1

TSCA NEW CHEMICALS REVIEW PROGRAM

STANDARD REVIEW RISK ASSESSMENT ON

MEDIUM-CHAIN CHLORINATED PARAFFINS 

(PMN P-12-0453)

AND

LONG-CHAIN CHLORINATED PARAFFINS 

(PMN P-12-0433)

This assessment was conducted under EPA’s TSCA Section 5 New Chemicals 

Review 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-12-0453-0433

Office of Chemical Safety and 

Pollution Prevention
2

CONCLUSIONS

Based on its assessment of the available surrogate hazard and exposure information on P-12-

0453 and P-12-0433, EPA/OPPT concludes the following pertaining to the manufacturing, 

processing and use of these PMN substances: 

1. Occupational Exposures: given the assumptions, data and scenarios evaluated in this 

assessment, there were low risks found for workers from either dermal or inhalation 

exposures.

2. General Population Exposures (from environmental releases): given the assumptions, data 

and scenarios evaluated in this assessment, there were low risks found to humans from 

environmental releases via exposure to either drinking water or fish ingestion.

3. Environmental Assessment:

a. Using estimated environmental concentrations, the PMN substances may present an 

unreasonable risk following acute and chronic exposures to aquatic organisms.

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

environment as supporting information, the PMN substances:

i. Are 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 

MCCP and LCCP congener groups that may present an unreasonable risk 

following acute and chronic exposures to aquatic organisms. 

4. PBT Assessment: The PMN substances may be very persistent and very bioaccumulative.
3

TABLE OF CONTENTS

TABLE OF CONTENTS ............................................................................................................. 3

1 INTRODUCTION.............................................................................................................. 6

1.1 PMNS RECEIVED ...........................................................................................................................................6

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

1.3 USES .............................................................................................................................................................9

2 ENVIRONMENTAL FATE .................................................................................................. 9

2.1 ENVIRONMNETAL PERSISTENCE...................................................................................................................9

2.2 BIOCONCENTRATION AND BIOACCUMULATION........................................................................................11

3 ECOLOGICAL HAZARD OVERVIEW ................................................................................. 12

4 HUMAN HEALTH HAZARD OVERVIEW ........................................................................... 15

4.1 MCCP HEALTH DATA REVIEW .....................................................................................................................15

4.2 LCCP HEALTH DATA REVIEW.......................................................................................................................17

5 EXPOSURE INFORMATION ............................................................................................ 18

5.1 ENVIRONMENTAL MONITORING................................................................................................................18

5.2 MODELED ENVIRONMENTAL RELEASES .....................................................................................................21

5.3 EXPOSURE ESTIMATES................................................................................................................................23

5.3.1 OCCUPATIONAL EXPOSURE ESTIMATES .................................................................................................23

5.3.2 CONSUMER EXPOSURE ESTIMATES........................................................................................................23

6 RISK ASSESSMENT ........................................................................................................ 24

6.1 ENVIRONMENTAL ASSESSMENT.................................................................................................................24

6.1.1 Risk Estimates Using Environmental Monitoring Concentrations..........................................................24

6.1.2 Risk Estimates Using Modeled Exposures...............................................................................................26

6.2 HUMAN HEALTH.........................................................................................................................................27

6.2.1 Workers ..................................................................................................................................................28

6.2.2 General Population.................................................................................................................................29

7 CONCLUSIONS .............................................................................................................. 31

8 REFERENCES ................................................................................................................. 32

9 APPENDICES ................................................................................................................. 46

Appendix A ENVIRONMENTAL FATE AND BIOACCUMULATION STUDY SUMMARIES ......... 47

A-1 ENVIRONMENTAL PERSISTENCE ...............................................................................................................47

A-1-1 Abiotic Degradation ..............................................................................................................................47

A-1-1-1 Fate in Air........................................................................................................................................................ 47

A-1-2 Biodegradation......................................................................................................................................48

A-1-2-1 Fate in Wastewater Treatment....................................................................................................................... 48

A-1-2-2 Fate in Surface and Groundwater................................................................................................................... 49

A-1-2-3 Fate in Soil ...................................................................................................................................................... 49

A-2 BIOCONCENTRATION AND BIOACCUMULATION......................................................................................53
4

Appendix B ECOTOXICITY STUDY SUMMARIES ................................................................. 58

B-1 MCCP ECOTOXICITY DATA.........................................................................................................................58

B-1-1 Acute Fish Toxicity..................................................................................................................................58

B-1-2 Acute Aquatic Invertebrate Toxicity.......................................................................................................59

B-1-3 Algae Toxicity.........................................................................................................................................62

B-1-4 Chronic Fish Toxicity...............................................................................................................................63

B-1-5 Chronic Aquatic Invertebrate Toxicity....................................................................................................65

B-1-6 Chronic Aquatic Sediment Invertebrate Toxicity....................................................................................68

B-1-7 Avian Toxicity.........................................................................................................................................70

B-1-8 Terrestrial Invertebrate Toxicity.............................................................................................................71

B-1-9 Terrestrial Plant Toxicity ........................................................................................................................72

B-1-10 Conclusions ..........................................................................................................................................73

B-2 LCCP ECOTOXICITY DATA ..........................................................................................................................74

B-2-1 Acute Fish Toxicity..................................................................................................................................74

B-2-2 Acute Aquatic Invertebrate Toxicity.......................................................................................................75

B-2-3 Aquatic Plant Toxicity ............................................................................................................................76

B-2-4 Chronic Fish Toxicity...............................................................................................................................77

B-2-5 Chronic Aquatic Invertebrate Toxicity....................................................................................................77

B-2-6 Chronic Aquatic Sediment Invertebrate Toxicity....................................................................................81

B-2-7 Avian Toxicity.........................................................................................................................................81

B-2-8 Terrestrial Invertebrate Toxicity.............................................................................................................81

B-2-9 Terrestrial Plant Toxicity ........................................................................................................................81

B-2-10 Conclusions ..........................................................................................................................................81

Appendix C HUMAN HEALTH HAZARD STUDY SUMMARIES.............................................. 83

C-1 MCCP HEALTH DATA REVIEW....................................................................................................................83

C-1-1 Metabolism............................................................................................................................................84

C-1-2 Acute Toxicity.........................................................................................................................................84

C-1-3 Irritation and Sensitization.....................................................................................................................85

C-1-4 Repeated-dose Toxicity ..........................................................................................................................85

C-1-5 Genotoxicity ...........................................................................................................................................86

C-1-6 Carcinogenicity.......................................................................................................................................86

C-1-7 Developmental Reproductive Toxicity Review .......................................................................................91

C-2 LCCP HEALTH DATA REVIEW .....................................................................................................................97

C-2-1 Metabolism............................................................................................................................................97

C-2-2 Acute Toxicity.........................................................................................................................................98

C-2-3 Irritation and Sensitization.....................................................................................................................98

C-2-4 Repeated-dose Toxicity ..........................................................................................................................98
5

C-2-5 Genotoxicity ...........................................................................................................................................99

C-2-6 Carcinogenicity.......................................................................................................................................99

C-2-7 Developmental Reproductive Toxicity Review .......................................................................................99

Appendix D ENVIRONMENTAL MONITORING................................................................. 102

D-1 MCCP MONITORING DATA .....................................................................................................................102

D-1-1 Surface Water.....................................................................................................................................102

D-1-2 Sediment.............................................................................................................................................104

D-1-3 Biosolids and Soil.................................................................................................................................113

D-1-4 Biota....................................................................................................................................................113

D-2 LCCP MONITORING DATA .......................................................................................................................116

Appendix E ENGINEERING (ChemSTEER) REPORTS ON P-12-0-0433 and P-12-0453........... 117

Appendix F EXPOSURE SCENARIO ESTIMATES.................................................................. 121

Appendix G SUPPLEMENTAL INFORMATION.................................................................... 123
6

1 INTRODUCTION

1.1 PMNS RECEIVED

INEOS Chlor Americas, Inc. (hereinafter “INEOS”) submitted two Premanufacture Notices 

(PMNs) identified by EPA/OPPT as either medium-chain chlorinated paraffins (MCCPs; P-12-

0453) of varying chain lengths with the formula CxH(2x-y+2)Cly and “x” equaling 14 to 17 and “y” 

equaling 6 to > 24 or long-chain chlorinated paraffins (LCCPs; P-12-0433) of varying chain 

lengths with the formula of CxH(2x-y+2)Cly and “x” equaling 18 to 20 and “y” equaling 6 to > 24.

Table 1 lists the basic information INEOS supplied on these two PMNs which are intended to be 

sold under the trade name “Cereclor®”.

Table 1: Identification, Production Volume and Use of P-12-0453 and P-12-0433

PMN Chemical Name

1st Year 

Production

Volume (kg)

% PMN 

in final 

Product

Uses Log 

KOW

Water 

Solubility

P-12-

0453

Alkanes, C14-17, chloro

(MCCPs; 40-60 wt% 

Cl)

CBI 5-20

74%: lubricant in 

MWFs;

17%: flame 

retardant/plasticizer in 

polymers; 

7%: plasticizer in 

adhesives; and 

2%: lubricant in 

sealants.

4.70 

(E)

< 0.03 

mg/L (E)

P-12-

0433

Alkanes, C18-20, chloro

(LCCPs; 40-55 weight 

percent chlorination 

[wt% Cl])

CBI 15

100%: lubricant in 

metal working fluids 

(MWFs).

7.46 

(E)

< 0.006 

mg/L (E)

E = Estimated

Though 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 longchain chlorinated paraffins (vLCCPs) analogs. The continuum of carbon chain length and 

degree/percent chlorination (wt% Cl) in all of the chlorinated paraffins (CPs) is important and 

the relationship among them needs to be kept in mind.

1.2 CHEMISTRY

Shown below are the structures and chlorine content of P-12-0453 and P-12-0433 products are 

shown below. 
7

P-12-0453: the average molecular formula ranges from C14H26Cl4 (low weight at ~40 wt% Cl) 

to C17H26Cl10 (high weight at ~60 wt% Cl). The parent hydrocarbon had the following 

measured composition: C14, 36 wt% (range 30-40 wt%); C15, 30 wt% (range 25-35 wt%); C16,

24 wt% (range 20-30 wt%); and C17, 10 wt% (8-18 wt%). 

P-12-0433: the average molecular formula ranges from C17H31Cl5 (low weight ~40 wt% Cl) to 

C20H34Cl8 (high weight ~50 wt% Cl). The parent hydrocarbon had the following measured 

composition: C17, 17.06 wt% (max 20 wt%); C18, 64.71 wt% (range 45-70 wt%); C19, 13.67 wt% 

(range 15-27 wt%); and C20, 2.13 wt% (range 4-12 wt%).

CPs have an unknown or variable composition (classified as UVCB1 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 chain length into the following 

categories:

1. SCCPs (C10-13) 

2. MCCPs (C14-17) 

3. LCCPs (C18-20) 

4. Very long-chain CPs (vLCCP, 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 variety 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 of 

 1 UVCB are chemical substances whose composition is Unknown or Variable compositions, Complex reaction 

products and Biological materials.
8

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 pressures and water solubilities, and 

greater LogKOW (logarithm of octanol:water partition coefficient). EPA/OPPT used the 

physicochemical properties listed in Table 2 for informing its evaluation of P-12-0453 and P-12-

0433. 

Table 2: Summary of Physiochemical Informationa,b

wt% Cl Melting Point Boiling Point Vapor Pressure Water 

Solubility Log KOW

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 > 8

a

Source: EURAR (ECB, 2008); EA (2009) 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/OPPT’s Estimation Programs Interface 

(EPI) Suite TM. This estimate was generated using a representative MCCP (i.e., C14H24Cl6, 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 liquid chromatography to measure a LogKOW between 5.5 and 8.2 

and then chose to use a Log KOW = 7 as a representative LogKOW for MCCPs with 45-52 wt% Cl.

Analytical challenges exist with evaluating CPs due to the sheer number of congener groups2

that may be present in CP products. The existence of multiple chain lengths in the UVCBs such 

as P-12-0453 and P-12-0433 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 of 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.

