% Cl at 100 mg/kg

- bw/day (NTP, 1986). Based on the liver toxicity, a LOAEL of 100 mg/kg

- bw/day is est ablished for the LCCPs with 43 wt% Cl and a NOAEL of 900 mg/kg

- bw/day is identified for the LCCPs with 70 wt% Cl.



### C

**- 2** 

**- 5**  **Genotoxicity**
Both LCCPs of C 22

- 26 , 43 wt% Cl and C 23, 43 wt% Cl were negative in several Salmonella strains of the Ames test with or without metabolic activation (IUCLID, 2003; NTP, 1986; cited in: UK, 2009). LCCPs with C 20

- 30 with 43 wt% Cl or C 22

- 26 with 70 wt% Cl did not induce significant increases of chromosomal or chromatid aberrations in bone marrow cells of rats (IRDC1983; cited in : Se rrone et al ., 1987).



### C

**- 2** 

**- 6**  **Carcinogenicity** The carcinogenicity of an LCCP (C 23, 43 wt% Cl) was studied by administering the chemical in corn oil by gavage to groups of 50 F344/N rats and 50 B6C3F1 mice of each sex, 5 days per week for 103 weeks. Male rats rece ived doses of 0, 1,875 or 3,750 mg/kg

- bw/day; female rats received 0, 2,500 or 5,000 mg/kg

- bw/day. An increased incidence of malignant lymphomas was reported for male mice given the LCCP; the incidences for the controls, low

- and high dosed groups are 6/50 , 12/50, and 16/50, respectively. However, mice are prone to the development of lymphomas with a range of tumor incidence of 2

- 73, and viruses may be the causative agent of lymphomas in laboratory strains of mice. It has been concluded that LCCPs of C 20

- 30 with 43 wt% Cl did not produce clear evidence of carcinogenicity when tested in rats and mice (Serrone et al ., 1987).



### C

**- 2** 

**- 7**  **Developmental Reproductive Toxicity Review** In a prenatal developmental toxicity study in rats, no treatment

- related effects were reporte d. The NOAEL for both maternal and developmental toxicity was 5000 mg/kg

- bw/day, the highest dose tested.

In a prenatal developmental toxicity study in rabbits, the LOAEL for maternal toxicity was 500 mg/kg

- bw/day (the lowest dose tested) based on incre ased incidence of clinical signs. The NOAEL for developmental toxicity was > 2000 mg/kg

- bw/day, the highest dose tested.

In a prenatal developmental toxicity study in rabbits, the NOAEL for both maternal and developmental toxicity was 1000 mg/kg

- bw/day, the highest dose tested.

Basis for Conclusions In a range

- finding study (Study # 438

- 033, October 27, 1981; Chlorinated Paraffin Consortium), in pregnant Charles River COBS rats administered LCCP (22

- 26 carbons, 43 wt% Cl) via gavage at dose levels of 0, 3000, and 5000 mg/kg

- bw/day on GD 6

- 19, the only effects reported occurred in dams and consisted of a slight increased incidence in anogenital matting during the latter portion of the treatment period at 5000 mg/kg

- bw/day. No a dverse treatment

- related effects were reported in offspring at doses up to 5000 mg/kg

- bw/day.111 Based on the findings of the range

- finding study, four groups of 25 pregnant Charles River COBS CD rats were administered LCCP (22

- 30 carbons, 70 wt% Cl) via ga vage at doses of 0, 500, 2000, and 5000 mg/kg

- bw/day on GD 6

- 19 (Study # 438

- 045, April 11, 1984; Chlorinated Paraffin Consortium). Unmated males and females were individually housed and acclimated for 15

- days in an environmentally controlled room. At the end of the acclimation period, all animals were weighed and subjected to a detailed physical examination. One female and one male rat were placed together for mating. Confirmation of mating was based on evidence of a copulatory plug. The day mating was con firmed was designated as day 0 of gestation. Test article was administered to pregnant females orally by gavage as a single daily dose on GD 6

- 19. During treatment, pregnant females were observed daily for mortality and clinical signs of toxicity. Any fema les not surviving to scheduled sacrifice were necropsied. Body weights were recorded on GD 0, 6, 9, 12, 16, and 20. All females were sacrificed on GD 20 and the uterus and ovaries excised for examination. The number and location of viable and nonviable fet uses, early and late resorptions, and the number of total implantations and corpora lutea were recorded. The uterus was weighed. The abdominal and thoracic cavities underwent gross examination. Maternal tissues were preserved for future histopathological a nalysis. Fetuses were weighed, sexed, tagged and examined for external malformations and variations, including the palate and the eyes. The fetuses underwent visceral and skeletal examinations for malformations and developmental variations.

No adverse treatment

- related effects were reported in the dams or offspring at doses up to 5000 mg/kg

- bw/day. The NOAEL for both maternal and developmental toxicity was 5000 mg/kg

bw/day, the highest dose tested.

In a range

- finding study (Study # 438

- 018, October 27, 1981; Chlorinated Paraffin Consortium) in pregnant Dutch Belted rabbits administered LCCP (22

- 26 carbons, 43 wt% Cl) via gavage at doses of 0, 500, 1000, 2000, 3000, and 5000 mg/kg

- bw/day on GD 6

- 27, a slight decrease in the amount of feces and a slig ht increase in matting and/or staining of the haircoat was reported in the dams at > 3000 mg/kg

- bw/day. No other effects were reported. Observations of offspring did not appear to be included.

Based on the results of the range

- finding study, four groups of 16 Dutch Belted pregnant rabbits were administered LCCP (22

- 26 carbons, 43 wt% Cl) via gavage at doses of 0, 500, 2000, and 5000 mg/kg

- bw/day on GD 6

- 27 (Study # 438

- 030, August 26, 1982; Chlo rinated Paraffin Consortium). Unmated males and females were individually housed and acclimated for 9

- weeks in an environmentally controlled room. As a result of a positive finding for parasites in stool samples collected during acclimation, all rabbits re ceived sodium sulfamethazine in their drinking water for 19 days during the acclimation period. This treatment was terminated 5 weeks prior to study initiation and only rabbits testing negative for parasites were placed on study. At the end of the acclimat ion period, all animals were weighed and subjected to a detailed examination. Females were impregnated via artificial insemination. Immediately after insemination, ovulation was induced via an injection of chorionic gonadotropin in a marginal ear vein. Sem en was collected from males of proven fertility and evaluated for motility. The day of artificial insemination was designated as day 0 of gestation. During treatment, pregnant females were observed for mortality and clinical signs of toxicity. Body weights were recorded on GD 0, 6, 12, 18, 24, and 28. Any females not surviving to scheduled sacrifice were necropsied. On GD 28, all surviving females were sacrificed and the uterus and ovaries excised for examination. The location and number of viable and nonvi able fetuses, early and late resorptions, and the number112 of total implantations and corpora lutea were recorded. The uterus was weighed. The thoracic and abdominal cavities underwent gross examination. Pooled samples of abdominal adipose tissue from 3 dams were frozen for future analysis. Each fetus was sexed, weighed, and examined for external malformations and variations, including the palate and the eyes, as well as visceral and skeletal examinations for malformations and developmental variations, includ ing examination of the brain and the heart.

A dose

- related trend in an increased incidence in soft stool and/or anogenital matting or staining was observed in the dams beginning at the low

- dose group. Three dams aborted and were sacrificed during treatmen t; one in the mid

- dose group, and two in the high

- dose group. Increases in postimplantation loss and corresponding decreases in viable fetuses, and increases in late resorptions were reported at the high

- dose group, however, these effects were not reported as being statistically significant. No signs of treatment

- related developmental toxicity were reported in the offspring at < 5000 mg/kg

- bw/day, although the sample size in the high

- dose group was limited, precluding any definitive conclusions. Therefore, the LOAEL for maternal toxicity was 500 mg/kg

- bw/day (the lowest dose tested), based on increased incidence of clinical signs. The NOAEL for developmental toxicity was > 2000 mg/kg

- bw/day.

In a range

- finding study (Study # 438

- 038, November 1, 1982; Chlorinated Paraffin Consortium) in pregnant Dutch Belted rabbits administered LCCP (22

- 30 carbons, 70 wt% Cl) via gavage at doses of 0, 2000, 3750, and 5000 mg/kg

- bw/day on GD 6

- 27, increases in abortions, reductions in maternal body weight, and increase s in post

- implantation losses were reported at > 2000 mg/kg

- bw/day. Therefore, a second range

- finding study was conducted at dose levels of 0, 50, 200, and 1000 mg/kg

- bw/day (Study # 438

- 040, November 4, 1982; Chlorinated Paraffin Consortium). A slight dec rease in viable fetuses and a slight increase in postimplantation loss were reported at 1000 mg/kg

- bw/day. No other effects were reported.

