## 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 20081 

### TABLE OF CONTENTS

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

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

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

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

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

2. IDENTITY AND PHYSICAL/C HEMICAL PROPERTIES........................................................ 14

2.1. I DENTITY

............................................................................................................................... ..... 14

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

2.2. P HYSICAL / CHEMICAL PROPERTIES .............................................................................................. 14

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

3.1. P RODUCTION

, IMPORTATION AND USE PATTERN ......................................................................... 15

3.2. R ELEASES TO THE ENVIRONMENT IN C ANADA ........................................................................... 16

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

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

4.1. E NVIRONMENTAL FATE .............................................................................................................. 17

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

4.1.2 MCCPs/L CCPs.................................................................................................................... . 17

4.2. P ERSISTENCE AND BIOACCUMULATION POTENTIAL .................................................................... 18

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

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

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

4.3. E NVIRONMENTAL C ONCENTRATIONS ......................................................................................... 31

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

4.3.2 Wastewater treatment effluents , sewage sludg e and soils .................................................... 32

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

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

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

4.4. E NVIRONMENTAL EFFECTS ......................................................................................................... 36

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

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

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

4.5. P OTENTIAL TO CAUSE ECOLOGICAL HARM ................................................................................. 39

4.6. U NCERTAINTIES ON THE ECOLOGICAL RISK ASSESSMENT ........................................................... 42

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

4.6.2 Persistence and Bioaccumulation Stat us and Risk Im plications .......................................... 43

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

5.1. P OPULATION EXPOSURE ............................................................................................................. 44

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

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

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

5.2. H AZARD CHARACTERIZATION AND DOSE

– RESPONSE ANALYSES ................................................ 52

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

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

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

5.3. H UMAN HEALTH RISK CHARACTERIZATION ................................................................................ 56

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

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

5.3.3 LCCPs.......................................................................................................................... ......... 622 5.4. U NCERTAINTIES AND DEGREE OF CONFIDENCE IN HUMAN HEALTH RISK CHARACTERIZATION ... 62

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

6.1. SCCP S

............................................................................................................................... ........ 63

6.2. MCCP S

............................................................................................................................... ....... 63

6.3. LCCP S

............................................................................................................................... ........ 64

REFERENCES ..................................................................................................................... ...................... 653 LIST OF TABLES

T ABLE 1. R ANGE OF PHYSICAL PROPERTIES OF CP S CONGENERS .................................................................. 14

T ABLE 2. S UMMARY OF PERSISTENCE AND BIOACCUMULATION INFORMATION ON SCCP S . .......................... 18

T ABLE 3. S UMMARY OF PERSISTENCE AND BIOACCUMULATION INFORMATION ON MCCP S .......................... 22

T ABLE 4: S UMMARY OF PERSISTENCE AND BIOACCUMULATION INFORMATION ON C18-20 LCCP S . ............. 25

T ABLE 5. S UMMARY OF PERSISTENCE AND BIOACCUMULATION INFORMATION ON C >20 LIQUID LCCP S ........ 28

T ABLE 6. S UMMARY OF PERSISTENCE AND BIOACCUMULATION INFORMATION ON C >20 SOLID LCCP S . ........ 30

T ABLE 7. L IST OF E STIMATED E XPOSURE V ALUES (EEV), C RITICAL T OXICITY V ALUES (CTV), A SSESSMENT

F ACTORS (AF), AND E STIMATED N O E XPOSURE V ALUES (ENEV) USED IN THE CALCULATION OF R ISK

Q UOTIENTS (RQ) FOR SCCP S

, MCCP S

, C18-20 LIQUID LCCP S

, C>20 LIQUID LCCP S AND C>20

SOLID LCCP S

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

T ABLE 8. C ONCENTRATIONS OF SHORT

- CHAIN

, MEDIUM

- CHAIN AND LONG

- CHAIN CHLORINATED PARAFFINS

IN FOODSTUFFS

............................................................................................................................... ..... 44

T ABLE 9. U PPER

- BOUNDING ESTIMATED AVERAGE DAILY INTAKE OF SHORT

- CHAIN CHLORINATED PARAFFINS

BY THE POPULATION OF C ANADA ........................................................................................................ 48

T ABLE 10. U PPER

- BOUNDING ESTIMATED AVERAGE DAILY INTAKE OF MEDIUM

- CHAIN CHLORINATED

PARAFFINS BY THE POPULATION OF C ANADA ....................................................................................... 50

T ABLE 11. U PPER

- BOUNDING ESTIMATED AVERAGE DAILY INTAKE OF LONG

- CHAIN CHLORINATED PARAFFINS

BY THE POPULATION OF C ANADA ........................................................................................................ 514 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 Envi ronmental Protection Act

CEPA 1999 Canadian Envir onmental Protection Act, 1999 CoA coenzyme A CP chlorinated paraffin

CPIA Chlorinated Paraffi ns 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 i on 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-resoluti on gas chromatography HRMS high-resolution mass spectrometry IC50 median inhibitory concentration KAW air–water partition coefficien t (or unitless Henry’s law constant) kg-bw kilogram body weight KOA octanol–air partition coefficient KOC organic carbon sorption coeffici ent (or organic carbon–water partition coefficient) Kow octanol/water partition coefficient

LC50 median lethal concentration LCCP long-chain chlorinated paraffin5 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 Pollut ant 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 polychlorinat ed dibenzodioxin PCDF polychlorinated dibenzofuran

PNEC Predicted No-Effect Concentration POC particulate organic carbon

PPAR α peroxisome proliferat or 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 associ ated 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 diphosphoglucose6 UDPGGT uridine diphosphoglucos e glucuronosyl transferase UDPGT uridine diphosphate glucuronosyl transferase UNECE United Nations Economic Commission for Europe VP vapour pressure wt. weight WWTP wastewater treatment plant7 PREFACE Following the assessment of chlorinated pa raffins conducted under the first Priority Substances List (PSL1), availa ble data were considered inad equate to evaluate whether medium and long chain chlorinated paraffins we re considered to be “toxic” as defined

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

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

considered insufficient to conclude whether they were “toxic” unde r Paragraph 11(a) of CEPA 1988, this group of substances was c onsidered “toxic” to human health under Paragraph 11(c) of CEPA 1988. In updating th e assessments of medium and long chain chlorinated paraffins, included herein, more recent data on the e ffects 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 se parately for the environmental and health effects, but cross-reference d, where appropriate. Informati on relevant to assessment of effects on the environment is presented ini tially, followed by information relevant to

assessment of effects on human health.8 SYNOPSIS

Chlorinated paraffins (CPs) are chlorinated derivatives of n-alka nes 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 constitu te or may constitute a danger in Canada to human life or health as defined under para graph 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 ha rmful 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 ga ps relevant to the assessmen t 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 Ca nada (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 pr oportions of SCCPs a nd 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, thei r production capacity wa s reported to be 8.5 kilotonnes; however, there is no pr oduction 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 Canadi an 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 in clude spills, facility washdown and drum rinsing/disposal. CPs in meta lworking/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 Assessment9 SCCPs have been detected in the following Ca nadian environmental media: Arctic air, sediments from remote northern lakes, sewa ge treatment plant effluents from southern Ontario, surface water, sediments, plankton, inve rtebrates and fish from Lake Ontario and

marine mammals from the Canadian Arctic a nd the St. Lawrence River. SCCPs have also been detected in plankton, invertebrates a nd fish from Lake Michigan. MCCPs have

been detected in effluent from a CPs manuf acturing 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 a quatic 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 fac ility in Australia.

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

SCCPs have been detected in Arctic biot a 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 ye ars 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 simila r to those of MCCPs, LCCPs are expected to be persistent in sediments. Several bi odegradation studies have also found that biodegradation is hindered by in creasing carbon chain length. It is, therefore, concluded that SCCPs, MCCPs and LCCPs are persiste nt as defined in the Persistence and Bioaccumulation Regulations of CEPA 1999.

Bioaccumulation factors (BAFs) of 9900–51200 we t 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-deri ved 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 calculate d 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) ar e 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 fact ors (BMFs) greater

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

LCCP C 18 H 30 Cl 7 had BMF values greater than one in rainbow trout in la boratory studies,

and its half-life in rainbow trout was found to be simila r to those of recalcitrant compounds that are known to accumulate in or ganisms and magnify in food chains. In addition, SCCPs, MCCPs and LCCPs have oc tanol–water partition coefficient (log K OW ) values greater than five. Elevated concen trations of MCCPs have been measured in aquatic biota from the St. Lawrence estuary, th e United States and Australia. While all of the available published BCF studies for LCCP have found values <5000, some elevated10 concentrations of LCCPs have been found in ma rine benthic organisms in Australia. In addition the Gobas BAF Model pr edicts that 44% of liquid C 18-20 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 evid ence, it is concluded that SCCPs, MCCPs, and liquid C 18-20 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 Persis tence and Bioaccumulation Regulations.

The available toxicity data indicate that SCCPs , MCCPs and C 18-20 LCCPs may be harmful to certain aquatic species (e.g., Daphnia magna ) at low concentrations (e.g.,

chronic NOECs < 100 μ g/L).

SCCPs, MCCPs and C 18-20 LCCPs are considered to be both highly persistent and bioaccumulative. In addition, there is evidence that SCCPs, MCCPs and C 18-20 LCCPs are released into the Canadian environment and have the potential to cause harm to sensitive aquatic organisms at relatively lo w concentrations. Substances that are persistent remain in the environment fo r a long time, increasing the magnitude and duration of exposure. Releases of small amount s 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 expos ures, 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 ha s 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 tum ours related to peroxisome proliferation and thyroid tumours related to thyroid–pituitary disruption). Complete documentation of

available studies and consideration in additi onal investigations of the reversibility of precursor lesions in the absence of continue d exposure is lacking. However, reported

data on mode of induction of tumours in addi tion 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-cance r 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 potentia l carcinogenicity.

For MCCPs and LCCPs, critical data releva nt 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 Decem ber 2000. Based upon these semi-quantitative11 data, upper-bounding estimates of daily intake for MCCPs and LCCPs are within the same order of magnitude of, or exce ed, the TDIs for these substances.

Conclusion

Based on the information available, it is c oncluded that CPs containing up to twenty carbon atoms are entering, or may enter, the e nvironment 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 pa raffins constitute or may constitute a danger in Canada to human lif e or health. CPs containing up to twenty carbon atoms are persistent and bioaccumula tive as defined in the Persistence and Bioaccumulation Regulations.12 1. INTRODUCTION

Chlorinated paraffins (CPs) are chlorinated derivatives of n-alka nes with carbon chain lengths from 10 to 38 carbon atoms, and with varying chlorine cont ents. Commercial products, of which there are ove r 2000 (Serrone et al. 1987), are complex mixtures of homologues and isomers. CP s with carbon chains cont aining 10–13 carbon atoms (C 10–13 ) are termed “short”, those with 14–17 carbon atoms (C 14–17 ) are called “medium” and those having 18 or more carbon atoms ( ≥ C 18 ) are called “long”. This report addresses the shortchain chlorinated paraffins (SCCPs), the me dium-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 perfor med 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, b ecause 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 consider ed “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 se ction 11 of CEPA 1988. CEPA 1999 places more emphasis on pollution prevention, and mandates th e application of a weight of evidence approach and the precautionary principle when conducting and interpreting th e results of risk assessments of existing substances. In addition, CEPA 1999 provides for special consideration of persistent and bioaccumulativ e substances. Substances that are shown to be both persistent and bioaccumulative, th erefore, may be assessed using a more precautionary approach than is used for other substances.

In 1997, a Scientific Justif ication 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 th e document stated: “O n the basis of the information reviewed, it is concluded that short chain chlorinated paraffins are predominantly anthropogenic, persistent, bi oaccumulative, and CEPA -toxic. Short chain chlorinated paraffins satisfy all four cr iteria outlined in the Toxic Substances Management Policy to identify substances for management under Track 1. Therefore, short chain chlorinated paraffins are propos ed for management under Track 1 of the Policy.” During the public comment period on the Scientific Justification, the Chlorinated Paraffins Industry Association (CPIA) review ed the information cited in the document proposing to list SCCPs as a Track 1 substan ce. They argued that the evidence did not13 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 crit eria for persistence. In order to further examine the persistence of SCCPs and their pot ential 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 Environmen t Canada and at the Freshwater Institute of the Department of Fisheries and Ocean s (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 th e 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 literatu re was also reviewed for new exposure and toxicological data on CPs on human and non-huma n organisms in Canada and elsewhere.

This new information is considered in this assessment report. Da ta acquired prior to February 2001 and December 2000 were cons idered 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 obtai ned as of July 2007 were considered as part of the ecological follow-up assessme nt 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 Substa nces Division of Environment Canada and Health Canada, as well as the National Wate r Research Institute of Environment Canada. The content of this report has been subject ed 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 de tailed information presented in a supporting

document. For additional information the re ader should consult this document. This assessment report and the associated envi ronmental 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 (C 10–13 ), MCCPs (C 14–17 ) and the lower chlorinated LC CPs are mixtures that are viscous, colourless or yellowish dense oils. C >20 highly chlorinated alkanes are waxy solids at ambient temperatures. The averag e chlorine content by weight is 30–52% for

C 18–20 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 c ontain stabilizers, which include epoxidized esters and soya bean oils, erythritol, thym ol, urea, glycidyl ethers, acetonitriles and organic phosphates (European Commi ssion, 2000; Schenker 1979; Houghton 1993).

Various stabilizers are often a dded 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 pr imarily responsible for the large differences

that are evident in measurements and estim ates 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 physic al properties for SCCPs, MCCPs and 3

subclasses of LCCPs.

Table 1. Range of physical properties of CPs congeners

.

CP Class Vapour

pressure a (Pa) Henry’s Law

Constant (Pa·m 3 /mol) Water solubility

(μg/L) log K OW log K OA Log K OC Reference b

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, 1115 CP Class Vapour

pressure a (Pa) Henry’s Law

Constant (Pa·m 3 /mol) Water solubility

(μg/L) log K OW log K OA Log K OC Reference b

C 18-20 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 –

12.12 c
(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 give n 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) – K OW correlation; 7. Fisk et al. ( 1998a), measured data; 8. U.K. Environment Agency (2003), measured data; 9. Campbell and Mc Connell (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 K OW /HLC (unitless).