Hüttig (2006) and Hüttig and Oehme (2006) reported that most commercial MCCP products 

were > 60 wt% of the C14 chain-length congener groups and the C15 chain-length congener 

groups comprised the majority of the remaining ∼ 40 wt%. The authors found < 15 wt% C16

chain-length congener groups present in most commercial MCCP samples and little to no C17

chain-length congener groups. Additional studies have reported that the C14 and C15 chain-length 

congener groups are the predominant MCCPs present in environmental media and in human 

breast milk (Bayen et al., 2006; Chen et al., 2011; Reth et al., 2006; Wang et al., 2013). Some 

variation is possible in commercial products where the C14 and C15 chain-length congener groups 

may not be the predominant congener groups in a specific MCCP product; however, even in 

these products, the C14 and C15 chain-length congener groups may serve as reasonable worse 

case surrogates for the C16-17 chain-length congener groups, due to their greater bioavailability 

and mobility in environmental media (ECB, 2005).

 2 For this report, congener groups are used to recognize the existence of different chain lengths and degrees of 

chlorination that could be present in any given CP product.
9

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 

(C14-17) are abundant in sediment has been explained using known water solubility and vapor 

pressure values in conjunction with predicted degradation pathways (de Boer, 2010). EPA/OPPT

determined that the toxic endpoints 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 (Cereclor S52®) is used as representative 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 were lacking for the 

PMNs, EPA/OPPT used measured data from SCCPs as surrogates for potential hazard and risks 

for the PMNs. 

1.3 USES

INEOS reported four uses in the PMN for P-12-0453, including:

1) 73.8% as a lubricant in metal working fluids (MWF) (no commercial or consumer uses). The 

notification substance is blended into the MWFs at 15%. 

2) 16.8% as a flame retardant/plasticizer in poly(vinyl chloride) (PVC) resins (no commercial or 

consumer uses). The notification substance is blended into the PVC resins at 15%. 

3) 7.4% as a plasticizer in adhesives (no commercial or consumer uses). The notification 

substance is blended into the adhesives at 5%. 

4) 2% as a lubricant in sealants (no commercial or consumer uses). The notification substance is 

blended into the sealants at 20%.

INEOS reported one category of use in the PMN for P-12-0433. This PMN substance is intended 

for use solely (i.e., 100% of the production volume) as a lubricant in MWF (no commercial or 

consumer uses). The PMN substance will be used at 15% in the final working formulation.

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 Canada (EC, 2008a)

and the EU (EA, 2009; ECB, 2005) to inform its assessment.

2.1 ENVIRONMNETAL PERSISTENCE

Abiotic 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 observed. The 

atmospheric half-life for MCCPs and LCCPs has been estimated at 1 - 2 days (EA, 2009; ECB, 

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, 
10

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 also 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 carbon 

chain cleavage may be possible 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 lesser-chlorinated 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, 2005). 

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 consistent with those 

provided by Canada and the EU. 

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

“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 states (ECB, 2005):

3

“No standard ready or inherent biodegradation tests results are available for mediumchain chlorinated paraffins. From the available information, medium-chain chlorinated 

paraffins can be considered to be not biodegradable in such test systems and so a 

biodegradation rate MCCPs of 0 day-1 is used in the risk assessment. 

 3

 Note, since the EU issued its assessment in 2005, standard inherent biodegradation studies were performed and are 

summarized in Appendix A.
11

There is evidence that some microorganisms may be capable of degrading MCCPs in the 

environment in acclimated or co-metabolic 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 

estimated for reaction with hydroxyl radicals. A value for the rate constant for the 

reaction (kOH) of 8 x 10-12 cm3 molecule-1 s-1 is used for the environmental modelling in 

the risk assessment.”

The UK assessment on LCCPs4 concluded the following (EA, 2009):

“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 with these characterizations. In 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 2 days and estimated half-lives in water, 

sediment and soil exceeding 2 months (60 days) (ECB, 2005; EA, 2009).

2.2 BIOCONCENTRATION 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 on 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 potential of MCCP and LCCP congener groups 

varies with the chain lengths and degree of chlorination and species evaluated. Shorter and less 

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 tissues. 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 factors (BCFs) or bioaccumulation factors (BAFs)

 4 Note, the UK assessment evaluated the following CPs under the term LCCP: C18-20 liquid products, C>20 liquid 

products, and C>20 solid products.
12

greater than 1000 or 5000 (EC, 2008a; ECB, 2005, 2008). This suggests that some congener 

groups in MCCP and LCCP products may be bioaccumulative or very bioaccumulative.

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

“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 LCCPs, the modelling results provided by the Modified Gobas BAF Model - which 

suggest that of all the LCCPs congeners only liquid C18-20 LCCPs have significant 

bioaccumulation potential -- are considered credible.” 

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

“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 cases. 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 to be minimal.” 

EPA/OPPT generally concurs with these characterizations. In 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 bioaccumulative to 

very bioaccumulative based on multiple lines of evidence, including: Log KOW values, modeled 

BCFs, laboratory-measured BCFs, field-measured BAFs, field-measured biomagnification 

factors (BMFs), laboratory-measured biota-sediment accumulation factors (BSAFs) and the 

presence of MCCPs and LCCPs in human and wildlife biota.

3 ECOLOGICAL 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 MCCPs, a high quality study was available on an SCCP product (C10-12, 58 wt% Cl) 

with similar physicochemical properties to MCCPs and was used for informing EPA/OPPT’s 

hazard evaluation (ECB, 2000). 

For LCCPs, ecotoxicity studies 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 hazard 

evaluation of LCCPs. This decision was considered a reasonable worst-case scenario because P12-0433 contains up to 20 wt% C17, a component of MCCPs (e.g., P-12-0453 contains between 
13

8-18 wt% C17). EPA/OPPT concludes that the studies summarized in Table 3 were the highest 

quality for assessing potential hazards in the aquatic, sediment and terrestrial compartments. 

Table 3: Summary of Aquatic, Sediment and Terrestrial Ecotoxicity Data for MCCPs and LCCPs

Test Substance Test Organism 

(Species) Test Guideline; Study type Endpoint Valuea Reference

Aquatic Invertebrates

Cereclor S-52 

(52 wt% Cl, 

C14-17)

Water flea

(Daphnia magna)

OECD 202, 1984; Acute 

immobilization test EC50 0.0059 CPA (1996)

Cereclor S-52

(52 wt% Cl, 

C14-17)

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, 

C14-17)

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 wt% Cl, 

C14-17)

Earthworm

(Eisenia fetida)

OECD Guideline-Draft, 

2000; 28-day reproductive 

toxicity test

NOEC

LOEC

MATC

79

280

149

Thompson et al. 

(2001d)

Terrestrial Vertebrates

Commercial CP 

(58 wt% Cl, 

C10-12)

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 3 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 represents

species sensitivity distributions (following US EPA, 2012). The COCs derived for aquatic-, 

sediment- and terrestrial-dwelling organisms are explained below and summarized thereafter in 

Table 4. 

The most reliable and acceptable toxicity studies to aquatic organisms indicate that for MCCPs 

and LCCPs, the toxicity to aquatic organisms are from the CPA (1996) study for acute toxicity 

and the Thompson et al. (1997) study for chronic toxicity. 

• Acute COC: The 48-hour EC50 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.
14

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 used to assess hazard. 

Using methods in US EPA (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 yield 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 Chronic 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/kg dry wt soil from this study will be used. Using methods in US EPA 

(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 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 COC = 298 mg/kg dry wt.

• 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. 

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 US EPA (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.
15

Table 4: COCs for Environmental Toxicity of MCCPs and LCCPs

Compartment Test organism Endpoint Value Assessment 

factor COC

Surface water Water flea

EC50 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. sediment

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 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, absorption following inhalation or dermal 

exposure is expected to be limited; previous evaluations 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 general, 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 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 

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 

histopathological 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 data 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 concluded based on an 

evaluation of the mechanistic data that the thyroid effects observed in rats is of little relevance to 

chronic toxicity in humans.
16

There is no information on the carcinogenicity of MCCPs; however, carcinogenicity studies on a 

SCCP and a vLCCP 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 (C12, 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 (details in 

ECB, 2008 and in Appendix C). 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 vLCCP 

(C23, 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. 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/kgbw/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 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 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 with a possible 

LOAEL for internal hemorrhaging and deaths in pups 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, 

suggesting that 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 
17

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 development is from cage-side observations in pups through 

lactation day 21. In these cases, no dose-related differences were reported in F1 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-bw/day from the rat study 

described above; CXR, 2005) was used to assess occupational and non-occupational (i.e., general 

population) risk of MCCPs. 

4.2 LCCP HEALTH 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 shorter 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 were observed in rats 

exposed to a C20-30 v/LCCP with 43 wt% Cl at dose levels of 100 mg/kg-bw and above. For 

another v/LCCP with C20-26 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 C23 vLCCP with 43 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 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 expected to pose a carcinogenic hazard to humans. 

Based on the LOAEL (100 mg/kg-bw) of the liver effects in female rats of repeated dose studies 

Health Canada calculated a tolerable daily intake (TDI) of 71 μg/kg-bw/day with LCCPs. Using 

upper bounding intake estimates ranging from 0.007 μg/kg-bw/day for 60+ age group to 0.024 

μg/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. 

The National Research Council (NRC, 2000) reviewed the toxicological risks of selected flame 

retardants, including a vLCCP containing C24 with 70 wt% Cl. Based on the NOAEL of 900 
18

mg/kg-bw/day (liver toxicity), 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-bw/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/day from the 90-day and two-year studies

described above) was used to assess potential occupational and non-occupational (i.e., general 

population) risk of LCCPs. 

5 EXPOSURE INFORMATION

EPA/OPPT used the information in this section and our standard PMN approaches to estimate 

potential worker exposures from activities associated with manufacturing, processing and use of 

P-12-0453 and P-12-0433. 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 MCCPs and LCCPs, which are not normally available for 

PMNs.

5.1 ENVIRONMENTAL MONITORING

For this assessment, environmental monitoring data consisting of measured 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 distribution of MCCPs and LCCPs. Appendix D 

contains information and 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 difficulties 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 ten) 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 ECNI-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 were collected in China. Over time and across countries, industrial practices and effluent 

pre-treatment have varied. Some of the monitoring studies only published their final measured 

concentrations and did not include the details of the analytical techniques and sampling 

locations. Generally, EPA/OPPT used studies sponsored by the environmental agencies, but full 
19

documentation is lacking for even these studies. The industrial sectors studied by other countries 

also are present in the US, suggesting that conditions 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 limits, 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, 

certain 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 analytical 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 used 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 5 and 6 for this assessment. When a 

limit of detection (LOD) value was reported for non-detectable results5

, EPA/OPPT used onehalf 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 5 and 6 represented the best available 

monitoring information for MCCPs and LCCPs, 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.

 5 Examples would be “not detected” (ND), negligible, or with a “less than” qualifier.
20

Table 5: 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/kgb 6.51 × 101 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 × 101 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/kgb 9.70 × 103 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); 

Kemmlein et al. (2002); 

Muir (2010); Muir et al. 

(2003); Muir et al. (2000); 

Reth et al. (2005,2006); 

Tomy et al., (1999a); 

USEPA (1988)

Biota 

(terrestrial) 8 5.00 × 10-3 mg/kg ww 3.70 × 10-1 mg/kg ww Reth et al. (2006)

a

Petersen et al. (2006) reported results for two water samples; EPA/OPPT assumed these were 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 MCCPs 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 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.

4. dw – dry weight and ww – wet weight.

Table 6 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 
21

based on the following information: 1) P-12-0433 commercial products contain up to 20% of 

C17, an MCCP congener (see Section 1.2 Chemistry), 2) LCCP congener groups are expected to 

behave in a manner similar to MCCP congener groups with comparable wt% Cl when released to 

the environment and 3) MCCP and LCCP commercial products have similar uses (see Table 1) 

and hence may have similar releases at facilities that process and use these chemicals. 

Table 6: Summary of Measured Concentrations of LCCPs in Environmental 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 above represent total LCCP C18-20 and not individual LCCP isomers.

2. The “n” value represents the number of media-specific LCCP C18-20 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 release estimates for P-12-

0453 and P-12-0433, 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 INEOS, 

published OECD Emission Scenario Documents, EPA Generic Scenarios and EPA models for 

estimating environmental releases. Table 7 provides a general summary of the release 

assessment.