Based on the results of the range

- finding studies, four groups of 16 Dutch Belted pregnant rabbits were administere d LCCP (22

- 30 carbons, 70 wt% Cl) via gavage at doses of 0, 100, 300, and 1000 mg/kg

- bw/day on GD 6

- 27 (Study # 438

- 039, July 18, 1983; Chlorinated Paraffin Consortium). Unmated males and females were individually housed and acclimated for 40

- days in an en vironmentally controlled room. At the end of the acclimation period, all animals were weighed and subjected to a detailed examination. Females were impregnated via artificial insemination. Immediately after insemination, ovulation was induced via an injection of chorionic gonadotropin in a marginal ear vein. Semen was collected from males of proven fertility and evaluated for motility. The day of artificial insemination was designated as day 0 of gestation. During treatment, pregnant females were observed for mortality and clinical signs of toxicity. Body weights were recorded on GD 0, 6, 12, 18, 24, and 28. Any females not surviving to scheduled sacrifice were necropsied. On GD 28, all surviving females were sacrificed and the uterus and ovaries e xcised for examination. The location and number of viable and nonviable fetuses, early and late resorptions, and the number of total implantations and corpora lutea were recorded. The uterus was weighed. The thoracic and abdominal cavities underwent gross examination. Pooled samples of abdominal adipose tissue from 3 dams were frozen for future analysis. Each fetus was sexed, weighed, and examined for external malformations and variations, including the palate and the eyes, as well as visceral and skeletal examinations for malformations and developmental variations, including examination of the brain and the heart.113 No treatment

- related adverse effects were observed in the dams or the offspring at any dose level. The increases in postimplantation losses noted in the two previous range

- finding studies reported above were not reproduced in the definitive study. Therefore, the NOAEL for both maternal and developmental toxicity was 1000 mg/kg

- bw/day, the highes t dose tested.114 

## Appendix D ENVIRONMENTAL MONITORING DATA

**D**

**- 1**  **MCCP MONITORING DATA D** 

**- 1** 

**- 1**  **Surface Water** It is known that over time, based on their molecular weight and physicochemical properties, MCCPs in surface water will partition to suspended particulates, sediment, sludge , or soil. Reported MCCP concentrations in surface water range from < 2.50 × 10

- 10 mg/L to 1.49 × 10

- 3 (Table D

- 1). Very little information is available on the specific sampling locations for many of the surface water measurements reported in Table D

- 1. Li mited documentation is available on two of the studies (Petersen et al., 2006 and Muir, 2003). Two sources provide a review of the literature with very little details (IPCS, 1996 and EC, 2008b). Two studies do provide detailed information on the sampling a pproach, including location (Houde et al., 2008 and USEPA, 1988). The Petersen et al. (2006) study, which had the highest published concentration, reported results for water samples collected from different Norwegian locations. EPA/OPPT assumes that these samples were collected in non

- marine waters. Three studies found were not used in this assessment (BUA, 1992; Hoechst, 1987; and Willis, 1994). However, all of the studies used in the assessment use modern analytical techniques, reference the specific CPs of interest, and provide, at a minimum, general information on the sampling location. Given the paucity of surface water data available, EPA/OPPT used measurements from the selected studies and used the minimum and maximum values in this assessment.

Measu rements of dissolved (filtered) concentrations were generally ND with few exceptions. Concentrations measured in surface water were largely from studies that measured total water concentrations which included MCCPs sorbed to particulates. More recent monit oring studies ( Tabl e D

- 1) have focused on measuring MCCPs in suspended solids, sediment pore water, and sediment.

Early analytical methods using thin layer chromatography (TLC) were used to measure CPs in surface water. However, this method has poor sensitivity and reproducibility, and provide false negative results. Current methods of quantification using gas or liquid chromatography coupled with a ra nge of detectors ( i.e. , mass spectrometry; MS) are more reliable. Nearly all of the water concentrations were measurements taken at a single point in time ( i.e. , the samples were not time series samples). Absent more extensive monitoring data, EPA/OPPT ass umed that the available data could be extrapolated to longer time periods for determination of a chronic exposure concentration.

MCCP concentrations in surface water, reported in Tabl e D

- 1, rely on test methods that filtered
or pre

- filtered samples before they were analyzed, which can underestimate environmental concentrations. Where appropriate, reported values were converted to a common unit, as
pres ented in the table. For the purposes of this assessment, in the studies considered acceptable, EPA/OPPT used the lowest and highest reported concentrations (< 2.50 × 10

- 10 mg/L to 1.49 × 10

- 3 ) to evaluate risks of potential concern to aquatic organisms.115 Tabl eApx_D

- 1

- 1 : Surface Water C once ntrations of MCCPs, Sorted by C ountry.

Media Country Location City, State or

Province Comments Converted Concentration Common Units Analytical Method References

Surface water Canada Lake Ontario Maximum 2.60×10

- 9 mg/L NR EC (2008b)

< 2.50×10

- 10 mg/L GC

- HRMS

- MAB Houde et al. (2008)

< 1.00×10

- 8 mg/L GC

- ECNI

- MS Muir et al. (2003)

Maximum 4.70×10

- 8 mg/L GC

- HRMS

- MAB Houde et al. (2008)

Mean 9.00×10

- 10 mg/L GC

- HRMS

- MAB Houde et al. (2008)

Germany River Lech at Langsweid --- 1.90×10

- 4 mg/L NR IPCS (1996)

River Lech at Rain

--- 1.70×10

- 4 mg/L NR IPCS (1996)

River L ech at

Gersthofen --- 9.00×10

- 5 mg/L NR IPCS (1996)

River Lech at

Augsburg < 2.50×10

- 5 mg/L NR IPCS (1996)

River Danube at Marxheim --- 7.00×10

- 5 mg/L NR IPCS (1996)

< 3.00×10

- 5 mg/L NR IPCS (1996) Norway NR

--- 1.49×10

- 3 mg/L GC

- ECNI

- MS Petersen et al. (2006)

United Kingdom Multiple locations < 1×10

- 4 mg/L GC

- ECNI

- MS Nicholls et al. (2001)

United States Sugar Creek, Ohio < 7.50×10

- 5 mg/L GC

- ECNI

- MS USEPA (1988) Central European Country NR < 5.00×10

- 5 mg/L GC

- ECNI

- MS Coelhan (2010)
NR: Not recorded. Location description was not provided in the study.

-- : Single sample value reported above the detection limit; therefore, no data qualifier required. GC

- HRMS

- MAB: Gas chromatography

- high resolution mass spectrometry with metastable atom bombardment ionization GC

- ECNI

- MS: Gas chromatography in combination with electron capture negative ion mass spectrometry N otes: 1. All values provided in the table above represent total MCCP and not individual MCCP isomers 2. In some cases, the minimum values in the table are preceded by “<”. This indicates that the value reported in article was rep orted as a non

- detect . In such cases, one half of the lowest reported detection limit was compiled as the ‘minimum’ reported monitoring data 3. All concentrations measured from impoundment lagoons and drainage ditches from the USEPA (1988) study have not been included as the y are not considered as surface water concentrations 4. All concentrations measured from suspended solid matter fraction from influents from the Coelhan (2010) study have not been i ncluded as they are not considered as surface
water concentr ations116 

### D

**- 1** 

**- 2**  **Sediment**
MCCP sediment concentrations from marine and non

- marine environments ranged from 5.00 × 10

- 3 to 1.64 × 10 1 mg/kg dw and from 2.00 × 10

- 3 to 6.51 × 10 1 mg/kg dw, respectively.

For the purposes of this assessment, in the studies considered acceptable, EPA/OPPT used the lowest and highest reported marine and non

- marine sediment concentrations (5.00 × 10

- 3 to 1.64 × 10 1 mg/kg dw and 2.00 × 10

- 3 to 6.51 × 10 1 , respectively) to evalu ate risks of potential concern to sediment organisms (Table D

- 2). Where appropriate, reported values were converted to a common unit, as presented in the table.117 Table Apx_ D

- 1

- 2 : Sediment C oncentrations of MCCPs, Sorted by C ountry.

