Canadian Environmental Protection Act, 1999

Follow-up Report on a PSL1 Assessment for Which 

 Data Were Insufficient to Conclude Whether the Substances 

 Were “Toxic” to the Environment and to the Human Health 

Chlorinated Paraffins 

August 2008 
1

TABLE OF CONTENTS 

LIST OF TABLES........................................................................................................................................ 3

LIST OF ACRONYMS AND ABBREVIATIONS .................................................................................... 4

PREFACE ..................................................................................................................................................... 7

SYNOPSIS .................................................................................................................................................... 8

1. INTRODUCTION............................................................................................................................. 12

2. IDENTITY AND PHYSICAL/CHEMICAL PROPERTIES........................................................ 14

2.1. IDENTITY.................................................................................................................................... 14

2.1.1 Composition of CP mixtures................................................................................................. 14

2.2. PHYSICAL/CHEMICAL PROPERTIES.............................................................................................. 14

3. ENTRY INTO THE ENVIRONMENT .......................................................................................... 15

3.1. PRODUCTION, IMPORTATION AND USE PATTERN......................................................................... 15

3.2. RELEASES TO THE ENVIRONMENT IN CANADA ........................................................................... 16

3.2.1 National Pollutant Release Inventory (NPRI) data .............................................................. 17

4. RISK ASSESSMENT OF ECOLOGICAL IMPACTS ................................................................. 17

4.1. ENVIRONMENTAL FATE .............................................................................................................. 17

4.1.1 SCCPs................................................................................................................................... 17

4.1.2 MCCPs/LCCPs..................................................................................................................... 17

4.2. PERSISTENCE AND BIOACCUMULATION POTENTIAL.................................................................... 18

4.2.1 SCCPs................................................................................................................................... 18

4.2.2 MCCPs ................................................................................................................................. 22

4.2.3 LCCPs................................................................................................................................... 25

4.3. ENVIRONMENTAL CONCENTRATIONS......................................................................................... 31

4.3.1 Atmospheric concentrations ................................................................................................. 32

4.3.2 Wastewater treatment effluents, sewage sludge and soils .................................................... 32

4.3.3 Surface waters ...................................................................................................................... 33

4.3.4 Sediments.............................................................................................................................. 33

4.3.5 Biota ..................................................................................................................................... 34

4.4. ENVIRONMENTAL EFFECTS......................................................................................................... 36

4.4.1 SCCPs................................................................................................................................... 36

4.4.2 MCCPs ................................................................................................................................. 37

4.4.3 LCCPs................................................................................................................................... 38

4.5. POTENTIAL TO CAUSE ECOLOGICAL HARM ................................................................................. 39

4.6. UNCERTAINTIES ON THE ECOLOGICAL RISK ASSESSMENT........................................................... 42

4.6.1 Exposure, Effects and Risk Quotient Calculations ............................................................... 43

4.6.2 Persistence and Bioaccumulation Status and Risk Implications .......................................... 43

5. HUMAN HEALTH RISK ASSESSMENT..................................................................................... 44

5.1. POPULATION EXPOSURE ............................................................................................................. 44

5.1.1 SCCPs................................................................................................................................... 47

5.1.2 MCCPs ................................................................................................................................. 50

5.1.3 LCCPs................................................................................................................................... 51

5.2. HAZARD CHARACTERIZATION AND DOSE–RESPONSE ANALYSES ................................................ 52

5.2.1 SCCPs................................................................................................................................... 53

5.2.2 MCCPs ................................................................................................................................. 55

5.2.3 LCCPs................................................................................................................................... 56

5.3. HUMAN HEALTH RISK CHARACTERIZATION................................................................................ 56

5.3.1 SCCPs................................................................................................................................... 56

5.3.2 MCCPs ................................................................................................................................. 61

5.3.3 LCCPs................................................................................................................................... 62
2

5.4. UNCERTAINTIES AND DEGREE OF CONFIDENCE IN HUMAN HEALTH RISK CHARACTERIZATION ... 62

6. CONCLUSIONS ............................................................................................................................... 63

6.1. SCCPS ....................................................................................................................................... 63

6.2. MCCPS ...................................................................................................................................... 63

6.3. LCCPS ....................................................................................................................................... 64

REFERENCES ........................................................................................................................................... 65
3

LIST OF TABLES 

TABLE 1. RANGE OF PHYSICAL PROPERTIES OF CPS CONGENERS .................................................................. 14

TABLE 2. SUMMARY OF PERSISTENCE AND BIOACCUMULATION INFORMATION ON SCCPS. .......................... 18

TABLE 3. SUMMARY OF PERSISTENCE AND BIOACCUMULATION INFORMATION ON MCCPS.......................... 22

TABLE 4: SUMMARY OF PERSISTENCE AND BIOACCUMULATION INFORMATION ON C18-20 LCCPS. ............. 25

TABLE 5. SUMMARY OF PERSISTENCE AND BIOACCUMULATION INFORMATION ON C>20 LIQUID LCCPS........ 28

TABLE 6. SUMMARY OF PERSISTENCE AND BIOACCUMULATION INFORMATION ON C>20 SOLID LCCPS. ........ 30

TABLE 7. LIST OF ESTIMATED EXPOSURE VALUES (EEV), CRITICAL TOXICITY VALUES (CTV), ASSESSMENT 

FACTORS (AF), AND ESTIMATED NO EXPOSURE VALUES (ENEV) USED IN THE CALCULATION OF RISK 

QUOTIENTS (RQ) FOR SCCPS, MCCPS, C18-20 LIQUID LCCPS, C>20 LIQUID LCCPS AND C>20

SOLID LCCPS. ..................................................................................................................................... 40

TABLE 8. CONCENTRATIONS OF SHORT-CHAIN, MEDIUM-CHAIN AND LONG-CHAIN CHLORINATED PARAFFINS 

IN FOODSTUFFS .................................................................................................................................... 44

TABLE 9. UPPER-BOUNDING ESTIMATED AVERAGE DAILY INTAKE OF SHORT-CHAIN CHLORINATED PARAFFINS 

BY THE POPULATION OF CANADA ........................................................................................................ 48

TABLE 10. UPPER-BOUNDING ESTIMATED AVERAGE DAILY INTAKE OF MEDIUM-CHAIN CHLORINATED 

PARAFFINS BY THE POPULATION OF CANADA....................................................................................... 50

TABLE 11. UPPER-BOUNDING ESTIMATED AVERAGE DAILY INTAKE OF LONG-CHAIN CHLORINATED PARAFFINS 

BY THE POPULATION OF CANADA ........................................................................................................ 51
4

LIST OF ACRONYMS AND ABBREVIATIONS 

ACP Arctic contamination potential 

BAF bioaccumulation factor 

BCF bioconcentration factor 

BMF biomagnification factor 

BOD biochemical oxygen demand 

BSAF biota–sediment accumulation factor 

CB chlorinated biphenyl 

CEPA 1988 Canadian Environmental Protection Act 

CEPA 1999 Canadian Environmental Protection Act, 1999 

CoA coenzyme A 

CP chlorinated paraffin 

CPIA Chlorinated Paraffins Industry Association 

CTV Critical Toxicity Value 

CYP cytochrome P-450 

DDT dichlorodiphenyltrichloroethane 

DFO Department of Fisheries and Oceans 

DNA deoxyribonucleic acid 

EC50 median effective concentration 

ECNI electron capture negative ion 

ECNI-HRMS electron capture negative ion high-resolution mass spectrometry 

EEV Estimated Exposure Value 

ENEV Estimated No-Effects Value 

EPA Environmental Protection Agency 

EQC Equilibrium Criterion 

EU European Union 

foc fraction of organic carbon 

GC gas chromatography 

GLP Good Laboratory Practice 

HCB hexachlorobenzene 

HCH hexachlorocyclohexane 

HLC Henry’s law constant 

HPLC high-pressure liquid chromatography 

HR high resolution 

HRGC high-resolution gas chromatography 

HRMS high-resolution mass spectrometry 

IC50 median inhibitory concentration 

KAW air–water partition coefficient (or unitless Henry’s law constant) 

kg-bw kilogram body weight 

KOA octanol–air partition coefficient 

KOC organic carbon sorption coefficient (or organic carbon–water partition 

coefficient) 

Kow octanol/water partition coefficient 

LC50 median lethal concentration 

LCCP long-chain chlorinated paraffin 
5

LCCPs long-chain chlorinated paraffins 

liquid LCCPs C18–20 and C>20 liquid LCCPs 

LOAEL Lowest-Observed-Adverse-Effect Level 

LOEC Lowest-Observed-Effect Concentration 

LOEL Lowest-Observed-Effect Level 

LR low resolution 

LRMS low-resolution mass spectromety 

LT50 time to 50% lethality 

MCCP medium-chain chlorinated paraffin 

MCCPs medium-chain chlorinated paraffins 

MS mass spectrometry 

NCI negative ion chemical ionization 

NIMS negative ion mass spectrometry 

NOAEL No-Observed-Adverse-Effect Level 

NOEC No-Observed-Effect Concentration 

NOEL No-Observed-Effect Level 

NPRI National Pollutant Release Inventory 

NTP National Toxicology Program (U.S.A.) 

NWRI National Water Research Institute 

OCDD octachlorodibenzodioxin 

OECD Organisation for Economic Co-operation and Development 

PAH polycyclic aromatic hydrocarbon 

PCB polychlorinated biphenyl 

PCDD polychlorinated dibenzodioxin 

PCDF polychlorinated dibenzofuran 

PNEC Predicted No-Effect Concentration 

POC particulate organic carbon 

PPARα peroxisome proliferator activated receptor, α isoform 

PSL1 first Priority Substances List 

PUF polyurethane foam 

PVC polyvinyl chloride 

QSAR quantitative structure–activity relationship 

SCCP short-chain chlorinated paraffin 

SCCPs short-chain chlorinated paraffins 

T3 triiodothyronine 

T4 thyroxine 

TCDD tetrachlorodibenzodioxin 

TD05 Tumorigenic Dose05, the dose associated with a 5% increase in tumour 

incidence 

TDI Tolerable Daily Intake 

TGD (European) Technical Guidance Document 

TLC thin-layer chromatography 

TSH thyroid stimulating hormone 

TSMP Toxic Substances Management Policy 

UDP uridine diphosphate 

UDPG uridine diphosphoglucose 
6

UDPGGT uridine diphosphoglucose glucuronosyl transferase 

UDPGT uridine diphosphate glucuronosyl transferase 

UNECE United Nations Economic Commission for Europe 

VP vapour pressure 

wt. weight 

WWTP wastewater treatment plant 
7

PREFACE 

Following the assessment of chlorinated paraffins conducted under the first Priority 

Substances List (PSL1), available data were considered inadequate to evaluate whether 

medium and long chain chlorinated paraffins were considered to be “toxic” as defined 

under section 11 of the 1988 Canadian Environmental Protection Act (CEPA 1988). 

While information on the environmental effects of short-chain chlorinated paraffins was 

considered insufficient to conclude whether they were “toxic” under Paragraph 11(a) of 

CEPA 1988, this group of substances was considered “toxic” to human health under 

Paragraph 11(c) of CEPA 1988. In updating the assessments of medium and long chain 

chlorinated paraffins, included herein, more recent data on the effects of short-chain 

chlorinated paraffins on human health and on the environment were also examined. 

In this report, the impact of critical new data on the initial assessment under CEPA 1988 

is considered. These data are presented separately for the environmental and health 

effects, but cross-referenced, where appropriate. Information relevant to assessment of 

effects on the environment is presented initially, followed by information relevant to 

assessment of effects on human health. 
8

SYNOPSIS 

Chlorinated paraffins (CPs) are chlorinated derivatives of n-alkanes with carbon chain 

lengths from 10 to 38 carbon atoms, and with varying chlorine contents. CPs include 

short chain chlorinated paraffins (SCCPs) (CPs with 10–13 carbon atoms), medium chain 

chlorinated paraffins (MCCPs) (CPs with 14–17 carbon atoms) and long chain 

chlorinated paraffins (LCCPs) (CPs with ≥18 carbon atoms). 

CPs were included on the first Priority Substances List (PSL1) under the 1988 Canadian 

Environmental Protection Act (CEPA 1988) for assessment of potential risks to the 

environment and human health. With the data available at that time, it was concluded 

that SCCPs were “toxic” because they constitute or may constitute a danger in Canada to 

human life or health as defined under paragraph 11(c) of CEPA 1988. However, as 

outlined in the PSL1 assessment report released in 1993, relevant data identified before 

August 1992 were considered insufficient to conclude whether SCCPs, MCCPs or 

LCCPs could have immediate or long-term harmful effects on the environment as defined 

under paragraph 11(a) of CEPA 1988 and whether MCCPs or LCCPS could be 

considered “toxic” to human health as defined under paragraph 11(c) of CEPA 1988. 

Research to address data gaps relevant to the assessment of impacts of CPs on the 

environment was funded; an industry survey on the Canadian manufacture, import and 

uses of CPs was conducted for the years 2000 and 2001 through a Canada Gazette Notice 

issued pursuant to section 71 of the Canadian Environmental Protection Act 1999 (CEPA 

1999); and literature was also reviewed for new exposure and toxicological data on CPs 

on human and non-human organisms in Canada and elsewhere. 

Total reported annual usage of CPs in Canada (production + imports – exports) was 

approximately 3,000 tonnes in 2000 and 2001. MCCPs accounted for a large majority of 

CP usage in Canada, followed by smaller proportions of SCCPs and LCCPs. The major 

uses of CPs in Canada are in plastics, in lubricating additives and in metalworking. 

There was only one manufacturer of CPs in Canada, and only MCCPs and LCCPs were 

produced at this facility. In 2000, their production capacity was reported to be 8.5 

kilotonnes; however, there is no production in Canada at present. 

There are no known natural sources of CPs. The major sources of release of CPs 

(SCCPs, MCCPs and LCCPs) into the Canadian environment are likely the formulation 

and manufacturing of products containing CPs and use in metalworking fluids. The 

possible sources of releases to water from manufacturing include spills, facility washdown and drum rinsing/disposal. CPs in metalworking/metal cutting fluids may also be 

released to aquatic environments from drum disposal, carry-off and spent bath. These 

releases are collected in sewer systems and often ultimately end up in the effluents of 

sewage treatment plants. When released to the environment, CPs tend to partition 

primarily to sediment or soil. 

Environmental Assessment 
9

SCCPs have been detected in the following Canadian environmental media: Arctic air, 

sediments from remote northern lakes, sewage treatment plant effluents from southern 

Ontario, surface water, sediments, plankton, invertebrates and fish from Lake Ontario and 

marine mammals from the Canadian Arctic and the St. Lawrence River. SCCPs have also 

been detected in plankton, invertebrates and fish from Lake Michigan. MCCPs have 

been detected in effluent from a CPs manufacturing facility near Cornwall, Ontario, and 

also in sediments near this facility (which has since ceased operation), in fish from Lake 

Ontario and in beluga from the St. Lawrence River. Maximum Canadian concentrations 

of SCCPs and MCCPs were observed in aquatic biota and sediments from the St. 

Lawrence River and also in sediments and fish from southwestern Ontario. No data on 

environmental concentrations in Canada exist for LCCPs. They have been detected in 

marine sediments, crabs and mussels near a CPs manufacturing facility in Australia. 

Atmospheric half-lives for many CPs are estimated to be greater than 2 days. In addition, 

SCCPs have been detected in Arctic biota and lake sediments in the absence of 

significant sources of SCCPs in this region, which suggests that long-range atmospheric 

transport of SCCPs is occurring. SCCP and MCCP residues have been detected in 

Canadian lake sediments dating back over 25 years at concentrations suggesting that the 

half-lives of SCCPs and MCCPs in sediment are greater than 1 year. There are no data 

available for LCCPs in Canadian lake sediments; however, based on their 

physical/chemical properties, which are similar to those of MCCPs, LCCPs are expected 

to be persistent in sediments. Several biodegradation studies have also found that 

biodegradation is hindered by increasing carbon chain length. It is, therefore, concluded 

that SCCPs, MCCPs and LCCPs are persistent as defined in the Persistence and 

Bioaccumulation Regulations of CEPA 1999. 

Bioaccumulation factors (BAFs) of 9900–51200 wet weight in sculpin, smelt and trout 

from Lake Ontario indicate that SCCPs are bioaccumulating to a high degree in aquatic 

biota in Canada, this is supported by very high laboratory-derived bioconcentration 

factors (BCFs). Despite the lack of valid studies of BCFs, MCCPs have been found to 

have significant potential to bioaccumulate in aquatic food webs: field BAFs for MCCPs 

in some Lake Ontario fish were calculated to be approximately 5450 wet weight. In 

addition BAF values calculated using the Modified Gobas BAF Model are >5000 for all 

SCCP and MCCP congeners. 

While biomagnification factors (BMFs) are not considered in the bioaccumulation 

Regulations, they are supporting evidence for bioaccumulation when substantially above 

1. Both SCCPs and MCCPs were found to have biomagnification factors (BMFs) greater 

than one in various food webs. MCCPs also had BMFs greater than one. The liquid 

LCCP C18H30Cl7 had BMF values greater than one in rainbow trout in laboratory studies, 

and its half-life in rainbow trout was found to be similar to those of recalcitrant 

compounds that are known to accumulate in organisms and magnify in food chains. In 

addition, SCCPs, MCCPs and LCCPs have octanol–water partition coefficient (log KOW) 

values greater than five. Elevated concentrations of MCCPs have been measured in 

aquatic biota from the St. Lawrence estuary, the United States and Australia. While all of 

the available published BCF studies for LCCP have found values <5000, some elevated 
10

concentrations of LCCPs have been found in marine benthic organisms in Australia. In 

addition the Gobas BAF Model predicts that 44% of liquid C18-20 LCCP congeners have 

BAF≥5000. On the other hand, none of the C>20 LCCP congeners had modeled BAFs 

≥5000. Therefore, based on the weight of evidence, it is concluded that SCCPs, MCCPs, 

and liquid C18-20 LCCPs are bioaccumulative as defined in the Persistence and 

Bioaccumulation Regulations of CEPA 1999. However based on the limited information 

available (particularly BAF estimates), C>20 liquid and solid LCCPs are not 

bioaccumulative as defined under the Persistence and Bioaccumulation Regulations. 

The available toxicity data indicate that SCCPs , MCCPs and C18-20 LCCPs may be 

harmful to certain aquatic species (e.g., Daphnia magna) at low concentrations (e.g., 

chronic NOECs < 100 μg/L). 

SCCPs, MCCPs and C18-20 LCCPs are considered to be both highly persistent and 

bioaccumulative. In addition, there is evidence that SCCPs, MCCPs and C18-20 LCCPs 

are released into the Canadian environment and have the potential to cause harm to 

sensitive aquatic organisms at relatively low concentrations. Substances that are 

persistent remain in the environment for a long time, increasing the magnitude and 

duration of exposure. Releases of small amounts of bioaccumulative substances may lead 

to high internal concentrations in exposed organisms. Highly bioaccumulative and 

persistent substances are of special concern, since they may biomagnify in food webs, 

resulting in very high internal exposures, especially for top predators. 

Human Health Assessment 

For SCCPs, critical data relevant to both estimation of exposure of the general population 

in Canada and assessment of the weight of evidence for the mode of induction of specific 

tumours were identified following release of the PSL1 assessment and prior to February 

2001, although most of this information has been reported in incomplete published 

summary accounts or abstracts. These data suggest that several tumours observed in 

carcinogenicity bioassays in rats and mice exposed to SCCPs are induced by modes of 

action either not relevant to humans (kidney tumours in male rats) or for which humans 

are likely less sensitive (in rats, liver tumours related to peroxisome proliferation and 

thyroid tumours related to thyroid–pituitary disruption). Complete documentation of 

available studies and consideration in additional investigations of the reversibility of 

precursor lesions in the absence of continued exposure is lacking. However, reported 

data on mode of induction of tumours in addition to the weight of evidence that SCCPs 

are not DNA reactive are at least sufficient as a basis for consideration of a Tolerable 

Daily Intake (TDI) for non-cancer effects as protective for carcinogenicity for observed 

tumours. Upper-bounding estimates of daily intake of SCCPs approach or exceed the 

TDI for these compounds, which, on the basis of available information, is likely also 

protective for potential carcinogenicity. 

For MCCPs and LCCPs, critical data relevant to both estimation of exposure of the 

general population in Canada and assessment of effects were identified following release 

of the PSL1 assessment and prior to December 2000. Based upon these semi-quantitative 
11

data, upper-bounding estimates of daily intake for MCCPs and LCCPs are within the 

same order of magnitude of, or exceed, the TDIs for these substances. 

Conclusion 

Based on the information available, it is concluded that CPs containing up to twenty 

carbon atoms are entering, or may enter, the environment in quantities or concentrations 

or under conditions that have or may have an immediate or long-term harmful effect on 

the environment or its biological diversity and that all chlorinated paraffins constitute or 

may constitute a danger in Canada to human life or health. CPs containing up to twenty 

carbon atoms are persistent and bioaccumulative as defined in the Persistence and 

Bioaccumulation Regulations. 
12

1. INTRODUCTION 

Chlorinated paraffins (CPs) are chlorinated derivatives of n-alkanes with carbon chain 

lengths from 10 to 38 carbon atoms, and with varying chlorine contents. Commercial 

products, of which there are over 2000 (Serrone et al. 1987), are complex mixtures of 

homologues and isomers. CPs with carbon chains containing 10–13 carbon atoms (C10–13) 

are termed “short”, those with 14–17 carbon atoms (C14–17) are called “medium” and those 

having 18 or more carbon atoms (≥C18) are called “long”. This report addresses the shortchain chlorinated paraffins (SCCPs), the medium-chain chlorinated paraffins (MCCPs) 

and the long-chain chlorinated paraffins (LCCPs). 