Table 7: Summary of Estimated Release to Water

PMN Chemical Name Manufacturing Processing (PROC) Use

P-12-0453 Alkanes, C14-17, chloro

(MCCPs; 40-60 wt% Cl) No, import only

PROC1: Formulation of 

MWFs (73.8% of PV)

USE1: Use of MWF 

(73.8% of PV)

PROC2: PVC compounding 

(16.8% of PV)

USE2: PVC converting 

(16.8% of PV)

PROC3: Formulation of 

adhesives and sealants 

(9.4% of PV)a

USE3: Use of 

adhesives and sealants 

(9.4% of PV)

P-12-0433 Alkanes, C18-20, chloro

(LCCPs; 40-55 wt% Cl) No, import only PROC1: Formulation of 

MWFs (100% of PV)

USE1: Use of MWFs 

(100% of PV)

a

Note, adhesives and sealants were assessed as a single process and use when evaluating potential environmental 

releases because of the comparable activities associated with each. 

Exposure pathways of interest for human health include drinking water, fish ingestion and air 

stack 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
22

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 applied an assumption of 90%

removal of MCCPs and LCCPs at the POTW. Water concentrations were estimated using EFAST V2.0’s probabilistic dilution model (PDM), which predicts downstream chemical 

concentrations from industrial discharges. These values were reported 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 aquatic invertebrates (see Table 4). Air stack emissions were 

estimated using generic scenarios, which assumed inhalation exposures occurring 100 meters

downwind of a facility. EPA/OPPT used these estimated values for calculating human health and 

environmental risks of P-12-0453 and P-12-0433.

The results of the E-FAST V2.0 modeling are provided in Appendix F. Table 8 presents the 

values used in the risk assessment. As explained in the footnotes in Table 8, the values represent

reasonable worst-case scenarios based on the processing and use scenarios presented in Table 7. 

However, these estimates require consideration of two important caveats: (1) limited 

environmental monitoring data are available and (2) MCCPs and LCCPs are expected to 

partition to particulates and sediment; however, E-FAST V2.0 models do not account for this 

partitioning.

Table 8: E-FAST Modeling Values Used for MCCPs 1

Scenario

Water Release - Human2 Air Release -

Human2

Water Release –

Aquatic Organisms3

Drinking Water

(mg/kg-bw/day)

Fish Ingestion

(mg/kg-bw/day)

Stack Air LADD

(mg/kg-bw/day)

Range of 

Concentrations 

(μg/L)

P-12-0453

PROC1: Formulation of 

MWFs 4.20 × 10-5 5.15 × 10-4 2.99 × 10-4 6.5-47 (240)

PROC2: PVC compounding 7.30 × 10-5 8.95 × 10-4 3.68 × 10-4 7-54 (278)

PROC3: Formulation of 

adhesives and sealants 1.30 × 10-4 1.59 × 10-3 6.50 × 10-4 7-150 (258)

USE1: Use of MWF 3.35 × 10-4 4.09 × 10-3 8.95 × 10-4 11-144 (1.30 × 103)

USE2: PVC converting 7.15 × 10-5 8.70 × 10-4 2.18 × 10-4 1-27 (148)

USE3: Use of adhesives and 

sealants 2.00 × 10-5 2.45 × 10-4 2.82 × 10-5 2-20 (210)

P-12-0433

PROC1: Formulation of 

MWFs 1.04 × 10-6 4.47 × 10-5 N/A 5-36 (184)

USE1: Use of MWFs 2.38 × 10-4 1.02 × 10-3 6.30 × 10-4 7-92 (829) 1 Taken from Appendix E. Values represent the highest concentrations/estimated doses (reported as the lifetime 

average daily dose, or LADD) for chronic (i.e., repeated exposure scenarios) for human health. 2 Estimated exposure values were corrected for absorption by the oral (50% absorption) and inhalation (50% 

absorption) routes of exposure (ECB 2005; EA 2009). 3 For PMNs, environmental risk was 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., 6.5 

μg/L) up to the central tendency for a low flow site (e.g., 47 μg/L) – this example was taken from the first row, 

last column, under P-12-0453. The central tendency was calculated using the harmonic mean flow. The value in 

parentheses represents the 7Q10 (e.g., 240 μg/L) value normally used to determine potential chronic risk to 

aquatic organisms.
23

5.3 EXPOSURE ESTIMATES

5.3.1 OCCUPATIONAL EXPOSURE ESTIMATES

EPA/OPPT calculated screening-level workplace exposure estimates with ChemSTEER ver. 2. 

Table 9 provides a summary of the exposure estimates used in this risk assessment for evaluating

worker exposures to P-12-0453 and P-12-0433. Detailed information is provided in Appendix E.

Table 9: Summary of Occupational Exposure Estimates Used.

Scenario1

Route of Exposure

Inhalation

(mg/day)

Dermal

(mg/day)

P-12-0453

PROC1: Formulation of MWFs N/A2 350 – 1,800

PROC2: PVC compounding 0.030 – 4.4 1,800

PROC3: Formulation of adhesives and sealants N/A2 530 – 1,800

USE1: Use of MWF 2.0 – 7.1 350 – 3,900

USE2: PVC converting 12 – 22 N/A2

USE3: Use of adhesives and sealants 23 530

P-12-0433

PROC1: Formulation of MWFs N/A2 350 – 1,800

USE1: Use of MWFs 2.0 – 7.1 350 – 3,900 1

The following represent the estimated number of sites, workers per scenario and exposure over the course of the 

stated number of days: P-12-0453 = PROC1: 59 sites, 472 workers and 84 days/year; PROC2: 8 sites, 192 

workers and 126 days/year; PROC3: 3 sites, 66 workers and 200 days/year; USE1: 207 sites, 9936 workers and 

247 days/year; USE2: 8 sites, 384 workers and 250 days/year; USE3: 58 sites, 2784 workers and 250 days/year; 

P-12-0433 = PROC1: 3 sites, 24 workers and 27 days/year; USE1: 4 sites, 192 workers and 247 days/year. 2

Not applicable, the use category does not result in exposures that are relevant for this route.

5.3.2 CONSUMER EXPOSURE ESTIMATES

INEOS did not identify consumer uses in its PMN applications for P-12-0453 and P-12-0433; 

therefore, EPA/OPPT did not perform an assessment for these types of exposures.
24

6 RISK ASSESSMENT

6.1 ENVIRONMENTAL 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-12-0453 and P-12-0433, 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 potential risks because the modeled exposure values 

were generated using exposure scenarios that are representative of the types of uses and releases 

that may occur with P-12-0453 and P-12-0433. 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 used 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 5 and 6) or estimated 

(see Table 8) 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 4). 

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

P-12-0453 and P-12-0433 may present a risk to aquatic- and sediment-dwelling organisms. 

6.1.1 Risk Estimates Using Environmental Monitoring Concentrations

The RQs shown in Table 10 suggest that measured concentrations of MCCPs and LCCPs in 

water and sediment may present a risk of acute and chronic injury 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 minimum 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 these 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 values as the sole source 
25

from which to inform potential environmental concentrations and risks that may result from the 

specific uses and releases associated with P-12-0453 and P-12-0433. 

Table 10: Risk Quotients Calculated for MCCPs from Environmental Monitoring Data for 

Surface Water, Sediment and the Terrestrial Environment.

Environmental Concentration Effect Level 

(i.e., COC) RQs

Acute Risk

Aquatic Species < 2.50×10-10 to 1.49×10-3 mg/L 0.001 mg/L < 2.50×10-7 to 1.491

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 dw2 Not calculated

1

Bolded values represent those that may present an unreasonable 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 as a potential means for calculating RQs 

for this compartment, once relevant data are available. 

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., C14-17). 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 congener 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 Petersen 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-

0453 and P-12-0433. 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) and 

Wang et al. (2013) reported measured levels of MCCPs in soil, the samples were collected from 
26

sites in remote alpine locations or industrialized areas, respectively. These data are relevant for 

assessing airborne deposition of MCCPs/LCCPs; however, the reported 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-12-0453 and P-12-0433. Specifically, these data were 

used as supporting information to inform the relevant pathways for estimating 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 a key basis for evaluating the potential risks of P-12-

0453 and P-12-0433 is presented in the following section. 

6.1.2 Risk Estimates Using Modeled Exposures

The RQs shown in Table 11 suggest that the intended processes and uses for P-12-0453 and P12-0433 are expected to result in releases to surface water at concentrations that may present a 

risk of injury to aquatic organisms. 

It is noteworthy that these estimated concentrations are within the range of measured surface 

water concentrations reported for MCCP congener groups (Table 5). Though there is uncertainty 

whether the form (i.e., dissolved or particle bound) of MCCP impacts the aquatic toxicity, the 

estimated values suggest that either form may exist. The median stream flow estimates are all 

below the reported water solubility for P-12-0453 and slightly above or below the water 

solubility reported for P-12-0433 (Table 1). Since the available aquatic 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 exceed the estimated water solubility of P-12-0453 and P-12-0433. 

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 pathway 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 determining that 

environmental releases of P-12-0453 and P-12-0433 may present a risk of injury to aquatic 

organisms. 
27

Table 11: Risk Assessment of Aquatic Organisms Using Modeled Exposures1

Scenario

Estimated Water Concentrations (μg/L)2

Median Stream RQs

Flow Scenario

Low Stream 

Flow Scenario

7Q10

Flow Scenario

P-12-0453

PROC1: Formulation of MWFs 6.5 47 240 6.5-240

PROC2: PVC compounding 7 54 278 7-278

PROC3: Formulation of 

adhesives and sealants 7 150 258 7-258

USE1: Use of MWF 11 144 1300 11-1300

USE2: PVC converting 1 27 148 1-148

USE3: Use of adhesives and 

sealants 2 20 210 2-210

P-12-0433

PROC1: Formulation of MWFs 5 36 184 5-184

USE1: Use of MWFs 7 92 829 7-829 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/tsca-screening-tools/e-fast-exposure-and-fate-assessment-screening-tool-2014-

documentation-manual

6.2 HUMAN 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 

1986). 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 (C14-17, 52 wt% Cl). In the second and third studies, NTP (1986) 

reported LOAELs of 100 mg/kg-bw/day based on granulomatous inflammation of the liver in 

female rats administered an LCCP congener (C23, 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 or 100 mg/kg-bw/day, EPA/OPPT performed routeto-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 previously reported estimated 

human exposure values for workers (Table 9) and the general population (Table 8). 

EPA/OPPT compared the MOEs to a benchmark value that consisted of a multiplicative 

composite of three possible uncertainty factors (UFs): intraspecies variability (UFH; default value 

= 10), interspecies variability (UFA; default value = 10), and LOAEL-to-NOAEL extrapolation 

uncertainty (UFL; default value =10). The UFH and UFA may each be subdivided to account for 
28

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 UFA to 1 (i.e., UFA = TK × TD = 1 × 3.16 ≈ 3).

EPA/OPPT interprets MOEs that were equal to or below a benchmark value (e.g., MOE < 1000

[UFH × UFA × UFL = 1000]) as an indication that the scenario(s) may present a risk of injury to 

human health, whereas MOEs that were above the benchmark 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 workers and the general population. 

6.2.1 Workers

EPA/OPPT performed route-to-route extrapolations to convert the oral NOAEL of 23 mg/kgbw/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., HEDINHAL-WORKER) using the 

following equation:

HEDINHAL-WORKER = NOAELORAL × (1 ÷ sRVRAT) × (ABSORAL-RAT ÷ ABSINHAL- HUMAN) × (sRVHUMAN ÷ wRV)

where, 

NOAELORAL = 23 or 100 mg/kg-bw/day

sRVRAT = rat standard respiratory volume for 8-hours = 0.38 m3

/kg bw

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

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

sRVHUMAN = human standard respiratory volume for 8-hours = 6.7 m3

wRV = worker respiratory volume for 8-hours = 10 m3

For the oral NOAEL of 23 mg/kg-bw/day and the oral LOAEL of 100 mg/kg-bw/day, the 

HEDINHAL-WORKER values equal 41 mg/m3 and 176 mg/m3

, respectively.