Media Country Location C omments Converted Concentration Common Units References City, State or Province

Sediment (Marine) Australia NR --- 1.11 mg/kg dw Kemmlein et al. (2002)

--- 1.17 mg/kg dw Kemmlein et al. (2002)

--- 3.11 mg/kg dw Kemmlein et al. (2002)

--- 1.64×10 1 mg/kg dw Kemmlein et al. (2002)

Canada Ha milton Harbour (Windemere basin) --- 2.90×10

- 1 mg/kg* Muir et al. (2000)

Germany German Bight, North Sea --- 5.00×10

- 3 mg/kg dw Hüttig et al. (2004)

--- 9.00×10

- 3 mg/kg dw Hüttig et al. (2004)

--- 9.00×10

- 3 mg/kg dw Hüttig et al. (2004)

--- 1.30×10

- 2 mg/kg dw Hüttig et al. (2004)

--- 2.80×10

- 2 mg/kg dw Hüttig et al. (2004)

--- 1.46×10

- 1 mg/kg dw Hüttig et al. (2004)

Baltic Sea --- 9.30×10

- 2 mg/kg dw Hüttig et al. (2004)

--- 1.15×10

- 1 mg/kg dw Hüttig et al. (2004)

--- 1.22×10

- 1 mg/kg dw Hüttig et al. (2004)

--- 2.11×10

- 1 mg/kg dw Hüttig et al. (2004)

--- 4.99×10

- 1 mg/kg dw Hüttig et al. (2004)118 Media Country Location C omments Converted Concentration Common Units References City, State or Province

North and Baltic Sea --- 2.20×10

- 2 mg/kg dw Hüt tig and Oehme (2006)

--- 2.30×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 3.30×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 3.40×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 3.70×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 3.90×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 4.30×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 4.30×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 4.80×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 5.40×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 5.80×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 6.10×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 7.20×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 7.60×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 7.70×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 8.10×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 8.50×10

- 2 mg/kg dw Hüttig and Oehme (2006)

--- 8.70×10

- 2 mg/kg dw Hüttig and Oehme (2006)119 Media Country Location C omments Converted Concentration Common Units References City, State or Province

--- 1.49×10

- 1 mg/kg dw Hüttig and Oehme (2006)

--- 1.49×10

- 1 mg/kg dw Hüttig and Oehme (2006)

--- 1.49×10

- 1 mg/kg dw Hüttig and Oehme (2006)

--- 2.75×10

- 1 mg/kg dw Hüttig and Oehme (2006)

--- 9.10×10

- 2 mg/kg dw Hüttig and Oehme (2005)

--- 4.80×10

- 2 mg/kg dw Hüttig and Oehme (2005)

--- 1.98×10

- 1 mg/kg dw Hüttig and Oehme (2005)

--- 1.31×10

- 1 mg/kg dw Hüttig and Oehme (2005)

--- 1.32×10

- 1 mg/kg dw Hüttig and Oehme (2005)

--- 3.03×10

- 1 mg/kg dw Hüttig and Oehme (2005)

--- 1.53×10

- 1 mg/kg dw Hüttig and Oehme (2005)

--- 1.14×10

- 1 mg/kg dw Hüttig and Oehme (2005)

--- 4.00×10

- 2 mg/kg dw Hüttig and Oehme (2005)

--- 2.70×10

- 2 mg/kg dw Hüttig and Oehme (2005)

--- 1.80×10

- 2 mg/kg dw Hüttig and Oehme (2005)

--- 1.90×10

- 2 mg/kg dw Hüttig and Oehme (2005)

--- 3.00×10

- 2 mg/kg dw Hüttig and Oehme (2005)

--- 3.20×10

- 2 mg/kg dw Hüttig and Oehme (2005)120 Media Country Location C omments Converted Concentration Common Units References City, State or Province

--- 1.80×10

- 2 mg/kg dw Hüttig and Oehme (2005)

--- 2.40×10

- 2 mg/kg dw Hüttig and Oehme (2005)

Sediment

(Non

- marine) Canada Lake Erie

--- 6.80×10

- 2 mg/kg dw Tomy and Stern (1999)

Lake St Francis Minimum 7.50×10

- 1 mg/kg dw EC (2008b)
Maximum 1.2 mg/kg dw EC (2008b)

Czech Republic NR Minimum 2.00×10

- 3 mg/kg* Pribylova et al. (2006)

Labe Sum 1.80×10

- 2 mg/kg dw Pribylova et al. (2006)

Sum 7.30×10

- 2 mg/kg dw Pribylova et al. (2006)

Libis

- Labe Sum 1.6 mg/kg dw Pribylova et al. (2006)

Bilina Sum 3.10×10

- 2 mg/kg dw Pribylova et al. (2006)

Mala Becva Sum 1.13×10

- 1 mg/kg dw Pribylova et al. (2006)

Becva Sum 1.20×10

- 1 mg/kg dw Pribylova et al. (2006)

Morava Sum 1.93×10

- 1 mg/kg dw Pribylova et al. (2006)

Ohre Sum 3.08×10

- 1 mg/kg dw Pribylova et al. (2006)

Sum 6.00×10

- 1 mg/kg dw Prib ylova et al. (2006)

Sum 5.58 mg/kg dw Pribylov a et al. (2006)

Morava Sum 4.16×10

- 1 mg/kg dw Pribylova et al. (2006) Dyje Sum 7.57×10

- 1 mg/kg dw Pribylova et al. (2006)121 Media Country Location C omments Converted Concentration Common Units References City, State or Province

Drevnice Sum 8.93×10

- 1 mg/kg dw Pribylova et al. (2006)

Germany Bodensee (middle) < 5.00×10

- 3 mg/kg dw IPCS (1996)

--- 7.00×10

- 2 mg/kg dw IPCS (1 996)

River Lech < 5.00×10

- 3 mg/kg dw IPCS (1996)

--- 3.25×10

- 1 mg/kg dw IPCS (1996)

Maximum 7.00×10

- 1 mg/kg* Tomy et al. (1998)

River Rhine --- 6.00×10

- 2 mg/kg dw IPCS (1996)

--- 8.50×10

- 2 mg/kg dw IPCS (1996)

--- 1.40×10

- 1 mg/kg dw IPCS (1996)

Minimum 1.45×10

- 1 mg/kg dw IPCS (1996)

Ma ximum 2.05×10

- 1 mg/kg dw IPCS (1996)

River Elbe at Hamburg Minimum 1.30×10

- 1 mg/kg dw IPCS (1996)

Maximum 2.30×10

- 1 mg/kg dw IPCS (1996)

River Main Minimum 1.60×10

- 1 mg/kg dw IPCS (1996)

Maximum 2.60×10

- 1 mg/kg dw IPCS (1996) Outer Alster, Hamburg

--- 3.70×10

- 1 mg/kg dw IPCS (1996)

Norway NR minimum 5.00×10

- 2 mg/kg dw Petersen et al. (2006)

maximum 3.24 mg/kg dw Petersen et al. (2006)

--- 2.7 mg/kg ww Borgen et al. (2003)

--- 1.14×10 1 mg/kg ww Borgen et al. (2003)

South China Pearl River Delta Minimum 8.80×10

- 1 mg/kg dw Chen et al. (2011)
Minimum 1.1 mg/kg dw Chen et al. (2011 )

Minimum 1.4 mg/kg dw Chen et al. (2011)

Maximum 1.4 mg/kg dw Chen et al. (2011)

Maximu m 3.8 mg/kg dw Chen et al. (2011)122 Media Country Location C omments Converted Concentration Common Units References City, State or Province

Mean 3.9 mg/kg dw Chen et al. (2011)

Mean 2.10×10 1 mg/kg dw Chen et al. (2011)

Maximum 3.80×10 1 mg/kg dw Chen et al. (2011)

Switzerland Lake Thun Minimum 5. 00×10

- 3 mg/kg dw Iozza et al. (2008)
Maximum 2.60×10

- 2 mg/kg dw Iozza et al. (2008)
Lake Zurich Maximum 5.00×10

- 3 mg/kg* Tomy et al. (1998)

United Kingdom NR < 1.00×10

- 1 mg/kg dw Nicholls et al. (2001)

South West Region: Grand Union Canal --- 3.00×10

- 1 mg/kg dw Nicholls et al. (2001)

--- 2.7 mg/kg dw Nicholls et al. (2001)

--- 2.8 mg/kg dw Nicholls et al. (2001 )

South West Region; Bristol Avon River --- 5.00×10

- 1 mg/kg dw Nicholls et al. (2001)

--- 6.00×10

- 1 mg/kg dw Nicholls et al. (2001)

--- 8.00×10

- 1 mg/kg dw Nicholls et al. (2001)

North East Region: Hull River --- 1.0 mg/kg dw Nicholls et al. (2001)

--- 1.35×10 1 mg/kg dw Nicholls et al. (2001)

--- 1.1 mg/kg dw Nicholls et al. (2001)

South West Region: Colne River --- 1.4 mg/kg dw Nicholls et al. (2001)

--- 2.0 mg/kg dw Nicholls et al. (2001)

West Midlands Region: Trent River --- 3.8 mg/kg dw Nicholls et al. (2001)

--- 6.02×10 1 mg/kg dw Nicholls et al. (2001)123 Media Country Location C omments Converted Concentration Common Units References City, State or Province

--- 6.51×10 1 mg/kg dw Nicholls et al. (2001)

North West Region:

Hornsmill brook --- 5.6 mg/kg dw Nicholls et al. (2001)

--- 1.25×10 1 mg/kg dw Nicholls et al. (2001)

--- 1.83×10 1 mg/kg dw Nicholls et al. (2001)

North East Region: Hull River --- 1.0 mg/kg dw Nicholls et al. (2001)

--- 1.1 mg/kg dw Nicholls et al. (2001)

--- 1.35×10 1 mg/kg dw Nicholls et al. (2001)

East Midlands Region: Idle River --- 1.62×10 1 mg/kg dw Nicholls et al. (2001)

--- 4.39×10 1 mg/kg dw Nicholls et al. (2001)

Northumberland Region: Skerne River --- 1.80×10 1 mg/kg dw Nicholls et al. (2001)

