SEPA
United States Premanufacture Notification
Environmental Protection Agency Number: P-12-0453-0433
Office of Chemical Safety and
Pollution Prevention
TSCA New Chemicals Review Program
Standard Review Risk Assessment on
Medium-Chain Chlorinated Paraffins
(PMN P-12-0453)
AND
Long-Chain Chlorinated Paraffins
(PMN P-12-0433)
This assessment was conducted under EPA's TSCA Section 5 New Chemicals
Review Program. EPA is assessing Medium-Chain Chlorinated Paraffin (MCCP)
and Long-Chain Chlorinated Paraffin (LCCP) chemicals as part of its New
Chemicals Review program. As with all Premanufacture Notice (PMN)
submissions, EPA followed the approaches, methods and statutory provisions of
TSCA section 5 for the chlorinated paraffin PMNs assessments.
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CONCLUSIONS
Based on its assessment of the available surrogate hazard and exposure information on P-12-
0453 and P-12-0433, EPA/OPPT concludes the following pertaining to the manufacturing,
processing and use of these PMN substances:
1. Occupational Exposures: given the assumptions, data and scenarios evaluated in this
assessment, there were low risks found for workers from either dermal or inhalation
exposures.
2. General Population Exposures (from environmental releases): given the assumptions, data
and scenarios evaluated in this assessment, there were low risks found to humans from
environmental releases via exposure to either drinking water or fish ingestion.
3. Environmental Assessment:
a. Using estimated environmental concentrations, the PMN substances may present an
unreasonable risk following acute and chronic exposures to aquatic organisms.
b. Using available measured concentrations of MCCP and LCCP congener groups in the
environment as supporting information, the PMN substances:
i. Are expected to partition to sediment and may partition to soil through land
application of biosolids and,
ii. May be released to the environment at levels at or above estimated concentrations of
MCCP and LCCP congener groups that may present an unreasonable risk
following acute and chronic exposures to aquatic organisms.
4. PBT Assessment: The PMN substances may be very persistent and very bioaccumulative.
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TABLE OF CONTENTS
TABLE OF CONTENTS 3
1 INTRODUCTION 6
1.1 PMNS RECEIVED 6
1.2 CHEMISTRY 6
1.3 USES 9
2 ENVIRONMENTAL FATE 9
2.1 ENVIRONMNETAL PERSISTENCE 9
2.2 BIOCONCENTRATION AND BIOACCUMULATION 11
3 ECOLOGICAL HAZARD OVERVIEW 12
4 HUMAN HEALTH HAZARD OVERVIEW 15
4.1 MCCP HEALTH DATA REVIEW 15
4.2 LCCP HEALTH DATA REVIEW 17
5 EXPOSURE INFORMATION 18
5.1 ENVIRONMENTAL MONITORING 18
5.2 MODELED ENVIRONMENTAL RELEASES 21
5.3 EXPOSURE ESTIMATES 23
5.3.1 OCCUPATIONAL EXPOSURE ESTIMATES 23
5.3.2 CONSUMER EXPOSURE ESTIMATES 23
6 RISK ASSESSMENT 24
6.1 ENVIRON MENTAL ASSESSMENT 24
6.1.1 Risk Estimates Using Environmental Monitoring Concentrations 24
6.1.2 Risk Estimates Using Modeled Exposures 26
6.2 HUMAN HEALTH 27
6.2.1 Workers 28
6.2.2 General Population 29
7 CONCLUSIONS 31
8 REFERENCES 32
9 APPENDICES 46
Appendix A ENVIRONMENTAL FATE AND BIOACCUMULATION STUDY SUMMARIES 47
A-l ENVIRONMENTAL PERSISTENCE 47
A-l-1 Abiotic Degradation 47
A-l-1-1 Fate in Air 47
A-l-2 Biodegradation 48
A-l-2-1 Fate in Wastewater Treatment 48
A-l-2-2 Fate in Surface and Groundwater 49
A-l-2-3 Fate in Soil 49
A-2 BIOCONCENTRATION AND BIOACCUMULATION 53
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Appendix B ECOTOXICITY STUDY SUMMARIES 58
B-l MCCP ECOTOXICITY DATA 58
B-l-1 Acute Fish Toxicity 58
B-l-2 Acute Aquatic Invertebrate Toxicity 59
B-l-3 Algae Toxicity 62
B-l-4 Chronic Fish Toxicity 63
B-l-5 Chronic Aquatic Invertebrate Toxicity 65
B-l-6 Chronic Aquatic Sediment Invertebrate Toxicity 68
B-l-7 Avian Toxicity 70
B-l-8 Terrestrial Invertebrate Toxicity 71
B-l-9 Terrestrial Plant Toxicity 72
B-l-10 Conclusions 73
B-2 LCCP ECOTOXICITY DATA 74
B-2-1 Acute Fish Toxicity 74
B-2-2 Acute Aquatic Invertebrate Toxicity 75
B-2-3 Aquatic Plant Toxicity 76
B-2-4 Chronic Fish Toxicity 77
B-2-5 Chronic Aquatic Invertebrate Toxicity 77
B-2-6 Chronic Aquatic Sediment Invertebrate Toxicity 81
B-2-7 Avian Toxicity 81
B-2-8 Terrestrial Invertebrate Toxicity 81
B-2-9 Terrestrial Plan t Toxicity 81
B-2-10 Conclusions 81
Appendix C HUMAN HEALTH HAZARD STUDY SUMMARIES 83
C-l MCCP HEALTH DATA REVIEW 83
C-l-1 Metabolism 84
C-l-2 Acute Toxicity 84
C-l-3 Irritation and Sensitization 85
C-l-4 Repeated-dose Toxicity 85
C-l-5 Genotoxicity 86
C-l-6 Carcinogenicity 86
C-l-7 Developmental Reproductive Toxicity Review 91
C-2 LCCP HEALTH DATA REVIEW 97
C-2-1 Metabolism 97
C-2-2 Acute Toxicity 98
C-2-3 Irritation and Sensitization 98
C-2-4 Repeated-dose Toxicity 98
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C-2-5 Genotoxicity 99
C-2-6 Carcinogenicity 99
C-2- 7 Developmental Reproductive Toxicity Review 99
Appendix D ENVIRONMENTAL MONITORING 102
D-l MCCP MONITORING DATA 102
D-l-1 Surface Water 102
D-l-2 Sediment 104
D-l-3 Biosolids and Soil 113
D-l-4 Biota 113
D-2 LCCP MONITORING DATA 116
Appendix E ENGINEERING (ChemSTEER) REPORTS ON P-12-0-0433 and P-12-0453 117
Appendix F EXPOSURE SCENARIO ESTIMATES 121
Appendix G SUPPLEMENTAL INFORMATION 123
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1 INTRODUCTION
1.1 PMNS RECEIVED
INEOS Chlor Americas, Inc. (hereinafter "INEOS") submitted two Premanufacture Notices
(PMNs) identified by EPA/OPPT as either medium-chain chlorinated paraffins (MCCPs; P-12-
0453) of varying chain lengths with the formula CxH(2x-y+2)Cly and "x" equaling 14 to 17 and "y"
equaling 6 to > 24 or long-chain chlorinated paraffins (LCCPs; P-12-0433) of varying chain
lengths with the formula of CxH(2x-y+2)Cly and "x" equaling 18 to 20 and "y" equaling 6 to > 24.
Table 1 lists the basic information INEOS supplied on these two PMNs which are intended to be
sold under the trade name "Cereclorฎ".
Table 1: Identification, Production Volume and Use of P-12-0453 and P-12-0433
PMN
Chemical Name
1st Year
Production
Volume (kg)
% PMN
in final
Product
Uses
Log
Kow
Water
Solubility
74%: lubricant in
MWFs;
17%: flame
P-12-
0453
Alkanes, C14-17, chloro
(MCCPs; 40-60 wt%
CI)
CBI
5-20
retardant/plasticizer in
polymers;
7%: plasticizer in
adhesives; and
2%: lubricant in
sealants.
4.70
(E)
<0.03
mg/L (E)
P-12-
0433
Alkanes, C18-20, chloro
(LCCPs; 40-55 weight
percent chlorination
[wt% CI])
CBI
15
100%: lubricant in
metal working fluids
(MWFs).
7.46
(E)
< 0.006
mg/L (E)
E = Estimated
Though the specific PMNs in this application include MCCPs and LCCPs, this standard review
presents data and information on short-chain chlorinated paraffins (SCCPs) and on very long-
chain chlorinated paraffins (vLCCPs) analogs. The continuum of carbon chain length and
degree/percent chlorination (wt% CI) in all of the chlorinated paraffins (CPs) is important and
the relationship among them needs to be kept in mind.
L2 CHEMISTRY
Shown below are the structures and chlorine content of P-12-0453 and P-12-0433 products are
shown below.
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CI CI CI CI CI CI CI CI CI
CI
P-12-0453: the average molecular formula ranges from C14H26C14 (low weight at -40 wt% CI)
to C17H26C110 (high weight at -60 wt% CI). The parent hydrocarbon had the following
measured composition: C14, 36 wt% (range 30-40 wt%); C15, 30 wt% (range 25-35 wt%); C16,
24 wt% (range 20-30 wt%); and C17, 10 wt% (8-18 wt%).
CI CI CI CI CI CI
P-12-0433: the average molecular formula ranges from C17H31CI5 (low weight -40 wt% CI) to
C20H34CI8 (high weight -50 wt% CI). The parent hydrocarbon had the following measured
composition: C17, 17.06 wt% (max 20 wt%); Ci8, 64.71 wt% (range 45-70 wt%); C19, 13.67 wt%
(range 15-27 wt%); and C20, 2.13 wt% (range 4-12 wt%).
CPs have an unknown or variable composition (classified as UVCB1 compounds for TSCA
Inventory purposes) of polychlorinated n-alkanes. The carbon chain length usually varies
between 10 and 30 carbon atoms and the degree of chlorination can vary between 30 and 75
wt%. EPA/OPPT subdivides CPs according to their carbon chain length into the following
categories:
1. SCCPs (C10-13)
2. MCCPs (C14-17)
3. LCCPs (C18-20)
4. Very long-chain CPs (vLCCP, C>2o)
SCCPs and MCCPs exist as liquids at standard temperature and pressure. CPs with a carbon
chain length > 18 are subdivided based on their physical state, which is a function of chain length
and chlorine content. The LCCPs and vLCCPs up to 70 wt% CI are typically liquids (40 - 55
wt% CI) while above 70 wt% CL they are waxy solids.
CP products contain a variety of carbon chain lengths that have been chlorinated to different
degrees {i.e., variation in the number and position of the chlorine atoms on the carbon chain).
The individual isomer content of commercial CPs is rarely identified because the number of
1 UVCB are chemical substances whose composition is Unknown or Variable compositions, Complex reaction
products and Biological materials.
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possible individual congener group is extremely large. Consequently, the physicochemical
properties of CPs vary by carbon chain length and chlorine content. Increased molecular weight
correlates to higher melting and boiling points, lower vapor pressures and water solubilities, and
greater LogKow (logarithm of octanol:water partition coefficient). EPA/OPPT used the
physicochemical properties listed in Table 2 for informing its evaluation of P-12-0453 and P-12-
0433.
Table 2: Summary of Physiochemical Information3'1*
wt% CI
Melting Point
Boiling Point
Vapor Pressure
Water
Solubility
Log Kow
MCCPs
>40
< 25ฐC
(pour point)
> 200ฐC (dec)
<0.036 Pa
at 20 ฐC
27 (ig/L
at 20ฐC
>5.5 (measured)
8.30 (estimated)c d
LCCPs
>40
< 25ฐC
(pour point)
> 200ฐC (dec)
< 2.7 x 10"4 Pa
at 20 ฐC
5 Hg/L
at 20ฐC
>8
"Source: EURAR (ECB, 2008); EA (2009)
bBecause most CP products are liquids and the CPs begin decomposing at 200 ฐC (via loss of HC1), melting point
and boiling points are considered less important in characterizing hazard and risk.
ฐValue calculated using the KOWWIN Program (vl.68) available in EPA/OPPT's Estimation Programs Interface
(EPI) Suite TM. This estimate was generated using a representative MCCP (i.e., C14H24CI6, 52 wt% CI) with the
following SMILES notation: CCC(C1)CC(C1)CCC1)CC(C1)CC(C1)CC(C1)C.
dThe EURAR (ECB, 2008) cited Renberg's liquid chromatography to measure a LogKow between 5.5 and 8.2
and then chose to use a Log K0w = 7 as a representative LogKow for MCCPs with 45-52 wt% CI.
Analytical challenges exist with evaluating CPs due to the sheer number of congener groups2
that may be present in CP products. The existence of multiple chain lengths in the UVCBs such
as P-12-0453 and P-12-0433 requires the use of analytical methods that separate congener
groups based on retention time in a column and mass spectrum of the respective peaks. Several
lines of evidence support the use of a representative SCCP or MCCP product as a surrogate for
congener groups present in MCCP or LCCP commercial products, respectively.
Hiittig (2006) and Hiittig and Oehme (2006) reported that most commercial MCCP products
were > 60 wt% of the C14 chain-length congener groups and the C15 chain-length congener
groups comprised the majority of the remaining ~ 40 wt%. The authors found <15 wt% Ci6
chain-length congener groups present in most commercial MCCP samples and little to no C17
chain-length congener groups. Additional studies have reported that the C14 and C15 chain-length
congener groups are the predominant MCCPs present in environmental media and in human
breast milk (Bayen et al., 2006; Chen et al., 2011; Reth et al., 2006; Wang et al., 2013). Some
variation is possible in commercial products where the C14 and C15 chain-length congener groups
may not be the predominant congener groups in a specific MCCP product; however, even in
these products, the C14 and C15 chain-length congener groups may serve as reasonable worse
case surrogates for the C16-17 chain-length congener groups, due to their greater bioavailability
and mobility in environmental media (ECB, 2005).
2 For this report, congener groups are used to recognize the existence of different chain lengths and degrees of
chlorination that could be present in any given CP product.
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These analyses, in conjunction with measured or estimated physicochemical and environmental
fate properties, allow for the reasonable use of associating commercial products with CP levels
found in environmental media and biota. For example, the experimental observation that MCCPs
(C14-17) are abundant in sediment has been explained using known water solubility and vapor
pressure values in conjunction with predicted degradation pathways (de Boer, 2010). EPA/OPPT
determined that the toxic endpoints of interest were measured for only one CP commercial
product {i.e., Cereclor S52ฎ). Therefore, this approach of relating one commercial product with
other commercial products is critical for attributing the hazard characterization to CPs of
different sources. Thus available information (hazard and environmental monitoring data on one
commercial product (Cereclor S52ฎ) is used as representative CP for this assessment.
Furthermore, experimental data on SCCPs show that these products are more toxic than the
longer chain CPs {e.g., MCCPs). Therefore, when endpoint specific data were lacking for the
PMNs, EPA/OPPT used measured data from SCCPs as surrogates for potential hazard and risks
for the PMNs.
1.3 USES
INEOS reported four uses in the PMN for P-12-0453, including:
1) 73.8% as a lubricant in metal working fluids (MWF) (no commercial or consumer uses). The
notification substance is blended into the MWFs at 15%.
2) 16.8%) as a flame retardant/plasticizer in poly(vinyl chloride) (PVC) resins (no commercial or
consumer uses). The notification substance is blended into the PVC resins at 15%.
3) 7.4%) as a plasticizer in adhesives (no commercial or consumer uses). The notification
substance is blended into the adhesives at 5%.
4) 2% as a lubricant in sealants (no commercial or consumer uses). The notification substance is
blended into the sealants at 20%.
INEOS reported one category of use in the PMN for P-12-043 3. This PMN substance is intended
for use solely {i.e., 100% of the production volume) as a lubricant in MWF (no commercial or
consumer uses). The PMN substance will be used at 15% in the final working formulation.
2 ENVIRONMENTAL FATE
EPA/OPPT reviewed available information on the environmental fate of MCCPs and LCCPs in
different environmental compartments and the properties that control transport (summarized in
Appendix A). In addition, EPA/OPPT reviewed assessments performed by Canada (EC, 2008a)
and the EU (EA, 2009; ECB, 2005) to inform its assessment.
2.1 ENVIRQNMNETAL PERSISTENCE
Abiotic studies have shown that MCCPs and LCCPs are stable to hydrolysis and to direct
photolysis in water and air. In laboratory studies using hydrocarbon solvents, CPs were shown to
poorly absorb ultraviolet (UV) light and no direct photodegradation was observed. The
atmospheric half-life for MCCPs and LCCPs has been estimated at 1 - 2 days (EA, 2009; ECB,
2005), based on estimated values for the second order rate constant for reaction with atmospheric
hydroxyl radicals for MCCPs (40-56 wt%> CI) and LCCPs (42-54 wt%> CI) (EA, 2009; ECB,
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2005). The persistence of MCCPs increases with carbon chain length and higher chlorine content
(EA, 2009; ECB, 2005).
Existing biotic degradation data suggest there are a number of microbial species that are capable
of degrading shorter chain, lower chlorinated MCCP congeners. Longer and higher chlorinated
chemicals also may be degraded, but at much slower rates (Allpress and Gowland, 1999; Muir,
2010; Omori et al., 1987). The results from laboratory studies of microbial metabolism, using
both isolated species and mixed cultures of acclimated microbes, show that MCCPs and LCCPs
may be degraded by direct metabolism or co-metabolism by some microbes and microbial
consortia in soil, wastewater treatment systems, sediment and other environmental media.
Overall, the existing studies suggest that with microbial degradation, dechlorination and carbon
chain cleavage may be possible in some media (see Table A-l); however, the degree of
degradation is generally low (Allpress and Gowland, 1999; Muir, 2010; Omori et al., 1987).
In general, MCCP and LCCP congeners with longer chain lengths and higher degrees of
chlorination are expected to be highly persistent in some environmental compartments. In
contrast, shorter and lesser-chlorinated congeners are likely to degrade rapidly, especially in
aerobic environments. Because persistence increases with chain length, LCCPs are generally
expected to be more persistent than MCCPs with comparable degrees of chlorination (EA, 2009;
ECB, 2005).
Based on the review of available literature and studies submitted by various manufacturers,
including confidential business information (CBI) not publicly available, EPA/OPPT's
conclusions regarding environmental persistence of MCCPs and LCCPs are consistent with those
provided by Canada and the EU.
The Canadian assessment on MCCPs and LCCPs states (EC, 2008a):
"Information on physical properties of MCCPs, and especially LCCPs, is limited. Values
used in this assessment are based on extrapolations mainly from SCCPs or QSARs. The
analysis of SCCPs and MCCPs in sediment cores and associated calculations provide
strong evidence for the persistence of these substances in the environment. Even though
there are no data for persistence of LCCPs in sediment, based on biodegradation data
which indicate increasing stability with increasing carbon chain length, it is reasonable
to conclude that LCCPs are persistent in sediment. "
The EU assessment on MCCPs states (ECB, 2005):3
"No standard ready or inherent biodegradation tests results are available for medium-
chain chlorinated paraffins. From the available information, medium-chain chlorinated
paraffins can be considered to be not biodegradable in such test systems and so a
biodegradation rate MCCPs of 0 day-1 is used in the risk assessment.
3 Note, since the EU issued its assessment in 2005, standard inherent biodegradation studies were performed and are
summarized in Appendix A.
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There is evidence that some microorganisms may be capable of degrading MCCPs in the
environment in acclimated or co-metabolic systems. The potential for biodegradation
appears to increase with decreasing chlorine content. However, it is not possible from
the available data to derive rate constants for biodegradation in soil, surface water and
sediment systems. As a worst case approach, no biodegradation will be assumed in these
media in the PEC calculations.
Hydrolysis is not expected to be a significant degradation process for medium-chain
chlorinated paraffins in the environment. An atmospheric half-life of 1-2 days is
estimatedfor reaction with hydroxy I radicals. A value for the rate constant for the
reaction (kon) of 8 x 10~12 cm3 molecule'1 s'1 is usedfor the environmental modelling in
the risk assessment."
The UK assessment on LCCPs4 concluded the following (EA, 2009):
"Based on the laboratory studies and other data available, LCCPs are unlikely to be
readily or inherently biodegradable. Although there is some evidence that they may
biodegrade in the environment, it is thought likely that this process will be sufficiently
slow that LCCPs meet the P or vP (very persistent) criteria. "
EPA/OPPT generally concurs with these characterizations. In the absence of information on
specific congener groups and data for MCCP or LCCP products, EPA/OPPT concludes that at
least some congener groups present in both MCCP and LCCP products are persistent to very
persistent; with estimated half-lives in air exceeding 2 days and estimated half-lives in water,
sediment and soil exceeding 2 months (60 days) (ECB, 2005; EA, 2009).
2.2 BIOCONCENTRATION AND BIOACCUMULATION
Recent reviews of the potential for MCCPs and LCCPs to bioaccumulate have shown that, while
data are limited, some congener groups are bioaccumulative or very bioaccumulative (EC,
2008a; ECB, 2005; Houde et al., 2008; Thompson and Vaughan, 2014). A summary of studies
reviewed by EPA/OPPT is provided in Appendix A.
Based on EPA/OPPT's review of existing studies (Bengtsson et al., 1979; CPC, 1980, 1983a,
1983b; Fisk et al., 1999; Fisk et al., 1998; Houde et al., 2008; Madeley and Maddock, 1983a,
1983b; Madeley and Thompson, 1983; Renberg et al., 1986; Thompson et al., 2000),
EPA/OPPT concludes that the bioconcentration potential of MCCP and LCCP congener groups
varies with the chain lengths and degree of chlorination and species evaluated. Shorter and less
chlorinated chemicals are readily taken up by organisms but also may be excreted or degraded
after absorption (Arnot, 2013). Longer and more highly chlorinated chemicals are typically not
absorbed across cellular membranes and are not accumulated in tissues. Some MCCP chemicals
with intermediate chain length and chlorination may be absorbed and retained. The available
evidence for MCCP and LCCP congener groups with intermediate chain lengths and chlorination
suggests that some may have bioconcentration factors (BCFs) or bioaccumulation factors (BAFs)
4 Note, the UK assessment evaluated the following CPs under the term LCCP: C18-20 liquid products, C>2o liquid
products, and C>2o solid products.
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greater than 1000 or 5000 (EC, 2008a; ECB, 2005, 2008). This suggests that some congener
groups in MCCP and LCCP products may be bioaccumulative or very bioaccumulative.
The Canadian assessment on MCCPs and LCCPs states (EC, 2008a):
"On the basis of the available information, and in particular the field BAF estimates, it is
concluded that MCCPs are bioaccumulative substances... "
"On the basis of the available information, and in particular the BAF model and
empirical BMF estimates, it is concluded that C18-20 liquid LCCPs are bioaccumulative
substances... " [page 27] and "... While there is a lack of empirical bioaccumulation data
for LCCPs, the modelling results provided by the Modified Gobas BAF Model - which
suggest that of all the LCCPs congeners only liquid C18-20 LCCPs have significant
bioaccumulation potential are considered credible. "
The UK assessment on LCCPs states (EA, 2009):
"The available data for LCCPs do show that uptake into fish from food occurs in the
laboratory, and that this uptake can be significant in some cases. The degree of uptake
appears to be highest for the C18-20 liquid chlorinated paraffins, but uptake of C>20
liquid chlorinated paraffins has also been demonstrated. The uptake of the highly
chlorinated C>20 solid chlorinated paraffins from food appears to be minimal. "
EPA/OPPT generally concurs with these characterizations. In the absence of information on
specific congener groups and data for MCCP or LCCP products, EPA/OPPT concludes that at
least some congener groups present in both MCCP and LCCP products are bioaccumulative to
very bioaccumulative based on multiple lines of evidence, including: Log Kow values, modeled
BCFs, laboratory-measured BCFs, field-measured BAFs, field-measured biomagnification
factors (BMFs), laboratory-measured biota-sediment accumulation factors (BSAFs) and the
presence of MCCPs and LCCPs in human and wildlife biota.
3 ECOLOGICAL HAZARD OVERVIEW
The available ecotoxicity data on MCCPs and LCCPs are summarized in Appendix B, along
with the criteria EPA/OPPT used for identifying the highest quality studies. Ecotoxicity studies
for MCCPs have been conducted in fish, aquatic invertebrates and plants, sediment and soil
invertebrates, and terrestrial plants and invertebrates. Though no avian reproduction studies were
available on MCCPs, a high quality study was available on an SCCP product (C10-12, 58 wt% CI)
with similar physicochemical properties to MCCPs and was used for informing EPA/OPPT's
hazard evaluation (ECB, 2000).
For LCCPs, ecotoxicity studies were only identified for aquatic invertebrates and vertebrates. No
data were available on sediment-dwelling or terrestrial organisms. Overall, the available data on
LCCPs were of low quality; therefore, the EPA/OPPT used data on MCCPs to inform its hazard
evaluation of LCCPs. This decision was considered a reasonable worst-case scenario because P-
12-0433 contains up to 20 wt% C17, a component of MCCPs (e.g., P-12-0453 contains between
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8-18 wt% C17). EPA/OPPT concludes that the studies summarized in Table 3 were the highest
quality for assessing potential hazards in the aquatic, sediment and terrestrial compartments.
Table 3: Summary of Aquatic, Sediment and Terrestrial Ecotoxicity Data for MCCPs and LCCPs
Test Substance
Test Organism
(Species)
Test Guideline; Study type
End-
point
Value3
Reference
A(|ii;ilic ln\erlcbrales
Cereclor S-52
(52 wt% CI,
C14-17)
Water flea
(Daphnia magna)
OECD 202, 1984; Acute
immobilization test
EC50
0.0059
CPA (1996)
Cereclor S-52
(52 wt% CI,
C14-17)
Water flea
(Daphnia magna)
OECD 202-Part II, 1984;
Reproduction test
NOEC
LOEC
MATC
0.01
0.018
0.013
Thompson,
Williams et al.
(1997)
Sediment-Dwelling Invertebrates
Cereclor S-52
(52 wt% CI,
C14-17)
Amphipod
(Hyalella azteca)
OECD 218- Draft, 2001; 28-
day prolonged sediment
toxicity study
NOEC
LOEC
MATC
130
270
187
Thompson et al.
(2002)
Terrestrial Invertebrates
Cereclor S-52
(52 wt% CI,
C14-17)
Earthworm
(Eisenia fetida)
OECD Guideline-Draft,
2000; 28-day reproductive
toxicity test
NOEC
LOEC
MATC
79
280
149
Thompson et al.
(200 Id)
Terrestrial Vertebrates
Commercial CP
(58 wt% CI,
C10-12)
Mallard duck
(Anas
platyrhynchos)
EPA 560/6-82-002; 22-week
reproduction test
NOEC
LOEC
168
1000
ECB (2000)
aUnits are mg/L for aquatic invertebrates, mg/kg dry weight (dw) sediment for sediment-dwelling invertebrates;
mg/kg dw soil for earthworm study; and mg/kg diet for the duck study.
Using the concentrations in the "value" column in Table 3 to represent hazard, EPA/OPPT
derived concentrations of concern (COCs) by applying assessment factors of five or ten for acute
or chronic exposures, respectively, which account for laboratory variability and represents
species sensitivity distributions (following US EPA, 2012). The COCs derived for aquatic-,
sediment- and terrestrial-dwelling organisms are explained below and summarized thereafter in
Table 4.
The most reliable and acceptable toxicity studies to aquatic organisms indicate that for MCCPs
and LCCPs, the toxicity to aquatic organisms are from the CPA (1996) study for acute toxicity
and the Thompson et al. (1997) study for chronic toxicity.
Acute COC: The 48-hour EC50 value 0.0059 mg/L is divided by an assessment factor of 5
to yield an acute concentration of concern (COC) of 0.00118 mg/L, or 0.001 mg/L, or 1
[igfL (1 ppb).
Aquatic Acute COC = 1 ppb.
Chronic COC: The chronic value 0.013 mg/L is divided by an assessment factor of 10 to
yield 0.0013 mg/L or 1.3 [j.g/L or 1.3 ppb.
Aquatic Chronic COC = 1 ppb.
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The most reliable and acceptable value for the acute toxicity to aquatic sediment invertebrate
organisms is based on the MCCP material from the Thompson et al. (2002) 28-d study. The 28-d
sediment invertebrate GMATC value of 187 mg/kg dry wt sediment is used to assess hazard.
Using methods in US EPA (2012):
Acute COC: The chronic value 187 mg/kg dry wt. is multiplied by an acute to chronic
ratio for invertebrates (10) to yield 1,870 mg/kg dry wt. This value is then divided by an
assessment factor of 5 to yield 374 mg/kg dry wt.
Aquatic Sediment Acute COC = 374 mg/kg dry wt sediment.
Chronic COC: The 28-d sediment invertebrate GMATC of 187 mg/kg dry wt sediment is
divided by an assessment factor of 10 to yield 18.7 mg/kg dry wt sediment.
Aquatic Sediment Chronic COC = 18.7 mg/kg dry wt sediment.
The most reliable and acceptable value for acute toxicity to terrestrial invertebrates is based on
the MCCP material from the Thompson et al. (2001a) study. The 28-d terrestrial invertebrate
GMATC value of 149 mg/kg dry wt soil from this study will be used. Using methods in US EPA
(2012):
Acute COC: To calculate an acute concern concentration from the chronic value the
value 149 mg/kg dry wt, is multiplied by an acute to chronic ratio for invertebrates (10)
to yield 1,490 mg/kg dry wt. This value is then divided by an assessment factor of 5 to
yield 298 mg/kg dry wt.
Terrestrial Invertebrate Acute COC = 298 mg/kg dry wt.
Chronic COC: The 28-d terrestrial invertebrate GMATC of 149 mg/kg dry wt is divided
by an assessment factor of 10 to yield 14.9 mg/kg dry wt.
Terrestrial Invertebrate Chronic COC = 14.9 mg/kg dry wt.
The most reliable and acceptable value for acute toxicity to terrestrial vertebrates is based on the
MCCP material from the ECB (2001) study. The 22-week terrestrial vertebrate NOEC value of
168 mg/kg dry wt soil from this study will be used. Using methods in US EPA (2012):
Acute COC: To calculate an acute concern concentration from the chronic value the
value 168 mg/kg diet is multiplied by an acute to chronic ratio for invertebrates (10) to
yield 1,680 mg/kg diet. This value is then divided by an assessment factor of 5 to yield
336 mg/kg diet.
Terrestrial Vertebrate Acute COC = 336 mg/kg diet.
Chronic COC: The 22-week terrestrial vertebrate NOEC of 168 mg/kg diet is divided by
an assessment factor of 10 to yield 16.8 mg/kg diet.
Terrestrial Vertebrate Chronic COC = 16.8 mg/kg diet.
14
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Table 4: COCs for Environmental Toxicity of MCCPs and LCCPs
Compartment
Test organism
Endpoint
Value
Assessment
factor
coc
Surface water
Water flea
ECso
0.0059 mg/L
5
0.001 mg/L
21-day MATC
0.013 mg/L
10
0.001 mg/L
Sediment
Amphipod
MATC
187 mg/kg dw
10
18.7 mg/kg dry
wt. sediment
Terrestrial
Earthworm
28-day MATC
149 mg/kg dw
10
14.9 mg/kg dry
wt. sediment
Mallard duck
22-week NOEC
168 mg/kg diet
10
16.8 mg/kg diet
4 HUMAN HEALTH HAZARD OVERVIEW
A summary of EPA/OPPT's evaluations on MCCPs and LCCPs is provided in sections 4.1 and
4.2, respectively; individual study reviews are provided in Appendix C.
4.1 MCCP HEALTH DATA REVIEW
There is no information on inhalation absorption of MCCPs in humans or in animals. Based on
their low vapor pressure and low water solubility, absorption following inhalation or dermal
exposure is expected to be limited; previous evaluations concluded that absorption by the
inhalation and dermal routes of exposure will not exceed 50 or one percent, respectively (ECB
2005; EA 2009). Some MCCPs demonstrated moderate absorption and metabolism following
oral exposure in animals. In general, absorption and metabolism are related to their carbon chain
length and degree of chlorination; the longer the carbon chain length and the higher the degree of
chlorination, the less absorption and metabolism.
No information is available on the toxicity of MCCPs in humans; however, the toxicology of
these compounds has been evaluated in experimental animals. Studies in rats and rabbits have
shown that MCCPs only caused slight skin irritation and have low eye irritation potential. No
evidence of skin sensitization was found when tested in guinea pigs. The liver, kidney and
thyroid are the target organs of MCCPs in oral repeated dose studies in experimental animals
(see Table C-l in Appendix C). MCCPs induced increased liver weight, enzyme activity and
histopathological changes at high dose levels. Some of these hepatic effects are likely related to
an increase in metabolic demand as an adaptive response, as well as to peroxisome proliferation,
which are considered of limited toxicological significance to humans. However, liver necrosis
was observed in a 90-day study in rats at 360 mg/kg-bw/day; this effect is considered relevant to
humans. The reported effects in the kidney may have been produced by the parent compound or
from metabolites. Mechanistic data cannot totally rule out that some kidney effects are relevant
to humans. From the data available, a LOAEL of 625 mg/kg-bw/day based on histopathological
changes in the kidneys of female rats is identified in a 90-day toxicity study and a NOAEL of 23
mg/kg-bw/day based on increased kidney weight at 222 mg/kg-bw/day is identified from another
90-day study in rats (CXR, 2005). Repeated dose studies in rats reported some changes in
histopathology and hormone levels of the thyroid. However, it may be concluded based on an
evaluation of the mechanistic data that the thyroid effects observed in rats is of little relevance to
chronic toxicity in humans.
15
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There is no information on the carcinogenicity of MCCPs; however, carcinogenicity studies on a
SCCP and a vLCCP are available. These studies, along with the genotoxicity data on MCCPs,
were used to inform the carcinogenic potential of MCCPs. When administered by gavage, a
SCCP (C12, 60 wt% CI) caused increased incidences of liver tumors in male and female rats,
kidney tumors in male rats and thyroid tumors in female rats. However, based on mechanistic
considerations, these tumors are considered to be of little or no relevance to humans (details in
ECB, 2008 and in Appendix C). An increased incidence of malignant lymphoma in male mice
was reported at the highest dose of 5,000 mg/kg-bw/day in carcinogenicity studies of a vLCCP
(C23, 43 wt% CI) in male and female rats and mice. However, malignant lymphoma is one of the
more variable tumors in mice and has a viral origin in many cases. No increased incidence of
malignant lymphoma was observed in the carcinogenicity study on a SCCP. Further, MCCPs are
non-genotoxic. Therefore, it may be concluded that MCCPs are unlikely to pose a carcinogenic
hazard to humans.
A series of range-finding and definitive prenatal developmental and reproductive toxicity studies
were conducted in rats and rabbits with MCCPs. These studies were conducted between 1981
and 1986. They appear to be valid toxicity studies, conducted according to the standard
methodologies available at the time.