EPA/OPPT calculated the HED values for dermal exposures to workers (i.e., HEDDERM-WORKER) 

based on the following equation:

HEDDERM-WORKER = NOAELORAL × (ABSORAL-RAT ÷ ABSDERMAL- HUMAN) × (BWRAT ÷ BWHUMAN)

1/4

where,

NOAELORAL = 23 or 100 mg/kg-bw/day

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

ABSDERM-HUMAN = percent absorption by the dermal route in humans = 1%

BWRAT = rat bodyweight = 0.250 kg

BWHUMAN = human bodyweight = 71.8 kg

The resulting HEDDERM-WORKER values equal 4600 mg/kg-bw/day for MCCP congener groups 

and 20,000 mg/kg-bw/day for the LCCP congener.
29

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, respectively. These values consisted of the following individual 

UFs. A default UFH of 10 was applied to each benchmark value, due to the absence of 

experimental data to inform the TK and TD subfactors of this UF. A reduced UFA of 3 was 

applied to each benchmark value, which accounted for a TK subfactor of 1 after converting the 

effect levels to HEDs. The UFA of 3 accounted for the remaining uncertainty associated with TD 

variability. A default UFL 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 HEDINHAL-WORKER and HEDDERM-WORKER values for calculating the 

respective MOEs using the estimated exposure values presented in Table 9. As shown in Table 

13, the MOEs for P-12-0453 and P-12-0433 all exceeded the respective benchmark values, 

which indicate a finding of low risk to workers for the processes and uses evaluated in this 

assessment.

Table 13: Occupational MOEs for P-12-0453 and P-12-0433

Exposure 

Route

Exposure Scenario Margin of Exposure

P-12-0453 Benchmark MOE = 30

Inhalation

PROC1: Formulation of MWFs NA

PROC2: PVC compounding 186 - 27,333

PROC3: Formulation of adhesives and sealants NA

USE1: Use of MWF 115 - 410

USE2: PVC converting 37

USE3: Use of adhesives and sealants 36

Dermal

PROC1: Formulation of MWFs 18,349 - 94,366

PROC2: PVC compounding 18,349

PROC3: Formulation of adhesives and sealants 18,349 - 62,317

USE1: Use of MWF 8,469 - 94,366

USE2: PVC converting NA

USE3: Use of adhesives and sealants 62,317

P-12-0433 Benchmark MOE = 300

Inhalation

PROC1: Formulation of MWFs NA

USE1: Use of MWFs 496 - 1,760

Dermal PROC1: Formulation of MWFs 79,778 - 410,286

USE1: Use of MWFs 36,821 - 410,286

6.2.2 General Population

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., HEDORAL-GENPOP) using the following equation:
30

HEDORAL-GENPOP = NOAELORAL × (ABSORAL-RAT ÷ ABSORAL- HUMAN) × (BWRAT ÷ BWHUMAN)

1/4 × (5 days ÷ 7 days)a

where, 

NOAELORAL = 23 or 100 mg/kg-bw/day

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

ABSORAL-HUMAN = percent absorption by the oral route in humans = 50%

BWRAT = rat bodyweight = 0.250 kg

BWHUMAN = human bodyweight = 71.8 kg a

A duration-specific adjustment was only applied to the oral LOAEL of 100 mg/kgbw/day because the animals were gavaged five days per week. 

For assessing inhalation exposures to the general population, EPA/OPPT performed route-toroute 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 population (i.e., HEDINHAL-GENPOP) using the following 

equation:

HEDINHAL-HUMAN = NOAELORAL × (1 ÷ sRVRAT) × (ABSORAL-RAT ÷ ABSINHAL- HUMAN) × (5 days ÷ 7 days)a

where, 

NOAELORAL = 23 or 100 mg/kg-bw/day

sRVRAT = rat standard respiratory volume for 8-hours = 1.15 m3

/kg bw

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

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

A duration-specific adjustment was only applied to the oral LOAEL of 100 mg/kgbw/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 

HEDINHAL-GENPOP values equal 20 mg/m3 and 62 mg/m3

, 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 HEDORAL-GENPOP and HEDINHAL-GENPOP values for calculating the respective 

MOEs using the estimated exposure values presented in Table 8. As shown in Table 14, the 

MOEs for P-12-0453 and P-12-0433 all exceeded 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. 
31

Table 14: General Population MOEs for P-12-0453 and P-12-04331

Scenario

Water Release Air Release 

Drinking Water 

MOE

Fish Ingestion 

MOE

Stack Air 

MOE

P-12-0453 (Benchmark MOE = 30)

PROC1: Formulation of MWFs 2.2 × 106 1.8 × 105 3.1 × 105

PROC2: PVC compounding 1.3 × 106 1.0 × 105 2.5 × 105

PROC3: Formulation of 

adhesives and sealants 7.1 × 105 5.8 × 104 1.4 × 105

USE1: Use of MWF 2.8 × 105 2.2 × 104 1.0 × 105

USE2: PVC converting 1.3 × 106 1.1 × 105 4.2 × 105

USE3: Use of adhesives and 

sealants 4.6 × 106 3.8 × 105 3.3 × 106

P-12-0433 (Benchmark MOE = 300)

PROC1: Formulation of MWFs 2.8 × 108 6.4 × 106 N/A

USE1: Use of MWFs 1.2 × 106 2.8 × 105 9.8 × 104

1

Taken from Appendix E. Values represent the highest concentrations/estimated doses (reported as the lifetime 

average daily dose, or LADD) for chronic (i.e., repeated exposure scenarios) for human health.

7 CONCLUSIONS

Based on its assessment of the available surrogate hazard and exposure information on P-12-

0453 and P-12-0433, EPA/OPPT concludes the following pertaining to the manufacturing, 

processing and use of these PMN substances: 

1. Occupational Exposures: given the assumptions, data and scenarios evaluated in this 

assessment, there were low risks found for workers from either dermal or inhalation 

exposures.

2. General Population Exposures (from environmental releases): given the assumptions, data 

and scenarios evaluated in this assessment, there were low risks found to humans from 

environmental releases via exposure to either drinking water or fish ingestion.

3. Environmental Assessment:

a. Using estimated environmental concentrations, the PMN substances may present an 

unreasonable risk following acute and chronic exposures to aquatic organisms.

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

environment as supporting information, the PMN substances:

iii. Are expected to partition to sediment and may partition to soil through land 

application of biosolids and,

iv. May be released to the environment at levels at or above estimated concentrations of 

MCCP and LCCP congener groups that may present an unreasonable risk 

following acute and chronic exposures to aquatic organisms. 

4. PBT Assessment: The PMN substances may be very persistent and very bioaccumulative.
32

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van Ginkel, C. G. 2010a. Biodegradability of a C14-17 Medium Chain Chlorinated Paraffin 

(63.2% Cl W/W) in the Closed Bottle Test. AkzoNobel confidential report: T10008c. 

AkzoNobel Technology and Engineering, Arnhem, The Netherlands.

van Ginkel, C. G. 2010b. Biodegradability of C14-17 Chlorinated Paraffin (45.6% Cl) in the 

Closed Bottle Test. AkzoNobel confidential report: T10007c. AkzoNobel Technology 

and Engineering, Arnhem, The Netherlands.

van Ginkel, C. G. 2010c. Biodegradability of Medium-Chain Chlorinated Paraffin (51.7% Cl 

W/W) in the Closed Bottle Test. AkzoNobel confidential report: T10031c. AkzoNobel 

Technology and Engineering, Arnhem, The Netherlands.

van Ginkel, C. G. 2010d. Biodegradability of Polychlorinated Tetradecane (45%) in the Closed 

Bottle Test. AkzoNobel confidential report 2.397.140. AkzoNobel Technology and 

Engineering, Arnhem, The Netherlands.

van Ginkel, C. G. 2014a. Biodegradability of C15 Chlorinated N-Alkane, 51% Cl (Wt.) in the 

Closed Bottle Test (OECD TG 301). AkzoNobel confidential report: F 14024 CG; Study 

number: T13029c. AkzoNobel Technology and Engineering, Arnhem, The Netherlands.

van Ginkel, C. G. 2014b. Biodegradability of C15 Chlorinated N-Alkane, 51% Cl (Wt.) in the 

Closed Bottle Test (OECD TG 301d). AkzoNobel confidential report: F 14025 CG; Study 

Number T13030c. AkzoNobel Technology and Engineering, Arnhem, The Netherlands.

van Ginkel, C. G., and A. Louwerse. 2010a. Biodegradability of Chlorinated Tetradecane in 

Closed Bottle Tests Incoluated with Activated Sludge and River Water. AkzoNobel 

confidential report 2.427.188. AkzoNobel Technology and Engineering, Arnhem, The 

Netherlands.

van Ginkel, C. G., and A. Louwerse. 2010b. Evaluation of the Ultimate Biodegradability of 

Chlorinated Tetradecanes Using in Sequencing Batch Reactors. AkzoNobel confidential 

report: 2.427.189. AkzoNobel Technology and Engineering, Arnhem, The Netherlands.

Wang, Y., J. Li, Z. Cheng, Q. Li, X. Pan, R. Zhang, D. Liu, C. Luo, X. Liu, A. Katsoyiannis, and 

G. Zhang. 2013. Short- and Medium-Chain Chlorinated Paraffins in Air and Soil of 

Subtropical Terrestrial Environment in the Pearl River Delta, South China: Distribution, 

Composition, Atmospheric Deposition Fluxes, and Environmental Fate. Environmental 

Science and Technology, 47(6), 2679-2687.

Wiegand, W. 1989. Determination of the Mutagenic Effects of Chloroparaffin 40g. AM-89/08. 

Huls AG, Huls AG, Marl, Germany.

Willis, B., M. J. Crookes, J. Diment, and S. D. Dobson. 1994. Environmental Hazard 

Assessment: Chlorinated Paraffins. TD170.9, G7ts, no. 19. Department of the 
45

Environment, Toxic Substances Division of the Directorate for Air, Climate and Toxic 

Substances, Garston, Watson, UK.

Zeng, L. X., H. J. Li, T. Wang, Y. Gao, K. Xiao, Y. G. Du, Y. W. Wang, and G. B. Jiang. 2013. 

Behavior, Fate, and Mass Loading of Short Chain Chlorinated Paraffins in an Advanced 

Municipal Sewage Treatment Plant. Environmental Science & Technology, 47(2), 732-

740.
46

9 APPENDICES
47

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 C10 to C24 including an MCCP product, Hoechst CP52, with chain lengths 

from C12 to C18 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 radiation 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 T1/2 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 numbers 

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. The atmospheric half-life has been estimated at 1 -

2 days (EA, 2009; ECB, 2005), based on estimated 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 wt% Cl) 

calculated using EPI SuiteTM/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 × m3/mol for C14-17 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; 

Hilger et al., 2013). Wide spread soil contamination and occurrence in artic samples suggest that 

MCCPs behave similarly to other chlorinated persistent organic pollutants (POPs) with high 
48

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 wt% Cl 70 wt% Cl 

C14 1.0 4.4

C15 0.8 3.0

C16 0.8 3.0

C17 0.8 2.9

A-1-2 Biodegradation

EPA/OPPT reviewed studies from the open literature 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 modified test conditions, C14

41.3% by wt. Cl and a C14 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 chlorinated and longer 

carbon chain CPs (C14-17 51.7% by wt. Cl, C14 55% by wt. Cl, C14 60.2% by wt. Cl, and C14 -17

63.2% by wt. Cl) biodegraded over a range of 2-54% in 28 days to 4-57% at up to 60 days. 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 chlorination 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 in 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 (Coelhan, 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 longer 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 contains 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 in sludge treatment and remain in the 

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 
49

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 transported 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 partition 

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 concentrations 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 keep 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 to 

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 receiving wastewater 

biosolids, manufacturing and processing facilities, and electronic waste processing and recycling 

facilities are shown to have higher levels (Wang et al., 2013). MCCPs are expected to be stable 

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 T1/2 of at least months to years.

Table_Apx A-2: Review of MCCP and LCCP Biodegradation Studies.

Biodegradation Studies on MCCPs (C14-17) and LCCPs (C>18) 

Study 

Authors

Publication 

Date

MCCP/LCCP 

Chemicals 

Evaluated (i.e., Clength, wt% Cl)

Method Study 

Duration Noteworthy Results and Implications

MCCPs

Van Ginkel 2010d C14 45 wt% Cl Closed bottle 28 days

Approximately 64% degraded in 28 

days

Based on oxygen demand

Van Ginkel 2010b C14-17 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 2010c C14-17 51.7 wt% Cl Closed bottle 28 days Approximately 27% degradation in 28 

days and 57% after 60 days 
50

Biodegradation Studies on MCCPs (C14-17) and LCCPs (C>18) 

Study 

Authors

Publication 

Date

MCCP/LCCP 

Chemicals 

Evaluated (i.e., Clength, wt% Cl)

Method Study 

Duration Noteworthy Results and Implications

Based on oxygen demand

Van Ginkel 2010a C14-17 63.2 wt% Cl Closed bottle 28 days Approximately 5% degradation after 28 

days and 10% after 60 days.