--- 2.56×10 1 mg/kg dw Nicholls et al. (2001)

--- 5.84×10 1 mg/kg dw Nicholls et al. (2001)

East Anglia Region: Lark River --- 3.22×10 1 mg/kg dw Nicholls et al. (2001)

--- 4.50×10 1 mg/kg dw Nicholls et al. (2001)

--- 6.04×10 1 mg/kg dw Nicholls et al. (2001)

United States Detroit River

--- 6.80×10

- 2 mg/kg dw Tomy et al. (1999)

Sugar Creek, Ohio Reported as trace with range of 1.5

- 5; used the average 3.25×10

- 3 mg/kg dw USEPA (1988)124 Media Country Location C omments Converted Concentration Common Units References City, State or Province

Reported as trace with range of 1 .5

- 5; used the average 3.25×10

- 3 mg/kg dw USEPA (1988)

--- 6.80×10

- 3 mg/kg dw USEPA (1988)

--- 8.20×10

- 3 mg/kg dw USEPA (1988)

--- 7.60×10

- 1 mg/kg dw USEPA (1988)

--- 2.10 ×10 1 mg/kg dw USEPA (1988)

--- 3.40×10 1 mg/kg dw USEPA (1988)

--- 5.00×10 1 mg/kg dw USEPA (1988)
Note: NR: Not recorded. Location description was not provided in the study.

-- : Single sample value reported above the detection limit; therefore, no data qualifier required. 1. All values provided in the table above represent total MCCP and not individual MCCP isomers 2. In some cases, the minimum values in the table are preceded by “< ”. This indicates that the value reported in article was reported as a non

- detect. In such cases, one half of the lowest reported detection limit was compiled as the ‘minimum’ reported monitoring data 3. dw.

– dry weight and ww.

– wet weigh t125 

### D

**- 1** 

**- 3**  **Biosolids and Soil** CPs are detected more frequently and at higher concentrations in treated sewage sludge ( i.e., biosolids) than in soil. MCCP concentrations ranged from 5.00 × 10

- 5 to 9.70 × 10 3 mg/kg dw in sludge and from 1.5 × 10

- 2 to 8.5 × 10

- 2 mg/k g dw in soil. It is unclear if the difference in MCCP concentrations in sludge and soil is related to the smaller sample sizes for these media compared to the typically larger data sets available for water and sediment. To determine the most reliable studi es for its consideration, EPA used the following criteria: designation of specific MCCP chain length and the appropriate analytical methodology. Thus, EPA/OPPT did not use information from other published studies reporting measured CPs in sludge and soil b ecause they did not distinguish the CPs measured (e.g., Nicholls et al., 2001) ; although they reported total CP concentrations at much lower levels ranging from 3.00 × 10

- 5 to 2.3 mg/kg dw.

.
Stevens et al. (2003) measured MCCP concentrations in sludge samples obtained from 14 WWTPs in the UK. MCCP concentrations ranged from 3.00 × 10 1 to 9.70 × 10 3 mg/kg dw. The authors concluded that these ver y high concentrations were likely the result of releases from numerous and ongoing diffuse

sources.

Although risk to terrestrial species was not calculated, EPA/OPPT notes that the lowest and highest reported biosolid and soil concentrations (5.00 × 10

- 5 to 9.70 × 10 3 mg/kg dw and 1.5 × 10

- 2 to 8.5 × 10

- 2 mg/kg dw, respectively) represents a very large range (up to eight orders of magnitude (Table D

- 3).

Table Apx_ D

- 1

- 3: Biosolid an d Soil C oncentrations of MCCPs

Location Media Concentration Units Referen ces Minimum Maximum

Switzerland Soil 1.5 × 10

- 2 8.5 × 10

- 2 mg/kg dw Iozza (2010) China Soil 2.1 × 10

- 6 1.53 × 10

- 3 mg/kg dw Wang et al. (2013) Czech Republic Sewage Sludge 7.36 × 10

- 1 2.30 mg/kg dw Pribylova et al. (2006) United Kingdom Sewage Sludge 3.00 × 10 1 9.70 × 10 3 mg/kg dw Stevens et al. (2003)

United States Sewage Sludge 5.00 × 10

- 5 5.00 × 10

- 5 mg/kg dw Prib ylova et al. (2006)



### D

**- 1** 

**- 4**  **Biota**
EPA/OPPT reviewed available published literature and summarized MCCP concentrations in tissues of aquatic and terrestrial biota (Table D

- 4). Measured tissue concentrations for aquatic biota ranged from ND to 2.63 mg/kg ww ( i.e. , beluga whales, seals, rainbow trout, carp, mackerel, arctic char, mussels, crustaceans, and plankton) and ranged from 5.00 × 10

- 3 to 3.70 × 10

- 1 mg/kg ww in terrestrial biota. The concentrations measured in the terrestrial studies did not designate t he specific CP congener groups measured.

As a result of EPA/OPPT’s evaluation, MCCPs were found in organisms across many different trophic levels indicating widespread environmental contamination (Table D

- 4). The data were insufficient for EPA/OPPT to dr aw conclusions about trends based on region, species, time, or other factors.126 While EPA/OPPT determined the concentrations of MCCPs in aquatic and terrestrial biota range from ND to 2.63 mg/kg ww and from 5.00 × 10

- 3 to 3.70 × 10

- 1 mg/kg ww, respectively, in this assessment, EPA/OPPT did not use tissue concentrations to determine risks of potential concern for biota. Rather, it used the risk quotient (RQ) method and the conventional PMN approach as described in Section 6127 Table_Apx D

- 1

- 4: MCCP Biota C oncentrations of MCCPs

Location Media Description Minimum Units Min Reference Maximum Units Max Reference

Aquatic Biota

Australia Invertebrates 2.32×10

- 5 mg/kg lw Kemmlein et al. (2002) 3.05×10

- 5 mg/kg lw K emmlein et al. (2002)

Canada Mammals 5.45×10

- 7 mg/kg ww Bennie et al. (2000) 8.00×10

- 5 mg/kg ww Bennie et al. (2000)

Fish 2.57×10

- 7 mg/kg ww Bennie et al. (2000) 2.63 mg/kg ww Muir et al. (2000)

Invertebrates ND 1 mg/kg ww EC (1993) ND 1 mg/kg ww EC (1993)

Total ND 1 mg/kg ww EC (1993) 2.63 mg/kg ww Muir et al. (2000) Europe Fish 7.00×10

- 3 mg/kg ww Reth et al. (2006) 4.70×10

- 2 mg/kg ww Reth et al. (2006)

North Sea/Balti c Sea Region 2 Fish ND 3 mg/kg ww IVL (2009) 2.6×10

- 1 mg/kg ww Reth et al. (200 5)

United States Fish 2.90×10

- 3 mg/kg ww Tomy and Stern (1999) 9.04×10

- 1 mg/kg ww Tomy and Stern (1999)

Invertebrates 3.50×10

- 3 mg/kg ww USEPA (1988) 1.70×10

- 1 mg/kg ww USEPA (1988)

Total 2.90×10

- 3 mg/kg ww Tomy and Stern (1999) 9.04×10

- 1 mg/kg ww Tomy and Stern (1999)

United States / Canada

- Great Lakes Fish 1.80×10

- 3 mg/kg ww Muir et al. (2003) 1.10×10

- 1 mg/kg ww Muir et al. (2003)

Invertebrates 2.40×10

- 3 mg/kg ww EC (2008a) 1.60×10

- 2 mg/kg ww Muir et al. (2003)

Total 1.80×10

- 3 mg/kg ww Muir et al. (2003) 1.10×10

- 1 mg/kg ww. Muir et al. (2003)

Terrestrial Biota Europe Birds 5.00×10

- 3 mg/kg ww Reth et al. (2006) 3.70×10

- 1 mg/kg ww Reth et al. (2006)
Notes: Summary values represent total MCCP and not individual MCCP isomers.

1 MCCPs were not detected in Invertebrates from Canada. Detection limit = 4.0 × 10

- 7 mg/kg; 1⁄2 DL = 2.0 × 10

- 7 (EC, 1993)

. 2 North Sea/Baltic Sea Region includes the following countries: Estonia, Latvia, Li thuania, Norway, Poland, and Sweden.
3 The minimum MCCP concentration value for fish from the North Sea/Baltic128 

### LCCP MONITORING DATA Kemmlein et al. (2002) optimized and tested the carbon skeleton reaction gas chromatography analytical method to analyze environmental samples for CPs. The optimized method was used for marine sediments, mussels and crabs taken from an area influenced b y a CP manufacturer in Yarraville, Australia. LCCP (C 18

- 20 ) concentrations in marine sediment ranged from 1.02 × 10

- 1 to 4.31 × 10

- 1 mg/kg dw ( Table_Apx D

- 2 ), those in mussels ranged from 4 × 10

- 1 to 1.9 mg/kg lw, and those in crab ranged from 3 × 10

- 2 to 4.4 mg/kg lw. The results presented in this paper show that bioaccumulation is evident. The mussel samples contained approximately two times and crab tissue around six times the concentration of CPs found in the most contaminated sediment sample. No other adequate studies were found to characterize LCCP concentrations in surface water, fresh wate r sediment or soil.