In several prenatal developmental toxicity studies with MCCPs conducted via gavage, no signs
of maternal toxicity were seen at doses as high as 500 mg/kg-bw/day in rats and 100 mg/kg-
bw/day in rabbits. Likewise, no signs of developmental toxicity were observed at doses as high
as 5000 mg/kg-bw/day in rats and 100 mg/kg-bw/day in rabbits.
Two reproductive toxicity studies with MCCPs in rats have been conducted. A one-generation
reproductive toxicity range-finding study showed that administration of approximately 100 and
400 mg/kg-bw/day MCCPs via the diet had no effect on fertility or other reproductive
parameters; however, internal hemorrhaging and deaths in pups were observed beginning from
74 mg/kg-bw/day (1000 ppm) up to approximately 400 mg/kg-bw/day (6250 ppm). These effects
in the pups were not seen in a more recent definitive one-generation reproductive toxicity study
with exposure to MCCPs for 11-12 weeks to doses as high as 100 mg/kg-bw/day (1200 ppm).
Internal hemorrhaging was not seen in the adult animals in either of these studies at doses as high
as 400 mg/kg-bw/day (6250 ppm), or in another study in non-pregnant female rats repeatedly
exposed to doses as high as 1000 mg/kg-bw/day. However, when dams were exposed to
approximately 500 mg/kg-bw/day (6250 ppm) MCCPs during cohabitation, gestation and
lactation, signs of hemorrhaging were observed in dams that died at the time of parturition.
Taken together, the results of these studies suggest that newborns during lactation and pregnant
females at the time of parturition are a potentially sensitive subpopulation with a possible
LOAEL for internal hemorrhaging and deaths in pups at an oral dose of 74/mg/kg-bw/day.
Additional studies with MCCPs have been conducted in an effort to clarify the possible causes of
the hemorrhaging in the pups. One (single-dose; 6250 ppm or 538 mg/kg-bw/day) study showed
maternal death during parturition due to low levels of vitamin K and related hemorrhaging,
suggesting that the act of parturition places dams at higher risk. It was concluded in data from
this study and a cross-fostering study that the fetus relies on clotting factors via mother's milk
16
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and severe deficiencies in vitamin K levels and related clotting factors in the pups results in
hemorrhaging.
No guideline developmental neurotoxicity studies on MCCPs were located. It is not clear if any
developmental neurotoxicity endpoints were measured in the available prenatal
developmental/reproductive toxicity studies; none were explicitly stated. The only information
available regarding behavior during development is from cage-side observations in pups through
lactation day 21. In these cases, no dose-related differences were reported in Fi post-weaning
appearance or cage-side behaviors. While thyroid hormone induced effects were observed in
adults, no data exist for developmental studies. Current studies do not evaluate developmental
neurotoxicity following perinatal exposures.
In this assessment, the lowest NOAEL (90-day value of 23 mg/kg-bw/day from the rat study
described above; CXR, 2005) was used to assess occupational and non-occupational (i.e., general
population) risk of MCCPs.
4.2 LCCP HEALTH DATA REVIEW
There is no information on inhalation absorption of v/LCCPs in humans or in animals. Based on
their low vapor pressure and water solubility, absorption following inhalation or dermal exposure
is expected to be limited. Some absorption and metabolism following oral exposure are possible
for v/LCCPs with shorter carbon chain length and lower degree of chlorination.
No information is available on the toxicity of v/LCCPs in humans. Acute oral toxicity data in
animals show that v/LCCPs are of very low acute toxicity. Studies in animals have shown that
some v/LCCPs may have the potential to cause slight skin irritation and sensitization but no eye
irritation potential. The liver is the main target organ of v/LCCPs in repeated dose studies in
experimental animals. Inflammatory and necrotic changes of the liver were observed in rats
exposed to a C20-30 v/LCCP with 43 wt% CI at dose levels of 100 mg/kg-bw and above. For
another v/LCCP with C20-26 70 wt% CI, effects in the liver occurred at a very high exposure level
of 3,750 mg/kg-bw/day; the NOAEL was 900 mg/kg-bw/day.
An increased incidence of malignant lymphoma in male mice was reported at the highest dose of
5,000 mg/kg-bw/day when tested using a C23 vLCCP with 43 wt% CI in carcinogenicity studies
in male and female rats and mice. However, malignant lymphoma is one of the more variable
tumors in mice and has a viral origin in many cases. Data on the analogous SCCPs have shown
no increase in the incidence of malignant lymphoma in a carcinogenicity study of SCCPs.
v/LCCPs are non-genotoxic and they are not expected to pose a carcinogenic hazard to humans.
Based on the LOAEL (100 mg/kg-bw) of the liver effects in female rats of repeated dose studies
Health Canada calculated a tolerable daily intake (TDI) of 71 [j.g/kg-bw/day with LCCPs. Using
upper bounding intake estimates ranging from 0.007 [j,g/kg-bw/day for 60+ age group to 0.024
[j,g/kg-bw/day for 0.5 years age group, Environment Canada determined that the exposure levels
are 10,000 and 3,000 times lower, respectively, than the TDI.
The National Research Council (NRC, 2000) reviewed the toxicological risks of selected flame
retardants, including a vLCCP containing C24 with 70 wt% CI. Based on the NOAEL of 900
17
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mg/kg-bw/day (liver toxicity), NRC derived an RfD of 0.3 mg/kg-bw/day. Using this RfD and
the worst case average daily exposure to be 0.16 mg/kg-bw/day, NRC concluded: "LCCP do not
pose a noncancer risk when incorporated into residential furniture at the estimated application
levels " Further, it was concluded that: "LCCP are not likely to be a human carcinogen and
derivation of a cancer potency factor is unnecessary
In this assessment, the LOAEL of 100 mg/kg-bw/day from the 90-day and two-year studies
described above) was used to assess potential occupational and non-occupational {i.e., general
population) risk of LCCPs.
5 EXPOSURE INFORMATION
EPA/OPPT used the information in this section and our standard PMN approaches to estimate
potential worker exposures from activities associated with manufacturing, processing and use of
P-12-0453 and P-12-0433. Environmental releases from these activities were also estimated for
use in assessing risk to both human health (general population) and the environment (aquatic
organisms). In addition, EPA/OPPT reviewed the available information on measured
environmental concentrations of MCCPs and LCCPs, which are not normally available for
PMNs.
5.1 ENVIRONMENTAL MONITORING
For this assessment, environmental monitoring data consisting of measured levels of MCCPs and
LCCPs in surface water, sediment and soil were used to characterize potential environmental
exposure to MCCPs and LCCPs. These data are not amenable to determining the ultimate release
source {i.e., manufacturing, processing, or use) into the environment; however, they provide
some insight on the geographical and temporal distribution of MCCPs and LCCPs. Appendix D
contains information and data used in this risk assessment.
Studies published between 1980 and 2013 that reported environmental concentrations of MCCPs
and/or LCCPs were reviewed for this assessment. Monitoring studies from the early 1980s could
not distinguish between the different chain lengths of CPs. The introduction of modern
techniques, such as electron capture negative ion mass spectrometry (ECNI-MS) allowed for the
detection of specific congeners, although difficulties with these methods have persisted {e.g.,
detection of low chlorination congeners in samples). Tomy (2010) performed a round robin
laboratory study of SCCPs that highlighted the inability of the ECNI-MS method to consistently
measure a reference sample, with concentrations varying up to a factor of six. Subsequent work
showed that significant errors (up to a factor of ten) could be introduced by the improper
selection of the calibration standards (Coelhan et al., 2000). A more recent inter-laboratory study
of SCCPs found good agreement amongst the laboratories that used ECNI-MS (Pellizzato et al.,
2009), but similar inter-laboratory studies for MCCPs or LCCPs have not been completed
(Tomy, 2010).
The majority of the monitoring data were collected in Europe and some more recent monitoring
data were collected in China. Over time and across countries, industrial practices and effluent
pre-treatment have varied. Some of the monitoring studies only published their final measured
concentrations and did not include the details of the analytical techniques and sampling
locations. Generally, EPA/OPPT used studies sponsored by the environmental agencies, but full
18
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documentation is lacking for even these studies. The industrial sectors studied by other countries
also are present in the US, suggesting that conditions in the US may be similar.
The level of detail provided in the studies varied. Some studies provided detailed information
regarding sampling locations (e.g., impacted sites), analytical methodology and final sample
results including detection limits, quantitation limits and estimated values. In contrast, other
studies provided only a summary of the results combined from a number of studies. These
summaries also did not provide details of the data analysis to obtain sample results. In addition,
certain studies reported concentrations within a given country but did not provide additional
details about the exact sampling location. Given the disparate conditions (i.e., number of sites
sampled, temporal period over which samples collected, differing analytical methods, etc.)
across the data sets, EPA/OPPT was unable to determine a central tendency or distribution for
the data sets and a range was used instead. Studies using older analytical techniques that did not
distinguish CP congeners were not used in this assessment. Other nations' assessments that used
newer, more reliable, analytical techniques were considered.
EPA/OPPT used the following selection criteria to identify the studies included in this
assessment:
Specific mention of MCCP/LCCP chain length;
Use of modern analytical techniques to distinguish categories of CPs; and
At a minimum, general information on sampling location.
EPA/OPPT used the monitoring data summarized in Tables 5 and 6 for this assessment. When a
limit of detection (LOD) value was reported for non-detectable results5, EPA/OPPT used one-
half of the LOD value.
Even though the existing monitoring data were limited in quality and quantity and it remains
unclear how well the measured data describe the potential range of US MCCP and LCCP use
scenarios, EPA/OPPT concluded that the data in Tables 5 and 6 represented the best available
monitoring information for MCCPs and LCCPs, respectively. These data provide some evidence
that MCCPs and LCCPs are released into the environment; however, these data reflect discrete
locations and times and the extent to which they are representative of the overall distribution of
MCCPs and LCCPs is unknown.
5 Examples would be "not detected" (ND), negligible, or with a "less than" qualifier.
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Table 5: Summary of Measured Concentrations of MCCPs in Environmental Media and
Biota.
Media Category
n
Min
Unit
Max
Unit
References
Surface water
(non-marine)
15
<2.50 x 10"10
mg/L
1.49 x 10"3
mg/L
Coelhan (2010); EC
(2008b); Houde et al.
(2008); IPCS (1996); Muir
et al. (2003); Petersen et al.
(2006)a; USEPA (1988)
Borgen et al. (2003); Chen
et al. (2011); EC (2008a);
Iozza et al. (2008); IPCS
Sediment
(non-marine)
78
2.00 x 10"3
mg/kgb
6.51 x 101
mg/kg
dw
(1996); Nicholls et al.
(2001); Petersen et al.
(2006); Pribylova et al.
(2006); Tomy et al. (1998);
Tomy et al. (1999);
USEPA (1988)
Sediment
(marine)
54
5.00 x 10"3
mg/kg dw
1.64 x 101
mg/kg
dw
Hiittig et al. (2004); Hiittig
and Oehme (2005, 2006);
Kemmlein et al. (2002);
Muir et al. (2000)
Sludge
9
5.00 x 10"5
mg/kgb
9.70 x 103
mg/kg
dw
Stevens et al. (2003);
Pribylova et al. (2006)
Soil
12
2.1 x 10"6
mg/kg dw
8.5 x 10"2
mg/kg
dw
Iozza (2010); Wang et al.
(2013)
Bennie et al. (2000); EC
Biota
(aquatic)
120
<2.00 x 10"7
mg/kg
2.63
mg/kg
WW
(1993, 2008a); Houde etal.
(2008); IVL (2009);
Kemmlein et al. (2002);
Muir (2010); Muir et al.
(2003); Muir et al. (2000);
Reth et al. (2005,2006);
Tomy et al., (1999a);
USEPA (1988)
Biota
(terrestrial)
8
5.00 x 10"3
mg/kg ww
3.70 x 10"1
mg/kg
WW
Reth et al. (2006)
"Petersen et al. (2006) reported results for two water samples; EPA/OPPT assumed these were non-marine
surface water samples.
bThe weight type was not reported (i.e., wet, dry, or lipid weight).
Notes:
1. All values provided in the table above represent total MCCPs and not individual MCCP isomers.
2. The "n" value represents the number of media-specific MCCP monitoring data values that were compiled from
various articles in the raw data table (provided in Appendix D).
3. In some cases, the minimum values in the table are preceded by "<". This indicates that the value reported in
article was reported as a non-detect. In such cases, one half of the lowest reported detection limit was compiled as
the 'minimum' reported monitoring data.
4. dw - dry weight and ww - wet weight.
Table 6 below summarizes the available environmental monitoring data for LCCPs.
Environmental data were available for marine sediment and aquatic invertebrates. Though no
data were available for other media categories (e.g., surface water, non-marine sediment,
terrestrial invertebrates), limited high quality data (from Table 6) were available for MCCPs
which could be used for informing concentrations of LCCPs in the environment. This decision is
20
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based on the following information: 1) P-12-0433 commercial products contain up to 20% of
C17, an MCCP congener (see Section 1.2 Chemistry), 2) LCCP congener groups are expected to
behave in a manner similar to MCCP congener groups with comparable wt% CI when released to
the environment and 3) MCCP and LCCP commercial products have similar uses (see Table 1)
and hence may have similar releases at facilities that process and use these chemicals.
Table 6: Summary of Measured Concentrations of LCCPs in Environmental Media and
Biota.
Media Category
n
Min
Units
Max
Units
References
Sediment (marine)
4
1.02X10"1
mg/kg dw
4.31X10"1
mg/kg dw
Kemmlein et al. (2002)
Biota (aquatic)
2
2.80X10"6
mg/kg lw
6.90xl0"6
mg/kg lw
Kemmlein et al. (2002)
Notes:
1. All values provided in the table above represent total LCCP C18-20 and not individual LCCP isomers.
2. The "n" value represents the number of media-specific LCCP C18-20 monitoring data values that were
compiled from various articles in the raw data table (provided in Appendix D).
3. dw - dry weight and lw - lipid weight.
5.2 MODELED ENVIRONMENTAL RELEASES
EPA/OPPT used screening-level models to generate environmental release estimates for P-12-
0453 and P-12-0433, which were used to calculate exposure concentrations for estimating risks
to humans and aquatic organisms. EPA/OPPT used the Chemical Screening Tool for Exposures
and Environmental Releases (ChemSTEER ver.2) to estimate environmental releases from
industrial processes; the results are provided in Appendix E. Inputs to the ChemSTEER ver. 2
release modeling were based on multiple sources, including information provided by INEOS,
published OECD Emission Scenario Documents, EPA Generic Scenarios and EPA models for
estimating environmental releases. Table 7 provides a general summary of the release
assessment.
Table 7: Summary of Estimated Release to Water
PMN
Chemical Name
Manufacturing
Processing (PROC)
Use
P-12-0453
Alkanes, C14-17, chloro
(MCCPs; 40-60 wt% CI)
No, import only
PROC1: Formulation of
MWFs (73.8% of PV)
USE 1: Use of MWF
(73.8% of PV)
PROC2: PVC compounding
(16.8% of PV)
USE2: PVC converting
(16.8% of PV)
PROC3: Formulation of
adhesives and sealants
(9.4% of PV)a
USE3: Use of
adhesives and sealants
(9.4% of PV)
P-12-0433
Alkanes, C18-20, chloro
(LCCPs; 40-55 wt% CI)
No, import only
PROC1: Formulation of
MWFs (100% of PV)
USE 1: Use of MWFs
(100% of PV)
aNote, adhesives and sealants were assessed as a single process and use when evaluating potential environmental
releases because of the comparable activities associated with each.
Exposure pathways of interest for human health include drinking water, fish ingestion and air
stack emissions. For aquatic organisms, the exposure pathway of concern is from direct releases
to water. EPA/OPPT assessed each of these pathways by using the ChemSTEER ver. 2 release
estimates as inputs to the Exposure and Fate Assessment Screening Tool (E-FAST V2.0) for
21
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estimating industrial releases and concentrations in the foregoing exposure pathways.
EPA/OPPT assumed that potential releases to water occurred from indirect discharges to publicly
owned treatment works (POTW). The E-FAST V2.0 modeling applied an assumption of 90%
removal of MCCPs and LCCPs at the POTW. Water concentrations were estimated using E-
FAST V2.0's probabilistic dilution model (PDM), which predicts downstream chemical
concentrations from industrial discharges. These values were reported as the central tendency for
a median flow site, a low flow site and the lowest seven-day average flow that occurs on average
once every ten years {i.e., 7Q10). These estimated water concentrations were compared to the
COC of 1 |ig/L for chronic aquatic invertebrates (see Table 4). Air stack emissions were
estimated using generic scenarios, which assumed inhalation exposures occurring 100 meters
downwind of a facility. EPA/OPPT used these estimated values for calculating human health and
environmental risks of P-12-0453 and P-12-0433.
The results of the E-FAST V2.0 modeling are provided in Appendix F. Table 8 presents the
values used in the risk assessment. As explained in the footnotes in Table 8, the values represent
reasonable worst-case scenarios based on the processing and use scenarios presented in Table 7.
However, these estimates require consideration of two important caveats: (1) limited
environmental monitoring data are available and (2) MCCPs and LCCPs are expected to
partition to particulates and sediment; however, E-FAST V2.0 models do not account for this
partitioning.
Table 8: E-FAST Modeling Values Used for MCCPs 1
Scenario
Water Release - Human2
Air Release -
Human2
Water Release -
Aquatic Organisms3
Drinking Water
(mg/kg-bw/day)
Fish Ingestion
(mg/kg-bw/day)
Stack Air LADD
(mg/kg-bw/day)
Range of
Concentrations
(lig/L)
P-12-0453
PROC1: Formulation of
MWFs
4.20 x 10"5
5.15 x 10"4
2.99 x 10"4
6.5-47 (240)
PROC2: PVC compounding
7.30 x 10"5
8.95 x 10"4
3.68 x 10"4
7-54 (278)
PROC3: Formulation of
adhesives and sealants
1.30 x 10"4
1.59 x 10"3
6.50 x 10"4
7-150 (258)
USE1: Use of MWF
3.35 x 10"4
4.09 x 10"3
8.95 x 10"4
11-144 (1.30 x 103)
USE2: PVC converting
7.15 x 10"5
8.70 x 10"4
2.18 x 10"4
1-27 (148)
USE3: Use of adhesives and
sealants
2.00 x 10"5
2.45 x 10"4
2.82 x 10"5
2-20 (210)
P-12-0433
PROC1: Formulation of
MWFs
1.04 x 10"6
4.47 x 10"5
N/A
5-36 (184)
USE1: Use of MWFs
2.38 x 10"4
1.02 x 10"3
6.30 x 10"4
7-92 (829)
1 Taken from Appendix E. Values represent the highest concentrations/estimated doses (reported as the lifetime
average daily dose, or LADD) for chronic (i.e., repeated exposure scenarios) for human health.
2 Estimated exposure values were corrected for absorption by the oral (50% absorption) and inhalation (50%
absorption) routes of exposure (ECB 2005; EA 2009).
3 For PMNs, environmental risk was evaluated by performing a PDM as described above and in the E-FAST
Manual (2007). The ranges encompass concentrations from the central tendency for a median flow site (e.g., 6.5
|ig/L) up to the central tendency for a low flow site (e.g., 47 |ig/L) - this example was taken from the first row,
last column, under P-12-0453. The central tendency was calculated using the harmonic mean flow. The value in
parentheses represents the 7Q10 (e.g., 240 |ig/L) value normally used to determine potential chronic risk to
aquatic organisms.
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5.3
EXPOSURE ESTIMATES
5.3.1 OCCUPATIONAL EXPOSURE ESTIMATES
EPA/OPPT calculated screening-level workplace exposure estimates with ChemSTEER ver. 2.
Table 9 provides a summary of the exposure estimates used in this risk assessment for evaluating
worker exposures to P-12-0453 and P-12-0433. Detailed information is provided in Appendix E.
Table 9: Summary of Occupational Exposure Estimates Used.
Route of Exposure
Scenario1
Inhalation
Dermal
(mg/day)
(mg/day)
P-12-0453
PROC1: Formulation ofMWFs
N/A2
350- 1,800
PROC2: PVC compounding
0.030-4.4
1,800
PROC3: Formulation of adhesives and sealants
N/A2
530- 1,800
USE1: Use of MWF
2.0-7.1
350-3,900
USE2: PVC converting
12-22
N/A2
USE3: Use of adhesives and sealants
23
530
P-12-0433
PROC1: Formulation ofMWFs
N/A2
350- 1,800
USE 1: Use ofMWFs
2.0-7.1
350-3,900
'The following represent the estimated number of sites, workers per scenario and exposure over the course of the
stated number of days: P-12-0453 = PROC1: 59 sites, 472 workers and 84 days/year; PROC2: 8 sites, 192
workers and 126 days/year; PROC3: 3 sites, 66 workers and 200 days/year; USE1: 207 sites, 9936 workers and
247 days/year; USE2: 8 sites, 384 workers and 250 days/year; USE3: 58 sites, 2784 workers and 250 days/year;
P-12-0433 = PROC1: 3 sites, 24 workers and 27 days/year; USE1: 4 sites, 192 workers and 247 days/year.
2Not applicable, the use category does not result in exposures that are relevant for this route.
5.3.2 CONSUMER EXPOSURE ESTIMATES
INEOS did not identify consumer uses in its PMN applications for P-12-0453 and P-12-0433;
therefore, EPA/OPPT did not perform an assessment for these types of exposures.
23
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6 RISK ASSESSMENT
6.1 ENVIRONMENTAL ASSESSMENT
PMN risk assessments typically use modeled exposure values because new chemical substances
are not in the stream of US commerce; however, for MCCPs and LCCPs, measured
environmental data are available for some locations in the US and abroad. Though these data are
not specific to P-12-0453 and P-12-0433, the data contain MCCP and LCCP congener groups
that may be present in the PMN substances. However, EPA/OPPT used modeled exposure values
as an important source for its assessment of potential risks because the modeled exposure values
were generated using exposure scenarios that are representative of the types of uses and releases
that may occur with P-12-0453 and P-12-0433. In contrast, the measured environmental data are
generally not amenable for identifying the types of uses or releases from which the measured
congeners originated. Therefore, EPA/OPPT used these measured data as supporting
information, along with modeled exposure values, to calculate potential environmental risks
using the risk quotient (RQ) method.
The RQ method integrates the results of exposure and ecotoxicity data (USEPA, 1998).
An RQ is defined as:
RQ = Environmental Concentration Effect Level
where, the environmental concentration represents measured (see Tables 5 and 6) or estimated
(see Table 8) values for each compartment (i.e., water, sediment and soil) and the effect level
represents the COC for aquatic, benthic, or terrestrial species (see Table 4).
An RQ greater than one serves as a benchmark for identifying whether aquatic concentrations of
P-12-0453 and P-12-0433 may present a risk to aquatic- and sediment-dwelling organisms.
6.1.1 Risk Estimates Using Environmental Monitoring Concentrations
The RQs shown in Table 10 suggest that measured concentrations of MCCPs and LCCPs in
water and sediment may present a risk of acute and chronic injury to aquatic organisms and may
present a risk of chronic injury to sediment-dwelling organisms. However, several limitations
must be noted about the monitoring studies and the level of uncertainty that they contribute to the
basis of these findings. First, the reported concentrations represent minimum and maximum
values that span, at a minimum, several orders of magnitude and translate to RQs of less than one
(i.e., low risk finding) or greater than one (i.e., risk finding), respectively. Second, the temporal
and geographical distributions of these data, along with the different types of uses and releases
that may have served as the originating sources, make it impossible to describe the central
tendency of these data. Finally, the frequency and magnitude of locations with relevant use and
release scenarios to the PMN substances, which may result in environmental releases of MCCPs
or LCCPs that exceed the relevant COCs, is unknown. In addition to these general limitations,
there are specific limitations and uncertainties that preclude using these values as the sole source
24
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from which to inform potential environmental concentrations and risks that may result from the
specific uses and releases associated with P-12-0453 and P-12-0433.
Table 10: Risk Quotients Calculated for MCCPs from Environmental Monitoring Data for
Surface Water, Sediment and the Terrestrial Environment.
Environmental Concentration
Effect Level
(i.e., COC)
RQs
Acute Risk
Aquatic Species
< 2.50xlO"10 to 1.49xl0"3 mg/L
0.001 mg/L
< 2.50xlO"7 to 1.491
Chronic Risk
Aquatic Species
< 2.50xlO"10 to 1.49xl0"3 mg/L
0.001 mg/L
<2.50xlO"7 to 1.49
Chronic Risk
Sediment-dwelling Species
Non-marine Environment
0.002 to 65 mg/kg dw
18.7 mg/kg dw
1.07xl0-4 to 3.5
Chronic Risk
Terrestrial Species
Insufficient Data
14.9 mg/kg dw2
Not calculated
folded values represent those that may present an unreasonable risk of injury.
2 The COCs for terrestrial invertebrates and vertebrates were 14.9 mg/kg dw and 16.8 mg/kg diet, respectively.
Since these values were comparable, EPA/OPPT used the lowest value as a potential means for calculating RQs
for this compartment, once relevant data are available.
For surface water, EPA/OPPT based the aquatic risk findings for MCCPs and LCCPs on the
highest concentration reported by Petersen et al. (2006). These authors collected two surface
water samples from an undisclosed location(s) in Norway and measured the concentration of
MCCP congener groups {i.e., C14-17). The authors reported a concentration of 1.49 x 10"3 mg/L
for MCCP congener groups in one sample; however, a numerical value was not provided for the
second sample, rather the distribution of congener groups was displayed in a bar graph. Based on
the ordinate scale, the concentration of MCCP congener groups in the second sample was greater
than zero, but less than 5.0 x 10"4 mg/L. Of the monitoring studies reviewed by EPA/OPPT (see
Appendix D), the Petersen et al. (2006) value of 1.49 x 10"3 is the only surface water
concentration that resulted in an RQ greater than one. All other surface water concentrations are
at least one order of magnitude below 1.49 x 10"3 mg/L {i.e., RQs < 1).
For sediment concentrations, EPA/OPPT reviewed multiple studies, some of which reported
values that exceeded the COC. Nicholls et al. (2001) reported the most relevant data for P-12-
0453 and P-12-0433. These authors measured concentrations of MCCPs at locations in the
United Kingdom where specific industries were known to employ MCCPs in the use categories
identified for the PMN substances {e.g., lubricant in MWFs, plasticizer in PVC resins, and
lubricant in sealants). Eight locations were sampled at three distances downstream {i.e., 100
meters, 300 meters, and 1-2 kilometers) from the respective sewage treatment works. At four of
the locations, at least one of the sampled downstream values exceeded the COC {i.e., RQs > 1,
risk finding). Though it is not possible to parse out the contribution of specific uses to the
measured values, these data support that releases occur at locations with relevant uses to the
PMN substances, which contribute to the environmental load of MCCP congener groups and in
some cases result in RQs greater than one.
For soil concentrations, EPA/OPPT was unable to calculate RQs for terrestrial organisms due to
the absence of relevant measured data from biosolid-amended soils. Though Iozza (2010) and
Wang et al. (2013) reported measured levels of MCCPs in soil, the samples were collected from
25
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sites in remote alpine locations or industrialized areas, respectively. These data are relevant for
assessing airborne deposition of MCCPs/LCCPs; however, the reported concentrations are of
questionable relevance with informing concentrations of MCCPs/LCCPs that may occur in
biosolid-amended agricultural soils.
Due to the foregoing limitations and resulting uncertainties with the measured environmental
data, EPA/OPPT used these data in a limited capacity for estimating potential risks associated
with the use categories identified for P-12-0453 and P-12-0433. Specifically, these data were
used as supporting information to inform the relevant pathways for estimating potential releases
from relevant use categories for the PMN substances. A summary of the estimated release values
and associated RQs that EPA/OPPT used as a key basis for evaluating the potential risks of P-12-
0453 and P-12-0433 is presented in the following section.
6.1.2 Risk Estimates Using Modeled Exposures
The RQs shown in Table 11 suggest that the intended processes and uses for P-12-0453 and P-
12-0433 are expected to result in releases to surface water at concentrations that may present a
risk of injury to aquatic organisms.
It is noteworthy that these estimated concentrations are within the range of measured surface
water concentrations reported for MCCP congener groups (Table 5). Though there is uncertainty
whether the form {i.e., dissolved or particle bound) of MCCP impacts the aquatic toxicity, the
estimated values suggest that either form may exist. The median stream flow estimates are all
below the reported water solubility for P-12-0453 and slightly above or below the water
solubility reported for P-12-0433 (Table 1). Since the available aquatic toxicity data support that
dissolved MCCP congener groups cause toxicity, the median stream flow values suggest that the
risk finding for this scenario is plausible. The low stream flow and 7Q10 flow scenarios estimate
water concentrations that far exceed the estimated water solubility of P-12-0453 and P-12-0433.
Under these scenarios, the MCCP or LCCP congener groups would likely be bound to
particulates and would eventually settle out in sediment. Nicholls et al. (2001) provided support
for this pathway and showed that sediment concentrations of MCCP congener groups generally
increased with distance downstream from the source outfall. Based on the foregoing information,
EPA/OPPT concludes that the median stream flow values were adequate for determining that
environmental releases of P-12-0453 and P-12-0433 may present a risk of injury to aquatic
organisms.
26
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Table 11: Risk Assessment of Aquatic Organisms Using Modeled Exposures1
Scenario
Estimated Water Concentrations (jig/L)2
RQs
Median Stream
Flow Scenario
Low Stream
Flow Scenario
7Q10
Flow Scenario
P-12-0453
PROC1: Formulation ofMWFs
6.5
47
240
6.5-240
PROC2: PVC compounding
7
54
278
7-278
PROC3: Formulation of
adhesives and sealants
7
150
258
7-258
USE1: Use of MWF
11
144
1300
11-1300
USE2: PVC converting
1
27
148
1-148
USE3: Use of adhesives and
sealants
2
20
210
2-210
P-12-0433
PROC1: Formulation ofMWFs
5
36
184
5-184
USE 1: Use ofMWFs
7
92
829
7-829
'Taken from full model run of summary data presented in Appendix E.
2For PMNs, EPA/OPPT evaluated potential environmental risks by performing a PDM as described above and in
the "Exposure and Fate Assessment Screening Tool (E-FAST) Version 2.0 Documentation Manual (2007)",
available at: httt>://www.era.eov/tsca-screenine-tools/e-fast-exi3osure-and-fate-assessment-screenine-tool-2014-
documentation-manual
6.2 HUMAN HEALTH
EPA/OPPT assessed potential risks to workers and the general population by calculating margins
of exposure (MOE). This approach is performed according to the following equation:
MOE = Point of Departure (POD) ^ Estimated human exposure
For the PODs, EPA/OPPT identified effect levels from three oral repeated dose toxicity studies,
which served as the basis for calculating human equivalent doses (HEDs) (CXR 2005; NTP
1986). In the first study, CXR (2005) reported a NOAEL of 23 mg/kg-bw/day based on
increased kidney weight at 222 mg/kg-bw/day in male rats exposed through diet for 90 days to
an MCCP congener group (C14-17, 52 wt% CI). In the second and third studies, NTP (1986)
reported LOAELs of 100 mg/kg-bw/day based on granulomatous inflammation of the liver in
female rats administered an LCCP congener (C23, 43 wt% CI) by gavage for 5 days/week for 12
months or two years.
Using the effect levels of 23 mg/kg-bw/day or 100 mg/kg-bw/day, EPA/OPPT performed route-
to-route extrapolations to develop HEDs for inhalation and dermal exposures in workers and for
inhalation and oral exposures in the general population. EPA/OPPT did not assess oral exposures
for workers, due to the unlikely nature of exposures occurring by this route. The respective
HEDs served as the PODs for calculating MOEs, along with the previously reported estimated
human exposure values for workers (Table 9) and the general population (Table 8).
EPA/OPPT compared the MOEs to a benchmark value that consisted of a multiplicative
composite of three possible uncertainty factors (UFs): intraspecies variability (UFh; default value
= 10), interspecies variability (UFa; default value = 10), and LOAEL-to-NOAEL extrapolation
uncertainty (UFl; default value =10). The UFh and UFa may each be subdivided to account for
27
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toxicokinetics (TK; default value = 3.16) and toxicodynamics (TD; default value = 3.16). When
effect levels from experimental animal studies are converted to HEDs, EPA/OPPT's default
approach is to reduce the TK subfactor of UFa to 1 {i.e., UFa = TK x TD = 1 x 3.16 ~ 3).
EPA/OPPT interprets MOEs that were equal to or below a benchmark value (e.g., MOE < 1000
[UFh x UFa x UFl = 1000]) as an indication that the scenario(s) may present a risk of injury to
human health, whereas MOEs that were above the benchmark value as a low risk finding. In the
following sections, more detailed descriptions are provided on: 1) converting effect levels to
route- and exposure-specific HEDs; 2) determining the appropriate UFs for the benchmark value,
and 3) evaluating risk estimates for workers and the general population.
6.2.1 Workers
EPA/OPPT performed route-to-route extrapolations to convert the oral NOAEL of 23 mg/kg-
bw/day (i.e., MCCP congener groups) and the oral LOAEL of 100 mg/kg-bw/day (i.e., LCCP
congener) to HED values for inhalation exposures to workers (i.e., HEDinhal-worker) using the
following equation:
HEDinhal-worker = NOAELoral x (1 sRVrat) x (ABSoral-rat ABSinhal- human) x (sRVhuman wRV)
where,
NOAELoral = 23 or 100 mg/kg-bw/day
sRVrat = rat standard respiratory volume for 8-hours = 0.38 m3/kg bw
ABSoral-rat = percent absorption by the oral route in rats = 50%
ABSinhal-human = percent absorption by inhalation in humans = 50%
sRVhuman = human standard respiratory volume for 8-hours = 6.7 m3
wRV = worker respiratory volume for 8-hours = 10 m3
For the oral NOAEL of 23 mg/kg-bw/day and the oral LOAEL of 100 mg/kg-bw/day, the
HEDinhal-worker values equal 41 mg/m3 and 176 mg/m3, respectively.
EPA/OPPT calculated the HED values for dermal exposures to workers (i.e., HEDderm-worker)
based on the following equation:
HEDderm-worker = NOAELoral x (ABSoral-rat ABSdermal-human) x (BWrat BWhuman)1'4
where,
NOAELoral = 23 or 100 mg/kg-bw/day
ABSoral-rat = percent absorption by the oral route in rats = 50%
ABSderm-human = percent absorption by the dermal route in humans = 1%
BWrat = rat bodyweight = 0.250 kg
BWhuman = human bodyweight = 71.8 kg
The resulting HEDderm-worker values equal 4600 mg/kg-bw/day for MCCP congener groups
and 20,000 mg/kg-bw/day for the LCCP congener.