Van Ginkel 

and 

Louwerse

2010a

C14 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

2010b C14 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 studies used modified test conditions to enhance or maximize biodegradation. Under these modified test 

conditions, C14 41.3 wt% Cl and a C14 45.5 wt% Cl substances are readily biodegradable. More highly chlorinated 

and longer carbon chain CPs (C14-17 51.7 wt% Cl, C14 55 wt% Cl, C14 60.2 wt% Cl, and C14-17 63.2 wt% Cl) 

biodegraded over a range of 2 – 54% in 28 days to 4 – 57% at up to 60 days. 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 chlorination can contribute to greater 

persistence under most environmental conditions.

Van Ginkel 2014b C15 51 wt% Cl Closed bottle 

(301D) 60 day 43% and 63% degradation at 28 and 60 days

Van Ginkel 2014a C15 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 C15 51 wt% Cl were found to be inherently degradable and possibly readily 

degradable in OECD 301 and 301D tests.

Madeley and 

Birtley 1980 C14-17 mixed 

product, 40 wt% Cl BOD test 25 days

Approximately 15.5% degradation as 

measured by theoretical BOD in nonacclimated samples and 22.5%

degradation in acclimated samples.1

Madeley and 

Birtley 1980 C14-17 mixed 

product, 45 wt% Cl BOD test 25 days

Approximately 10% degradation as 

measured by theoretical BOD in nonacclimated samples and 30% 

degradation with acclimated soil 

microbes added.1

Madeley and 

Birtley 1980 C14-17 mixed 

product, 52 wt% Cl BOD test 25 days

Approximately 4% degradation as 

measured by theoretical BOD in nonacclimated samples and 6% degradation 

with acclimated soil microbes added.1
51

Biodegradation Studies on MCCPs (C14-17) and LCCPs (C>18) 

Study 

Authors

Publication 

Date

MCCP/LCCP 

Chemicals 

Evaluated (i.e., Clength, wt% Cl)

Method Study 

Duration Noteworthy Results and Implications

Madeley and 

Birtley 1980 C14-17 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 C20-30 42% of > 50% between day 20 

and 25.

1

The ThOD (theoretical oxygen demand) was estimated (ThOD (g O2/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. 

LCCPs

Madeley and 

Birtley 1980 C20-30 mixed 

product, 42 wt% Cl BOD test 25 days

Approximately 7.5% degradation as 

measured by theoretical BOD in nonacclimated samples and 23%

degradation with acclimated soil 

microbes added.1

Madeley and 

Birtley 1980 C25 “chlorinated pentacosane”

14C on central 

carbon

8 weeks 

(mean)

11% of 14C- released as CO2. Nonacclimated 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 C20-30 42% of > 50% between day 20 

and 25.

1

The ThOD (theoretical oxygen demand) was estimated (ThOD (g O2/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 C20-30 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 > C20-30 mixed 

product, 70 wt% Cl BOD test 5 days

2% 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 > C20-30 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. 
52

Biodegradation Studies on MCCPs (C14-17) and LCCPs (C>18) 

Study 

Authors

Publication 

Date

MCCP/LCCP 

Chemicals 

Evaluated (i.e., Clength, wt% Cl)

Method Study 

Duration Noteworthy Results and Implications

Substances other than the chlorinated 

paraffin contributed to this total carbon 

content.

Conclusions: As described 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 

surfactant, 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 CO2 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 to draw definite conclusions 

as to the degradability of the chlorinated paraffins in these tests” (EA, 2009).

Hoechst AG 1976 and 

1977 C18-20, 35 wt% Cl BOD test 5 days 0.7% degradation

Hoechst AG 1976 and 

1977 C18-20, 44 wt% Cl BOD test 5 days < 1.2% degradation

Hoechst AG 1976 and 

1977 C18-20, 49 wt% Cl BOD test 5 days < 2.3% degradation

Hoechst AG 1976 and 

1977 C18-20, 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 that 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 nonacclimated sludge microbes so degradation may have been possible but had not yet occurred.

Omori et al. 1987 C24.5H44.5Cl6.5, 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 (HK3, HK-6, HK-8 and HK-10)

Omori et al. 1987 C24.5H41Cl10, 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 C24.5H30Cl21, 70 wt% 

Cl

Chloride 

release 48 hours

2.6% degradation using bacterial strain 

HK-3;

12% H15-4;

1.4% HK-6;

1.7% HK-8; 

15% using mixed bacterial culture (HK3, HK-6, HK-8 and HK-10)

Conclusions: Omori et 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 
53

Biodegradation Studies on MCCPs (C14-17) and LCCPs (C>18) 

Study 

Authors

Publication 

Date

MCCP/LCCP 

Chemicals 

Evaluated (i.e., Clength, wt% Cl)

Method Study 

Duration Noteworthy Results and Implications

on the products were reported. Loss of Cl suggests dechlorination can occur and that lower Cl content or shorter 

chain lengths may be produced.

Allpress and 

Gowland 1999 C18-20, 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 O2/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 data on bioaccumulation of MCCPs are somewhat limited and 

conclusions vary with type of CP mixture 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., 1986; 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 congeners with intermediate chain lengths and 

chlorination suggests that some may have BCFs or BAFs greater than 1000 or 5000 (EC, 2008b; 

ECB, 2008). This suggests that some congeners in MCCP product mixtures may be 

bioaccumulative or very bioaccumulative. In conclusion, some MCCP congeners present in the 

mixtures are both very persistent and very bioaccumulative.

Additional evidence for bioaccumulation of MCCPs is provided by Houde et al. (2008). Fieldderived log BAFs for MCCPs (C14-15), 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 

with 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 
54

specifically, large BMFs were observed for all MCCP chain lengths in Lake Ontario, and for C14

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-order 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. 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 are bioaccumulative (e.g., bleak and rainbow trout (32-2856) and BCF of 6920 for 

common mussel).
55

Table_Apx A-3: Review of MCCP and LCCP Bioaccumulation Studies

Bioaccumulation studies on MCCPs (C14-17) and LCCPs (> C18) 

MCCPs

Study Authors Publication 

Date

MCCP/LCCP 

Chemicals 

Evaluated (i.e., Clength,% Cl)

Method Study Duration Noteworthy Results and Implications

Houde et al. 2008 C14-15 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)

Thompson et al.; 

as summarized in 

ECB, 2005

2000 n-pentadecane-8-

14C, 

51% Cl mixed with a 

non-radio-labelled 

C14-17, 51% Cl 

chlorinated paraffin

Freshwater; flowthrough; 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 ug/L

BCF = 265 L/kg at 35 days when exposed at 4.9 ug/L

kinetic BCF = 1,087 L/kg at 35 days when exposed at 0.9 

ug/L

kinetic BCF = 349 L/kg at 35 days when exposed at 4.9 

ug/L
56

CPC (Madeley et 

al.); as 

summarized in 

ECB, 2005

1983 commercial product 

mixed with a npentadecane-8-

14C 

chlorinated to a 

similar degree

freshwater; flowthrough; rainbow trout; 

acetone solvent used

60 days Concentrations above water solubility; hence, water 

concentrations may be overestimated and BCF 

underestimated. 

Uncertainty as to whether 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 C14-17, 45% Cl freshwater; flowthrough; 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 in 

ECB, 2005

1983 Total MCCPs Bioconcentration 

factors. MCCPs 

concentrations were 

above the water 

solubility limit, using 

acetone as the cosolvent in the test 

solutions, and hence are 

not in compliance with 

OECD guideline 

requirements

No Information BCF = 32 – 2856 for common mussel, bleak and rainbow 

trout. May not have reached steady-state.

Fisk et al. 1999 Average formula: 

C14H23.3Cl6.7, 

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 14C16 35% Cl and 14C16 69% ClLake sediments were 

spiked and worms 

added after 18 and 32

No Information Kinetic BAF probably represents the

upper limit of the true bioaccumulation factor

14C16 35% Cl

14-day BSAFss = 0.7

Kinetic BSAF = 4.4

14C16 69% Cl 
57

14-day BSAFss = 0.2

Kinetic BSAF = 0.6

Bengtsson et al. 1979 C14-17, 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 in 

ECB, 2005

1983 commercial C14-17, 

52% Cl

seawater; flow-through; 

mussel

acetone solvent used

60 days BCFs = 2,182 L/kg (parent compound analysis) or 2,856 

L/kg (14C-measurements) when exposed to 0.22 mg/l

BCF = 339 L/kg (parent compound analysis) or 429 l/kg 

(

14C measurements)

when exposed to 3.8 mg/l.

Renberg et al. 1985 C16H30.7Cl3.3 (34% 

Cl) and C12H16Cl9.8

(68.5% Cl) mixture 

synthesized with 14C 

radiolabel

Flow through exposure 

to mussel (Mytilus 

edulis)

C16 - 0.13 and 5.0 μg/L

C16 - 28 day uptake

C12 - 21 day uptake 

followed by 28 day 

depuration

Steady state BCF about 7000 for C16 and 140,000 for C12 

based on 14C 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 C18–26 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/kg
58

Appendix B ECOTOXICITY STUDY SUMMARIES

B-1 MCCP ECOTOXICITY 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, C15H26Cl6) similarly to ASTM (1980). Bluegill sunfish (Lepomis 

macrochirus) and yellow perch (Perca flavescens) were exposed to the test substance in a flowthrough 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 trout, and yellow perch and 20 oC for channel catfish. 

Dilution water hardness was 44 mg CaCO3/L in the rainbow trout and channel catfish test system 

and 314 mg CaCO3/L for the bluegill sunfish and yellow perch test system. Reported effect 

levels are considered to be nominal with LC50 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 LC50 = 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 unspecified concentrations of Cereclor 

S52® (C14-17, 52% Cl), Chloroparaffin huls 40G (C15.5, 40% Cl), and Witaclor 50 (C14-17, 50% 

Cl) under static test conditions. Salinity was 7 ppt, pH was 7.8, temperature was 10 oC, and 

dissolve oxygen was considered by study 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 LC50 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 LC50 = 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 

two 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 
59

were not fed during the exposure period. Six groups of 15 fish were used for both the exposure 

and control solutions. No mortality or effect on behavior was seen in fish exposed to the 

medium-chain chlorinated paraffin during the 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 study 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 C14-17 chlorinated paraffin with 52% chlorination 

that contained 0.3% epoxy soya bean oil stabilizer as well as a small amount of radiolabelled npentadecane-8-

14C (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. Appearance 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 dissolved 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 CaCO3/L. At 48 hours, 0%, 45%, 90%, 

75%, 85%, 100%, and 100% immobilization was observed 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 EC50 = 0.0059 mg/L

(2) A 48-hour acute Daphnia magna toxicity study was conducted the University of Bremen, 

Department of Physical and Environmental Chemistry with CP 52 (C12-18, 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 CaCO3/L. 

Ambient laboratory air temperature 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 EC50 = 0.052 mg/L
60

(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 C14-17 chlorinated paraffin with 52% chlorination 

that was mixed with an equal weight of radiolabeled n-pentadecane-8-

14C (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 containing 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 and 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 CaCO3/L. 

Observed immobilization was limited to the highest test concentration (100% solution) with 55% 

immobilization. 

EPA/OPPT Conclusion

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 C14-17 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, 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 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 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 around 0.071-0.142 

mg/L). In experiments using the water soluble fraction from the 10,000 mg/L stock solution, an 

EC0 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 Steinhäuser (1994) as EC0 = 0.140 mg/L and EC25 = 0.339 mg/L, and it was noted that some 
61

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 EC50 of 0.037 mg/L and an EC0 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 (approximately EC0 ≥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 only 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 EC50 = 0.037 mg/L; 100 mg/L stock (Frank and Steinhäuser, 

1994)

5) The following study summary (Thompson and Gore, 1999) provided in the 2005 European 

Chemical Bureau Risk Assessment of MCCP was considered supportive of the aquatic 

invertebrate hazard determination. The acute toxicity of C14-17, 52% wt. Cl substance 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. magna 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 the study had a hardness of 220 mg/L as CaCO3 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 using the 

same test system at 20 oC over 48 hours, where complete immobilization was seen at the lowest 

test 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 EC50 (Gammarus pulex) > 1 mg/L

48-hr EC50 (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 supportive of the 

aquatic invertebrate 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. 