3. ENTRY INTO TH E 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 Chemi cals Inc. (formerly ICI Canada), Cornwall, Ontario, was the only Canadian producer of CP s. However, this plant was recently sold to Dover Chemical Corporation and it is currently not pr oducing chlorinated paraffins. This Cornwall plant previously produced M CCPs 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 correspond ing 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 Ca nada (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 th e 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 cu stomers, and users of CPs also reported on16 how they use CPs and the end uses for products that they formulate. There were some differences in reported usage volumes for CP s by distributors and users, but the uses generally were in agreement.

Nearly all usage of SCCPs wa s reported to be in metalwor king applications. Minor uses included use as a flame retardant in plastics and rubber.

The majority of uses for MCCPs as reporte d 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 retard ant in plastics or rubber.

The major uses of LCCPs are in lubricati ng additives, metalworki ng fluids and paints. Minor uses were in plastics and as flame reta rdants, engine oil, fabric adhesive and rock drilling fluid. Additional information on us es 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 s ource of CPs (U.K. Environment Agency 2003). Anthropogenic rele ases of CPs into the environment may occur during production, storage, transportati on, industrial and consumer usage of CPcontaining products, disposal a nd burning of waste, and land filling of products (Tomy et al. 1998a).

The two major sources of release of SCCP s, MCCPs and LCCPs into the Canadian environment are likely use in metalworking applications and manu facturing of products containing these CPs. The possible sources of releases to water from manufacturing include spills, facility wash-down and stor m water runoff. CPs in metalworking/metal cutting fluids may also be released into a quatic environments from drum disposal, carryoff and spent bath use (Government of Cana da 1993a). These releas es 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 mi ning, fluids and equipment used in oil and gas exploration, manufacture of seamless pi pe, metalworking and operation of turbines on ships (CPIA 2002; Environment Canada 2003b).

Landfilling is a major disposal route for pol ymeric 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 fo r centuries after disposal (IPCS 1996).

Polymer-incorporated CPs could also be releas ed during recycling of plastics, which may involve processes such as c hopping, grinding and washing. If re leased as dust from these17 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 cont aining CP polymers (PVC, other plastics, paints, sealants, etc.) (European Commission, 2000; U.K. Enviro nment 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 th e reporting criteria. Based on information collected by the NPRI, very small amounts of CPs are bein g released to the Canadian environment bycompanies that meet the NPRI reporting requirements. In 2002, small transfers of shortchain CPs for disposal to landfill (1.45 tonne s) and recycling by recovery of organics (1.94 tonnes) have been reporte d to the NPRI from only two companies, both located in Ontario. Less than 5 kg of rel eases and/or transfers of C 6–18 CPs have been reported by a third company in Ontario. In 2001, the same three companies mentioned above reported similar quantities of releases/t ransfers of CPs to the NPRI. It should be noted, however, that CPs are likely to be released from sour ces other than the indus trial sector s included in the NPRI, and releases to the Canadian e nvironment 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 s hown 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 (C 14-17 ) and a liquid C 18 LCCPs was estimated using the Equilibrium Criterion (E QC) Level III fugacity model of Trent University’s Canadian Environmental Model ling 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 orinputs and removal processes (advection, vola tilization, deposition, photol ysis, 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 pote ntial emissions of CP s mainly associated with landfills, land application of sewage sl udge and consumer uses. Results from the18 Level III EQC model suggest th at these CPs would achieve th eir highest concentrations in sediment and soil. Concentrations in water and air were extremely low for all compounds. The environmental residence time of the three C 16–18 CPs were estimated to be greater than 500 days comp ared with 250 days for the C 14 CP. However, these results should be viewed with caution because the degr adations rates, used as input parameters for the CPs, were highly uncertain. Simila r results were obtained using a Level III fugacity calculation with a C 14–17 MCCPs (U.K. Environment Agency 2003).

4.2. Persistence and bioaccumulation potential

When evaluating persistence in this secti on, 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 poten tial for long-range transport in this medium. Although soil is potentia lly an important compartment fo r CPs, there are too few data available to permit meaningful eval uation 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 bi oaccumulation information on SCCPs.

Medium or
parameter CEPA criteria 1 SCCPs Information

Air t 1/2 ≥ 2 days

or it is subject to atmospheric transport

from its source to a remote area Estimated t 1/2 of many SCCPs are ≥ 2 days

SCCPs have been detected in ai r, sediment and biota in the Arctic in the absence of significant sources, indicating long

range transport

Sediment t 1/2 ≥ 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 t 1/2 ≥ 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 K OW ≥ 5 4.39 – 8.69 (measured and modeled)

1 Government of Canada 2000

A- Persistence

Air and Long-Range Transport19 Estimated atmospheric half-lives for SCCP s based on reaction with hydroxyl radicals range from 0.81 to 10.5 days, using the default atmospheric hydroxyl radical

concentration of 1.5 × 10 6 molecules/cm 3 during sunlight hours in AOPWIN (v. 1.86) computer program (Meylan and Howard, 1993; Atkinson 1986, 1987). Using a lower hydroxyl radical concen tration of 5 × 10 5 molecules/cm 3 , which is generally used as a daily (24-hour) average in relatively unpollute d air in the EU, atmospheric half-lives ranged from 1.2 to 15.7 days. Tomy (1997) al so estimated atmospheric half-lives of greater than 2 days for the ma jor 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·m 3 /mol for C 10–12 congeners) that are in the range of VPs and HLCs for some persistent organic pollu tants that are known to undergo long-range atmospheric transport under the 1979 UNECE Convention on Long Range Transboundary Air Pollution (e.g., hexachlorocycl ohexane [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. Con centrations ranged from <1 to 8.5 pg/m 3 in gasphase samples. Borgen et al. (2000) measured SCCPs in Arctic air samples taken at Mt. Zeppelin, Svalbard, Norway, in 1999. Concen trations ranging from 9.0 to 57 pg/m 3 were detected. Borgen et al. (2002) found much hi gher SCCPs concentrations in air at Bear Island, a small isolated island between Sv albard and mainland Norway. Total SCCPs concentrations ranged from 1,800 to 10,600 pg/m 3 . SCCPs residues were found in the

surficial sediments in three remote Arct ic 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 c oncentration profiles for the Arctic marine mammals show a predominance of the shorter carbon chain length congeners, i.e., the C 10 and C 11 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 a nd 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). Concentrati ons ranged from 0.005 to 0.088 mg/kg wet weight.

Estimated atmospheric half-lives of many SC CPs are greater than 2 days for a large percentage (61% using hydroxyl radical concentration of 1.5 × 10 6 molecules/cm 3 and 83% using hydroxyl radical concentration of 5 × 10 5 molecules/cm 3 ) 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

· m 3 /mol, respectively.20 structures. Therefore, SCCPs meet the CEPA 1999 half-life criteri on for persistence in air specified in the Persistence and Bioaccu mulation 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 se diments in the absence of significant sources of SCCPs in this region suggests that thes e residues are present due to long-range atmospheric transport.

On the basis of the available information, it is concluded that estim ated atmospheric halflives of SCCPs exceed the criterion of 2 days and SCCPs are subject to long-range atmospheric transport. Hence, SCCPs are pe rsistent in air according to the criteria stipulated in the Persistence and Bi oaccumulation Regulations of CEPA 1999 (Government of Canada 2000).

Sediments and Soils

There is limited evidence for the biodeg radation or removal of SCCPs from soil following sewage sludge application. Ni cholls et al. (2001) did not detect

SCCPs/MCCPs in farm soils amended with sl udges containing mg/kg concentrations of CPs. However, worms living in these same so ils 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 th e highly chlorinated CPs, which included two SCCPs (60% and 70% chlo rine). On the other hand, Fi sk et al. (1998a) found that

two 14 C-labelled C 12 chloro-n-alkanes (56% and 69% ch lorine) were degraded at 12°C in aerobic sediments used for a study of the bioava ilability 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 biodegrad ation of SCCPs in both freshwater and marine sediments was undertaken by Thompson and Noble (2007). Using 14 C-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 car bon dioxide or methane production) over 98 days was determined. The mean half-lives for mineralization for a C 10 - 13 , 65% chlorine by weight product, calculated as the aver age for the two individual products, were estimated to be 1630 days in freshwater se diments and 450 days in marine sediments

under aerobic conditions. Little or no mineralizat ion 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-liv es 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 Lake21 DV09 (0.0176). The profile from Lake DV09 gene rally follows the pattern of historical usage of SCCPs (Stern and Evans 2003). Con centration profiles of SCCPs in sediments from Lake Winnipeg (Manitoba), Fox Lake (Yukon Territory), the west basin of Lake Ontario (Ontario) and Lake DV09 (Devon Isla nd, 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 a ny 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 Pers istence and Bioaccumulation Regulations of CEPA 1999 (Government of Canada 2000). However, the fact that SCCPs residues were detected in sediment cores dating back to th e 1940s at these locations is evidence that SCCPs can persist for more than 50 year s in subsurface anaerobic sediments. Environment Canada (2008) used first orde r decay equations in a back calculation

method to determine that SCCPs have a half-l ife in sediments longer than 1 year. The equation used for these calculations are st andard 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 availa ble information, it is thus concluded that SCCPs are persistent in sediments according to the criterion stipulated in the Persistence and Bioaccumulation Regulations of CE PA 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 filtere d 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). SCCP s 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 fo r fish estimated BAF values greater than 5000 for all possible SCCPs (Arnot and Gobas 2003).

Bioconcentration factors (BCFs) calculated fr om 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 factor s (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 biom agnification 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 like ly to have high BCF/BAF

values.

• Concentrations of SCCPs in fish colle cted 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 aver age 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 concentr ations suggest that SCCPs have the potential to bioaccumulate in aquatic organisms.

• Tomy (1997) found that SCCPs (ar ound 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 findi ngs are indicative of bioaccumulation through the food chain since food woul d be the major or only source of environmental exposure for the Inuit.

On the basis of the available information, it is concluded that SCCP s are bioaccumulative substances according to the criteria stipul ated in the Persistence and Bioaccumulation Regulations of CEPA 1999 (G overnment of Canada 2000).

4.2.2 MCCPs

Table 3 summarizes persistence and bioaccumulation information for MCCPs in comparison with criteria in the Persistence and Bioaccumu lation Regulations of CEPA 1999 (Government of Canada 2000).

Table 3. Summary of persistence and bi oaccumulation information on MCCPs.

Medium or
parameter CEPA criteria 1 MCCPs Information

Air t 1/2 ≥ 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 par ticles likely to be much slower

Sediment t 1/2 ≥ 1 year Back calculation using concentrations from sediment cores shows half-life >1 year

Soil t 1/2 ≥ 6 months Limited evidence for rapid biodegradation or removal following sludge applications23 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 K OW ≥ 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 ×

10 5 molecules/cm 3 . Half-lives for vapour phase MCCPs ranged from 2.7 to 7.1 days, with the longest half-lives for MCCP s 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·m 3 /mol for C 14-17 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 Conve ntion on Long Range Transboundary Air Pollution, such as lindane, heptachlor and mirex.

On the basis of the available information, it is concluded that estim ated atmospheric halflives of MCCPs exceed the criteri on of 2 days and hence are pers istent in air according to the criteria stipulated in th e Persistence and Bioaccumulati on Regulations of CEPA 1999 (Government of Canada 2000).

Sediment and Soil

There is limited evidence for the biodeg radation or removal of MCCPs from soil following sewage sludge application. Ni cholls et al. (2001) did not detect

SCCPs/MCCPs in farm soils amended with sl udges containing mg/kg concentrations of CPs. However, worms living in these same so ils 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 highe st concentrations estimated to have been deposited in 1972 (M uir et al. 2002). Envir onment 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 y ear. The equation used for these calculations are standard first order decay equations. Mo reover, the fact that MCCPs residues were detected in sediment cores dating back to th e 1970s at these locations is evidence that24 SCCPs can persist for more than 30 years in subsurface anaerobic sediments. Persistence in sediment is particularly important as Le vel III fugacity calculations show that MCCPs are expected to par tition primarily to sediment and soil.

On the basis of the available information, it is concluded that MCCP s are persistent in sediments according to the criteria stipulat ed in the Persistence and Bioaccumulation Regulations of CEPA 1999 (G overnment of Canada 2000).

B- Bioaccumulation

Bioaccumulation factors (BAFs) for MCCPs chai n 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 con centrations collected in 2001 using data in Houde et al. (2006). C 14-15 MCCPs were found in all components of this food chain and BAFs ranged from 9.99 x 10 6 to 7.15 x 10 8 . In addition, bioaccumulation factors (BAF) for 21 MCCPs congeners using the M odified Gobas BAF Model (assuming no metabolism) were all above the bioaccumulative criteria ( ≥ 5000 BAF) (Arnot and Gobas 2003).

Most of the laboratory-based BCF studi es conducted on aquatic organisms may underestimate the true BCF, because th e 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 complian ce with OECD guideline requirements. Estimated BCF values for common mussel, bl eak 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 Thomps on 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 (T hompson 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 factor s (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 C 14 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-normaliz ed 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 co nsidered in the Persistence and Bioaccumulation Regulations (Govern ment of Canada 2000), BMFs are25 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 li pid 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 c ontaining around double and crab meat

around six times the amount of ch loroparaffins 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 C 10–20 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). Thes e findings qualitatively indicate potential for bioaccumulation of MCCPs through the human food chain.

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

concluded that MCCPs are bioaccumulative substances according to the criteria stipulated in the Persistence and Bi oaccumulation 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 C 18-20 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 C 18-20 LCCPs.

Medium or
parameter CEPA criteria 1 C 18-20 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 K OW ≥ 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 ×

10 5 molecules/cm 3 . Half-lives for liquid LCCPs range d from 2.4 to 10.5 days, with many example structures having half-lives greater than 2 days. However, C 18-20 liquid LCCPs have very low partitioning to air.

C 18-20 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 organi c pollutants that are known to undergo longrange atmospheric transport under th e 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 estim ated atmospheric halflives of C 18-20 liquid LCCPs exceed the criterion of 2 da ys and hence are persistent in air according to the criteria stipulat ed 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 th e CEPA criteria. However, given that both SCCPs and MCCPs are expected to be persiste nt in sediment (half lives > 1 year), and that resistance to microbial degradation ha s been observed to ge nerally 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 info rmation, it is concluded that C 18–20 liquid LCCPs are persistent in sediments according to the cr iteria 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%) C 18-20 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 appl icable for LCCPs, as they are simple hydrophobic and persistent chemicals.