28
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EPA/OPPT used the foregoing HED values to inform the appropriate application of UFs to
derive benchmark values. For MCCP congener groups and the LCCP congener, a benchmark
value of 30 or 300 was applied, respectively. These values consisted of the following individual
UFs. A default UFh of 10 was applied to each benchmark value, due to the absence of
experimental data to inform the TK and TD subfactors of this UF. A reduced UFa of 3 was
applied to each benchmark value, which accounted for a TK subfactor of 1 after converting the
effect levels to HEDs. The UFa of 3 accounted for the remaining uncertainty associated with TD
variability. A default UFl of 10 was only used for the benchmark value compared to the MOE
derived from an LCCP congener because the underlying study reported a LOAEL, not a
NOAEL.
EPA/OPPT used the HEDinhal-worker and HEDderm-worker values for calculating the
respective MOEs using the estimated exposure values presented in Table 9. As shown in Table
13, the MOEs for P-12-0453 and P-12-0433 all exceeded the respective benchmark values,
which indicate a finding of low risk to workers for the processes and uses evaluated in this
assessment.
Table 13: Occupational MOEs for P-12-0453 and P-12-0433
Exposure
Route
Exposure Scenario
Margin of Exposure
P-12-0453
Benchmark MOE = 30
Inhalation
PROC1: Formulation of MWFs
NA
PROC2: PVC compounding
186 -27,333
PROC3: Formulation of adhesives and sealants
NA
USE 1: Use of MWF
115-410
USE2: PVC converting
37
USE3: Use of adhesives and sealants
36
Dermal
PROC1: Formulation of MWFs
18,349 -94,366
PROC2: PVC compounding
18,349
PROC3: Formulation of adhesives and sealants
18,349 -62,317
USE 1: Use of MWF
8,469 - 94,366
USE2: PVC converting
NA
USE3: Use of adhesives and sealants
62,317
P-12-0433
Benchmark MOE = 300
Inhalation
PROC1: Formulation of MWFs
NA
USE1: Use of MWFs
496 - 1,760
Dermal
PROC1: Formulation of MWFs
79,778 -410,286
USE1: Use of MWFs
36,821 -410,286
6.2.2 General Population
EPA/OPPT converted the oral NOAEL of 23 mg/kg-bw/day (i.e., MCCP congener groups) and
the oral LOAEL of 100 mg/kg-bw/day (i.e., LCCP congener) to HED values for oral exposures
to the general population (i.e., HEDoral-genpop) using the following equation:
29
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HEDoral-genpop = NOAELoral x (ABSoral-rat ABSoral- human) x (BWrat BWhuman)1'4 x (5 days 7 days)3
where,
NOAELoral = 23 or 100 mg/kg-bw/day
ABSoral-rat = percent absorption by the oral route in rats = 50%
ABSoral-human = percent absorption by the oral route in humans = 50%
BWrat = rat bodyweight = 0.250 kg
BWhuman = human bodyweight = 71.8 kg
aA duration-specific adjustment was only applied to the oral LOAEL of 100 mg/kg-
bw/day because the animals were gavaged five days per week.
For assessing inhalation exposures to the general population, EPA/OPPT performed route-to-
route extrapolations to convert the oral NOAEL of 23 mg/kg-bw/day {i.e., MCCP congener
groups) and the oral LOAEL of 100 mg/kg-bw/day {i.e., LCCP congener) to HED values for
inhalation exposures to the general population {i.e., HEDinhal-genpop) using the following
equation:
HEDinhal-human = NOAELoral x (1 sRVrat) x (ABSoral-rat ABSmhal- human) x (5 days 7 days)3
where,
NOAELoral = 23 or 100 mg/kg-bw/day
sRVrat = rat standard respiratory volume for 8-hours =1.15 m3/kg bw
ABSoral-rat = percent absorption by the oral route in rats = 50%
ABSinhal-human = percent absorption by inhalation in humans = 50%
aA duration-specific adjustment was only applied to the oral LOAEL of 100 mg/kg-
bw/day because the animals were gavaged five days per week.
For the oral NOAEL of 23 mg/kg-bw/day and the oral LOAEL of 100 mg/kg-bw/day, the
HEDinhal-genpop values equal 20 mg/m3 and 62 mg/m3, respectively.
The same benchmark values of 30 or 300 were used for evaluating the general population MOEs.
These benchmark values consisted of the same individual UFs and rationale discussed previously
for workers.
EPA/OPPT used the HEDoral -genpop and HEDinhal-genpop values for calculating the respective
MOEs using the estimated exposure values presented in Table 8. As shown in Table 14, the
MOEs for P-12-0453 and P-12-0433 all exceeded the respective benchmark values, which
indicate a finding of low risk to the general population for environmental exposures that may
occur due to the processes and uses evaluated in this assessment.
30
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Table 14: General Population MOEs for P-12-0453 and P-12-04331
Scenario
Water Release
Air Release
Drinking Water
MOE
Fish Ingestion
MOE
Stack Air
MOE
P-12-0453 (Benchmark IV
[OE = 30)
PROC1: Formulation of MWFs
2.2 x 106
1.8 x 105
3.1 x io5
PROC2: PVC compounding
1.3 x 106
1.0 x 105
2.5 x 105
PROC3: Formulation of
adhesives and sealants
7.1 x 105
5.8 x 104
1.4 x 105
USE1: Use of MWF
2.8 x 105
2.2 x 104
1.0 x io5
USE2: PVC converting
1.3 x 106
1.1 x io5
4.2 x 105
USE3: Use of adhesives and
sealants
4.6 x 106
3.8 x 105
3.3 x 106
P-12-0433 (Benchmark MOE = 300)
PROC1: Formulation of MWFs
2.8 x 108
6.4 x 106
N/A
USE1: Use of MWFs
1.2 x 106
2.8 x 105
9.8 x 104
'Taken from Appendix E. Values represent the highest concentrations/estimated doses (reported as the lifetime
average daily dose, or LADD) for chronic (i.e., repeated exposure scenarios) for human health.
7 CONCLUSIONS
Based on its assessment of the available surrogate hazard and exposure information on P-12-
0453 and P-12-0433, EPA/OPPT concludes the following pertaining to the manufacturing,
processing and use of these PMN substances:
1. Occupational Exposures: given the assumptions, data and scenarios evaluated in this
assessment, there were low risks found for workers from either dermal or inhalation
exposures.
2. General Population Exposures (from environmental releases): given the assumptions, data
and scenarios evaluated in this assessment, there were low risks found to humans from
environmental releases via exposure to either drinking water or fish ingestion.
3. Environmental Assessment:
a. Using estimated environmental concentrations, the PMN substances may present an
unreasonable risk following acute and chronic exposures to aquatic organisms.
b. Using available measured concentrations of MCCP and LCCP congener groups in the
environment as supporting information, the PMN substances:
iii. Are expected to partition to sediment and may partition to soil through land
application of biosolids and,
iv. May be released to the environment at levels at or above estimated concentrations of
MCCP and LCCP congener groups that may present an unreasonable risk
following acute and chronic exposures to aquatic organisms.
4. PBT Assessment: The PMN substances may be very persistent and very bioaccumulative.
31
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001. http//www.epa.gov/opt/sf/pubs/sf-p2-manual.html
USEPA (U.S. Environmental Protection Agency). 2014. Framework for Human Health Risk
Assessment to Inform Decision Making. EPA/100/R-14/001. Office of the Science
Advisor, Risk Assessment Forum, Washington, DC. http://www.epa.gov/raf.
43
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van Ginkel, C. G. 2010a. Biodegradability of a C14-17 Medium Chain Chlorinated Paraffin
(63.2% CI W/W) in the Closed Bottle Test. AkzoNobel confidential report: T10008c.
AkzoNobel Technology and Engineering, Arnhem, The Netherlands.
van Ginkel, C. G. 2010b. Biodegradability of C14-17 Chlorinated Paraffin (45.6% CI) in the
Closed Bottle Test. AkzoNobel confidential report: T 10007c. AkzoNobel Technology
and Engineering, Arnhem, The Netherlands.
van Ginkel, C. G. 2010c. Biodegradability of Medium-Chain Chlorinated Paraffin (51.7% CI
W/W) in the Closed Bottle Test. AkzoNobel confidential report: T 10031c. AkzoNobel
Technology and Engineering, Arnhem, The Netherlands.
van Ginkel, C. G. 2010d. Biodegradability of Polychlorinated Tetradecane (45%) in the Closed
Bottle Test. AkzoNobel confidential report 2.397.140. AkzoNobel Technology and
Engineering, Arnhem, The Netherlands.
van Ginkel, C. G. 2014a. Biodegradability of C15 ChlorinatedN-Alkane, 51% CI (Wt.) in the
Closed Bottle Test (OECD TG 301). AkzoNobel confidential report: F 14024 CG; Study
number: T13029c. AkzoNobel Technology and Engineering, Arnhem, The Netherlands.
van Ginkel, C. G. 2014b. Biodegradability of CI 5 Chlorinated N-Alkane, 51% CI (Wt.) in the
Closed Bottle Test (OECD TG 301d). AkzoNobel confidential report: F 14025 CG; Study
Number T13030c. AkzoNobel Technology and Engineering, Arnhem, The Netherlands.
van Ginkel, C. G., and A. Louwerse. 2010a. Biodegradability of Chlorinated Tetradecane in
Closed Bottle Tests Incoluated with Activated Sludge and River Water. AkzoNobel
confidential report 2.427.188. AkzoNobel Technology and Engineering, Arnhem, The
Netherlands.
van Ginkel, C. G., and A. Louwerse. 2010b. Evaluation of the Ultimate Biodegradability of
Chlorinated Tetradecanes Using in Sequencing Batch Reactors. AkzoNobel confidential
report: 2.427.189. AkzoNobel Technology and Engineering, Arnhem, The Netherlands.
Wang, Y., J. Li, Z. Cheng, Q. Li, X. Pan, R. Zhang, D. Liu, C. Luo, X. Liu, A. Katsoyiannis, and
G. Zhang. 2013. Short- and Medium-Chain Chlorinated Paraffins in Air and Soil of
Subtropical Terrestrial Environment in the Pearl River Delta, South China: Distribution,
Composition, Atmospheric Deposition Fluxes, and Environmental Fate. Environmental
Science and Technology, 47(6), 2679-2687.
Wiegand, W. 1989. Determination of the Mutagenic Effects of Chloroparaffin 40g. AM-89/08.
Huls AG, Huls AG, Marl, Germany.
Willis, B., M. J. Crookes, J. Diment, and S. D. Dobson. 1994. Environmental Hazard
Assessment: Chlorinated Paraffins. TD170.9, G7ts, no. 19. Department of the
44
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Environment, Toxic Substances Division of the Directorate for Air, Climate and Toxic
Substances, Garston, Watson, UK.
Zeng, L. X., H. J. Li, T. Wang, Y. Gao, K. Xiao, Y. G. Du, Y. W. Wang, and G. B. Jiang. 2013.
Behavior, Fate, and Mass Loading of Short Chain Chlorinated Paraffins in an Advanced
Municipal Sewage Treatment Plant. Environmental Science & Technology, 47(2), 732-
740.
45
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9 APPENDICES
46
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Appendix A ENVIRONMENTAL FATE AND
BIOACCUMULATION STUDY SUMMARIES
A-l ENVIRONMENTAL PERSISTENCE
A-l-1 Abiotic Degradation
Generally, CPs are stable to hydrolysis and to direct photolysis in air and water, though very
limited data exist on hydrolysis and direct and indirect photolysis in soil, water, or air. In studies
using aliphatic hydrocarbon solvents, CPs were shown to be poor absorbers of UV light and no
direct photodegradation was observed (Friedman and Lombardo, 1975; Lombardo et al., 1975).
Koh and Thiemann (2001) studied photolysis of aqueous solutions for CPs products with chain
lengths ranging from Cio to C24 including an MCCP product, Hoechst CP52, with chain lengths
from C12 to Ci8 and an average of 52 wt% CI. A mercury vapor lamp with main radiation
wavelengths of 254, 302, 313, 366,405/408, and 436 was used in batch experiments. Following a
5 hour radiation time, estimated atmospheric degradation rates showed photolysis half-lives of
less than 20 hours based on measurement of free chloride and analysis of degradation products.
The MCCP product had a T1/2 of 12.8 hour in aqueous solution. The addition of peroxide or
acetone increased the photolysis rate suggesting that indirect photolysis may be significant. The
authors also reported that longer chain CPs were formed during this study and speculated that
recombination of smaller alkyl radicals could occur under some conditions.
Thermal degradation data for MCCPs and LCCPs are limited, but studies of SCCPs and
Polyvinyl chlorides suggest MCCPs are degraded rapidly at 250 - 350 ฐC (Bergman et al., 1984).
Dehydrohalogenation may lead to the formation of a large number of aliphatic and aromatic
compounds. Chlorine radical formation can lead to production of highly chlorinated aromatics
including polychlorinated biphenyls. Higher CI content results in production of greater numbers
and amounts of chlorinated aromatics (Bergman et al., 1984).
A-l-1-1 Fate in Air
As noted above, CPs lack structural components that absorb light in the UV or visible spectrum,
so direct photolysis is not expected to occur. The atmospheric half-life has been estimated at 1 -
2 days (EA, 2009; ECB, 2005), based on estimated values for the second order rate constant for
reaction with atmospheric hydroxyl radicals for MCCPs with lower chlorine contents between 40
and 56 wt%. EPA/OPPT also estimated atmospheric half-lives for MCCPs (40 and 70 wt% CI)
calculated using EPI Suite/AOPWIN (v. 1.92a) that range from about 1 to > 4 days (see
Table Apx A-l). MCCPs with the shorter chain lengths and higher chlorine contents were
calculated to be more persistent.
MCCPs have low estimated vapor pressures (4.5 x 10"8 to 2.27 x 10"3 Pa at 20 - 25ฐC) and a
Henry's law constant (HLC) (0.014 - 51.3 Pa x m3/mol for C14-17 congener groups) and are not
expected to partition to air. They may be transported associated with particulate matter, and have
been reported in indoor and outdoor air and house dust (Barber et al., 2005; Friden et al., 2011;
Hilger et al., 2013). Wide spread soil contamination and occurrence in artic samples suggest that
MCCPs behave similarly to other chlorinated persistent organic pollutants (POPs) with high
47
-------
production volumes and releases, and are subject to long range transport (Dick et al., 2010;
Medeiros et al., 2011; Tomy et al., 2000).
TableApx A-l: Estimated Atmospheric Half
Lives Using EPI Suite/AOPWIN (v. 1.92a)
for Varying MCCP Chain Length and
Chlorination Percents Based on Wt.
Chain Length
40 wt% CI
70 wt% CI
Cl4
1.0
4.4
Cl5
0.8
3.0
Cl6
0.8
3.0
Cl7
0.8
2.9
A-1-2 Biodegradation
EPA/OPPT reviewed studies from the open literature and submitted to EPA/OPPT including
those described in the Canada and EU assessments and referenced in Table Apx A-2 (EC,
2008a; ECB, 2005) to determine biodegradation under a variety of environmental conditions.
Some of these studies used modified test conditions to enhance or maximize biodegradation.
EPA/OPPT concurs with the EU's conclusions that under these modified test conditions, Ci4
41.3% by wt. CI and a Ci4 45.5% by wt. CI substances are readily biodegradable. C 15 51% by
wt. CI were found to be inherently degradable and possibly readily degradable in modified
OECD 301 and 301D tests. This suggests that CPs with these chain lengths and shorter, and this
degree of chlorination and lower, are inherently degradable. More highly chlorinated and longer
carbon chain CPs (C14-17 51.7% by wt. CI, C14 55% by wt. CI, C14 60.2% by wt. CI, and C14 -17
63.2% by wt. CI) biodegraded over a range of 2-54% in 28 days to 4-57% at up to 60 days. The
most highly chlorinated, (C14-17, 63.2 wt% CI) biodegraded 5% in 28 days and 10% at 60 days
in the enhanced biodegradation studies, suggesting that longer chain and higher chlorination can
contribute to greater persistence under most environmental conditions. (Van Ginkel, 2014 a and
b; Van Ginkel 2010 a-d; Van Ginkel and Louwerse 2010 a and b).
A-l-2-1 Fate in Wastewater Treatment
In its review of the available measured data on MCCPs in wastewater treatment from data in
from other countries, EPA/OPPT determined that CPs are present in the majority of municipal
waste water treatment plant (WWTP) influent (Coelhan, 2010; Nicholls et al., 2001; Stevens et
al., 2003; Zeng et al., 2013). Low water solubility and relatively high partitioning coefficients
suggest that most of the MCCPs and LCCPs entering WWTP systems will associate with solids.
Some biodegradation of shorter chain, lower chlorinated MCCP congener groups may occur,
while longer chain length, more chlorinated congener groups will be resistant to aerobic and
anaerobic degradation. Shorter and lower chlorinated congener groups have higher vapor
pressure and may be lost to the vapor phase during aeration. WWTP effluent also contains some
particulate-associated MCCPs. Because of their low water solubility, little MCCP or LCCP will
be in the dissolved phase, and the majority will be removed along with settled sludge. Once
associated with the sludge, the CPs will generally be stable in sludge treatment and remain in the
residual biosolids. Land application of biosolids will transfer the MCCPs and LCCPs to
agricultural and other soils (Nicholls et al., 2001; Stevens et al., 2003). Because 50 - 60% of
48
-------
biosolids in the US are land applied, the majority of MCCPs and LCCPs entering WWTPs may
be released to the environment via application to soil, and may be transported from contaminated
soil to other locations and media by soil erosion, runoff, and wind borne particulates, and
volatilization.
A-l-2-2 Fate in Surface and Groundwater
Because they generally have low water solubility, high sorption coefficients, and tend to partition
to solids, MCCPs and LCCPs released to surface water will partition to surficial sediment where
they may be buried and removed from potential degradation processes. This explains what is
found in the limited monitoring data that exist - MCCP concentrations in surface water are
generally in the low pg/L range, while sediment concentrations are several orders of magnitude
higher (EC, 2008a).
MCCPs may leach from soil and be transported to groundwater, but low solubility and high
sorption will act to keep dissolved concentrations very low. Facilitated transport with colloids
and particulates may occur so that MCCPs can be transported in groundwater, but in general,
concentrations in this compartment are expected to be very low. MCCPs that are introduced to
groundwater will tend to partition to the solid phase and not be mobile.
A-l-2-3 Fate in Soil
Existing monitoring data suggest that MCCPs are present in soil, probably as a result of
atmospheric transport and deposition. Areas near sources, such as land receiving wastewater
biosolids, manufacturing and processing facilities, and electronic waste processing and recycling
facilities are shown to have higher levels (Wang et al., 2013). MCCPs are expected to be stable
in soil, and once deposited, could remain/persist in the soil for years or decades. Burial and
advective transport away from the site of deposition are the major dissipation processes. No data
are available on soil photolysis, although aqueous photolysis data suggest that indirect photolysis
may result in degradation to shorter and less chlorinated CP congener groups. No soil
biodegradation data exists, but some strains of bacteria that can co-metabolize MCCPs have been
identified (Allpress and Gowland, 1999). If degradation does occur, it is expected to be slow
with T1/2 of at least months to years.
Table Apx A-2: Review of MCCP and LCCP Biodegradation Studies.
Biodegradation Studies on MCCPs (C14-17) and LCCPs (C>ix)
Study
Authors
Publication
Date
MCCP/LCCP
Chemicals
Evaluated (Le., C-
length, wt% CI)
Method
Study
Duration
Noteworthy Results and Implications
MCCPs
Van Ginkel
2010d
C14 45 wt% CI
Closed bottle
28 days
Approximately 64% degraded in 28
days
Based on oxygen demand
Van Ginkel
2010b
C14-17 45.6 wt% CI
Closed bottle
28 days
Approximately 51% in 28 days and
63% in 42 days degradation
Based on oxygen demand
Van Ginkel
2010c
C14-17 51.7 wt%Cl
Closed bottle
28 days
Approximately 27% degradation in 28
days and 57% after 60 days
49
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Biodegradation Studies on MCCPs (C14-17) and LCCPs (C>ix)
Study
Authors
Publication
Date
MCCP/LCCP
Chemicals
Evaluated (Le., C-
length, wt% CI)
Method
Study
Duration
Noteworthy Results and Implications
Based on oxygen demand
Van Ginkel
2010a
C14-17 63.2 wt% CI
Closed bottle
28 days
Approximately 5% degradation after 28
days and 10% after 60 days.
Van Ginkel
and
Louwerse
2010a
C14 41.3-60.2 wt%
Cl%
Closed bottle
with river
water and
sludge
inoculum
28 days
Approximately 66% (41.3 wt% CI) to
11 (60.2 wt% CI) degradation in 28
days respectively
Van Ginkel
and
Louwerse
2010b
C14 41.3-50 wt% CI
Batch reactor
21 and
105 days
41.3 wt% CI: 79% degradation in 21
days and 94% at 105 days
50 wt% CI: 14% degradation by 21
days 5 wt% CI in 80 days based on
quantitation of released chloride
Conclusions: Quantification of degradation was by oxygen uptake or chloride release. No information on the
chemical distribution in the test material or degradates was provided.
These studies used modified test conditions to enhance or maximize biodegradation. Under these modified test
conditions, Cm 41.3 wt% CI and a C14 45.5 wt% CI substances are readily biodegradable. More highly chlorinated
and longer carbon chain CPs (C14-17 51.7 wt% CI, C14 55 wt% CI, C14 60.2 wt% CI, and C14-17 63.2 wt% CI)
biodegraded over a range of 2 - 54% in 28 days to 4 - 57% at up to 60 days. The most highly
chlorinated, (C14-17 63.2 wt% CI) biodegraded 5% in 28 days and 10% at 60 days in the enhanced
biodegradation studies, suggesting that longer chain and higher chlorination can contribute to greater
persistence undermost environmental conditions.
Van Ginkel
2014b
C15 51 wt% CI
Closed bottle
(30 ID)
60 day
43% and 63% degradation at 28 and 60
days
Van Ginkel
2014a
C15 51wt% CI
Closed bottle
(301)
60 day
37% and 57% degradation at 28 and 60
days
Conclusions: Unlike the 2010 series of studies, these most recent biodegradation studies did not have significant
protocol modifications and the C15 51 wt% CI were found to be inherently degradable and possibly readily
degradable in OECD 301 and 30ID tests.
Madeley and
Birtley
1980
C14-17 mixed
product, 40 wt% CI
BOD test
25 days
Approximately 15.5% degradation as
measured by theoretical BOD in non-
acclimated samples and 22.5%
degradation in acclimated samples.1
Madeley and
Birtley
1980
C14-17 mixed
product, 45 wt% CI
BOD test
25 days
Approximately 10% degradation as
measured by theoretical BOD in non-
acclimated samples and 30%
degradation with acclimated soil
microbes added.1
Madeley and
Birtley
1980
C14-17 mixed
product, 52 wt% CI
BOD test
25 days
Approximately 4% degradation as
measured by theoretical BOD in non-
acclimated samples and 6% degradation
with acclimated soil microbes added.1
50
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Biodegradation Studies on MCCPs (C14-17) and LCCPs (C>ix)
Study
Authors
Publication
Date
MCCP/LCCP
Chemicals
Evaluated (Le., C-
length, wt% CI)
Method
Study
Duration
Noteworthy Results and Implications
Madeley and
Birtley
1980
C14-17 mixed
product, 58 wt% CI
BOD test
25 days
No significant degradation
Conclusions: The data from Madeley and Birtley suggests the potential for biodegradation but has significant
limitations. The BOD studies were done on mixed products. No attempt was made to determine which specific
congeners were degraded or the reaction products. No identification of the congeners present was provided. The
degradation was estimated from the BOD but other compounds may have contributed to the ThBOD in the bottles.
BOD measurements are highly variable as evidenced by the decrease in the C20-30 42% of > 50% between day 20
and 25.
'The ThOD (theoretical oxygen demand) was estimated (ThOD (g 02/g substance) = 16[2><(h-cl)]/mw;
where c=number of carbon atoms, h=number of hydrogen atoms, cl=number of chlorine atoms and MW =
molecular weight). This is questionable for a product containing mixture of congeners as was used in all studies.
I.CCPs
Madeley and
Birtley
1980
C20-30 mixed
product, 42 wt% CI
BOD test
25 days
Approximately 7.5% degradation as
measured by theoretical BOD in non-
acclimated samples and 23%
degradation with acclimated soil
microbes added.1
Madeley and
Birtley
1980
C25 "chlorinated
pentacosane"
14C on central
carbon
8 weeks
(mean)
11% of 14C- released as CO2 Non-
acclimated microbes.
Conclusions: The data from Madeley and Birtley suggests the potential for biodegradation but has significant
limitations. The BOD studies were done on mixed products. No attempt was made to determine which specific
congeners were degraded or the reaction products. No identification of the congeners present was provided. The
degradation was estimated from the BOD but other compounds may have contributed to the ThBOD in the bottles.
BOD measurements are highly variable as evidenced by the decrease in the C20-30 42% of > 50% between day 20
and 25.
1 The ThOD (theoretical oxygen demand) was estimated (ThOD (g 02/g substance) = 16[2> C20-30 mixed
product, 70 wt% CI
BOD test
5 days
2% degradation. Degradation was
estimated by the authors as the% of the
theoretical BOD based on the total
carbon content of the test solution.
Substances other than the chlorinated
paraffin contributed to this total carbon
content.
Hildebrecht
1972
> C20-30 mixed
product, 70 wt% CI
BOD test
5 days
65% degradation. Degradation was
estimated by the authors as the% of the
theoretical BOD based on the total
carbon content of the test solution.
51
-------
Biodegradation Studies on MCCPs (C14-17) and LCCPs (C>ix)
Study
Authors
Publication
Date
MCCP/LCCP
Chemicals
Evaluated (Le., C-
length, wt% CI)
Method
Study
Duration
Noteworthy Results and Implications
Substances other than the chlorinated
paraffin contributed to this total carbon
content.
Conclusions: As described by the (EA, 2009), Hildebrecht's results are questionable (Hildebrecht, 1972). This
report is not available so it cannot be reviewed directly, but others have reported that it provided limited details. A
surfactant, other carbon sources, and nutrients were added that may have contributed BOD. The extent of
degradation was determined by the comparing the oxygen consumption in the test with the theoretical oxygen
demand (ThOD) based on oxidation to CO2 of the total organic carbon present in the solution from all sources. This
estimation of ThOD does not take into account oxygen consumption by other compounds or the unknown
composition. Of the CPs in the mixture, as the UK report concludes, "It is not possible to draw definite conclusions
as to the degradability of the chlorinated paraffins in these tests" (EA, 2009).
Hoechst AG
1976 and
1977
C18-20, 35 wt% CI
BOD test
5 days
0.7% degradation
Hoechst AG
1976 and
1977
C18-20, 44 wt% CI
BOD test
5 days
< 1.2% degradation
Hoechst AG
1976 and
1977
C18-20, 49 wt% CI
BOD test
5 days
<2.3% degradation
Hoechst AG
1976 and
1977
C18-20, 52 wt% CI
BOD test
5 days
< 0.6% degradation
Conclusions: The Hoechst reports from early industry studies are not available so it is not possible to directly
review the data (Hoechst, 1976, 1977). Others (EA, 2009) have reported the limitations of the studies. Limited
details of the studies were apparently reported by Hoechst. These tests were done on mixtures of congeners with
unknown composition. They reported that the majority of the CPs were removed by sorption on to the solids so no
degradation may have occurred that would have been detected as BOD. The tests were run for 5 days using non-
acclimated sludge microbes so degradation may have been possible but had not yet occurred.
Omori et al.
1987
C24.5H44.5CI6.5, 40.5
wt% CI
Chloride
release
48 hours
9.9% degradation using bacterial strain
HK-3;
13%H15-4;
2.2% HK-6;
3.5% HK-8;
33% using mixed bacterial culture (HK-
3, HK-6, HK-8 and HK-10)
Omori et al.
1987
C24.5H41CI10,50 wt%
CI
Chloride
release
48 hours
3% degradation using bacterial strain
HK-3;
9% H15-4;
1.8% HK-6;
2.6% HK-8
Omori et al.
1987
C24.5H3oCl2i,70wt%
CI
Chloride
release
48 hours
2.6% degradation using bacterial strain
HK-3;
12%H15-4;
1.4% HK-6;
1.7% HK-8;
15% using mixed bacterial culture (HK-
3, HK-6, HK-8 and HK-10)
Conclusions: Omori et al. (1987) showed the potential for biodegradation using pure and mixed cultures in short
(48 hour) incubations. No information on the starting mixtures were provided except average compositions. No data
52
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Biodegradation Studies on MCCPs (C14-17) and LCCPs (C>ix)
Study
Authors
Publication
Date
MCCP/LCCP
Chemicals
Evaluated (Le., C-
length, wt% CI)
Method
Study
Duration
Noteworthy Results and Implications
on the products were reported. Loss of CI suggests dechlorination can occur and that lower CI content or shorter
chain lengths may be produced.
Allpress and
Gowland
1999
C18-20, 48 wt% CI
Chloride
release
71 days
11% degradation using Rhodococcus sp.
bacteria
Allpress and
Gowland
1999
C> 20, 42 wt% CI
Chloride
release
71 days
14% degradation using Rhodococcus sp.
bacteria
Conclusions: Allpress and Gowland (1999) also showed that CPs have the potential to biodegrade using pure
culture. They used mixed congener products and did not provide any information on the composition of the starting
material or the degradation products. They found that the Rhodococcus sp. was able to use CPs as carbon source as
well as an energy source.
1 The ThOD (theoretical oxygen demand) was estimated (ThOD (g Ch/g substance) = 16[2>
-------
specifically, large BMFs were observed for all MCCP chain lengths in Lake Ontario, and for Cu
MCCPs in Lake Michigan, indicating biomagnification. BMFs (2.4 - 7.7) were also above 1 for
smelt and lake trout in Lake Michigan.
In laboratory studies with rainbow trout and oligochaetes, lipid-normalized equilibrium BMFs
estimated from a first-order bioaccumulation model for constant dietary exposure ranged from
0.4 - 5.0 (Fisk et al., 1996; Fisk et al., 2000; Fisk et al., 1998).
Most of the laboratory-based BCF studies (Bengtsson et al., 1979; CPC, 1980, 1983a, 1983b;
Fisk et al., 1999; Fisk et al., 1998; Houde et al., 2008; Madeley and Maddock, 1983a, 1983b;
Madeley and Thompson, 1983; Renberg et al., 1986; Thompson et al., 2000), were reported to
have been conducted at MCCPs concentrations above the water solubility limit and hence likely
underestimate the true BCF. Furthermore, acetone as a solvent in these tests, so they do not
adhere to OECD guidelines. Nonetheless, some BCF values estimated from these studies indicate
MCCPs are bioaccumulative (e.g., bleak and rainbow trout (32-2856) and BCF of 6920 for
common mussel).
54
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Table Apx A-3: Review of MCCP and LCCP Bioaccumulation Studies
Bioaccumulation studies on MCCPs (Ci417) and LCCPs (> Cis)
MCCPs
Study Authors
Publication
Date
MCCP/LCCP
Chemicals
Evaluated (i.e., C-
length,% CI)
Method
Study Duration
Noteworthy Results and Implications
Houde el al.
2008
Ci 4-15 Only
BAF =
([predator]/[water
(filtered)]);
BMF) =
[predator]/[prey])
where the
concentrations in
predator and prey are
on a lipid basis
Three sampling
periods: October
2000, June 2002,
and
July 2004.
Issues related to the temporal variability of water
concentrations over the period of biota sampling (1999 -
2004) in this study have been raised (ECB, 2005; EC, 2008)
contributing to uncertainties associated with the reported
BAF values.
Log BAF =
Plankton: C14=6.2; C15=6.6; ฃ=6.5
Alewife: C14=7.0; C15=6.8; ฃ=6.9
Sculpin: C14=7.4; C15=7.2; ฃ=7.3
Rainbow smelt: C14=7.4; C15=7.1; X=7.2
Lake trout: C14=6.8; C15=6.5; ฃ=6.6
BMF (Lake Ontario) =
0.25 (lake trout - alewife);
0.14 (lake trout - smelt);
8.7 (sculpin-Diporeia)
BMF (Lake Michigan) =
0.22 (lake trout - alewife);
0.94 (lake trout - sculpin);
0.88 (sculpin - Diporeia)
Thompson et al.;
as summarized in
ECB, 2005
2000
n-pentadecane-8-14C,
51% CI mixed with a
non-radio-labelled
C14-17, 51% CI
chlorinated paraffin
Freshwater; flow-
through; acetone
solvent used;
35 days
Steady-state may not have been achieved, so kinetic BCF
data considered more reliable.
BCF = 860 L/kg at 35 days when exposed at 0.9 ug/L
BCF = 265 L/kg at 35 days when exposed at 4.9 ug/L
kinetic BCF = 1,087 L/kg at 35 days when exposed at 0.9
ug/L
kinetic BCF = 349 L/kg at 35 days when exposed at 4.9
ug/L
55
-------
CPC (Madeley et
al); as
summarized in
ECB, 2005
1983
commercial product
mixed with a n-
pentadecane-8-14C
chlorinated to a
similar degree
freshwater; flow-
through; rainbow trout;
acetone solvent used
60 days
Concentrations above water solubility; hence, water
concentrations may be overestimated and BCF
underestimated.
Uncertainty as to whether steady-state was reached.
BCF = 32-45 1/kg on a wet weight basis when exposed at
1.05 mg/1;
BCF = 42-67 1/kg on a wet weight basis when exposed at
the 4.5 mg/1
CPC (Madeley
and Pearson); as
summarized in
ECB, 2005
1980
C14-17, 45% CI
freshwater; flow-
through; rainbow trout;
28 days
Measured water concentrations questionable; water
concentrations may be overestimated and BCF
underestimated.
BCF = 50-60 1/kg based on nominal exposure
concentrations
BCF = 280-600 1/kg based on measured water
concentrations
Madeley and
Maddock; as
summarized in
ECB, 2005
1983
Total MCCPs
Bioconcentration
factors. MCCPs
concentrations were
above the water
solubility limit, using
acetone as the co-
solvent in the test
solutions, and hence are
not in compliance with
OECD guideline
requirements
No Information
BCF = 32 - 2856 for common mussel, bleak and rainbow
trout. May not have reached steady-state.
Fisk et al.
1999
Average formula:
C14H23.3C16.7,
55% CI
freshwater; medaka
eggs
20-days
Uncertainty as to whether steady-state was reached; hence
BCFs probably represent lower limit of true value
BCF = 32- 680 L/kg
Fisk et al.
1998
14Ci6 35% CI and
14Ci6 69% CI
Lake sediments were
spiked and worms
added after 18 and 32
No Information
Kinetic BAF probably represents the
upper limit of the true bioaccumulation factor
14Ci6 35% CI
14-day BSAFss = 0.7
Kinetic BSAF = 4.4
14Cis 69% CI
56
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14-day BSAFss = 0.2
Kinetic BSAF = 0.6
Bengtsson et al.