CP waxes appeared on the first Priority Substances List (PSL1) of the 1988 Canadian 

Environmental Protection Act (CEPA 1988), published in the Canada Gazette, Part I, on 

February 11, 1989. An assessment was performed to determine whether CPs should be 

considered “toxic” as defined under CEPA 1988 and was completed in 1993 

(Government of Canada 1993a). As a result of this assessment, SCCPs were declared 

“toxic” under Paragraph 11(c) of CEPA 1988, because they were found to constitute a 

danger to human health. The conclusion of this assessment, published in the Canada 

Gazette, Part I, on January 22, 1994, also indicates that available data were considered 

insufficient to determine whether SCCPs, MCCPs or LCCPs could have immediate or 

long-term harmful effects on the environment as defined under paragraph 11(a) of CEPA 

1988 and whether MCCPs or LCCPS could be considered “toxic” to human health as 

defined under paragraph 11(c) of CEPA 1988. 

Subsequent to the completion of the PSL1 assessments, a revised CEPA, CEPA 1999, 

came into effect on March 31, 2000. Section 64 of CEPA 1999 has a definition of 

“toxic” that is similar to that in section 11 of CEPA 1988. CEPA 1999 places more 

emphasis on pollution prevention, and mandates the application of a weight of evidence 

approach and the precautionary principle when conducting and interpreting the results of 

risk assessments of existing substances. In addition, CEPA 1999 provides for special 

consideration of persistent and bioaccumulative substances. Substances that are shown to 

be both persistent and bioaccumulative, therefore, may be assessed using a more 

precautionary approach than is used for other substances. 

In 1997, a Scientific Justification document recommending that SCCPs be candidate 

substances for management under Track 1 (virtual elimination) of the Toxic Substances 

Management Policy (TSMP) (Government of Canada 1995) was published (Environment 

Canada 1997). The overall conclusion of the document stated: “On the basis of the 

information reviewed, it is concluded that short chain chlorinated paraffins are 

predominantly anthropogenic, persistent, bioaccumulative, and CEPA-toxic. Short chain 

chlorinated paraffins satisfy all four criteria outlined in the Toxic Substances 

Management Policy to identify substances for management under Track 1. Therefore, 

short chain chlorinated paraffins are proposed for management under Track 1 of the 

Policy.” During the public comment period on the Scientific Justification, the Chlorinated 

Paraffins Industry Association (CPIA) reviewed the information cited in the document 

proposing to list SCCPs as a Track 1 substance. They argued that the evidence did not 
13

constitute a scientifically credible basis to determine CEPA toxicity. Additionally, it was 

stated that the Scientific Justification document offered no persuasive evidence that 

SCCPs met the TSMP’s prescribed half-life criteria for persistence. In order to further 

examine the persistence of SCCPs and their potential to cause ecological harm, as well as 

to reassess MCCPs and LCCPs based on new information, scientists at the National 

Water Research Institute (NWRI) of Environment Canada and at the Freshwater Institute 

of the Department of Fisheries and Oceans (DFO) have generated new scientific 

information to address data gaps relevant to the assessment of impacts of CPs on the 

environment. 

To set further context for the update of the CPs assessment, an industry survey on the 

Canadian manufacture, import and uses of CPs was conducted for the years 2000 and 

2001 through a Canada Gazette Notice issued pursuant to section 71 of CEPA 1999 

(Environment Canada 2003a). Recent literature was also reviewed for new exposure and 

toxicological data on CPs on human and non-human organisms in Canada and elsewhere. 

This new information is considered in this assessment report. Data acquired prior to 

February 2001 and December 2000 were considered in the follow-up assessment of 

whether SCCPs and MCCPs/LCCPs, respectively, constitute or may constitute a danger 

in Canada to human life or health. Data obtained as of July 2007 were considered as part 

of the ecological follow-up assessment of SCCPs, MCCPs, and LCCPs. 

This assessment report was prepared under the authority of Section 68 of CEPA 1999. It 

was written by the staff of the Existing Substances Division of Environment Canada and 

Health Canada, as well as the National Water Research Institute of Environment Canada. 

The content of this report has been subjected to external review by Canadian and 

international experts selected from government and academia, and also to a 60-day public 

comment period. However, the conclusions presented in this report are those of 

Environment Canada and Health Canada and do not necessarily reflect the opinions of 

the external reviewers. 

This report represents a summary of more detailed information presented in a supporting 

document. For additional information the reader should consult this document. This 

assessment report and the associated environmental supporting document are available 

upon request by e-mail from existing.substances.existantes@ec.gc.ca. Information on 

assessments under CEPA 1999 is available at 

http:www.chemicalsubstanceschimiques.gc.ca. 
14

2. IDENTITY AND PHYSICAL/CHEMICAL PROPERTIES 

2.1. Identity 

As was the case for the PSL1 assessment, SCCPs, MCCPs and LCCPs are assessed 

separately in this report. 

2.1.1 Composition of CP mixtures 

SCCPs (C10–13), MCCPs (C14–17) and the lower chlorinated LCCPs are mixtures that are 

viscous, colourless or yellowish dense oils. C>20 highly chlorinated alkanes are waxy 

solids at ambient temperatures. The average chlorine content by weight is 30–52% for 

C18–20 liquid products, 40–54% for C>20 liquid products, and 70–72% for C>20 solid 

products. 

Impurities in commercial CPs are likely to be related to those present in the n-alkane 

feedstocks, which consist of a mixture of homologues. Furthermore, the n-alkanes may 

contain branched alkanes (usually <1%) and aromatics (<0.1%), which could be 

chlorinated. Commercial mixtures also contain stabilizers, which include epoxidized 

esters and soya bean oils, erythritol, thymol, urea, glycidyl ethers, acetonitriles and 

organic phosphates (European Commission, 2000; Schenker 1979; Houghton 1993). 

Various stabilizers are often added to commercial CP products in order to improve their 

thermal stability or light stability. 

2.2. Physical/chemical properties 

The large difference in chlorine content is primarily responsible for the large differences 

that are evident in measurements and estimates of physical/chemical properties. The 

approximate range of molecular weights for SCCPs is 320–500 (European Commission, 

2000), for MCCPs is 235–825 (U.K. Environment Agency 2003) and for LCCPs is 325–

1355 (U.K. Environment Agency 2001). 

Presented in Table 1 are ranges of physical properties for SCCPs, MCCPs and 3 

subclasses of LCCPs. 

Table 1. Range of physical properties of CPs congeners. 

CP Class Vapour 

pressurea

 (Pa)

Henry’s Law 

Constant 

(Pa·m3

/mol) 

Water 

solubility 

(μg/L) 

log KOW log KOA Log KOC Referenceb 

SCCPs 

2.8 x 10-7 –

0.028 

(48 – 71% Cl) 

0.68 – 17.7 

(48 – 56% Cl) 

6.4 – 2370 

(48 – 71% Cl) 

4.39 – 8.69 

(48 – 71% 

Cl) 

8.2 – 9.8 

(48 – 56% 

Cl) 

4.1 – 5.44 1-7, 14, 15 

MCCPs 

4.5 x 10-8 – 2.27 

x 10-3 

(42 – 58% Cl) 

0.014 – 51.3 

(37 – 56% Cl) 

9.6 x 10-2 – 50 

(37 – 56% Cl) 

5.47 – 8.21 

(32 – 68% 

Cl) 

8.81 – 12.96 

(32 – 68% 

Cl) 

5.0 – 6.23 4, 5, 6, 7, 8, 9, 

10, 11 
15

CP Class Vapour 

pressurea

 (Pa)

Henry’s Law 

Constant 

(Pa·m3

/mol) 

Water 

solubility 

(μg/L) 

log KOW log KOA Log KOC Referenceb 

C18-20 liquid

LCCPs 

2 × 10-8 – 5 × 

10-4 

(40 – 52% Cl) 

0.021 – 54.8 

(34 – 54% Cl) 

0.017 – 6.1 

(34 – 54% Cl) 

7.34 – 7.57 

(34 – 54% 

Cl) 

9.21 – 

12.12c

(34 – 54% 

Cl) 

- 4, 10, 11, 12, 

C>20 liquid 

LCCPs 

3 × 10-15 - 2.7 × 

10-3 

(40 – 54% Cl) 

0.003 

(50% Cl) 

1.6 × 10–6 – 6.6

(41.9 – 50% 

Cl) 

7.46 – 12.83 

(42 – 49% 

Cl) 

- - 4, 5, 6, 9, 10, 

12, 13 

C>20 solid 

LCCPs 

1 × 10-23 – 3 × 

10-14 

(70% Cl) 

3.6 × 10–7 – 5.6 

× 10–6 

(70 – 71.3% 

Cl) 

1.6 × 10–11 – 

5.9 

(70 – 71.3% 

Cl) 

- - - 4, 5, 12 

a Vapour pressure values not given at a consistent temperature. b References: 1. Drouillard et al. (1998a), measured data; 2. Drouillard et al. (1998b), estimated data; 3. Sijm and 

Sinnige (1995), measured data; 4. BUA (1992), estimated data; 5. Madeley et al. (1983a), measured data; 6. Renberg 

et al. (1980), thin-layer chromatography (TLC) – KOW correlation; 7. Fisk et al. (1998a), measured data; 8. U.K. 

Environment Agency (2003), measured data; 9. Campbell and McConnell (1980a), measured data; 10. BUA (1989), 

measured data; 11. Sijm and Sinnige (1995), estimated data; 12. U.K. Environment Agency (2001), estimated data; 

13. Howard et al. (1975), estimated data; 14. Drouillard (1996), measured and estimated; 15. Thompson et al. (1998), 

measured. c Octanol–air partition coefficients, estimated from ratio of KOW/HLC (unitless). 

3. ENTRY INTO THE ENVIRONMENT 

3.1. Production, importation and use pattern 

Canadian production and usage data for CPs were collected by means of a Notice, issued 

pursuant to section 71 of CEPA 1999, that was published in the Canada Gazette

(Environment Canada 2003a). CPs are no longer produced in Canada (Camford 

Information Services, 2001). Pioneer Chemicals Inc. (formerly ICI Canada), Cornwall, 

Ontario, was the only Canadian producer of CPs. However, this plant was recently sold 

to Dover Chemical Corporation and it is currently not producing chlorinated paraffins. 

This Cornwall plant previously produced MCCPs and LCCPs with a chlorine content of 

up to 56% under the trade name Cereclor (Camford Information Services 2001). The 

capacity for production was 5.0, 5.0, 8.5 and 8.5 kilotonnes in 1997, 1998, 1999 and 

2000, respectively; the corresponding imports to Canada in these years were 2.0, 2.0, 1.7 

and 1.8 kilotonnes, respectively. 

Total reported annual usage of CPs in Canada (production + imports – exports) was 

approximately 3,000 tonnes in 2000 and 2001 (Environment Canada 2003a). Whether 

the amount in use is the same at present is not known. North American demand for CPs 

fluctuates depending on the strength of the economy (Camford Information Services 

2001). 

Canadian use pattern data were obtained in two ways in the Notice issued pursuant to 

section 71 of CEPA 1999 (Environment Canada 2003a); distributors of CPs reported their 

sales volumes and intended usages for their customers, and users of CPs also reported on 
16

how they use CPs and the end uses for products that they formulate. There were some 

differences in reported usage volumes for CPs by distributors and users, but the uses 

generally were in agreement. 

Nearly all usage of SCCPs was reported to be in metalworking applications. Minor uses 

included use as a flame retardant in plastics and rubber. 

The majority of uses for MCCPs as reported by distributors were in plastics and as 

lubricating additives. Minor uses were as an additive in sealants and caulking, in rubber 

and paints, and as a flame retardant in plastics or rubber. 

The major uses of LCCPs are in lubricating additives, metalworking fluids and paints. 

Minor uses were in plastics and as flame retardants, engine oil, fabric adhesive and rock 

drilling fluid. Additional information on uses is available in the supporting document 

(Environment Canada 2008). 

3.2. Releases to the environment in Canada 

There is currently no evidence of any significant natural source of CPs (U.K. 

Environment Agency 2003). Anthropogenic releases of CPs into the environment may 

occur during production, storage, transportation, industrial and consumer usage of CPcontaining products, disposal and burning of waste, and land filling of products (Tomy et 

al. 1998a). 

The two major sources of release of SCCPs, MCCPs and LCCPs into the Canadian 

environment are likely use in metalworking applications and manufacturing of products 

containing these CPs. The possible sources of releases to water from manufacturing 

include spills, facility wash-down and storm water runoff. CPs in metalworking/metal 

cutting fluids may also be released into aquatic environments from drum disposal, carryoff and spent bath use (Government of Canada 1993a). These releases are collected in 

sewer systems and ultimately end up in the effluents of sewage treatment plants. 

Other releases could be associated with use of gear oil packages, fluids used in hard rock 

mining and equipment use in other types of mining, fluids and equipment used in oil and 

gas exploration, manufacture of seamless pipe, metalworking and operation of turbines 

on ships (CPIA 2002; Environment Canada 2003b). 

Landfilling is a major disposal route for polymeric products in Canada. CPs would be 

expected to remain stabilized in these products, with minor losses to washoff from 

percolating water. Leaching from landfill sites is likely to be negligible owing to strong 

binding of CPs to soils. Minor emissions of these products, which are effectively 

dissolved in polymers, could occur for centuries after disposal (IPCS 1996). 

Polymer-incorporated CPs could also be released during recycling of plastics, which may 

involve processes such as chopping, grinding and washing. If released as dust from these 
17

operations, the CPs would be adsorbed to particles because of high sorption and octanol–

air partition coefficients. 

Another significant source of release of CPs to the environment is from losses during the 

service life of products containing CP polymers (PVC, other plastics, paints, sealants, 

etc.) (European Commission, 2000; U.K. Environment Agency 2003). These releases are 

predicted to be mainly to urban/industrial soil and to wastewater. 

3.2.1 National Pollutant Release Inventory (NPRI) data 

Since 1999, on-site environmental releases of CPs (alkanes, C10-13, chloro; alkanes, C6-

18, chloro) in Canada must be reported to the National Pollutant Release Inventory 

(NPRI) by companies meeting the reporting criteria. Based on information collected by 

the NPRI, very small amounts of CPs are being released to the Canadian environment by 

companies that meet the NPRI reporting requirements. In 2002, small transfers of shortchain CPs for disposal to landfill (1.45 tonnes) and recycling by recovery of organics 

(1.94 tonnes) have been reported to the NPRI from only two companies, both located in 

Ontario. Less than 5 kg of releases and/or transfers of C6–18 CPs have been reported by a 

third company in Ontario. In 2001, the same three companies mentioned above reported 

similar quantities of releases/transfers of CPs to the NPRI. It should be noted, however, 

that CPs are likely to be released from sources other than the industrial sectors included 

in the NPRI, and releases to the Canadian environment could thus be considerably higher 

than those reported to this inventory. 

4. RISK ASSESSMENT OF ECOLOGICAL IMPACTS 

4.1. Environmental fate 

4.1.1 SCCPs 

Level III fugacity modelling of SCCPs has shown that they would achieve their highest 

concentrations in sediment and soil (Muir et al. 2001). 

4.1.2 MCCPs/LCCPs 

The environmental distribution of three MCCPs (C14-17) and a liquid C18 LCCPs was 

estimated using the Equilibrium Criterion (EQC) Level III fugacity model of Trent 

University’s Canadian Environmental Modelling Centre (Mackay et al. 1996). Level III 

represents a steady-state, non-equilibrium system comprised of soil, sediment, air and 

water compartments, with the chemical undergoing reactions or inputs and removal 

processes (advection, volatilization, deposition, photolysis, hydrolysis and 

biodegradation). Inputs of 100 kg/hour to soil, 1.6–6.4 kg/hour to air and 2.2–8.8 kg/hour 

to surface water were designed to reflect potential emissions of CPs mainly associated 

with landfills, land application of sewage sludge and consumer uses. Results from the 
18

Level III EQC model suggest that these CPs would achieve their highest concentrations 

in sediment and soil. Concentrations in water and air were extremely low for all 

compounds. The environmental residence time of the three C16–18 CPs were estimated to 

be greater than 500 days compared with 250 days for the C14 CP. However, these results 

should be viewed with caution because the degradations rates, used as input parameters 

for the CPs, were highly uncertain. Similar results were obtained using a Level III 

fugacity calculation with a C14–17 MCCPs (U.K. Environment Agency 2003). 

4.2. Persistence and bioaccumulation potential 

When evaluating persistence in this section, the focus is on sediment as results of 

fugacity modelling indicate that this is an important compartment for all CPs. Persistence 

in air is also evaluated, because of the potential for long-range transport in this medium. 

Although soil is potentially an important compartment for CPs, there are too few data 

available to permit meaningful evaluation of persistence in soil. 

4.2.1 SCCPs 

Table 2 summarizes persistence and bioaccumulation information for SCCPs in 

comparison with criteria of the Persistence and Bioaccumulation Regulations of CEPA 

1999 (Government of Canada 2000). 

Table 2. Summary of persistence and bioaccumulation information on SCCPs. 

Medium or 

parameter 

CEPA criteria 1 SCCPs Information 

Air t1/2 ≥ 2 days 

or

it is subject to atmospheric transport 

from its source to a remote area 

Estimated t1/2 of many SCCPs are ≥ 2 days 

SCCPs have been detected in air, sediment and biota in the 

Arctic in the absence of significant sources, indicating long 

range transport 

Sediment t1/2 ≥ 1 year Back calculation using concentrations from sediment cores 

shows half-life >1 year. Biodegradation test following 

OECD standard methods indicates half-lives >1 year in 

aerobic and anaerobic freshwater and marine sediments. 

Soil t1/2 ≥ 6 months Limited evidence for rapid biodegradation or removal 

following sludge applications 

BAF ≥5000 Field BAFs >5000 in sculpin, smelt and trout; BMF values 

approaching or >1; Modified Gobas Model predicts BAF 

>5000 for some SCCPs 

BCF ≥5000 BCFs>5000 in trout and mussels. 

Log KOW ≥5 4.39 – 8.69 (measured and modeled) 1

 Government of Canada 2000 

A- Persistence 

Air and Long-Range Transport 
19

Estimated atmospheric half-lives for SCCPs based on reaction with hydroxyl radicals 

range from 0.81 to 10.5 days, using the default atmospheric hydroxyl radical 

concentration of 1.5 × 106 molecules/cm3 during sunlight hours in AOPWIN (v. 1.86) 

computer program (Meylan and Howard, 1993; Atkinson 1986, 1987). Using a lower 

hydroxyl radical concentration of 5 × 105 molecules/cm3, which is generally used as a 

daily (24-hour) average in relatively unpolluted air in the EU, atmospheric half-lives 

ranged from 1.2 to 15.7 days. Tomy (1997) also estimated atmospheric half-lives of 

greater than 2 days for the major SCCPs detected in the Great Lakes and Arctic air and 

biota. 

SCCPs have vapour pressures (VPs) (2.8 × 10–7 to 0.028 Pa) and Henry’s Law Constants 

(HLCs) (0.68–17.7 Pa·m3

/mol for C10–12 congeners) that are in the range of VPs and 

HLCs for some persistent organic pollutants that are known to undergo long-range 

atmospheric transport under the 1979 UNECE Convention on Long Range 

Transboundary Air Pollution (e.g., hexachlorocyclohexane [lindane], heptachlor, mirex).1

The HLC values imply partitioning from water to air or from moist soils to air, depending 

on environmental conditions and prevailing concentrations in each compartment. 

SCCPs were detected in four air samples collected at Alert at the northern tip of 

Ellesmere Island in the high Arctic. Concentrations ranged from <1 to 8.5 pg/m3 in gasphase samples. Borgen et al. (2000) measured SCCPs in Arctic air samples taken at Mt. 

Zeppelin, Svalbard, Norway, in 1999. Concentrations ranging from 9.0 to 57 pg/m3 were 

detected. Borgen et al. (2002) found much higher SCCPs concentrations in air at Bear 

Island, a small isolated island between Svalbard and mainland Norway. Total SCCPs 

concentrations ranged from 1,800 to 10,600 pg/m3. SCCPs residues were found in the 

surficial sediments in three remote Arctic lakes including Yaya Lake, Hazen Lake and 

Lake DV-09. Concentrations ranged from 0.0016 to 0.0176 mg/kg dry wt. (Tomy et al. 

1998a; Stern and Evans 2003). 

SCCPs have been found at concentrations ranging from 0.095 to 0.626 mg/kg wet wt. in 

the blubber of marine mammals, including beluga (Delphinapterus leucas), ringed seal 

(Phoca hispida) and walrus (Odobenus rosmarus) from several locations in the Arctic 

(Tomy et al. 1998b;2000). Tomy et al. (2000) observed that the concentration profiles for 

the Arctic marine mammals show a predominance of the shorter carbon chain length 

congeners, i.e., the C10 and C11 formula groups. Drouillard et al. (1998a) showed that 

these congeners are the more volatile components of SCCPs mixtures, which show a 

trend of decreasing VPs with increasing carbon chain length and degree of chlorination. 