(1986), where a static method was employed using water of salinity 7‰ at a temperature of 20-

22 oC without aeration, probably using acetone as cosolvent. 

EPA/OPPT Conclusion
62

The results are considered supportive to address aquatic invertebrate acute 

toxicity. 

96-hour LC50 = 9 mg/L (C14-17, 45% wt Cl)

96-hour LC50 >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 (C12-18, 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. 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 CaCO3/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 C14-17 chlorinated paraffin with 52% chlorination that 

contained 0.3% epoxy soya bean oil stabilizer as well as a small amount of radiolabeled npentadecane-8-

14C (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 were 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 to 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 EC50 = >3.2 mg/L (nominal); 1.2 mg/L (mean measured).

96-hour EC50 = >3.2 mg/L (nominal); 1.2 mg/L (mean measured).

72-hr NOECb = 0.1 mg/L 

72-hr LOECb = 0.18 mg/L 

72-hr GMATC = 0.134 mg/L 
63

B-1-4 Chronic Fish Toxicity

(1) A 60-day fish toxicity study was conducted by Brixham Laboratories in 1983 with radiolabeled 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-

8332184 (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 were 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 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 during 

embryonic development” (U.S. 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 C14-17

chlorinated 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 concentrations 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 specified 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, swimming 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 considered 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.

(3) A 20-day Japanese medaka (Oryzias latipes) embryo toxicity study was conducted with the 

formulation C14H24.9Cl5.1, 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 
64

starting after fertilization 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 14C-C14H23.3Cl6.7, 55% Cl (composition: 0.2% C14H26Cl4 (42.3% Cl), 4.4% 

C14H25Cl5 (47.7% Cl), 34% C14H24Cl6 (52.6% Cl), 45% C14H23Cl7 (56.4% Cl), 14% C14H22Cl8

(59.9% Cl), and 1.9% C14H21Cl9 (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 exception 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 Assessment 

of MCCP was considered supportive, but did not characterize all fish life-cycle stages. Cooley et 

al. (2001) studied the toxicity of C14H24.9Cl5.1, 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 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 were 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 parameters 

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. 

EPA/OPPT Conclusion

This study is considered unacceptable to characterize chronic mortality in 

fish because it did not characterize life stages but instead characterized 

physiological effects. 
65

(6) Cooley et al. (2001) studied the toxicity of another medium chain chlorinated paraffin with a 

slightly different chemical composition and at slightly different concentration levels. The 

chemical formula was 14C-C14H23.3Cl6.7, 55% Cl (as described in Fisk et al., 1999) and was 

juvenile rainbow trout (Oncorhynchus mykiss) were exposed 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 were 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 parameters 

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 characterize 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 C14-17 chlorinated 

paraffin with 52% chlorination that contained 0.3% epoxy soya bean oil stabilizer as well as a 

small amount of radiolabeled n-pentadecane-8-

14C (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 concentrations 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 

selecting 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 static-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 concentrations 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 water control, 20% mortality was observed. Overall, dilution water control and solvent 

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 dissolved 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%, 
66

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 significant effect level. 

Thus, using a weight of evidence approach, the study is considered 

acceptable for this endpoint.

21-day LC50 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 radiolabeled 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-

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 dilution 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 concentrations 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 dissolved 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 considered 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 NOEC 

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 supportive 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 C14-17 chlorinated paraffin with 52% chlorination, which was tested 

as a water 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 
67

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 were 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 experiments 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 

average 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 rates 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 nominal 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 acceptability 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 considered supportive 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 

C14-17 chlorinated paraffin 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. The 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. 

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 
68

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) Thompson 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 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) that was mixed with a small amount of a 

radio labeled n-pentadecane-8-

14C, 51% wt. Cl substance (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% calcium 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 dry 

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 control), 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 NOEC = 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 described in Phipps et al. (1993) using a static test 

system. The substance used in the test 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) that was mixed with a small amount of a radio labelled npentadecane-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% 

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. 
69

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 termination 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 pvalue 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 the 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 available 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-chain chlorinated paraffin 

product (C14-17, 52.5% wt. Cl) mixed with a small amount of a radiolabeled chlorinated npentadecane-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 overnight 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 weight 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 exposure 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 CaCO3/L, and 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 endpoint, 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 in 

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 

acceptable.
70

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 ringnecked pheasants (Phasianus colchicus) were exposed by gavage to 0 (control) or 24,606 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 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/OPPT Conclusion

Acute LD50 > 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 LD50 > 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 weight gain was comparable to controls. Based on reported 

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 LD50 > 24,063 ppm
71

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 14Clabelled 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 tissue 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 dry 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. Nominal 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 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. 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 risk assessment. A 28-day earthworm reproductive toxicity test was 

conducted according to OECD guidelines. 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). 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 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-
72

7.09. 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. 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 days; 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 dry 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 vessels 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. Statistical 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 soil

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 C14-17, 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-labelled 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, evaporating 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, 158, 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 stability 

determined at the 50, 500, and 5,000 mg/kg dry wt. concentrations (corresponding 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 significant 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.
73

EPA/OPPT Conclusion

Since soil control and solvent control means were equal (two tailed TTest) 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

28-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 preparation 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 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) physical-chemical properties of MCCPs (water solubility values of 

approximately 30 μg/L and Log Kow 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 toxicity to aquatic 

organisms for acute endpoints are from the Thompson et al. 1996 study for aquatic invertebrates. 

Where the 48-hour EC50 value = 0.0059 mg/L. Using the methods described in the Sustainable 

Futures/P2 Manual (US EPA, 2012), the acute and chronic concentrations of concern (CoC) are 

derived as follows:

• Acute CoC: The 48-hour EC50 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 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

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 invertebrate GMATC value of 187 mg/kg dry wt sediment is used to 

assess hazard. Again, using methods in US EPA (2012): 
74

• 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. 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 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 Chronic Concern Concentration = 19 mg/kg dry wt sediment.

The most reliable and acceptable value for the toxicity to terrestrial 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): 

• 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 for 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.

• 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.

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; Chlorowax 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 used 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. Dilution water 

hardness ranged from 40 to 50 mg/L as CaCO3. Reported effect levels are considered to be 

nominal with LC50 values of >300 mg/L for bluegill sunfish and rainbow trout for all LCCPs 

tested; all values are above the limit of solubility.
75

EPA/OPPT Conclusions

Using a weight-of-evidence approach, these studies were considered 

acceptable to characterize the acute fish toxicity endpoint.

96-hr LC50 = NES (> 300 mg/L)

(2) Bengtsson et al. 1979

A 96-hour acute fish toxicity study as part of a bioaccumulation study was published by 

Bengtsson et al. (1979). Bleak (Alburnus alburnus) were exposed to a long-chain chlorinated 

paraffin, Witaclor 549 (C18-26, 49% Cl). The test was performed at 10 °C under semi-static 

testing conditions. A nominal concentration of 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 though the test 

duration was 14-days, no mortality occurred within 96-hours or 14-days, thus the 96-hour LC50

was > 0.125 mg/L. 

EPA/OPPT Conclusions

This study was considered supplemental to characterize the acute 

saltwater fish toxicity endpoint using a weight-of-evidence approach. 

96-hr LC50 = 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 C14-17 product 

with a 52 wt% Cl. 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, 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 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. 

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 the water soluble fraction from the 10,000 mg/L stock 

solution, an EC0 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 EC0 = 0.140 mg/L and EC25 = 0.339 mg/L, 
76

and it was noted that some 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 EC50 of 0.037 mg/L and an EC0 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 

(approximately EC0 ≥ 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 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 therefore 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-solvent.

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 study. However, the harpaticoid, Nitocra spinipes, is not a standard 

test species and the test concentrations greatly exceed the limit of 

solubility. The 96-hour LC50 values for both chlorinated paraffins were 

identified, but the reliability of the results cannot 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. 

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 (C17-24, 35% 

chlorination), CP 40 (C17-20, 44% chlorination) and Hordaflex LC50 (C17-20, 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 
77

(Bremen, Germany) was used without adjustments 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 CaCO3/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 C17 and C18

constituents which are considered to have LCCP-like properties. 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 EC50 (growth rate) = NES (> 0.250 mg/L; CP 30 and CP 40)

72-hr EC50 (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 as C18-20 chlorinated paraffin with 52% chlorination. Frank 

(1993) carried out a series of acute and longer-term studies with Daphnia magna using a 

commercial C18-20, 52% wt. Cl product. The tests were carried out using dilutions of the 

water-soluble fraction 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 determined by AOX (adsorbable organic halogen) 

analysis. The detection limit of the method used was around 10 μg Cl/L, which is 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 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 paraffin 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).
78

Long-term (21-day) reproduction studies were also performed using dilutions of the watersoluble 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 offspring 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 response 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 

dilution, 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 reported 

(estimated) water solubility of LCCPs. It was also noted in the U.K. 

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 Thompson (2001) in the statistical method. They determined that the 
79

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 significant 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 

questionable.

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 C18-20 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 CaCO3. 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 acetone. 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. 

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 for 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 
80

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 for 

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 C18–20, 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 treatment 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 C18–20, 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) Environmental 

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 controls (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 C18-20, 52 wt% chlorination 

treatment groups 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 Daphnia 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. 

21-day NOEC = 0.002 mg/L (reproduction Hoechst 56 Flüssig)
81

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 above 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 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 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) physical-chemical properties of LCCPs and MCCPs (water solubility values 

between 2 and 30 μg/L and Log Kow 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 sediment, 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 C14-17 which is 

defined as a MCCP material. These data are used in this assessment to fill data gaps for the C18-20

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 the 
82

absence of data for the LCCP materials 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 EC50 value = 0.0059 mg/L. Using the methods described in the Sustainable 

Futures/P2 Manual (US EPA, 2012), the acute and chronic concentrations of concern (CoC) are 

derived as follows:

• Acute CoC: The 48-hour EC50 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 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 data 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 

10 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 from 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 hazard. 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. 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. (MCCP and LCCP)

• 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 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 endpoint 

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 analog 

approach to assess hazard. 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): 
83

• 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 for 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. 

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, absorption 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 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 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 

thyroid 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 increase 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 is 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 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. 

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 chronic toxicity in humans.
84

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 genotoxicity data on MCCPs, may be used to inform the 

carcinogenic potential of MCCPs. When administered by gavage, a SCCP (C12, 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 

carcinogenicity studies of a LCCP (C23, 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 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 repeateddose toxicology 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 further 

information and/or testing or for risk reduction measures beyond those which are being applied 

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 is expected to be limited. An in vitro study using human skin showed that after 24 

hours, approximately 0.7% of a C15 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-

14C]-labeled C15 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 body 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 limited, 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 (C14-17, 51-60 wt% Cl); the LD50

was reported to be > 4,000 mg/kg bw (Birtley et al., 1980; cited in: IPCS, 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.
85

C-1-3 Irritation and Sensitization

No signs of skin irritation were seen with MCCPs (C14-17, 45 wt% Cl), and only slight erythema 

on the shaved skin was reported in one rabbit at 24 hours exposed to MCCPs (C14-17, 40 wt% Cl) 

(Chater, 1978; cited in: EURAR, 2008). A mild skin irritancy response was reported in one of 

nine unpublished skin irritation studies of MCCPs (C14-17, 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 (C14-17, 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 (C14-17

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. 

MCCPs caused an increase in liver weight in male rats at exposure levels of > 100 mg/kgbw/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 trace 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 study, 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, despite 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 by 

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 (C14-17 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 dilation 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:
86

EURAR, 2008). Though this effect is considered relevant to humans, the study suffers from a 

number of limitations, which preclude 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 (C14-17 52 wt% Cl) have been reported to cause minimal to mild adaptive 

histopathological changes in the thyroid (i.e., follicular cell hypertrophy 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 T4 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 dose 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 T3 in male rats, but not total T3 or 

free/total T4, 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 T4 depletion following 

increased excretion of this hormone. This depletion of plasma T4 results from the induction of 

hepatic UDPG-transferase, increased glucuronidation, and ultimately excretion of T4 following 

exposure to MCCPs. The pituitary responds to the decreased levels of T4 by releasing more TSH, 

which in turn leads to increased production of T4 by the thyroid. The continuous stimulation of 

the thyroid in response to the increased excretion of plasma T4 is predicted to ultimately give rise 

to hypertrophy and hyperplasia in this organ. Humans, unlike rodents, possess T4-globulin 

binding protein and are therefore less susceptible to plasma T4 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 these changes may be relevant for 

assessing potential adverse outcomes during reproduction and development, as discussed under 

section. 