1979
C14-17, 50% CI
seawater; semi-static;
bleak;
acetone solvent used
14 days
Measured water concentrations questionable; water
concentrations may be overestimated and BCF
underestimated.
BCF ~ 40 L/kg
Madeley and
Thompson; as
summarized in
ECB, 2005
1983
commercial C14-17,
52% CI
seawater; flow-through;
mussel
acetone solvent used
60 days
BCFs = 2,182 L/kg (parent compound analysis) or 2,856
L/kg (14C-measurements) when exposed to 0.22 mg/1
BCF = 339 L/kg (parent compound analysis) or 429 l/kg
(14C measurements)
when exposed to 3.8 mg/1.
Renberg et al.
1985
C16H30.7CI3.3 (34%
CI) and C12H16CI9.8
(68.5% CI) mixture
synthesized with 14C
radiolabel
Flow through exposure
to mussel (Mytilus
edulis)
Ci6 - 0.13 and 5.0 ng/L
Ci6 - 28 day uptake
C12 - 21 day uptake
followed by 28 day
depuration
Steady state BCF about 7000 for C16 and 140,000 for C12
based on 14C quantification.
No chemical specific analysis for CPs. Metabolism and
accumulation of degradation products may have accounted
for high values.
I.CCPs
Bengtsson etal.;
as summarized in
ECB, 2005
1979
C18-26
concentrations were
above the water
solubility limit and
hence are not in
compliance with OECD
guideline requirements
No Information
Concentrations above water solubility; hence, water
concentrations may be overestimated and BCF
underestimated. Uncertainty as to whether steady-state was
reached.
BCF reported =8-16 L/kg
57
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Appendix B ECOTOXICITY STUDY SUMMARIES
B-l MCCP ECOTOXICITY DATA
B-l-1 Acute Fish Toxicity
(1) A series of 96-hour acute fish toxicity studies were conducted by Mayer an Ellersieck (1986)
with Paroil 1048 (50-52% CI, C15H26C16) similarly to ASTM (1980). Bluegill sunfish (Lepomis
macrochirus) and yellow perch (Perca flavescens) were exposed to the test substance in a flow-
through test system and channel catfish (Ictaluruspunctatus) and rainbow trout (Oncorhynchus
mykiss) were exposed to the test substance in a static test system. Solvent use was not specified
for this compound. The average pH level was between 7.4 and 7.5 for all tests. Test temperature
was 12 ฐC for bluegill sunfish, rainbow trout, and yellow perch and 20 ฐC for channel catfish.
Dilution water hardness was 44 mg CaCCb/L in the rainbow trout and channel catfish test system
and 314 mg CaCCb/L for the bluegill sunfish and yellow perch test system. Reported effect
levels are considered to be nominal with LC50 values of >10 mg/L for bluegill sunfish, channel
catfish, and yellow perch and >0.011 mg/L for rainbow trout; all values are greatly above the
limit of solubility.
EPA/OPPT Conclusion
Using a weight-of-evidence approach, these studies were considered
acceptable to characterize the acute fish toxicity endpoint.
96-hr LCso = NES
(2) A 96-hour acute fish toxicity study was published by Linden et al. (1979). Groups of 10
Bleak (Alburnus alburnus) were exposed to six nominal unspecified concentrations of Cereclor
S52ฎ (C14-17, 52% CI), Chloroparaffin huls 40G (C15.5, 40% CI), and Witaclor 50 (C14-17, 50%
CI) under static test conditions. Salinity was 7 ppt, pH was 7.8, temperature was 10 ฐC, and
dissolve oxygen was considered by study authors to be satisfactory. EPA/OPPT requires
reporting of dissolved oxygen concentrations to determine study adequacy. EPA/OPPT also does
not consider the test species, the bleak, a standard test species. The 96-hour fish LC50 values
were >10,000 mg/L, >5,000, and >5,000 for Cereclor S52ฎ, Chloroparaffin huls 40G, and
Witaclor 50.
EPA/OPPT Conclusion
Given effect levels observed in Mayer and Ellersieck (1986) and the
reported water solubility of medium chain paraffins, these studies were
considered acceptable using a weight-of-evidence approach to
characterize the acute saltwater fish toxicity endpoint.
96-hr LCso = NES
(3) Bengtsson et al. (1979) also studied the toxicity of a medium-chain chlorinated paraffin to
Bleak (Alburnus alburnus) as part of a bioaccumulation study. The chlorinated paraffin tested
was a C14-17, 50% wt. CI substance. The tests were performed at 10 ฐC using a semi-static
procedure in which the test solutions containing 125 [j,g/L of the substance were renewed every
two to three days over the 14-day exposure period. The water used in the experiment was Baltic
Sea water with a salinity of 7%o, and acetone was present in all aquaria, including controls at a
concentration of 0.1 ml/1. The fish used in the experiment had an average weight of 4.5 g and
58
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were not fed during the exposure period. Six groups of 15 fish were used for both the exposure
and control solutions. No mortality or effect on behavior was seen in fish exposed to the
medium-chain chlorinated paraffin during the test.
EPA/OPPT Conclusion
This data review was part of a BAF study and as such will be used as a
weight of evidence to support other data for this category of organisms.
B-l-2 Acute Aquatic Invertebrate Toxicity
(1) A 48-hour acute Daphnia magna toxicity study was conducted by Thompson et al. (1996)
according to OECD TG 202 (1984) with GLP compliance using a static test system. The test
substance was identified as Cereclor S52ฎ, a C14-17 chlorinated paraffin with 52% chlorination
that contained 0.3% epoxy soya bean oil stabilizer as well as a small amount of radiolabeled n-
pentadecane-8-14C (51% chlorinated). Four replicates of 5 Daphnia magna Straus (<24 hours
old) were exposed to nominal concentrations of 0 (dilution water control), 0 (solvent control),
0.0032, 0.0056, 0.01, 0.018, 0.032, 0.056, and 0.1 mg/Ltest substance in acetone (0.1 mL/L).
Test solutions were prepared by adding the appropriate stock solution to dilution water while
continuously and vigorously stirring with a magnetic follower. Appearance of test solutions was
not provided. Corresponding mean measured concentrations determined by radiochemical
methods were 0.0025, 0.0041, 0.0094, 0.015, 0.024, 0.047, and 0.095 mg/L. Daphnid loading
was 25 daphnids/L. Over the course of the study dissolved oxygen concentrations remained
between 9 and 9.2 mg/L, pH remained within 8 and 8.1, and temperatures were 20 ฑ1 ฐC.
Dilution water had a total water hardness of 248 mg CaCCb/L. At 48 hours, 0%, 45%, 90%,
75%), 85%), 100%), and 100% immobilization was observed at the mean measured concentrations
of 0.0025, 0.0041, 0.0094, 0.015, 0.024, 0.047, and 0.095 mg/L, respectively. Red coloration on
parts of the exoskeleton was observed in animals exposed to each of the test substance
treatments, which the laboratory notes as being of an uncertain significance.
EPA/OPPT Conclusion
The study is acceptable.
48-hr EC50 = 0.0059 mg/L
(2) A 48-hour acute Daphnia magna toxicity study was conducted the University of Bremen,
Department of Physical and Environmental Chemistry with CP 52 (C12-18, 52% chlorination)
according to DIN 38412 by Koh and Thiemann (2001). Study methods were not fully
characterized. Additional communications with the study author Wolfram Thiemann clarified
that a static test system was used with nominal test concentrations. Based on communications
with the study author, local (Bremen, Germany) tap water was used without adjustments.
Presumed pH was between 5 and 6 and water hardness was between 35.7 and 53.5 mg CaC03/L.
Ambient laboratory air temperature was around 21 ฐC. The solvent acetone was used to maintain
test substance in solution. Floating effects at the surface of the water were observed in individual
cases due to undissolved oil slicks, but communications with the study author noted that there
was no significant loss of daphnids due to mechanical trapping since most daphnid swam away
from these occasional slicks observed at the higher test concentrations.
EPA/OPPT Conclusion
This study is considered acceptable.
48-hr EC50 = 0.052 mg/L
59
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(3) A 48-hour acute Daphnia magna toxicity study was conducted by Thompson et al. (1994)
according to OECD TG 202 (1984) with GLP compliance using a static test system. The test
substance was identified as Cereclor S52ฎ, a C14-17 chlorinated paraffin with 52% chlorination
that was mixed with an equal weight of radiolabeled n-pentadecane-8-14C (51% chlorinated).
Four replicates of 5 Daphnia magna (<24 hours old) were exposed to 0% (dilution water
control), 6.3%), 12.5%, 25%, 50%, and 100% of stock solution containing test substance. The test
substance was prepared in solution by (1) combining 0.75 g test substance and 25 mL acetone to
a borosilicate glass conical flask, (2) evaporation of the acetone using a stream of nitrogen, (3)
addition of 1.5 L dilution water, (4) stirring for three days, and (5) filtration of the aqueous
phase. Radiochemical methods were used to determine the concentration of test substance in
solution. Nominal concentrations of 0 (dilution water control), 0.14, 0.28, 0.55, 1.1, and 2.2
mg/L were within 86-100%) of measured concentrations. Concerns regarding test solution
preparation methods and analytical technique were identified by the submitter that included
increasing the level of more soluble impurities (i.e., short chain chlorinated paraffins),
questionable analytical monitoring results due to the presence of radio-labeled impurities, and
abnormally low recovery of the chlorinated paraffin into hexane. Over the course of the study
dissolved oxygen concentrations remained between 8 and 9 mg/L, pH remained within 8 and 8.1,
and temperatures were 20 ฑ1 ฐC. Dilution water had a total water hardness of 237 mg CaCCb/L.
Observed immobilization was limited to the highest test concentration (100% solution) with 55%
immobilization.
EPA/OPPT Conclusion
The study is unacceptable since EPA/OPPT agrees that methods used to
prepare the test solution and analyze the test concentrations were
questionable.
4) The following study summary (Frank, 1993; Frank and Steinhauser, 1994) provided in the
2005 European Chemical Bureau Risk Assessment of MCCP was considered supportive of the
aquatic invertebrate hazard determination. The chlorinated paraffin used in these studies was a
commercial C14-17 product with a 52% by weight chlorine content. Daphnia magna were exposed
to nominal concentrations of either 100 mg/L or 10,000 mg/L. The 100 mg/L solution was
sonicated for 1 hour and then left to stand in the dark for 48 hours before use. The 10,000 mg/L
solution also stood for 48 hours in the dark before use, but this time without sonication. After
this period, both solutions were filtered firstly with glass filters and then with membrane filters
to remove undissolved test substance. The concentrations of medium-chain chlorinated paraffin
in the water soluble fractions were then determined by AOX (adsorbable organic halogen)
analysis (detection limit of 10 p,g/L CI was equivalent to around 20 p,g/L of the chlorinated
paraffin). This analysis showed that the concentration of chlorinated paraffin present in the water
soluble fraction was around 0.404-0.500 mg/L for the 10,000 mg/L nominal solution and 0.071-
0.142 mg/L L for the 100 mg/L stock solution. The acute (48-hour) toxicity tests were carried
out using dilutions of the two prepared water soluble fractions. The method used was DIN 38
412, Teil 11, which is equivalent to OECD 202.
In the tests using the water soluble fraction from the 100 mg/L nominal solutions no toxicity was
seen at concentrations up to the undiluted stock solution (i.e. no effects up to around 0.071-0.142
mg/L). In experiments using the water soluble fraction from the 10,000 mg/L stock solution, an
ECo of 0.140 mg/L (also reported as 0.100-0.110 mg/L in the paper) and an EC25 of 0.423 mg/L
(also reported as 0.420-0.470 mg/L in the paper) was determined (maximum mortality seen was
25%) (Frank, 1993). The latter results for the 10,000 mg/L stock solution were reported by Frank
and Steinhauser (1994) as ECo = 0.140 mg/L and EC25 = 0.339 mg/L, and it was noted that some
60
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of the Daphnia were floating on the surface of the test solution. In the later study (Frank and
Steinhauser, 1994), the results of further acute toxicity studies were reported using the same test
method. An EC50 of 0.037 mg/L and an ECo of 0.009 mg/L were determined using the water
soluble fraction from the 100 mg/L stock solution and no toxic effects were seen in tests with the
water soluble fraction from the 10,000 mg/L stock solution (approximately ECo >0.525 mg/L).
The authors noted that the effects seen in the acute tests showed poor reproducibility, probably
because effects were seen only around the water solubility limit of the substance. However, the
authors thought that the possibility of undissolved droplets affecting the results could be ruled
out, as floating Daphnia were only sporadically observed in the test.
EPA/OPPT Conclusion
Tthe results of these studies should be treated with caution, as the effects
were mainly seen in the saturated solutions only.
48-hour EC50 = 0.037 mg/L; 100 mg/L stock (Frank and Steinhauser,
1994)
5) The following study summary (Thompson and Gore, 1999) provided in the 2005 European
Chemical Bureau Risk Assessment of MCCP was considered supportive of the aquatic
invertebrate hazard determination. The acute toxicity of C14-17, 52% wt. CI substance was
tested using the freshwater crustacean Gammarus pulex and the freshwater daphnid, Daphnia
magna. The medium-chain chlorinated paraffin used was dissolved in acetone and then added to
beakers in two separate studies containing either Gammarus or D. magna to give nominal
concentrations of 0.1, 0.32, and 1.0 mg/L. A control and solvent control (containing 0.1 mL/L
acetone) were also run. The tests were carried out for 96 hours at 15 ฐC, with the solutions being
renewed after 48 hours. The water used in the study had a hardness of 220 mg/L as CaCCb and
had a pH of 8.0-9.2. No mortalities of the Gammarus were seen in any of the test substance
solutions or control. One animal died in the solvent control. Therefore, no significant toxic
effects were seen with the medium-chain chlorinated paraffin over the concentration range
tested. This contrasted markedly to the situation when Daphnia magna were exposed using the
same test system at 20 ฐC over 48 hours, where complete immobilization was seen at the lowest
test concentration (0.1 mg/L).
EPA/OPPT Conclusion
The high immobilization rate observed in Daphnia magna in this study
appears consistent with the other studies and Gammarus pulex appear to
be a less sensitive to medium chained chlorinated paraffins then Daphnia
magna. EPA/OPPT reserves judgment on the acceptability of this study
until further details become available.
96-hr EC50 (Gammarus pulex) > 1 mg/L
48-hr EC50 (Daphnia magna) <0.1 mg/L
6) The following study summary (Tarkpea et al., 1981; as quoted in WHO, 1996) provided in the
2005 European Chemical Bureau Risk Assessment of MCCP was considered supportive of the
aquatic invertebrate hazard determination. The results of tests with the brackish water
harpacticoid Nitocra spinipes have been reported (Tarkpea et al., 1981). No other details of the
test were reported but the test method was probably the same as reported by Tarkpea et al.
(1986), where a static method was employed using water of salinity 7%o at a temperature of 20-
22 ฐC without aeration, probably using acetone as cosolvent.
EPA/OPPT Conclusion
61
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The results are considered supportive to address aquatic invertebrate acute
toxicity.
96-hour LCso= 9 mg/L (C14-17, 45% wt CI)
96-hour LCso >10,000 mg/L (C14-17, 52% wt. CI)
B-l-3 Algae Toxicity
(1) A 72-hour algae toxicity study was conducted by the University of Bremen, Department of
Physical and Environmental Chemistry with CP 52 (C12-18, 52% chlorination) according to DIN
38412 by Koh and Thiemann (2001). Study methods were not fully characterized. Additional
communications with the study author Wolfram Thiemann clarified that a static test system was
used with nominal test concentrations. Scenedesmus subspicatus were exposed to the test
substance and cell density was determined using a particle counter. Based on communications
with the study author, local (Bremen, Germany) tap water was used without adjustments.
Presumed pH was between 5 and 6 and water hardness was between 35.7 and 53.5 mg CaCCb/L.
Ambient laboratory air temperature was around 21 ฐC. The solvent acetone was used to maintain
test substance in solution. Effects were calculated based on growth rate. No effects were
observed up to 0.1 mg/L.
EPA/OPPT Conclusion
Due to deficiencies/missing details in the study methods, the study alone
was not acceptable to characterize aquatic toxicity to plants.
72-hr NOEC = 0.1 mg/L (Highest Test Concentration)
(2) A 96-hour algae toxicity study was conducted according to OECD TG 201 (2006). The test
substance was a commercial product of a C14-17 chlorinated paraffin with 52% chlorination that
contained 0.3% epoxy soya bean oil stabilizer as well as a small amount of radiolabeled n-
pentadecane-8-14C (51% chlorinated). Selenastrum capricornutum were exposed to nominal
concentrations of 0 (dilution water control), 0 (solvent control), 0.1, 0.18, 0.32, 0.56, 1, 1.8, and
3.2 mg/L test substance in the solvent acetone. Six replicates were tested for each control and
three replicates were tested for each treatment. A mean measured concentration of 0.49, 0.77,
and 1.2 mg/L was determined using radiochemical analysis for the nominal concentrations of 1,
1.8, and 3.2 mg/L, respectively, but effects were reported based on nominal test concentrations.
At the start of the test, the pH was 7.4-7.5, but had reached 10.0-10.3 by the end of the test. The
shift in pH was thought to be a function of the high control growth rates observed in the test
according to the study summary. The section-by-section coefficient of variation for the solvent
control remained below 35% indicating acceptable control growth rates throughout the duration
of the study.
EPA/OPPT Conclusion
The maximum inhibition in the growth rate and biomass seen was 3% and
18%), respectively, but a dose response relationship was not seen. The
nominal NOEC was 0.1 mg/L and the nominal LOEC based on 18%
biomass inhibition was 0.18 mg/L. A GMATC of 0.134 mg/L was
calculated. The study was considered acceptable.
72-hour EC50= >3.2 mg/L (nominal); 1.2 mg/L (mean measured).
96-hour EC50= >3.2 mg/L (nominal); 1.2 mg/L (mean measured).
72-hr NOECb = 0.1 mg/L
72-hr LOECb = 0.18 mg/L
72-hr GMATC = 0.134 mg/L
62
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B-l-4 Chronic Fish Toxicity
(1) A 60-day fish toxicity study was conducted by Brixham Laboratories in 1983 with radio-
labeled chlorinated (52%) n-pentadecane (Trade Name: Cereclor S52ฎ) under flow through
testing conditions. A full non-CBI study report was submitted under TSCA in 1983 as DCN 40-
8332184 (OTS Fiche 0507258). Two replicates of 3 immature rainbow trout (Salmo gairdneri)
per concentration were exposed to nominal concentrations of 0 (dilution water control), 0
(acetone control, 500 ppm), 1, or 5.6 mg/L in 500 ppm acetone. Corresponding mean measured
concentrations were 0, 0, 1.05, and 4.5 mg/L. Test concentrations were determined by radio
activity measurements. Flow rate of the test system was 0.25 mL/minute for exposure
concentrations. No mortality or adverse sub-lethal behavioral effects were observed for the
duration of the 60-day exposure period. Effects observed were limited to the highest test
concentration and involved sluggish movements. The measured NOEC was identified as 4.5
mg/L. In addition to the hazard assessment, the submitter provided an assessment of
bioconcentration which indicated that analytically determined exposure concentrations of 1.05
and 4.5 mg/L resulted in fish tissue concentrations of 34 and 190 |ig/g wet weight, respectively.
EPA/OPPT Conclusion
This study appears to have been previously reviewed by EPA in 1985. The
previous conclusion that "a fish full life cycle toxicity test or modification
thereof is needed to address the effects of CPs present in fish eggs during
embryonic development" (U.S. EPA Memorandum, 1991) is still relevant
for MCCPs. Thus, this study is considered unacceptable to characterize
chronic population-level effects in fish.
60-day NOEC = 4.5 mg/L
(2) A 28-day fish toxicity study was conducted by Brixham Laboratories in 1978 with a C14-17
chlorinated paraffin having 45% chlorination under unspecified testing conditions. The study
report was submitted under TSCA in 1992 as DCN 88920006972 (OTS Fiche 0545375).
Rainbow trout {Salmo gairdneri) per concentration were exposed to nominal concentrations of 0
(dilution water control), 0.1, or 1 mg/L in acetone. The age and size of the rainbow trout used in
the study were not specified in the study. The specific environmental conditions of the test, such
as pH, temperature and water quality were not specified in the report. Concentrations of the
chlorinated paraffin in the test water were measured using TLC analytical procedures resulting in
mean measured concentrations of 0.01 and 0.18 mg/L. Mortality and behavior (response to food,
general behavior, swimming behavior and pigmentation) were assessed during the 28-day study.
Survival was 96.6 and 100% for the mean measured exposures of 0.01 and 0.18. No behavioral
effects were seen over the course of the study.
EPA/OPPT Conclusion
The study was considered unacceptable to characterize the chronic fish
toxicity endpoint since insufficient study details were provided including
the age and/or the life-stages of the exposed organisms.
(3) A 20-day Japanese medaka (Oryzias latipes) embryo toxicity study was conducted with the
formulation C14H24.9C15.1,48%>C1 (composition: 10.5% 1, 2, 13, 14-tetrachlorotetradecane (42.3%
CI); 74.3%) x, 1, 2, 13, 14-pentachlorotetradecane (47.7% CI); 14.2%> x, y, 1, 2, 13, 13-
hexachlorotetradecane (52.6%> CI); 1.0% x, y, z, 1, 2, 13, 14-heptachlorotetradecane (56.4% CI))
by Fisk et al. (1999) under static testing conditions. Five sets of 10 vials containing 1 egg each
were exposed to nominal concentrations of 0.001, 0.010, 0.100, 1, or 10 mg/L test substance
63
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starting after fertilization and terminating approximately 3 days post-hatch. No adverse effects
were reported in exposed embryos.
EPA/OPPT Conclusion
The study was considered unacceptable primarily due to insufficient
exposure duration and insufficient number of eggs per exposure
concentration.
(4) A 20-day Japanese medaka (Oryzias latipes) embryo toxicity study was conducted with the
formulation 14C-Ci4H23.3Cl6.7, 55% CI (composition: 0.2% C14H26CI4 (42.3% CI), 4.4%
C14H25CI5 (47.7% CI), 34% C14H24CI6 (52.6% CI), 45% C14H23CI7 (56.4% CI), 14% C14H22CI8
(59.9%) CI), and 1.9% C14H21CI9 (62.8% CI)) by Fisk et al. (1999) under static testing conditions.
Five sets of 10 vials containing 1 egg each were exposed to measured concentrations of 0.0014,
0.012, 0.120, 0.420, or 1.6 mg/L test substance starting after fertilization and terminating
approximately 3 days post-hatch. No adverse effects were reported in exposed embryos.
Concentrations of the test substance were found in larvae and eggs in a dose-dependent manner
(with exception of the highest concentration) suggesting that the substance can diffuse through
the egg. Corresponding measured concentrations in eggs were 0.04, 8.4, 63, 110, and 72 mg/kg
and corresponding measured concentrations in larvae were 0.24, 8.2, 45, 84, and 51 mg/L.
EPA/OPPT Conclusion
The study was considered unacceptable primarily due to insufficient
exposure duration and insufficient number of eggs per exposure
concentration.
(5) The following summary provided in the 2005 European Chemical Bureau Risk Assessment
of MCCP was considered supportive, but did not characterize all fish life-cycle stages. Cooley et
al. (2001) studied the toxicity of C14H24.9C15.1,48%C1 (as described in Fisk et al., 1999) to
juvenile rainbow trout (1Oncorhynchus mykiss) through dietary exposure. Treatment groups of 10
fish were exposed to 0.78 and 2.9 mg/kg for 21 days and 0.082 mg/kg for 85 days. Three control
groups were also run. Histological examination and analysis of the chlorinated paraffin
concentration was performed in five fish per treatment after 21 days in the two higher test
concentrations and in three fish per treatment after 85 days in the lowest test concentration.
Three fish were also sacrificed from each low exposure group and the remaining control group
(but were not analyzed) after 21 days of exposure. Quantitative histomorphological
measurements were also carried out on livers and thyroid of the exposed fish in the middle
exposure group after 21 days, and also the low exposure group after 85 days. The parameters
investigated included hepatocyte nuclear diameter, hepatocyte volume index, nucleusxytoplasm
area ratio and thyroid epithelium cell height. Livers displaying mild hepatocyte necrosis and
moderate to severe depletion of glycogen/lipids were reported for the 0.78 mg/kg exposure. At
2.9 mg/L abnormal behavior was observed from day 3 onwards. Quantitative effects following
21 days of exposure were limited to a significantly (p=0.05) reduced mean hepatocyte volume in
2.9 mg/L exposure group.
EPA/OPPT Conclusion
This study is considered unacceptable to characterize chronic mortality in
fish because it did not characterize life stages but instead characterized
physiological effects.
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(6) Cooley et al. (2001) studied the toxicity of another medium chain chlorinated paraffin with a
slightly different chemical composition and at slightly different concentration levels. The
chemical formula was 14C-Ci4H23.3Cl6.7, 55% CI (as described in Fisk et al., 1999) and was
juvenile rainbow trout (Oncorhynchus mykiss) were exposed through the diet. Treatment groups
of 10 fish were exposed to 29 and 78 mg/kg for 21 days and 5.7 mg/kg for 85 days. Three
control groups were also run. Histological examination and analysis of the chlorinated paraffin
concentration was performed in five fish per treatment after 21 days in the two higher test
concentrations and in three fish per treatment after 85 days in the lowest test concentration.
Three fish were also sacrificed from each low exposure group and the remaining control group
(but were not analyzed) after 21 days of exposure. Quantitative histomorphological
measurements were also carried out on livers and thyroid of the exposed fish in the middle
exposure group after 21 days, and also the low exposure group after 85 days. The parameters
investigated included hepatocyte nuclear diameter, hepatocyte volume index, nucleusxytoplasm
area ratio and thyroid epithelium cell height. At 29 mg/kg abnormal behavior was observed from
day 2 onwards and livers exhibited mild to moderate hepatocyte necrosis and moderate to severe
depletion of glycogen lipids. Abnormal behavior from day 3 onward was also observed at 78
mg/kg.
EPA/OPPT Conclusion
This study is considered unacceptable to characterize chronic mortality in
fish because it did not characterize life stages but instead characterized
physiological effects.
B-l-5 Chronic Aquatic Invertebrate Toxicity
(1) A 21-day chronic Daphnia magna reproduction toxicity study was conducted by Thompson
et al. (1997b) according to OECD 202, Part II (1984) using a static-renewal test system with
renewal 3 times/week. The test substance was identified as Cereclor S52ฎ , a C14-17 chlorinated
paraffin with 52% chlorination that contained 0.3% epoxy soya bean oil stabilizer as well as a
small amount of radiolabeled n-pentadecane-8-14C (51% chlorinated). Ten replicates of 1
Daphnia magna Straus (<24 hours old) were tested per exposure concentration, which did not
comply with OECD 202, Part II requirements that at least 40 daphnid be tested per
concentration. Nominal concentrations were 0 (dilution water control), 0 (solvent control),
0.0056, 0.01, 0.018, 0.032, 0.056, and 0.1 mg/L test substance in acetone (0.025 mL/L). Results
from the acute daphnid study by the same author do not appear to have been considered when
selecting concentrations for this study. Test solutions were prepared by adding the appropriate
stock solution to dilution water while continuously and vigorously stirring with a magnetic
follower. The submitter does not indicate whether renewal of the static-renewal test systems was
carried out at regular intervals (e.g, Monday-Wednesday-Friday). Corresponding mean measured
concentrations determined by radiochemical methods were 0.0037, 0.005, 0.01, 0.018, 0.032,
and 0.065 mg/L and were 78-94%) of nominal concentrations at the start of the renewal period
and 7.3-61%) of nominal concentrations at the end of the renewal period indicating a notable loss
of test substance. Also, analysis of test concentrations appears to be at irregular intervals. In the
dilution water control, 20%> mortality was observed. Overall, dilution water control and solvent
control results were significantly different for reproductive parameters. The test was carried out
at temperatures of 19.5-20.3 ฐC, at pH levels of 7.41-8.13, and at dissolved oxygen
concentrations of 6.2-9.2 mg/L. A significant decrease in the number of live offspring was
reported at the mean measured concentration of 0.018 mg/L and delayed release of first offspring
was observed at higher concentrations. Percentage dead offspring reported was 0%>, 0%>, 5.9%>,
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20.4%, and 18.5% for the 0.0037, 0.005, 0.01, 0.018, 0.032, and 0.065 mg/L mean measured
exposures.
EPA/OPPT Conclusion
The inability to maintain test concentrations, unspecified renewal periods,
and a smaller population size may have affected subsequent reproductive
results. Given the uncertainties of the test, reported effect levels may not
represent a worst case scenario but do exhibit a clear dose response
relationship with a clearly defined statistically significant effect level.
Thus, using a weight of evidence approach, the study is considered
acceptable for this endpoint.
21-day LCso value of 0.025 mg/L (parent mortality)
21-d NOEC = 0.01 mg/L
21-d LOEC = 0.018 mg/L
21-d GMATC = 0.013 mg/L
(2) A 60-day mussel toxicity study was conducted by Brixham Laboratories in 1983 with radio-
labeled chlorinated (52%) n-pentadecane (Trade Name: Cereclor S52ฎ) under flow through
testing conditions. A full non-CBI study report was submitted under TSCA in 1983 as DCN 40-
8332184 (OTS Fiche 0507258). Blue mussels (Mytilus edulis) were exposed to nominal
concentrations of 0 (dilution sea water control), 0 (acetone control, 500 ppm), 0.56, or 5.6 mg/L
in 500 ppm acetone. Two replicates of 50 mussels were tested for the dilution water and solvent
controls and a single replicate of 50 mussels was exposed for each treatment concentration.
Corresponding mean measured concentrations were 0, 0, 0.22, and 3.8 mg/L. Test solutions were
cloudy at the higher test concentration. Test concentrations were determined by radio activity
measurements. Flow rate of the test system was 0.25 mL/minute for exposure concentrations.
Over the course of the study, water temperature ranged from 14.6 - 15.6 ฐC, pH ranged from 8.0
- 8.3, and the dissolved oxygen concentrations ranged from 6.1 - 8.25 mg/L. Dilution water
salinity was 34-35.5 ppb, which is high by OCSPP standards.
EPA/OPPT Conclusion
One mussel exposed to 0.56 mg/L died, and two mussels exposed to the
controls died; this was not considered to be a test substance related effect.
Decreases in filter feeding were observed at 5.6 mg/L. In addition, the
submitter provided an assessment of bioconcentration, but this assessment
does not appear to include a depuration phase. Overall, the 60 day NOEC
and LOEC were 0.22 and 3.8 mg/L based on reduced filtration. The study
was acceptable to characterize mussel toxicity, but mussels are not
considered a standard species to fulfill the chronic aquatic invertebrate
toxicity endpoint.
60-d NOEC = 0.22 mg/L
60-d LOEC = 3.8 mg/L (reduced filtration)
(3) The following study summary (Frank, 1993; Frank and Steinhauser, 1994) provided in the
2005 European Chemical Bureau Risk Assessment of MCCP was considered supportive of the
aquatic invertebrate hazard determination. A 21-day chronic Daphnia magna reproduction
toxicity study was conducted using a static-renewal test system (renewal 3 times/week). The test
substance was identified as C14-17 chlorinated paraffin with 52% chlorination, which was tested
as a water soluble fraction of two stock solutions (dilutions used 1:2 to 1:32). Nominal test
concentrations prepared from the 100 mg/L stock solution were 3.125, 6.25, 12.5, 25, and 50
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mg/L. Nominal test concentrations prepared from the 10,000 mg/L stock solution were 312.5,
625, 1250, 2500, and 5000 mg/L. Analytical monitoring of test concentrations was conducted,
but only the final effect levels were presented as measured concentrations. Methods for test
solution preparation were not provided in the summary. The tests were carried out at 20 ฐC and
at pH 7.79-8.44.
In the experiments using the 100 mg/L stock solution the mortality seen in the exposed
populations was 0% at 3.125 mg/L, 0% at 6.25 mg/L, 20% at 12.5 mg/L, 90% at 25 mg/L, and
100%) at 50 mg/L. In the experiments using the 10,000 mg/L stock solution the mortality seen in
the exposed populations was 0% at 312.5 mg/L, 30% at 625 mg/L, 70% at 1250 mg/L, and 100%
at lower dilutions (>1250 mg/L). In the experiments using the 100 mg/L stock solution the
average number of young/adult was 82 at 3.125 mg/L, 89 at 6.25 mg/L, 80 at 12.5 mg/L, 15 at
25 mg/L and 0 at 50 mg/L (all parents died). Similarly in the experiments using the 10,000 mg/L
stock solution the average number of young/adult was 74 at 312.5 mg/L, 64 at 625 mg/L, 43 at
1250 mg/L, and 0 at 2,500 and 5,000 mg/L (all parents died). Based on these effects,
survivability/mortality appears to be the more sensitive endpoint. Based on the known measured
concentrations in the stock solutions and the dilution rates used the NOEC for mortality was
around 0.0044-0.0089 mg/L for the 100 mg/L nominal stock solution experiments and 0.0126-
0.0156 mg/L for the 10000 mg/L nominal stock solution experiments. The corresponding
LOECs were 0.0089- 0.0178 mg/L (100 mg/L nominal stock) and 0.0253-0.0313 (J,g/L (10 g/L
nominal stock). The GMATC of 0.006 mg/L was calculated using the geometric mean from the
most conservative NOEC (0.0044 mg/L) and LOEC (0.0089 mg/).
EPA/OPPT Conclusion
EPA/OPPT reserves judgment on the acceptability of this study until
further details become available.
21-d NOEC = 0.0044 mg/L
21-d LOEC = 0.0089 mg/L
21-d GMATC = 0.006 mg/L
3) The following study summary (TNO, 1993) provided in the 2005 European Chemical Bureau
Risk Assessment of MCCP was considered supportive of the aquatic invertebrate hazard
determination. A 21-day chronic Daphnia magna reproduction toxicity study was conducted
using a static-renewal test system (renewal 3 times/week). The test substance was identified as
C14-17 chlorinated paraffin with 52% chlorination. The test solutions were prepared by stirring 20
g of the test substance in 2 litres of heated water (60 ฐC) with stirring and then filtration through
a 0.8 |im and 0.2 |im filter. This resulting stock solution was referred to as a water soluble
fraction, but given that each concentration was not independently prepared, the test solutions is
considered by EPA/OPPT to be merely a mixed and filtered solution that was subsequently
diluted. Following dilution of the stock solution, exposure concentrations were analytically
determined using the extractable organic halogen method, but not provided in the study
summary. The test was carried out at 20 ฑ 1 ฐC and solutions were gently aerated from day 9
onwards. The pH of the test water varied between 7.7 and 8.3, the dissolved oxygen
concentration was > 7 mg/L, and the hardness was 214 mg CaC03/L. Test solutions were clear.