Reth et al. (2006) measured SCCPs in liver and muscle from seabirds (little auk and 

kittiwake) collected at Bear Island (European Arctic). Concentrations ranged from 0.005 

to 0.088 mg/kg wet weight.

Estimated atmospheric half-lives of many SCCPs are greater than 2 days for a large 

percentage (61% using hydroxyl radical concentration of 1.5 × 106

 molecules/cm3 and 

83% using hydroxyl radical concentration of 5 × 105

 molecules/cm3) of example 

 

1

 The VP of lindane is 4.3 × 10-3 Pa (IPCS 1991), the VP of heptachlor is 3.0 × 10-6 Pa (IPCS 1984a) and the VP of 

mirex is 2.3 × 10-9 Pa (IPCS 1984b). The HLCs of lindane and heptachlor are 0.13 and 0.02 Pa·m3

/mol, respectively. 
20

structures. Therefore, SCCPs meet the CEPA 1999 half-life criterion for persistence in 

air specified in the Persistence and Bioaccumulation Regulations (Government of Canada 

2000). The detection of the more volatile shorter carbon chain length congeners of 

SCCPs in Arctic biota and in Arctic lake sediments in the absence of significant sources 

of SCCPs in this region suggests that these residues are present due to long-range 

atmospheric transport. 

On the basis of the available information, it is concluded that estimated atmospheric halflives of SCCPs exceed the criterion of 2 days and SCCPs are subject to long-range 

atmospheric transport. Hence, SCCPs are persistent in air according to the criteria 

stipulated in the Persistence and Bioaccumulation Regulations of CEPA 1999 

(Government of Canada 2000). 

Sediments and Soils 

There is limited evidence for the biodegradation or removal of SCCPs from soil 

following sewage sludge application. Nicholls et al. (2001) did not detect 

SCCPs/MCCPs in farm soils amended with sludges containing mg/kg concentrations of 

CPs. However, worms living in these same soils did contain low mg/kg wet wt. levels of 

CPs. 

Using 25-day biochemical oxygen demand (BOD) tests, Madeley and Birtley (1980) 

found that SCCPs composed of 49% chlorine appeared to be rapidly and completely 

degraded by acclimatized micro-organisms after 25 days. However, no significant 

oxygen uptake was observed in tests using the highly chlorinated CPs, which included 

two SCCPs (60% and 70% chlorine). On the other hand, Fisk et al. (1998a) found that 

two 14C-labelled C12 chloro-n-alkanes (56% and 69% chlorine) were degraded at 12°C in 

aerobic sediments used for a study of the bioavailability of SCCPs to oligochaetes. Halflives in sediment were 12 ± 3.6 days and 30 ± 2.6 days for the 56% and 69% chlorine 

products, respectively. 

A study on the aerobic and anaerobic biodegradation of SCCPs in both freshwater and 

marine sediments was undertaken by Thompson and Noble (2007). Using 14C-labelled ndecane and n-tridecane 65% chlorine by weight products and basing their experiments on 

the OECD 308 Test Guideline (aerobic and anaerobic transformation in aquatic sediment 

systems), mineralization (as measured by carbon dioxide or methane production) over 98 

days was determined. The mean half-lives for mineralization for a C10-13, 65% chlorine 

by weight product, calculated as the average for the two individual products, were 

estimated to be 1630 days in freshwater sediments and 450 days in marine sediments 

under aerobic conditions. Little or no mineralization was noted in anaerobic sediments. It 

should be noted that these half-lives were calculated based on degradation observed after 

the 40-50 day lag phase, and that the half-lives were extrapolated beyond the available 

data. 

SCCPs residues were found in the surficial sediments of the following remote Arctic 

lakes (reported in mg/kg dry wt.): Yaya Lake (0.0016), Hazen Lake (0.0045) and Lake 
21

DV09 (0.0176). The profile from Lake DV09 generally follows the pattern of historical 

usage of SCCPs (Stern and Evans 2003). Concentration profiles of SCCPs in sediments 

from Lake Winnipeg (Manitoba), Fox Lake (Yukon Territory), the west basin of Lake 

Ontario (Ontario) and Lake DV09 (Devon Island, Nunavut) indicate that SCCPs residues 

were present in the 1940s (Muir et al. 1999a; Tomy et al. 1999). The highest 

concentration in Lake Ontario (800 ng/g dry wt.) was observed in the slice dated at 1971 

(Muir et al. 1999a). 

In the absence of information on loading for any of the years at any of these locations, it 

is not possible to calculate discrete half-life values from these data for comparison with 

the criteria for persistence in the Persistence and Bioaccumulation Regulations of CEPA 

1999 (Government of Canada 2000). However, the fact that SCCPs residues were 

detected in sediment cores dating back to the 1940s at these locations is evidence that 

SCCPs can persist for more than 50 years in subsurface anaerobic sediments. 

Environment Canada (2008) used first order decay equations in a back calculation 

method to determine that SCCPs have a half-life in sediments longer than 1 year. The 

equation used for these calculations are standard first order decay equations. 

Several government assessments and published reviews have concluded that only slow 

biodegradation in sediment may be expected to occur, even in the presence of adapted 

micro-organisms (Government of Canada 1993a,b; Tomy et al. 1998a; European 

Commission, 2000). On the basis of the available information, it is thus concluded that 

SCCPs are persistent in sediments according to the criterion stipulated in the Persistence 

and Bioaccumulation Regulations of CEPA 1999 (Government of Canada 2000). 

B- Bioaccumulation 

Bioaccumulation factors (BAFs) for SCCPs chain length groups in Lake Ontario 

plankton, alewife (Alosa pseudoharengus), slimy sculpin (Cottus cognatus), rainbow 

smelt (Osmerus mordax) and lake trout (Salvelinus namaycush) were determined based 

on a whole organism (wet weight) and filtered water concentrations using data from 

Houde et al. (2006). SCCPs were found in all components of the food chain and BAFs 

ranged from 9,900 to 51,200 (wet weight). SCCPs bioaccumulated to the greatest extent 

in fish, with the highest BAFs (51,200) in sculpin, smelt and trout. Assuming no 

metabolism, the Modified Gobas BAF model for fish estimated BAF values greater than 

5000 for all possible SCCPs (Arnot and Gobas 2003). 

Bioconcentration factors (BCFs) calculated from laboratory studies for SCCPs have been 

reviewed in Government of Canada (1993b) and were found to vary dramatically among 

different species. Relatively low BCF values have been determined in freshwater and 

marine algae (<1–7.6). BCF values of up to 7816 wet wt. have been measured in rainbow 

trout (Oncorhynchus mykiss) (Madeley and Maddock 1983a,b) and 5785–138 000 wet 

wt. in the common mussel (Mytilus edulis) (Madeley et al. 1983b, Madeley and 

Thompson 1983d, Renberg et al. 1986). 
22

Other evidence that SCCPs are bioaccumulative is as follows: 

• Reported log Kow values for SCCPs range from 4.39 – 8.69 (Table 1). 

• Lipid normalized biomagnification factors (BMFs) were also determined by 

Houde et al. (2006) for pelagic food webs in both Lakes Ontario and Michigan. 

BMFs ranged from 0.3 to 3.2. While biomagnification factors (BMF) are not a 

parameter considered in the Persistence and Bioaccumulation Regulations 

(Government of Canada 2000), BMFs are important supplemental information. If 

a substance has a BMF greater than one, it is more likely to have high BCF/BAF 

values. 

• Concentrations of SCCPs in fish collected around the Great Lakes between 1996 

and 2001 ranged from 0.0046 to 2.63 mg/kg wet weight (Muir et al. 2001; and 

Houde et al. 2006). SCCPs have also been detected in the blubber of belugas 

from the St. Lawrence River at an average concentration of 0.785 mg/kg wet wt. 

(Tomy et al. 1998b; 2000) and blubber of ringed seal from several Arctic 

locations. Concentrations in these mammals from the Arctic and the St. Lawrence 

River ranged from 0.095 to 0.626 mg/kg wet wt. (Jansson et al. 1993; Tomy et al. 

1998b; 2000). These relatively high concentrations suggest that SCCPs have the 

potential to bioaccumulate in aquatic organisms. 

• Tomy (1997) found that SCCPs (around 60–70% chlorine by weight) were 

present at a concentration of 0.011–0.017 mg/kg lipid (mean concentration 0.013 

mg/kg lipid) in human breast milk from Inuit women living on the Hudson Strait 

in northern Quebec, Canada. These findings are indicative of bioaccumulation 

through the food chain since food would be the major or only source of 

environmental exposure for the Inuit. 

On the basis of the available information, it is concluded that SCCPs are bioaccumulative 

substances according to the criteria stipulated in the Persistence and Bioaccumulation 

Regulations of CEPA 1999 (Government of Canada 2000). 

4.2.2 MCCPs 

Table 3 summarizes persistence and bioaccumulation information for MCCPs in 

comparison with criteria in the Persistence and Bioaccumulation Regulations of CEPA 

1999 (Government of Canada 2000). 

Table 3. Summary of persistence and bioaccumulation information on MCCPs. 

Medium or 

parameter 

CEPA criteria 1 MCCPs Information

Air t1/2 ≥ 2 days Estimated to be 2.7–7.1 days for vapor phase, but it should be 

noted that the extent of partitioning for MCCPs to air is low 

Degradation rate on airborne particles likely to be much slower 

Sediment t1/2 ≥ 1 year Back calculation using concentrations from sediment cores shows 

half-life >1 year 

Soil t1/2 ≥ 6 months Limited evidence for rapid biodegradation or removal following 

sludge applications 
23

Medium or 

parameter 

CEPA criteria 1 MCCPs Information

BAF ≥5000 Field BAFs for fish >5000 in Lake Ontario; high BMFs found in 

laboratory studies and a food web study in Lake Ontario and 

Lake Michigan; Modified Gobas Model predicts BAF>5000 for 

all congeners 

BCF ≥5000 Laboratory BCFs <5000; however, the BCF was probably 

underestimated due to CP concentrations exceeding solubility 

Log KOW ≥5 5.47–8.21 (measured and modeled) 1

 Government of Canada 2000 

A- Persistence 

Air 

Atmospheric half-lives for MCCPs were calculated using the Syracuse Research 

Corporation AOPWIN (v. 1.91) program using a hydroxyl radical concentration of 5 × 

105

 molecules/cm3. Half-lives for vapour phase MCCPs ranged from 2.7 to 7.1 days, with 

the longest half-lives for MCCPs with the highest chlorine contents and also with the 

shorter chain lengths. However, MCCPs have very low partitioning to air. 

MCCPs have estimated VP (4.5 × 10–8 to 2.27 × 10–3 to Pa at 20–25°C) and HLC (0.014 

– 51.3 Pa·m3/mol for C14-17 congeners) values that are in the range of VPs and HLCs for 

some persistent organic pollutants that are known to undergo long-range atmospheric 

transport under the 1979 UNECE Convention on Long Range Transboundary Air 

Pollution, such as lindane, heptachlor and mirex. 

On the basis of the available information, it is concluded that estimated atmospheric halflives of MCCPs exceed the criterion of 2 days and hence are persistent in air according to 

the criteria stipulated in the Persistence and Bioaccumulation Regulations of CEPA 1999 

(Government of Canada 2000). 

Sediment and Soil 

There is limited evidence for the biodegradation or removal of MCCPs from soil 

following sewage sludge application. Nicholls et al. (2001) did not detect 

SCCPs/MCCPs in farm soils amended with sludges containing mg/kg concentrations of 

CPs. However, worms living in these same soils did contain low mg/kg wet wt. levels of 

CPs. 

Concentrations of total MCCPs in a sediment core from Lake St. Francis, downstream of 

Cornwall, Ontario, ranged from 0.75 to 1.2 mg/kg dry wt, with the highest concentrations 

estimated to have been deposited in 1972 (Muir et al. 2002). Environment Canada (2008) 

used first order decay equations in a back calculation method to determine that MCCPs 

have a half-life in sediments longer than 1 year. The equation used for these calculations 

are standard first order decay equations. Moreover, the fact that MCCPs residues were 

detected in sediment cores dating back to the 1970s at these locations is evidence that 
24

SCCPs can persist for more than 30 years in subsurface anaerobic sediments. Persistence 

in sediment is particularly important as Level III fugacity calculations show that MCCPs 

are expected to partition primarily to sediment and soil. 

On the basis of the available information, it is concluded that MCCPs are persistent in 

sediments according to the criteria stipulated in the Persistence and Bioaccumulation 

Regulations of CEPA 1999 (Government of Canada 2000). 

B- Bioaccumulation 

Bioaccumulation factors (BAFs) for MCCPs chain length groups in Lake Ontario alewife 

(Alosa pseudoharengus), slimy sculpin (Cottus cognatus), rainbow smelt (Osmerus 

mordax) and lake trout (Salvelinus namaycush) were determined based on a whole 

organism (wet weight) and filtered water concentrations collected in 2001 using data in 

Houde et al. (2006). C14-15 MCCPs were found in all components of this food chain and 

BAFs ranged from 9.99 x 106

 to 7.15 x 108. In addition, bioaccumulation factors (BAF) 

for 21 MCCPs congeners using the Modified Gobas BAF Model (assuming no 

metabolism) were all above the bioaccumulative criteria (≥5000 BAF) (Arnot and Gobas 

2003). 

Most of the laboratory-based BCF studies conducted on aquatic organisms may 

underestimate the true BCF, because the studies were performed at MCCPs 

concentrations above the water solubility limit, using acetone as the co-solvent in the test 

solutions, and hence are not in compliance with OECD guideline requirements. 

Estimated BCF values for common mussel, bleak and rainbow trout (32-2856) are all 

below the BCF criterion of 5000 (Madeley et al. 1983b; Madeley and Maddock 1983a; 

Bengtsson et al. 1979; Madeley and Thompson 1983a), except for one common mussel 

study which reported a BCF of 6920 (Renberg et al. 1986). The only BCF study that did 

not use acetone reported BCFs values of 349 to 1087 for rainbow trout following the 

OECD test method 305 (Thompson et al. 2000). 

Other evidence that MCCPs are bioaccumulative is as follows: 

• Reported log Kow values for MCCPs range from 5.47 – 8.21 (Table 1). 

• Lipid normalized biomagnification factors (BMFs) were also determined by 

Houde et al. (2006) between Diporeia and sculpin in Lake Ontario and Lake 

Michigan. BMFs ranged from 1 to 15. Large BMFs were observed for these 

species for all chain lengths in Lake Ontario, and for C14 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 (Bruggeman et al. 1981) 

ranged from 0.4-5.0 (Fisk et al. 1996; 1998b;2000). While biomagnifications 

factors (BMF) are not a criterion considered in the Persistence and 

Bioaccumulation Regulations (Government of Canada 2000), BMFs are 
25

important supplemental information. If a substance has a BMF greater than 

one, it is more likely to have high BCF/BAF values. 

• Oligochaetes were found to have biota-sediment accumulation factors 

(BASFs) ranging from 0.6 to 4.4 (Fisk et al. 1998a). These BASFs, reflecting 

bioaccumulation from sediment at levels above that expected at equilibrium, 

imply significant food chain transfer. 

• Elevated levels of MCCPs were found in catfish from the Detroit River (0.904 

mg/kg wet wt.), and in crab and mussel (up to 38.7 mg/kg lipid wt.) located 

near a CPs manufacturing plant in Australia (Tomy and Stern 1999; 

Kemmlein et al. 2002). Kemmlein et al. (2002) stated: “Bioaccumulation is 

clearly evident, the mussel meat containing around double and crab meat 

around six times the amount of chloroparaffins found in the most 

contaminated sediment sample.” 

• MCCPs have been found in a breast milk sample (0.061 mg/kg lipid) from the 

United Kingdom (Thomas and Jones 2002), and C10–20 CPs were detected in 

liver, adipose and kidney tissues from human cadavers at up to 1.5 mg/kg wet 

wt. (Campbell and McConnell 1980a). These findings qualitatively indicate 

potential for bioaccumulation of MCCPs through the human food chain. 

On the basis of the available information, and in particular the field BAF estimates, it is 

concluded that MCCPs are bioaccumulative substances according to the criteria 

stipulated in the Persistence and Bioaccumulation Regulations of CEPA 1999 

(Government of Canada 2000). 

4.2.3 LCCPs 

4.2.3.1 C18–20 liquid LCCPs 

Table 4 summarizes persistence and bioaccumulation information for C18-20 liquid LCCPs 

in comparison with criteria in the Persistence and Bioaccumulation Regulations of CEPA 

1999 (Government of Canada 2000). 

Table 4: Summary of persistence and bioaccumulation information on C18-20 LCCPs. 

Medium or 

parameter CEPA criteria 1 C18-20 LCCPs Information 

Air t1/2 ≥ 2 days Estimated to be 2.4–10.5 days but it should be noted that 

the extent of partitioning to air for LCCPs is low 

Sediment t1/2 ≥ 1 year Unknown, but half-life likely > 1 year 

Soil t1/2 ≥ 6 months Unknown 

BAF ≥5000 

Laboratory diet studies suggest highly chlorinated C18 has 

high BMF from food; insufficient information on field 

BAFs; Modified Gobas Model finds nearly half of the C18-

20 congeners examined have BAF≥5000 (see section 

4.4.3.3.) 

BCF ≥5000 Laboratory BCFs <5000; BCF probably underestimated due 

to CP concentrations exceeding solubility 

Log KOW ≥5 7.34 – 7.57 (modeled) 
26

1 Government of Canada 2000 

A- Persistence 

Air 

Atmospheric half-lives for liquid LCCPs were calculated using the Syracuse Research 

Corporation AOPWIN (v. 1.91) program using a hydroxyl radical concentration of 5 × 

105 molecules/cm3. Half-lives for liquid LCCPs ranged from 2.4 to 10.5 days, with many 

example structures having half-lives greater than 2 days. However, C18-20 liquid LCCPs 

have very low partitioning to air. 

C18-20 liquid LCCPs have estimated VP (5 × 10–4 to 2 × 10–8 Pa at 25  ̊C) values that are in 

the range of VPs for some persistent organic pollutants that are known to undergo longrange atmospheric transport under the 1979 UNECE Convention on Long Range 

Transboundary Air Pollution, such as lindane, heptachlor and mirex. 

On the basis of the available information, it is concluded that estimated atmospheric halflives of C18-20 liquid LCCPs exceed the criterion of 2 days and hence are persistent in air 

according to the criteria stipulated in the Persistence and Bioaccumulation Regulations of 

CEPA 1999 (Government of Canada 2000). 

Sediment and Soil 

There is no empirical information available on the fate (i.e., half-lives) of LCCPs in soil 

or sediment with which to compare with the CEPA criteria. However, given that both 

SCCPs and MCCPs are expected to be persistent in sediment (half lives > 1 year), and 

that resistance to microbial degradation has been observed to generally increase with 

increases in carbon chain length (Allpress and Gowland 1999; Omori et al. 1987), it is 

likely that LCCPs also have half lives of more than 1 year in sediment. 

On the basis of the available information, it is concluded that C18–20 liquid LCCPs are 

persistent in sediments according to the criteria stipulated in the Persistence and 

Bioaccumulation Regulations of CEPA 1999 (Government of Canada 2000). 

B- Bioaccumulation 

Assuming no metabolism the Modified Gobas BAF Model predicts that 12 out of 27 

(44%) C18-20 congeners meet the bioaccumulation criteria of BAF ≥5000 (Arnot and 

Gobas 2003). As confirmed by personal communication with Frank Gobas (Simon Fraser 

University, Burnaby, BC), the model is applicable for LCCPs, as they are simple 

hydrophobic and persistent chemicals. 

On the other hand, BCF values for C18-26 liquid LCCPs were estimated by U.K. 

Environmental Agency (2001), using the data of Bengtsson et al. (1979) and were found 

to range from 8 to 16 in bleak; these values are below the BCF criterion of 5000 
27

(Government of Canada 2000). However, this study may underestimate the true BCF 

values, because the study was performed at LCCPs concentrations above the solubility 

limit for water and hence was not in compliance with OECD guidelines. As well, the 

study did not indicate if steady state was reached during the uptake phase of the test. 

Other evidence that LCCPs are bioaccumulative is as follows: 

• Reported log Kow values for C18–20 liquid LCCPs range from 7.34 – 7.57 

(Table 1). 

• Biomagnification factors (BMFs) were determined by Fisk et al (2000) in a 

dietary accumulation study involving juvenile rainbow trout exposed to 

C18H31Cl7. Lipid normalized BMFs ranged from 2.1 to 2.8. While 

biomagnifications factors (BMF) are not a criterion considered in the 

Persistence and Bioaccumulation Regulations (Government of Canada 2000), 

BMFs are important supplemental information. If a substance has a BMF 

greater than one, it is more likely to have high BCF/BAF values. 

• Fisk et al. (2000) also found that C18H31Cl7 has similar biotransformation halflives in rainbow trout compared to half-lives of recalcitrant organochlorines 

(Fisk et al. 1998c). This suggests limited biotransformation or metabolism. 

• Limited biotransformation of LCCPs was also observed during an 

uptake/elimination study with bleak. Bengtsson and Baumann-Ofstad (1982) 

found that a C18–26 LCCPs had a low uptake efficiency, but 50% of this 

compound remained in the fish tissues after a 316-day elimination period, 

which suggests that some of the LCCPs isomers in this formulation are 

bioaccumulative (Bengtsson and Baumann-Ofstad 1982). 