Table_Apx D

- 2

- 1 : Marine Sediment Concentrations of LCCPs

Media Country Location

Comments Concentration Units References City, State or Province

Sediment (Marine) Australia NR Sum 1.02×10

- 1 mg/kg dw Kemmlein et al. (2002)

Sum 1.28×10

- 1 mg/kg dw Kemmlein et al. (2002)

Sum 3.04×10

- 1 mg/kg dw Kemmlein et al. (2002)

Sum 4.3 1×10

- 1 mg/kg dw Kemmlein et al. (2002)
Notes: 1. Values provided in the table above represent total LCCP (C 18

- 20 ) a nd not individual isomers. 2. dw

– dry weight
Table_Apx D

- 2

- 2 : LCCP Biota Concentrations of LCCPs

Media Country Location

Minimum Maximum Units References City, State

or Province Aquatic

Biota Australia NR 2.89×10

- 6 6.90×10

- 6 mg/kg lw Kemmlein et al. (2002)

Notes: 1. Minimum and Maximum concentrations provided in the table above rep resent total LCCP (C 18

- 20 ) and not individual isomers. 2. lw

– lipid weight129 

## Appendix E ENGINEERING (ChemSTEER) REPORTS ON P

- 12

- 0282, P

- 12

- 0283 and P

- 12

- 0284 (Used for both identifying potential releases to the environment and for estimated occupational
exposures . SEE AP PENDIX G FOR REFERENCE TO FULL REPORTS UNDER SEPARATE COVER )

P

- 12

- 0282

Scenario # Sites # Workers MFG: Batch Manufacturing of PMN (92.3% of PV) 1 17 Exposure mg/day day/yr Inhalation

- Liquid Inhalation exposure is expectd to be negligible (VP < 0.001 torr) Dermal

- Liquid High End 8.8E+2 250 Dermal

- Liquid High End 1.8E+3 250 Release kg/site/day day/yr Incineration Relese from 0357 / Release from all 3 sites 1.4E+1 / 1.4E+1 251

Incineration or Landfill Conservative 4.1E+2 1 Landfill Release from 0282/0283 / Release from all 3 sites 2.2E+0 / 2.2E+0 251

Landfill Relese from 0357 / Release from all 3 sites 6.0E+0 / 6.0E+0 251

Scenario # Sites # Workers PROC1: Plastics Compounding (21.5% of PV) 32 768 Exposure mg/day day/yr Inhalation

- Particulate High End of Range / Low end of range 4.4E+0 / 3.0E

- 2 250

Dermal

- Particulate High End 1.8E+3 250 Release kg/site/day day/yr Water Output 2 8.4E

- 3 265 Water Output 2 4.2E

- 2 265 Water or Incineration or Landfill GS Estimate 6.2E+0 128 Water or Incineration or Landfill EPA/OPPT Model / Submitter Estimate 4.2E+0 / 6.5E+0 94

Water or Incineration or Landfill Daily cleaning / cleaning 1x per 5 bt 1.7E+1 / 1.7E+1 265

Water or Incineration or Landfill Suggested Estimate 8.4E

- 2 265 Air Output 2 8.4E

- 3 265 Air Output 2 4.2E

- 2 265130 P

- 12

- 0282

Scenario # Sites # Workers USE1: Plastics Converting (21.5% of PV) 86 4128 Exposure mg/day day/yr Inhalation

- Particulate Output 2 / Output 1 1.2E+1 / 10.0E

- 2 250

Dermal

- Particulate PMN will be encapsulated in plastic pellets during handling of plastic raw material. While some surface contact may occur, dermal exposure to solids in this form are

non

- quantifiable (2004 ESD; Cast Solids, CEB Method for Screening

- Level Assessments of Dermal Exposure). Release kg/site/day day/yr Water Suggested Estimate 1.7E

- 2 250

Water or Incineration or Landfill Suggested Estimate 3.3E

- 2 250

Water or Incineration or Landfill Output 2 3.3E+0 250

Water or Incineration or Landfill Cleaning 1x per wk / Daily Cleaning 6.7E+0 / 6.7E+0 50

Water or Landfill Suggested Estimate 8.3E+0 250

Water or Landfill Output 2 3.3E

- 2 250 Air Suggested Estimate 1.7E

- 2 250

Scenario # Sites # Workers PROC2: Rubber Compounding (34.4% of PV) 33 792 Exposure mg/day day/yr Inhalation

- Particulate High End of Range / Low End of Range 7.0E

- 1 / 10.0E

- 2 153

Dermal

- Particulate High End 1.8E+3 153 Release kg/site/day day/yr Water Output 2 2.3E

- 2 153 Water Output 2 1.1E

- 1 153 Water or Incineration or Landfill Submitter Estimate / EPA/OPPT Model 1.9E+0 / 8.1E+0 153

Water or Incineration or Landfill EPA/OPPT Model 4.2E+0 146 Water or Incineration or Landfill Cleaning 1x per week / Cleaning each day 4.5E+1 / 4.5E+1 30

Water or Incineration or Landfill Suggested Estimate 2.3E

- 1 153 Air Output 2 2.3E

- 2 153 Air Output 2 1.1E

- 1 153131 P

- 12

- 0282

Scenario # Sites # Workers USE2: Rubber Converting (34.4% of PV) 33 1584 Exposure mg/day day/yr Inhalation

- Vapor High end of range / Low end of range 7.0E

- 1 / 10.0E

2 250

Dermal

- Vapor PMN will be encapsulated in rubber sheets during handling of rubber raw material. While some surface contact may occur, dermal exposure to solids in this form are

non

- quantifiable (2004 ESD; Cast Solids, CEB Method for Screening

- Level Assessments of Dermal Exposure). Release kg/site/day day/yr Water Suggested Estimate 6.9E

- 2 250

Water or Incineration or Landfill Suggested Estimate 1.4E

- 1 250

Water or Incineration or Landfill Cleaning 1x per wk / Daily Cleaning 2.8E+1 / 2.8E+1 50

Water or Incineration or Landfill Output 2 1.4E+1 250

Water or Landfill Suggested Estimate 3.4E+1 250

Water or Landfill Output 2 1.4E

- 1 250 Air Suggested Estimate 6.9E

- 2 250

Scenario # Sites # Workers PROC3: Formulation of Metalworking Fluids (43.3% of PV) 110 1430

Exposure mg/day day/yr Inhalation

- Liquid negligible (VP < 0.001 torr) Dermal

- Liquid High End 1.8E+3 143 Dermal

- Liquid High End 3.5E+2 143 Release kg/site/day day/yr Water or Incineration or Landfill High End 6.2E+0 138

Water or Incineration or Landfill High End 4.2E+0 49

Water or Incineration or Landfill Conservative 9.1E+0 143

Water or Incineration or Landfill Output 2 2.3E+0 143

Air High End (Suggested Est) 4.6E

- 1 143

Incineration or Landfill Output 2 1.8E+1 143132 P

- 12

- 0282

Scenario # Sites # Workers USE3: Use of Metalworking Fluids (43.3% of PV) (GS Estimates

- "Large Shop") 1576 75648

Exposure mg/day day/yr Inhalation

- Mist High End of Range / Typical 7.1E+0 / 6.6E

- 1 247

Dermal

- Mist High End 3.5E+2 247 Dermal

- Mist / / 3.9E+3 Release kg/site/day day/yr Water GS Estimate 3.9E+0 247 Water or Incineration or Landfill High End 1.2E+0 218 Water or Incineration or Landfill Output 2 1.3E+1 247 Water or Incineration or Landfill GS est 1/day 1.6E+1 247 Air 1% submission estimate / 5% 2002 EU RA 3.6E

- 1 / 1.8E+0 247

Scenario # Sites # Workers USE3: Use of Metalworking Fluids (43.3% of PV) (Submission Estimates

- "Small Shop") 28200 253800

Exposure mg/day day/yr Inhalation

- Mist Output 2 / Output 1 7.1E+0 / 6.6E

- 1 250 Dermal

- Mist High End 3.5E+2 250 Dermal

- Mist Output 2 3.9E+3 250 Release kg/site/day day/yr Water GS Estimate 1.8E

- 1 300 Water or Incineration or Landfill High End 1.2E+0 12 Water or Incineration or Landfill Output 2 5.9E

- 1 300 Water or Incineration or Landfill Output 2 7.6E

- 1 300 Air Output 1 / Output 2 1.6E

- 2 / 8.2E

- 2 300133 P

- 12

- 0283

Scenario # Sites # Workers MFG: Batch Manufacturing of PMN 2 8 Exposure mg/day day/yr Inhalation

- Liquid Inhalation exposure is expectd to be negligible (VP < 0.001 torr) Dermal