EPA/OPPT Conclusion
Although analytical results obtained were considered to be too erratic to
allow precise determination of concentrations (according to the study
summary in ECB, 2005), the NOEC was reported based on survivability
and/or reproductive effects. A LOEC in mg/L was not reported in the
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summary, nor could EPA/OPPT extrapolate one. EPA/OPPT reserves
judgment on the acceptability of this study until further details become
available.
21-d NOEC = 0.004-0.008 mg/L
B-l-6 Chronic Aquatic Sediment Invertebrate Toxicity
1) Thompson et al. (2001c) conducted a 28-day prolonged sediment invertebrate toxicity study
with spiked sediment was conducted according to OECD 218 draft guideline (February 2000
version) using a static test system. The substance used in the test was commercial C14-17, 52% wt.
CI substance containing no stabilizers (the substance was reported to have a C14-17 content of
99.06% with 0.67%) of C10-13 chain length substances) that was mixed with a small amount of a
radio labeled n-pentadecane-8-14C, 51% wt. CI substance (radiochemical purity >96.6%). The
sediment used in the test was an artificial sediment that did not fully adhere to the final OECD
TG recommendations, but the composition of 10% sphagnum moss peat, 70% quartz sand, 20%
kaolinite clay, and <0.1% calcium carbonate are not considered to be significantly different. The
sediment had a mean organic carbon content of 4.9% and a pH of 6.0. The sediment was spiked
with the test substance by firstly mixing a solution of the test substance in acetone with the dry
sand component of the sediment and allowing the acetone to evaporate overnight under an air
stream. Measured concentrations were determined using radiochemical analysis. Over the course
of the study, temperature was maintained at 20 ฐC, pH levels were 6.2-7.6, and dissolved oxygen
in overlying water was maintained at 7.3-8.6 mg/L. Three replicates of 15 midge (Chironomus
riparius) larvae (<48 hours post hatch) were exposed to mean measured concentrations of 0
(sediment control), 0 (solvent sediment control), 36, 110, 370, 1200, 3800, or 13000 mg/kg dry
wt. sediment. Time to first emergence, mean emergence time, mean number emerged per
replicate, and sex ratio was assessed for each exposure group. Statistically significant effects (p
= 0.05) were limited to a decrease in mean number emerged per replicated in the 13,000 mg/kg
dry wt. sediment exposed midges.
EPA/OPPT Conclusion
The overall NOEC of 3,800 mg/kg dry wt. sediment corresponded to
1,460 mg/kg on a wet weight basis. The study is acceptable.
28-d NOEC = 3,800 mg/kg dry wt sediment
28-d LOEC = 13,000 mg/kg dry wt sediment
28-d GMATC = 7,029 mg/kg dry wt sediment
2) Thompson et al. (200Id) conducted a 28-day prolonged sediment invertebrate toxicity study
with spiked sediment according to methods described in Phipps et al. (1993) using a static test
system. The substance used in the test was commercial C14-17, 52% wt. CI substance containing
no stabilizers (the substance was reported to have a C14-17 content of 99.06% with 0.61% of C10-
13 chain length substances) that was mixed with a small amount of a radio labelled n-
pentadecane-8-14C, 51% wt. CI substance (radiochemical purity >96.6%). The sediment used in
the test was an artificial sediment consisting of 10% sphagnum moss peat, 70% quartz sand, 20%
kaolinite clay, and <0.1% calcium carbonate. The sediment had a mean organic carbon content
of 4.9% and a pH of 6.0. The test sediments were made up by adding the test substance to the
sand phase as a solution in acetone, evaporating the acetone overnight and mixing the spiked
sand with the rest of the sediment for 16 hours. Six replicates of 10 oligochaete (Lumbriculus
variegatus) adults were exposed to mean measured concentrations of 0 (sediment control), 0
(solvent sediment control), 39, 130, 410, 1300, 4000, or 13000 mg/kg dry wt. sediment.
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Throughout the duration of the study, water temperature was maintained at 20 ฑ 1 Cฐ, and pH
remained between 6.3 and 7.9. Mortality and reproductive success were determined by total
number of worms at study termination since differentiation of adult and young worms is
difficult. Mean number of worms per replicate and mean total dry weight of worms per replicate
was significantly different from controls (p = 0.01) at mean measured concentrations of 410
mg/kg dry weight sediment and greater. Statistical methods used were not reported, and a p-
value of 0.01 was used.
EPA/OPPT Conclusion
Given that a clear decline in the mean number of worms per replicate and
mean total dry weight of worms per replicate was observed at the lowest
test concentration (39 mg/kg dry weight sediment) a conservative LOEC
of 39 mg/kg dry weight sediment will be used based on noticeable
differences. EPA/OPPT reserves judgment on the acceptability of this
study until further details become available regarding the analytical
measurements of the chlorinated paraffin mixture.
28-d NOEC = 130 mg/kg dry wt sediment
28-d LOEC = 410 mg/kg dry wt sediment
GMATC = 230.9 mg/kg dry wt sediment
3) Thompson et al. (2002) conducted a 28-day prolonged sediment toxicity study with amphipod
Hyalella azteca in spiked sediment using a static-renewal test system with weekly renewals. The
substance used in the test was a mixture of a commercial medium-chain chlorinated paraffin
product (C14-17, 52.5% wt. CI) mixed with a small amount of a radiolabeled chlorinated n-
pentadecane-8-14C (51% wt. CI). The sediment used in the test was an artificial sediment
consisting of 10% sphagnum moss peat, 70% quartz sand, 20% kaolinite clay, and <0.1%
calcium carbonate. The sediment had a mean organic carbon content of 4.9% and a pH of 6.0.
The test sediments were made up by adding the test substance to the sand phase as a solution in
acetone, evaporating the acetone overnight and mixing the spiked sand with the rest of the
sediment and water. Six replicates per concentration of ten juvenile Hyalella azteca (~7-day-old)
were exposed to 0 (sediment control), 0 (acetone sediment control), 38, 75, 150, 300, or 600
mg/kg dry weight sediment. The concentration of the test substance was measured in the
sediment phase by radiochemical analysis with concentrations at the start of the exposure period
of 85-97%) of the nominal values and concentrations at the end of the 29-day exposure period of
78-90%) of nominal. Results of the test were expressed as the arithmetic mean concentration.
Over the course of the study, dissolved oxygen ranged from 7.7 to 8.4 mg/L, pH ranged from 7.0
to 7.6, water hardness ranged from 41 to 42 mg CaC03/L, and temperature ranged from 22.4-
23.2ฐC. The endpoints investigated in the study included survival, growth (dry weight) and
sexual development of females (proportion of gravid females).
EPA/OPPT Conclusion
Controls responded adequately. For the survival endpoint, a statistically
significant (p=0.05) reduction in survival was seen at 470 mg/kg dry
weight. A statistically significant reduction in mean weight was seen in
females only at exposures of 470 g/kg dry weight sediment and a
statistically significant (p=0.05) reduction in mean weight was seen at 270
mg/kg dry weight. For the sexual development endpoint, there was a
statistically significant (p=0.05) reduction in the proportion of gravid
females in the 470 mg/kg dry weight treatment. This study was
acceptable.
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28-d NOEC = 130 mg/kg dry wt sediment
28-d LOEC = 270 mg/kg dry wt sediment
28-d GMATC = 187 mg/kg dry wt sediment
B-l-7 Avian Toxicity
(1) An acute avian toxicity study conducted according to OPPTS guidelines was published by
Madeley & Birtley (1980). Following a range-finding study, groups of 5 male and 5 female ring-
necked pheasants (Phasianus colchicus) were exposed by gavage to 0 (control) or 24,606 mg/kg
Cereclor S52ฎ (C14-17, 52% CI) and then observed for 14 days. Based on reported tissue
concentrations, the test substance is believed to have been absorbed by the ring-necked pheasant.
Doses up to 24,606 mg/kg failed to produce any abnormal clinical signs or mortality.
EPA/OPPT Conclusion
Acute LD50 > 24,606 ppm
(2) An acute avian toxicity study conducted according to OCSPP guidelines was published by
Madeley & Birtley (1980). Following a range-finding study, groups of 5 male and 5 female
mallard ducks (Anasplatyrynchos) were exposed by gavage to 0 (control) or 10,280 mg/kg
Cereclor S52ฎ (C14-17, 52% CI) and then observed for 14 days. Based on reported tissue
concentrations, the test substance is believed to have been absorbed by the mallard ducks. Doses
up to 10,280 mg/kg failed to produce any abnormal clinical signs or mortality.
EPA/OPPT Conclusion
Acute LD50 > 10,280 ppm
(3) A sub-acute dietary avian toxicity study conducted according to OPPTS guidelines was
published by Madeley & Birtley (1980). Following a range-finding study, groups of 5 male and
5 female ring-necked pheasants (Phasianus colchicus) were exposed to diets containing 0
(control), 1,000, or 24,063 ppm Cereclor S52ฎ (C14-17, 52% CI) for 5 days. Three groups were
exposed to the negative control and two groups were exposed to each of the treatment
concentrations. Based on reported tissue concentrations, the test substance is believed to have
been absorbed by the ring-necked pheasant. Good health was noted in all control and treatment
groups. No abnormal effects were noted at necropsy.
EPA/OPPT Conclusion
5-day LD50 > 24,063 ppm
(3) A sub-acute dietary avian toxicity study conducted according to OPPTS guidelines was
published by Madeley & Birtley (1980). Following a range-finding study, groups of 5 male and
5 female mallard ducks (Anas platyrynchos) were exposed to diets containing 0 (control), 1,000,
or 24,063 ppm Cereclor S52ฎ (C14-17, 52% CI) for 5 days. Three groups were exposed to the
negative control and two groups were exposed to each of the treatment concentrations. Inferior
food intake was noted for ducks, but weight gain was comparable to controls. Based on reported
tissue concentrations, the test substance is believed to have been absorbed by the mallard ducks.
Good health was noted in all control and treatment groups. No abnormal effects were noted at
necropsy.
EPA/OPPT Conclusion
5-day LD50 > 24,063 ppm
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B-l-8 Terrestrial Invertebrate Toxicity
1) A 28-day earthworm reproductive toxicity test was conducted by Thompson et al., 2001a
according to OECD guideline (2000 draft version). The substance tested was commercial C14-
17, 52% wt. CI substance containing no stabilizers (the substance was reported to have a C14-17
content of 99.06% with 0.61% of C10-13 chain length substances) and a small amount of re-
labelled n-pentadecane, 51% wt. CI substance. Four replicates per concentration of 10 adult
earthworm (Eisenia fetida) were exposed to nominal concentrations of 0 (soil control), 0 (solvent
soil control), 100, 320, 1000, 3200, or 10,000 mg/kg dry wt. soil. Corresponding mean measured
concentrations of 0 (soil control), 0 (solvent soil control), 79, -280, 900, -2,800, or 9,300 mg/kg
dry wt. was determined using radiochemical analysis; concentrations identified as approximate
(~) were approximated using the mean% of nominal (87%) determined in other treatments.
Measured tissue concentrations in adults on day 28 were 169, 802, and 732 mg/kg wet weight for
the 79, 900, and 9,300 mg/kg dry weight exposure groups. Measured tissue concentrations in
juveniles on day 56 were 140 and 1,011 mg/kg wet weight for the 79 and 900 mg/kg dry weight
exposure groups. The soil used in the test was an artificial soil consisting of 10% sphagnum
moss peat, 70% quartz sand, 20% kaolinite clay, and 0.25% calcium carbonate. The soil had an
organic carbon content of 4.7% and a pH of 6.66-7.09. Nominal test temperatures remained at 20
ฑ 1 ฐC. The soils were prepared up by firstly adding the test substance in solution with acetone to
a small portion of soil, evaporating out the acetone overnight under a stream of compressed air,
and then mixing with the remainder of the soil. Before use, distilled water was added to the dry
soil to provide a soil wet:dry ratio of 1.35. Following the 28-day parental exposure period, adult
earthworms were removed, and vessels were incubated for an additional 28 days to allow
hatching of any egg cocoons produced by parent. Effects assessed were parental survival, growth
as determined by change in weight of parents, and reproduction as determined by number of live
offspring. A statistically significant (p = 0.05) reduction in parental survival (85%) was observed
at 9,300 mg/kg dry wt. soil. A statistically significant (recalculated with Dunnett's Procedure, P
= 0.05) reduction in parental weight was reported at 2800 mg/kg dry wt. soil. A statistically
significant (recalculated with Dunnett's Procedure, P = 0.05) reduction in number of live
offspring was reported at 280 mg/kg dry wt. soil. In addition, the submitter assesses
corresponding tissue concentrations in earthworms and determines that at nominal
concentrations of 100 mg/kg dry wt. soil the concentration in parental earthworm tissue after 28
days is 850 mg/kg dry wt.and in juvenile worms after 56 days was 703 mg/kg dry wt.
EPA/OPPT Conclusion
The study was acceptable.
28-d NOEC = 79 mg/kg dry wt soil
28-d LOEC = 280 mg/kg dry wt soil
28-d ChV = 149 mg/kg dry wt soil
2) The following study summary (Thompson et al., 2001a) provided in the 2005 European
Chemical Bureau Risk Assessment of MCCP was used to characterize terrestrial invertebrate
hazard in a screening level risk assessment. A 28-day earthworm reproductive toxicity test was
conducted according to OECD guidelines. The substance tested was commercial C14-17, 52% wt.
CI substance containing no stabilizers (the substance was reported to have a C14-17 content of
99.06% with 0.61% of C10-13 chain length substances). The test substance contained a small
amount of 14C-labelled n-pentadecane, 51% wt. CI substance so that radiochemical analysis
could be used to analytically determine test concentrations in soil. The soil used in the test was
an artificial soil consisting of 10% sphagnum moss peat, 70% quartz sand, 20% kaolinite clay,
and 0.25%) calcium carbonate. The soil had an organic carbon content of 4.7% and a pH of 6.66-
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7.09. The soils were prepared up by firstly adding the test substance in solution with acetone to a
small portion of soil, evaporating out the acetone overnight under a stream of compressed air,
and then mixing with the remainder of the soil. Before use, distilled water was added to the dry
soil to provide a soil wet:dry ratio of 1.35. Four replicates per concentration of 10 adult
earthworm (Eisenia fetida) were exposed to mean measured concentrations of 0 (soil control), 0
(solvent soil control), 79, -280, 900, -2800, or 9300 mg/kg dry wt. for 28 days; concentrations
identified as approximate (~) were approximated using the mean% of nominal (87%) determined
in other treatments. Measured tissue concentrations in adults on day 28 were 169, 802, and 732
mg/kg wet weight for the 79, 900, and 9300 mg/kg dry weight exposure groups. Measured tissue
concentrations in juveniles on day 56 were 140 and 1011 mg/kg wet weight for the 79 and 900
mg/kg dry weight exposure groups. Following the 28-day parental exposure period, adult
earthworm were removed, and vessels were incubated for an additional 28 days to allow
hatching of any egg cocoons produced by parent. Effects assessed were parental survival, growth
as determined by change in weight of parents, and reproduction as determined by number of live
offspring. Statistical methods used to calculate significance were not provided. A statistically
significant (p = 0.05) reduction in parental survival (85%) was observed at 9300 mg/kg dry wt.
soil. A statistically significant (p = 0.01) reduction in parental weight was reported at 2800
mg/kg dry wt. soil and a noticeable reduction in parental weight was reported at 280 mg/kg dry
wt. soil. A statistically significant (p = 0.01) number of live offspring was reported at 1000
mg/kg dry wt. soil and a noticeable reduction in number of live offspring was reported at 320
mg/kg dry wt. soil. A clear decline in parental weight and number of live offspring was observed
at 280 mg/kg dry wt. soil.
EPA/OPPT Conclusion
This study is acceptable.
28-d NOEC = 79 mg/kg dry wt soil
28-d LOEC = 280 mg/kg dry wt soil
28-d GMATC = 149 mg/kg dry wt soil
B-l-9 Terrestrial Plant Toxicity
1) Thompson et al., 200lab conducted a 28-day seed germintation and vegetative vigor study.
The toxicity of a C14-17, 52% wt. CI substance (99.06%) purity) to wheat {Triticum aestivum\
monocotyledon), oilseed rape (Brassica napus; ditcotyledon), and mung bean (Phaseolus
aureus; dicotyledonous legume) has been studied using OECD guideline 208 (July, 2000
Revision). The test substance contained a small amount of 14C-labelled n-pentadecane, 51%> wt.
CI substance so that radiochemical analysis could be used to analytically determine test
concentrations in soil. The soils were prepared up by firstly adding the test substance in solution
with acetone to dry silver sand, evaporating the acetone overnight, and mixing the spiked sand
with the soil. Four replicate pots per exposure concentration each containing 9 seeds were
exposed for 28 days to nominal exposure concentrations of 0 (soil control), 0 (solvent soil
control), 50, 158, 500, 1,580, or 5,000 mg/kg dry wt.. According to the E.U. Risk Assessment,
use of nominal test concentrations to determine effect levels is based on test substance stability
determined at the 50, 500, and 5,000 mg/kg dry wt. concentrations (corresponding mean
measured concentrations: 49, 520, and 5,800 mg/kg dry wt.). Effects assessed were seed
germination, emergence (%> emerged plants on Day 14), vegetative growth (mean shoot dry
weight per plant), and visual appearance of seedling. No statistically significant differences (p =
0.05) were observed in wheat, oilseed rape. A statistically significant reduction in growth was
seen at 1,580 and 5,000 mg/kg dry wt. for mungbean when compared to soil control results.
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EPA/OPPT Conclusion
Since soil control and solvent control means were equal (two tailed T-
Test) indicating no solvent interference, comparison of treatments was
made to the soil control. Thus, the NOEC and LOEC for terrestrial plants
was 500 and 1,580 mg/kg dry wt. soil and the GMATC (geometric mean
of the NOEC and LOEC) was 888.8 mg/kg dry wt. soil. The study was
acceptable to characterize both monocot (wheat) and dicot (mung bean)
seed germination and vegetative vigor; reproductive effects remain
uncharacterized.
28-d NOEC = 500 mg/kg dry wt soil
28-d LOEC = 1,580 mg/kg dry wt soil
28-d GMATC = 888.8 mg/kg dry wt soil
B-l-10 Conclusions
Sufficient data were available to characterize the acute fish, the acute aquatic invertebrate, the
chronic aquatic invertebrate, the chronic aquatic sediment invertebrate, avian, and terrestrial
plant toxicity endpoints for MCCPs. Data for other toxicity endpoints (i.e., chronic fish, aquatic
plant, etc.) were inconclusive due to lack of study details, uncertainties in analytical methods, or
test material preparation methods; thus, these data are included in order to characterize risk in a
qualitative manner, but are used as supportive for the categories under which they are provided.
Supporting data were included in order to provide a weight-of-evidence approach used to
characterize some endpoints.
Most of the data provided in this review indicated several difficulties were encountered when
testing in an aquatic environment. These included: (1) getting the material into solution, (2)
measuring the material in solution, and (3) characterizing the effects for each study listed. Often
there were many details of a given study omitted, prohibiting a full and robust review of the data.
The (estimated) physical-chemical properties of MCCPs (water solubility values of
approximately 30 [j,g/L and Log Kowvalues between 4-8) suggest these materials may not
partition to the aquatic media or elicit toxicity to aquatic organisms within the water column.
The most reliable and acceptable studies indicate that for MCCPs, the toxicity to aquatic
organisms for acute endpoints are from the Thompson et al. 1996 study for aquatic invertebrates.
Where the 48-hour ECso value = 0.0059 mg/L. Using the methods described in the Sustainable
Futures/P2 Manual (US EPA, 2012), the acute and chronic concentrations of concern (CoC) are
derived as follows:
Acute CoC: The 48-hour ECso value = 0.0059 mg/L is divided by an assessment factor of
5 to yield an acute concentration of concern (CoC) of 0.00118 mg/L, or 0.001 mg/L, or 1
[j,g/L (1 ppb). Aquatic Acute Concern Concentration= 1 ppb
Chronic CoC: The aquatic invertebrate chronic value of 0.013 mg/L from the 1997
Thompson et al. study based on a MCCP material is divided by an assessment factor of
10 to yield 0.0013 mg/L or 1.3 (J,g/L or 1.3 ppb. Aquatic Chronic Concern
Concentration = 1 ppb
The most reliable and acceptable value for the toxicity to aquatic sediment invertebrate
organisms acute endpoint is based on the MCCP material from the Thompson et al. 2002 28-d
study. The 28-d sediment invertebrate GMATC value of 187 mg/kg dry wt sediment is used to
assess hazard. Again, using methods in US EPA (2012):
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Acute CoC: Calculating an acute concern concentration from the chronic value of 187
mg/kg dry wt. The 187 value is multiplied by an acute to chronic ratio for invertebrates
(10) to yield 1,870 mg/kg dry wt. This value is then divided by an assessment factor of 5
to yield 374 mg/kg dry wt. Aquatic Sediment Acute Concern Concentration = 374
mg/kg dry wt sediment.
Chronic CoC: The 28-d sediment invertebrate GMATC of 187 mg/kg dry wt sediment is
divided by an assessment factor of 10 to yield 18.7 mg/kg dry wt sediment. Aquatic
Sediment Chronic Concern Concentration = 19 mg/kg dry wt sediment.
The most reliable and acceptable value for the toxicity to terrestrial invertebrates acute endpoint
is based on the MCCP material from the Thompson et al. 2001a study. The 28-d terrestrial
invertebrate GMATC value of 149 mg/kg dry wt sediment from this study will be used.
Again, using methods in US EPA (2012):
Acute CoC: Calculating an acute concern concentration from the chronic value of 149
mg/kg dry wt, this value is multiplied by an acute to chronic ratio for invertebrates (10)
to yield 1,490 mg/kg dry wt. This value is then divided by an assessment factor of 5 to
yield 298 mg/kg dry wt. Terrestrial Invertebrate Acute Concern Concentration = 298
mg/kg dry wt sediment.
Chronic CoC: The 28-d terrestrial invertebrate GMATC of 149 mg/kg dry wt is divided
by an assessment factor of 10 to yield 14.9 mg/kg dry wt sediment. Terrestrial
Invertebrate Chronic Concern Concentration = 15 mg/kg dry wt sediment.
The most reliable and acceptable value for the toxicity to terrestrial plants is based on the MCCP
material from the Thompson et al. 2001ab study. For LCCPs, the analog approach using the
values from this study may be used. However, there is no OPPT guidance regarding assessing
concern concentrations for terrestrial plants.
B-2 LCCP ECQTQXICITY DATA
Where data are absent for long chain chlorinated paraffins for ecotoxicity endpoints, data from
sources using medium chain chlorinated paraffins will be used to inform the hazard for these
endpoints.
B-2-1 Acute Fish Toxicity
(1) Johnson and Finely 1980
A series of 96-hour acute fish toxicity studies were conducted by the United States Geological
Survey's Columbia National Fisheries Research Laboratory, over the years 1965 - 1978, with
several long-chain chlorinated paraffins (Chlorowax LV, C>i7, 39% CI; Chlorowax 40, C>20, 40 -
42% CI; Chlorowax 50, C>20, 48 - 54% CI; Chlorowax 70, C>20, 70% CI) published by Johnson
and Finely (1980). Bluegill sunfish (Lepomis macrochirus) and rainbow trout (Oncorhynchus
mykiss) were exposed to the test substance (100% commercial formulation) in a static test
system. Stock solutions were prepared immediately before each test with acetone used as a
carrier solvent. The average pH level was between 7.2 and 7.5 for all tests and the test
temperature was 20 ฑ 1ฐC for Bluegill sunfish and 10 ฑ 1 ฐC for Rainbow trout. Dilution water
hardness ranged from 40 to 50 mg/L as CaCCb. Reported effect levels are considered to be
nominal with LCso values of >300 mg/L for bluegill sunfish and rainbow trout for all LCCPs
tested; all values are above the limit of solubility.
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EPA/OPPT Conclusions
Using a weight-of-evidence approach, these studies were considered
acceptable to characterize the acute fish toxicity endpoint.
96-hr LCso = NES (> 300 mg/L)
(2) Bengtsson et al. 1979
A 96-hour acute fish toxicity study as part of a bioaccumulation study was published by
Bengtsson et al. (1979). Bleak (,Alburnus alburnus) were exposed to a long-chain chlorinated
paraffin, Witaclor 549 (Cis-26, 49% CI). The test was performed at 10 ฐC under semi-static
testing conditions. A nominal concentration of 0.125 mg/L was prepared by first dissolving
Witaclor in acetone and then added to the dilution water. The treatment vessels and acetone
control vessels consisted of six groups of 15 fish (average 4.5 g per fish). The acetone in all
treatments did not exceed 0.1 ml/L. The test solutions were renewed every two to three days over
the 14-day exposure period. Specific details of the test conditions were not provided other than
the tests were performed at 10ฐC in seawater with a salinity of 7 ppt. Even though the test
duration was 14-days, no mortality occurred within 96-hours or 14-days, thus the 96-hour LCso
was > 0.125 mg/L.
EPA/OPPT Conclusions
This study was considered supplemental to characterize the acute
saltwater fish toxicity endpoint using a weight-of-evidence approach.
96-hr LCso = NES (> 0.125 mg/L)
B-2-2 Acute Aquatic Invertebrate Toxicity
(1) Frank (1993) and Frank and Steinhauser (1994)
The following study summary (Frank, 1993; Frank and Steinhauser, 1994), provided in the ECB
(2005) MCCP risk assessment, was considered supportive of the aquatic invertebrate hazard
determination. The chlorinated paraffin used in these studies was a commercial C14-17 product
with a 52 wt% CI. Daphnia magna were exposed to nominal concentrations of either 100 mg/L
or 10,000 mg/L. The 100 mg/L solution was sonicated for 1 hour and then left to stand in the
dark for 48 hours before use. The 10,000 mg/L solution also stood for 48 hours in the dark
before use, but this time without sonication. After this period, both solutions were filtered firstly
with glass filters and then with membrane filters to remove undissolved test substance. The
concentrations of medium-chain chlorinated paraffin in the water soluble fractions were then
determined by AOX (adsorbable organic halogen) analysis (detection limit of 10 [j,g/L CI was
equivalent to around 20 [j,g/L of the chlorinated paraffin). This analysis showed that the
concentration of chlorinated paraffin present in the water soluble fraction was around 0.404 -
0.500 mg/L for the 10,000 mg/L nominal solution and 0.071 - 0.142 mg/L for the 100 mg/L
stock solution. The acute (48-hour) toxicity tests were carried out using dilutions of the two
prepared water soluble fractions. The method used was DIN 38 412, Teil 11, which is equivalent
to OECD 202.
In the tests using the water soluble fraction from the 100 mg/L nominal solutions no toxicity was
seen at concentrations up to the undiluted stock solution (i.e., no effects up to around 0.071-
0.142 mg/L). In experiments using the water soluble fraction from the 10,000 mg/L stock
solution, an ECo of 0.140 mg/L (also reported as 0.100 - 0.110 mg/L in the paper) and an EC25
of 0.423 mg/L (also reported as 0.420-0.470 mg/L in the paper) was determined (maximum
mortality seen was 25%) (Frank, 1993). The latter results for the 10,000 mg/L stock solution
were reported by Frank and Steinhauser (1994) as ECo = 0.140 mg/L and EC25 = 0.339 mg/L,
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and it was noted that some of the Daphnia were floating on the surface of the test solution. In the
later study (Frank and Steinhauser, 1994), the results of further acute toxicity studies were
reported using the same test method. An EC50 of 0.037 mg/L and an ECo of 0.009 mg/L were
determined using the water soluble fraction from the 100 mg/L stock solution and no toxic
effects were seen in tests with the water soluble fraction from the 10,000 mg/L stock solution
(approximately ECo > 0.525 mg/L). The authors noted that the effects seen in the acute tests
showed poor reproducibility, probably because effects were seen only around the water
solubility limit of the substance. However, the authors thought that the possibility of undissolved
droplets affecting the results could be ruled out, as floating Daphnia were only sporadically
observed in the test.
EPA/OPPT Conclusions
This study could not be adequately assessed due to inconsistencies in the
hazard data for the different dilutions of the test material. In addition, the
authors noted that the effects seen showed poor reproducibility, most
likely due to effects observed only around the solubility limit in the test
system used. The results of this test should therefore be treated with
caution, as the effects were mainly seen in the saturated solutions only.
(2) Tarkpeaetal. 1981
The Tarkpea et al. (1981); as quoted in IPCS (1996) summary, provided in the ECB (2005)
MCCP risk assessment, was considered supportive of the aquatic invertebrates hazard
determination.
The results of tests with the brackish water harpacticoid Nitocra spinipes have been reported
(Tarkpea et al., 1981). No other details of the test were reported but the test method was
probably the same as reported by Tarkpea et al. (1981), where a static method was employed
using water of salinity 7%o at a temperature of20 - 22ฐC without aeration, probably using
acetone as co-solvent.
EPA/OPPT Conclusions
This study could not be adequately assessed due to the lack of details
provided regarding specific conditions of the test and the preparation of
the test solutions. The lack of detail prohibits a review and adequacy of
the study. However, the harpaticoid, Nitocra spinipes, is not a standard
test species and the test concentrations greatly exceed the limit of
solubility. The 96-hour LC50 values for both chlorinated paraffins were
identified, but the reliability of the results cannot be determined. Overall,
EPA/OPPT reserves judgment on the acceptability of this study.
Therefore, data from sources using medium chain chlorinated paraffins
will be used to inform the hazard for these endpoints.
B-2-3 Aquatic Plant Toxicity
(1) Koh and Thiemann 2001
A 72-hour algae toxicity study using Scenedesmus subspicatus, was conducted at the University
of Bremen, Department of Physical and Environmental Chemistry with CP 30 (C17-24, 35%
chlorination), CP 40 (C17-20, 44% chlorination) and Hordaflex LC50 (C17-20, 52% chlorination)
according to DIN 38412 by Koh and Thiemann (2001). Additional communications with the
study author Wolfram Thiemann clarified that nominal test concentrations and local tap water
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(Bremen, Germany) was used without adjustments however, the study methods were not fully
characterized. The stock solutions were prepared to a concentration 6 mg/mL of test substance in
acetone. The individual stock solutions were diluted with distilled water to prepare individual
standard solutions of 0.250 mg/L for CP 30 and CP 40 and 0.125 mg/L for Hordaflex LC 50. pH
was between 5 and 6 and water hardness was between 35.7 and 53.5 mg CaCCb/L. Ambient
laboratory air temperature was ~ 21 ฐC. The solvent acetone was used to maintain test substance
in solution. Effects were calculated based on growth rate; no effects were observed up to 0.250
and 0.125 mg/L.
EPA/OPPT Conclusions
All three of the chlorinated paraffins tested contain C17 and Ci8
constituents which are considered to have LCCP-like properties. More
information concerning composition would be needed to accept these
results for long-chain chlorinated paraffins. Overall, EPA/OPPT reserves
judgment on the acceptability of this study.
72-hr ECso (growth rate) = NES (> 0.250 mg/L; CP 30 and CP 40)
72-hr EC50 (growth rate) = NES (> 0.125 mg/L; Hordaflex LC 50)
72-hr NOEC > 0.250 mg/L (CP 30 and CP 40)
72-hr NOEC > 0.125 mg/L (Hordaflex LC 50)
B-2-4 Chronic Fish Toxicity
No data are available.
B-2-5 Chronic Aquatic Invertebrate Toxicity
(1) Frank 1993 and Frank Steinhauser 1994
The following data review of Frank (1993) and Frank and Steinhauser (1994) were directly
excerpted from the U.K. Environmental Risk Assessment of long-chain chlorinated paraffins:
The test substance was identified as C18-20 chlorinated paraffin with 52% chlorination. Frank
(1993) carried out a series of acute and longer-term studies with Daphnia magna using a
commercial C18-20, 52% wt. CI product. The tests were carried out using dilutions of the
water-soluble fraction of the chlorinated paraffin. Stock solutions of the chlorinated paraffin
were made up in water to give nominal concentrations of either 100 mg/L or 10 g/L. The 100
mg/L solution was sonicated for one hour and then left to stand in the dark for 48 hours
before use. The 10 g/L solution also stood for 48 hours in the dark before use, but this time
without sonication. After this period, both solutions were filtered firstly with glass filters and
then with membrane filters to remove undissolved test material (microscopic and
spectroscopic investigation of the filtered solutions gave no indication of the presence of
droplets) to give the respective water-soluble fractions. The concentration of the chlorinated
paraffin in the water-soluble fractions was determined by AOX (adsorbable organic halogen)
analysis. The detection limit of the method used was around 10 jag Cl/L, which is equivalent
to around 20 [j,g/L of the chlorinated paraffin. This analysis showed that the concentration of
chlorinated paraffin present in the water-soluble fraction was around 462-519 [j,g/L for the
10 g/L nominal solution but was not detectable in the 100 mg/L solution (i.e. <20 (J,g/L).
Experiments were carried out to show that in the test vessels, although the concentration of
chlorinated paraffin present fell over time, it remained within 80 per cent of the initial
concentration over 2-3 days. This time period was used in the long term tests as the renewal
period for the solution (semi-static method).
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Long-term (21-day) reproduction studies were also performed using dilutions of the water-
soluble fractions of the two stock solutions. The dilutions used were 1:2, 1:4, 1:8 and 1:16
for the 100 mg/L loading and 1:4, 1:8, 1:16, 1:32 and 1:64 for the 10 g/L loading. In these
experiments, the test medium was changed three times per week and 10 animals were used
per concentration. The tests were carried out at 20ฐC and a pH of 7.79-8.44. Two endpoints
were determined in the study: effects on parent mortality and effects on reproduction
(number of offspring per adult). Parent mortality in the controls was 0 per cent in the test
carried out with the 10 g/L nominal stock solution and 10 per cent in the test carried out with
the 100 mg/L nominal stock solution.
Elevated mortality was seen in the exposed populations. For the 10 g/L stock solution the
LOEC was determined as the 1:8 dilution (approximately 58-65 (J,g/L) and the NOEC was
determined as the 1:16 dilution (approximately 29-32 (J,g/L). For the 100 mg/L nominal
stock solution the LOEC was determined as the 1:4 dilution and the NOEC was determined
as the 1:8 dilution. These dilutions are based on the detection limit for the analysis of the 100
mg/L stock solution, and equate to LOEC and NOEC [values] of <5 and <2.5 [j,g/L
respectively). From the dose response curves it appears that 100% parent mortality occurred
at a concentrations of around <10 [j.g/L in the 100 mg/L nominal stock solution experiments
and around 125 [j,g/L in the 10 g/L nominal stock solution experiments.