• Elevated levels of C18–29 LCCPs were found in crab and mussel (9.3 and 14.3 

mg/kg lipid wt., respectively) located near a CPs manufacturing plant in 

Australia (Kemmlein et al. 2002). Kemmlein et al. (2002) stated: 

“Bioaccumulation is clearly evident, the mussel meat containing around 

double and crab meat around six times the amount of chloroparaffins found in 

the most contaminated sediment sample.” However it is unclear if 

bioaccumulation of C18-20 or C>20 congeners was responsible for the elevated 

concentrations. 

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 

according to the criteria stipulated in the Persistence and Bioaccumulation Regulations of 

CEPA 1999 (Government of Canada 2000). 

4.2.3.2 C>20 liquid LCCPs 

Table 5 summarizes persistence and bioaccumulation information for C>20 liquid LCCPs 

in comparison with criteria in the CEPA 1999 Persistence and Bioaccumulation 

Regulations (Government of Canada 2000). 
28

Table 5. Summary of persistence and bioaccumulation information on C>20 liquid LCCPs. 

Medium or 

parameter 

CEPA criteria 1 C>20 liquid LCCPs Information 

Air t1/2 ≥ 2 days Estimated to be 1.8–8.4 days but it should be noted that the 

extent of partitioning to air for LCCPs is low 

Sediment t1/2 ≥ 1 year Unknown, but half life likely > 1 year 

Soil t1/2 ≥ 6 months Unknown 

BAF ≥5000 Insufficient information on field BAFs; Modified Gobas 

Model finds none of the C>20 congeners examined have 

BAF≥5000 

BCF ≥5000 Laboratory BCFs <5000; BCF probably underestimated due 

to CP concentrations exceeding solubility 

Log KOW ≥5 >7.46 – 12.83 (estimated) 1

 Government of Canada 2000 

A- Persistence 

Air 

Atmospheric half-lives for liquid LCCPs were calculated using the Syracuse Research 

Corporation AOPWIN (v. 1.91) program using a hydroxyl radical concentration of 5 × 

105

 molecules/cm3. Half-lives for liquid LCCPs ranged from 1.8 to 8.4 days, with many 

example structures having half-lives greater than 2 days. However, LCCPs have very 

low partitioning to air. 

C>20 liquid LCCPs have estimated VPs (5 × 10–5 to 3 × 10–15 Pa at 25  ̊C) that are in the 

range of VPs for some persistent organic pollutants that are known to undergo long-range 

atmospheric transport under the 1979 UNECE Convention on Long Range 

Transboundary Air Pollution, such as heptachlor and mirex. 

On the basis of the available information, it is concluded that estimated atmospheric halflives of C>20 liquid LCCPs exceed the criterion of 2 days and hence are persistent in air 

according to the criteria stipulated in the Persistence and Bioaccumulation Regulations of 

CEPA 1999 (Government of Canada 2000). 

Sediment and Soil 

There is no empirical information available on the fate (i.e., half-lives) of LCCPs in soil 

or sediment with which to compare with the CEPA criteria. However, given that both 

SCCPs and MCCPs are expected to be persistent in sediment (half lives > 1 year), and 

that resistance to microbial degradation has been observed to generally increase with 

increases in carbon chain length (Allpress and Gowland 1999; Omori et al. 1987), it is 

likely that LCCPs also have half lives of more than 1 year in sediment. 
29

On the basis of the available information, it is concluded that C>20 liquid LCCPs are 

persistent in sediments according to the criteria stipulated in the Persistence and 

Bioaccumulation Regulations of CEPA 1999 (Government of Canada 2000). 

B- Bioaccumulation 

Although C>20 liquid LCCPs may have some potential to bioaccumulate, the Modified 

Gobas BAF Model predicts that none of the C>20 congeners meet the bioaccumulation 

criteria of BAF ≥5000. Thus, these very high molecular weight LCCPs are not expected 

to be bioaccumulative. 

BCF values were found to range from 8-16 for C18-26 liquid LCCPs in bleak, and 18-1158 

for liquid C>20 LCCPs in rainbow trout and common mussel (Madeley and Maddock 

1983b; Bengtsson et al. 1979; Madeley and Thompson 1983b; U.K. Environment Agency 

2001). However, these values may underestimate the true BCF values, because the 

studies were performed at LCCPs concentrations above the solubility limit for water and 

hence were not in compliance with OECD guidelines. As well, the studies did not 

indicate if steady state was reached during the uptake phase of the tests. BCF values for 

these species were below the BCF criterion of 5000. 

On the other hand there is some evidence to suggest that C>20 LCCPs may be 

bioaccumulative: 

• Reported log Kow values for C>20 liquid LCCPs range from 7.46 – 12.83 

(Table 1). 

• Limited biotransformation of LCCPs was also observed during an 

uptake/elimination study with bleak. Bengtsson and Baumann-Ofstad (1982) 

found that a C18–26 LCCPs had a low uptake efficiency, but 50% of this 

compound remained in the fish tissues after a 316-day elimination period, 

which suggests that some of the LCCPs isomers in this formulation are 

bioaccumulative (Bengtsson and Baumann-Ofstad 1982). 

• Elevated levels of C18–29 LCCPs were found in crab and mussel (9.3 and 14.3 

mg/kg lipid wt., respectively) located near a CPs manufacturing plant in 

Australia (Kemmlein et al. 2002). Kemmlein et al. (2002) stated: 

“Bioaccumulation is clearly evident, the mussel meat containing around 

double and crab meat around six times the amount of chloroparaffins found in 

the most contaminated sediment sample.” However it is unclear if 

bioaccumulation of C18-20 or C>20 congeners was responsible for the elevated 

concentrations. 

• C20–30 CPs were detected in fat and liver of some postmortem human tissues 

from the United Kingdom at concentrations between 0.080 and 3.5 mg/kg wet 

wt. These findings qualitatively indicate the potential for bioaccumulation of 

LCCPs in the human food chain. 
30

Although there are noteable uncertainties, based mainly on the available BAF 

information, it is concluded that C>20 liquid LCCPs are not bioaccumulative substances 

according to the criteria stipulated in the Persistence and Bioaccumulation Regulations of 

CEPA 1999 (Government of Canada 2000). 

4.2.3.3 C>20 solid LCCPs 

Table 6 summarizes persistence and bioaccumulation information for C>20 solid LCCPs 

in comparison with criteria in the CEPA 1999 Persistence and Bioaccumulation 

Regulations (Government of Canada 2000). 

Table 6. Summary of persistence and bioaccumulation information on C>20 solid LCCPs. 

Medium or 

parameter 

CEPA criteria 1 C>20 solid LCCPs Information 

Air t1/2 ≥ 2 days Estimated to be ≥7.8 days but it should be noted that the 

extent of partitioning to air for LCCPs is low 

Sediment t1/2 ≥ 1 year Unknown, but half life likely > 1 year 

Soil t1/2 ≥ 6 months Unknown 

BAF ≥5000 Low accumulation by salmon; poor absorption and high 

excretion via feces by rats; Modified Gobas Model predicts 

BAF <5000 

BCF ≥5000 Laboratory BCFs <5000; BCF probably underestimated due 

to CP concentrations exceeding solubility 

Log KOW ≥5 Unknown 1

 Government of Canada 2000 

A- Persistence 

Air 

Atmospheric half-lives for solid LCCPs were calculated using the Syracuse Research 

Corporation AOPWIN (v. 1.91) program using a hydroxyl radical concentration of 5 × 

105

 molecules/cm3. Half-lives for solid C18-25 LCCPs ranged from 7.8 to 15.5 days. 

However, LCCPs have very low partitioning to air. 

On the basis of the available information, it is concluded that estimated atmospheric halflives of C>20 solid exceed the criterion of 2 days and hence are persistent in air according 

to the criteria stipulated in the Persistence and Bioaccumulation Regulations of CEPA 

1999 (Government of Canada 2000). 

Sediment and Soil 

No soil or sediment half-life data exist for the C>20 solid LCCPs. However, given that 

both SCCPs and MCCPs are expected to be persistent in sediment (half lives > 1 year), 

and that resistance to microbial degradation has been observed to generally increase with 
31

increases in carbon chain length (Allpress and Gowland 1999; Omori et al. 1987), it is 

likely that LCCPs also have half lives of more than 1 year in sediment. 

On the basis of the available information, it is concluded that C>20 solid LCCPs are 

persistent in sediments according to the criteria stipulated in the Persistence and 

Bioaccumulation Regulations of CEPA 1999 (Government of Canada 2000). 

B- Bioaccumulation 

Although C>20 LCCPs may have some potential to bioaccumulate, the Modified Gobas 

BAF Model predicts that none of the C>20 congeners meet the bioaccumulation criterion 

of BAF ≥5000. Thus, these very high molecular weight LCCPs are not expected to be 

bioaccumulative. 

Measured BCF values for solid LCCPs were found to range from 5.7 to 341 in fish and 

common mussels (Madeley and Maddock 1983c, Madeley and Thompson 1983c). 

However, these studies may underestimate the true BCF values, because the studies were 

performed at LCCPs concentrations above the solubility limit for water and hence were 

not in compliance with OECD guidelines. As well, the studies did not indicate if steady 

state was reached during the uptake phase of the tests. Estimated BCF values for these 

species were below the BCF criterion of 5000 (Madeley and Maddock 1983b,c; 

Bengtsson et al. 1979; Madeley and Thompson 1983b,c). 

Log Kow values are not available for C>20 solid LCCPs. 

Other evidence that C>20 LCCPs may not be bioaccumulative is as follows: 

• One aquatic BAF study was identified for C>20 solid LCCPs. Zitko (1974) 

observed very low accumulation of a 70% chlorine LCCPs by juvenile Atlantic 

salmon fed a diet that had high CP concentrations (10 and 100 μg/g) during a 181-

day exposure period. 

• Two rat bioaccumulation studies with LCCPs, including C>20 solid LCCPs, 

showed high rates of excretion via feces and poor absorption of the LCCPs, 

Section 4.4.3.2, supporting document (Environment Canada, 2008).. No BMF 

data exist for C>20 solid LCCPs. 

Although there are noteable uncertainties, based on the available information it is 

concluded that C>20 solid LCCPs are not bioaccumulative substances according to the 

criteria stipulated in the Persistence and Bioaccumulation Regulations of CEPA 1999 

(Government of Canada 2000). 

4.3. Environmental Concentrations 
32

This section describes the results of recent monitoring of CPs in environmental samples 

using analytical techniques having higher specificity for SCCPsS. Data on environmental 

levels of MCCPs and LCCPs are very limited. Due to the non-volatile and hydrophobic 

characteristics of these CP groups, the majority of results are for sediments and sewage 

sludges. 

Data presented in this section focus on Canadian concentrations. In situations where 

Canadian data are lacking or are few, concentrations measured in other countries are 

presented. Additional information on ambient concentrations may be found in the 

Supporting Document (Environment Canada, 2008). 

4.3.1 Atmospheric concentrations 

SCCPs were detected in air in Canada, United Kingdon and Norway. They have also 

been detected in arctic air and in air of other remote areas (Section 4.2.1). Concentrations 

of SCCPs in air samples collected at Egbert, Ontario, Canada, in 1990 ranged from 65 to 

924 pg/m3 (Tomy 1997; Tomy et. al. 1998a). Concentrations of SCCPs over Lake Ontario 

in 1999 and 2000 ranged from 120 to 1,510 pg/m3

 (Muir et al. 2001). 

No atmospheric concentration data are available for MCCPs and LCCPs, either in 

Canada or elsewhere. 

4.3.2 Wastewater treatment effluents, sewage sludge and soils 

SCCPs were detected in wastewater effluents in Canada, the United States and Germany. 

SCCPs were detected in all eight sewage treatment plant final effluents from southern 

Ontario, Canada, sampled in 1996. Total SCCPs (dissolved and particulate C10-13) ranged 

from 59 to 448 ng/L. The highest concentrations were found in samples from treatment 

plants in industrialized areas, including Hamilton, St. Catharine’s and Galt. No 

wastewater treatment effluent concentration data are available for MCCPs and LCCPs, 

either in Canada or elsewhere. 

Concentrations of CPs have also been detected in sewage sludge in several European 

countries and the United States. Nicholls et al. (2001) found total CPs (SCCPs + 

MCCPs) concentrations in digested sewage sludge ranging from 1.8 to 93.1 mg/kg dry 

wt. in England and Wales. Similarly, Stevens et al. (2002) found SCCPs concentrations 

ranging from 6.9 to 200 mg/kg dry wt. in sewage sludge from 14 WWTPs in the United 

Kingdom. Highest concentrations of SCCPs were in sludge from industrial catchments. 

However, a rural catchment with zero industrial effluent had significant levels (590 

mg/kg) of total SCCPs/MCCPs in sludge (Stevens et al. 2002). Total concentrations of 

MCCPs in sewage sludges from 15 WWTPs in the United Kingdom ranged from 30 to 

9,700 mg/kg dry wt. (Stevens et al. 2002). Agricultural soils may also be a potentially 

major reservoir of CPs due to sewage sludge application (Stevens et al. 2002; Nicholls et 

al. 2001). No values in sewage sludge or soil were identified for LCCPs. Concentrations 

of CPs in sewage sludge in Canada are not available. 
33

4.3.3 Surface waters 

SCCPs were detected in surface waters in Canada and the United Kingdom. Low levels 

of dissolved total (C10-13) SCCPs were measured in western Lake Ontario between 1999 

and 2004 (Muir et al. 2001, Houde et al. 2006). The concentration of total SCCPs was 

1.75 ng/L in 1999. Concentrations of total SCCPs ranged from 0.606 – 1.935 ng/L over 

the 2000 – 2004 sampling period. Concentrations were generally greater in western Lake 

Ontario, likely due to the proximity of large urban areas (Houde et al. 2006). SCCPs 

concentrations of 30 ± 14 ng/L were measured in the Red River in Selkirk, Manitoba, 

over a 6-month period in 1995 (Tomy 1997). 

MCCPs were detected in surface waters in Canada, the United States, the United 

Kingdom and Germany. Metcalfe-Smith et al. (1995) reported C14–17 MCCPs 

concentrations in a 24-hour composite sample of effluent from the only manufacturing 

plant in Canada, ICI Canada (now PCI Canada), on the St. Lawrence River at Cornwall, 

Ontario, to be 12,700 ng/L. This plant is not currently manufacturing CPs. Houde et al. 

(2006) collected water samples from various sites in Lake Ontario in 2002 and 2004. 

Total MCCPs concentrations ranged from <0.0005 to 0.0026 ng/L in filtered samples. 

Concentrations of MCCPs in an impoundment ditch that received effluent from a CP 

production plant in Dover, Ohio, were <150 – 3,800 ng/L (Murray et al. 1988). MCCPs 

were found in all the samples taken in 16 rivers, canals and reservoirs in the United 

Kingdom (ICI 1992). Concentrations ranged from 620 to 3,750 ng/L. The majority of the 

samples appear to have been collected in urban/industrial areas. Levels of MCCPs have 

been measured at several sites in Germany (Hoechst AG 1987; Ballschmiter 1994). The 

levels measured in 1987 ranged from 4,000 to 20,000 ng/L while those of 1994 were 

substantially lower and ranged from < 50 to 185 ng/L. 

There are no Canadian measured water concentrations of LCCPs and very few 

measurements of LCCPs in surface waters from other countries. Nicholls et al. (2001) 

reported <100 ng/L of any CP group in all sites near sewage treatment works in the 

United Kingdom except for one (Darwen, U.K.). Only one study was identified 

measuring surface water concentrations of LCCPs. Murray et al. (1988) conducted a 

study near a CPs production plant in Dover, Ohio, reporting total concentrations of C20–30, 

40–50% chlorine LCCPs of 8,300 ng/L in the middle of the impoundment lagoon at this 

site. In a drainage ditch leading from the impoundment lagoon, a concentration of 4,200 

ng/L total LCCPs (3,700 ng/L particulate, 500 ng/L dissolved) was measured just above 

its discharge to Sugar Creek. A concentration of 620 ng/L particulates (<50 ng/L 

dissolved) was found in water from Sugar Creek, just downstream of the outlet of the 

drainage ditch. 

4.3.4 Sediments 

SCCPs were detected in sediments around the Great Lakes, St. Lawrence River, and other 

lakes in Canada, as well as in Germany, Czech Republic and the United Kingdom. They 

have also been detected in arctic sediment (Section 4.2.1). Concentrations of SCCPs in 

Lake Winnipeg and Lake Nippigon ranged from 0.008 to 0.176 mg/kg dry wt. (Tomy et 

al. 1999; Stern and Evans 2003). Tomy et al. (1997) measured SCCPs at concentrations 
34

around 0.245 mg/kg dry wt. in sediment from the mouth of the Detroit River at Lake Erie 

and Middle Sister Island in western Lake Erie, in 1995. SCCPs were also detected in all 

surface sediment samples from harbour areas along Lake Ontario at concentrations 

ranging from 0.0059 to 0.290 mg/kg dry wt. in 1996 (Muir et al. 2001). The highest 

concentrations were found at the most industrialized site (Windermere Basin, Hamilton 

Harbour), which has well-documented heavy metal, PAH and PCB contamination. 

Similarly, Marvin et al. (2003) reported a SCCPs concentration of 0.410 mg/kg dry wt. in 

Lake Ontario sediments near an industrialized area. SCCPs were detected in all 26 

samples from Lake Ontario, and the average SCCPs concentration was 0.049 mg/kg dry 

wt., which is much higher than sediment concentrations reported for lakes (Yaya, DV09, 

Hazen, Nipigon) influenced primarily by atmospheric sources (Tomy et al. 1999; Stern 

and Evans 2003). 

MCCPs were detected in sediments around the Great Lakes in Canada, as well as in the 

United States, Germany, Wales, Switzerland, Australia and the United Kingdom. 

Metcalfe-Smith et al. (1995) were unable to detect (<3.5 mg/kg dry wt.) SCCPs + 

MCCPs in sediments from the St. Lawrence River downstream of a CP manufacturing 

plant. Tomy and Stern (1999) reported concentrations of C14–17 MCCPs of 0.068 mg/kg 

dry wt. in sediment samples collected near the mouth of the Detroit River in western 

Lake Erie. Muir et al. (2002) reported concentrations of total MCCPs in a sediment core 

from Lake St. Francis downstream of Cornwall, Ontario, of 0.75 –1.2 mg/kg dry wt. The 

highest concentrations of MCCPs detected in sediments were found downstream from 

sewage treatment works in the United Kingdom. Concentrations of MCCPs ranged from 

<0.2 to 65.1 mg/kg dry wt. (Nicholls et al. 2001). Similar concentrations were found at 

several other locations downstream from sewage treatment plants in the United Kingdom 

(Nicholls et al. 2001).

No LCCPs were measured in sediments in Canada, but they have been detected in the 

United States, Australia and Germany near CP manufacturing plants. Concentrations of 

LCCPs in these countries ranged from 0.0081 to 170 mg/kg dry wt. (Rotard et al. 1998; 

Murray et al. 1988; Kemmlein et al. 2002). 

4.3.5 Biota 

A- Aquatic Biota 

SCCPs were detected in biota in Canada, England, Norway, Chile, Greece, Germany, 

Iceland, France, the United States, and the North and Baltic Seas. Muir et al. (2001) and 

Houde et al. (2006) measured SCCPs in fish collected from Lake Ontario and Lake 

Michigan, between 1996 and 2001. Concentrations of total SCCPs ranged from 0.0046 

to 2.63 mg/kg wet wt. The highest concentration was measured in carp collected at 

Hamilton harbour (Muir et al. 2001). Houde et al. (2006) determined the concentration 

of SCCPs in plankton, Diporeia sp. and Mysis sp. from Lakes Ontario and Michigan. In 

Lake Ontario, total SCCPs concentrations in plankton, Diporeia and Mysis were 0.0055, 

0.0063, and 0.0028 mg/kg wet wt., respectively, and in Lake Michigan they were 0.023, 

0.024, and 0.0075 mg/kg wet wt., respectively. 
35

MCCPs have been measured in fish in Canada, the United Kingdon, Norway, Chile, 

Greece and Germany amongst others. Houde et al. (2006) also measured the 

concentrations of MCCPs in fish in Lake Ontario and Lake Michigan in 1999 and 2001. 

Concentrations of total MCCPs ranged from 0.0028 to 0.109 mg/kg weight wt. MCCPs 

were also detected in Diporeia at concentrations ranging from 0.0024 to 0.0041 mg/kg 

(Houde et al. 2006). The highest concentration in fish measured in Canada was 0.904 

mg/kg weight wt. for catfish in the Detroit River (Tomy and Stern 1999). 

Murray et al. (1988) measured concentrations of C20–30, 42% chlorine LCCPs in zebra 

mussels from Sugar Creek, Ohio, near a CPs manufacturing plant. Concentrations ranged 

from <0.007 upstream to 0.18 mg/kg downstream of where the drainage ditch from the 

plant emptied into Sugar Creek. Kemmlein et al. (2002) found high levels of C18–29

LCCPs in marine mussels and crabs (9.3 and 14.3 mg/kg lipid wt., respectively) near a 

CPs manufacturing plant in Australia. 