On the other hand, BCF values for C 18-26 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 va lues are below the BCF criterion of 500027 (Government of Canada 2000). However, th is 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 complia nce with OECD guidelines. As well, the study did not indicate if steady state was reach ed during the uptake phase of the test.

Other evidence that LCCPs are bioaccumulative is as follows:

• Reported log Kow values for C 18–20 liquid LCCPs range from 7.34 – 7.57 (Table 1).

• Biomagnification factors (BMFs) were de termined by Fisk et al (2000) in a dietary accumulation study involving j uvenile rainbow trout exposed to

C 18 H 31 Cl 7 . Lipid normalized BMFs ra nged from 2.1 to 2.8. While biomagnifications factors (BMF) are not a criterion considered in the Persistence and Bioaccumulation Regulat ions (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 C 18 H 31 Cl 7 has similar biotransformation halflives in rainbow trout compared to half -lives of recalcitrant organochlorines (Fisk et al. 1998c). This suggests limite d biotransformation or metabolism.

• Limited biotransformation of L CCPs was also observed during an uptake/elimination study with bleak. Bengtsson and Baumann-Ofstad (1982)

found that a C 18–26 LCCPs had a low uptake efficiency, but 50% of this compound remained in the fish tissue s after a 316-day elimination period, which suggests that some of the LCCP s isomers in this formulation are bioaccumulative (Bengtsson and Baumann-Ofstad 1982).

• Elevated levels of C 18–29 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 th e amount of chloroparaffins found in the most contaminated sediment sample.” However it is unclear if

bioaccumulation of C 18-20 or C >20 congeners was responsible for the elevated

concentrations.

On the basis of the available information, a nd in particular the BA F model and empirical BMF estimates, it is concluded that C 18–20 liquid LCCPs are bioaccumulative substances according to the criteria stipul ated 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 CE PA 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 t 1/2 ≥ 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 t 1/2 ≥ 1 year Unknown, but ha lf life likely > 1 year

Soil t 1/2 ≥ 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 K OW ≥ 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 ×

10 5 molecules/cm 3 . Half-lives for liquid LCCPs range d 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 estim ated atmospheric halflives of C >20 liquid LCCPs exceed the criterion of 2 days and hence are persistent in air according to the criteria stipul ated 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 th e CEPA criteria. However, given that both SCCPs and MCCPs are expected to be persiste nt in sediment (half lives > 1 year), and that resistance to microbial degradation ha s been observed to ge nerally 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 info rmation, it is concluded that C >20 liquid LCCPs are persistent in sediments according to the cr iteria stipulated in the Persistence and Bioaccumulation Regulations of CEPA 1999 (Government of Canada 2000).

B- Bioaccumulation

Although C >20 liquid LCCPs may have some potentia l 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 C 18-26 liquid LCCPs in bleak, and 18-1158 for liquid C >20 LCCPs in rainbow trout and co mmon mussel (Madeley and Maddock 1983b; Bengtsson et al. 1979; Ma deley and Thompson 1983b; U.K. Environment Agency 2001). However, these values may underest imate the true BCF values, because the studies were performed at LCCPs concentratio ns 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 th e uptake phase of the tests. BCF values for these species were below th e 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 L CCPs was also observed during an uptake/elimination study with bleak. Bengtsson and Baumann-Ofstad (1982)

found that a C 18–26 LCCPs had a low uptake efficiency, but 50% of this compound remained in the fish tissue s after a 316-day elimination period, which suggests that some of the LCCP s isomers in this formulation are bioaccumulative (Bengtsson and Baumann-Ofstad 1982).

• Elevated levels of C 18–29 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 th e amount of chloroparaffins found in the most contaminated sediment sample.” However it is unclear if

bioaccumulation of C 18-20 or C >20 congeners was responsible for the elevated

concentrations.

• C 20–30 CPs were detected in fat and live r of some postmortem human tissues from the United Kingdom at concentr ations between 0.080 and 3.5 mg/kg wet wt. These findings qualitatively indicat e 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 stipul ated 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 CE PA 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 t 1/2 ≥ 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 t 1/2 ≥ 1 year Unknown, but ha lf life likely > 1 year

Soil t 1/2 ≥ 6 months Unknown

BAF ≥ 5000 Low accumulation by salmon; poor absorption and high excretion via feces by rats; M odified Gobas Model predicts

BAF <5000

BCF ≥ 5000 Laboratory BCFs <5000; BCF probably underestimated due to CP concentrations exceeding solubility

Log K OW ≥ 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 ×

10 5 molecules/cm 3 . Half-lives for solid C 18-25 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 estim ated atmospheric halflives of C >20 solid exceed the criterion of 2 days a nd hence are persistent in air according to the criteria stipulated in the Persiste nce 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 pers istent in sediment (half lives > 1 year),

and that resistance to microbial degradation has been observed to generally increase with31 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 info rmation, it is concluded that C >20 solid LCCPs are persistent in sediments according to the cr iteria 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 we ight 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 Maddoc k 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 criteri on of 5000 (Madeley and Maddock 1983b,c; Bengtsson et al. 1979; Made ley 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% ch lorine LCCPs by juvenile Atlantic salmon fed a diet that had high CP con centrations (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 fe ces and poor absorption of the LCCPs, Section 4.4.3.2, supporting document (E nvironment Canada, 2008).. No BMF

data exist for C >20 solid LCCPs.

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

concluded that C >20 solid LCCPs are not bioaccumula tive substances according to the criteria stipulated in th e Persistence and Bioaccumula tion Regulations of CEPA 1999 (Government of Canada 2000).

4.3. Environmental Concentrations32 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 lim ited. Due to the non- volatile and hydrophobic characteristics of these CP groups, the major ity of results are for sediments and sewage sludges.

Data presented in this section focus on Cana dian concentrations. In situations where Canadian data are lacking or are few, con centrations measured in other countries are presented. Additional information on ambi ent concentrations may be found in the Supporting Document (Envi ronment 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/m 3 (Tomy 1997; Tomy et. al. 1998a). Concentr ations of SCCPs over Lake Ontario in 1999 and 2000 ranged from 120 to 1,510 pg/m 3 (Muir et al. 2001).

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

Canada or elsewhere.

4.3.2 Wastewater treatment effluent s, sewage sludge and soils

SCCPs were detected in wastew ater effluents in Canada, th e United States and Germany. SCCPs were detected in all eight sewage tr eatment plant final effluents from southern Ontario, Canada, sampled in 1996. Total SCCPs (dissolved and particulate C 10-13 ) 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 concentrati on data are available for MCCPs and LCCPs,

either in Canada or elsewhere.

Concentrations of CPs have also been detect ed in sewage sludge in several European

countries and the United States. Nicho lls 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, Steven s et al. (2002) found SCCPs concentrations ranging from 6.9 to 200 mg/kg dry wt. in se wage sludge from 14 WWTPs in the United Kingdom. Highest concentrati ons of SCCPs were in sludge from industrial catchments. However, a rural catchment with zero indus trial effluent had si gnificant levels (590 mg/kg) of total SCCPs/MCCPs in sludge (Steve ns et al. 2002). Tota l 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). Agri cultural soils may also be a potentially major reservoir of CPs due to sewage sludge ap plication (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 Ca nada and the United Kingdom. Low levels of dissolved total (C 10-13 ) SCCPs were measured in west ern 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 tota l 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 larg e 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-S mith et al. (1995) reported C 14–17 MCCPs concentrations in a 24-hour composite sample of effluent from the only manufacturing plant in Canada, ICI Canada (now PCI Cana da), 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 ri vers, 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 co llected in urban/industrial areas. Levels of MCCPs have

been measured at several sites in Germ any (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 si tes 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 C 20–30 , 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 concentr ation of 620 ng/L part iculates (<50 ng/L dissolved) was found in water from Sugar Cr eek, 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 fr om 0.008 to 0.176 mg/kg dry wt. (Tomy et al. 1999; Stern and Evans 2003). Tomy et al. (1997) measured SCCPs at concentrations34 around 0.245 mg/kg dry wt. in sediment from the mouth of the Detroit River at Lake Erie

and Middle Sister Island in we stern Lake Erie, in 1995. SCCPs were also detected in all surface sediment samples from harbour area s 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 indus trialized site (Windermere Basin, Hamilton Harbour), which has well-documented heavy metal, PAH and PCB contamination. Similarly, Marvin et al. (2003) reported a S CCPs concentration of 0.410 mg/kg dry wt. in

Lake Ontario sediments near an industriali zed area. SCCPs were detected in all 26 samples from Lake Ontario, and the aver age SCCPs concentration was 0.049 mg/kg dry wt., which is much higher than sediment c oncentrations reported for lakes (Yaya, DV09, Hazen, Nipigon) influenced primarily by atmo spheric 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, Switzerla nd, 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. Lawren ce River downstream of a CP manufacturing plant. Tomy and Stern (1999) reported concentrations of C 14–17 MCCPs of 0.068 mg/kg dry wt. in sediment samples collected near the mouth of the Detr oit River in western Lake Erie. Muir et al. (2002) reported concentrations of total MCCPs in a sediment core

from Lake St. Francis downstream of Corn wall, 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 Kingdo m. 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 sewa ge treatment plants in the United Kingdom (Nicholls et al. 2001).

No LCCPs were measured in sediments in Ca nada, but they have been detected in the United States, Australia and Ge rmany near CP manufacturing pl ants. Concentrations of LCCPs in these countries ranged from 0.0081 to 170 mg/kg dry wt. (R otard 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 No rth and Baltic Seas. Mu ir et al. (2001) and Houde et al. (2006) measured SCCPs in fish collected from Lake Ontario and Lake Michigan, between 1996 and 2001. Concentrat ions of total SCCPs ranged from 0.0046 to 2.63 mg/kg wet wt. The highest concentra tion 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 c oncentrations 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 fr om 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 concentra tion in fish measured in Canada was 0.904 mg/kg weight wt. for catfish in th e Detroit River (Tomy and Stern 1999).

Murray et al. (1988) measur ed concentrations of C 20–30 , 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 C 18–29 LCCPs in marine mussels and crabs (9.3 a nd 14.3 mg/kg lipid wt., re spectively) 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 bel uga whales from northwest Greenland, Sanikiluaq (Hudson Bay), Pangnirtung (C umberland Sound), Kimmirut and the Mackenzie Delta; and in walrus from north west 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 blubbe r in the St-Lawrence ranged from 1.8 – 80.0 mg/kg weight wt. (Bennie et al . 2000). However, results ob tained by Bennie et al. (2000) may not be reliable due to interf erences 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) re ported CP concentrations (unspecified chain length) in rabbit (Revingeshed, Skåne), moos e (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 th e 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 C 10–20 CPs in birds in the United Kingdom. The C 10–20 levels were likely to be dominated by contributions from the

SCCPs and MCCPs. Concentrations of C 10–20 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 C 20–

30 CPs ranged from not detected to 1.5 mg/kg we ight 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 from36 the seabirds, little auk ( Alle alle ) and kittiwake ( Rissa tridactyla ) collected at Bear Island (European Arctic). Concentrations be tween 0.005 and 0.088 mg/kg wet weight were measured. Reth et al. (2006) de termined the concentration of C 14-15 MCCPs in seabirds from the European Arctic. Concentrati ons ranged from 0.005 to 0.370 mg/kg wet wt.

4.4. Environmental effects

Overall, toxicity studies are few for effect s 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 M CCPs. In particular, the acute and chronic toxicity of MCCPs has been studied in algae, invertebrates and seve ral 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 so il-dwelling organisms. LOECs for sedimentdwelling biota ranged from 270 to 410 mg/ kg dry weight. A re productive 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. Additiona l 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 a quatic 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 availabl e for a sediment-dwelling invertebrate. As a result, an equilibrium partitioning appr oach (Di Toro et al . 1991) using the most sensitive chronic measurement endpoint identif ied for a pelagic fres hwater invertebrate aquatic species (8,900 ng/L) was used to estimate the toxicity to benthic organisms. The

LOEC benthic 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 th e 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, us ing an OECD artifi cial 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 LC 50 value

for adult survival and EC 50 and EC 10 values for reproduction of 5733, 1230, and 660 mg/kg dry wt. (nominal), respectively. Th e 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 LOA EL for mammals. Interspecies scaling using data for a typical adu lt 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 decrea se 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 statis tically 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) LOEC38 of 383 mg/kg dry wt. in soil w ith an organic carbon conten t of 2%, for reproduction in earthworms (Thompson et al. 2001a).

D- Mammals

The lowest effect level observed for mamma ls 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 typi cal adult otter was used to extrapolate to a food concentra tion for this species. This resulted in a CTV of 42 mg/kg food. See Table 7 of this report and Appe ndix 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 Fr ank (1993) and Frank and Steinhäuser (1994). The most sensitive aquatic toxicity value for liquid C 18-20 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 pelagi c freshwater species (68,000 ng/L) was used to estimate the toxicity of liquid C 18-20 LCCPs to soil-dwelling organisms . The LOEC soil for C 18-20 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 wa s 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 usi ng data for a typical adult otter will be used to extrapolate to a food concentration for th is 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 (C 20–30 , 43% chlorine) during a 90- day study. In addition, mild nephrosis was obser ved in the kidneys of male rats, as was mineralization in the kidneys of female ra ts administered 3750 mg/kg-bw per day. A

NOEL could not be established for the female s (LOEL = 100 mg/kg- food). Interspecies scaling using data for a typical adult otte r will be used to ex trapolate 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 ma y 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 ri sks for persistent and bioaccumulative substances - such as SCCPs, MCCPs, and C 18-20 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 st ate concentrations in sediment and soil, EEVs based on monitoring data will be t oo 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 ma ximum steady state concentrations in the tissues of laboratory organisms, ENEVs based on standard toxicity tests may

underestimate effect thresholds if test dur ations are insufficient to achieve maximum internal organism concentrations. Furtherm ore, since food consumption is usually the primary route of exposure to persistent and bioaccumulativ e substances in the field – especially for top predators - ENEVs may unde restimate effect thresholds if the food pathway is not considered in key toxicity studies. These factors are exacerbated when40 available effects and exposure data are limited, as is the case for the chlorinated paraffins.

Risk quotients were calculated for SCCPs, MCCPs, C 18-20 LCCPs and C >20 LCCPs (Table 7). For each identified class of risk re ceptors (e.g., pelagic organisms, benthic organisms), an EEV was selected based on empirical data. The maximum reported field

concentration which is relevant to the Ca nadian environment was used as the EEV.