For the reproduction endpoint, the average number of young per adult in controls was 72.3 in
the 100 mg/L nominal stock solution series of experiments and 73.5 in the 10 g/L nominal
stock solution experiments. A significant reduction in the number of young per adult was
seen in some of the exposed organisms. For the 100 mg/L nominal stock solution, this effect
on reproduction was significantly different from the control groups at the lowest
concentration tested (a 1:16 dilution which is equivalent to a chlorinated paraffin
concentration of <1.2 (J,g/L, based on the detection limit of the analytical method used). Thus
the NOEC/LOEC for this series of experiments was <1.2 (J,g/L. Similarly, for the 10 g/L
nominal stock solution effects were again seen at the lowest concentration tested (a 1:64
dilution, which is equivalent to a chlorinated paraffin concentration of 7.3-8.1 (J,g/L). This
value is treated as the LOEC for this series of experiments. The report also indicates that the
NOEC is very close to this value, since using a different statistical method (Dunnett's Test
rather than Williams' Test), the effects seen at this concentration were not statistically
significantly different from controls.
EPA/OPPT Conclusions
The interpretation of the results is complicated by the difficulties
interpreting the effects from the different loading and dilution
concentrations used. The actual exposure concentration in the 100 mg/L
nominal stock solution is unknown and the measured concentration in the
10 g/L nominal stock solution (500 |ig/L) is above the reported
(estimated) water solubility of LCCPs. It was also noted in the U.K.
environmental risk assessment, that the data for the 10 g/L loading were
reanalyzed by Thompson (2001). The reanalysis suggests that the
statistical significance of parent mortality is questionable and that the 1:8
dilution (20% mortality) considered as the LOEC is a marginal effect at
best. The NOEC for parent mortality could be considered to be the 1:8
dilution (0.058 - 0.065 mg/L). In addition, there was a serious error found
by Thompson (2001) in the statistical method. They determined that the
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statistical software misinterpreted increasing dilutions as increasing
concentrations based on how the data were entered. Re-analysis of the
data showed that effects were statistically significant compared with the
controls only at the 1:8 dilution, leading to a NOEC at the 1:16 dilution
(0.029 - 0.032 mg/L). This problem with the statistical analysis may also
explain why the NOEC and LOEC from the 100 mg/L stock were
observed below the lowest concentration tested; thus, the NOEC/LOEC of
< 0.0012 mg/L is questionable. The data from these studies should be
approached with caution due to deficiencies and uncertainties with the
statistical analysis. The NOEC and LOEC values that follow are
questionable.
21-day NOEC = 0.029 - 0.032 mg/L (10 g/L solution; parent
mortality)
21-day LOEC = 0.058 - 0.065 mg/L (10 g/L solution; parent
mortality)
21-day NOEC < 0.0025 mg/L (below detection limit; 100 mg/L
solution; parent mortality)
21-day LOEC < 0.0050 mg/L (below detection limit; 100 mg/L
solution; parent mortality)
21-day NOEC < 0.0073 mg/L (10 mg/L solution; reproduction)
21-day LOEC < 0.0073 mg/L (10 mg/L solution; reproduction)
21-day NOEC < 0.0012 mg/L (100 mg/L solution; reproduction)
21-day LOEC < 0.0012 mg/L (100 mg/L solution; reproduction)
(2) TNO 1993
The test was carried out according to the OECD [211] methodology using a semi-static test
procedure (test solution renewal was carried out every 48-72 hours). The test substance was
identified as C18-20 chlorinated paraffin with 52% chlorination (Chloroparaffin Hoechst 56
Fliissig, Chloroparaffin Hoechst 52 Fliissig, and Hordaflex LC50. The dilution water used was a
synthetic medium (DSWL) prepared by the addition of various salts to ground water. The
hardness of the medium was 214 mg/L as CaC03. The test was carried out using saturated
solutions of the chlorinated paraffin using a column technique. The column was prepared by
firstly dissolving/suspending 0.1 g of the test substance in 25 mL of acetone. This solution was
then added to 10 g of the packing material for the column (chromosorb 60/80 mesh) and the
acetone removed by rotation evaporation. The coated packing material was stored at room
temperature in the dark until needed. The columns were stainless steel (25 cm long with an
internal diameter of 4.3 mm) filled with 1 g of the coated packing material. The column was
conditioned by pumping dilution water through at a flow rate of 6.2 mL every three minutes; the
first 500 mL was collected and discarded. Around 18 litres of dilution water was then collected
in a bottle and continually re-circulated through the column at a flow rate of 3.4 mL per minute
throughout the test. The required amount of the saturated solution needed for the start of the test,
and at each renewal period, was then taken from the bottle.
Four replicates (10 daphnids in 400 mL of test solution) were carried out for each treatment. The
tests were performed in 600 mL beakers and these were conditioned to the test solutions for two
days prior to the start of the test. The solutions were renewed every Monday, Wednesday and
Friday during the test.
The concentration of test substance was determined at each renewal time in both the "fresh"
solution and the "spent" solution. The analytical method used was based on extractable organic
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halogen (EOX; similar in principle to AOX) analysis. The mean EOX measured in the test
solution over the course of the test was around 1 [j,g/L (the range found in the "fresh" solutions
was 1.0-1.5 [j,g/L and the range in the "spent" solutions was 0.5-1.5 (J,g/L. The EOX
concentrations in the control solutions were generally <0.5 [j,g/L in the "fresh" solution but the
range found in the spent solutions was <0.5-1.0 [j,g/L for the blank control and 0.5 2.0 [j,g/L for
the column control.
The temperature, DO, and pH during the test were 19.2 - 20.3 ฐC, >7.1 mg/L and 7.6 - 8.5,
respectively. The parent survival in both of the control groups was 97.5%. In the C18-20, 52% wt.
CI treatment group the parent survival was 92.5%. This survival was not significantly different
(at the p=0.05 level) from the control group. Therefore, it was concluded that no treatment
related effects on parent morality were seen in the study. The reproduction rate (expressed as the
cumulative number of young per living female) in the study was 113.6 in the blank control and
127.3 in the column control. The response of the two controls was not significantly different (at
the p=0.05 level). The reproduction rate in the C18-20, 52% wt. CI treatment group was 100.8.
This was 88.7 per cent of the blank control response and 79.1 per cent of the column control
response. These responses were analyzed statistically in TNO (1993) using the two-tailed
Dunnett-test. No statistically significant differences were found between the C18-20, 52% wt. CI
treatment group and the blank control, or the column control.
EPA/OPPT Conclusions
Thompson (2007) conducted a further analysis of the results, due to the
author's perception of discrepancies in the UK's (2009) Environmental
Risk Assessment. The discrepancy was based on a seemingly
contradictory statement from pg. 7 of the TNO report where the
reproduction rate of the Hodaflex LC50 and Hoechst 52 Fliissig was stated
as being significantly different from the controls (Pg.7 of the report) but
not significantly different from the controls (Pg. 16 of the report). The
data were re-analyzed by Thompson (2007) and no differences in the
reproduction rate were observed between the C18-20, 52 wt% chlorination
treatment groups Hodaflex LC50 and Hoechst 52 Fliissig, the column
control, blank control or the pooled control group (combined blank control
and column control). An additional statistical analysis was conducted by
OPPT and no differences were determined for reproduction rate in the
treatment groups versus controls using SAS (v. 9.3). Regardless of the
discrepancy in the TNO report, statistically significant differences were
reported by Thompson (2007) for adult mortality, reproduction rate, and
mortality in general for Daphnia exposed to Hoechst 52 Fliissig compared
to both controls. However, it is still unclear as to the levels of saturation of
each solution the organisms were exposed to based on the extraction
technique and nominal loading rates. Therefore, EPA/OPPT will reserve
judgment on the acceptability of this study based on a weight of evidence
approach.
21-day NOEC = 0.002 mg/L (reproduction Hoechst 56 Fliissig)
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B-2 -6 Chronic Aquatic Sediment Invertebrate Toxicity
No data are available. Data from secondary sources using medium-chain chlorinated paraffins
for chronic aquatic sediment invertebrate toxicity can be used to fill data gaps for the long-chain
chlorinated paraffin; and these data were described above in the MCCP section.
B-2-7 Avian Toxicity
No data are available. Data from secondary sources using medium-chain chlorinated paraffins
for avian toxicity can be used to fill data gaps for the long-chain chlorinated paraffins; and these
data were described above in the MCCP section.
B-2-8 Terrestrial Invertebrate Toxicity
No data are available. Data from secondary sources using medium-chain chlorinated paraffins
for terrestrial invertebrate toxicity can be used to fill data gaps for the long-chain chlorinated
paraffins; and these data were described above in the MCCP section.
B-2-9 Terrestrial Plant Toxicity
No data are available. Data from secondary sources using medium-chain chlorinated paraffins
for terrestrial plant toxicity can be used to fill data gaps for the long-chain chlorinated paraffins;
and these data were described above in the MCCP section.
B-2-10 Conclusions
Sufficient data were available to characterize the acute fish, the acute aquatic invertebrate, the
chronic aquatic invertebrate, the chronic aquatic sediment invertebrate, avian, and terrestrial
plant toxicity endpoints for MCCPs and LCCPs by read across in some instances. Data for other
toxicity endpoints (i.e., chronic fish, aquatic plant, etc.) were inconclusive due to lack of study
details, uncertainties in analytical methods, or test material preparation methods; thus, these data
are included in order to characterize risk in a qualitative manner, but are used as supportive for
the categories under which they are provided. Supporting data were included in order to provide
a weight-of-evidence approach used to characterize some endpoints.
Most of the data provided in this review indicated several difficulties were encountered when
testing in an aquatic environment. These included: (1) getting the material into solution, (2)
measuring the material in solution, and (3) characterizing the effects for each study listed. Often
there were many details of a given study omitted, prohibiting a full and robust review of the data.
The (estimated) physical-chemical properties of LCCPs and MCCPs (water solubility values
between 2 and 30 [j.g/L and Log Kowvalues between 4-8) suggest these materials may not
partition to the aquatic media or elicit toxicity to aquatic organisms within the water column.
Specifically for the chronic and acute aquatic invertebrate, aquatic sediment, avian, and
terrestrial plant endpoints for LCCPs, other analog data provided was acceptable using
compounds with chlorination percentage of 52 wt% and carbon chain lengths of C14-17 which is
defined as a MCCP material. These data are used in this assessment to fill data gaps for the C18-20
LCCPs as this would be a conservative approach to charactering hazard in the absence of data.
Concern concentrations based on these data are again a very conservative approach in the
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absence of data for the LCCP materials themselves and therefore may not inherently characterize
toxicity to LCCPs directly.
The most reliable and acceptable studies indicate that for MCCPs, the toxicity to aquatic
organisms for acute endpoints are from the Thompson et al. 1996 study for aquatic invertebrates.
Where the 48-hour EC50 value = 0.0059 mg/L. Using the methods described in the Sustainable
Futures/P2 Manual (US EPA, 2012), the acute and chronic concentrations of concern (CoC) are
derived as follows:
Acute CoC: The 48-hour EC50 value = 0.0059 mg/L is divided by an assessment factor of
5 to yield an acute concentration of concern (CoC) of 0.00118 mg/L, or 0.001 mg/L, or 1
[j,g/L (1 ppb). Aquatic Acute Concern Concentration= 1 ppb
Chronic CoC: The aquatic invertebrate chronic value of 0.013 mg/L from the 1997
Thompson et al. study based on a MCCP material is divided by an assessment factor of
10 to yield 0.0013 mg/L or 1.3 (J,g/L or 1.3 ppb. Aquatic Chronic Concern
Concentration = 1 ppb
For LCCPs, the acute concern concentration may be derived from the Johnson and Finely (1980)
studies. For the chronic concern concentration, the results from the Thompson et al. 1997 study
based on a MCCP material will be used as a conservative qualitative assessment due to the lack
of overly reliable data for this endpoint for LCCPs.
Acute CoC: Aquatic Acute Concern Concentration= NES
Chronic CoC: The aquatic invertebrate chronic value of 0.013 mg/L from the 1997
Thompson et al. study based on a MCCP material is divided by an assessment factor of
10 to yield 0.0013 mg/L or 1.3 (J,g/L or 1.3 ppb. Aquatic Chronic Concern
Concentration = 1 ppb (MCCP and LCCP)
The most reliable and acceptable value for the toxicity to aquatic sediment invertebrate
organisms acute endpoint is based on the MCCP material from the Thompson et al. 2002 28-d
study. For both MCCPs and LCCPs, the 28-d sediment invertebrate GMATC value of 187
mg/kg dry wt sediment is used to assess hazard. The 28-d sediment invertebrate GMATC value
of 187 mg/kg dry wt sediment is used to assess hazard. Again, using methods in US EPA (2012):
Acute CoC: Calculating an acute concern concentration from the chronic value of 187
mg/kg dry wt. The 187 value is multiplied by an acute to chronic ratio for invertebrates
(10) to yield 1,870 mg/kg dry wt. This value is then divided by an assessment factor of 5
to yield 374 mg/kg dry wt. Aquatic Sediment Acute Concern Concentration = 374
mg/kg dry wt sediment. (MCCP and LCCP)
Chronic CoC: The 28-d sediment invertebrate GMATC of 187 mg/kg dry wt sediment is
divided by an assessment factor of 10 to yield 18.7 mg/kg dry wt sediment. Aquatic
Sediment Chronic Concern Concentration = 19 mg/kg dry wt sediment. (MCCP and
LCCP)
The most reliable and acceptable value for the toxicity to terrestrial invertebrates acute endpoint
is based on the MCCP material from the Thompson et al. 2001a study. For LCCPs, the 28-d
terrestrial invertebrate GMATC value of 149 mg/kg dry wt sediment will be used as an analog
approach to assess hazard. The 28-d terrestrial invertebrate GMATC value of 149 mg/kg dry wt
sediment from this study will be used. Again, using methods in US EPA (2012):
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Acute CoC: Calculating an acute concern concentration from the chronic value of 149
mg/kg dry wt, this value is multiplied by an acute to chronic ratio for invertebrates (10)
to yield 1,490 mg/kg dry wt. This value is then divided by an assessment factor of 5 to
yield 298 mg/kg dry wt. Terrestrial Invertebrate Acute Concern Concentration = 298
mg/kg dry wt sediment. (MCCP and LCCP)
Chronic CoC: The 28-d terrestrial invertebrate GMATC of 149 mg/kg dry wt is divided
by an assessment factor of 10 to yield 14.9 mg/kg dry wt sediment. Terrestrial
Invertebrate Chronic Concern Concentration = 15 mg/kg dry wt sediment. (MCCP
and LCCP)
The most reliable and acceptable value for the toxicity to terrestrial plants is based on the MCCP
material from the Thompson et al. 2001ab study. For LCCPs, the analog approach using the
values from this study may be used. However, there is no OPPT guidance regarding assessing
concern concentrations for terrestrial plants.
Appendix C HUMAN HEALTH HAZARD STUDY
SUMMARIES
C-l MCCP HEALTH DATA REVIEW
There is no information on inhalation absorption of MCCPs in humans or in animals. Based on
their low vapor pressure and low water solubility, absorption following inhalation or dermal
exposure is expected to be limited. Some MCCPs demonstrated moderate absorption and
metabolism following oral exposure in animals. In general, absorption and metabolism are related
to their carbon chain length and degree of chlorination; the longer the carbon chain length and the
higher the degree of chlorination, the less absorption and metabolism.
No information is available on the toxicity of MCCPs in humans; however, the toxicology of
these compounds has been evaluated in experimental animals. Studies in rats and rabbits have
shown that MCCPs only caused slight skin irritation and have low eye irritation potential. No
evidence of skin sensitization was found when tested in guinea pigs. The liver, kidney and
thyroid are the target organs of MCCPs in oral repeated dose studies in experimental animals.
MCCPs induced increased liver weight, enzyme activity, and histopathological changes at high
dose levels. Some of these hepatic effects are likely related to an increase in metabolic demand
as an adaptive response, as well as to peroxisome proliferation, which are considered of limited
toxicological significant to humans. However, liver necrosis was observed in a 90-day study in
rats at 360 mg/kg-bw/day; this effect is considered relevant to humans. The reported effects in
the kidney may have been produced by the parent compound or from metabolites. Mechanistic
data cannot totally rule out that some kidney effects are relevant to humans. From the data
available, a LOAEL of 625 mg/kg-bw/day based on histopathological changes in the kidneys of
female rats is identified in a 90-day toxicity study, and a NOAEL of 23 mg/kg-bw/day based on
increased kidney weight at 222 mg/kg-bw/day is identified from another 90-day study in rats.
Repeated dose studies in rats reported some changes in histopathology and hormone levels of the
thyroid. However, it may be concluded based on an evaluation of the mechanistic data that the
thyroid effects observed in rats is of little relevance to chronic toxicity in humans.
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There is no information on the carcinogenicity of MCCPs; however, carcinogenicity studies on a
short-chain chlorinated paraffin (SCCP) and a long-chain chlorinated paraffin (LCCP) are
available. These studies, along with the genotoxicity data on MCCPs, may be used to inform the
carcinogenic potential of MCCPs. When administered by gavage, a SCCP (C12, 60 wt% CI)
caused increased incidences of liver tumors in male and female rats, kidney tumors in male rats,
and thyroid tumors in female rats. However, based on mechanistic considerations, these tumors
are considered to be of little or no relevance to humans. An increased incidence of malignant
lymphoma in male mice was reported at the highest dose of 5,000 mg/kg-bw/day in
carcinogenicity studies of a LCCP (C23, 43 wt% CI) in male and female rats and mice. However,
malignant lymphoma is one of the more variable tumors in mice and has a viral origin in many
cases. No increased incidence of malignant lymphoma was observed in the carcinogenicity study
on an SCCP. Further, MCCPs are non-genotoxic. Therefore, it may be concluded that MCCPs
are unlikely to pose a carcinogenic hazard to humans.
When evaluating the risks of workers from exposure to MCCPs based on the available repeated-
dose toxicology studies, the EU's draft Risk Assessment Report (RAR) on MCCPs concluded
that except metal working fluids (MWF) use, "There is at present no needfor further
information and/or testing or for risk reduction measures beyond those which are being applied
already' (EURAR, 2008).
Using the NOAEL (23 mg/kg-bw/day) of kidney toxicity identified in the 90-day oral study in
rats (CXR, 2005), aMOS of 70,000 was estimated for dermal exposure of consumers resulting
from wearing leather clothes treated with MCCP. For inhalation exposure of consumers using
metal fluids containing MCCP, a MOS of 2,875 was obtained. Therefore, it was also concluded
that: "There is present no needfor further information and/or testing or for risk reduction
measures beyond those which are being applied already" (EURAR, 2008).
C-l-1 Metabolism
There is no information on inhalation absorption of MCCPs in humans or in animals. Based on
their low vapor pressure and low water solubility, absorption following inhalation or dermal
exposure is expected to be limited. An in vitro study using human skin showed that after 24
hours, approximately 0.7% of a C15 chlorinated paraffin was absorbed (Scott, 1984; cited in:
EURAR, 2008). Oral studies (IRDC, 1984, CXR, 2005; cited in: EURAR, 2008) showed that
approximately 50% of a single dose of [8-14C]-labeled C15 chlorinated paraffin (52 wt% CI) was
absorbed from the GI tract in rats. Excretion via feces was the major route of elimination of
radiolabeled material. Elimination of radioactivity from body tissues occurred with an
elimination half-life of approximately 2-5 days (liver and kidney) or approximately 2 weeks
(adipose tissue).
C-l-2 Acute Toxicity
There is no information on the effects of a single exposure to MCCPs in humans. No deaths and
only limited, non-specific clinical signs of toxicity resulting from exposure of rats to very high
doses were observed in an acute oral toxicity study of MCCPs (C14-17,51-60 wt% CI); the LD50
was reported to be > 4,000 mg/kg bw (Birtley et al., 1980; cited in: IPCS, 1996). Though no
acute toxicity data are available for MCCP by the inhalation or dermal route of exposure, the low
acute toxicity data for SCCPs by these routes suggest that MCCPs are likely to have low acute
inhalation and dermal toxicity.
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C-l-3 Irritation and Sensitization
No signs of skin irritation were seen with MCCPs (C14-17, 45 wt% CI), and only slight erythema
on the shaved skin was reported in one rabbit at 24 hours exposed to MCCPs (C14-17, 40 wt% CI)
(Chater, 1978; cited in: EURAR, 2008). A mild skin irritancy response was reported in one of
nine unpublished skin irritation studies of MCCPs (C14-17, 51-60% CI) in rats (Birtley el al.,
1980; cited in: EURAR, 2008). The material caused slight, transient eye irritation in rabbits
(Birtley et al., 1980; Kuhnert et al., 1986; cited in: EURAR, 2008).
No skin sensitization reactions were produced in guinea pig maximization tests conducted on
MCCPs (C14-17, 40-45 wt% CI) (Murmann, 1988; Chater, 1978; cited in: EURAR, 2008).
C-l-4 Repeated-dose Toxicity
There are a number of repeated dose toxicity studies (up to 90-days duration) of MCCPs (C14-17
40 wt% CI or 52 wt% CI) in rats by oral exposure (CXR, 2005; Poon et al., 1995; IRDC, 1984;
Birtley et al., 1980; and Wyatt et al., 1997; cited in: EURAR, 2008). Though the quality and
reliability of these studies differs, the liver, kidney, and thyroid were consistently established as
the target organs. A summary of the results from these studies is provided in Table 1.
MCCPs caused an increase in liver weight in male rats at exposure levels of > 100 mg/kg-
bw/day) and in female rats at exposure levels of > 32 mg/kg-bw/day. Liver enzyme induction
was reported in male and female rats starting from 222 and 100 mg/kg-bw/day, respectively.
Liver hypertrophy of trace to minimal severity was reported in male rats at dose levels of > 100
mg/kg-bw/day and higher. Collectively, these changes are likely to be related to an increase in
metabolic demand as an adaptive response and to peroxisome proliferation, both of which are
considered of no or limited toxicological significance to humans. Though Poon et al. (1995)
reported various histopathological effects in the liver of male and female rats at dose levels > 36
mg/kg-bw/day, there are a number of deficiencies with this study, including the scoring and
classification of histopathological findings and limited reporting of data, which preclude its
utility in hazard evaluation. This conclusion is consistent with previous evaluations of this study
(EURAR, 2008). Further, despite the consistency of findings reported in the review article by
Birtley et al. (1980) with other 90-day studies, these findings should be viewed cautiously
because the original full study report is not available. Based on the available data, the studies by
IRDC (1984) and CXR (2005) provide the most reliable data for identifying effect levels of
MCCPs on the liver. For the purposes of this assessment, a NOAEL of 100 mg/kg-bw/day was
chosen based on increases in absolute liver weight (i.e., 22-26%), liver hypertrophy of trace
severity, and enzyme induction (i.e., 30% increase).
Kidney effects have been reported in a number of studies, with effect levels typically being
observed at the limit doses. MCCPs (C14-17 52 wt% CI) caused significant increases (9-13%) in
kidney weight at 222 mg/kg-bw/day (CXR, 2005; cited in: EURAR, 2008), as well as "chronic
nephritis" and tubular pigmentation in the kidney of female rats at 625 mg/kg-bw/day (IRDC,
1984, cited in: EURAR, 2008). One study reported a dose-related increase in congestion starting
at 32 mg/kg-bw/day; however, no information was provided on the incidence or severity of this
effect (Birtley et al., 1980; cited in: EURAR, 2008). An additional study reported minimal to
mild hyaline-droplet like cytoplasmic inclusions, starting at > 0.4 mg/kg-bw/day in male rats.
This effect is considered of limited relevance to humans. The authors also reported minimal
dose-related increases with inner medullary tubular dilation at an incidence of 1/10, 4/10, and
8/10 female rats at 4, 42, and 420 mg/kg-bw/day, respectively (Poon et al., 1995; cited in:
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EURAR, 2008). Though this effect is considered relevant to humans, the study suffers from a
number of limitations, which preclude utilizing it for hazard evaluation. However, based on the
incidence reported by the authors, the NOAEL of 42 mg/kg-bw/day for kidney effects is
consistent with the NOAEL of 23 mg/kg-bw/day reported in the CXR (2005) study. Therefore, a
NOAEL of 23 mg/kg-bw/day was chosen for the kidney, based on increases in organ weight at
the next highest dose level.
MCCPs (C14-17 52 wt% CI) have been reported to cause minimal to mild adaptive
histopathological changes in the thyroid {i.e., follicular cell hypertrophy and hyperplasia) in two
studies in rats starting at 50 ppm (4 mg/kg-bw/day) and above (Poon et al., 1995; IRDC, 1985).
Decreased T4 levels and increased TSH levels in the plasma were also seen at similar dose
levels. As noted previously, these results have been drawn into question based on the scoring and
classification for histopathology, the limited reporting of data, and the inconsistent findings from
other more robust studies (EURAR, 2008). Therefore, these studies will not be considered
further for hazard identification. IRDC (1985) reported mild to moderate hypertrophy and
hyperplasia in male rats at dose levels of > 10 mg/kg-bw/day and higher, whereas changes in
absolute organ weights of male and female rats were not observed except at the limit dose of 625
mg/kg-bw/day (IRDC, 1985; cited in: EURAR, 2008). The remaining studies that evaluated
thyroid hormone levels identified a decrease in plasma free T3 in male rats, but not total T3 or
free/total T4, and an increase in TSH in female rats at dose levels of 24.6 or 242 mg/kg-bw/day,
respectively (CXR, 2005; cited in: EURAR, 2008), or fluctuations in thyroid hormones in male
or female rats at doses of > 312 mg/kg-bw/day or higher (Wyatt et al., 1997; cited in: EURAR,
2008). There is evidence that the thyroid effects observed are attributable to stimulation of this
organ arising from a negative feedback effect arising from plasma T4 depletion following
increased excretion of this hormone. This depletion of plasma T4 results from the induction of
hepatic UDPG-transferase, increased glucuronidation, and ultimately excretion of T4 following
exposure to MCCPs. The pituitary responds to the decreased levels of T4 by releasing more TSH,
which in turn leads to increased production of T4 by the thyroid. The continuous stimulation of
the thyroid in response to the increased excretion of plasma T4 is predicted to ultimately give rise
to hypertrophy and hyperplasia in this organ. Humans, unlike rodents, possess T4-globulin
binding protein and are therefore less susceptible to plasma T4 depletion and hence any resultant
thyroid stimulation. The thyroid effects observed in rats are not considered to be of relevance to
chronic human health at relevant levels of exposure, although these changes may be relevant for
assessing potential adverse outcomes during reproduction and development, as discussed under
section.
C-l-5 Genotoxicity
MCCPs (C14-17 40-52 wt% CI) are not mutagenic to bacteria. Three in vivo bone marrow studies
also show that MCCPs are not clastogenic {cited in: EURAR, 2008). Therefore, it may be
concluded that MCCPs possess a low potential to cause genotoxic effects.
C-l-6 Carcinogenicity
There is no information on the carcinogenicity of MCCPs. When administered by gavage, a
SCCP (C12, 60 wt% CI) caused increased incidences liver tumors in male and female rats, kidney
tumors in male rats, and thyroid tumors in female rats (NTP, 1986). However, on mechanistic
considerations, these tumors are considered to be of little or no relevance to humans. This
conclusion is consistent with previous carcinogenicity evaluations (EURAR, 2008). An
increased incidence of malignant lymphoma in male mice was reported at the highest dose of
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5,000 mg/kg-bw/day in carcinogenicity studies of a LCCP (C23, 43 wt% CI) in male and female
rats and mice. However, malignant lymphoma is one of the more variable tumors in mice and
has a viral origin in many cases. No increased incidence of malignant lymphoma was observed
in the carcinogenicity study on a SCCP (C12, 60 wt% CI). Based on structure-activity
relationships and the absence of positive genotoxicity data on MCCPs, the available
carcinogenicity studies on a SCCP and a LCCP suggest that MCCPs are not expected to pose a
carcinogenic hazard to humans.
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Table Apx C-l: Summary of Results from 90-day Studies in Rats Administered MCCPs
Strain (sample size)
Test substance and
dose levels
Target
organ
Effect levels
3 at 222 and $ at 242 mg/kg-bw/day, 13-31% t in organ weight
3 at 222 mg/kg-bw/day, minimal centrilobular hypertrophy in 9/10 animals
Liver
(51 at 222 mg/kg-bw/day, 82% f in microsomal T4-UDPGA-glucuronyl transferase activity
Cm-17, 52 wt% CI
$ at 100, 300, and 300 mg/kg-bw/day, 30, 30, and 252% f in microsomal T4-UDPGA-glucuronyl transferase activity,
respectively
F-344
(10 rats/sex/group)1
Dietary intake for '': 0,
2.38,9.34, 23.0, or 222
mg/kg-bw/day.
Kidney
3 at 222 and $ at 242 mg/kg-bw/day, 9-13% j in organ weight
3 at > 222 and $ at 242 mg/kg-bw/day, no treatment-related histopathology
Dietary intake for $: 0,
2.51,9.70, 24.6, or 242
mg/kg-bw/day.
Thyroid
3 at 222 mg/kg-bw/day, 17% | in plasma TSH
(51 at 23.0 and 222 mg/kg-bw/day, 26% or 22% j in plasma free T3, respectively, but no effects on total T3 or on free/total
T4 at any dose
$ at > 242 mg/kg-bw/day, no effects on free/total T3 or T4
at 24.6 and 242 mg/kg-bw/day, 20 and 39% f in plasma TSH
at > 222 and $ at 242 mg/kg-bw/day, no treatment-related histopathology
3 at 360 and $ at 420 mg/kg-bw/day, 28 and 48% f in absolute and relative weights, respectively
3 and at ฃ 4 mg/kg-bw/day, no treatment-related histopathology
Ci4-n, 52 wt% CI
'at 36 and $ at 42 mg/kg-bw/day, minimal increase in anisokaryosis and vesiculation of the nuclei
Sprague-Dawley
(10 rats/sex/group)2
Dietary intake for '': 0,
0.4, 4, 36, or 360
mg/kg-bw/day.
Dietary intake for $: 0,
0.4, 4, 42, or 420
mg/kg-bw/day.
Liver
3 at 360 and $ at 420 mg/kg-bw/day, mild increase in anisokaryosis and vesiculation of the nuclei (7-10 animals)
3 at 360 mg/kg-bw/day, t in perivenous homogeneity
$ at 42 and 420 mg/kg-bw/day, t in perivenous homogeneity
(51 at 360 and $ at 420 mg/kg-bw/day, t in single cell necrosis (incidence not reported)
Kidney
3 at 360 and $ at 420 mg/kg-bw/day, 11% t in absolute and relative weights
3 at > 0.4 mg/kg-bw/day, minimal to mild hyaline-droplet like cytoplasmic inclusions, with significant accumulation at
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Table Apx C-l: Summary of Results from 90-day Studies in Rats Administered MCCPs
Strain (sample size)
Test substance and
dose levels
Target
organ
Effect levels
the limit dose
at > 4 mg/kg-bw/day, minimal dose-related inner medullary tubular dilation seen in 0/10, 0/10, 1/10, 4/10, and 8/10
animals
Thyroid
3 at 36 and 360 mg/kg-bw/day, minimal to mild morphological changes affecting the architecture (i.e., reduced follicle
sizes and collapsed angularity) and the epithelium (i.e., increased height, cytoplasmic vacuolation, and nuclear
vesiculation)
at > 4 mg/kg-bw/day, minimal to mild morphological changes affecting the architecture (i.e., reduced follicle sizes and
collapsed angularity) and the epithelium (i.e., increased height, cytoplasmic vacuolation, and nuclear vesiculation)
F-344
(15 rats/sex/group)3
Ci4-i7, 52 wt% CI
Dietary intake for (51
and $: 0, 10, 100, or
625 mg/kg-bw/day.
Liver
(51 and $ at 100 and 625 mg/kg-bw/day, 22-26% and 64-92% f in absolute weight values, respectively
(51 at 100 and 625 mg/kg-bw/day, hypertrophy of trace severity seen in 1/15 and 13/15 animals, respectively
$ at 625 mg/kg-bw/day, hypertrophy of trace severity seen in 13/15 animals
Kidney
3 and $ at 625 mg/kg-bw/day, 18% f in absolute weight values
3 at > 10 mg/kg-bw/day, trace to mild nephritis seen in 1/15, 3/15, 4/15, and 10/15 animals
$ at 625 mg/kg-bw/day, tubular pigmentation (9/14 animals)
Thyroid
S at 625 mg/kg-bw/day, 50% f in absolute weight values
(51 at > 10 mg/kg-bw/day, mild to moderate hypertrophy observed in controls with a dose-dependent trend towards t
severity in treated animals
3 at > 10 mg/kg-bw/day, trace to mild hyperplasia with a dose-dependent trend towards t severity
Adrenal
(51 and $ at 625 mg/kg-bw/day, 25% f in absolute weight values
Wistar-derived
(24 rats/sex/group)4
Cm-17, 52 wt% CI,
containing epoxidized
vegetable oil as a
stabilizer
Dietary intake for '': 0,
33, 167, or 333 mg/kg-
bw/day.
Liver
3 at 167 and 333 mg/kg-bw/day, 15 and 22% f in relative weight values, respectively
$ at 32, 160, and 320 mg/kg-bw/day, 11, 21, and 48% f in relative weight values, respectively
S at 333 and $ at 320 mg/kg-bw/day, no histopathological abnormalities
3 at > 33 and at > 32 mg/kg-bw/day, dose-related t in proliferation of smooth endoplasmic reticulum (electron
microscopy)
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Table Apx C-l: Summary of Results from 90-day Studies in Rats Administered MCCPs
Strain (sample size)
Test substance and
dose levels
Target
organ
Effect levels
Dietary intake for $: 0,
32, 160, or 320 mg/kg-
bw/day.
Kidney
3 and $ at limit dose, 15% f in relative weight
3 and . dose-related t in congestion (incidence and severity not reported)
(51 and $ at > limit dose, no histopathological abnormalities
3 and $ at 312 and 625 mg/kg-bw/day, 37 and 72% f in relative weight, respectively, (absolute weight and bodyweight
not presented)
Liver
3 and $, dose-related t in centrilobular hypertrophy (incidence and severity not reported)
3 and $ at 312 and 625 mg/kg-bw/day, dose-related t in (3-oxidation from day 29 onwards (-2.1- and 3.3-fold t,
respectively, at study termination)
C14-17, 40 wt% CI
3 and $ at 312 and 625 mg/kg-bw/day, dose-related t in UDPG-transferase activity from day 15 onwards (up to 100% f,
respectively)
F-344
(10 rats/sex/group)5
Oral gavage for (51 and
?: 0,312, or 625
mg/kg-bw/day
Thyroid
S and $ at 312 and 625 mg/kg-bw/day, | in levels of free and plasma T3, which reached statistical significance on days
15 and 57
3 at 312 and 625 mg/kg-bw/day, t TSH up to 2-fold on day 8 only
$ at 312 and 625 mg/kg-bw/day, T3 significantly t by day 91
$ at 312 and 625 mg/kg-bw/day, total plasma T4 significantly | by up to 25% on day 57
S and $ at 312 and 625 mg/kg-bw/day, t follicular cell hypertrophy throughout the study, and accompanied by follicular
cell hyperplasia on days 55 and 91 (incidence and severity not reported)
3 and $ at 312 and 625 mg/kg-bw/day, significantly t replicative DNA synthesis on day 29, but not on day 91
1 CXR (2005), cited in: EURAR (2008).
2 Poon et al. (1995), cited in: EURAR (2008).
3IRDC (1984), cited in: EURAR (2008).