B- Marine Mammals 

SCCPs have been detected in the blubber of belugas from the St. Lawrence River at an 

average concentration of 0.785 mg/kg weight wt. SCCPs have also been detected in the 

blubber of ringed seal from southwest Ellesmere Island (Eureka), Pangnirtung 

(Cumberland Sound) and Svalbard; in beluga whales from northwest Greenland, 

Sanikiluaq (Hudson Bay), Pangnirtung (Cumberland Sound), Kimmirut and the 

Mackenzie Delta; and in walrus from northwest Greenland. Concentrations of SCCPs 

from these areas ranged from 0.095 to 0.626 mg/kg weight wt. (Jansson et al. 1993; 

Tomy et al. 1998b; 2000).

Concentrations of MCCPs in beluga blubber in the St-Lawrence ranged from 1.8 – 80.0 

mg/kg weight wt. (Bennie et al. 2000). However, results obtained by Bennie et al. (2000) 

may not be reliable due to interferences in the analytical method. 

C- Terrestrial and Avian Wildlife 

Very limited information is available on SCCPs concentrations in tissues of terrestrial 

wildlife. In Sweden, Jansson et al. (1993) reported CP concentrations (unspecified chain 

length) in rabbit (Revingeshed, Skåne), moose (Grismsö, Västmanland), reindeer (Ottsjö, 

Jaämtland) and osprey (from various regions in Sweden) to be 2.9, 4.4, 0.14 and 0.53 

mg/kg lipid wt., respectively. Nicholls et al. (2001) reported the concentrations of 

SCCPs and MCCPs in earthworms residing in fields on which sludge had been applied 

ranging from <0.1 to 0.7 mg/kg dry wt. in the United Kingdom in the summer of 1998. 

Campbell and McConnell (1980a) determined levels of C10–20 CPs in birds in the United 

Kingdom. The C10–20 levels were likely to be dominated by contributions from the 

SCCPs and MCCPs. Concentrations of C10–20 CPs ranged from 0.1 to 1.2 mg/kg weight 

wt. in liver of birds and from <0.05 to >6 mg/kg in seabird eggs. Concentrations of C20–

30 CPs ranged from not detected to 1.5 mg/kg weight wt. in liver of birds and from <0.05 

to 1 mg/kg in seabird eggs. Reth et al. (2006) quantified SCCPs in liver and muscle from 
36

the seabirds, little auk (Alle alle) and kittiwake (Rissa tridactyla) collected at Bear Island 

(European Arctic). Concentrations between 0.005 and 0.088 mg/kg wet weight were 

measured. Reth et al. (2006) determined the concentration of C14-15 MCCPs in seabirds 

from the European Arctic. Concentrations ranged from 0.005 to 0.370 mg/kg wet wt. 

4.4. Environmental effects 

Overall, toxicity studies are few for effects of SCCPs to pelagic biota and mammals. 

LOECs (i.e., survival, reproduction and growth) ranged from 8,900 to 10,000 ng/L for 

pelagic biota. Effects of SCCPs to benthic and soil-dwelling organisms are not available. 

More toxicological data are available for MCCPs. In particular, the acute and chronic 

toxicity of MCCPs has been studied in algae, invertebrates and several species of fish. 

The range of acute effects is 5,900 ng/L to > 10g/L (10 000 000 000 ng/L). LOECs for 

pelagic biota ranged from 18,000 to 31,000 ng/L. Contrary to SCCPs, toxicity studies, 

albeit few, are available for benthic and soil-dwelling organisms. LOECs for sedimentdwelling biota ranged from 270 to 410 mg/kg dry weight. A reproductive LOEC for 

earthworm was reported to be 383 mg/kg dry weight. Few studies are available for 

effects of MCCPs to mammals; LOAELs ranged from 4.2 to 5.7 mg/kg bw/day for 

effects to rats. Similarly, limited number of studies is available for effects to pelagic 

biota. Acute effects were observed at greater than 3 800 000 ng/L. Very few 

toxicological data are available for the three types of LCCPs. These data are presented 

below. 

This section will focus on the most sensitive toxicological information used to derive the 

critical toxicity values (CTV) only. Additional toxicity information is available in the 

supporting document. 

4.4.1 SCCPs 

A- Pelagic aquatic organisms 

The lowest toxic effect level identified for a pelagic freshwater aquatic species is 8,900 

ng/L, which is the 21-day chronic LOEC for Daphnia magna (Thompson and Madeley 

1983a). The effect was for mortality of the offspring. The NOEC is 5,000 ng/L. 

B- Benthic organisms 

A valid measurement endpoint was not available for a sediment-dwelling invertebrate. 

As a result, an equilibrium partitioning approach (Di Toro et al. 1991) using the most 

sensitive chronic measurement endpoint identified for a pelagic freshwater invertebrate 

aquatic species (8,900 ng/L) was used to estimate the toxicity to benthic organisms. The 

LOECbenthic was estimated to be 35.5 mg/kg dry wt. for sediment containing 2% organic 

carbon (Environment Canada, 2008). 
37

 

C- Soil-dwelling organisms 

 

Bezchlebová et al. (2007) investigated the effects of SCCPs on the survival and 

reproduction of five species of soil organisms (Fosomia candida, Eisenia fetida, 

Enchytraeus albidus, Enchytraeus crypticus, and Caenorhabditis elegans). All tests were 

preformed following international methods, using an OECD artificial soil (70% sand, 

20% clay, 10% peat) with an organic carbon content of approximately 2.7%. Folsomia 

candida (collembola) was identified as the most sensitive organism, with an LC50 value 

for adult survival and EC50 and EC10 values for reproduction of 5733, 1230, and 660 

mg/kg dry wt. (nominal), respectively. The soil CTV for SCCPs is 660 mg/kg dry wt. 

D- Mammals 

In a 13-week oral (gavage) rat study by IRDC (1984), increases in liver and kidney 

weight and hypertrophy of the liver and thyroid occurred at doses of 100 mg/kg-bw per 

day. This value was the most sensitive LOAEL for mammals. Interspecies scaling using 

data for a typical adult otter was used to extrapolate to a food concentration for this 

species. This resulted in a CTV of 1,000 mg/kg food. See Table 7 of this report and 

Appendix 2 of supporting document for additional information (Environment Canada, 

2008). 

4.4.2 MCCPs 

A- Pelagic aquatic organisms 

In a 21-day chronic study with Daphnia, Thompson et al. (1997) reported a LOEC of 

18,000 ng/L and a NOEC of 10,000 ng/L for a decrease in the number of live offspring 

and the length of the parent organisms. This LOEC is the most sensitive toxicity value 

for aquatic organisms. 

B- Benthic organisms 

The most sensitive value for sediment toxicity of MCCPs is the LOEC for growth from a 

28-day study with the amphipod Hyalella azteca using sediment that contained 5% 

organic carbon (Thompson et al. 2003). A statistically significant (p = 0.05) reduction in 

the mean dry weights of survivors in the treatment groups was seen at exposure 

concentrations of 270 mg/kg dry wt. and above when compared with the solvent control. 

C- Soil-dwelling organisms 

The most sensitive toxicity value for terrestrial organisms is the chronic (28-day) LOEC 
38

of 383 mg/kg dry wt. in soil with an organic carbon content of 2%, for reproduction in 

earthworms (Thompson et al. 2001a). 

D- Mammals 

The lowest effect level observed for mammals is the LOAEL of 4.2 mg/kg-bw per day 

for mild effects on the kidney and thyroid of female rats during a 13-week feeding study 

(Poon et al. 1995). Interspecies scaling using data for a typical adult otter was used to 

extrapolate to a food concentration for this species. This resulted in a CTV of 42 mg/kg 

food. See Table 7 of this report and Appendix 2 of supporting document for additional 

information. 

4.4.3 LCCPs 

4.4.3.1 LCCPs (C18-20 liquid) 

A- Pelagic aquatic organisms

A chronic 21-day Daphnia magna study was carried out by Frank (1993) and Frank and 

Steinhäuser (1994). The most sensitive aquatic toxicity value for liquid C18-20 LCCPs is 

the 21-day (chronic) LOEC of 68,000 ng/L. 

B- Soil-dwelling organisms 

There are no studies available on the toxicity of either liquid or solid LCCPs to terrestrial 

plants, earthworms or other soil-dwelling organisms. Therefore, an equilibrium 

partitioning approach (Di Toro et al. 1991) using the most sensitive measurement 

endpoint identified for a pelagic freshwater species (68,000 ng/L) was used to estimate 

the toxicity of liquid C18-20 LCCPs to soil-dwelling organisms. The LOECsoil for C18-20

LCCPs was estimated to be 2,035 mg/kg dry wt. for a soil containg 2% organic carbon 

(Environment Canada, 2008). 

4.4.3.2 LCCPs (C>20 liquid) 

There is no relevant exposure or toxicity data available for C>20 liquid LCCPs in pelagic 

organisms, benthic organisms, or soil dwelling organisms. 

A- Mammals 

In 90-day and 2-year feeding studies with rats with C>20 (43% chlorine by weight) 
39

LCCPs, the lowest LOAEL in the studies was 100 mg/kg-bw per day (Serrone et al. 

1987; Bucher et al. 1987; NTP 1986). This LOAEL was the most sensitive toxicity value. 

The main effects were seen on the liver, and in both studies effects were seen at the 

lowest concentrations. Interspecies scaling using data for a typical adult otter will be used 

to extrapolate to a food concentration for this species. This resulted in a CTV of 1,000 

mg/kg food. See Table 7 of this report and Appendix 2 of supporting document for 

additional information. 

4.4.3.3 LCCPs (C>20 solid) 

There is no relevant exposure or toxicity data available for C>20 solid LCCPs in pelagic 

organisms, benthic organisms, or soil dwelling organisms. 

A- Mammals 

Serrone et al. (1987) reported a LOAEL for hepatic lesions in female rats following 

administration by gavage of another long-chain CP (C20–30, 43% chlorine) during a 90-

day study. In addition, mild nephrosis was observed in the kidneys of male rats, as was 

mineralization in the kidneys of female rats administered 3750 mg/kg-bw per day. A 

NOEL could not be established for the females (LOEL = 100 mg/kg- food). Interspecies 

scaling using data for a typical adult otter will be used to extrapolate to a food 

concentration for this species. This resulted in a CTV of 100 mg/kg food. See Table 7 of 

this report and Appendix 2 of supporting document for additional information. 

4.5. Potential to cause ecological harm 

Potential to cause environmental harm may be estimated quantitatively using risk 

quotients (RQs). When RQs exceed 1 (i.e., in this case when Estimated Exposure Values 

(EEVs) exceed Estimated No-Effect Values (ENEVs)) this is an indication of potential 

for risk. 

It is acknowledged, however, that when risks for persistent and bioaccumulative 

substances - such as SCCPs, MCCPs, and C18-20 LCCPs - are determined using standard 

methods, the risks may be underestimated. For example, since it can take decades for 

persistent substances to achieve maximum steady state concentrations in sediment and 

soil, EEVs based on monitoring data will be too low if steady state concentrations have 

not been achieved in these media. Similarly, since it can take a long time for persistent 

and bioaccumulative substances to reach maximum steady state concentrations in the 

tissues of laboratory organisms, ENEVs based on standard toxicity tests may 

underestimate effect thresholds if test durations are insufficient to achieve maximum 

internal organism concentrations. Furthermore, since food consumption is usually the 

primary route of exposure to persistent and bioaccumulative substances in the field – 

especially for top predators - ENEVs may underestimate effect thresholds if the food 

pathway is not considered in key toxicity studies. These factors are exacerbated when 
40

available effects and exposure data are limited, as is the case for the chlorinated paraffins. 

Risk quotients were calculated for SCCPs, MCCPs, C18-20 LCCPs and C>20 LCCPs (Table 

7). For each identified class of risk receptors (e.g., pelagic organisms, benthic 

organisms), an EEV was selected based on empirical data. The maximum reported field 

concentration which is relevant to the Canadian environment was used as the EEV. 

Chemical concentrations from the Canadian environment were preferably used for EEVs; 

however, data from other regions in the world were used in the absence of suitable 

Canadian data. Section 8.2 of the supporting document (Environment Canada, 2008) 

further discusses this point. An ENEV was determined by dividing a Critical Toxicity 

Value (CTV) by an assessment factor. CTVs, a detailed description is provide in Section 

8.0 of the supporting document (Environment Canada, 2008), typically represent the 

lowest chronic ecotoxicity value from an available and acceptable data set. Assessment 

factors were used to reduce the CTV to account for extrapolation from a sometimes 

limited set of effects data for laboratory organisms, to estimates of effect thresholds for 

sensitive species in the field. Note that an extra assessment factor was not used to 

account for the tendency for conventional RQs to underestimate potential for harm for 

persistent and bioaccumulative substances. Results are summarized in Table 7. 

Concentrations of C18-20 liquid LCCPs in sediments representative of Canadian 

environments are not available. In addition, no toxicity data were available for the effects 

of C18-20 liquid LCCPs on secondary consumers. Therefore, risk quotients could not be 

calculated for exposure of benthic organisms and secondary consumers to C18-20 liquid 

LCCPs. Furthermore there are no relevant exposure and toxicity data available for C>20

liquid and C>20 solid LCCPs in pelagic organisms, benthic organisms, or soil dwelling 

organisms. As such, risk quotients were not calculated for these groups. 

Table 7. List of Estimated Exposure Values (EEV), Critical Toxicity Values (CTV), 

Assessment Factors (AF), and Estimated No Exposure Values (ENEV) used in the 

calculation of Risk Quotients (RQ) for SCCPs, MCCPs, C18-20 liquid LCCPs, C>20 

liquid LCCPs and C>20 solid LCCPs. 

Organism EEV CTV AF ENEV RQ 

(EEV/ENEV) 

SCCPs 

Pelagic 44.8a ng/L 8,900b ng/L 10 (lab/field) 890 ng/L 0.05 

Benthic 0.41c mg/kg 35.5d mg/kg 10 (lab/field) 3.55 mg/kg 0.12 

Soil-dwelling 0.64emg/kg 660d mg/kg 10 (lab/field) 66.0 mg/kg 0.01 

Secondary Consumer 2.63f mg/kg 1,000g

 mg/kg 

food 

100 (lab/field & 

species variations) 10 mg/kg 0.26 

MCCPs 

Pelagic 0.0026h ng/L 18,000i ng/L 10 (lab/field) 1,800 ng/L 0.0000014 

Benthic 65.1j mg/kg 270k mg/kg 10 (lab/field) 27 mg/kg 2.40 

Soil-dwelling 31.0l mg/kg 383m mg/kg 10 (lab/field) 38.3 mg/kg 0.81 

Secondary Consumer 0.904n mg/kg 42o mg/kg food 100 (lab/field & 

species variations) 0.42 mg/kg 2.15 

C18-20 liquid LCCPs 

Pelagic 100p ng/L 68,000q ng/L 10 (lab/field) 6,800 ng/L 0.02 
41

Soil-dwelling 3.1r mg/kg 2,035s mg/kg 10 (lab/field) 203.5 mg/kg 0.02 

C>20 liquid LCCPs 

Secondary Consumer 0.0465t mg/kg 1,000u mg/kg 10 (lab/field) 100 mg/kg 0.0005 

C>20 solid LCCPs 

Secondary Consumer 0.0465v mg/kg 100w mg/kg 10 (lab/field) 100 mg/kg 0.000465 a

 The highest concentration of SCCPs observed in final effluent of sewage treatment plants in southern Ontario was 448 

ng/L at the Woodward Avenue sewage treatment plant in Hamilton, Ontario. A dilution factor of 10 was used to 

calculate the EEV which results in an EEV of 44.8 ng/L. b

 21-day LOEC for Daphnia magna. c Highest concentration in surface sediments observed from Lake Ontario, Niagara (or west) basin, in 1998. d

 EC10 for F. candida reproduction. e

 The maximum allowable rate for sewage biosolid application to agricultural lands is 8 tonnes of solids per hectare per 

5 years (MOE 1998). The soil mass is 5,000 tonnes/ha (assuming that the biosolids are incorporated into the top 20 

cm of the soil having a dry soil bulk density of 2500 kg/m3

 (EU 2003)). Using a SCCPs concentration in sewage 

sludge of 200 mg/kg dry wt. and assuming that SCCPs-containing sludge is applied to the land for 10 years and that 

no or little biodegradation of the SCCPs occurs, a soil concentration of 0.64 mg/kg dry wt is estimated. f

 Concentration of total SCCPs found in carp from Hamilton Harbour in Lake Ontario. 

g

 The LOAEL for the 13-week oral (gavage) rat study is 100 mg/kg bw/day (IRDC 1984). Interspecies scaling using 

this LOAEL for a typical adult otter (lutra canadensis) (adult body weight of 8 kg and average daily food ingestion 

rate of 0.8 kg wet wt. per day) was used to extrapolate to a food concentration for this species (CCME 1998). The 

resulting CTV was 1,000 mg/kg food wet wt. h

 Concentration measured in Lake Ontario. 

i

 21-day LOEC for Daphnia magna. 

j

Concentration found downstream from sewage treatment works in the United Kingdom. k

 28-day LOEC for growth for the amphipod Hyalella azteca. l

 The maximum allowable rate for sewage biosolid application to agricultural lands is 8 tonnes of solids per hectare per 

5 years (MOE 1998). The soil mass is 5,000 tonnes/ha (assuming that the biosolids are incorporated into the top 20 

cm of the soil having a dry soil bulk density of 2500 kg/m3 (EU 2003)). Using a MCCPs concentration in sewage 

sludge of 9,700 mg/kg dry wt. and assuming that SCCPs-containing sludge is applied to the land for 10 years and that 

no or little biodegradation of the SCCPs occurs, a soil concentration of 31 mg/kg dry wt is estimated. m 28-day LOEC in soil with an organic carbon content of 2% for reproduction in earthworms. n

 Concentration of MCCPs in catfish collected from the Detroit River, Michigan, and southern Ontario. o

 The LOAEL for the 13-week oral (gavage) rat study is 4.2 mg/kg bw/day (Poon et al. 1995). Interspecies scaling 

using this LOAEL for a typical adult otter (lutra canadensis) (adult body weight of 8 kg and average daily food 

ingestion rate of 0.8 kg wet wt. per day) was used to extrapolate to a food concentration for this species (CCME 

1998). The resulting CTV was 42 mg/kg food wet wt. p

 Detection limit for sewage treatment works in the United Kingdom. q

 21-day LOEC for Daphnia magna. r The maximum allowable rate for sewage biosolid application to agricultural lands is 8 tonnes of solids per hectare per 

5 years (MOE 1998). The soil mass is 5,000 tonnes/ha (assuming that the biosolids are incorporated into the top 20 

cm of the soil having a dry soil bulk density of 2500 kg/m3

 (EU 2003)). Using a MCCPs concentration in sewage 

sludge of 9,700 mg/kg dry wt. (worst-case concentration in the absence of exposure data for C18-20 liquid LCCPs) and 

assuming that SCCPs-containing sludge is applied to the land for 10 years and that no or little biodegradation of the 

SCCPs occurs, a soil concentration of 31 mg/kg dry wt is estimated. Since LCCPs usage is about 10% of MCCPs 

usage, this would result in a C18-20 liquid LCCPs soil concentration of 3.1 mg/kg dry wt. s

 Value calculated using an equilibrium partitioning approach using the Daphnia magna LOEC. t

 C18–29 CPs in mussel near a manufacturing plant in Australia had a lipid wt. concentration of 9.3 mg/kg (Kemmlein et 

al. 2002). Using an average lipid content of zebra mussels and other North American exotic mussel species in the 

Great Lakes of 0.5% wet wt. (Marvin 2003), and assuming that the Australian mussel had a similar lipid content to 

zebra mussels and that all of the LCCPs measured in the Australian mussel were of the C>20 liquid type, the 

concentration of LCCPs in mussels was estimated to be 0.0465 mg/kg on a wet wt. u

 The LOAEL for a 90-day (Serrone et al. 1987) and 2-year feeding (Bucher et al. 1987) studies with rats with C>20

(43% chlorine by weight) LCCPs is 100 mg/kg bw/day. Interspecies scaling using this LOAEL for a typical adult 

otter (lutra canadensis) (adult body weight of 8 kg and average daily food ingestion rate of 0.8 kg wet wt. per day) 

was used to extrapolate to a food concentration for this species (CCME 1998). The resulting CTV was 1,000 mg/kg 

food wet wt. v

 C18–29 CPs in mussel near a manufacturing plant in Australia had a lipid wt. concentration of 9.3 mg/kg (Kemmlein et 

al. 2002). Using an average lipid content of zebra mussels and other North American exotic mussel species in the 

Great Lakes of 0.5% wet wt. (Marvin 2003), and assuming that the Australian mussel had a similar lipid content to 

zebra mussels and that all of the LCCPs measured in the Australian mussel were of the C>20 solid type, the 
42

concentration of LCCPs in mussels was estimated to be 0.0465 mg/kg on a wet wt. w Effects were seen in the liver and kidney of rats at a concentration of 3750 mg/kg-bw per day in a 90-day dietary 

study (Serrone et al. 1987). The LOAEL for a 90-day dietary study with rats was 3,750 mg/kg bw/day (Serrone et al. 

1987). Interspecies scaling using this LOAEL for a typical adult otter (lutra canadensis) (adult body weight of 8 kg 

and average daily food ingestion rate of 0.8 kg wet wt. per day) was used to extrapolate to a food concentration for 

this species (CCME 1998). The resulting CTV was 37,500 mg/kg food wet wt. 