Chemical concentrations from the Canadian environment were preferably used for EEVs; however, data from other regions in the wo rld were used in the absence of suitable Canadian data. Section 8.2 of the suppor ting 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 (Environmen t Canada, 2008), typically represent the lowest chronic ecotoxicity value from an av ailable 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 organi sms, 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 RQ s to underestimate potential for harm for persistent and bioaccumulative substances. Results are summarized in Table 7.

Concentrations of C 18-20 liquid LCCPs in sediments representative of Canadian environments are not available. In addition, no toxicity data were av ailable for the effects

of C 18-20 liquid LCCPs on secondary consumers. Therefore, risk quotients could not be calculated for exposure of benthic orga nisms and secondary consumers to C 18-20 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 we re not calculated for these groups.

Table 7. List of Estimated Exposure Values (EEV), Critical Toxicity Values (CTV), Assessment Factors (AF), a nd 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 44.8 a ng/L 8,900 b ng/L 10 (lab/field) 890 ng/L 0.05

Benthic 0.41 c mg/kg 35.5 d mg/kg 10 (lab/field) 3.55 mg/kg 0.12

Soil-dwelling 0.64 e mg/kg 660 d mg/kg 10 (lab/field) 66.0 mg/kg 0.01

Secondary Consumer 2.63 f mg/kg 1,000 g mg/kg

food 100 (lab/field & species variations) 10 mg/kg 0.26

MCCPs
Pelagic 0.0026 h ng/L 18,000 i ng/L 10 (lab/field) 1,800 ng/L 0.0000014

Benthic 65.1 j mg/kg 270 k mg/kg 10 (lab/field) 27 mg/kg 2.40

Soil-dwelling 31.0 l mg/kg 383 m mg/kg 10 (lab/field) 38.3 mg/kg 0.81

Secondary Consumer 0.904 n mg/kg 42 o mg/kg food 100 (lab/field & species variations) 0.42 mg/kg 2.15

C

18-20 liquid LCCPs
Pelagic 100 p ng/L 68,000 qng/L 10 (lab/field) 6,800 ng/L 0.0241 Soil-dwelling 3.1 r mg/kg 2,035 s mg/kg 10 (lab/field) 203.5 mg/kg 0.02

C >20 liquid LCCPs
Secondary Consumer 0.0465 t mg/kg 1,000 u mg/kg 10 (lab/field) 100 mg/kg 0.0005

C >20 solid LCCPs
Secondary Consumer 0.0465 v mg/kg 100 w mg/kg 10 (lab/field) 100 mg/kg 0.000465

a The highest concentration of SCCPs obser ved in final effluent of sewage treatm ent plants in southern Ontario was 448 ng/L at the Woodward Avenue sewage treatment plant in Ha milton, 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 sedi ments observed from Lake Ontario, Niagara (or west) basin, in 1998.

d EC 10 for F. candida reproduction.

e The maximum allowable rate for sewage biosolid application to agricultural land s is 8 tonnes of solids per hectare per

5 years (MOE 1998). The soil mass is 5,000 tonnes/ha (assu ming that the biosolids are incorporated into the top 20 cm of the soil having a dry soil bulk density of 2500 kg/m 3 (EU 2003)). Using a SCCPs concentration in sewage sludge of 200 mg/kg dry wt. and assuming that SCCPs-contai ning sludge is applied to the land for 10 years and that no or little biodegrad ation of the SCCPs occurs, a soil concentrat ion 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 concentr ation 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 land s is 8 tonnes of solids per hectare per

5 years (MOE 1998). The soil mass is 5,000 tonnes/ha (assu ming that the biosolids are incorporated into the top 20 cm of the soil having a dry soil bulk density of 2500 kg/m 3 (EU 2003)). Using a MCCPs concentration in sewage sludge of 9,700 mg/kg dry wt. and assuming that SCCPs-contai ning sludge is applied to the land for 10 years and that no or little biodegrad ation of the SCCPs occurs, a soil concentr ation 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 th e 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 ex trapolate to a food concentration for this species (CCME 1998). The resulting CTV was 42 mg/kg food wet wt.

p Detection limit for se wage treatment works in the United Kingdom.

q21-day LOEC for Daphnia magna.

r The maximum allowable rate for sewage bi osolid application to agricultural lands is 8 tonnes of solids per hectare per

5 years (MOE 1998). The soil mass is 5,000 tonnes/ha (assu ming that the biosolids are incorporated into the top 20 cm of the soil having a dry soil bulk density of 2500 kg/m 3 (EU 2003)). Using a MCCPs concentration in sewage sludge of 9,700 mg/kg dry wt. (worst-case concentr ation in the absence of exposure data for C 18-20 liquid LCCPs) and assuming that SCCPs-containing sludge is applied to the la nd 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 C 18-20 liquid LCCPs soil concentration of 3.1 mg/kg dry wt.

s Value calculated using an equilibr ium partitioning approach using the Daphnia magna LOEC.

t C 18–29 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 ex otic 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 es timated 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. Inte rspecies scaling using this LOAEL for a typical adult otter ( lutra canadensis ) (adult body weight of 8 kg a nd average daily food ingestion ra te of 0.8 kg wet wt. per day) was used to extrapolate to a food concentration for th is species (CCME 1998). The resulting CTV was 1,000 mg/kg

food wet wt.

v C 18–29 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 ex otic 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, the42 concentration of LCCPs in mussels was es timated 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 dietar y 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 extrapolat e 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 th an 1. The MCCPs risk quotient for benthic organisms (RQ=2.40) and the MCCP s 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 da ta mentioned above and explained in more detail in Section 8.2 of the supporting docum ent, the absence of RQs above 1 for SCCPs

and C 18-20 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.. A lthough 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 poi nt 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 potenti al for toxicity to organisms, provides a significant indication of its pot ential 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. Releas es of small amounts of bioaccumulative substances may lead to high internal con centrations 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 C 18-20 LCCPs are considered to be both highly persistent and

bioaccumulative. The limited availa ble 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

C 18-20 LCCPs are released into the Canadian envi ronment and have the potential to cause harm to sensitive aquatic organisms at relati vely low concentrations (i.e., chronic NOECs for pelagic organisms < 100 ng/L).

In light of this evidence, it is conclu ded that SCCPs, MCCPs and LCCPs up to C 20 may be causing long term ecological harm in Canada.

4.6. Uncertainties on the ecological risk assessment

This risk assessment contains several sour ces of uncertainty. Uncertainties in the43 exposure and effects assessment can influence th e characterization of risks. Below is a

brief discussion of these uncertainties. Add itonal 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 Cana dian conditions. Concentrations of CPs in various media were often only available for ce rtain areas, and were onl y representative of a short time period, in Canada and other c ountries. As a result, it is unkown how concentrations of CPs vary temporally and sp atially. Moreover, concentrations were often not available near po tential point sources such as metalworking operations (primary source of CPs) and other formulati ng/manufacturing sites that use CPs.

Uncertainties with the toxicity informati on used to drive ENEVs in this assessment

include:

• The use of an equilibrium partitioning appro ach 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 risk s from exposure to SCCPs, MCCPs, C 18-20 LCCPs in Canada have likely been unde restimated 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 ri sks to other types of organism s cannot be estimated at all.

4.6.2 Persistence and Bioaccumulation St atus and Risk Implications

Information on physical properties of MCCPs, and especially LCCPs, is limited. Values used in this assessment are based on extrapol ations 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 s ubstances in the envir onment. Even though there are no data for persistence of LCCPs in sediment, based on biodegradation data which indicate increasing stability with incr easing carbon chain length, it is reasonable to conclude that LCCPs are pe rsistent in sediment.

The empirical and modelled bioaccumulation data for SCCPs and MCCPs are very robust

and indicate the substances are bioaccumulative. While th ere is a lack of empirical44 bioaccumulation data for LCCPs, the modelling results provided by the Modified Gobas BAF Model - which suggest that of all the LCCPs congeners only liquid C 18-20 LCCPs have significant bioaccumulation poten tial – 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 persiste nt 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 iden tified recent data cons idered critical to quantitative estimation of exposure of the ge neral 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 chlo rinated 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)/elect ron capture negative ion high-resolution mass spectrometry (ECNI-HRMS), due to its higher mass resolving power and selectivity. However, such information is lim ited 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 so il, respectively, in estimating intake.

Indeed, data on concentrations of chlorina ted paraffins in foodstuffs are extremely

limited. While additional data on the concentr ations of SCCPs, MCCPs and LCCPs in foods in the United Kingdom (Campbell a nd McConnell, 1980b) re ported in an early investigation reviewed in the PSL1 asse ssment (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 me thodology available at that time. Analysis was based on liquid–solid adsorpti on chromatography, which has now largely been replaced by micro-analytical techniques and quantification by visual re ference to spots appearing on thin-layer chromatographic plates.

Table 8. Concentrations of short-chain, medium -chain and long-chain chlorinated paraffins

in foodstuffs45 Concentration used to represent food group Food

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.

C 10–20 (SCCPs and MCCPs) (Campbell and McConnell, 1980a) 0.19 μg/g

1 sample of cheese in U.K.

C 20–30 (Campbell and McConnell, 1980a)

Fats 0.15 μg/g

mean of 6 samples of vegetable oils and derivatives

C 10–20 (SCCPs and MCCPs) (Campbell and McConnell, 1980a) 0.05 μg/g

detection limit in analysis of 1 sample of

lard in U.K.

C 20–30 (Campbell and McConnell, 1980b)

Fruits 0.025 μg/g

mean of 16 samples of fruits and vegetables in

U.K.

C 10–20 (SCCPs and MCCPs) (Campbell and McConnell, 1980a) 0.025 μg/g

1 sample of peach fruit in U.K.

C 20–30 (Campbell and McConnell, 1980a)

Vegetables 0.025 μg/g

mean of 16 samples of fruits and vegetables in

U.K.

C 10–20 (SCCPs and MCCPs) (Campbell and McConnell, 1980a) 0.025 μg/g

1 sample of potato crisps in U.K.

C 20–30 (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 Pes ticides Team, 1995); average molecular formula is C 12 H 19 Cl 7 , 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.

C 10–20 (SCCPs and MCCPs) (Campbell and McConnell, 1980b) 0.05 μg/g

detection limit in analys es of corn flakes in

U.K.

C 20–30 (Campbell and McConnell, 1980b)

Meat and poultry 0.099 μg/g

1 sample of bacon in U.K.

C 10–20 (SCCPs and MCCPs) (Campbell and McConnell, 1980b) 0.05 μg/g

detection limit in anal ysis of 1 sample each

of ox liver and beef in U.K.

C 20–30 (Campbell and McConnell, 1980b)46 Concentration used to represent food group Food

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; C 10 –C 13 (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.

C 10–20 (SCCPs and MCCPs) (Campbell and McConnell, 1980b) 0.05 μg/g

detection limit in 1 sample of strawberry jam in U.K.

C 20–30 (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 anal ysis of 1 sample each

of beer and tea in U.K.

C 20–3047 Concentration used to represent food group Food

group Short- and medium-chain chlorinated paraffins Long-chain chlorinated paraffins
(Campbell and McConnell, 1980b)

5.1.1 SCCPs

Tomy (1997) determined SCCPs (C 10–13 , 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 Tor onto,” by HRGC/ECNI-HRMS (Muir et al ., 1999). Concentrations ranged from 65 to 924 pg/m 3 . Although a summary statistic of 543 pg/m 3
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 identif ied 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 (C 10–13 , 52% chlorine) ranged from 11 to 17 μg/kg in human breast milk in Canada (Tomy, 1997). Analys es were carried out by HRGC/ECNI-HRMS. No additional details were reported. 2

Muir et al . (1999) analysed whole fish samples for SCCPs (C 10–13 ) and detected 2630 ng/g (wet weight) in carp from Hamilton Har bour, 58.8 ng/g (wet weight) in lake trout from Niagara-on-the-Lake and 72.6 ng/g (wet wei ght) 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 approximat ely 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 confir med by unspecified analysis.

Concentrations of SCCPs have been identifi ed in blubber of aquatic mammals such as ringed seal, beluga and walrus (Tomy et al ., 2000 5 ; Bennie et al ., 2000 6 ). 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. Concen trations 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 repo rted 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 C 12 H 19 Cl 7 , 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 drinki ng water in Canada or elsewhere were not identified. The maximum concentration of SCCPs (C 10–13 , 50–70% chlorine ) in the Red River, at a site remote from industr ialized areas, was 0.05 μg/L (Tomy, 1997). 7 Analyses were by HRGC/ECNI-HRMS. A lower concentra tion was reported in surface water from Lake Ontario (Muir et al ., 2001).

Concentrations of SCCPs in soil in Cana da or elsewhere were not identified. The

concentrations in surface sediment in ha rbours 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 fo r 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 th e estimated intake is from food. The upperbound estimated intake of breast-fed infant s 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 foods 8 ).

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 sh ort-chain chlorinated paraffins by various

age groups

0–6 months 1 Route of

exposure

breast

fed 2 formul

a fed 3 n ot formula

fed 4 0.5–4

years 5 5–11

years 6 12–19

years 7 20–59

years 8 60+

years 9

Ambient air 10 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001

Indoor air 11 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001

Drinking

water 12 0.001 0.001 0.001 <0.001 <0.001 <0.001

Food 13 1.7 0.005

25.96 24.26 16.44 9.02 7.18 5.14

Soil 14 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 m 3 per day (21 hours indoors, 3 hours outdoors) (EHD, 1998).

2 Concentrations of SCCP (C 10–13 , 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 m ilk per day (EHD, 1998).

3 For formula-fed infants, intake fr om 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 repo rted 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 m 3 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 m 3 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 m 3 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 m 3 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 m 3 per day (21 hours indoors, 3 hours outdoors) and to drink 0.4 L of water per day. Consumption of food groups re ported by Health Canada (EHD, 1998).

10 The maximum concentration of C 10 –C 13 (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/m 3 (Muir et al

., 1999).