C-1-5 Genotoxicity

MCCPs (C14-17 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 (C12, 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 relevance 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 
87

5,000 mg/kg-bw/day in carcinogenicity studies of a LCCP (C23, 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 (C12, 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. 
88

Table_Apx C-1: Summary of Results from 90-day Studies in Rats Administered MCCPs

Strain (sample size) Test substance and 

dose levels

Target 

organ Effect levels

F-344 

(10 rats/sex/group)1

C14-17, 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 animals

♂ at 222 mg/kg-bw/day, 82% ↑ in microsomal T4-UDPGA-glucuronyl transferase activity

♀ at 100, 300, and 300 mg/kg-bw/day, 30, 30, and 252% ↑ in microsomal T4-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 T3, respectively, but no effects on total T3 or on free/total 

T4 at any dose

♀ at > 242 mg/kg-bw/day, no effects on free/total T3 or T4

♀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

C14-17, 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 weights, 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 at 
89

Table_Apx C-1: Summary of Results from 90-day Studies in Rats Administered 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 angularity) and the epithelium (i.e., increased height, cytoplasmic vacuolation, and nuclear vesiculation)

F-344 

(15 rats/sex/group)3

C14-17, 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 animals

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-dependent trend towards ↑ severity 

Adrenal ♂ and ♀ at 625 mg/kg-bw/day, 25% ↑ in absolute weight values

Wistar-derived

(24 rats/sex/group)4

C14-17, 52 wt% Cl, 

containing epoxidized 

vegetable oil as a 

stabilizer

Dietary intake for ♂: 0, 

33, 167, or 333 mg/kgbw/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, respectively

♂ 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)
90

Table_Apx C-1: Summary of Results from 90-day Studies in Rats Administered MCCPs

Strain (sample size) Test substance and 

dose levels

Target 

organ Effect levels

Dietary intake for ♀: 0, 

32, 160, or 320 mg/kgbw/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

C14-17, 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 T3, 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, T3 significantly ↑ by day 91

♀ at 312 and 625 mg/kg-bw/day, total plasma T4 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 and 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).
91

C-1-7 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 the 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/kgbw/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 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 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 UK 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/kgbw/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 100 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. 

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 
92

maternal death during parturition due to low levels of vitamin K and related hemorrhaging, 

suggesting that 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 in 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 toxicity 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 F1 post-weaning appearance 

or cage-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, the 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 for 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 River COBS CD 

rats administered MCCPs (C14-17, 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 (C14-17, 52 wt% Cl) via gavage at doses of 0, 500, 2000, and 5000 mg/kgbw/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 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 or by vaginal smear for sperm. The day 
93

mating was confirmed 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 females 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 histopathological 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 only 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/kgbw/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 (C14-17, 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 observed 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 prenatal 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 (C14-17, 52 wt% Cl) via gavage on 

GD 6-27. Unmated 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 weighed 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 superovulation. 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 excised 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 
94

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.

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/kgbw/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 (C14-17, 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). F0

animals were exposed to test substance from 28 days prior to mating until sacrifice; F1 animals 

were treated from weaning until sacrifice, with additional potential exposures occurring in utero

and during lactation. All F0 males were sacrificed after the mating period. Following the 

premating period, each male was cohabited with two females for 10 days. Females 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 F0 parents and at 21 days of age for the 

F1 weanlings. The F0 and F1 animals were observed for clinical signs of toxicity, changes in 

general appearance and behavior, and mortality. In the F0 adults, body weights and food

consumption were measured weekly; in addition, body weights were measured in F0 females on 

GD 0, 7, 14, and 20; and on lactation days (LD) 0, 7, 14, and 21. Estrous cyclicity was 

determined in F0 females prior to mating, during mating, and prior to dosing. All F0 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 size) 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. F0 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 

F1 weanlings/sex/dose were sacrificed and necropsied. Five F1 males and ten F1 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 F1 pups, the surviving F1 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 F1 pups in the high-dose group survived until lactation day 21); and 

slight (11%, not statistically significant) decreases in pup survival, and labored breathing, 

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 F1 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 F1 post-weaning appearance, behavior, food consumption, or clinical or 
95

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 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/kgbw/day). The LOAEL for developmental toxicity was 1000 ppm (~62/74 mg/kg-bw/day) based 

on pup mortality due to hemorrhaging. 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 mg/kg-bw/day) MCCPs (C14-17, 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 pups 

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 (lower 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 from 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 lactating 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 (C14-17, 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. 

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, respectively) MCCPs (C14-17, 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 from 

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 
96

deaths were considered to be treatment-related as there was no indication 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 gestation 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 turn 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 

insufficient 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 crossfostering 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 conducted 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/day in females) MCCPs (C14-17, 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-dose 

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 condition, 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 weights 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 of 
97

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/kgbw/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 toxicity 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 C20-30 LCCP with 43 wt% Cl at dose levels of 100 mg/kg-bw and above. For 

another LCCP with C20-26 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 C23 LCCP with 43 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 tolerable daily intake 

(TDI) of 71 μg/kg-bw/day. Using upper bounding intake estimates ranging from 0.007 μg/kgbw/day for 60+ age group to 0.024 μg/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 environmental Protection Act.” 

The National Research Council (NRC, 2000) reviewed the toxicological risks of selected flame 

retardant, including a LCCP containing C24 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.
98

Oral (gavage) studies (IRDC, 1981; cited in: UK, 2009) showed that approximately 82-95% of a 

single dose of [8-

14C]-labeled C22-26 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 C22-26 chlorinated paraffin (70 wt% Cl), it was found that 61-88% of the administrated 

radioactivity was recovered 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: C20-30 , 41-50 

wt% Cl; C22-26, 42 wt% Cl, C23, 43 wt% Cl; C20-30, 61-70 wt% Cl; C24, 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: C19 ,44 wt% Cl; C22-26, 42 wt% Cl; 

C20-30, 41-50 wt% Cl; C20-30, 70 wt% Cl. No evidence of irritation was seen in three of the four 

LCCPs. For the C22-26, 42 wt% Cl product, erythema was observed in two of six animals tested; 

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 C22-26, 42 wt% Cl product. However, the 

criteria for classification as an eye irritant were not met (IUCLID, 2003; cited in: UK, 2009).

In a maximization test and a Buehler test using guinea pig, a C22-26, 42 wt% Cl product was 

negative (IUCLID, 2003; Bailey and Sheldon, 1998; cited in: UK, 2009). A C18-27, 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 C20-30, 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 and necrosis. Nephrosis was observed in the kidney of male rats and mineralization in 

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 with C22-26, 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 C23, 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 established 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.
99

C-2-5 Genotoxicity

Both LCCPs of C22-26, 43 wt% Cl and C23, 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 C20-30 with 43 wt% Cl or C22-26 with 70 wt% Cl did not induce significant 

increases of chromosomal or chromatid aberrations in bone marrow cells of rats (IRDC1983; 

cited in: Serrone et al., 1987). 

C-2-6 Carcinogenicity

The carcinogenicity of an LCCP (C23, 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 received 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 C20-30 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 reported. 

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 increased 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 adverse treatment-related effects 

were reported in offspring at doses up to 5000 mg/kg-bw/day. 

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 gavage 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 
100

plug. The day mating was confirmed 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 

females 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 histopathological 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. 

 

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/kgbw/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 slight 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; Chlorinated 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 received 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 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 excised 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.

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 
101

sacrificed during treatment; 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 increases 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 decrease 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 administered 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 environmentally 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 excised 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.

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 highest dose tested. 
102

Appendix D ENVIRONMENTAL MONITORING 

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 to 1.49 × 10-3 mg/L 

(Table_Apx D-1-1). Very little information is available on the specific sampling locations for 

many of the surface water measurements reported in Table_Apx D-1-1. Limited 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 approach, 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.

Measurements of dissolved (filtered) concentrations were generally non-detectable (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 

monitoring studies (Table_Apx D-1-1) 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 range 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 assumed 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 Table_Apx D-1-1, 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 presented 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 to 

1.49 × 10-3 mg/L; Table_Apx D-1-1) to evaluate risks of potential concern to aquatic organisms. 
103

Table_Apx D-1-1: Surface Water Concentrations of MCCPs, Sorted by Country

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 Lech 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

Notes:

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. All concentrations measured from impoundment lagoons and drainage ditches from the USEPA (1988) study have not been included as they 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 included as they are not considered as surface 

water concentrations
104

D-1-2 Sediment

MCCP sediment concentrations from marine and non-marine environments ranged from 5.00 × 

10-3 to 1.64 × 101mg/kg dw and from 2.0 × 10-3 to 6.51 × 101 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 × 

101mg/kg dw and 2.00 × 10-3 to 6.51 × 101 mg/kg dw, respectively) to evaluate risks of potential 

concern to sediment organisms (Table_Apx D-1-2). Where appropriate, reported values were 

converted to a common unit, as presented in the table. 
105

Table_Apx D-1-2: Sediment Concentrations of MCCPs, Sorted by Country

Media Country

Location

Comments 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)

--- 1.64×101 mg/kg dw Kemmlein et al. 

(2002)

Canada Hamilton 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)

North and Baltic Sea --- 2.20×10-2 mg/kg dw Hüttig and Oehme 

(2006)
106

Media Country

Location

Comments Converted 

Concentration

Common 

Units References

City, State or Province

--- 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)

--- 1.49×10

-1 mg/kg dw Hüttig and Oehme 

(2006)
107

Media Country

Location

Comments 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)

--- 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)

--- 1.80×10

-2 mg/kg dw Hüttig and Oehme 

(2005)
108

Media Country

Location

Comments Converted 

Concentration

Common 

Units References

City, State or Province

--- 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 Pribylova et al. 

(2006)

Sum 5.58 mg/kg dw Pribylova 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)

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)
109

Media Country

Location

Comments Converted 

Concentration

Common 

Units References

City, State or Province

--- 7.00×10

-2 mg/kg dw IPCS (1996)

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)

Maximum 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)

Maximum 3.8 mg/kg dw Chen et al. (2011)

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)
110

Media Country

Location

Comments Converted 

Concentration

Common 

Units References

City, State or Province

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)

--- 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)
111

Media Country

Location

Comments Converted 

Concentration

Common 

Units References

City, State or Province

--- 1.25×101 mg/kg dw Nicholls et al. 

(2001)

--- 1.83×101 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)

--- 1.35×101 mg/kg dw Nicholls et al. 

(2001)

East Midlands Region: Idle 

River

--- 1.62×101 mg/kg dw Nicholls et al. 

(2001)

--- 4.39×101 mg/kg dw Nicholls et al. 

(2001)

Northumberland Region: 

Skerne River

--- 1.80×101 mg/kg dw Nicholls et al. 

(2001)

--- 2.56×101 mg/kg dw Nicholls et al. 

(2001)

--- 5.84×101 mg/kg dw Nicholls et al. 

(2001)

East Anglia Region: Lark 

River

--- 3.22×101 mg/kg dw Nicholls et al. 

(2001)

--- 4.50×101 mg/kg dw Nicholls et al. 

(2001)

--- 6.04×101 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)

Reported as trace with range of 

1.5-5; used the average 3.25×10-3 mg/kg dw

USEPA (1988)
112

Media Country

Location

Comments Converted 

Concentration

Common 

Units References

City, State or Province

--- 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)

--- 2.10×101 mg/kg dw USEPA (1988)

--- 3.40×101 mg/kg dw USEPA (1988)

--- 5.00×101 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 weight
113

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 × 103 mg/kg dw in 

sludge and from 1.5 × 10-2 to 8.5 × 10-2 mg/kg 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 

studies for its consideration, EPA/OPPT used the following criteria: designation of specific 

MCCP chain length and the appropriate analytical methodology. Based these criteria, 

EPA/OPPT determined that the data reported by Stevens et al. (2003) were the most reliable for 

its use in this assessment: data are summarized below. 