4 Birtley et al. (1980), cited in: EURAR (2008); note, this study was only summarized in the review by Birtley et al. (1980). The underlying original study report was not available.
5 Wyatt et al. (1997), cited in: EURAR (2008).
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C-l-7 Developmental Reproductive Toxicity Review
A series of range-finding and definitive prenatal developmental and reproductive toxicity studies
were conducted in rats and rabbits with medium-chain chlorinated paraffins (MCCPs). These
studies were conducted between 1981 and 1986. They appear to be valid toxicity studies,
conducted according to the standard methodologies available at the time. More recently,
additional studies with MCCPs have been conducted in an attempt to determine the cause of
hemorrhaging in the pups observed in a one-generation reproductive toxicity range-finding
study.
In several prenatal developmental toxicity studies with MCCPs conducted via gavage, no signs
of maternal toxicity were seen at doses as high as 500 mg/kg-bw/day in rats and 100 mg/kg-
bw/day in rabbits. Likewise, no signs of developmental toxicity were observed at doses as high
as 5000 mg/kg-bw/day in rats and 100 mg/kg-bw/day in rabbits.
Two reproductive toxicity studies with MCCPs in rats have been conducted. A one-generation
reproductive toxicity range-finding study showed that administration of approximately 100 and
400 mg/kg-bw/day MCCPs via the diet had no effect on fertility or other reproductive
parameters; however, internal hemorrhaging and deaths in pups were observed beginning from
74 mg/kg-bw/day (1000 ppm) up to approximately 400 mg/kg-bw/day (6250 ppm). These effects
in the pups were not seen in a more recent definitive one-generation reproductive toxicity study
with exposure to MCCPs for 11-12 weeks to doses as high as 100 mg/kg-bw/day (1200 ppm).
Internal hemorrhaging was not seen in the adult animals in either of these studies at doses as
high as 400 mg/kg-bw/day (6250 ppm), or in another study in non-pregnant female rats
repeatedly exposed to doses as high as 1000 mg/kg-bw/day. However, when dams were exposed
to approximately 500 mg/kg-bw/day (6250 ppm) MCCPs during cohabitation, gestation, and
lactation, signs of hemorrhaging were observed in dams that died at the time of parturition.
Taken together, the results of these studies suggest that newborns during lactation and pregnant
females at the time of parturition are a potentially sensitive subpopulation.
The UK Risk Assessment (February, 2008) did not use the LOAEL of 74 mg/kg-bw/day (1000
ppm) from the one-generation reproductive toxicity range-finder study as a point of departure
because the pup deaths at that dose were not statistically significant. The study itself used a
limited number of animals and was intended for dose range-finding purposes only and, more
importantly, the pup deaths were not repeated in a more recently conducted definitive study.
With respect to developmental/reproductive toxicity, the UK Risk Assessment identified two
subpopulations at risk: offspring during lactation and pregnant dams at parturition. The NOAELs
from the definitive one-generation reproductive toxicity study (a maternal NOAEL ~ 47 mg/kg-
bw/day (600 ppm) for effects on the offspring mediated via lactation; and a maternal NOAEL ~
100 mg/kg-bw/day (1200 ppm) for effects on the dam during the time of parturition) were used
to calculate risk. Assuming a conservative value of 50% oral absorption, the margin of safety
(MOS) for effects on the offspring mediated via lactation and effects on the dam during the time
of parturition were calculated for workers, consumers, and other scenarios. In all but one
scenario (oil-based metal working fluids), the margins of safety were above 100 and in many
cases, several fold above. In addition, margins of exposure were calculated for infants exposed
via breast milk and via cow's milk, and in both instances, large MOEs (i.e., > 100) were
calculated.
Additional studies with MCCPs have been conducted in an effort to clarify the possible causes of
the hemorrhaging in the pups. One (single-dose; 6250 ppm or 538 mg/kg-bw/day) study showed
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maternal death during parturition due to low levels of vitamin K and related hemorrhaging,
suggesting that the act of parturition places dams at higher risk. It was concluded in data from
this study and a cross-fostering study that the fetus relies on clotting factors via mother's milk
and severe deficiencies in vitamin K levels and related clotting factors in the pups results in
hemorrhaging.
No definitive developmental neurotoxicity studies on MCCPs were located. It is not clear if any
developmental neurotoxicity endpoints were actually measured in the available prenatal
developmental/reproductive toxicity studies; none were explicitly stated. The only information
available regarding behavior during development is from cage-side observations in pups through
LD 21. In these cases, no dose-related differences were reported in Fi post-weaning appearance
or cage-side behaviors.
In the prenatal developmental toxicity study in rats, the LOAEL for maternal toxicity was
2000 mg/kg-bw/day based on clinical signs. The NOAEL for maternal toxicity was 500
mg/kg/day. The NOAEL for developmental toxicity was 5000 mg/kg-bw/day, the highest
dose tested.
In the prenatal developmental toxicity study in rabbits, no adverse, treatment-related effects were
reported in the dams or the offspring. The NOAEL for both maternal and developmental
toxicity was 100 mg/kg-bw/day, the highest dose tested.
In the reproduction range-finding study in rats, the LOAEL for maternal toxicity was 6250
ppm (463 mg/kg-bw/day) based on reductions in body weight gains. The NOAEL for
maternal toxicity was 1000 ppm (74 mg/kg-bw/day). The LOAEL for developmental
toxicity was 1000 ppm (62/74 mg/kg/day) based on pup mortality associated with internal
hemorrhages. The NOAEL for developmental toxicity was 100 ppm (6/8 mg/kg-bw/day).
No effects on any reproductive parameters were reported. The NOAEL for reproductive
toxicity was 6250 ppm (384/463 mg/kg-bw/day).
In the one-generation reproduction toxicity study in rats, the LOAEL for maternal toxicity was
1200 ppm (-100 mg/kg-bw/day) based on increases in liver weight; the NOAEL for
maternal toxicity was 600 ppm (~ 47 mg/kg-bw/day). The NOAEL for developmental and
reproductive toxicity was 1200 ppm (~ 84/99 mg/kg-bw/day), the highest dose tested.
Basis for Conclusions
In a range-finding prenatal developmental toxicity study in pregnant Charles River COBS CD
rats administered MCCPs (C14-17, 52 wt% CI) via gavage at dose levels of 0, 1000, 1500, and
2500 mg/kg-bw/day on gestation days (GD) 6-20, no effects were observed in the dams at doses
up to 2500 mg/kg-bw/day (IRDC, 1983, 1984; cited in: EURAR, 2008). As a result, doses
greater than 2500 mg/kg-bw/day were selected for the definitive study.
In the definitive study, four groups of 25 pregnant Charles River COBS CD rats were
administered MCCPs (C14-17, 52 wt% CI) via gavage at doses of 0, 500, 2000, and 5000 mg/kg-
bw/day on GD 6-19 (IRDC, 1984; cited in: EURAR, 2008). Unmated males and females were
individually housed and acclimated for 21-days in an environmentally controlled room. At the
end of the acclimation period, all animals were weighed and subjected to a detailed physical
examination. One female and one male rat were placed together for mating. Confirmation of
mating was based on evidence of a copulatory plug or by vaginal smear for sperm. The day
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mating was confirmed was designated as day 0 of gestation. Test article was administered to
pregnant females orally by gavage as a single daily dose on GD 6-19. During treatment, pregnant
females were observed daily for mortality and clinical signs of toxicity. Any females not
surviving to scheduled sacrifice were necropsied. Body weights were recorded on GD 0, 6, 9, 12,
16, and 20. All females were sacrificed on GD 20 and the uterus and ovaries excised for
examination. The number and location of viable and nonviable fetuses, early and late
resorptions, and the number of total implantations and corpora lutea were recorded. The uterus
was weighed. The abdominal and thoracic cavities underwent gross examination. Maternal
tissues were preserved for future histopathological analysis. Fetuses were weighed, sexed,
tagged, and examined for external malformations and variations, including the palate and the
eyes. The fetuses underwent visceral and skeletal examinations for malformations and
developmental variations.
The only effects reported in dams consisted of an increased incidence in wet matted and yellow
stained haircoat in the anogenital area at 5000 mg/kg-bw/day, and soft stool at > 2000 mg/kg-
bw/day. No treatment-related adverse effects were reported in offspring at doses up to 5000
mg/kg-bw/day. The LOAEL for maternal toxicity was 2000 mg/kg-bw/day based on clinical
signs; the NOAEL for maternal toxicity was 500 mg/kg-bw/day. The NOAEL for developmental
toxicity was 5000 mg/kg-bw/day, the highest dose tested.
In a range-finding prenatal developmental toxicity study in pregnant Dutch Belted rabbits
administered MCCPs (C14-17, 52 wt% CI) via gavage at dose levels of 0, 100, 300, 1000, 2000,
and 3000 mg/kg/day on GD 6-27, an increase in the number of abortions was observed at > 1000
mg/kg/day (IRDC, 1982a; cited in: EURAR, 2008). Body weight reductions in the dams were
reported at 100 and 300 mg/kg/day. As a result, another range-finding prenatal developmental
toxicity study in rabbits was initiated. This second range-finding study showed decreases in
maternal weight gain at 80 and 160 mg/kg-bw/day (IRDC, 1982b; cited in: EURAR, 2008).
Based on the results of these range-finding studies, dose levels of 10, 30, and 100 mg/kg-bw/day
were selected for the definitive prenatal oral gavage developmental toxicity study (IRDC, 1983;
cited in: EURAR, 2008). In the definitive study, four groups of 16 pregnant Dutch Belted rabbits
were administered 0, 10, 30, and 100 mg/kg-bw/day MCCPs (C14-17, 52 wt% CI) via gavage on
GD 6-27. Unmated males and females were individually housed and acclimated for 50-days in
an environmentally controlled room. As a result of a positive finding for parasites in stool
samples collected during acclimation, all rabbits received sodium sulfamethazine in their
drinking water for 16 days during the acclimation period. This treatment was terminated 4 weeks
prior to study initiation and only rabbits testing negative for parasites were placed on study. At
the end of the acclimation period, all animals were weighed and subjected to a detailed
examination. Females were impregnated via artificial insemination. Three weeks prior to
artificial insemination, females were given chorionic gonadotropin via an injection in a marginal
ear vein in order to induce superovulation. Semen was collected from males of proven fertility
and evaluated for motility. The day of artificial insemination was designated as day 0 of
gestation. During treatment, pregnant females were observed for mortality and clinical signs of
toxicity. Body weights were recorded on GD 0, 6, 12, 18, 24, and 28. Any females not surviving
to scheduled sacrifice were necropsied. On GD 28, all surviving females were sacrificed and the
uterus and ovaries excised for examination. The location and number of viable and nonviable
fetuses, early and late resorptions, and the number of total implantations and corpora lutea were
recorded. The uterus was weighed. The thoracic and abdominal cavities underwent gross
examination. Pooled samples of abdominal adipose tissue from 3 dams were frozen for future
analysis. Each fetus was sexed, weighed, and examined for external malformations and
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variations, including the palate and the eyes, as well as visceral and skeletal examinations for
malformations and developmental variations, including examination of the brain and the heart.
No adverse, treatment-related effects were reported in the dams or the offspring at doses up to
100 mg/kg-bw/day. The NOAEL for both maternal and developmental toxicity was 100 mg/kg-
bw/day, the highest dose tested.
In a one-generation reproductive toxicity range-finding study, four groups of 5 male and 10
female Charles River COBS SC rats were administered MCCP (C14-17, 52 wt% CI) via the diet at
0, 100, 1000, and 6250 ppm (~ 0, 6, 62, and 384 mg/kg-bw/day, respectively, in males; and 0, 8,
74, or 463 mg/kg-bw/day, respectively, in females) (IRDC, 1985; cited in: EURAR, 2008). Fo
animals were exposed to test substance from 28 days prior to mating until sacrifice; Fi animals
were treated from weaning until sacrifice, with additional potential exposures occurring in utero
and during lactation. All Fo males were sacrificed after the mating period. Following the
premating period, each male was cohabited with two females for 10 days. Females were
examined for evidence of copulation by means of vaginal smears and/or the appearance of a
vaginal plug. The day evidence of copulation was determined was designated as day 0 of
gestation. Direct dosing began at 83 days of age for the Fo parents and at 21 days of age for the
Fi weanlings. The Fo and Fi animals were observed for clinical signs of toxicity, changes in
general appearance and behavior, and mortality. In the Fo adults, body weights and food
consumption were measured weekly; in addition, body weights were measured in Fo females on
GD 0, 7, 14, and 20; and on lactation days (LD) 0, 7, 14, and 21. Estrous cyclicity was
determined in Fo females prior to mating, during mating, and prior to dosing. All Fo females
were allowed to deliver. The day the entire litter was found and delivery was judged to be
complete was designated as LD 0. Gestation duration was calculated. Following delivery, all
pups were examined for external malformations and the numbers of live births and stillbirths
(litter size) was recorded for each dam. Pups were weighed, sexed, and examined externally on
LD 0, 7, 14, and 21. Litter size was determined on LD 0, 4, 10, and 21. The number of male and
female pups was recorded on LD 4. Litters were examined daily for survival. Fo females were
examined for behaviors in nesting and nursing. On LD 21, all dams were sacrificed and a gross
necropsy performed, including examination of the uterine contents for implantation sites; and ten
Fi weanlings/sex/dose were sacrificed and necropsied. Five Fi males and ten Fi females/dose
group were retained after LD 21 and sacrificed at 70 days (10 weeks) of age and necropsied. Due
to high mortality in high-dose Fi pups, the surviving Fi pups in the high-dose group and an equal
number of control pups were sacrificed on LD 6 and 7 and necropsied. Blood was collected via
heart puncture and complete blood counts performed. Bone marrow smears were collected from
the femur, and the abdominal contents of the pups with milk in the stomach were collected and
frozen for future analyses.
Effects in the adults consisted of isolated reductions food consumption and body weight in the
dams at 6250 ppm. Effects in the offspring consisted of significant reductions in pup survival at
the high dose (none of the Fi pups in the high-dose group survived until lactation day 21); and
slight (11%, not statistically significant) decreases in pup survival, and labored breathing,
subcutaneous hematoma, pale discoloration, blood around the orifices, pale liver, kidney, and
spleen, and blood in the cranial cavity and brain beginning at the mid-dose. No dose-response,
treatment-related adverse effects were reported in the offspring in the low dose group.
Reductions in body weight in Fi male and female pups occurred during LD 7, 14, and 21, but
these reductions were not statistically significantly different from controls, and were seen only in
the low- and mid-dose groups but not the high-dose group. There were no dose-related
differences in Fi post-weaning appearance, behavior, food consumption, or clinical or
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anatomical pathology in the low- and mid-dose groups. Based on the results of this study, it was
recommended that dosage levels in a two-generation reproduction toxicity study not exceed 1000
ppm. The LOAEL for maternal toxicity was 6250 ppm (-463 mg/kg-bw/day) based on
reductions in body weight gains. The NOAEL for maternal toxicity was 1000 ppm (-74 mg/kg-
bw/day). The LOAEL for developmental toxicity was 1000 ppm (-62/74 mg/kg-bw/day) based
on pup mortality due to hemorrhaging. The NOAEL for developmental toxicity was 100 ppm
(-6/8 mg/kg-bw/day). No effects on any of the reproductive parameters were reported. The
NOAEL for reproductive toxicity was 6250 ppm (-384/463 mg/kg-bw/day).
In an effort to determine the cause of hemorrhaging in the pups at the high dose from the
reproductive toxicity range-finding study, a screening level cross-fostering developmental
toxicity study was conducted in Charles River COBS Wistar rats fed diets containing either 0 or
6250 ppm (~ 3125 mg/kg-bw/day) MCCPs (C14-17, 52 wt% CI) for 4 weeks prior to mating and
throughout pregnancy in a series of groups (Hart et al., 1985; cited in: EURAR, 2008). Offspring
from two of these groups (pups from control females reared from treated females, and pups
reared from their treated mothers) showed high-pup mortality associated with internal
hemorrhages. Hematological assays in the pups from these two groups showed decreases in
factor X, resulting in a disruption of a vitamin K-dependent clotting system (lower plasma
vitamin K levels). It was concluded that the pup mortalities were due to internal hemorrhages
caused by a decrease in the vitamin K-dependent hemostatic mechanism (not examined in this
study), induced during lactational exposures via the milk from mothers receiving MCCPs.
Additional studies have been conducted to investigate two hypotheses in an effort to clarify the
possible causes of the hemorrhaging in the pups.
The first hypothesis proposes that MCCPs induce a catabolism of vitamin K in lactating rats
leading to decreased plasma concentrations and ultimately low levels of vitamin K in the milk
pups receive (vitamin K controls the formation of several clotting factors in the liver). In order to
test this hypothesis, a preliminary study (CXR Biosciences Ltd., 2003; cited in: EURAR, 2008)
was conducted in which three groups of 6 female adult Sprague-Dawley were administered
MCCPs (C14-17, 52 wt% CI) via gavage at doses of 0, 500, or 1000 mg/kg-bw/day for 21 days
while being fed a normal diet or a vitamin K-deficient diet. Following exposures to MCCPs,
significant decreases in plasma concentrations of a clotting factor were seen in rats fed a normal
diet; however, these decreases did not affect prothrombin clotting times. Reductions of a clotting
factor in both treated and control groups were also seen in animals fed a vitamin-K deficient diet.
Plasma vitamin K levels were not affected by treatment in the normal diets, but they were lower
in high-dose animals fed vitamin K-deficient diets. The results from this study suggested that
MCCPs did not adversely affect the blood clotting system in adult female rats treated for 3
weeks up to a dose of 1000 mg/kg-bw/day; and the hemorrhaging effects in pups are unlikely to
be mediated by reduced vitamin K levels in breast milk.
The second hypothesis proposes that MCCPs transferred to the pups through breast milk causes
disruption of the pup clotting system. In order to test this hypothesis, a study (CXR Biosciences
Ltd., 2004; cited in: EURAR, 2008) was conducted in two groups of 16 male and 32 female
Sprague-Dawley rats administered 0 or 6250 ppm (~ 0 and 513 and 538 mg/kg-bw/day in males
and females, respectively) MCCPs (C14-17, 52 wt% CI) for 4 weeks prior to mating, during
cohabitation, gestation, and lactation until study termination (at about 2 weeks after the first
litters were born, due to high rate of pup mortality). Milk, blood, and liver samples from
lactating dams, and blood and liver samples from lactating pups were assessed for plasma
vitamin K levels. Five dams died or were killed at the time of parturition (16% mortality). These
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deaths were considered to be treatment-related as there was no indication of obstruction or
hindrance to delivery. The clinical necropsy of these dams showed effects suggestive of
hemorrhaging in 3 out of the 5 dams and one male who died. Slight reductions in food
consumption and body weight gains were observed during gestation and lactation. There were no
effects on mating performance or duration of gestation. Concentrations of plasma vitamin K
levels in adult females having gone through lactation and pregnancy was markedly decreased by
treatment with MCCPs, which in turn produced a decrease in activity of the plasma clotting
factors in treated dams. Prothrombin clotting times were not affected in the dams, suggesting that
the functional reserve in these adult animals was sufficient. Pup plasma volumes were reportedly
insufficient to measure vitamin K directly, but clotting factor activities were possible to analyze.
No effects on litter size at birth or on pup mortality from birth to LD 4 were reported; however,
after pup mortality increased significantly after LD 4. The majority of these pups showed
internal hemorrhages at necropsy. It was concluded that data from this study and the cross-
fostering study performed by Hart et al. (1985) suggest that the fetus receives sufficient vitamin
K via the placenta, but after birth becomes severely deficient in vitamin K and related clotting
factors and relies on these factors via mother's milk. In addition, the pups also receive
considerable levels of MCCPs via lactation (through mother's milk) which may also contribute
to further reducing the vitamin K levels. These severe deficiencies in vitamin K levels and
related clotting factors in the pups results in hemorrhaging. It was also concluded that the act of
parturition places dams at higher risk.
More recently, a definitive one-generation reproductive toxicity study was conducted to refine
the NOAEL for effects in the offspring and to further explore the mechanisms of hemorrhaging
(CXR, 2006; cited in: EURAR, 2008). This study was reportedly conducted in compliance with
OECD TG 421 and Good Laboratory Practice standards. Four groups of 12-17 male and female
Sprague-Dawley rats were administered 0, 300, 600, and 1200 ppm (~ 0 and 21, 44, and 84
mg/kg-bw/day in males; and 0, 23, 47, and 99 mg/kg-bw/day in females) MCCPs (C14-17, 52
wt% CI) for 4 weeks prior to mating, during cohabitation, gestation, and lactation until study
termination (for a total treatment of 11-12 weeks). Males were terminated on LD 4 (9 weeks of
treatment) and females were allowed to litter and rear their offspring until PND 21. Females
were sacrificed on LD 21. Adult males were assessed for signs of clinical toxicity, body weight,
food consumption, and macropathology. Adult females were assessed for signs of clinical
toxicity, body weight, food consumption, gestation length, parturition, liver weights, and
macropathology. Mating performance and fertility were also evaluated. Offspring evaluations
included clinical signs of toxicity, litter size, survival, sex ratio, body weight, and pathological
examinations at necropsy. Milk, blood, and liver samples were obtained from selected offspring
at specific time points between birth of litters and PND 21. In addition, blood, liver, and milk
samples from a satellite group of five females and their litters from the control and high-dose
group (1200 ppm) were collected for future analysis. Analysis of these samples was still pending
at the time of the UK assessment.
No adverse effects were reported in the adult animals for clinical condition, body weight, body
weight gain, food consumption, estrous cycling, mating performance, pre-coital interval, fertility,
number of implantations, gestation lengths, or parturition. The only effect reported was for
higher absolute and relative liver weights in high-dose females (1200 ppm; 99 mg/kg-bw/day).
Likewise, no adverse effects were in the offspring at any dose level for litter size, sex ratio,
offspring survival, body weights, body weight gains, macropathology and liver weights. No
adverse effects were reported on pre- and post-natal survival and growth up to sacrifice
(weaning). Though no histopathology was performed, the body cavity and cranial cavity were
opened and examined for any signs of hemorrhaging. None was reported. Based on the results of
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this study, the LOAEL for maternal toxicity was 1200 ppm (-100 mg/kg-bw/day) based on
increases in liver weight; the NOAEL for maternal toxicity was 600 ppm (~ 47 mg/kg-bw/day).
The NOAEL for developmental and reproductive toxicity was 1200 ppm (~ 84/99 mg/kg-
bw/day), the highest dose tested.
C-2 LCCP HEALTH DATA REVIEW
There is no information on inhalation absorption of LCCPs in humans or in animals. Based on
their low vapor pressure and water solubility, absorption following inhalation or dermal
exposure is expected to be limited. Some absorption and metabolism following oral exposure are
possible for LCCPs with shorter carbon chain length and lower degree of chlorination.
No information is available on the toxicity of LCCPs in humans. Acute oral toxicity data in
animals show that LCCPs are of very low acute toxicity. Studies in animals have shown that
some LCCPs may have the potential to cause slight skin irritation and sensitization but no eye
irritation potential. The liver is the main target organ of LCCPs in repeated dose studies in
experimental animals. Inflammatory and necrotic changes of the liver were observed in rats
exposed to a C20-30 LCCP with 43 wt% CI at dose levels of 100 mg/kg-bw and above. For
another LCCP with C20-26 70 wt% CI, effects in the liver occurred at a very high exposure level
of 3,750 mg/kg-bw/day; the NOAEL was 900 mg/kg-bw/day.
An increased incidence of malignant lymphoma in male mice was reported at the highest dose of
5,000 mg/kg-bw/day when tested using a C23 LCCP with 43 wt% CI in carcinogenicity studies in
male and female rats and mice. However, malignant lymphoma is one of the more variable
tumors in mice and has a viral origin in many cases. Data on the analogous short-chain
chlorinated paraffins (SCCPs) have shown no increase in the incidence of malignant lymphoma
in a carcinogenicity study of SCCPs. LCCPs are non-genotoxic and they are not expected to pose
a carcinogenic hazard to humans.
No testing is needed for the PMN substances. Based on the LOAEL (100 mg/kg-bw) of the liver
effects in female rats of repeated dose studies, Health Canada calculated a tolerable daily intake
(TDI) of 71 [j,g/kg-bw/day. Using upper bounding intake estimates ranging from 0.007 (J,g/kg-
bw/day for 60+ age group to 0.024 [j,g/kg-bw/day for 0.5 years age group, Environment Canada
determined that the exposure levels are 10,000 and 3,000 times lower, respectively, than the
TDI. Based on these evaluations, Health Canada concluded: "LCCP are not harmful to human
health as defined in the Canadian environmental Protection Act.'"
The National Research Council (NRC, 2000) reviewed the toxicological risks of selected flame
retardant, including a LCCP containing C24 with 70 wt% CI. Based on the NOAEL of 900
mg/kg-bw/day (liver toxicity), the NRC derived an RfD of 0.3 mg/kg-bw/day. Using this RfD
and the worst case average daily exposure to be 0.16 mg/kg/day, NRC concluded: "LCCP do not
pose a noncancer risk when incorporated into residential furniture at the estimated application
levels." Further, it was concluded that: "LCCP are not likely to be a human carcinogen and
derivation of a cancer potency factor is unnecessary
C-2-1 Metabolism
There is no information on inhalation absorption of LCCPs in humans or in animals. Based on
their low vapor pressure and low water solubility, absorption following inhalation or dermal
exposure is expected to be limited.
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Oral (gavage) studies (IRDC, 1981; cited in: UK, 2009) showed that approximately 82-95% of a
single dose of [8-14C]-labeled C22 -26 chlorinated paraffin (43% CI) was recovered in the feces in
rats during the seven-day collection period. Only 0.1-0.8%) of the radiolabel was excreted in the
urine. For C22-26 chlorinated paraffin (70 wt%> CI), it was found that 61-88%) of the administrated
radioactivity was recovered in the feces in rats during the seven-day collection period. Less than
0.1-1%) was excreted in the urine.
C-2-2 Acute Toxicity
Acute toxicity studies have been conducted in rats, mice or dogs on five LCCPs: C20-30,41-50
wt%> CI; C22-26, 42 wt%> CI, C23,43 wt%> CI; C20-30,61-70 wt%> CI; C24,70% CI. The maximum
dose levels used in these studies ranged from 4-50 g/kg-bw. No deaths were reported in any of
the studies (IUCLID, 2003; cited in: UK, 2009).
C-2-3 Irritation and Sensitization
Skin irritation testing has been conducted on four LCCPs: C19 44 wt%> CI; C22-26, 42 wt%> CI;
C20-30,41-50 wt%> CI; C 20-30,70 wt%> CI. No evidence of irritation was seen in three of the four
LCCPs. For the C22-26, 42 wt%> CI product, erythema was observed in two of six animals tested;
the severity threshold was below the classification of the EU system (IUCLID, 2003; cited in:
UK, 2009).
Evidence of slight eye irritation was seen in a test of a C22-26, 42 wt%> CI product. However, the
criteria for classification as an eye irritant were not met (IUCLID, 2003; cited in: UK, 2009).
In a maximization test and a Buehler test using guinea pig, a C22-26, 42 wt%> CI product was
negative (IUCLID, 2003; Bailey and Sheldon, 1998; cited in: UK, 2009). A C18-27, 40 wt%> CI
product was reported to elicit a positive response in the guinea pig maximization test (IUCLID,
2000; cited in: UK, 2009). However, no information is available on the quality of this study.
C-2-4 Repeated-dose Toxicity
LCCPs with C20-30,43 wt% CI were dissolved in corn oil and given by gavage at 100, 900 or
3,750 mg/kg-bw/day to groups of 15 male and female Fisher 344 rats in a 14- and a 90-day
studies (IRDC1981, 1984; cited in: Serront el a/., 1987). There was a treatment-related effect on
the liver of female rats at all dose levels, but no liver effects were seen in the males. Female liver
weights were increased and a multifocal granulomatous hepatitis characterized by inflammatory
changes and necrosis. Nephrosis was observed in the kidney of male rats and mineralization in
the kidneys of female rats at 3,750 mg/kg-bw/day. Similar liver effects were observed in the
high-dose (3,750 mg/kg-bw/day) rats of both sexes in a 90-day study on LCCPs with C22-26,70
wt%> CI (IRDC 1981; cited in: Serront et al., 1987) and in a 90-day study as well as a 2-year
bioassay on a LCCP with an average of C23,43 wt%> CI at 100 mg/kg-bw/day (NTP, 1986).
Based on the liver toxicity, a LOAEL of 100 mg/kg-bw/day is established for the LCCPs with 43
wt%> CI and a NOAEL of 900 mg/kg-bw/day is identified for the LCCPs with 70 wt%> CI.
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C-2-5 Genotoxicity
Both LCCPs of C22 -26, 43 wt% CI and C23,43 \vt"o CI were negative in several SalfHouella strains
of the Ames test with or without metabolic activation (IUCLID, 2003; NTP, 1986; cited in: UK,
2009). LCCPs with C20-30 with 43 wt% CI or C22-26 with 70 wt% CI did not induce significant
increases of chromosomal or chromatid aberrations in bone marrow cells of rats (IRDC1983;
cited in. Serrone etal., 1987).
C-2-6 Carcinogenicity
The carcinogenicity of an LCCP (C23,43 wt% CI) was studied by administering the chemical in
corn oil by gavage to groups of 50 F344/N rats and 50 B6C3F1 mice of each sex, 5 days per
week for 103 weeks. Male rats received doses of 0, 1,875 or 3,750 mg/kg-bw/day; female rats
received 0, 2,500 or 5,000 mg/kg-bw/day. An increased incidence of malignant lymphomas was
reported for male mice given the LCCP; the incidences for the controls, low- and high dosed
groups are 6/50, 12/50, and 16/50, respectively. However, mice are prone to the development of
lymphomas with a range of tumor incidence of 2-73, and viruses may be the causative agent of
lymphomas in laboratory strains of mice. It has been concluded that LCCPs of C20-30 with 43
wt% CI did not produce clear evidence of carcinogenicity when tested in rats and mice (Serrone
etal., 1987).
C-2-7 Developmental Reproductive Toxicity Review
In a prenatal developmental toxicity study in rats, no treatment-related effects were reported.
The NOAEL for both maternal and developmental toxicity was 5000 mg/kg-bw/day, the
highest dose tested.
In a prenatal developmental toxicity study in rabbits, the LOAEL for maternal toxicity was
500 mg/kg-bw/day (the lowest dose tested) based on increased incidence of clinical signs.
The NOAEL for developmental toxicity was > 2000 mg/kg-bw/day, the highest dose tested.
In a prenatal developmental toxicity study in rabbits, the NOAEL for both maternal and
developmental toxicity was 1000 mg/kg-bw/day, the highest dose tested.
Basis for Conclusions
In a range-finding study (Study # 438-033, October 27, 1981; Chlorinated Paraffin Consortium),
in pregnant Charles River COBS rats administered LCCP (22-26 carbons, 43 wt% CI) via
gavage at dose levels of 0, 3000, and 5000 mg/kg-bw/day on GD 6-19, the only effects reported
occurred in dams and consisted of a slight increased incidence in anogenital matting during the
latter portion of the treatment period at 5000 mg/kg-bw/day. No adverse treatment-related effects
were reported in offspring at doses up to 5000 mg/kg-bw/day.
Based on the findings of the range-finding study, four groups of 25 pregnant Charles River
COBS CD rats were administered LCCP (22-30 carbons, 70 wt% CI) via gavage at doses of 0,
500, 2000, and 5000 mg/kg-bw/day on GD 6-19 (Study # 438-045, April 11, 1984; Chlorinated
Paraffin Consortium). Unmated males and females were individually housed and acclimated for
15-days in an environmentally controlled room. At the end of the acclimation period, all animals
were weighed and subjected to a detailed physical examination. One female and one male rat
were placed together for mating. Confirmation of mating was based on evidence of a copulatory
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plug. The day mating was confirmed was designated as day 0 of gestation. Test article was
administered to pregnant females orally by gavage as a single daily dose on GD 6-19. During
treatment, pregnant females were observed daily for mortality and clinical signs of toxicity. Any
females not surviving to scheduled sacrifice were necropsied. Body weights were recorded on
GD 0, 6, 9, 12, 16, and 20. All females were sacrificed on GD 20 and the uterus and ovaries
excised for examination. The number and location of viable and nonviable fetuses, early and late
resorptions, and the number of total implantations and corpora lutea were recorded. The uterus
was weighed. The abdominal and thoracic cavities underwent gross examination. Maternal
tissues were preserved for future histopathological analysis. Fetuses were weighed, sexed, tagged
and examined for external malformations and variations, including the palate and the eyes. The
fetuses underwent visceral and skeletal examinations for malformations and developmental
variations.
No adverse treatment-related effects were reported in the dams or offspring at doses up to 5000
mg/kg-bw/day. The NOAEL for both maternal and developmental toxicity was 5000 mg/kg-
bw/day, the highest dose tested.
In a range-finding study (Study # 438-018, October 27, 1981; Chlorinated Paraffin Consortium)
in pregnant Dutch Belted rabbits administered LCCP (22-26 carbons, 43 wt% CI) via gavage at
doses of 0, 500, 1000, 2000, 3000, and 5000 mg/kg-bw/day on GD 6-27, a slight decrease in the
amount of feces and a slight increase in matting and/or staining of the haircoat was reported in
the dams at > 3000 mg/kg-bw/day. No other effects were reported. Observations of offspring did
not appear to be included.