11 Concentrations of SCCP in indoor air in Canada or else where were not identified. Th e 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 no t identified. The maximum concentration of SCCP (C 10–13 , 50– 70% chlorine) identified in the Red River, at a site remo te 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.; C 10–20 (SCCP and MCCP) (Campbell and McConnell, 1980a) Fats: 0.15 μg/g; mean of 6 samples of vegetable oils and derivatives; C 10–20 (SCCP and MCCP) (Campbell and McConnell, 1980a) Fruits: 0.025 μg/g; mean of 16 sample s of fruits and vegetables in U.K.; C 10–20 (SCCP and MCCP) (Campbell and McConnell, 1980a) Vegetables: 0.025 μg/g; mean of 16 samples of fruits and vegetables in U.K.; C 10–20 (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 Admi nistration (KAN-DO Office and Pesticides Team, 1995); average molecular formula is C 12 H 19 Cl 7 , with 60–65% chlorine content (w/w) (IPCS, 1996) Meat and poultry: 0.099 μg/g; 1 sample of bacon in U.K.; C 10–20 (SCCP and MCCP) (Campbell and McConnell, 1980b) Fish: 2.630 μg/g (wet weight); analysis of whole samples of carp from Hamilton Harbour; C 10 –C 13 (Muir et al

., 1999) Eggs: no data identified Foods primarily sugar: 0.025 μg/g; 1 sample of strawberry jam in U.K.; C 10–20 (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 On tario was 290 ng/g dry weight (Muir et al ., 2001).

Canadian data incorporated within this es timate include high-confidence values in fish (whole carp determined by GC/ECNI-HRMS) an d data on breast milk, for which details of sampling and analysis were not reported. Esti mated intake of SCCPs in fish represents up to 58% of the total daily intake. The in take from dairy products, which accounts for

89.9% of the intake of infants not form ula fed, is based upon limited sampling and analysis — considered semi-quantitative only — of dairy products in the United Kingdom, reported in 1980. Proba bly 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 constitutes50 <0.1% of total estimated intake, and anal ytical methods were not specified.

Intake of SCCPs by a potentially highe r-exposure subgroup of Inuit for whom the primary source of food is subsistence hun ting and fishing (Kuhnlein, 1989; Kinloch et al

., 1992) was also estimated, based on data on c oncentrations 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 estimat ed intake of an In uit 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 manufactur ing plant in Canada in 1993, but not in

surface water or sedime nt (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 MCCP s and the assumptions on which they are based are presented in Table 10. For each age gr oup, virtually all of the estimated intake is from food, which, in turn, is based al most 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 months 1 Route of

exposure

formula

fed 2 not

formula

fed 3 6 months– 4 years 4 5–11

years 5 12–19

years 6 20–59

years 7 60+

years 8

Ambient air 9

– – – – – – –
Indoor air 10

– – – – – – – Drinking

water 11 0.01 0.01 0.01 <0.01 <0.01 <0.01

Food 12 0.05

25.48 18.48 11.64 6.3 4.69 3.47

Soil 13 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 m 3 per day (21 hours indoors, 3 hours outdoors) (EHD, 1998).

2 For formula-fed infants, intake fr om 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. Consumpti on of food groups reported by Health Canada (EHD, 1998).

4 Assumed to weigh 15.5 kg, to breathe 9.3 m 3 per day (21 hours indoors, 3 hours outdoors), to ingest 100 mg of51 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 m 3 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 m 3 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 m 3 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 m 3 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.; C 10–20 (SCCP and MCCP) (Campbell and McConnell, 1980a) Fats: 0.15 μg/g; mean of 6 samples of vegetable oils and derivatives; C 10–20 (SCCP and MCCP) (Campbell and McConnell, 1980a) Fruits: 0.025 μg/g; mean of 16 sample s of fruits and vegetables in U.K.; C 10–20 (SCCP and MCCP) (Campbell and McConnell, 1980a) Vegetables: 0.025 μg/g; mean of 16 samples of fruits and vegetables in U.K.; C 10–20 (SCCP and MCCP) (Campbell and McConnell, 1980a) Cereal products: 0.05 μg/g, detecti on limit in analyses of corn flakes in U.K. (C ampbell and McConnell, 1980b) Meat and poultry: 0.099 μg/g; 1 sample of bacon in U.K.; C 10–20 (SCCP and MCCP) (Campbell and McConnell, 1980b) Fish: 1.23 μg/g (wet weight); mean of 10 sample s 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.; C 10–20 (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 li mit 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 to tal intake of LCCPs and a ssociated assumptions are presented in Table 11. As for SCCPs and MCCP s, for each age group, virtually all of the estimated intake is from food. The highest in take 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 lim ited in that estimates for five of the eight food groups are based upon the limit of de tection 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 months 1 6 months– 5–11 12–19 20–59 60+ years 852 formula

fed 2 not

formula

fed 3 4years 4 years 5 years 6 years 7

Ambient air 9

– – – – – – –
Indoor air 10

– – – – – – – Drinking

water 11 0.01 0.01 0.01 <0.01 <0.01 <0.01

Food 12 0.05

16.81 9.66 5.61 3.04 2.12 1.73

Soil 13 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 m 3 per day (21 hours indoors, 3 hours outdoors) (EHD, 1998).

2 For formula-fed infants, intake fr om 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. C onsumption of food groups reported in EHD (1998).

4 Assumed to weigh 15.5 kg, to breathe 9.3 m 3 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 m 3 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 m 3 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 m 3 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 m 3 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. (Cam pbell 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.; C 20–30 (Campbell and McConnell, 1980a) Fats: 0.05 μg/g; detection limit in analysis of 1 sample of lard in U.K.; C 20–30 (Campbell and McConnell, 1980a) Fruits: 0.025 μg/g; 1 sample of peach fruit in U.K.; C 20–30 (Campbell and McConnell, 1980a) Vegetables: 0.025 μg/g; 1 sample of potato crisps in U.K.; C 20–30 (Campbell and McConnell, 1980a) Cereal products: 0.05 μg/g, detecti on limit in analysis of corn flakes in U.K. (Campbell and McConnell, 1980b) Meat and poultry: 0.05 μg/g; detecti on limit in analysis of 1 sample each of ox liver and beef in U.K.; C 20–30 (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.; C 20–30 (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.;

C 20–30 (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. (Cam pbell and McConnell, 1980a).

5.2. Hazard characterization and dose–response analyses

A limited number of studies on th e toxicity of SCCPs have been reported in the period following release of the PSL1 assessment. Mo st of these studies were conducted to53 investigate the mode of acti on of carcinogenicity for the tumours observed in the NTP (1986a) bioassay, which were liv er 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, re sults 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 onl y one of the relevant i nvestigations has a full published account been identified (Wyatt et al ., 1993). While secondary accounts of (possibly) other studies inves tigating mode of action of tu mour induction in assessments have been reported by the European Comm ission (2000), the U.S. National Research Council (U.S. NRC, 2000) and the Australian National Industrial Ch emicals Notification and Assessment Scheme (NICNAS, 2001), they ar e 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 th e release of the PSL1 assessmen t report. The following presentation is limited to those considered critical to hazard characterization or dose– response analyses for effects in the general popul ation 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 toxicologica l data that impact on critical aspects, the dose–response analyses for MCCPs and LCCP s 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 hype rtrophy, 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 admi nistered were not specified, and quantitative dose- or sex-specific data and analyses were not presented.

Elcombe et al

. (2000) administered Chlorowax 500C (C 10–13 ; 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 (T 4 )–uridine diphosphoglucose glucuronosyl transferase (UDPGGT). (The e ffects were, presumably, observed at both dose levels.) These effects were not obser ved 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 specif ied, and quantitative dose- or sex-specific data and analyses were not pres ented in this summary account.54 Wyatt et al . (1993) exposed groups of five male ra ts (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: C 10–13 , 58% chlorine; or Cereclor 56L, C 10–13 : 56% chlorine). For the 58% chlorine SCCPs, both absolute and relative liver weights were significantly increased in a dose-related manner, at dos es 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 (irre gular dose–response). For the 56% chlorine SCCPs, the pattern of response for absolute liver weight was irregular; however, relativ e liver weight was increased in a dose-related manner, significantly at 50 mg/kg-bw per day a nd 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 relativ e liver weight and pa lmitoyl 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 li ver weights were signifi cantly 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 dos e-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 interc ellular 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 α 2u globulin) and regenerative focal basophi lic tubules, as we ll 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 eff ects 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 protei n nephropathy, associated w ith regenerative hyperplasia and increased DNA synthesis (S-phase activit y), presumably at both dose levels. There

was “some limited evidence” for an involvement of α 2u globulin. These changes were not observed in the guinea pigs. Again, neither quan titative data nor statis tical 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 (C 12 ; 60% chlorine)/kg-bw per day. There was an increase in α 2u globulin and cell proliferation in the kidney of males only. Da ta from individual ra ts indicated that increased cell proliferation was directly correlated with the increase in α 2u globulin. Five

different isoelectric isoforms of α 2u globulin were identified by Western blotting in the55 control male kidney, and all five were increas ed in the treated males. These observations were reported in an abstract; neither quantit ative 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 T 4 –glucuronosyl transferase activity, accompanied by a decrease in plasma T 4 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 maxi mum 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 an alyses 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 T 4 , 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 C 10–13 (71% chlorine) to female Sprague-Dawley rats for 14 days had no effect upon thyroid hormonal T 4 levels or microsomal enzyme activity (Hallgren and Darnerud, 1998).

In male rats (Alpk:APfSD strain) expos ed by gavage for 14 days to two SCCPs (Chlorowax 500C: C 10–13 , 58% chlorine; or Cereclor 56L, C 10–13 : 56% chlorine), for which examination of thyroid function was rest ricted to the contro l and high-dose groups (1000 mg/kg-bw per day), both free and total T 4 were significantly reduced, TSH was significantly increased and the capability of liver microsomes to glucuronidate T 4 was significantly increased in exposed animals (Wyatt et al

., 1993). No differences in levels of free or total triiodothyronine (T 3 ) 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 C 10–13 (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 identifi ed in the PSL1 assessment of chlorinated paraffins (Government of Canada , 1993a). Sprague-Dawley rats (10 per sex per group) were fed diets c ontaining 0, 5, 50, 500 or 5000 ppm for 13 weeks. The dose levels calculated by the authors on the basi s 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 bi ochemistry, hematology, hepatic enzyme activities, urinary enzyme activity, organ weights and histopathology. Mild, adaptive histological changes were dete cted 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 per56 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 obs erved in the renal pr oximal tubules of males at the highest dose and in the inner me dulla 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 wei ght of evidence for hypothesized modes of

induction of tumours addressed below incl ude the criterion that SCCPs are not DNAreactive. Recent data on genot oxicity reported since the PSL1 assessment was released

have not been identified. Limited availabl e data reviewed with in 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, incl uding two additional unpublished studies in

which no increases in revertant colonies in five strains of Salmonella 9 and no increases in

mutant colonies in Chinese hamster V79 cells 10 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 prol iferation involves activation of a nuclear receptor in rodent liver, the peroxiso me proliferator activated receptor, α isoform (PPAR α ). The activated PPAR α interacts with regulatory el ements of the DNA to initiate transcription of genes for increased peroxisomal enzyme activity and cell proliferation characterized by morphological and biochemi cal 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 involve d in peroxisomal fatty acid oxidation) and induction of microsomal fatty acid oxid ation through the CYP4A subfamily of cytochrome P-450 isozymes. Minimum cr iteria for characterizing peroxisome proliferation are considered to include hepa tomegaly, enhanced cell proliferation and an increase in hepatic acyl-CoA oxidase a nd/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 obs erved 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 pero xisome proliferating he patocarcinogens, such as di(2-ethylhexyl)phthalate.

Available data on the role of peroxisome pro liferation in the etiol ogy of hepatic effects and liver tumours induced by SCCPs are restri cted to one study for which there is a published manuscript (Wyatt et al ., 1993) and two investig ations reported only in summary (Elcombe et al ., 2000) or abstract form (Elcombe et al

., 1994). Significant,

dose-related increases in both absolute and re lative liver weights accompanied at higher doses by increases in palmitoyl CoA oxi dation in male Alpk:APfSD rats and Alpk:APfCD-1 mice exposed to two SCCPs, reported by Wyatt et al

. (1993), are

consistent with the observa tions 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 characteri zation 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 invest igations are also consistent (increases in relative liver weight in rats were significant at ≥ 50 mg/kg-bw per day and pa lmitoyl 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 character ization of exposure–response was limited in the NTP bioassay to only two dose levels, evidence to date indicates th at tumours in both rats and mice occur only at doses at which peroxisome proliferation and associated morphological

and biochemical effects have been obs erved in shorter-term studies (Wyatt et al ., 1993;

Elcombe et al ., 1994, 2000).

Additional weight of evidence for concor dance might have been afforded through

consideration of sex-related differences in peroxisome pr oliferation 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 peroxiso me proliferation is in itiated rapidly after treatment with a prolif erator 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 bioass ays 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 observe d for other peroxisome proliferators. Rats and mice are uniquely responsive to the morphological and biochemical effects of peroxisome proliferators, while Syrian hams ters exhibit intermediate responsiveness. This is consistent with marked intersp ecies variations in the expression of PPAR α

.

Additional published documenta tion of existing relevant st udies is desirable. Also, investigation of additional aspects of concordance would strengthen the weight of evidence for causality for the purported associ ation between peroxisome proliferation and liver tumours induced by SCCPs. Howeve r, although there ar e limitations of the identified information, data are strongly sugge stive that peroxisome proliferation plays a role in the etiology of liver damage and he patic 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 se x-specific response attributable to α 2u globulin nephropathy and hence not releva nt to humans. This mode of induction of renal tumours, which is relatively well char acterized, involves binding to α 2u globulin, a protein specific to male rats. This binding renders the protei n more resistant to proteolytic degradation, which causes its accumulation in renal prox imal tubule cells (manifested as hyaline droplets on histopathol ogical examination), resulting in cell death and regenerative proliferation. Sustained cell proliferation lead s to a low but significant incidence of renal

tubular tumours.

Minimum criteria for establishment of α 2u globulin nephropathy as a basis for tumour development include lack of genotoxicity a nd observation of requisi te 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 regenerativ e tubular cell proliferation in the presence of continued exposure, but also on explicit id entification of the pr otein accumulating in

tubule cells as α 2u globulin, along with demonstrated re versible binding of the relevant

chemical or metabolite to α 2u 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 wa s 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 investigati ons 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 proxima l tubular renal cells were observed in male, but not female rats a nd considered by the authors to constitute

“limited evidence” of the role of α 2u globulin. More recentl y, the presence of α 2u globulin was confirmed using immunohistochemical techniques, although no details of methodology were provided (Warnasuriya et al

., 2000).