Only two of the 12 calculated risk quotients are larger than 1. The MCCPs risk quotient 

for benthic organisms (RQ=2.40) and the MCCPs risk quotient for secondary consumers 

(RQ=2.15) both suggest that MCCPs pose a risk to these receptors. However, because of 

limitations in available exposure and effects data mentioned above and explained in more 

detail in Section 8.2 of the supporting document, the absence of RQs above 1 for SCCPs 

and C18-20 LCCPs cannot be considered proof that these persistent and bioaccumulative 

substances do not cause ecological harm. 

Because data available for C>20 LCCPs are very limited, only one RQ could be calculated 

for each of the solid and liquid subgroups.. Although the resulting RQs are very low, this 

too is likely an underestimate of possible high-end risks, in part because of limitations in 

information on environmental concentrations close to relevant point sources (Section 8.2 

of the supporting document). 

Evidence that a substance is very persistent and bioaccumulative as defined in the 

Persistence and Bioaccumulation Regulations of CEPA 1999, when taken together with 

potential for environmental release and potential for toxicity to organisms, provides a 

significant indication of its potential to cause harmful long term ecological effects. 

Substances that are persistent remain in the environment for a long time, increasing the 

magnitude and duration of exposure. Releases of small amounts of bioaccumulative 

substances may lead to high internal concentrations in exposed organisms. Highly 

bioaccumulative and persistent substances are of special concern, since they may 

biomagnify in food webs, resulting in very high internal exposures, especially for top 

predators. 

SCCPs, MCCPs and C18-20 LCCPs are considered to be both highly persistent and 

bioaccumulative. The limited available evidence suggests that although C>20 LCCPs are 

persistent, they are not bioaccumulative. 

In addition, there is evidence (including some monitoring data), that SCCPs, MCCPs and 

C18-20 LCCPs are released into the Canadian environment and have the potential to cause 

harm to sensitive aquatic organisms at relatively low concentrations (i.e., chronic NOECs 

for pelagic organisms < 100 ng/L). 

In light of this evidence, it is concluded that SCCPs, MCCPs and LCCPs up to C20 may 

be causing long term ecological harm in Canada. 

4.6. Uncertainties on the ecological risk assessment 

This risk assessment contains several sources of uncertainty. Uncertainties in the 
43

exposure and effects assessment can influence the characterization of risks. Below is a 

brief discussion of these uncertainties. Additonal details can be found in Section 8.2 of 

the supporting document. 

4.6.1 Exposure, Effects and Risk Quotient Calculations 

When Canadian exposure data were lacking, data from other countries were used as 

EEVs and assumed to be representative of Canadian conditions. Concentrations of CPs in 

various media were often only available for certain areas, and were only representative of 

a short time period, in Canada and other countries. As a result, it is unkown how 

concentrations of CPs vary temporally and spatially. Moreover, concentrations were often 

not available near potential point sources such as metalworking operations (primary 

source of CPs) and other formulating/manufacturing sites that use CPs. 

Uncertainties with the toxicity information used to drive ENEVs in this assessment 

include: 

• The use of an equilibrium partitioning approach to estimate toxicity to benthic and 

soil organisms for SCCPs and LCCPs. 

• The lack of aquatic toxicity tests for C>20 solid LCCPs, particularly with daphnids, 

a species that was found to be the most sensitive to SCCPs, MCCPs and liquid 

LCCPs. 

• The use of test substance concentrations in excess of their water solubility for all 

fish toxicity tests. 

Additional assessment factors were not used to account for these limitations when 

deriving ENEVs from CTVs. 

Because of the above-mentioned limitations - and the fact that in general risks of 

persistent and bioaccumulative substances are likely to be underestimated using standard 

assessment approaches – ecological risks from exposure to SCCPs, MCCPs, C18-20

LCCPs in Canada have likely been underestimated by risk quotient calculations, 

especially close to industrial sources. In the case of C>20 LCCPs, limitations in the 

available exposure and effects data mean that risks to secondary consumers have likely 

been underestimated, and that risks to other types of organisms cannot be estimated at all. 

4.6.2 Persistence and Bioaccumulation Status and Risk Implications 

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 empirical and modelled bioaccumulation data for SCCPs and MCCPs are very robust 

and indicate the substances are bioaccumulative. While there is a lack of empirical 
44

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. 

Lastly, there are uncertainties associated with extrapolating from evidence that a 

substance is both persistent and bioaccumulative to a conclusion that it may be causing 

ecological harm. However, given that persistent and bioaccumulative substances have the 

potential to cause widespread harm that is difficult to reverse, a precautionary assessment 

approach is justified. 

5. HUMAN HEALTH RISK ASSESSMENT 

5.1. Population exposure 

The following presentation is limited to identified recent data considered critical to 

quantitative estimation of exposure of the general population in Canada to chlorinated 

paraffins and, hence, to assessment of “toxic” under Paragraph 64(c) of CEPA 1999.

Other sources of data that were also identified but were not directly relevant to estimation 

of exposure in Canada include Peters et al. (2000), Borgen et al. (2000, 2002) and 

Lahaniatis et al. (2000). 

The degree of confidence in data on the concentrations of chlorinated paraffins in various 

media varies considerably, depending upon the nature of the analysis. To the extent 

possible, estimates of intake have been based on higher-confidence analyses by highresolution gas chromatography (HRGC)/electron capture negative ion high-resolution 

mass spectrometry (ECNI-HRMS), due to its higher mass resolving power and 

selectivity. However, such information is limited solely to determination of SCCPs in 

human breast milk (Tomy, 1997), fish (Muir et al., 1999) and media that contribute less 

to human exposure, including ambient air (Tomy, 1997), surface water (Tomy, 1997) and 

sediment (Muir et al., 2001). For all chlorinated paraffins, either concentrations in surface 

water and sediment, or the limits of detection for these media, were used as surrogates for 

concentrations in drinking water and soil, respectively, in estimating intake. 

Indeed, data on concentrations of chlorinated paraffins in foodstuffs are extremely 

limited. While additional data on the concentrations of SCCPs, MCCPs and LCCPs in 

foods in the United Kingdom (Campbell and McConnell, 1980b) reported in an early 

investigation reviewed in the PSL1 assessment (Campbell and McConnell, 1980a) were 

acquired and are presented in Table 8, they are considered, at best, to be semiquantitative, owing to limitations of the methodology available at that time. Analysis was 

based on liquid–solid adsorption chromatography, which has now largely been replaced 

by micro-analytical techniques and quantification by visual reference to spots appearing 

on thin-layer chromatographic plates. 

Table 8. Concentrations of short-chain, medium-chain and long-chain chlorinated paraffins 

in foodstuffs 
45

Food Concentration used to represent food group 

group Short- and medium-chain chlorinated paraffins Long-chain chlorinated paraffins 

Dairy 0.3 μg/g 

mean of 13 samples of dairy products in U.K. 

C10–20 (SCCPs and MCCPs) 

(Campbell and McConnell, 1980a) 

0.19 μg/g 

1 sample of cheese in U.K. 

C20–30

(Campbell and McConnell, 1980a) 

Fats 0.15 μg/g 

mean of 6 samples of vegetable oils and derivatives 

C10–20 (SCCPs and MCCPs) 

(Campbell and McConnell, 1980a) 

0.05 μg/g 

detection limit in analysis of 1 sample of 

lard in U.K. 

C20–30 

(Campbell and McConnell, 1980b) 

Fruits 0.025 μg/g 

mean of 16 samples of fruits and vegetables in 

U.K. 

C10–20 (SCCPs and MCCPs) 

(Campbell and McConnell, 1980a) 

0.025 μg/g 

1 sample of peach fruit in U.K. 

C20–30 

(Campbell and McConnell, 1980a) 

Vegetables 0.025 μg/g 

mean of 16 samples of fruits and vegetables in 

U.K. 

C10–20 (SCCPs and MCCPs) 

(Campbell and McConnell, 1980a) 

0.025 μg/g 

1 sample of potato crisps in U.K. 

C20–30 

(Campbell and McConnell, 1980a) 

Cereal 

products 

SCCPs 

0.13 μg/g 

one reported concentration for “Chlorowax 500C” 

in enriched white bread in market basket survey 

carried out by U.S. Food and Drug Administration 

(KAN-DO Office and Pesticides Team, 1995); 

average molecular formula is C12H19Cl7, with 60–

65% chlorine content (w/w) (IPCS, 1996) 

SCCPs/MCCPs 

0.05 μg/g 

detection limit in analysis of 1 sample of corn 

flakes in U.K. 

C10–20 (SCCPs and MCCPs) 

(Campbell and McConnell, 1980b) 

0.05 μg/g 

detection limit in analyses of corn flakes in 

U.K. 

C20–30 

(Campbell and McConnell, 1980b) 

Meat and 

poultry 

0.099 μg/g 

1 sample of bacon in U.K. 

C10–20 (SCCPs and MCCPs) 

(Campbell and McConnell, 1980b) 

0.05 μg/g 

detection limit in analysis of 1 sample each 

of ox liver and beef in U.K. 

C20–30 

(Campbell and McConnell, 1980b) 
46

Food Concentration used to represent food group 

group Short- and medium-chain chlorinated paraffins Long-chain chlorinated paraffins 

Fish Note: Campbell and McConnell (1980b) presented 

data for combined SCCPs and MCCPs. Data for 

fish identified in Bennie et al. (2000), Muir et al.

(1999) and Tomy and Stern (1999) were presented 

as separate analyses. 

SCCPs 

2.630 μg/g (wet weight); analysis of whole samples 

of carp from Hamilton Harbour; C10–C13 (Muir et 

al., 1999) 

0.0588 μg/g; lake trout, Niagara-on-the-Lake (Muir

et al., 1999) 

0.0726 μg/g; lake trout, Port Credit (Muir et al., 

1999) 

 

0.502 μg/g; carp (n = 3) (Bennie et al., 2000) 

1.47 μg/g; trout (n = 10) (Bennie et al., 2000) 

1.8 μg/g (estimated); perch, Detroit River (Tomy 

and Stern, 2000) 

MCCPs 

1.23 μg/g; mean of 10 samples of whole trout from 

western Lake Ontario (Bennie et al., 2000) 

0.393 μg/g; carp (n = 3) (Bennie et al., 2000) 

82 ng/g in perch; 904 ng/g in catfish (Tomy and 

Stern, 1999) 

0.008 μg/g (estimated); perch, Detroit River (Tomy 

and Stern, 2000) 

no data identified 

Eggs no data identified no data identified 

Foods 

primarily 

sugar 

0.025 μg/g 

1 sample of strawberry jam in U.K. 

C10–20 (SCCPs and MCCPs) 

(Campbell and McConnell, 1980b) 

0.05 μg/g 

detection limit in 1 sample of strawberry 

jam in U.K. 

C20–30 

(Campbell and McConnell, 1980b) 

Mixed 

dishes 

no data identified no data identified 

Nuts and 

seeds 

no data identified no data identified 

Soft 

drinks, 

alcohol, 

coffee, tea 

0.05 μg/g 

detection limit in analyses of beverages in U.K. 

(Campbell and McConnell, 1980a) 

0.05 μg/g 

detection limit in analysis of 1 sample each 

of beer and tea in U.K. 

C20–30 
47

Food Concentration used to represent food group 

group Short- and medium-chain chlorinated paraffins Long-chain chlorinated paraffins 

(Campbell and McConnell, 1980b) 

5.1.1 SCCPs 

Tomy (1997) determined SCCPs (C10–13, 60–70% chlorine) in 24-hour air samples 

collected daily during a 4-month period in the summer of 1990 in Egbert, Ontario, a 

“rural site northwest of Toronto,” by HRGC/ECNI-HRMS (Muir et al., 1999). 

Concentrations ranged from 65 to 924 pg/m3

. Although a summary statistic of 543 pg/m3

was reported, it was not specified whether this was a mean or median value. Egbert has 

also been reported to be near an “industrialized area” (Muir et al., 2000). Lower 

concentrations of SCCPs have been identified at other sites in Canada (Halsall et al., 

1998; Stern et al., 1998; Bidleman et al., 1999, 2000, 2001; Muir et al., 2001). 

Concentrations of SCCPs (C10–13, 52% chlorine) ranged from 11 to 17 μg/kg in human 

breast milk in Canada (Tomy, 1997). Analyses were carried out by HRGC/ECNI-HRMS. 

No additional details were reported.2 

Muir et al. (1999) analysed whole fish samples for SCCPs (C10–13) and detected 2630 

ng/g (wet weight) in carp from Hamilton Harbour, 58.8 ng/g (wet weight) in lake trout 

from Niagara-on-the-Lake and 72.6 ng/g (wet weight) in lake trout from Port Credit. The 

quantification was by GC/ECNI-HRMS. Lower concentrations were reported in an 

earlier study (Muir et al., 1996). 

In a market basket survey (KAN-DO Office and Pesticides Team, 1995)3 of 234 ready-toeat foods, which represented approximately 5000 food types in American diets, 

“Chlorowax 500C”4 was detected once, in enriched white bread, at a concentration of 

0.13 μg/g. Food items were screened by gas or liquid chromatography using ion-selective 

detectors. Findings were confirmed by unspecified analysis. 

Concentrations of SCCPs have been identified in blubber of aquatic mammals such as 

ringed seal, beluga and walrus (Tomy et al., 20005

; Bennie et al., 20006). The samples 

were from animals in Greenland, the Canadian Arctic and the St. Lawrence River. A 

mean concentration of 46 100 ng/g (n = 15) was reported for beluga from the St. 

Lawrence River/Gulf of St. Lawrence. Concentrations in ringed seals from Ellesmere 

Island ranged from 370 to 770 ng/g. Jansson et al. (1993) detected SCCPs in biota in 

 

2

 These data were identified in a secondary source and were originally reported in a Ph.D. thesis. 

3

 Reported as a summary of results from 1982 to 1991. 

4

 The average molecular formula for Chlorowax 500C is C12H19Cl7, with 60–65% chlorine content (w/w) 

(IPCS, 1996). 

5

 Analysis by HRGC/ECNI-HRMS.

6

 Analysis by GC/low-resolution negative chemical ionization mass spectrometry. 
48

Sweden, including fish and both terrestrial and marine mammals. Analysis was by 

GC/MS. 

Data on concentrations of SCCPs in drinking water in Canada or elsewhere were not 

identified. The maximum concentration of SCCPs (C10–13, 50–70% chlorine) in the Red 

River, at a site remote from industrialized areas, was 0.05 μg/L (Tomy, 1997).7 Analyses 

were by HRGC/ECNI-HRMS. A lower concentration was reported in surface water from 

Lake Ontario (Muir et al., 2001). 

Concentrations of SCCPs in soil in Canada or elsewhere were not identified. The 

concentrations in surface sediment in harbours in Lake Ontario ranged from 5.9 to 290 

ng/g dry weight (Muir et al., 2001). Analyses were by HRGC/ECNI-HRMS.

Upper-bound estimates of intake of SCCPs for the general Canadian population and the 

assumptions upon which they are based are presented in Table 9. For each age group in 

the Canadian population, virtually all of the estimated intake is from food. The upperbound estimated intake of breast-fed infants was 1.7 μg/kg-bw per day, and that of 

formula-fed infants was 0.01 μg/kg-bw per day. For the remaining age groups, intakes 

ranged from 5.1 μg/kg bw per day for adults over 60 years of age to 26.0 μg/kg-bw per 

day for infants who were not formula fed (i.e., those being introduced to solid foods8

). 

Table 9. Upper-bounding estimated average daily intake of short-chain chlorinated 

paraffins by the population of Canada 

Estimated intake (μg/kg-bw per day) of short-chain chlorinated paraffins by various 

age groups 

0–6 months1

Route of 

exposure 

breast 

fed2

formul

a fed3

n

ot formula 

fed4

0.5–4 

years5

5–11 

years6

12–19 

years7

20–59 

years8

60+ 

years9

Ambient air10 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001 

Indoor air11 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001 

Drinking 

water12

0.001 0.001 0.001 <0.001 <0.001 <0.001 

Food13

1.7 0.005 

25.96 24.26 16.44 9.02 7.18 5.14 

Soil14 0.001 0.001 0.001 0.002 0.001 <0.001 <0.001 <0.001 

Total intake 1.7 0.01 25.97 24.26 16.44 9.02 7.18 5.14 

1

 Assumed to weigh 7.5 kg and to breathe 2.1 m3 per day (21 hours indoors, 3 hours outdoors) (EHD, 1998). 2

 Concentrations of SCCP (C10–13, 52% chlorine) ranged from 11 to 17 μg/kg in human breast milk in Canada 

(Tomy, 1997). No additional details were reported. These data were identified in a secondary source and were 

originally reported in a Ph.D. thesis. Assumed to consume 0.75 kg breast milk per day (EHD, 1998). 3

 For formula-fed infants, intake from water is synonymous with intake from food. Assumed to consume 0.8 L 

 

7

 These data were identified in a secondary source and were originally reported in a Ph.D. thesis. 

8

 Solid foods are introduced to approximately 50% of infants by 4 months of age and to 90% by 6 months 

of age (NHW, 1990).
49

reconstituted formula daily (EHD, 1998). 4

 Assumed to drink 0.2 L of water per day. Consumption of food groups reported by Health Canada (EHD, 1998). 

5 Assumed to weigh 15.5 kg, to breathe 9.3 m3 per day (21 hours indoors, 3 hours outdoors) and to drink 0.2 L of 

water per day. Consumption of food groups reported by Health Canada (EHD, 1998). 

6 Assumed to weigh 31.0 kg, to breathe 14.5 m3 per day (21 hours indoors, 3 hours outdoors) and to drink 0.4 L of 

water per day. Consumption of food groups reported by Health Canada (EHD, 1998). 

7 Assumed to weigh 59.4 kg, to breathe 15.8 m3 per day (21 hours indoors, 3 hours outdoors) and to drink 0.4 L of 

water per day. Consumption of food groups reported by Health Canada (EHD, 1998). 

8 Assumed to weigh 70.9 kg, to breathe 16.2 m3 per day (21 hours indoors, 3 hours outdoors) and to drink 0.4 L of 

water per day. Consumption of food groups reported by Health Canada (EHD, 1998). 

9 Assumed to weigh 72.0 kg, to breathe 14.3 m3 per day (21 hours indoors, 3 hours outdoors) and to drink 0.4 L of 

water per day. Consumption of food groups reported by Health Canada (EHD, 1998). 10 The maximum concentration of C10–C13 (60–70% chlorine) in gas-phase air samples collected every day over a 4-

month period in the summer of 1990 at Egbert, a rural site northwest of Toronto, was 924 pg/m3 (Muir et al., 

1999). 11 Concentrations of SCCP in indoor air in Canada or elsewhere were not identified. The value used for calculating 

intake here is the above concentration identified for ambient air (Muir et al., 1999). 12 Concentrations of SCCP in drinking water were not identified. The maximum concentration of SCCP (C10–13, 50–

70% chlorine) identified in the Red River, at a site remote from industrialized areas, was 0.05 μg/L (Tomy, 1997). 13 Estimates of intake from food are based upon concentrations in foods that are selected to represent the food groups 

addressed in calculating exposure to Priority Substances (EHD, 1998): 

 Dairy: 0.3 μg/g; mean of 13 samples of dairy products in U.K.; C10–20 (SCCP and MCCP) (Campbell and 

McConnell, 1980a) 

 Fats: 0.15 μg/g; mean of 6 samples of vegetable oils and derivatives; C10–20 (SCCP and MCCP) (Campbell and 

McConnell, 1980a) 

 Fruits: 0.025 μg/g; mean of 16 samples of fruits and vegetables in U.K.; C10–20 (SCCP and MCCP) (Campbell and 

McConnell, 1980a) 

 Vegetables: 0.025 μg/g; mean of 16 samples of fruits and vegetables in U.K.; C10–20 (SCCP and MCCP) 

(Campbell and McConnell, 1980a) 

 Cereal products: 0.13 μg/g; one reported concentration for “Chlorowax 500C” in enriched white bread in market 

basket survey carried out by U.S. Food and Drug Administration (KAN-DO Office and Pesticides Team, 1995); 

average molecular formula is C12H19Cl7, with 60–65% chlorine content (w/w) (IPCS, 1996) 

 Meat and poultry: 0.099 μg/g; 1 sample of bacon in U.K.; C10–20 (SCCP and MCCP) (Campbell and McConnell, 

1980b) 

 Fish: 2.630 μg/g (wet weight); analysis of whole samples of carp from Hamilton Harbour; C10–C13 (Muir et al., 

1999) 

 Eggs: no data identified 

 Foods primarily sugar: 0.025 μg/g; 1 sample of strawberry jam in U.K.; C10–20 (SCCP and MCCP) (Campbell and 

McConnell, 1980b) 

 Mixed dishes: no data identified 

 Nuts and seeds: no data identified 

 Soft drinks, alcohol, coffee, tea: 0.05 μg/g; detection limit in analyses of beverages in U.K. (Campbell and 

McConnell, 1980a) 

 

 Amounts of foods consumed on a daily basis by each age group are described by Health Canada (EHD, 1998). 

14 No data were identified on concentrations of SCCP in soil in Canada. The maximum concentration in surface 

sediment in harbours in Lake Ontario was 290 ng/g dry weight (Muir et al., 2001). 

Canadian data incorporated within this estimate include high-confidence values in fish 

(whole carp determined by GC/ECNI-HRMS) and data on breast milk, for which details 

of sampling and analysis were not reported. Estimated intake of SCCPs in fish represents 

up to 58% of the total daily intake. The intake from dairy products, which accounts for 

89.9% of the intake of infants not formula fed, is based upon limited sampling and 

analysis — considered semi-quantitative only — of dairy products in the United 

Kingdom, reported in 1980. Probably the most representative estimates of intake are 

those from cereals, which are based upon data reported in an American market basket 

survey, carried out from 1982 to 1991; however, intake from this foodstuff constitutes 
50

<0.1% of total estimated intake, and analytical methods were not specified. 