- Liquid High End 8.8E+2 93 Dermal

- Liquid High End 1.8E+3 93 Release kg/site/day day/yr Incineration or Landfill Conservative 4.1E+2 1 Landfill Output 2 2.2E+0 93

Scenario # Sites # Workers PROC1: Plastics Compounding (43% of PV) 32 768 Exposure mg/day day/yr Inhalation

- Particulate High End of Range / Low End of Range 4.4E+0 / 3.0E

- 2 60

Dermal

- Particulate High End 1.8E+3 60 Release kg/site/day day/yr Water Output 2 8.5E

- 3 60 Water or Incineration Output 1 1.7E+1 60 Water or Incineration or Landfill Output 1 6.2E+0 56 Water or Incineration or Landfill Output 1 4.2E+0 19 Water or Incineration or Landfill Suggested Value 8.5E

- 2 60 Air Output 2 8.5E

- 3 60

Scenario # Sites # Workers USE1: Plastics Converting (43% of PV) 32 1536 Exposure mg/day day/yr Inhalation

- Particulate OSHA PEL / 2008 EU RA Data 4.5E+1 / 1.2E+1 250

Dermal

- Particulate PMN will be encapsulated in plastic pellets during handling of plastic raw material. While some surface contact may occur, dermal exposure to solids in this form are

non

- quantifiable (2004 ESD; Cast Solids, CEB Method for Screening

- Level Assessments of Dermal Exposure). Release kg/site/day day/yr134 Water Suggested Estimate 1.5E

- 1 250

Water or Incineration or Landfill Suggested Estimate 2.0E

- 2 250

Water or Incineration or Landfill Output 2 2.0E+0 250

Water or Incineration or Landfill Conservative 4.1E+0 250

Water or Landfill Suggested Estimate 5.1E+0 250

Air Suggested Estimate 1.5E

- 1 250

Air Output 2 4.6E

- 2 250 Landfill Output 2 4.5E+0 250

P

- 12

- 0283

Scenario # Sites # Workers PROC2: Rubber Compounding (24% of PV) 33 792 Exposure mg/day day/yr Inhalation

- Particulate High End of Range / Low End of Range 7.0E

- 1 / 10.0E

- 2 31

Dermal

- Particulate High End 1.8E+3 31 Release kg/site/day day/yr Water Output 2 8.9E

- 3 31 Water or Incineration or Landfill Submitter Estimate / EPA/OPPT Model 1.5E+0 / 6.2E+0 30

Water or Incineration or Landfill Submitter Estimate / EPA/OPPT Model 3.1E+0 / 4.2E+0 10

Water or Incineration or Landfill Output 2 1.8E+1 31 Water or Incineration or Landfill Suggested Estimate 8.9E

- 2 31 Air Output 2 8.9E

- 3 31

Scenario # Sites # Workers USE2: Rubber Converting (24% of PV) 33 1584 Exposure mg/day day/yr Inhalation

- Particulate OSHA PEL / 2008 EU RA Data 1.5E+1 / 7.0E

1 250

Dermal

- Particulate PMN will be encapsulated in rubber sheets during handling of rubber raw material. While some surface contact may occur, dermal exposure to solids in this form are

non

- quantifiable (2004 ESD; Cast Solids, CEB Method for Screening

- Level Assessments of Dermal Exposure). Release kg/site/day day/yr135 Water Suggested Estimate 2.7E

- 1 250

Water or Incineration or Landfill Suggested Estimate 1.1E

- 2 250

Water or Incineration or Landfill Conservative 2.2E+0 250

Water or Incineration or Landfill Output 2 1.1E+0 250

Water or Landfill Suggested Estimate 2.7E+0 250

Air Suggested Estimate 2.7E

- 1 250

Air Output 2 1.2E

- 3 250 Landfill Output 2 1.2E

- 1 250

P

- 12

- 0283

Scenario # Sites # Workers PROC3: Formulation of Metalworking Fluids (30% of PV) 110 440

Exposure mg/day day/yr Inhalation

- Liquid negligible (VP < 0.001 torr) Dermal

- Liquid High End 1.8E+3 12 Dermal

- Liquid High End 1.8E+2 12 Release kg/site/day day/yr Water or Incineration or Landfill High End 6.2E+0 11

Water or Incineration or Landfill High End 4.2E+0 4

Water or Incineration or Landfill Conservative 1.7E+1 12

Air Suggested Estimate 4.3E

- 1 12

Incineration or Landfill Suggested Estimate 1.7E+1 12

Scenario # Sites # Workers USE3: Use of Metalworking Fluids (30% of PV) 250 12000 Exposure mg/day day/yr Inhalation

- Mist High End of Range / Typical 3.6E+0 / 9.8E

- 1 247

Dermal

- Mist High End 1.8E+2 247 Dermal

- Mist / / 4.9E+2 Release kg/site/day day/yr Water Output 2 2.0E+0 247 Water or Incineration or Landfill High End 6.2E

- 1 218 Water or Incineration or Landfill Output 2 6.4E+0 247 Water or Incineration or Landfill Output 2 8.2E+0 247 Air Output 2 8.9E

- 1 247

Scenario # Sites # Workers136 PROC4: Formulation of Engine Lubricants (2% of PV) 15 225

Exposure mg/day day/yr Inhalation

- Liquid negligible (VP < 0.001 torr) Dermal

- Liquid High End 1.8E+3 12 Dermal

- Liquid High End 7.1E+1 12 Release kg/site/day day/yr Water Output 2 2.8E+0 12 Water or Incineration or Landfill High End 6.2E+0 6

Water or Incineration or Landfill Submitter Estimate 9.2E+0 12

Water or Incineration or Landfill High End 4.2E+0 2

P

- 12

- 0283

Scenario # Sites # Workers USE4: Commerical Use of Engine Lubricants (1.6% of PV) 16 80

Exposure mg/day day/yr Inhalation

- Liquid negligible (VP < 0.001 torr) Dermal

- Liquid High End 7.1E+1 250 Release kg/site/day day/yr Incineration Output 2 1.3E+1 250 Incineration or Landfill High End 9.1E

- 2 250 Landfill Output 2 1.2E+0 250

Scenario # Sites # Workers PROC5: Formulation of Paints, Sealants, and Adhesives (0.5% of PV) 33 1287

Exposure mg/day day/yr Inhalation

- Liquid negligible (VP < 0.001 torr) Dermal

- Liquid High End 1.8E+3 2 Dermal

- Liquid High End 2.6E+2 2 Release kg/site/day day/yr Water Output 2 2.9E

- 2 2 Water or Incineration or Landfill High End 6.2E+0 1

Water or Incineration or Landfill Submitter Estimate 7.2E+0 2

Incineration or Landfill Output 2 2.6E

- 1 2 Incineration or Landfill Output 2 2.9E+0 2 Incineration or Landfill 2002 EU RA Estimate 1.4E+1 2

Scenario # Sites # Workers137 USE5: Application of Paints, Sealants, and Adhesives (0.5% of PV) 38 114

Exposure mg/day day/yr Inhalation

- Mist What

- If 9.0E+1 249 Dermal

- Mist High End 2.6E+2 249 Dermal

- Mist High End 1.3E+3 249 Release kg/site/day day/yr Water or Incineration or Landfill High End 1.7E

- 2 175 Water or Incineration or Landfill ESD Estimate 4.0E

- 2 249 Air Output 2 1.5E

- 1 249 Landfill Output 2 1.3E+0 249

P

- 12

- 0283

Scenario # Sites # Workers USE6: Plasticizer/Flame Retardant in Textiles (0.5% of PV) 11 33

Exposure mg/day day/yr Inhalation

- Liquid negligible (VP < 0.001 torr) Dermal

- Liquid High End 4.3E+2 95 Dermal

- Liquid High End 1.8E+3 95 Release kg/site/day day/yr Water Output 2 3.6E+0 95 Water or Incineration or Landfill High End 6.2E+0 2

Water or Incineration or Landfill Submitter Estimate 4.6E+0 1

P

- 12

- 0284

Scenario # Sites # Workers MFG: Batch Manufacturing of PMN 1 4 Exposure mg/day day/yr Inhalation

- Liquid Inhalation exposure is expectd to be negligible (VP < 0.001 torr) Dermal

- Liquid High End 8.8E+2 3 Dermal

- Liquid High End 1.8E+3 3 Release kg/site/day day/yr138 Incineration or Landfill EPA/OPPT Model / Submitter Estimate 3.0E+2 / 4.1E+2 1

Landfill Output 2 2.2E+0 3

Scenario # Sites # Workers PROC1: Formulation of Metalworking Fluids 24 312

Exposure mg/day day/yr Inhalation

- Liquid negligible (VP < 0.001 torr) Dermal

- Liquid High End 1.8E+3 6 Dermal

- Liquid High End 2.2E+2 6 Release kg/site/day day/yr Water or Incineration or Landfill High End 7.1E+0 6