Owing to the inadequate characte rization 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 α 2u 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 formati on, 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 de sirable, 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 thyroi d tumours in rats associated with decreased circulating thyroi d hormone levels due to increased hepatic metabolism (particularly Phase II conjugati ng enzymes such as uridine diphosphate (UDP) glucuronosyl transferases [UDPGTs] and glutathione S-transferases) and clearance. These compounds induce hepati c glucuronidation of thyroid hormones and increase biliary excretion of the conjugated hormones, resul ting in decreased circulating

T 3 and T 4 levels. As a result of the hypothyroid state, TSH levels increase and cause sustained thyroid follicular cell hyperpl asia, leading to tumour formation.

While the basic physiology and feedback m echanisms of the hypothalamic–pituitary– thyroid axis are qualitatively similar across sp ecies, quantitative differences make rodents more sensitive than humans to development of thyroid cancer for which the sole mode of action is thyroid–pituitary disr uption (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 horm one. With a more rapid turnover of T 4 , there is a generalized increased activity of the pituitary–thyroi d axis in rats compared with humans,

which correlates with increased susc eptibility to thyroi d 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 (the60 latter including reduction in circulating serum T 4 and T 3 and an increase in TSH levels within days or a few weeks of exposure). Ev idence of increases in thyroid growth is provided by measured increases in absolute or relative thyroid weight, histological indication of cellular hypertrophy and hyperpla sia, morphometric determination of alteration in thyroid cellular components and changes in pro liferation of fo llicular 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 carci nomas (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 highes t dose groups; the administered dose for the latter was considerably greater than those in the NTP bioassay asso ciated with thyroid tumours (i.e., 1000 mg/kg-bw per day vers us 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 gava ge 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 concentr ation and thyroid follicular cell hypertrophy and hyperplasia in both sexes.” There are extrem ely 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, su ch 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 T 4 levels or microsomal enzyme activity were not observed were much less than those administered in the NTP bioassay; as a resu lt, these are not additi onally 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 thyr oid disruption in part, it

should be noted that these data are insuffici ent as a basis for analysis of dose–response for concordance with that for thyroid tumours. Also, recovery in th e absence of continued exposure has not been investig ated. In view of the limitati ons of both reporting and dose– response analyses, therefore, there is considerable uncerta inty in attributing observed thyroid tumours to thyroid–pitu itary disruption, to which rode nts are more sensitive than

humans.

B- Risk characterization

Available data relevant to consideration of the weight of evidence for proposed modes of61 induction of liver, kidney and thyroid tumours associated with exposure to SCCPs, although limited, are suggestive that tolerabl e intakes that prot ect for non-neoplastic precursor effects will likely also be protective for cancer. However, owing principally to limited investigation of asp ects such as recovery and inadequate documentation of relevant studies, there is cons iderable uncertainty in drawing this conclusion, particularly for the thyroid tumours. In recognition ofthis 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-Obser ved-Effect Level (NO EL) 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 in creases in liver a nd kidney weight and hypertrophy of the liver and t hyroid. In IPCS (1996), an unc ertainty factor of 100 was applied in the development of the TDI to ac count 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 just ification 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 estimat ed the dose associated with a 5% increase in tumour incidence (Tumorigenic Dose 05 [TD 05 ]) to be 11 mg/kg-bw per day (amortized for period of administration).

The upper-bound estimate of exposure for the ag e group with greatest exposure to SCCPs (i.e., 26 μg/kg-bw per day) is within the rang e 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 estimat e of exposure for the age group with greatest exposure to SCCPs and the Tumorigenic Dose (TD 05 ) (i.e., 440) is also considered inadequate in view of the un certainty concerning m ode 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 assessm ent (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 genera l population of Canada exceed the TDI (6 μg/kg-bw per day) for non-neoplastic effect s. Indeed, for infants not formula fed, the62 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 inta ke of LCCPs from drinking water, food and soil for the genera l 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 es timates 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 an alysis 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 indi vidual groups of chlorinated paraffins (SCCPs, MCCPs and LCCPs) from these sources has been overestimated.

The estimates of intake for SCCPs are base d in part upon the results of more recent surveys, for which methods of analysis we re 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 ex tent of overestimation of exposure based on the earlier, likely less selec tive 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 SC CPs 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 uppe r-bounding estimates of exposure to MCCPs. These estimates are based in large part upon concentrations reporte d in a limited number of foodstuffs in the United Kingdom, which were published in 1980. More recent, although limited, data on concentr ations in trout analysed by LRMS were included in the calculation of upper-bounding estimates.

There is minimal confidence in the uppe r-bounding estimates of exposure to LCCPs. These estimates are based entirely upon con centrations reported in a limited number of63 foodstuffs in the United Kingdom, which were published in 1980. Furthermore, concentrations in foods were represented by th e limits of detection for five of eight food groups in the calculati ons of daily intake.

There is a low degree of confidence in the databa se of toxicological st udies 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 pub lished complete report (Wyatt et al ., 1993) is available and for which it has not been possible to identify published acc ounts for reported prepublication manuscripts reviewed in previ ous assessments. Results in the only fully documented study provide most meaningful suppor t 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 t oxicological studies upon which the TDI for MCCPs is based, for which studies on chronic toxicity or carcinogenicity are lacking. The database fo r 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 s hort-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 medi um-chain chlorinated paraffins (MCCPs) are entering, or may en ter, 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 C 20 are entering, or may enter, th e 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 C 20 are “toxic” as defined in paragraphs 64 (a) of the Canadian Environmental Protection Act, 1999.

In addition, on the basis of the limited availa ble data, it is conc luded that long-chain chlorinated paraffins are ente ring, 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 paraffi ns are “toxic” as defined in paragraph 64 (c) of the Canadian Environmental Protection Act, 1999.65 REFERENCES

Allpress, J.D. and P.C. Gowland. 1999. Biodegr adation of chlorinated paraffins and longchain chloroalkanes by Rhodococcus sp. S45-1. Int. Biodeter. Biodegr. 43: 173–179.

Arnot, J.A., and F.A.P.C. Gobas. 2003. A generic QSAR for assessing the bioaccumulation potential of organic chemi cals in aquatic food webs. QSAR Comb Sci 22(3): 337-345.

Atkinson, R. 1986. Kinetics and mechanisms of gas phase reactions of the hydroxyl radical with organic compounds under atmo spheric conditions. Chem. Rev. 86: 69–201.

Atkinson, R. 1987. Estimation of gas-phase hydro xyl radical rate c onstants for organic

chemicals. Environ. Toxicol. Chem. 7: 435–442.

Ballschmiter, K. 1994. [Determination of s hort and medium chain length chlorinated paraffins in samples of water and sedime nt from surface water.] Department of Analytical and Environmental Chemistry, Univ ersity of Ulm, Ulm, Germany, May 10 (in German).

Bengtsson, B. and E. Baumann-Ofstad. 1982. Long-term studies of uptake and elimination of some chlorinated paraffins in the bleak, Alburnus alburnus. Ambio 11: 38–

40.

Bengtsson, B., O. Svenberg, E. Lindén, G. Lunde and E. Baumann-Ofstad. 1979. Structure related uptake of chlorinated paraffi ns in bleaks (Alburnus alburnus). Ambio 8:

121–122.

Bennie, D.T., C.A. Sullivan and R.J. Maguire. 2000. Occurrence of chlorinated paraffins in beluga whales (Delphinapterus leucas) fr om the St. Lawrence River and rainbow trout (Oncorhynchus mykiss) and carp (Cyprinus carp io) from Lake Ontario. Water Qual. Res.

J. Can. 35: 263–281.

Bezchlebová, J., J. Č ernohláková, K. Kobeti č ová, J. Lána, I. Sochová, J. Hofman. 2007.

Effects of short-chain chlori nated paraffins on soil organism s. Ecotox. & Envir. Safety

67:206-211.

Bidleman, T., D. Muir and G. Stern. 1999. Ne w persistent chemicals in the Arctic environment. In: S. Kalhok (ed.), Synops is of research conducted under the 1998–99 Northern Contaminants Program. Indian and Northern Affairs Canada, Ottawa, Ontario. pp. 17–25 [cited in Indian and Northern Affairs Canada, 2003].

Bidleman, T., M. Alaee and G. Stern. 2000. Ne w persistent chemicals in the Arctic environment. In: S. Kalhok (ed.), Synops is of research conducted under the 1999–2000 Northern Contaminants Program. Indian and Northern Affairs Canada, Ottawa, Ontario. pp. 87–93 (QS-8602-000-EF-A1) [cited in Indian and Northern Affairs Canada, 2003].66 Bidleman, T., M. Alaee and G. Stern. 2001. Ne w persistent chemicals in the Arctic environment. In: S. Kalhok (ed.), Synops is of research conducted under the 2000–2001 Northern Contaminants Program. Indian and Northern Affairs Canada, Ottawa, Ontario [cited in Indian and Northern Affairs Canada, 2003].

Borgen, A., M. Schlabach and H. Gundersen. 200 0. Polychlorinated alkanes in Arctic air. In: M.S. Denison (ed.)., Dioxin 2000. 20th International Symposium on Halogenated Environmental Organic Pollutants and Persis tent Organic Pollutant s (POPs), Monterey, California, August 13–17, 2000. Or ganohalogen Compd. 47: 272–274.

Borgen, A., M. Schlabach, R. Kallenbor n, G. Christensen and T. Skotvold. 2002. Polychlorinated alkanes in ambient air from Bear Island. In: Dioxin 2002. 22nd International Symposium on Halogenated Environmental Organic Pollutants and Persistent Organic Polluta nts (POPs), Barcelona. Or ganohalogen Compd. 59: 303–306.

Bruggeman, W.A., L.B.J.M. Martron, D. Kooiman and O. Hutzinger. 1981. Accumulation and elimination of di-, tri- and tetra-chlorobiphe nyls by goldfish after dietary and aqueous expo sure. Chemosphere 10: 811–832.

BUA (Beratergremium für Umweltrelevante Al stoffe). 1989. Chlorinated paraffins. Draft report of January 1989. German Chemical Society (GDCh) Advisory Committee on Existing Chemicals of Environmental Relevance.

BUA (Beratergremium für Umweltrelevante Alstoffe). 1992. Chlorinated paraffins. German Chemical Society (GDCh) Advi sory Committee on Exis ting Chemicals of Environmental Relevance, June (BUA Report 93).

Bucher, J.R., R.H. Alison, C.A. Montgomery, J. Huff, J.K. Haseman, D. Farnell, R. Thompson and J.D. Prejean. 1987. Comparativ e toxicity and carci nogenicity of two chlorinated paraffins in F 344/N and B6C3F1 mice. Fundam. Appl. Toxicol. 9: 454–468.

Camford Information Services. 2001. Chlo rinated paraffins. CPI Product Profile. Toronto, Ontario. 2 pp.

Campbell, I. and G. McConnell. 1980a. Chlo rinated paraffins and the environment. 1.

Environmental occurrence. Envi ron. Sci. Technol. 14(10): 1209–1214.

Campbell, I. and G. McConnell. 1980b. Extended versions of Tables I, II, and IV–VI describing analysis of human post-mortem organ and tissue samples for chlorinated paraffins; Figures 4–6 showing locations of sampling points; and an appendix of substances not interfering with the analyt ical method. Business Operations, Books and Journals Division, American Chem ical Society, Washington, D.C.

CCME (Canadian Council of Ministers of the Environment). 1998. Protocol for the derivation of Canadian tissue residue guide lines for the protecti on of wildlife that67 consume aquatic biota. Wi nnipeg, Manitoba [reprinted in CCME. 1999. Canadian environmental quality guidelines . Chapter 8. Winnipeg, Manitoba].

CPIA (Chlorinated Paraffins Industry Asso ciation). 2002. Comments on the draft report “Short chain chlorinated parafffins (SCCPs) substance dossier” (draft March 2). Correspondence to G. Filyk, Environment Canada, from R. Fensterheim, CPIA, May 17.

Di Toro, D.M., C.S. Zarba, D.J. Hansen, W.J. Berry, R.C. Swartz, C.E. Cowan, S.P. Pavlou, H.E. Allen, N.A. Thomas and P.R. Paquin. 1991. Technical basis for establishing sediment quality criteria for nonionic organi c chemicals using equilibrium partitioning.

Environ. Toxicol. Chem. 10: 1541–1583.

Dohler, K., C. Wong and A. von zur Muhlen. 1979. The rat as a model for the study of drug effects on thyroid function: considerati on of methodological problems. Pharmacol.

Ther. 5: 305–318.

Drouillard, K.G. 1996. Physico-chemical propert y determination on chlorinated n-alkanes (C10 to C12). Parameters for estimation of the environmental fate of chlorinated nparaffins. M.Sc. Thesis, University of Manitoba, Winnipeg, Manitoba.

Drouillard, K.G., G.T. Tomy, D.C.G. Muir and K.J. Friesen. 1998a. Volatility of chlorinated n-alkanes (C10–12): vapour pressu res and Henry’s law constants. Environ.

Toxicol. Chem. 17: 1252–1260.

Drouillard, K.G., T. Hiebert, P. Tran, G.T. Tomy, D.C.G. Muir and K.J. Friesen. 1998b. Estimating the aqueous solubilities of indi vidual chlorinated n-alkanes (C10–12) from

measurements of chlorinated alkane mi xtures. Environ. Toxicol. Chem. 17: 1261–1267.

DuPont (DuPont Specialty Chemicals). 1995. Huma n skin irritation test — five subjects.

Submission to Office of Prevention & Pest icides and Toxics, U.S. Environmental Protection Agency, by Haskell Laboratory for Toxicology and I ndustrial Medicine (EPA/OTS Document No. 86960000221S; NTIS Microfiche No. NTIS/OTS0572882).

EHD (Environmental Health Directorate). 1998. Exposure factors for assessing total daily intake of Priority Substances by the genera l population of Canada. Priority Substances Section, Environmental Health Director ate, Health Canada, Ottawa, Ontario (unpublished).

Elcombe, C., S. Watson, I. Wyatt and J. Foster. 1994. Chlorinated paraffins (CP): Mechanisms of carcinogenesis. Toxi cologist 14(1): 276 (Abstract 1056).

Elcombe, C., G. Warnasuriya and J. Foster . 2000. Chlorinated para ffins: mechanisms of non-genotoxic carcinogenesis. In: M.S. Denison (ed.)., Dioxin 2000. 20th International Symposium on Halogenated Envi ronmental Organic Pollutants and Persistent Organic Pollutants (POPs), Monterey, Califor nia, August 13–17, 2000. Organohalogen Compd.

47: 143–146.68 Environment Canada. 1997. Scientific justif ication. Short chain chlorinated paraffins. Candidate substance for management unde r Track 1 of the Toxic Substances Management Policy (ISBN 662-25391-4; Catalogue No. En40-230/5-1997E).