Intake of SCCPs by a potentially higher-exposure subgroup of Inuit for whom the 

primary source of food is subsistence hunting and fishing (Kuhnlein, 1989; Kinloch et al., 

1992) was also estimated, based on data on concentrations of SCCPs in blubber from 

marine mammals in Canada (Tomy et al. 2000) and less specific data (including both 

SCCPs and MCCPs) for terrestrial and marine mammals from Sweden (Jansson et al., 

1993). On the basis of these data, the estimated intake of an Inuit adult, namely 1.47 

μg/kg-bw per day, is well within the range of values estimated above for the general 

population. 

5.1.2 MCCPs 

MCCPs were detected by HRGC/low-resolution mass spectrometry (LRMS) in effluent 

(13 μg/L) from a chlorinated paraffin manufacturing plant in Canada in 1993, but not in 

surface water or sediment (Metcalfe-Smith et al., 1995). MCCPs were detected in three 

samples of carp from Hamilton Harbour in 1996 by low-resolution GC/MS (mean 0.393 

μg/g; range 0.276–0.563 μg/g) (Bennie et al., 2000). Similarly, MCCPs were detected in 

the homogenized (whole) samples of 10 trout collected from western Lake Ontario in 

1996 (mean 1.23 μg/g; range 0.257–4.39 μg/g) (Bennie et al., 2000). 

Upper-bounding estimates of intake for MCCPs and the assumptions on which they are 

based are presented in Table 10. For each age group, virtually all of the estimated intake 

is from food, which, in turn, is based almost entirely upon the limited data reported by 

Campbell and McConnell (1980a,b). The highest intake estimated (25.5 μg/kg-bw per 

day) was for infants not formula fed. 

Table 10. Upper-bounding estimated average daily intake of medium-chain chlorinated 

paraffins by the population of Canada 

Estimated intake (μg/kg-bw per day) of medium-chain chlorinated paraffins by 

various age groups 

0–6 months1

Route of 

exposure 

formula 

fed2

not 

formula 

fed3

6 months–

4 years4

5–11 

years5

12–19 

years6

20–59 

years7

60+ 

years8

Ambient air9 – – – – – – – 

Indoor air10 – – – – – – – 

Drinking 

water11

0.01 0.01 0.01 <0.01 <0.01 <0.01 

Food12

0.05 

25.48 18.48 11.64 6.3 4.69 3.47 

Soil13 0.01 0.01 0.02 0.01 <0.01 <0.01 <0.01 

Total intake 0.07 25.51 18.51 11.65 6.3 4.69 3.47 

1

 Assumed to weigh 7.5 kg and to breathe 2.1 m3 per day (21 hours indoors, 3 hours outdoors) (EHD, 1998). 2

 For formula-fed infants, intake from water is synonymous with intake from food. Assumed to consume 0.8 L 

reconstituted formula daily. No data on concentrations of MCCP in formula were identified for Canada. 3

 Assumed to drink 0.2 L of water per day. Consumption of food groups reported by Health Canada (EHD, 1998). 4

 Assumed to weigh 15.5 kg, to breathe 9.3 m3 per day (21 hours indoors, 3 hours outdoors), to ingest 100 mg of 
51

soil per day and to drink 0.2 L of water per day. Consumption of food groups reported in EHD (1998). 5

 Assumed to weigh 31.0 kg, to breathe 14.5 m3 per day (21 hours indoors, 3 hours outdoors), to ingest 65 mg of 

soil per day and to drink 0.4 L of water per day. Consumption of food groups reported in EHD (1998). 6

 Assumed to weigh 59.4 kg, to breathe 15.8 m3 per day (21 hours indoors, 3 hours outdoors), to ingest 30 mg of 

soil per day and to drink 0.4 L of water per day. Consumption of food groups reported in EHD (1998). 7

 Assumed to weigh 70.9 kg, to breathe 16.2 m3 per day (21 hours indoors, 3 hours outdoors), to ingest 30 mg of 

soil per day and to drink 0.4 L of water per day. Consumption of food groups reported in EHD (1998). 8

 Assumed to weigh 72.0 kg, to breathe 14.3 m3 per day (21 hours indoors, 3 hours outdoors), to ingest 30 mg of 

soil per day and to drink 0.4 L of water per day. Consumption of food groups reported in EHD (1998). 9

 Concentrations of MCCP in ambient air in Canada or elsewhere were not identified. 10 Concentrations of MCCP in indoor air in Canada or elsewhere were not identified. 11 Concentrations of MCCP in Canadian drinking water were not identified. Intakes are based upon the limit of 

detection (0.5 μg/L) in a survey of drinking water in reservoirs in the U.K. (Campbell and McConnell, 1980a). 12 Estimates of intake from food are based upon concentrations in foods that are selected to represent the food groups 

addressed in calculating exposure to Priority Substances (EHD, 1998): 

 Dairy: 0.3 μg/g; mean of 13 samples of dairy products in U.K.; C10–20 (SCCP and MCCP) (Campbell and 

McConnell, 1980a) 

 Fats: 0.15 μg/g; mean of 6 samples of vegetable oils and derivatives; C10–20 (SCCP and MCCP) (Campbell and 

McConnell, 1980a) 

 Fruits: 0.025 μg/g; mean of 16 samples of fruits and vegetables in U.K.; C10–20 (SCCP and MCCP) (Campbell and 

McConnell, 1980a) 

 Vegetables: 0.025 μg/g; mean of 16 samples of fruits and vegetables in U.K.; C10–20 (SCCP and MCCP) 

(Campbell and McConnell, 1980a) 

 Cereal products: 0.05 μg/g, detection limit in analyses of corn flakes in U.K. (Campbell and McConnell, 1980b) 

 Meat and poultry: 0.099 μg/g; 1 sample of bacon in U.K.; C10–20 (SCCP and MCCP) (Campbell and McConnell, 

1980b) 

 Fish: 1.23 μg/g (wet weight); mean of 10 samples of whole trout from western Lake Ontario (Bennie et al., 2000) 

 Eggs: no data identified 

 Foods primarily sugar: 0.025 μg/g; 1 sample of strawberry jam in U.K.; C10–20 (SCCP and MCCP) (Campbell and 

McConnell, 1980b) 

 Mixed dishes: no data identified 

 Nuts and seeds: no data identified 

 Soft drinks, alcohol, coffee, tea: 0.05 μg/g; detection limit in analyses of beverages in U.K. (Campbell and 

McConnell, 1980a) 

 

 Amounts of foods consumed on a daily basis by each age group are described by Health Canada (EHD, 1998). 

13 The value used for calculating intake from soil is the limit of quantification (3.5 μg/g) in a survey of sediment 

from the St. Lawrence River (Metcalfe-Smith et al., 1995). 

5.1.3 LCCPs 

Upper-bounding estimates of total intake of LCCPs and associated assumptions are 

presented in Table 11. As for SCCPs and MCCPs, for each age group, virtually all of the 

estimated intake is from food. The highest intake estimated (16.8 μg/kg-bw per day) was 

for infants not formula fed. In addition to the limitations of the analytical methodology 

noted previously, these estimates are further limited in that estimates for five of the eight 

food groups are based upon the limit of detection in that survey (Campbell and 

McConnell, 1980a,b). 

Table 11. Upper-bounding estimated average daily intake of long-chain chlorinated 

paraffins by the population of Canada

Estimated intake (μg/kg-bw per day) of long-chain chlorinated paraffins by 

various age groups 

Route of 

exposure 

0–6 months1

 6 months– 5–11 12–19 20–59 60+ years8
52

formula 

fed2

not 

formula 

fed3 

4years4 years5 years6 years7

Ambient air9 – – – – – – – 

Indoor air10 – – – – – – – 

Drinking 

water11

0.01 0.01 0.01 <0.01 <0.01 <0.01 

Food12

0.05 

16.81 9.66 5.61 3.04 2.12 1.73 

Soil13 0.01 0.01 0.02 0.01 <0.01 <0.01 <0.01 

Total intake 0.07 16.83 9.69 5.63 3.04 2.12 1.73 

1

 Assumed to weigh 7.5 kg and to breathe 2.1 m3 per day (21 hours indoors, 3 hours outdoors) (EHD, 1998). 2

 For formula-fed infants, intake from water is synonymous with intake from food. Assumed to consume 0.8 L 

reconstituted formula daily. No data on concentrations of LCCP in formula were identified for Canada. 3

 Assumed to drink 0.2 L of water per day. Consumption of food groups reported in EHD (1998). 4

 Assumed to weigh 15.5 kg, to breathe 9.3 m3 per day (21 hours indoors, 3 hours outdoors), to ingest 100 mg of 

soil per day and to drink 0.2 L of water per day. Consumption of food groups reported in EHD (1998). 5

 Assumed to weigh 31.0 kg, to breathe 14.5 m3 per day (21 hours indoors, 3 hours outdoors), to ingest 65 mg of 

soil per day and to drink 0.4 L of water per day. Consumption of food groups reported in EHD (1998). 6

 Assumed to weigh 59.4 kg, to breathe 15.8 m3 per day (21 hours indoors, 3 hours outdoors), to ingest 30 mg of 

soil per day and to drink 0.4 L of water per day. Consumption of food groups reported in EHD (1998). 7

 Assumed to weigh 70.9 kg, to breathe 16.2 m3 per day (21 hours indoors, 3 hours outdoors), to ingest 30 mg of 

soil per day and to drink 0.4 L of water per day. Consumption of food groups reported in EHD (1998). 8

 Assumed to weigh 72.0 kg, to breathe 14.3 m3 per day (21 hours indoors, 3 hours outdoors), to ingest 30 mg of 

soil per day and to drink 0.4 L of water per day. Consumption of food groups reported in EHD (1998). 9

 Concentrations of LCCP in ambient air in Canada or elsewhere were not identified. 

10 Concentrations of LCCP in indoor air in Canada or elsewhere were not identified. 11 Concentrations of LCCP in Canadian drinking water were not identified. Intakes are based upon the limit of 

detection (0.5 μg/L) in a survey of drinking water in reservoirs in U.K. (Campbell and McConnell, 1980a). 12 Estimates of intake from food are based upon concentrations in foods that are selected to represent the food groups 

addressed in calculating exposure to Priority Substances (EHD, 1998): 

 Dairy: 0.19 μg/g; 1 sample of cheese in U.K.; C20–30 (Campbell and McConnell, 1980a) 

 Fats: 0.05 μg/g; detection limit in analysis of 1 sample of lard in U.K.; C20–30 (Campbell and McConnell, 1980a) 

 Fruits: 0.025 μg/g; 1 sample of peach fruit in U.K.; C20–30 (Campbell and McConnell, 1980a) 

 Vegetables: 0.025 μg/g; 1 sample of potato crisps in U.K.; C20–30 (Campbell and McConnell, 1980a) 

 Cereal products: 0.05 μg/g, detection limit in analysis of corn flakes in U.K. (Campbell and McConnell, 1980b) 

 Meat and poultry: 0.05 μg/g; detection limit in analysis of 1 sample each of ox liver and beef in U.K.; C20–30

(Campbell and McConnell, 1980b) 

 Fish: no data identified 

 Eggs: no data identified 

 Foods primarily sugar: 0.05 μg/g; detection limit in analysis of 1 sample of strawberry jam in U.K.; C20–30

(Campbell and McConnell, 1980b) 

 Mixed dishes: no data identified 

 Nuts and seeds: no data identified 

 Soft drinks, alcohol, coffee, tea: 0.05 μg/g; detection limit in analysis of 1 sample each of beer and tea in U.K.; 

C20–30 (Campbell and McConnell, 1980b) 

 

 Amounts of foods consumed on a daily basis by each age group are described by Health Canada (EHD, 1998). 

13 The value used for calculating intake from soil is the maximum concentration (3.2 μg/g) reported in a survey of 

sediment in the U.K. (Campbell and McConnell, 1980a). 

5.2. Hazard characterization and dose–response analyses 

A limited number of studies on the toxicity of SCCPs have been reported in the period 

following release of the PSL1 assessment. Most of these studies were conducted to 
53

investigate the mode of action of carcinogenicity for the tumours observed in the NTP 

(1986a) bioassay, which were liver tumours in both sexes of rats and mice, kidney 

tumours in male, but not female, rats and thyroid tumours in rats and mice (females only). 

For several of these more recent studies, results have been reported in abstracts or 

summaries only: Elcombe et al. (1994) (abstract), Elcombe et al. (2000) (summary) and 

Warnasuriya et al. (2000) (abstract). For only one of the relevant investigations has a full 

published account been identified (Wyatt et al., 1993). While secondary accounts of 

(possibly) other studies investigating mode of action of tumour induction in assessments 

have been reported by the European Commission (2000), the U.S. National Research 

Council (U.S. NRC, 2000) and the Australian National Industrial Chemicals Notification 

and Assessment Scheme (NICNAS, 2001), they are not further considered here, owing to 

lack of availability or confirmation of subsequent publication (Jackson, 2001). 

Few data relevant to the assessment of the toxicity of either MCCPs or LCCPs were 

identified for the period to the release of the PSL1 assessment report. The following 

presentation is limited to those considered critical to hazard characterization or dose–

response analyses for effects in the general population and, hence, to assessment of 

“toxic” under Paragraph 64(c) of CEPA 1999. Other sources of non-critical data 

identified but not included were DuPont (1995), Kato and Kenne (1996) and Warngard 

(1996). 

In view of the absence of recent toxicological data that impact on critical aspects, the 

dose–response analyses for MCCPs and LCCPs presented here reflect primarily those 

developed in the PSL1 Assessment Report released under CEPA 1988. 

5.2.1 SCCPs 

A- Liver 

Increased liver weight, hepatocellular hypertrophy, peroxisomal proliferation and 

increased S-phase activity in hepatocytes were reported in Fischer 344 rats administered 

SCCPs for up to 90 days (presumably by gavage) at dose levels up to 1000 mg/kg-bw per 

day (Elcombe et al., 1994; abstract). Lower doses administered were not specified, and 

quantitative dose- or sex-specific data and analyses were not presented. 

Elcombe et al. (2000) administered Chlorowax 500C (C10–13; 58% chlorine) to male and 

female Fischer 344 rats by gavage in corn oil for up to 90 days, at dose levels of 0, 312 or 

625 mg/kg-bw per day. In both sexes, liver weight was increased, accompanied by 

peroxisomal proliferation (as indicated by an increase in cyanide-insensitive palmitoyl 

coenzyme A [CoA] oxidation) and increased thyroxine (T4)–uridine diphosphoglucose 

glucuronosyl transferase (UDPGGT). (The effects were, presumably, observed at both 

dose levels.) These effects were not observed in male Dunkin Hartley guinea pigs 

similarly administered 0, 500 or 1000 mg/kg-bw per day for 14 consecutive days. The 

numbers of animals exposed were not specified, and quantitative dose- or sex-specific 

data and analyses were not presented in this summary account. 
54

Wyatt et al. (1993) exposed groups of five male rats (Alpk:APfSD strain) each by gavage 

for 14 days to 0, 10, 50, 100, 250, 500 or 1000 mg/kg-bw per day to two SCCPs 

(Chlorowax 500C: C10–13, 58% chlorine; or Cereclor 56L, C10–13: 56% chlorine). For the 

58% chlorine SCCPs, both absolute and relative liver weights were significantly 

increased in a dose-related manner, at doses of 100 mg/kg-bw per day or greater. 

Peroxisomal fatty acid β-oxidation activity (indicated by palmitoyl CoA oxidation) was 

significantly increased at 250 mg/kg-bw per day and greater (irregular dose–response). 

For the 56% chlorine SCCPs, the pattern of response for absolute liver weight was 

irregular; however, relative liver weight was increased in a dose-related manner, 

significantly at 50 mg/kg-bw per day and greater. Palmitoyl CoA oxidation was 

significantly increased only at the highest dose. 

In similarly exposed male mice (Alpk:APfCD-1 strain), for the 58% chlorine SCCPs, 

there was a dose-related increase in relative liver weight and palmitoyl CoA oxidation, 

both significant at 250 mg/kg-bw per day and greater (Wyatt et al., 1993). For the 56% 

chlorine SCCPs, both absolute and relative liver weights were significantly increased in a 

dose-related manner at doses of 100 mg/kg-bw per day or greater. Palmitoyl CoA 

oxidation was significantly increased in a dose-related manner at 250 mg/kg-bw per day 

and greater. 

The only other relevant investigation identified was an in vitro study in which SCCPs 

inhibited gap junction intercellular communication in rat liver cells (Kato and Kenne, 

1996; Warngard et al., 1996). 

B-Kidney 

Increased proximal tubular cell eosinophilia (suggestive of a protein overload, but not 

necessarily α2u globulin) and regenerative focal basophilic tubules, as well as increased Sphase activity in the proximal tubular cells, were reported in male, but not female, rats 

administered up to 1000 mg SCCPs/kg-bw per day for up to 90 days (other dose levels 

were not specified) (Elcombe et al., 1994). These observations were reported in an 

abstract; neither quantitative data nor statistical analyses were presented. 

Elcombe et al. (2000) also investigated renal effects in F344 rats and guinea pigs 

administered 0, 312 or 625 mg SCCPs/kg-bw per day for up to 90 days. In the male rats 

only, there was chronic protein nephropathy, associated with regenerative hyperplasia 

and increased DNA synthesis (S-phase activity), presumably at both dose levels. There 

was “some limited evidence” for an involvement of α2u globulin. These changes were not 

observed in the guinea pigs. Again, neither quantitative data nor statistical analyses were 

presented in this summary account. 

Warnasuriya et al. (2000) exposed male and female rats by gavage for 28 days to 625 mg 

SCCPs (C12; 60% chlorine)/kg-bw per day. There was an increase in α2u globulin and cell 

proliferation in the kidney of males only. Data from individual rats indicated that 

increased cell proliferation was directly correlated with the increase in α2u globulin. Five 

different isoelectric isoforms of α2u globulin were identified by Western blotting in the 
55

control male kidney, and all five were increased in the treated males. These observations 

were reported in an abstract; neither quantitative data nor statistical analyses were 

presented. 

C- Thyroid 

Elcombe et al. (1994) reported that exposure of rats to SCCPs for up to 90 days resulted 

in induction of T4–glucuronosyl transferase activity, accompanied by a decrease in 

plasma T4 and an increase in thyroid stimulating hormone (TSH). Thyroid follicular cell 

hypertrophy and hyperplasia were also observed. Increased S-phase activity in the thyroid 

follicular cells was also reported. The maximum dose was 1000 mg/kg-bw per day; other 

dose levels were not specified. This study was reported as an abstract; neither quantitative 

data nor statistical analyses were presented. 

In male and female Fischer 344 rats exposed by gavage in corn oil to 0, 312 or 625 

mg/kg-bw per day for up to 90 days, there were decreases in plasma T4, increases in 

plasma TSH and thyroid follicular cell hypertrophy and hyperplasia in both sexes, 

changes that were not observed in male guinea pigs (Elcombe et al., 2000). Quantitative 

data and statistical analyses were not presented in this summary account. 

Gavage administration of 6.8 mg/kg-bw per day commercial C10–13 (71% chlorine) to 

female Sprague-Dawley rats for 14 days had no effect upon thyroid hormonal T4 levels or 

microsomal enzyme activity (Hallgren and Darnerud, 1998). 

In male rats (Alpk:APfSD strain) exposed by gavage for 14 days to two SCCPs 

(Chlorowax 500C: C10–13, 58% chlorine; or Cereclor 56L, C10–13: 56% chlorine), for 

which examination of thyroid function was restricted to the control and high-dose groups 

(1000 mg/kg-bw per day), both free and total T4 were significantly reduced, TSH was 

significantly increased and the capability of liver microsomes to glucuronidate T4 was 

significantly increased in exposed animals (Wyatt et al., 1993). No differences in levels 

of free or total triiodothyronine (T3) were observed for either SCCPs. A significant 

increase in glucuronosyl transferase activity with p-nitrophenol was observed only from 

microsomes from rats exposed to the C10–13 (58% chlorine) compound. 

5.2.2 MCCPs 

A subchronic dietary study with MCCPs in rats (Poon et al., 1995) was initiated by 

Health Canada in response to the research needs identified in the PSL1 assessment of 

chlorinated paraffins (Government of Canada, 1993a). Sprague-Dawley rats (10 per sex 

per group) were fed diets containing 0, 5, 50, 500 or 5000 ppm for 13 weeks. The dose 

levels calculated by the authors on the basis of weekly food consumption were 0, 0.4, 3.6, 

36 and 363 mg/kg-bw per day for males and 0, 0.4, 4.2, 42 and 419 mg/kg-bw per day for 

females. The protocol included serum biochemistry, hematology, hepatic enzyme 

activities, urinary enzyme activity, organ weights and histopathology. Mild, adaptive 

histological changes were detected in the liver of rats of both sexes at the two highest 

doses (LOEL = 36 mg/kg-bw per day) and in the thyroid of males at 36 mg/kg-bw per 
56

day and greater and of females at 4.2 mg/kg-bw per day and greater (NOAEL = 0.4 

mg/kg-bw per day). Minimal changes were observed in the renal proximal tubules of 

males at the highest dose and in the inner medulla of females at the two highest doses. 