Water or Incineration or Landfill High End 4.2E+0 1

Water or Incineration or Landfill Conservative 4.6E+0 6

Water or Incineration or Landfill Output 2 1.1E+0 6

Air 2010 Japanese RA Estimate 2.3E

- 2 6

Incineration or Landfill 2002 EU RA Low End Est / 2002 EU RA High End Est 4.6E+0 / 9.2E+0 6

P

- 12

- 0284

Scenario # Sites # Workers USE1: Use of Metalworking Fluids (GS Throughput

- "Large shop") 12 576

Exposure mg/day day/yr Inhalation

- Mist High End of Range / Typical 4.4E+0 / 1.2E+0 247

Dermal

- Mist High End 2.2E+2 247 Dermal

- Mist / / 2.4E+3 Release kg/site/day day/yr Water GS Estimate 2.4E+0 247 Water or Incineration or Landfill High End 7.8E

- 1 220 Water or Incineration or Landfill Output 2 8.0E+0 247 Water or Incineration or Landfill GS est 1/day 1.0E+1 247 Air 1% submission estimate / 5% 2002 EU RA 2.2E

- 1 / 1.1E+0 247

Scenario # Sites # Workers USE1: Use of Metalworking Fluids (Submission Throughput

- "Small Shop") 364 3276139 Exposure mg/day day/yr Inhalation

- Mist Output 2 / Output 1 4.4E+0 / 1.2E+0 250 Dermal

- Mist High End 2.2E+2 250 Dermal

- Mist High End 2.4E+3 250 Release kg/site/day day/yr Water GS Estimate 6.7E

- 2 300 Water or Incineration or Landfill High End 7.8E

- 1 7 Water or Incineration or Landfill Output 2 2.2E

- 1 300 Water or Incineration or Landfill GS est 1/day 2.8E

- 1 300 Air 1% submission estimate / 5% 2002 EU RA 6.1E

- 3 / 3.0E

- 2 300140 

## Appendix F EXPOSURE SCENARIO ESTIMATES

(E

- FAST Model Run

. SEE A PPENDIX G FOR REFERENCE TO FULL REPORTS UNDER SEPARATE COVER )

INITIAL REVIEW EXPOSURE REPORT (IRExR) Chemical ID: P

- 12

- 0282

Results Table: Dose, Concentration, and Days Exceeded Results Summary Exposure Scenario 1 Water Landfill Stack Air Fugitive Air

Release activity(ies) 2 ; exposure calculation(s) 3 Drinking Water Fish Ingestion 7Q10 4 CC=1 PDM Days Exceede d LADD ADR LADD ADR LADD ADR LADD ADR LADD

mg/kg/day mg/kg/da
y mg/kg/d
ay mg/kg/d
ay μg/l # Days mg/kg/da
y mg/kg/da
y mg/kg/da
y mg/kg/da
y mg/kg/da
y Mfg: Max ADR & Max LADD --- --- --- --- --- --- 2.17E

- 04 0.51 1.11E

- 04

--- --- Proc1: Max ADR, & PDM1 2.09E

- 02

--- 0.43

--- 427.32 58

--- 4.04E

- 02

--- 1.38E

- 04

---

Proc1: PDM2

--- --- --- --- 343.56 174

--- --- --- --- ---

Proc1: PDM3

--- --- --- --- 219.86 288

--- --- --- --- ---

Proc1: Max LADD

--- 3.76E

- 04

--- 4.60E

- 03

--- --- 4.30E

- 04

--- 1.89E

- 03

--- 3.49E

- 06

Proc2: Max ADR 1.79E

- 02

--- 0.37

--- --- --- --- 3.49E

- 02

--- 2.24E

- 04

---

Proc2: PDM1

--- --- --- --- 367.24 32

--- --- --- --- ---

Proc2: PDM2

--- --- --- --- 313.12 94

--- --- --- --- ---

Proc2: PDM3

--- --- --- --- 233.22 96

--- --- --- --- ---

Proc2: Max LADD

--- 1.32E

- 04

--- 1.62E

- 03

--- --- 5.11E

- 04

--- 6.66E

- 04

--- 1.88E

- 06

Proc3: Max ADR 1.72E

- 02

--- 0.36

--- --- --- --- 5.52E

- 02

--- 1.06E

- 02

---

Proc3: PDM1

--- --- --- --- 298.97 35

--- --- --- --- ---

Proc3: PDM2

--- --- --- --- 219.07 38

--- --- --- --- ---

Proc3: Max LADD

--- 4.93E

- 05

--- 6.03E

- 04

--- --- 3.25E

- 04

--- 4.24E

- 04

--- 3.52E

- 05

Proc4: Max ADR 1.41E

- 02

--- 0.29

--- --- --- --- 2.39E

- 02

--- --- ---

Proc4: PDM1

--- --- --- --- 154.64 34

--- --- --- --- ---

Proc4: Max LADD

--- 2.91E

- 05

--- 3.56E

- 04

--- --- 9.23E

- 05

--- 1.21E

- 04

--- ---

Proc5: Max ADR 1.10E

- 02

--- 0.23

--- --- --- --- 4.96E

- 02

--- --- ---

Proc5: Max LADD

--- 2.82E

- 06

--- 3.45E

- 05

--- --- 3.37E

- 05

--- 4.39E

- 05

--- ---

Use1: Max ADR 2.18E

- 02

--- 0.41

--- --- --- --- 2.39E

- 02

--- 1.53E

- 02

---

Use1: PDM1

--- --- --- --- 446.80 250

--- --- --- --- ---

Use1: Max LADD

--- 4.30E

- 04

--- 5.26E

- 03

--- --- 2.56E

- 03

--- 1.32E

- 03

--- 3.36E

- 04

Use2: Max ADR 3.65E

- 02

--- 0.46

--- --- --- --- 1.27E

- 02

--- 2.21E

- 02

---

Use2: PDM1

--- --- --- --- 777.14 250

--- --- --- --- ---141 Exposure Scenario 1 Water Landfill Stack Air Fugitive Air

Release activity(ies) 2 ; exposure calculation(s) 3 Drinking Water Fish Ingestion 7Q10 4 CC=1 PDM Days Exceede d LADD ADR LADD ADR LADD ADR LADD ADR LADD

mg/kg/day mg/kg/da
y mg/kg/d
ay mg/kg/d
ay μg/l # Days mg/kg/da
y mg/kg/da
y mg/kg/da
y mg/kg/da
y mg/kg/da
y
Use2: Max LADD

--- 4.80E

- 04

--- 5.88E

- 03

--- --- 1.17E

- 03

--- 6.94E

- 04

--- 4.89E

- 04

Use3: Max ADR 3.07E

02 --- 0.37

--- --- --- --- 1.84E

- 02

--- 2.21E

- 02

---

Use3: PDM1

--- --- --- --- 652.27 218

--- --- --- --- ---

Use3: PDM2

--- --- --- --- 628.79 247

--- --- --- --- ---

Use3: Max LADD

--- 3.81E

- 04

--- 4.67E

- 03

--- --- 7.79E

- 04

--- 1.01E

- 03

--- 4.84E

- 04

Use4: Max ADR

--- --- --- --- --- --- --- 1.62E

- 02

--- --- ---

Use4: Max LADD

--- --- --- --- --- --- 6.74E

- 05

--- 8.90E

- 04

--- ---

Use5: Max ADR 2.46E

04 --- 2.63E

03 --- --- --- --- 7.06E

- 05

--- 3.77E

- 03

---

Use5: PDM1

--- --- --- --- 5.38 60

--- --- --- --- ---

Use5: PDM2

--- --- --- --- 3.77 68

--- --- --- --- ---

Use5: Max LADD

--- 2.51E

06 --- 3.07E

05 --- --- 7.59E

- 05

--- 3.54E

- 06

--- 8.31E

- 05

Use6: Max ADR 8.80E

03 --- 0.18

--- --- --- --- 1.29E

- 02

--- --- ---

Use6: PDM1

--- --- --- --- 46.39 289

--- --- --- --- ---

Use6: Max LADD

--- 5.96E

05 --- 7.29E

04 --- --- 9.50E

- 06

--- 1.24E

- 05

--- ---

1 Exposure scenario titles consist of release activity followed by exposure calculation abbreviation. 2 Release activities are from engineering report's Manufacturing (Mfg), Processing (Proc) and Use release activity labels. Multiple release activities are combined in one exposure scenario if their releases occur at same location. 3 Exposure calculations are Acute Dose Rate (ADR), Life time Average Daily Dose (LADD), and Probabilistic Dilution Model (PDM). There may be one, two, or all three exposure calculations per exposure scenario. CC is the aquatic concentration of concern. 4 This column displays concentration values for the 7Q10 st reamflow, which is defined as the average streamflow of the 7 consecutive days of lowes t flow within a 10 year period.