Environment Canada. 2003a. Data collected from “Notice with Respect to Short-, Medium- and Long-chain Chlorinated Paraffins .” Canada Gazette, Part I, November 30,

2002.

Environment Canada. 2003b. Short chain ch lorinated paraffins (SCCPs) substance dossier. Final draft II, revised May 16. Prepared for United Nations Economic Commission for Europe Ad hoc Expert Gr oup on Persistent Or ganic Pollutants.

Environment Canada. 2008. Follow-up report on PSL1 substance for which there was

insufficient information to conclude whethe r the substance constitutes a danger to the environment; Chlorinated Paraffins Supporting Working Document.

EU (European Union). 2003. Technical guidanc e document on risk assessment, Part II.

Institute for Health and Consumer Protec tion, European Chemicals Bureau, EU Joint Research Centre (EUR 20418 EN/2).

European Commission. 2000. European Union ri sk assessment report. Vol. 4. Alkanes,

C 10–13 , chloro. Joint Research Centre, Instit ute for Health and Consumer Protection, European Chemicals Bureau, Eur opean Commission (ISBN 92-828-8451-1).

Fisk, A., C. Cymbalisty, A. Bergman a nd D.C.G. Muir. 1996. Dietary accumulation of

C 12 - and C 16 -chlorinated alkanes by juvenile rainbow trout (Oncorhynchus mykiss). Environ. Toxicol. Chem. 15(10): 1775–1782.

Fisk, A.T., S.C. Wiens, G.R.B. Webste r, A. Bergman and D.C.G. Muir. 1998a. Accumulation and depuration of sediment-sorbed C 12 and C 16 polychlorinated alkanes by oligochaetes ( Lumbriculus variegatus ). Environ. Toxicol. Chem. 17: 2019–2026.

Fisk, A.T., C.D. Cymbalisty, G.T. Tomy and D.C.G. Muir. 1998b. Dietary accumulation and depuration of C10-, C11- and C14-polychlorinated alkane s by juvenile rainbow trout (Oncorhynchus mykiss). Aquat. Toxicol. 43: 209–221.

Fisk, A.T., R.J. Norstrom, C.D. Cymb alisty and D.C.G. Muir. 1998c. Dietary accumulation and depuration of hydrophobic organochlorines: Bioaccumulation parameters and their relationship with Kow. Environ. Toxicol. Chem. 17: 951–961.

Fisk, A.T., G.T. Tomy, C.D. Cymbalisty and D.C.G. Muir. 2000. Dietary accumulation and quantitative structure activity relationshi ps for depuration and biotransformation of short, medium and long carbon chain polychlor inated alkanes by juvenile rainbow trout (Oncorhynchus mykiss). Environ. Toxicol. Chem. 19: 1508–1516.

Frank, U. 1993. Ökotoxizität von Chloroparaffinen. Institut für Wasser- Boden und69 Lufthygiene, November 23 [cited in U.K. Environment Agency 2001].

Frank, U. and F.G. Steinhäuser. 1994. [Eco toxicity of poorly soluble substances exemplarily tested by Daphnia toxicity of chlorinated paraffins.] Vom Wasser 83: 203– 211 (in German with English summary).

Government of Canada. 1993a. Priority Substa nces List assessment report. Chlorinated paraffins. Minister of Supply and Serv ices, Ottawa, Ontario (ISBN 0-662-20515-4; Catalogue No. En40-215/17E).

Government of Canada. 1993b. Canadian E nvironmental Protection Act. Priority Substances List supporting document. Chlori nated paraffins. Environment Canada and Health and Welfare Canada. 66 pp.

Government of Canada. 1995. Toxic Substan ces Management Policy, persistence and bioaccumulation criteria. Minister of Suppl y and Services, Ottawa, Ontario (ISBN 0-662- 23524-X; Catalogue No. En 40-499/2-1995E).

Government of Canada. 2000. Persistence and Bioaccumulation Regulations. Canada Gazette, Part II. 134 (7): 607-612.

Hallgren, S. and P. Darnerud, P. 1998. Effect s of polybrominated diphenyl ethers (PBDEs), polychlorinated biphenyls (PCBs) a nd chlorinated paraffins (CPs) on thyroid hormone levels and enzyme ac tivities in rats. In : N. Johansson, Å. Bergman, D. Broman,
H. Håkansson, B. Jansson, E. Klasson Wehler , L. Poellinger and B. Wahlström (eds.), Dioxin ’98. 18th International Symposiu m on Chlorinated Dioxins and Related Compounds, Stockholm, Sweden, Augus t 17–21, 1998. Organohalogen Compd. 35: 391–

394.

Halsall, C., R. Bailey, G. Stern, L. Barrie, P. Fellin, D. Muir, B. Rosenberg, F. Rovinsky, E. Kononov and B. Pastukhov. 1998. Multi-yea r observations of organohalogen pesticides in the Arctic atmosphere. E nviron. Pollut. 102: 51–62 [cited in Indian and Northern Affairs Canada, 2003].

Hoechst AG. 1987. Chloroparaffin 56 liquid — De tection of gene mutations in somatic

cells in culture. HGPRT-test with V79 cells . Hoechst AG, Frankfurt/Main, Germany. 23 pp. (Report No. 87.1719) [cited in U.K. Environment Agency 2003].

Houde, M., M. Whittle, G. Tomy, C. Teixeira, S. Moore, and D.C.G. Muir. 2006. Short and medium chain chlorinated paraffins in food webs from Lake Ontario and Lake Michigan. Unpublished report, Water Scien ce and Technology Directorate, Environment Canada, Burlington, Ontario.

Houghton, K.L. 1993. Chlorinated paraffins. In: Kirk-Othmer encyclopedia of chemical technology. 4th edition. Vol. 6. John Wile y and Sons, New York, NY. pp. 78–87.70 Howard, P., J. Santodonato and J. Saxena. 1975. Investigation of selected potential

environmental contaminants: Chlorinated paraffins. U.S. Environmental Protection Agency, Washington, D.C. 109 pp. (EPA- 560/2-75-007; NTIS Reports PB-248634 and PB-243075).

IARC (International Agency for Research on Cancer). 1999. Consensus report. Species differences in thyroid, kidney and urinary bladder carcinogenesis. IARC Sci. Publ. 147:

1–14.

ICI. 1992. Personal communication from ICI Cana da [cited in U.K. Environment Agency 2003].

Indian and Northern Affairs Canada. 2003. Ca nadian Arctic contaminants assessment report II. Sources, occurrence, trends and pathways in the physical environment. Northern Contaminants Program, Indian and Northern Affairs Canada, Ottawa, Ontario.

IPCS (International Programm e on Chemical Safety). 1984a. Heptachlor. World Health Organization, Geneva (Environmen tal Health Criteria 38).

IPCS (International Progra mme on Chemical Safety). 1984b. Mirex. World Health Organization, Geneva (Environm ental Health Criteria 44).

IPCS (International Progra mme on Chemical Safety). 1991. Lindane. World Health Organization, Geneva (Environmen tal Health Criteria 124).

IPCS (International Programm e on Chemical Safety). 1996. Ch lorinated paraffins. World Health Organization, Geneva. 181 pp. (Envi ronmental Health Criteria 181).

IRDC (1984) 13-week dietary t oxicity study in rats with combined excretion, tissue level and elimination studies/determination of excretion, tissue level and elimination after single oral (gavage) administra tion to rats. Chlorinated para ffin: 70% chlorination of long chain length n-paraffins; 14C labeled CP. Mattawan, Michigan, International Research and Development Corporation, 316 pp (Report No. 438-027/024).

IRDC (International Research and Devel opment Corporation). 1984a. 13-week oral (gavage) toxicity study in rats with comb ined excretion, tissue level and elimination studies; determination of excre tion, tissue level and elimination after single oral (gavage) administration to rats. Chlorinated paraffin: 58% chlorination of short chain length nparaffins; 14C labeled CP. Mattawan, Mi chigan. 350 pp. (Report No. 438-029/022) [cited in IPCS, 1996].

Jackson, K. 2001. Personal communicati on. Memorandum regarding article by C. Elcombe from K. Jackson, Environmental He alth Centre Library, to G. Long, Existing Substances Division, Health Cana da, Ottawa, Ontario, April 4, 2001.

Jansson, B., R. Andersson, L. Asplund, K. L itzen, K. Nylund, U. Sellstrom, U. Uvemo,71 C. Wahlberg, U. Wideqvist, T. Odsjo a nd M. Olsson. 1993. Chlorinated and brominated persistent organic compounds in biological samples from the environment. Environ. Toxicol. Chem. 12(7): 1163–1174.

KAN-DO Office and Pesticides Team. 1995. Accumulated pesticide and industrial chemical findings from a ten-year study of r eady-to-eat foods. J. Assoc. Off. Anal. Chem. Int. 78(3): 614–631.

Kato, Y. and K. Kenne. 1996. Inhibition of cell–cell communication by commercial chlorinated paraffins in rat liver epithelial IAR 20 cells. Pharmacol. Toxicol. 79(1): 23–

28.

Kemmlein, S., A. Hermeneit and W. Ro tard. 2002. Carbon skeleton analysis of chloroparaffins in sediment, mussels and crabs. Organohalo gen Compd. 59: 279–282.

Kinloch, D., H. Kuhnlein and D. Muir. 1992. Inuit foods and diet: a preliminary

assessment of benefits and risks. Sci. Total Environ. 122(1/2): 247–278.

Kuhnlein, H. 1989. Nutritional and toxicological components of Inuit diets in Broughton Island, Northwest Territories. October 1989. Cont ract report for E. Berthelet, Assistant Deputy Minister, Department of Healt h, Yellowknife, Northwest Territories.

Lahaniatis, M., M. Coelhan and H. Parlar. 2000. Clean-up and quantif ication of short and medium chain polychlorinated n-alkanes in fis h, fish oil and fish feed. In: M.S. Denison (ed.)., Dioxin 2000. 20th International Sym posium on Halogenated Environmental Organic Pollutants and Persis tent Organic Pollutants (P OPs), Monterey, California, August 13–17, 2000. Organohalogen Compd. 47: 276–279.

Mackay, D., A. Di Guardo, S. Paterson and C.E. Cowan. 1996. Evaluating the environmental fate of a variety of types of chemicals using the EQC model. Environ.

Toxicol. Chem. 15: 1627–1637.

Madeley, J. and R. Birtley. 1980. Chlorinated paraffins and the environment. 2. Aquatic and avian toxicology. Environ. Sci. Tec hnol. 14: 1215–1221 [cited in U.K. Environment Agency 2003].

Madeley, J.R. and B.G. Maddock. 1983a. The bi oconcentration of a chlorinated paraffin in the tissues and organs of rainbow trout (Salmo gairdneri). Imperial Chemical Industries PLC, Devon, U.K. (B rixham Report No. BL/B/2310).

Madeley, J.R. and B.G. Maddock. 1983b. Toxicity of a chlorinated paraffin to rainbow trout over 60 days. Imperial Chemical Indus tries PLC, Devon, U.K. (Brixham Report No. BL/B/2203).

Madeley, J.R. and B.G. Maddock. 1983c. Toxici ty of a chlorinated paraffin to rainbow trout over 60 days. Imperial Chemical Indus tries PLC, Devon, U.K. (Brixham Report No.72 BL/B/2202).

Madeley, J.R. and R.S. Thompson. 1983a. Toxic ity of chlorinated paraffin to mussels (Mytilus edulis) over 60 days. Imperial Chemical Industries PLC, Devon, U.K. (Brixham Report No. BL/B/2289).

Madeley, J.R. and R.S. Thompson. 1983b. Toxic ity of chlorinated paraffin to mussels (Mytilus edulis) over 60 days. Imperial Chemical Industries PLC, Devon, U.K. (Brixham Report No. BL/B/2288).

Madeley, J.R. and R.S. Thompson. 1983c. Toxic ity of chlorinated paraffin to mussels (Mytilus edulis) over 60 days. Imperial Chemical Industries PLC, Devon, U.K. (Brixham Report No. BL/B/2290).

Madeley, J.R. and R.S. Thompson. 1983d. Toxic ity of chlorinated paraffin to mussels (Mytilus edulis) over 60 days. (iv) Chlorina ted paraffin – 58% chlorination of short chain length n-paraffins. Imperial Chemical I ndustries PLC, Devon, U.K. (Brixham Report No. BL/B/2291).

Madeley, J.R., E. Gillings and L.F. Reynolds. 1983a. The determination of the solubility of four chlorinated paraffins in water. Im perial Chemical Industries PLC, Devon, U.K. (Brixham Report No. BL/B/2301).

Madeley J.R., R.S. Thompson and D. Brown 1983b. The bioconcentration of a chlorinated paraffin by the common mussel (Mytilus edulis). Imperial Chemical Industries PLC, Devon, U.K. (B rixham Report No. BL/B/2351).

Marvin, C.H. 2003. Personal communicati on. National Water Research Institute, Burlington, Ontario.

Marvin, C.H., S. Painter, G.T. Tomy, G.A. Stern, E. Braekvelt and D.C.G. Muir. 2003. Spatial and temporal trends in short-chai n chlorinated paraffins in Lake Ontario sediments. Environ. Sci. Technol. 37(20): 4561–4568.

Metcalfe-Smith, J.L., R.J. Maguire, S.P. Ba tchelor and D.T. Bennie. 1995. Occurrence of chlorinated paraffins in the St. Lawrence Ri ver near a manufacturing plant in Cornwall, ON. Aquatic Ecosystem Protection Branch, National Water Research Institute, Environment Canada, Burlington, Ontario.

Meylan, W.M. and P.H. Howard. 1993. Comput er estimation of the atmospheric gasphase reaction rate of organic com pounds with hydroxyl radicals and ozone. Chemosphere 12: 2293–2299.

MOE (Ontario Ministry of th e Environment). 1998. Guidelin es for the utilization of biosolids and other wastes on agri cultural land. January revision.73 Muir, D.C.G., G.T. Tomy, G.A. Stern and A.T. Fisk. 1996. Preliminary report on concentrations of chlorinated n-paraffins in sediment and fish from the Detroit River. Contract report prepared for the U.S. Envi ronmental Protection Agency. University of Manitoba, Winnipeg, Manitoba (U.S. EP A OTS Document No. FYI-OTS-0496-1268; Microfiche No. NTIS/OTS0001268).