5.2.3 LCCPs 

No critical data relevant to the assessment of the toxicity of LCCPs were identified for 

the period since the PSL1 assessment was released. 

5.3. Human health risk characterization 

5.3.1 SCCPs 

A- Hazard characterization 

Genotoxicity 

Requisite criteria for assessing the weight of evidence for hypothesized modes of 

induction of tumours addressed below include the criterion that SCCPs are not DNAreactive. Recent data on genotoxicity reported since the PSL1 assessment was released 

have not been identified. Limited available data reviewed within the PSL1 assessment 

indicated that SCCPs were clastogenic in in vitro assays, although they had not been 

clastogenic or mutagenic in a limited number of in vivo assays. 

Based on review of the available data, including two additional unpublished studies in 

which no increases in revertant colonies in five strains of Salmonella9

 and no increases in 

mutant colonies in Chinese hamster V79 cells10 were reported in the secondary account, it 

was concluded that “as a group, SCCPs are not mutagenic” (European Commission, 

2000). 

Liver 

It has been hypothesized that SCCPs cause liver tumours in rodents secondary to 

peroxisome proliferation. Peroxisome proliferation involves activation of a nuclear 

receptor in rodent liver, the peroxisome proliferator activated receptor, α isoform 

(PPARα). The activated PPARα interacts with regulatory elements of the DNA to initiate 

transcription of genes for increased peroxisomal enzyme activity and cell proliferation 

characterized by morphological and biochemical changes in the liver. These changes 

include increased liver weight through both hepatocyte hypertrophy and hyperplasia, 

 

9

 Cited by the European Commission (2000) as: Unpublished Report 86, Hoechst AG, Unpublished study, 

88.0099, 1988. 

10 Cited by the European Commission (2000) as: Unpublished Report 92, Hoechst AG, Unpublished study, 

87.1719, 1987.
57

increased number and size of peroxisomes, increased activity (up to 40-fold) of 

peroxisomal enzymes (especially those involved in peroxisomal fatty acid oxidation) and 

induction of microsomal fatty acid oxidation through the CYP4A subfamily of 

cytochrome P-450 isozymes. Minimum criteria for characterizing peroxisome 

proliferation are considered to include hepatomegaly, enhanced cell proliferation and an 

increase in hepatic acyl-CoA oxidase and/or palmitoyl-CoA oxidation levels. 

In the NTP bioassay (NTP, 1986a; Bucher et al., 1987) reported in the PSL1 assessment, 

increases in benign liver tumours were observed in both SCCPs-exposed rats (312 and 

625 mg/kg-bw per day) and mice (125 and 250 mg/kg-bw per day), with males of both 

species being considerably more sensitive. This pattern of induction of liver tumours by 

SCCPs is consistent with that for other peroxisome proliferating hepatocarcinogens, such 

as di(2-ethylhexyl)phthalate. 

Available data on the role of peroxisome proliferation in the etiology of hepatic effects 

and liver tumours induced by SCCPs are restricted to one study for which there is a 

published manuscript (Wyatt et al., 1993) and two investigations reported only in 

summary (Elcombe et al., 2000) or abstract form (Elcombe et al., 1994). Significant, 

dose-related increases in both absolute and relative liver weights accompanied at higher 

doses by increases in palmitoyl CoA oxidation in male Alpk:APfSD rats and 

Alpk:APfCD-1 mice exposed to two SCCPs, reported by Wyatt et al. (1993), are 

consistent with the observations in rats of Elcombe et al. (1994, 2000). Also, to the extent 

to which the more recent and better-documented study of Wyatt et al. (1993), with more 

extensive characterization of dose–response, can be compared with the earlier 

investigations of Elcombe et al. (1994, 2000), for which only summary reports are 

available, observations on dose–response for increases in liver weight and palmitoyl CoA 

oxidation in rats in these investigations are also consistent (increases in relative liver 

weight in rats were significant at ≥50 mg/kg-bw per day and palmitoyl CoA oxidation at 

≥250 mg/kg-bw per day; comparable values for mice were 100 mg/kg-bw per day and 

250 mg/kg-bw per day). 

Therefore, although characterization of exposure–response was limited in the NTP 

bioassay to only two dose levels, evidence to date indicates that tumours in both rats and 

mice occur only at doses at which peroxisome proliferation and associated morphological 

and biochemical effects have been observed in shorter-term studies (Wyatt et al., 1993; 

Elcombe et al., 1994, 2000). 

Additional weight of evidence for concordance might have been afforded through 

consideration of sex-related differences in peroxisome proliferation in shorter-term 

mechanistic studies. Unfortunately, this aspect was not investigated in the well-reported 

study by Wyatt et al. (1993) in which only male rats and mice were exposed; moreover, 

the limited extent of reporting in Elcombe et al. (1994, 2000) precludes consideration of 

relevant data in this context, if such data were, indeed, collected. Recovery studies would 

also have been informative, since peroxisome proliferation is initiated rapidly after 

treatment with a proliferator begins, attains a maximal response in a few weeks and is 

maintained only in the continued presence of the proliferator. Consistent with a receptor-
58

mediated response, the process is reversible. 

While there have been no carcinogenesis bioassays for SCCPs in species other than rats 

and mice, the variation in species sensitivity to peroxisome proliferation reported by 

Elcombe et al. (2000) is consistent with that observed for other peroxisome proliferators. 

Rats and mice are uniquely responsive to the morphological and biochemical effects of 

peroxisome proliferators, while Syrian hamsters exhibit intermediate responsiveness. 

This is consistent with marked interspecies variations in the expression of PPARα. 

Additional published documentation of existing relevant studies is desirable. Also, 

investigation of additional aspects of concordance would strengthen the weight of 

evidence for causality for the purported association between peroxisome proliferation and 

liver tumours induced by SCCPs. However, although there are limitations of the 

identified information, data are strongly suggestive that peroxisome proliferation plays a 

role in the etiology of liver damage and hepatic tumours associated with exposure to 

SCCPs. Although additional evidence for the weight of causality for liver tumours is 

desirable, a TDI based on hepatic effects in experimental animals is considered to be 

protective for carcinogenicity. 

Kidney 

It has been hypothesized that the kidney tumours observed following exposure of male 

rats to SCCPs are a species- and sex-specific response attributable to α2u globulin 

nephropathy and hence not relevant to humans. This mode of induction of renal tumours, 

which is relatively well characterized, involves binding to α2u globulin, a protein specific 

to male rats. This binding renders the protein more resistant to proteolytic degradation, 

which causes its accumulation in renal proximal tubule cells (manifested as hyaline 

droplets on histopathological examination), resulting in cell death and regenerative 

proliferation. Sustained cell proliferation leads to a low but significant incidence of renal 

tubular tumours. 

Minimum criteria for establishment of α2u globulin nephropathy as a basis for tumour 

development include lack of genotoxicity and observation of requisite precursor lesions 

and tumours in male rats only. Confirmation of requisite precursor lesions is based not 

only on histopathological observations such as excessive accumulation of hyaline 

droplets in renal proximal tubule cells, subsequent cytotoxicity and single-cell necrosis of 

the tubular epithelium and sustained regenerative tubular cell proliferation in the presence 

of continued exposure, but also on explicit identification of the protein accumulating in 

tubule cells as α2u globulin, along with demonstrated reversible binding of the relevant 

chemical or metabolite to α2u globulin (U.S. EPA, 1991; IARC, 1999). 

In the NTP bioassay (NTP, 1986a; Bucher et al., 1987) reported in the PSL1 assessment, 

renal tubular cell adenomas were observed in male rats at both doses (312 and 625 

mg/kg-bw per day), although the increase was significant (p < 0.05) only at the lower 

dose. Characterization of exposure–response was limited, therefore, in the NTP bioassay 

to only two dose levels. 
59

Available data on the mode of induction of kidney tumours in male rats by SCCPs are 

restricted to three investigations reported only in summary or abstract format (Elcombe et 

al., 1994, 2000; Warnasuriya et al., 2000). In Elcombe et al. (1994, 2000), regenerative 

focal basophilic tubules and increased S-phase activity in the proximal tubular renal cells 

were observed in male, but not female rats and considered by the authors to constitute 

“limited evidence” of the role of α2u globulin. More recently, the presence of α2u globulin 

was confirmed using immunohistochemical techniques, although no details of 

methodology were provided (Warnasuriya et al., 2000). 

Owing to the inadequate characterization in abstracts of even administered doses, in some 

cases with quantitative data on effects and analyses not being reported, there is very 

limited documentation to serve as a basis for conclusion that renal tumours occur only at 

doses at which either chronic protein nephropathy associated with regenerative 

hyperplasia and increased DNA synthesis (Elcombe et al., 2000) or α2u globulin is 

observed (Warnasuriya et al., 2000). 

While information is strongly suggestive that the kidney tumours observed in male rats 

are attributable to hyaline droplet formation, a male rat-specific phenomenon not relevant 

to humans, additional published documentation of available studies is clearly desirable as 

a basis for consideration of the weight of evidence of mode of induction of kidney 

tumours. Although additional confirmation is desirable, a TDI based on renal effects in 

experimental animals is considered to be protective for carcinogenicity. 

Thyroid 

There are a variety of non-DNA-reactive compounds that cause thyroid tumours in rats 

associated with decreased circulating thyroid hormone levels due to increased hepatic 

metabolism (particularly Phase II conjugating enzymes such as uridine diphosphate 

(UDP) glucuronosyl transferases [UDPGTs] and glutathione S-transferases) and 

clearance. These compounds induce hepatic glucuronidation of thyroid hormones and 

increase biliary excretion of the conjugated hormones, resulting in decreased circulating 

T3 and T4 levels. As a result of the hypothyroid state, TSH levels increase and cause 

sustained thyroid follicular cell hyperplasia, leading to tumour formation. 

While the basic physiology and feedback mechanisms of the hypothalamic–pituitary–

thyroid axis are qualitatively similar across species, quantitative differences make rodents 

more sensitive than humans to development of thyroid cancer for which the sole mode of 

action is thyroid–pituitary disruption (U.S. EPA, 1998). These include the lack of a highaffinity thyroid binding globulin in rats relative to humans (Dohler et al., 1979), which 

likely affects the turnover of the hormone. With a more rapid turnover of T4, there is a 

generalized increased activity of the pituitary–thyroid axis in rats compared with humans, 

which correlates with increased susceptibility to thyroid gland neoplasia. 

Minimum criteria for establishment of this mode of action as a basis for tumour 

development include evidence of increases in thyroid growth and hormonal changes (the 
60

latter including reduction in circulating serum T4 and T3 and an increase in TSH levels 

within days or a few weeks of exposure). Evidence of increases in thyroid growth is 

provided by measured increases in absolute or relative thyroid weight, histological 

indication of cellular hypertrophy and hyperplasia, morphometric determination of 

alteration in thyroid cellular components and changes in proliferation of follicular cells 

detected by DNA labelling or mitotic indices (U.S. EPA, 1998). 

In the NTP bioassay (NTP, 1986a; Bucher et al., 1987) reported in the PSL1 assessment, 

increases in follicular cell adenomas and carcinomas (combined) were observed in female 

rats only, at 312 and 625 mg/kg-bw per day, and in female mice only, at 250 mg/kg-bw 

per day. 

Available data relevant to assessment of the weight of evidence of induction of thyroid 

tumours in rats by SCCPs are limited to one study for which there is a published 

manuscript (Wyatt et al., 1993) and two investigations for which only a published 

summary report (Elcombe et al., 2000) or abstract (Elcombe et al., 1994) is available. In 

the study for which a complete account was published, effects on the thyroid were 

considered only in the control and highest dose groups; the administered dose for the 

latter was considerably greater than those in the NTP bioassay associated with thyroid 

tumours (i.e., 1000 mg/kg-bw per day versus 312 and 625 mg/kg-bw per day). In 

addition, in the abstract and summary accounts, quantitative data on effects or analyses 

were not presented. For example, Elcombe et al. (2000) reported only that male and 

female Fischer 344 rats were exposed by gavage in corn oil for up to 90 days at dose 

levels of 0, 312 or 625 mg/kg-bw per day and that “there were decreases in plasma 

thyroxine, increases in plasma TSH concentration and thyroid follicular cell hypertrophy 

and hyperplasia in both sexes.” There are extremely limited data, therefore, to serve as a 

basis for consideration of concordance of dose–response between thyroid tumour 

induction and precursor effects in shorter-term studies, such as thyroid growth and 

hormonal changes. In a single additional study for which a full account is available 

(Hallgren and Darnerud, 1998), the dose level at which effects on thyroid hormonal T4

levels or microsomal enzyme activity were not observed were much less than those 

administered in the NTP bioassay; as a result, these are not additionally meaningful in 

this context. 

As a result, although data from the studies reported by Elcombe et al. (1994, 2000) and 

Wyatt et al. (1993) fulfil the criteria for tumour induction by thyroid disruption in part, it 

should be noted that these data are insufficient as a basis for analysis of dose–response 

for concordance with that for thyroid tumours. Also, recovery in the absence of continued 

exposure has not been investigated. In view of the limitations of both reporting and dose–

response analyses, therefore, there is considerable uncertainty in attributing observed 

thyroid tumours to thyroid–pituitary disruption, to which rodents are more sensitive than 

humans.

B- Risk characterization 

Available data relevant to consideration of the weight of evidence for proposed modes of 
61

induction of liver, kidney and thyroid tumours associated with exposure to SCCPs, 

although limited, are suggestive that tolerable intakes that protect for non-neoplastic 

precursor effects will likely also be protective for cancer. However, owing principally to 

limited investigation of aspects such as recovery and inadequate documentation of 

relevant studies, there is considerable uncertainty in drawing this conclusion, particularly 

for the thyroid tumours. In recognition of this uncertainty, both neoplastic and nonneoplastic effects are considered here. 

IPCS (1996) derived a TDI of 100 μg/kg-bw per day for non-neoplastic effects of SCCPs 

on the basis of the lowest reported No-Observed-Effect Level (NOEL) of 10 mg/kg-bw 

per day in a 13-week study in rats (IRDC, 1984a). At the next higher dose in the critical 

study (100 mg/kg-bw per day), there were increases in liver and kidney weight and 

hypertrophy of the liver and thyroid. In IPCS (1996), an uncertainty factor of 100 was 

applied in the development of the TDI to account for interspecies variation (×10) and 

intraspecies variation (×10). The potential for progression of lesions following longerterm exposure was not explicitly addressed in the development of the TDI. This is 

balanced to some degree by the relatively large margin between the NOEL and the LOEL 

(10-fold) in the critical study and the minimal severity of the effects at the next higher 

concentration; however, there is some justification for considering a somewhat lower 

value for the TDI. 

On the basis of multistage modelling of the tumours with highest incidence 

(hepatocellular adenomas or carcinomas [combined] in male mice) in the carcinogenesis 

bioassay with SCCPs, IPCS (1996) also estimated the dose associated with a 5% increase 

in tumour incidence (Tumorigenic Dose05 [TD05]) to be 11 mg/kg-bw per day (amortized 

for period of administration). 

The upper-bound estimate of exposure for the age group with greatest exposure to SCCPs 

(i.e., 26 μg/kg-bw per day) is within the range of the IPCS (1996) TDI, for which there is 

some justification for considering a somewhat lower value, to take into account potential 

progression of the lesions in longer-term studies. 

The margin between the upper-bound estimate of exposure for the age group with 

greatest exposure to SCCPs and the Tumorigenic Dose (TD05) (i.e., 440) is also 

considered inadequate in view of the uncertainty concerning mode of induction of 

tumours. 

5.3.2 MCCPs 

A TDI developed on the basis of the NOAEL (0.4 mg/kg-bw per day) in the more recent 

subchronic study conducted by Health Canada (Poon et al., 1995) would be similar to 

that derived for the PSL1 assessment (i.e., 6 μg/kg-bw per day). 

Several of the highly uncertain bounding estimates of total daily intake of MCCPs from 

drinking water, food and soil for the general population of Canada exceed the TDI (6 

μg/kg-bw per day) for non-neoplastic effects. Indeed, for infants not formula fed, the 
62

total daily intake of MCCPs (i.e., 25.5 μg/kg-bw per day) exceeds the TDI by up to 4-

fold. 

5.3.3 LCCPs 

None of the highly uncertain bounding estimates of total daily intake of LCCPs from 

drinking water, food and soil for the general population of Canada exceeds the TDI (71 

μg/kg-bw per day) for non-neoplastic effects. However, for infants not formula fed, the 

total daily intake of LCCPs (16.8 μg/kg-bw per day) is within the same order of 

magnitude as the TDI. 

5.4. Uncertainties and degree of confidence in human health risk characterization 

There is low confidence in the upper-bounding estimates of exposure to all chlorinated 

paraffins. The estimates of intake for most age groups in the general Canadian population 

are based almost entirely upon limited sampling of foodstuffs in the United Kingdom, 

which were published in 1980. Methodology for analysis in this study is considered 

inadequate by present-day standards, and, as such, the data can be regarded at best as 

semi-quantitative. Reported concentrations represented both SCCPs and MCCPs, and, as 

a result, intake of the individual groups of chlorinated paraffins (SCCPs, MCCPs and 

LCCPs) from these sources has been overestimated. 

The estimates of intake for SCCPs are based in part upon the results of more recent 

surveys, for which methods of analysis were more reliable (i.e., quantification by 

GC/ECNI-HRMS). Concentrations of SCCPs determined by HRMS were available for 

ambient air, water and samples of carp from Hamilton Harbour (intake from fish 

represented 38–58% of estimated total intake of SCCPs, although fish accounts for, at 

most, 4% of the total daily intake of food across the six age groups). 

However, it is not possible to quantify the extent of overestimation of exposure based on 

the earlier, likely less selective analytical methodology, owing to lack of comparable 

data. Moreover, results based on analysis of the same samples by LRMS versus HRMS 

have been inconsistent, with levels of SCCPs being 1–2 orders of magnitude less for the 

latter in samples of whale blubber (Bennie et al., 2000; Tomy et al., 2000) and trout 

(Muir et al., 1999; Bennie et al., 2000) but slightly greater for the high-resolution 

analysis in carp (Muir et al., 1999; Bennie et al., 2000). 

There is minimal confidence in the upper-bounding estimates of exposure to MCCPs. 

These estimates are based in large part upon concentrations reported in a limited number 

of foodstuffs in the United Kingdom, which were published in 1980. More recent, 

although limited, data on concentrations in trout analysed by LRMS were included in the 

calculation of upper-bounding estimates. 

There is minimal confidence in the upper-bounding estimates of exposure to LCCPs. 

These estimates are based entirely upon concentrations reported in a limited number of 
63

foodstuffs in the United Kingdom, which were published in 1980. Furthermore, 

concentrations in foods were represented by the limits of detection for five of eight food 

groups in the calculations of daily intake. 

There is a low degree of confidence in the database of toxicological studies that serves as 

the basis for the assessment of the weight of evidence for mode of induction of tumours 

by SCCPs, for which only one published complete report (Wyatt et al., 1993) is available 

and for which it has not been possible to identify published accounts for reported prepublication manuscripts reviewed in previous assessments. Results in the only fully 

documented study provide most meaningful support for the purported role of peroxisome 

proliferation in induction of liver tumours in rats and mice. 

There is a moderate degree of confidence in the database of toxicological studies upon 

which the TDI for MCCPs is based, for which studies on chronic toxicity or 

carcinogenicity are lacking. The database for LCCPs is more complete, including a welldocumented carcinogenicity bioassay in rats and mice. 

 

6. CONCLUSIONS 

6.1. SCCPs 

On the basis of the available information, it is concluded that short-chain chlorinated 

paraffins (SCCPs) are entering, or may enter, the environment in a quantity or 

concentration or under conditions that: 

• have or may have an immediate or long-term harmful effect on the environment 

or its biological diversity; and, 

• constitute or may constitute a danger in Canada to human life or health. 

Therefore, it is concluded that short-chain chlorinated paraffins are “toxic” as defined 

under paragraphs 64 (a) and (c) of the Canadian Environmental Protection Act, 1999. 

6.2. MCCPs 

On the basis of the available information, it is concluded that medium-chain chlorinated 

paraffins (MCCPs) are entering, or may enter, the environment in quantities or 

concentrations or under conditions that: 

• have or may have an immediate or long-term harmful effect on the environment 

or its biological diversity;and, 

• constitute or may constitute a danger in Canada to human life or health. 

Therefore, it is concluded that medium-chain chlorinated paraffins are “toxic” as defined 

in paragraphs 64 (a) and (c) of the Canadian Environmental Protection Act, 1999.
64

6.3. LCCPs 

On the basis of the limited available data, it is concluded that long-chain chlorinated 

paraffins up to C20 are entering, or may enter, the environment in quantities or 

concentrations or under conditions that: 

• have or may have an immediate or long-term harmful effect on the environment 

or its biological diversity. 

Therefore, it is concluded that long-chain chlorinated paraffins up to C20 are “toxic” as 

defined in paragraphs 64 (a) of the Canadian Environmental Protection Act, 1999.

In addition, on the basis of the limited available data, it is concluded that long-chain 

chlorinated paraffins are entering, or may enter, the environment in quantities or 

concentrations or under conditions that: 

• constitute or may constitute a danger in Canada to human life or health. 

Therefore, it is concluded that long-chain chlorinated paraffins are “toxic” as defined in 

paragraph 64 (c) of the Canadian Environmental Protection Act, 1999.
65

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