Stevens et al. (2003) measured MCCP concentrations in sludge samples obtained from 14 

WWTPs in the UK. MCCP concentrations ranged from 3.00 × 101 to 9.70 × 103 mg/kg dw. The 

authors concluded that these very high concentrations were likely the result of releases from 

numerous and ongoing diffuse sources. EPA/OPPT did not use information from other published 

studies reporting measured CPs in sludge and soil because they did not distinguish the CPs 

measured (Nicholls et al., 2001). These studies reported total CP concentrations at much lower 

levels ranging from 3.00 × 10-5 to 2.3 mg/kg dw. 

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 × 103 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_Apx D-1-3).

Table_Apx D-1-3: Biosolid and Soil Concentrations of MCCPs 

Location Media

Concentration

Units References 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 × 101 9.70 × 103 mg/kg dw Stevens et al. (2003)

United States Sewage 

Sludge 5.00 × 10-5 5.00 × 10-5 mg/kg dw Pribylova 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_Apx D-1-4). 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 the specific CP congeners measured. 
114

As a result of EPA/OPPT’s evaluation, MCCPs were found in organisms across many different 

trophic levels indicating widespread environmental contamination (Table_Apx D-1-4). The data 

were insufficient for EPA/OPPT to draw conclusions about trends based on region, species, time, 

or other factors.

While EPA/OPPT determined the concentrations of MCCPs in aquatic and terrestrial biota range 

from ND to 2.63 mg/kg ww 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 Table_Apx D-1-4. Rather, it used the risk quotient (RQ) method as described in 

Section 6.
115

Table_Apx D-1-4: Biota Concentrations 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 Kemmlein 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 ND1 mg/kg ww EC (1993) ND1 mg/kg ww EC (1993)

Total ND1 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/Baltic Sea 

Region2 Fish ND3 mg/kg ww IVL (2009) 2.6×10-1 mg/kg ww Reth et al. (2005)

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, Lithuania, Norway, Poland, and Sweden. 3 The minimum MCCP concentration value for fish from the North Sea/Baltic Sea Region was non-detect. The detection limit = 2.5 × 10-4 mg/kg; 1⁄2 DL = 1.25 × 10-4 (IVL 2009)
116

D-2 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 by a CP manufacturer in 

Yarraville, Australia. LCCP (C18-20) concentrations in marine sediment ranged from 1.02 × 10-1

to 4.31 × 10-1 mg/kg dw (Table_Apx D-2-1), 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 water 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 4.31×10-1 mg/kg dw Kemmlein et 

al. (2002)

Notes:

1. Values provided in the table above represent total LCCP (C18-20) and not individual isomers.

2. dw – dry weight

Table_Apx D-2-2: 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 represent total LCCP (C18-20) and not individual 

isomers.

2. lw – lipid weight
117

Appendix E ENGINEERING (ChemSTEER) REPORTS ON P-12-

0-0433 and P-12-0453

(Used for both identifying potential releases to the environment and for estimated 

occupational exposures. SEE APPENDIX G FOR REFERENCE TO FULL REPORTS 

UNDER SEPARATE COVER)

P-12-0433

Scenario # Sites # Workers

PROC: Formulation of 

Metalworking Fluids

3 24

Exposure mg/day day/yr

Inhalation - negligible (VP < 0.001 torr)

Dermal - Liquid High End 1.8E+3 27

Dermal - Liquid High End 3.5E+2 27

Release kg/site/day day/yr

Water or Incineration or 

Landfill

High End 8.6E+0 27

Water or Incineration or 

Landfill

Conservative 5.7E+0 27

Air Suggested 

Estimate

5.7E-1 27

Incineration or Landfill Suggested 

Estimate

5.7E+0 27

Scenario # Sites # Workers

USE: Use of Metalworking Fluids 4 192

Exposure mg/day day/yr

Inhalation - Mist High End of Range / 

Typical

7.1E+0 / 2.0E+0 247

Dermal - Liquid High End 3.5E+2 247

Dermal - Mist / / 3.9E+3 247

Release kg/site/day day/yr

Water Output 2 2.5E+0 247

Water or Incineration or Landfill High End 1.2E+0 139

Water or Incineration or Landfill Output 2 8.2E+0 247

Water or Incineration or Landfill Output 2 1.0E+1 247

Air Output 2 1.1E+0 247
118

P-12-0453

Scenario # Sites # Workers

PROC2: PVC Compounding (16.8% of PV) 8 192

Exposure mg/day day/yr

Inhalation - Particulate High End of Range / 

Low End of Range

4.4E+0 / 3.0E-2 126

Dermal - Liquid High End 1.8E+3 126

Release kg/site/day day/yr

Water Output 2 4.3E-3 126

Water or Incineration Output 1 8.5E+0 126

Water or Incineration or Landfill Output 1 1.3E+1 126

Water or Incineration or Landfill Suggested Value 4.3E-2 126

Air Output 2 4.3E-3 126

Scenario # Sites # Workers

USE2: PVC Converting (16.8% of 

PV)

8 384

Exposure mg/day day/yr

Inhalation - Particulate OSHA PEL / 

EU Data from 

Submitter

2.2E+1 / 

1.2E+1

250

Dermal - 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

1.6E-1 250

Water or Incineration or 

Landfill

Suggested 

Estimate

2.1E-2 250

Water or Incineration or 

Landfill

Output 2 2.1E+0 250

Water or Incineration or 

Landfill

Conservative 4.3E+0 250

Water or Landfill Suggested 

Estimate

5.3E+0 250

Air Suggested 

Estimate

1.6E-1 250

Air Output 2 2.4E-2 250

Landfill Output 2 2.4E+0 250
119

Scenario # Sites # Workers

PROC1: Formulation of 

Metalworking Fluids (73.8% of PV)

59 472

Exposure mg/day day/yr

Inhalation - negligible (VP < 0.001 torr)

Dermal - Liquid High End 1.8E+3 84

Dermal - Liquid High End 3.5E+2 84

Release kg/site/day day/yr

Water or Incineration or 

Landfill

High End 1.1E+1 84

Water or Incineration or 

Landfill

Conservative 7.6E+0 84

Air Suggested 

Estimate

1.9E-1 84

Incineration or Landfill Suggested 

Estimate

7.6E+0 84

Scenario # Sites # Workers

USE1: Use of Metalworking Fluids (73.8% of PV) 207 9936

Exposure mg/day day/yr

Inhalation - Mist High End of Range / 

Typical

7.1E+0 / 2.0E+0 247

Dermal - Liquid High End 3.5E+2 247

Dermal - Mist / / 3.9E+3 247

Release kg/site/day day/yr

Water Output 2 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 Output 2 1.6E+1 247

Air Output 2 1.8E+0 247

Scenario # Sites # Workers

PROC3: Formulation of Adhesives 

and Sealants (9.4% of PV)

3 66

Exposure mg/day day/yr

Inhalation - negligible (VP < 0.001 torr)

Dermal - Liquid High End 1.8E+3 200

Dermal - Liquid High End 5.3E+2 200

Release kg/site/day day/yr

Water or Incineration or

Landfill

High End 1.2E+1 200

Water or Incineration or 

Landfill

Output 2 8.0E+0 200

Water or Incineration or 

Landfill

Output 2 2.0E+2 4
120

Scenario # Sites # Workers

USE3: Use of Adhesives and Sealants (9.4% of PV) 58 2784

Exposure mg/day day/yr

Inhalation - Mist What-If 2.3E+1 250

Dermal - Liquid High End 5.3E+2 250

Release kg/site/day day/yr

Water or Incineration or Landfill High End 1.9E+0 66

Water or Incineration or Landfill Conservative 3.3E-1 250

Air Output 2 1.2E+0 250

Landfill Output 2 1.1E+1 250
121

Appendix F EXPOSURE SCENARIO ESTIMATES

(E-FAST Model Run. SEE APPENDIX G FOR REFERENCE TO FULL REPORTS 

UNDER SEPARATE COVER)

INEOS: P12-0433 Exposure 1

Exposure 

Scenario1

Water Landfil

l

Stack Air Fugitive Air

Drinking Water Fish Ingestion

ADR LADD ADR LADD 7Q104

COC = 

1

PDM 

Days 

Exceede

d

LADD ADR LADD ADR LADD

Release 

activity(ies)

2; exposure 

calculation(

s)3

mg/kg/day mg/kg/da

y

mg/kg/da

y

mg/kg/da

y

μg/l # Days mg/kg/da

y

mg/kg/da

y

mg/kg/da

y

mg/kg/da

y

mg/kg/da

y

Proc: Max 

ADR, 

PDM, 

LADD

8.99E-03

2.08E056.49E028.93E05

184.28 27 1.12E-04 --- --- 1.43E02

3.39E05

USE: Max 

ADR, 

PDM1

3.90E-02 --- 0.17 --- 829.55 139 --- --- --- 2.76E02 ---

USE: 

PDM2 --- --- --- --- 784.09 247 --- --- --- --- ---

USE: Max 

LADD --- 4.75E04 --- 2.04E03 --- --- 9.71E04 --- 1.26E03 --- 5.99E04

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 Model (PDM). There may be one, two, or all three exposure calculations per exposure scenario. COC is 

the aquatic concentration of concern.

4

This column displays concentration values for the 7Q10 streamflow, which is defined as the average streamflow 

of the 7 consecutive days of lowest flow within a 10 year period.
122

INEOS: P12-0453 Exposure 1

Exposure 

Scenario1

Water Landfill Stack Air Fugitive Air

Drinking Water Fish Ingestion

ADR LADD ADR LADD 7Q104

COC 

= 1

PDM 

Days 

Exceeded

LADD ADR LADD ADR LADD

Release 

activity(ies)2; 

exposure 

calculation(s)3

mg/kg/day mg/kg/day mg/kg/day mg/kg/day μg/l # Days mg/kg/day mg/kg/day mg/kg/day mg/kg/day mg/kg/day

PROC1: Max 

ADR, PDM, 

LADD

1.17E-02 8.43E-05 0.24 1.03E-03 239.69 84 4.58E-04 3.31E-02 5.97E-04 4.78E-03

3.52E05

USE1: Max 

ADR, PDM1

6.08E-02 --- 0.74 --- 1291.67 218 --- 3.68E-02 --- 4.60E-02 ---

USE1: PDM2 --- --- --- --- 1246.21 247 --- --- --- --- ---

USE1: Max 

LADD --- 7.55E-04 --- 9.24E-03 --- --- 1.55E-03 --- 2.01E-03 --- 9.78E04

PROC2: Max 

ADR, PDM, 

LADD

1.35E02

1.46E-04 0.28 1.79E-03 277.67 126 3.42E-04 2.76E-02 7.35E-04 1.08E-04

1.19E06

USE2: Max 

ADR, PDM, 

LADD

7.24E03

1.43E-04 0.14 1.74E-03 148.14 248 7.35E-04 7.94E-03 4.35E-04 4.61E-03

1.01E04

PROC3: Max 

ADR, Max 

acute eco

0.14 --- 2.85 --- 2835.05 --- --- 0.27 --- --- ---

PROC3: PDM --- --- --- --- 257.73 200 --- --- --- --- ---

PROC3: Max 

LADD --- 2.59E-04 --- 3.17E-03 --- --- 1.00E-03 --- 1.30E-03 --- ---

USE3: Max 

ADR, PDM1

9.61E03 --- 0.10 --- 210.38 64 --- 2.76E-03 --- 2.94E-02 ---

USE3: PDM2 --- --- --- ---- 31.13 186 --- --- --- --- ---

USE3: Max 

LADD --- 4.00E-05 --- 4.89E-04 --- --- 6.16E-04 --- 5.64E-05 --- 6.60E04

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 

Model (PDM). There may be one, two, or all three exposure calculations per exposure scenario. COC is the aquatic 

concentration of concern.

4

This column displays concentration values for the 7Q10 streamflow, which is defined as the average streamflow of the 

7 consecutive days of lowest flow within a 10 year period.
123

Appendix G SUPPLEMENTAL INFORMATION

SUPPLEMENTAL DATA FOR APPENDIX E (ChemSteer Engineering Reports):

p120433.ceb – 23 page pdf file

p120453.ceb - 60 page pdf file

SUPPLEMENTAL DATA FOR APPENDIX F (E-FAST Exposure Reports):

P-12-0433.exp1_Draft Final_REVISED_022013 – 9-page pdf file

P-12-0453.exp1_Draft Final_102512 – 21-page pdf file