Muir, D., R. Wilkinson, C. Teixeira, D. Be nnie, M. Whittle, G. Tomy and G. Stern. 1999. Polychlorinated (C 10 –C 13 )-alkanes in the Great Lakes. In : P. Mocarelli (ed.)., Dioxin ’99. 19th International Symposium on Halogenate d Environmental Organic Pollutants and Persistent Organic Pollutants (POP s), Venice, Italy, September 12–17, 1999. Organohalogen Compd. 43: 93–96.

Muir, D.C.G., M. Alaee and G.A. Stern. 1999a. Polychlorinated (C 10 –C 13 ) n-alkanes (SCCPs) and brominated diphenyl ethers (BDP Es) in the Canadian environment. Paper presented at Workshop on Persistent Orga nic Pollutants and Heavy Metals, Durham, North Carolina, October 5-7, 1999.

Muir, D., G. Stern and G. Tomy. 2000. Chlorina ted paraffins. In: J. Paasivirta (ed.), The handbook of environmental chemistry. Vol. 3, Part K. New types of persistent halogenated compounds. Springer- Verlag, New York, N.Y. pp. 203–236.

Muir, D.C.G., D. Bennie, C. Teixeira, A.T. Fisk, G.T. Tomy, G.A. Stern and M. Whittle. 2001. Short chain chlorinated paraffins: Are they persistent and bioaccumulative? In: R. Lipnick, B. Jansson, D. Mackay and M. Patr eas (eds.), Persistent , bioaccumulative and

toxic substances. Vol. 2. AC S Books, Washington, D.C. pp. 184–202.

Muir, D., E. Braekevelt, G. Tomy and M. Whittle. 2002. Analysis of medium chain chlorinated paraffins in Great Lakes food webs and in a dated sediment core from Lake St. Francis in the St. Lawrence River system . Preliminary report to Existing Substances Branch, Environment Canada, Hull, Quebec. 9 pp.

Murray, T.M., D.H. Frankenberry, D.H. Steele and R.G. Heath. 1988. Chlorinated paraffins: A report on the findings from two fi eld studies, Sugar Creek, Ohio and Tinkers Creek, Ohio. Vol. 1. Technical report. U.S. Environmental Protection Agency, Washington, D.C. 150 pp. (EPA/560/5 87/012).

NHW (Department of National Health and We lfare). 1990. Present patte rns and trends in infant feeding in Canada. Department of Na tional Health and Welfare, Ottawa, Ontario [cited in EHD, 1998].

Nicholls, C.R., C.R. Allchin and R.J. La w. 2001. Levels of short and medium chain length polychlorinated n-alkanes in environmen tal samples from selected industrial areas in England and Wales. Environ. Pollut. 114: 415–430.

NICNAS (National Industrial Chemicals Notification and Assessment Scheme). 2001. Short chain chlorinated paraffins (SCCP s). Department of Health and Ageing,74 Government of Australia (Priority Exis ting Chemical Assessment Report No. 16) (<http://www.nicnas.gov.a u/publications/CAR/PEC/P EC16/PEC16index.htm>).

NTP (National Toxicology Program). 1986. NT P technical report on the toxicology and carcinogenesis studies of chlorinated paraffins (C 23 , 43% Cl) (CAS No. 63449-39-8) in F344/N rats and B6C3F1 mice (gavage studies). National Institutes of Health, Research Triangle Park, North Carolina (NTP TR 305; NIH Publication No. 86-2561).

NTP (National Toxicology Program). 1986a. NT P technical report on the toxicology and carcinogenesis studies of chlorinated paraffins (C 12 , 60% chlorine) (CAS No. 63449-39- 8) in F344/N rats and B6C3F1 mice (gavage studies). National In stitutes of Health, Research Triangle Park, North Carolina (N TP TR 308; NIH Publication No. 86-2564).

Omori, T., T. Kimura and T. Kodama. 1987. Bacterial cometabolic degradation of chlorinated paraffins. Appl. Mi crobiol. Biotechnol. 25: 553–557.

Peters, A., G. Tomy, K. Jones, P. Cole man and G. Stern. 2000. Occurrence of C10–C13 polychlorinated n-alkanes in the atmosphe re of the United Kingdom. Atmos. Environ. 34(19): 3085–3090 [cited in ISI, 2003].

Poon, R., P. Lecavalier, P. Chan, C. Viau, H. Hakansson, I. Chu and V.E. Valli. 1995. Subchronic toxicity of a medium -chain chlorinated paraffin in the rat. J. Appl. Toxicol. 15(6): 455–463.

Renberg, L., G. Sundström and K. Sundh-Nygå rd. 1980. Partition coefficients of organic chemicals derived from reversed phase thin layer chromatography. Evaluation of methods and application on phosphate esters, polychlorinated paraffins and some PCBsubstitutes. Chemosphere 9: 683–691.

Renberg, L., M. Tarkpea and G. Sundström. 1986. The use of the bivalve Mytilus edulis as a test organism for bioconcentration studies. Ecotoxicol. Environ. Saf. 11: 361–372.

Reth, M., Ciric, A., Christensen, G.N., Heim stad, E.S., and M. Oehme. 2006. Short- and medium-chain chlorinated paraffins in biota from the European Arctic- differences in homologue group patterns. Sci. Tot. Environ. 367: 252-260.

Rotard, W., W. Mailahn, S. Kuhn, A. Hermene it, S. Hildebrecht, S. Kemmlein and S. Mentzel. 1998. Analyse von Chloroparaffinen in Sediment, Muscheln und Krabben aus der Umgebung einer Produktionsanlage. Umed-Info 7, 19-22, Landesgesundheitsamt BadenWürttemberg [cited in U.K. Environment Agency 2001].

Schenker, B.A. 1979. Chlorinated paraffins. In : Kirk-Othmer encyclopedia of chemical technology. 3rd edition. Vol. 5. John W iley and Sons, New York, N.Y. pp. 786–791.

Serrone, D.M., R.D.N. Birtley, W. Weiga nd and R. Millischer. 1987. Summaries of toxicological data. Toxicology of chlorina ted paraffins. Food Chem. Toxicol. 25: 553–75 562.

Sijm, D.T.H.M. and T.L. Sinnige. 1995. Experime ntal octanol/water partition coefficients of chlorinated paraffins. Chemosphere 31: 4427–4435.

Stern, G.A. and M. Evans. 2003. Persistent organic pollutants in marine and lake

sediments. In: Canadian Arctic Contaminan ts Assessment Report II. Sources, occurrence, trends and pathways in th e physical environment. Northern Contaminants Program, Department of Indian Affairs and Nort hern Development, Ottawa, Ontario. pp. 100–115.

Stern, G., G. Tomy, D. Muir, J. West more, E. Dewailly and E. Rosenberg. 1998. Polychlorinated n-alkanes in aquatic biota and human milk. Presented at the American Society for Mass Spectrometry and Allied T opics, 45th Annual Conference, May, Palm Springs, California [cited in Muir et al., 2000].

Stevens, J.L., G.L. Northcott, G.A. Ster n, G.T. Tomy and K.C. Jones. 2002. PAHs, PCBs, PCNs, organochlorine pesticides, synt hetic musks and polychlorinated n-alkanes in UK sewage sludge: survey results and im plications. Environ. Sci. Technol. 37: 462–

467.

Thomas, G.O. and K.C. Jones. 2002. Chlorina ted paraffins in huma n and bovine milk-fat. A report on a research project funded by the Euro Chlor Chlorinated Paraffins Sector Group. Department of Environmental Sciences , Lancaster Universit y, Lancaster, U.K. [cited in U.K. Environment Agency 2003].

Thompson, R.S. and J.R. Madeley. 1983a. The ac ute and chronic toxicity of a chlorinated paraffin to Daphnia magna. Imperial Chemi cal Industries PLC, Devon, U.K. (Brixham Report BL/B/2358).

Thompson R. S. and Noble H. (2007). Short-chain chlori nated paraffins (C 10 - 13 , 65% chlorinated): Aerobic and anaerobic transfor mation in marine and freshwater sediment systems. Draft Report No BL8405/B. Brixha m Environmental Laboratory, AstraZeneca UK Limited. (as cited in U.K. Environment Agency, 2007)

Thompson, R.S., N.J. Williams and E. Gillings. 1997. Chlorinated paraffin (52% chlorinated, C 14–17 ): Chronic toxicity to Daphnia magna. Zeneca Ltd., Devon, U.K. (Zeneca Confidential Report BL5875/B).

Thompson, R.S., E. Gillings and R.I. Cumm ing. 1998. Short-chain chlorinated paraffin (55% chlorinated): Determination of organi c carbon partition coefficient. Zeneca Ltd., Devon, U.K (Zeneca Confidential Report BL6426/B).

Thompson, R.S., J.E. Caunter and E. Gillings . 2000. Medium-chain chlorinated paraffin (51% chlorinated n-pentadecane-8-14C): Bi oconcentration and elimination by rainbow trout (Oncorhynchus mykiss). AstraZeneca, U.K. (AstraZeneca Confidential Report BL6869/B) [cited in U.K. Environment Agency 2003].76 Thompson, R.S., A.J. Windeatt and E. G illings. 2001a. Medium-chain chlorinated paraffin (52% chlorinated, C 14

– 17 ): Effects in soil on seed germination and vegetative growth of wheat (Triticum aestivum), oils eed rape (Brassica napus) and mung bean (Phaseolus aureus). AstraZeneca, U.K. (AstraZeneca Confiden tial Report BL7128/B) [cited in U.K. Environment Agency 2003].

Thompson, R.S., D.V. Smyth and E. Gillings . 2003. Medium-chain chlorinated paraffin (52% chlorinated, C 14

– 17 ): Effects in sediment on th e survival, growth and sexual development of the freshwater amphipod, Hyalella azteca. AstraZeneca, U.K. (AstraZeneca Confidential Report BL7469/B) [cited in U.K. Environment Agency 2003].

Tomy, G. 1997. The mass spectrometric char acterization of polychl orinated n-alkanes and the methodology for their analysis in the environment. Ph.D. thesis, University of Manitoba, Winnipeg, Manitoba [c ited in Tomy et al., 1998].

Tomy, G.T. and G.A. Stern. 1999. Analysis of C 14 –C 17 polychloro-n-alkanes in environmental matrixes by accelerated so lvent extraction–high-resolution gas chromatography/electron capture negative ion high-resolution mass spectrometry. Anal.

Chem. 71: 4860–

Tomy, G. and G. Stern. 2000. Simultaneous quantitation of short and medium chain polychlorinated n-alkanes in environmental samples by HRGC/ECNI-HRMS. In: M.S. Denison (ed.)., Dioxin 2000. 20th Inte rnational Symposium on Halogenated Environmental Organic

Tomy, G.T., G.A. Stern, D.C.G. Muir, A.T. Fisk, D. Cymbalisty and J.B. Westmore. 1997. Quantifying C 10 –C 13 polychloroalkanes in environmental samples by high resolution gas chromatography/electron captu re negative ion mass spectrometry. Anal.

Chem. 69: 27

Tomy, G.T., A.T. Fisk, J.B. Westmore and D.C.G. Muir. 1998a. Environmental chemistry and toxicology of polychlorinated nalkanes. Rev. Environ. Contam. Toxicol.

158: 53–128.

Tomy, G., G. Stern, K. Koczanski and T. Halldorson. 1998b. Polychloro-n-alkanes in beluga whales from the Arctic and the St. Lawrence River estuary. Organohalogen Compd. 35: 399–401.

Tomy, G.T., G.A. Stern, W.L. Lockhart and D.C.G. Muir. 1999. Occurrence of C 10 –C 13 polychlorinated n-alkanes in Ca nadian mid-latitude and Arc tic lake sediments. Environ.

Sci. Technol. 33: 2858–2863.

Tomy, G.T., D.C.G. Muir, G.A. Stern and J.B. Westmore. 2000. Levels of C 10 –C 13 polychloro-n-alkanes in mari ne mammals from the Arctic and the St. Lawrence River estuary. Environ. Sci. Technol. 34: 1615–1619.77 U.K. Environment Agency. 2001. Long-chain ch lorinated paraffins. Environmental risk assessment report. Draft, November. Prepar ed by Building Research Establishment Ltd. for Chemicals Assessment Section, U. K. Environment Agency, Wallingford, Oxfordshire, U.K. 184 pp.

U.K. Environment Agency. 2003. Risk assessment of alkanes, C 14–17 , chloro. Draft document, February. Prepared by Building Research Establishment Ltd. for Chemicals Assessment Section, U.K. Environment Agency, Wallingford, Oxfordshire, U.K. 326 pp.

U.K. Environment Agency. 2007. Risk assessment of alkanes, C 14–17 , chloro. Draft document, August. Prepared by Building Re search Establishment Ltd. for Chemicals Assessment Unit, U.K. Environment Agenc y, Wallingford, Oxfordshire, U.K. 146 pp.

U.S. EPA (United States Environmental Protection Agency). 1991. Alpha-2u-globulin: association with chemically-indu ced renal toxicity and neoplasia in the male rat. Prepared for the Risk Assessment Forum (EPA/625/3-91/019A).

U.S. EPA (United States Environmental Protection Agency). 1998. Assessment of thyroid follicular cell tumors. Risk Assessment Forum (EPA/630/R-97/002).

U.S. NRC (National Research Council). 2000. Chlorinated paraffins. In: Toxicological

risks of selected flame-retardant chemicals. Board on Environmental Studies and Toxicology, Committee on Toxicology, Subcom mittee on Flame-Retardant Chemicals. National Academy Press, Washington, D.C. pp. 440–491.

Warnasuriya, G., J. Foster, B. Elcombe and C. Elcombe. 2000. Mechanisms of nephrocarcinogenicity of a short chain chlo rinated paraffin. Toxicol. Sci. 54(1): 270 (Abstract 1266).

Warngard, L., Y. Bager, Y. Kato, K. Kenne and U.G. Ahlborg. 1996. Mechanistical

studies of the inhibition of intercellular communication by organochlorine compounds. Arch. Toxicol. 18(Suppl.): 149–159.

Wyatt, I., C.T. Coutts and C.R. Elcombe. 1993. The effect of chlorinated paraffins on hepatic enzymes and thyroid hor mones. Toxicology 77(1/2): 81–90.

Zitko, V. 1974. Uptake of chlorinated paraffi ns and PCB from suspended solids and food by juvenile Atlantic salmon. Bull. Environ. Contam. Toxicol. 12: 406–412.