Results Table: Consumer Exposure Results Summary

Scenario Water (DtD) Dermal Inhalation Drinking Water Fish Ingestion
7Q10 PDM Days Exceede d ADR LADD ADR LADD ADR LADD ADR LADD

mg/kg/d

ay mg/kg/d
ay mg/kg/d
ay mg/kg/d
ay μg/l # Days mg/kg/d
ay mg/kg/da
y mg/kg/ day mg/kg/d
ay CEM, Used Motor Oil --- --- --- --- --- --- 2.46E+1 2.05E

- 01

--- ---142 INITIAL REVIEW EXPOSURE REPORT (IRExR)

Chemical ID: P

- 12

- 0283

Exposure Scenario 1 Water Landfill Stack Air Fugitive Air

Release activity(ies) 2 ; exposure calculation(s) 3 Drinking Water Fish Ingestion 7Q10 4 CC=1 PDM Days Exceede d LADD ADR LADD ADR LADD ADR LADD ADR LADD

mg/kg/day mg/kg/da
y mg/kg/d
ay mg/kg/d
ay μg/l # Days mg/kg/da
y mg/kg/da
y mg/kg/da
y mg/kg/da
y mg/kg/da
y
MFG: max ADR

--- --- --- --- --- --- --- 0.51

--- --- --- MFG: max LADD --- --- --- --- --- --- 1.28E

- 04

--- 1.11E

- 04

--- --- PROC1: max ADR, acute eco 1.73E

- 02

--- 0.36

--- 354.30

--- --- 3.49E

- 02

--- 2.13E

- 04

PROC1: PDM1

--- --- --- --- 300.17 56

--- --- --- --- ---

PROC1: PDM2

--- --- --- --- 220.28 60

--- --- --- --- --- PROC1: max LADD --- 7.84E

- 05

--- 9.59E

- 04

--- --- 9.00E

- 05

--- 3.94E

- 04

--- 1.13E

- 06

PROC2: max ADR, acute eco 1.79E

- 02

--- 0.37

--- 367.24

--- --- 3.49E

- 02

--- 2.24E

- 04

---

PROC2: PDM1

--- --- --- --- 313.12 30

--- --- --- --- ---
PROC2: PDM2

--- --- --- --- 233.22 31

--- --- --- --- --- PROC2: max LADD --- 4.26E

- 05

--- 5.21E

- 04

--- --- 1.64E

- 04

--- 2.14E

- 04

--- 6.07E

- 07

PROC3: max ADR, acut eco 1.72E

- 02

--- 0.36

--- 353.09

--- --- 5.52E

- 02

--- 1.08E

- 02

--- PROC3: max LADD --- 1.56E

- 05

--- 1.91E

- 04

--- --- 1.03E

- 04

--- 1.34E

- 04

--- 1.14E

- 05

PROC4: Max ADR, acute eco 1.41E

- 02

--- 0.29

--- 288.66

--- --- 2.39E

- 02

--- --- --- PROC4: max LADD --- 1.02E

- 05

--- 1.25E

- 04

--- --- 3.25E

- 05

--- 4.24E

- 05

--- --- PROC5: max ADR, acute eco 8.44E

- 03

--- 0.17

--- 173.05

--- --- 3.86E

- 02

--- --- --- PROC5: max LADD --- 1.11E

- 06

--- 1.36E

- 05

--- --- 1.14E

- 05

--- 1.49E

- 05

--- --- USE1: max ADR, PDM, LADD 6.93E

- 03 1.36E

- 04 0.13 1.67E

- 03 141.77 248 8.18E

- 04 7.57E

- 03 4.15E

- 04 4.93E

- 03 1.08E

- 04
USE2: max ADR, PDM, LADD 1.17E

- 02 1.54E

- 04 0.15 1.89E

- 03 249.25 247 3.19E

- 04 4.10E

- 03 2.25E

- 04 6.80E

- 03 1.49E

- 04
USE3: max ADR, PDM1 3.07E

- 02

--- 0.37

--- 652.27 218

--- 1.84E

- 02

--- 2.21E

- 02

---

USE3: PDM2

--- --- --- --- 628.79 247

--- --- --- --- --- USE3: max LADD --- 3.81E

- 04

--- 4.67E

- 03

--- --- 7.79E

- 04

--- 1.01E

- 03

--- 4.84E

- 04
USE4: max ADR, LADD --- --- --- --- --- --- 6.72E

- 05 1.62E

- 02 8.90E

- 04

--- --- USE5: max ADR, PDM1 2.46E

- 04

--- 2.63E

03 --- 5.38 60

--- 7.06E

- 05

--- 3.77E

- 03

--- USE5: max LADD --- 2.49E

- 06

--- 3.05E

- 05

--- --- 7.01E

- 05

--- 3.51E

- 06

--- 8.24E

- 05

USE5: PDM2

--- --- --- --- 3.77 67

--- --- --- --- --- USE6: max ADR,
acute eco 9.05E

- 03

--- 0.19

--- 185.57

--- --- 1.34E

- 02

--- --- ---143 USE6: PDM

--- --- --- --- 46.39 93

--- --- --- --- --- USE6: max LADD --- 1.94E

- 05

--- 2.37E

- 04

--- --- 3.54E

- 06

--- 4.61E

- 06

--- ---
1 Exposure scenario titles consist of release activity followed by exposure calculation abbreviation.
2 Release activities are from engineering report's Manufacturing (Mfg), Processing (Proc) and Use release activity labels. Multiple release activities are combined in one exposure scenario if their releases occur at same location.
3 Exposure calculations are Acute Dose Rate (ADR), Lifetime Average Daily Dose (LADD), and Probabilistic Dilution Mode l (PDM). There may be one, two, or all three exposure calculations per exposure scenario. CC is the aquatic concentration of concern.
4 This column displays concentration values for the 7Q10 streamflow, which is defined as the average streamflow the 7 consecutive days of lowest flow within a 10 year period.

Results Table: Consumer Exposure Results Summary

Scenario Water (DtD) Dermal Inhalation Drinking Water Fish Ingestion
7Q10 PDM Days Exceede d ADR LADD ADR LADD ADR LADD ADR LADD

mg/kg/d

ay mg/kg/d
ay mg/kg/d
ay mg/kg/d
ay μg/l # Days mg/kg/d
ay mg/kg/da
y mg/kg/ day mg/kg/d
ay CEM 1 : Used

Motor Oil (dermal) --- --- --- --- --- --- 2.46E+1 2.05E

- 01

--- ---

1 Consumer Exposure Module within NCEM2 model.144 Chemical ID: P

- 12

- 0284

Results Table: Dose, Concentration, and Days Exceeded Results Summary

Exposure
Scenario 1 Water Landfil l Stack Air Fugitive Air

Drinking Water Fish Ingestion

ADR LADD ADR LADD 7Q10 4
CC =

1 PDM Days Exceede d LADD ADR LADD ADR LADD

Release activity(ies) 2 ; exposure calculation(s) 3 mg/kg/da

y mg/kg/da

y mg/kg/da

y mg/kg/da

y μg/l # Days mg/kg/day mg/kg/da

y mg/kg/da

y mg/kg/da

y mg/kg/da
y

Mfg: Max ADR (stack);
LADD (landfill) --- --- --- --- --- --- 8.68E

- 05 0.51

--- --- ---

Proc1: Max ADR, Max

acute eco 1.47E

02 0.11

--- 301.55

--- --- 4.60E

- 02

--- --- ---

Proc1: Max
LADD --- 3.00E

- 06

--- 1.29E

- 05 --- --- 1.97E

- 05

--- --- --- ---

Use1: Max ADR, PDM1 1.05E

02 --- 6.95E

- 02

--- 214.71 218

--- --- --- 2.21E

- 02 ---

Use1: PDM2

--- --- --- --- 206.98 246

--- --- --- --- ---

Use1: Max
LADD --- 2.03E

- 04

--- 8.71E

- 04 --- --- 7.79E

- 04

--- 1.01E

- 03

--- 4.84E

04
1 Exposure scenario titles consist of release activity followed by exposure calculation abbreviation. 2 Release activities are from engineering report's Manufacturing (Mfg), Processing (Proc) and Use release activity labels. Multiple release activities are combined i n one exposure scenario if their releases occur at same location. 3 Exposure calculations are Acute Dose Rate (ADR), Lifetime Average Daily Dose (LADD), and Probabilistic Dilution Model (PDM). There may be one, two, or all three exposure calculations per

e xposure scenario. CC is the aquatic concentration of concern. 4 This column displays concentration values for the 7Q10 streamflow, which is defined as the average daily streamflow of the seven consecutive days of lowest flow within a ten year period.145 

## Appendix G SUPPLEMENTAL INFORMATION

SUPPLEMENTAL DATA FOR APPENDIX E (ChemSteer Engineering Reports):

p120282.ceb(ceb reviewed draft)

– 41 page pdf file p120283.ceb (ceb reviewed draft)

- 46 page pdf file p120284.ceb (ceb reviewed draft)

- 13 page pdf file

SUPPLEMENTAL DATA FOR APPENDIX F (E

- FAST Exposure Reports):

P

- 12

- 0282.exp1_Draft Final_072512

– 157 page pdf file P

- 12

- 0283.exp1_Draft Final_072512

– 122 page pdf file P

- 12

- 0284.exp1_Draft Final_041813

– 26 page pdf file