Based on the results of the range-finding study, four groups of 16 Dutch Belted pregnant rabbits
were administered LCCP (22-26 carbons, 43 wt% CI) via gavage at doses of 0, 500, 2000, and
5000 mg/kg-bw/day on GD 6-27 (Study # 438-030, August 26, 1982; Chlorinated Paraffin
Consortium). Unmated males and females were individually housed and acclimated for 9-weeks
in an environmentally controlled room. As a result of a positive finding for parasites in stool
samples collected during acclimation, all rabbits received sodium sulfamethazine in their
drinking water for 19 days during the acclimation period. This treatment was terminated 5 weeks
prior to study initiation and only rabbits testing negative for parasites were placed on study. At
the end of the acclimation period, all animals were weighed and subjected to a detailed
examination. Females were impregnated via artificial insemination. Immediately after
insemination, ovulation was induced via an injection of chorionic gonadotropin in a marginal ear
vein. Semen was collected from males of proven fertility and evaluated for motility. The day of
artificial insemination was designated as day 0 of gestation. During treatment, pregnant females
were observed for mortality and clinical signs of toxicity. Body weights were recorded on GD 0,
6, 12, 18, 24, and 28. Any females not surviving to scheduled sacrifice were necropsied. On GD
28, all surviving females were sacrificed and the uterus and ovaries excised for examination. The
location and number of viable and nonviable fetuses, early and late resorptions, and the number
of total implantations and corpora lutea were recorded. The uterus was weighed. The thoracic
and abdominal cavities underwent gross examination. Pooled samples of abdominal adipose
tissue from 3 dams were frozen for future analysis. Each fetus was sexed, weighed, and
examined for external malformations and variations, including the palate and the eyes, as well as
visceral and skeletal examinations for malformations and developmental variations, including
examination of the brain and the heart.
A dose-related trend in an increased incidence in soft stool and/or anogenital matting or staining
was observed in the dams beginning at the low-dose group. Three dams aborted and were
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sacrificed during treatment; one in the mid-dose group, and two in the high-dose group.
Increases in postimplantation loss and corresponding decreases in viable fetuses, and increases in
late resorptions were reported at the high-dose group, however, these effects were not reported as
being statistically significant. No signs of treatment-related developmental toxicity were
reported in the offspring at < 5000 mg/kg-bw/day, although the sample size in the high-dose
group was limited, precluding any definitive conclusions. Therefore, the LOAEL for maternal
toxicity was 500 mg/kg-bw/day (the lowest dose tested), based on increased incidence of clinical
signs. The NOAEL for developmental toxicity was > 2000 mg/kg-bw/day.
In a range-finding study (Study # 438-038, November 1, 1982; Chlorinated Paraffin Consortium)
in pregnant Dutch Belted rabbits administered LCCP (22-30 carbons, 70 wt% CI) via gavage at
doses of 0, 2000, 3750, and 5000 mg/kg-bw/day on GD 6-27, increases in abortions, reductions
in maternal body weight, and increases in post-implantation losses were reported at > 2000
mg/kg-bw/day. Therefore, a second range-finding study was conducted at dose levels of 0, 50,
200, and 1000 mg/kg-bw/day (Study # 438-040, November 4, 1982; Chlorinated Paraffin
Consortium). A slight decrease in viable fetuses and a slight increase in postimplantation loss
were reported at 1000 mg/kg-bw/day. No other effects were reported.
Based on the results of the range-finding studies, four groups of 16 Dutch Belted pregnant
rabbits were administered LCCP (22-30 carbons, 70 wt% CI) via gavage at doses of 0, 100, 300,
and 1000 mg/kg-bw/day on GD 6-27 (Study # 438-039, July 18, 1983; Chlorinated Paraffin
Consortium). Unmated males and females were individually housed and acclimated for 40-days
in an environmentally controlled room. At the end of the acclimation period, all animals were
weighed and subjected to a detailed examination. Females were impregnated via artificial
insemination. Immediately after insemination, ovulation was induced via an injection of
chorionic gonadotropin in a marginal ear vein. Semen was collected from males of proven
fertility and evaluated for motility. The day of artificial insemination was designated as day 0 of
gestation. During treatment, pregnant females were observed for mortality and clinical signs of
toxicity. Body weights were recorded on GD 0, 6, 12, 18, 24, and 28. Any females not surviving
to scheduled sacrifice were necropsied. On GD 28, all surviving females were sacrificed and the
uterus and ovaries excised for examination. The location and number of viable and nonviable
fetuses, early and late resorptions, and the number of total implantations and corpora lutea were
recorded. The uterus was weighed. The thoracic and abdominal cavities underwent gross
examination. Pooled samples of abdominal adipose tissue from 3 dams were frozen for future
analysis. Each fetus was sexed, weighed, and examined for external malformations and
variations, including the palate and the eyes, as well as visceral and skeletal examinations for
malformations and developmental variations, including examination of the brain and the heart.
No treatment-related adverse effects were observed in the dams or the offspring at any dose
level. The increases in postimplantation losses noted in the two previous range-finding studies
reported above were not reproduced in the definitive study. Therefore, the NOAEL for both
maternal and developmental toxicity was 1000 mg/kg-bw/day, the highest dose tested.
101
-------
Appendix D ENVIRONMENTAL MONITORING
D-l MCCP MONITORING DATA
D-l-1 Surface Water
It is known that over time, based on their molecular weight and physicochemical properties,
MCCPs in surface water will partition to suspended particulates, sediment, sludge, or soil.
Reported MCCP concentrations in surface water range from < 2.50 x 10"10to 1.49 x 10"3 mg/L
(TableApx D-l-1). Very little information is available on the specific sampling locations for
many of the surface water measurements reported in TableApx D-l-1. Limited documentation
is available on two of the studies (Petersen et al., 2006 and Muir, 2003). Two sources provide a
review of the literature with very little details (IPCS, 1996 and EC, 2008b). Two studies do
provide detailed information on the sampling approach, including location (Houde et al., 2008
and USEPA, 1988). The Petersen et al. (2006) study, which had the highest published
concentration, reported results for water samples collected from different Norwegian locations.
EPA/OPPT assumes that these samples were collected in non-marine waters. Three studies found
were not used in this assessment (BUA, 1992; Hoechst, 1987; and Willis, 1994). However, all of
the studies used in the assessment use modern analytical techniques, reference the specific CPs
of interest, and provide, at a minimum, general information on the sampling location. Given the
paucity of surface water data available, EPA/OPPT used measurements from the selected studies
and used the minimum and maximum values in this assessment.
Measurements of dissolved (filtered) concentrations were generally non-detectable (ND) with
few exceptions. Concentrations measured in surface water were largely from studies that
measured total water concentrations which included MCCPs sorbed to particulates. More recent
monitoring studies (Table Apx D-l-1) have focused on measuring MCCPs in suspended solids,
sediment pore water, and sediment.
Early analytical methods using thin layer chromatography (TLC) were used to measure CPs in
surface water. However, this method has poor sensitivity and reproducibility, and provide false
negative results. Current methods of quantification using gas or liquid chromatography coupled
with a range of detectors {i.e., mass spectrometry; MS) are more reliable. Nearly all of the water
concentrations were measurements taken at a single point in time {i.e., the samples were not time
series samples). Absent more extensive monitoring data, EPA/OPPT assumed that the available
data could be extrapolated to longer time periods for determination of a chronic exposure
concentration.
MCCP concentrations in surface water, reported in Table Apx D-l-1, rely on test methods that
filtered or pre-filtered samples before they were analyzed, which can underestimate
environmental concentrations. Where appropriate, reported values were converted to a common
unit, as presented in the table. For the purposes of this assessment, in the studies considered
acceptable, EPA/OPPT used the lowest and highest reported concentrations (< 2.50 x 10"10to
1.49 x 10"3 mg/L; Table Apx D-l-1) to evaluate risks of potential concern to aquatic organisms.
102
-------
Table Apx D-l-1: Surface Water Concentrations of MCCPs, Sorted by Country
Media
Country
Location
City, State or
Province
Comments
Converted
Concentration
Common Units
Analytical Method
References
Maximum
2.60xl0"9
mg/L
NR
EC (2008b)
<
2.50xl0"10
mg/L
GC-HRMS-MAB
Houde et al. (2008)
Canada
Lake Ontario
<
l.OOxlO"8
mg/L
GC-ECNI-MS
Muiretal. (2003)
Maximum
4.70xl0"8
mg/L
GC-HRMS-MAB
Houde et al. (2008)
Mean
9.00xl0"10
mg/L
GC-HRMS-MAB
Houde et al. (2008)
River Lech at
Langsweid
1.90xl0"4
mg/L
NR
IPCS (1996)
River Lech at Rain
1.70xl0"4
mg/L
NR
IPCS (1996)
Surface Water
Germany
River Lech at
Gersthofen
9.00xl0"5
mg/L
NR
IPCS (1996)
River Lech at
Augsburg
<
2.50xl0"5
mg/L
NR
IPCS (1996)
River Danube at
...
7.00xl0"5
mg/L
NR
IPCS (1996)
Marxheim
<
3.00xl0"5
mg/L
NR
IPCS (1996)
Norway
NR
...
1.49xl0"3
mg/L
GC-ECNI-MS
Petersen et al. (2006)
United Kingdom
Multiple locations
<
lxlO"4
mg/L
GC-ECNI-MS
Nicholls et al. (2001)
United States
Sugar Creek, Ohio
<
7.50xl0"5
mg/L
GC-ECNI-MS
USEPA (1988)
Central European
Country
NR
<
5.00xl0"5
mg/L
GC-ECNI-MS
Coelhan (2010)
NR: Not recorded. Location description was not provided in the study.
: Single sample value reported above the detection limit; therefore, no data qualifier required.
GC-HRMS-MAB: Gas chromatography-high resolution mass spectrometry with metastable atom bombardment ionization
GC-ECNI-MS: Gas chromatography in combination with electron capture negative ion mass spectrometry
Notes:
1. All values provided in the table above represent total MCCP and not individual MCCP isomers
2. In some cases, the minimum values in the table are preceded by This indicates that the value reported in article was reported as a non-detect. In such cases, one half of the
lowest reported detection limit was compiled as the 'minimum' reported monitoring data
3. All concentrations measured from impoundment lagoons and drainage ditches from the USEPA (1988) study have not been included as they are not considered as surface water
concentrations
4. All concentrations measured from suspended solid matter fraction from influents from the Coelhan (2010) study have not been included as they are not considered as surface
water concentrations
103
-------
I)-1-2 Sediment
MCCP sediment concentrations from marine and non-marine environments ranged from 5.00 x
10"3 to 1.64 x 101 mg/kg dw and from 2.0 x 10"3 to 6.51 x 101 mg/kg dw, respectively.
For the purposes of this assessment, in the studies considered acceptable, EPA/OPPT used the
lowest and highest reported marine and non-marine sediment concentrations (5.00xl0"3 to 1.64 x
101 mg/kg dw and 2.00 x 10"3 to 6.51 x 101 mg/kg dw, respectively) to evaluate risks of potential
concern to sediment organisms (Table Apx D-l-2). Where appropriate, reported values were
converted to a common unit, as presented in the table.
104
-------
Table Apx D-1-2: Sediment Concentrations of MCCPs, Sorted hv Country
1 AVIIIIOII
(mi\ cried
( oininon
Mod i;i
( on mi r\
( oninieiils
References
( il>. Sialc or I'nn nice
( oiiceiiiraiioii
1 mis
...
1.11
mg/kg dw
Kemmlein et al.
(2002)
Australia
NR
...
1.17
mg/kg dw
Kemmlein et al.
(2002)
...
3.11
mg/kg dw
Kemmlein et al.
(2002)
...
1.64X101
mg/kg dw
Kemmlein et al.
(2002)
Canada
Hamilton Harbour
(Windemere basin)
...
2.90X10"1
mg/kg*
Muir et al. (2000)
...
5.00xl0"3
mg/kg dw
Hiittig et al.
(2004)
...
9.00xl0"3
mg/kg dw
Hiittig et al.
(2004)
Sediment
(Marine)
German Bight, North Sea
...
9.00X10"3
mg/kg dw
Hiittig et al.
(2004)
...
1.30xl0"2
mg/kg dw
Hiittig et al.
(2004)
...
2.80xl0"2
mg/kg dw
Hiittig et al.
(2004)
Germany
...
1.46X10"1
mg/kg dw
Hiittig et al.
(2004)
...
9.30xl0"2
mg/kg dw
Hiittig et al.
(2004)
...
1.15X10"1
mg/kg dw
Hiittig et al.
(2004)
Baltic Sea
...
1.22X10"1
mg/kg dw
Hiittig et al.
(2004)
...
2.11X10"1
mg/kg dw
Hiittig et al.
(2004)
...
4.99X10"1
mg/kg dw
Hiittig et al.
(2004)
North and Baltic Sea
...
2.20xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
105
-------
1 .OCIIIIOII
(mi\ cried
(\. Siaic or h\>\ ince
( oinmenls
References
2.30xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
3.30xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
3.40xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
3.70xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
3.90xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
4.30xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
4.30xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
4.80xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
5.40xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
5.80xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
6.10xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
7.20xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
7.60xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
7.70xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
8.10xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
8.50xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
8.70xl0"2
mg/kg dw
Hiittig and Oehme
(2006)
1.49X101
mg/kg dw
Hiittig and Oehme
(2006)
106
-------
1 .OCIIIIOII
(mi\ cried
(\. Siaic or h\>\ ince
( oinmenls
References
1.49X10"1
mg/kg dw
Hiittig and Oehme
(2006)
1.49X10"1
mg/kg dw
Hiittig and Oehme
(2006)
2.75 xlO"1
mg/kg dw
Hiittig and Oehme
(2006)
9.10xl0"2
mg/kg dw
Hiittig and Oehme
(2005)
4.80xl0"2
mg/kg dw
Hiittig and Oehme
(2005)
1.98X10"1
mg/kg dw
Hiittig and Oehme
(2005)
1.31X10"1
mg/kg dw
Hiittig and Oehme
(2005)
1.32X10"1
mg/kg dw
Hiittig and Oehme
(2005)
3.03X10"1
mg/kg dw
Hiittig and Oehme
(2005)
1.53X10"1
mg/kg dw
Hiittig and Oehme
(2005)
1.14X10"1
mg/kg dw
Hiittig and Oehme
(2005)
4.00xl0"2
mg/kg dw
Hiittig and Oehme
(2005)
2.70X10"2
mg/kg dw
Hiittig and Oehme
(2005)
1.80xl0"2
mg/kg dw
Hiittig and Oehme
(2005)
1.90xl0"2
mg/kg dw
Hiittig and Oehme
(2005)
3.00X10"2
mg/kg dw
Hiittig and Oehme
(2005)
3.20xl0"2
mg/kg dw
Hiittig and Oehme
(2005)
1.80xl0"2
mg/kg dw
Hiittig and Oehme
(2005)
107
-------
1 .ocalion
('omened
( oninioii
Malm
( on mi r\
( omiiiciils
References
( il>. Siale or l'ro\ iiice
('oiieeiiiraiion
1 mis
2.40xl0"2
mg/kg dw
Hiittig and Oehme
(2005)
Canada
Lake Erie
...
6.80xl0"2
mg/kg dw
Tomy and Stern
(1999)
Lake St Francis
Minimum
7.50X10"1
mg/kg dw
EC (2008b)
Maximum
1.2
mg/kg dw
EC (2008b)
NR
Minimum
2.00xl0"3
mg/kg*
Pribylova et al.
(2006)
Labe
Sum
1.80xl0"2
mg/kg dw
Pribylova et al.
(2006)
Sum
7.30X10-2
mg/kg dw
Pribylova et al.
(2006)
Libis-Labe
Sum
1.6
mg/kg dw
Pribylova et al.
(2006)
Bilina
Sum
3.10xl0"2
mg/kg dw
Pribylova et al.
(2006)
Sediment
Mala Becva
Sum
1.13X10"1
mg/kg dw
Pribylova et al.
(2006)
(Non-
marine)
Czech
Becva
Sum
1.20X10"1
mg/kg dw
Pribylova et al.
(2006)
Republic
Morava
Sum
1.93 xlO"1
mg/kg dw
Pribylova et al.
(2006)
Sum
3.08X10"1
mg/kg dw
Pribylova et al.
(2006)
Ohre
Sum
6.00X10"1
mg/kg dw
Pribylova et al.
(2006)
Sum
5.58
mg/kg dw
Pribylova et al.
(2006)
Morava
Sum
4.16X10"1
mg/kg dw
Pribylova et al.
(2006)
Dyje
Sum
7.57X10"1
mg/kg dw
Pribylova et al.
(2006)
Drevnice
Sum
8.93 xlO"1
mg/kg dw
Pribylova et al.
(2006)
Germany
Bodensee (middle)
<
5.00xl0"3
mg/kg dw
IPCS (1996)
108
-------
1 .ocalion
('omened
( omiiioii
Malm
( on mi r\
( omiiiciils
References
( il>. Sialc or ho\ incc
( oiicciilialioii
1 mis
7.00xl0"2
mg/kg dw
IPCS (1996)
<
5.00X10"3
mg/kg dw
IPCS (1996)
River Lech
3.25X10"1
mg/kg dw
IPCS (1996)
Maximum
7.00X10"1
mg/kg*
Tomy et al. (1998)
6.00xl0"2
mg/kg dw
IPCS (1996)
8.50X10"2
mg/kg dw
IPCS (1996)
River Rhine
1.40X10"1
mg/kg dw
IPCS (1996)
Minimum
1.45X10"1
mg/kg dw
IPCS (1996)
Maximum
2.05X10"1
mg/kg dw
IPCS (1996)
River Elbe at Hamburg
Minimum
1.30X10"1
mg/kg dw
IPCS (1996)
Maximum
2.30X10"1
mg/kg dw
IPCS (1996)
River Main
Minimum
1.60X10"1
mg/kg dw
IPCS (1996)
Maximum
2.60X10"1
mg/kg dw
IPCS (1996)
Outer Alster, Hamburg
3.70X10"1
mg/kg dw
IPCS (1996)
minimum
5.00xl0"2
mg/kg dw
Petersen et al.
(2006)
Norway
NR
maximum
3.24
mg/kg dw
Petersen et al.
(2006)
2.7
mg/kg ww
Borgen et al.
(2003)
1.14X101
mg/kg ww
Borgen et al.
(2003)
Minimum
8.80X10"1
mg/kg dw
Chen et al. (2011)
Minimum
1.1
mg/kg dw
Chen et al. (2011)
Minimum
1.4
mg/kg dw
Chen et al. (2011)
South China
Pearl River Delta
Maximum
1.4
mg/kg dw
Chen et al. (2011)
Maximum
3.8
mg/kg dw
Chen et al. (2011)
Mean
3.9
mg/kg dw
Chen et al. (2011)
Mean
2.10X101
mg/kg dw
Chen et al. (2011)
Maximum
3.80X101
mg/kg dw
Chen et al. (2011)
109
-------
1 .ocalion
(oii\eried
('oiiceiil ration
( oninioii
1 mis
Media
( on mi r\
( il>. Siale or ho\ iiiee
( oinmenls
References
Lake Thun
Minimum
5.00X10"3
mg/kg dw
Iozza et al. (2008)
Switzerland
Maximum
2.60xl0"2
mg/kg dw
Iozza et al. (2008)
Lake Zurich
Maximum
5.00X10"3
mg/kg*
Tomy et al. (1998)
NR
<
l.OOxlO"1
mg/kg dw
Nicholls et al.
(2001)
...
3.00X10"1
mg/kg dw
Nicholls et al.
(2001)
South West Region: Grand
Union Canal
...
2.7
mg/kg dw
Nicholls et al.
(2001)
...
2.8
mg/kg dw
Nicholls et al.
(2001)
...
5.00X10"1
mg/kg dw
Nicholls et al.
(2001)
South West Region; Bristol
Avon River
...
6.00X10"1
mg/kg dw
Nicholls et al.
(2001)
...
8.00X10"1
mg/kg dw
Nicholls et al.
(2001)
United
...
1.0
mg/kg dw
Nicholls et al.
(2001)
Kingdom
North East Region: Hull
River
...
1.35X101
mg/kg dw
Nicholls et al.
(2001)
...
1.1
mg/kg dw
Nicholls et al.
(2001)
South West Region: Colne
...
1.4
mg/kg dw
Nicholls et al.
(2001)
River
...
2.0
mg/kg dw
Nicholls et al.
(2001)
...
3.8
mg/kg dw
Nicholls et al.
(2001)
West Midlands Region:
Trent River
...
6.02X101
mg/kg dw
Nicholls et al.
(2001)
...
6.51X101
mg/kg dw
Nicholls et al.
(2001)
North West Region:
5.6
mg/kg dw
Nicholls et al.
Hornsmill brook
(2001)
110
-------
1 .OCIIIIOII
('nmeried
( oninioii
Media
( on mi r\
( oinmenls
References
( il>. Siale or h\>\ iiiee
(\iiicciiiralkiii
1 mis
1.25X101
mg/kg dw
Nicholls et al.
(2001)
1.83X101
mg/kg dw
Nicholls et al.
(2001)
1.0
mg/kg dw
Nicholls et al.
(2001)
North East Region: Hull
River
1.1
mg/kg dw
Nicholls et al.
(2001)
1.35X101
mg/kg dw
Nicholls et al.
(2001)
East Midlands Region: Idle
1.62X101
mg/kg dw
Nicholls et al.
(2001)
River
4.39X101
mg/kg dw
Nicholls et al.
(2001)
1.80X101
mg/kg dw
Nicholls et al.
(2001)
Northumberland Region:
Skerne River
2.56X101
mg/kg dw
Nicholls et al.
(2001)
5.84X101
mg/kg dw
Nicholls et al.
(2001)
3.22X101
mg/kg dw
Nicholls et al.
(2001)
East Anglia Region: Lark
River
4.50X101
mg/kg dw
Nicholls et al.
(2001)
6.04X101
mg/kg dw
Nicholls et al.
(2001)
Detroit River
6.80xl0"2
mg/kg dw
Tomy et al. (1999)
United States
Sugar Creek, Ohio
Reported as trace with range of
1.5-5; used the average
3.25 xlO"3
mg/kg dw
USEPA (1988)
Reported as trace with range of
1.5-5; used the average
3.25xl0"3
mg/kg dw
USEPA (1988)
Ill
-------
1 .OCIIIIOII
(mi\ cried
(\. Si;iic or h\>\ incc
( oinmenls
References
6.80xl0"3
mg/kg dw
USEPA (1988)
8.20xl0"3
mg/kg dw
USEPA (1988)
7.60X10"1
mg/kg dw
USEPA (1988)
2.10X101
mg/kg dw
USEPA (1988)
3.40X101
mg/kg dw
USEPA (1988)
5.00X101
mg/kg dw
USEPA (1988)
Note:
NR: Not recorded. Location description was not provided in the study.
: Single sample value reported above the detection limit; therefore, no data qualifier required.
1. All values provided in the table above represent total MCCP and not individual MCCP isomers
2. In some cases, the minimum values in the table are preceded by "<". This indicates that the value reported in article was reported as a non-detect. In such cases, one half of the
lowest reported detection limit was compiled as the 'minimum' reported monitoring data
3. dw - dry weight and ww - wet weight
112
-------
D-l-3 Biosolids and Soil
CPs are detected more frequently and at higher concentrations in treated sewage sludge {i.e.,
biosolids) than in soil. MCCP concentrations ranged from 5.00 x 10"5 to 9.70 x 103 mg/kg dw in
sludge and from 1.5 x 10"2 to 8.5 x 10"2 mg/kg dw in soil. It is unclear if the difference in MCCP
concentrations in sludge and soil is related to the smaller sample sizes for these media compared
to the typically larger data sets available for water and sediment. To determine the most reliable
studies for its consideration, EPA/OPPT used the following criteria: designation of specific
MCCP chain length and the appropriate analytical methodology. Based these criteria,
EPA/OPPT determined that the data reported by Stevens et al. (2003) were the most reliable for
its use in this assessment: data are summarized below.
Stevens et al. (2003) measured MCCP concentrations in sludge samples obtained from 14
WWTPs in the UK. MCCP concentrations ranged from 3.00 x 101 to 9.70 x 103 mg/kg dw. The
authors concluded that these very high concentrations were likely the result of releases from
numerous and ongoing diffuse sources. EPA/OPPT did not use information from other published
studies reporting measured CPs in sludge and soil because they did not distinguish the CPs
measured (Nicholls et al., 2001). These studies reported total CP concentrations at much lower
levels ranging from 3.00 x 10"5 to 2.3 mg/kg dw.
Although risk to terrestrial species was not calculated, EPA/OPPT notes that the lowest and
highest reported biosolid and soil concentrations (5.00 x 10"5 to 9.70 x 103 mg/kg dw and 1.5 x
10"2 to 8.5 x 10"2 mg/kg dw, respectively) represents a very large range (up to eight orders of
magnitude (TableApx D-l-3).
Table Apx D-l-3:
Jiosolid and Soil Concentrations of MCCPs
Location
Media
Concen
Minimum
tration
Maximum
Units
References
Switzerland
Soil
1.5 x 10"2
8.5 x 10"2
mg/kg dw
Iozza (2010)
China
Soil
2.1 x 10"6
1.53 x 10"3
mg/kg dw
Wang et al. (2013)
Czech Republic
Sewage
Sludge
7.36 x 10"1
2.30
mg/kg dw
Pribylova et al.
(2006)
United
Kingdom
Sewage
Sludge
3.00 x 101
9.70 x 103
mg/kg dw
Stevens et al. (2003)
United States
Sewage
Sludge
5.00 x 10"5
5.00 x 10"5
mg/kg dw
Pribylova et al.
(2006)
D-l-4
Biota
EPA/OPPT reviewed available published literature and summarized MCCP concentrations in
tissues of aquatic and terrestrial biota (Table Apx D-l-4). Measured tissue concentrations for
aquatic biota ranged from ND to 2.63 mg/kg ww (i.e., beluga whales, seals, rainbow trout, carp,
mackerel, arctic char, mussels, crustaceans, and plankton) and ranged from 5.00 x 10"3 to 3.70 x
10"1 mg/kg ww in terrestrial biota. The concentrations measured in the terrestrial studies did not
designate the specific CP congeners measured.
113
-------
As a result of EPA/OPPT's evaluation, MCCPs were found in organisms across many different
trophic levels indicating widespread environmental contamination (TableApx D-l-4). The data
were insufficient for EPA/OPPT to draw conclusions about trends based on region, species, time,
or other factors.
While EPA/OPPT determined the concentrations of MCCPs in aquatic and terrestrial biota range
from ND to 2.63 mg/kg ww from 5.00 x 10"3 to 3.70 x 10"1 mg/kg ww, respectively, in this
assessment, EPA/OPPT did not use tissue concentrations to determine risks of potential concern
for biota Table Apx D-l-4. Rather, it used the risk quotient (RQ) method as described in
Section 6.
114
-------
Table Apx D-l-4: Biota Concentrations of MCCPs
Location
Media
Description
Minimum
Units
Min Reference
Maximum
Units
Max Reference
Aquatic Biota
Australia
Invertebrates
2.32xl0"5
mg/kg lw
Kemmlein et al. (2002)
3.05xl0"5
mg/kg lw
Kemmlein et al. (2002)
Canada
Mammals
5.45xl0"7
mg/kg ww
Bennie et al. (2000)
8.00xl0"5
mg/kg ww
Bennie et al. (2000)
Fish
2.57xl0"7
mg/kg ww
Bennie et al. (2000)
2.63
mg/kg ww
Muir et al. (2000)
Invertebrates
ND1
mg/kg ww
EC (1993)
ND1
mg/kg ww
EC (1993)
Total
ND1
mg/kg ww
EC (1993)
2.63
mg/kg ww
Muir et al. (2000)
Europe
Fish
7.00X10-3
mg/kg ww
Reth et al. (2006)
4.70xl0"2
mg/kg ww
Reth et al. (2006)
North Sea/Baltic Sea
Region2
Fish
ND3
mg/kg ww
IVL (2009)
2.6X10"1
mg/kg ww
Reth et al. (2005)
United States
Fish
2.90xl0"3
mg/kg ww
Tomy and Stern (1999)
9.04X10"1
mg/kg ww
Tomy and Stern (1999)
Invertebrates
3.50X10-3
mg/kg ww
USEPA (1988)
1.70X10"1
mg/kg ww
USEPA (1988)
Total
2.90X10-3
mg/kg ww
Tomy and Stern (1999)
9.04X10"1
mg/kg ww
Tomy and Stern (1999)
United States / Canada
- Great Lakes
Fish
1.80x10-3
mg/kg ww
Muir et al. (2003)
l.lOxlO"1
mg/kg ww
Muir et al. (2003)
Invertebrates
2.40x10-3
mg/kg ww
EC (2008a)
1.60xl0"2
mg/kg ww
Muir et al. (2003)
Total
1.80x10-3
mg/kg ww
Muir et al. (2003)
l.lOxlO"1
mg/kg ww.
Muir et al. (2003)
Terrestrial Biota
Europe
Birds
5.00x10-3
mg/kg ww
Reth et al. (2006)
3.70X10"1
mg/kg ww
Reth et al. (2006)
Notes:
Summary values represent total MCCP and not individual MCCP isomers.
1 MCCPs were not detected in Invertebrates from Canada. Detection limit = 4.0 x 10"7 mg/kg; 'A DL = 2.0 x 10"7 (EC, 1993).
2 North Sea/Baltic Sea Region includes the following countries: Estonia, Latvia, Lithuania, Norway, Poland, and Sweden.
3 The minimum MCCP concentration value for fish from the North Sea/Baltic Sea Region was non-detect. The detection limit = 2.5 x 10"4 mg/kg; Vi DL = 1.25 x 10"4 (IVL 2009)
115
-------
D-2
LCCP MONITORING DATA
Kemmlein et al. (2002) optimized and tested the carbon skeleton reaction gas chromatography
analytical method to analyze environmental samples for CPs. The optimized method was used
for marine sediments, mussels and crabs taken from an area influenced by a CP manufacturer in
Yarraville, Australia. LCCP (C18-20) concentrations in marine sediment ranged from 1.02 x 10"1
to 4.31 x 10"1 mg/kg dw (Table_Apx D-2-1), those in mussels ranged from 4 x 10"1 to 1.9 mg/kg
lw, and those in crab ranged from 3 x 10"2 to 4.4 mg/kg lw. The results presented in this paper
show that bioaccumulation is evident. The mussel samples contained approximately two times
and crab tissue around six times the concentration of CPs found in the most contaminated
sediment sample. No other adequate studies were found to characterize LCCP concentrations in
surface water, fresh water sediment or soil.
Table Apx D-2-1: Marine Sediment Concentrations of LCCPs
Location
Media
Country
City, State
or
Province
Comments
Concentration
Units
References
Sum
1.02X10"1
mg/kg dw
Kemmlein et
al. (2002)
Sediment
Australia
NR
Sum
1.28X10"1
mg/kg dw
Kemmlein et
al. (2002)
(Marine)
Sum
3.04X10"1
mg/kg dw
Kemmlein et
al. (2002)
Sum
4.31X10"1
mg/kg dw
Kemmlein et
al. (2002)
Motes:
1. Values provided in the table above represent total LCCP (C18-20) and not individual isomers.
2. dw-dry weight
Table Apx
)-2-2: Biota Concentrations of LCCPs
Location
Media
Country
City, State
or
Province
Minimum
Maximum
Units
References
Aquatic
Biota
Australia
NR
2.89xl0"6
6.90xl0"6
mg/kg lw
Kemmlein et
al. (2002)
Notes:
1. Minimum and Maximum concentrations provided in the table above represent total LCCP (C18-20) and not individual
isomers.
2. lw - lipid weight
116
-------
Appendix E ENGINEERING (ChemSTEER) REPORTS ON P-12-
0-0433 and P-12-0453
(Used for both identifying potential releases to the environment and for estimated
occupational exposures. SEE APPENDIX G FOR REFERENCE TO FULL REPORTS
UNDER SEPARATE COVER)
P-12-0433
Smiiirio
# Siics
# Workers
PROC: Formulation of
Metalworking Fluids
3
24
l''.\|)OSIIIV
inu/(l:i>
(l;i\/\ r
Inhalation -
negligible (VP < 0.001 torr)
Dermal - Liquid
High End
1.8E+3
27
Dermal - Liquid
High End
3.5E+2
27
Kok'iisc
kti/sik'/(l;i\
r
Water or Incineration or
Landfill
High End
8.6E+0
27
Water or Incineration or
Landfill
Conservative
5.7E+0
27
Air
Suggested
Estimate
5.7E-1
27
Incineration or Landfill
Suggested
Estimate
5.7E+0
27
Scenario
# Silcs
# Workers
I SI! I so of \1clal\uii'kiim I'luids
4
i<>:
r.\i)(isuiv
-------
P-12-0453
Scenario
# Silos
# Workers
PROC2. PYC ( niiipuMiidiim ( l<> S" ol P\ )
s
i<>:
r.\i)(isuiv
d;i\/\ r
Inhalation - Particulate
High End of Range /
Low End of Range
4.4E+0 / 3.0E-2
126
Dermal - Liquid
High End
1.8E+3
126
Kcloiiso
Uli/sik'/(l;i\
(l.i\/\ r
Water
Output 2
4.3E-3
126
Water or Incineration
Output 1
8.5E+0
126
Water or Incineration or Landfill
Output 1
1.3E+1
126
Water or Incineration or Landfill
Suggested Value
4.3E-2
126
Air
Output 2
4.3E-3
126
Scenario
# Silos
# Workers
I SI!2 I'N'C (\iii\ei'liim ( l(>
l'\ )
s
'X4
l'l\|)osiire
(l;i\/\ r
Inhalation - Particulate
OSHA PEL /
EU Data from
Submitter
2.2E+1 /
1.2E+1
250
Dermal -
PMN will be encapsulated in plastic pellets during handling of plastic raw material.
While some surface contact may occur, dermal exposure to solids in this form are
non-quantifiable (2004 ESD; Cast Solids, CEB Method for Screening-Level
Assessments of Dermal Exposure).
liok'iiso
kii/sik'/(l;i\
(l;i\/\ r
Water
Suggested
Estimate
1.6E-1
250
Water or Incineration or
Suggested
2.1E-2
250
Landfill
Estimate
Water or Incineration or
Output 2
2.1E+0
250
Landfill
Water or Incineration or
Conservative
4.3E+0
250
Landfill
Water or Landfill
Suggested
Estimate
5.3E+0
250
Air
Suggested
Estimate
1.6E-1
250
Air
Output 2
2.4E-2
250
Landfill
Output 2
2.4E+0
250
118
-------
Scenario
# Siics
# Workers
PROC1: Formulation of
Metal working Fluids (73.8% of PV)
59
472
I'1\|)osiiiv
lปU/(lil>
(l;i\/\ r
Inhalation -
negligible (VP < 0.001 torr)
Dermal - Liquid
High End
1.8E+3
84
Dermal - Liquid
High End
3.5E+2
84
Kolciisc
Uli/sik'/(l;i\
(l;i\/\ r
Water or Incineration or
High End
1.1E+1
84
Landfill
Water or Incineration or
Conservative
7.6E+0
84
Landfill
Air
Suggested
Estimate
1.9E-1
84
Incineration or Landfill
Suggested
Estimate
7.6E+0
84
Scenario
# Silos
# Workers
I SI! 1 l se of Melalwoikiim I'luids iS",,of I'N'i
:
-------
Scenario
# Silos
# Workers
I sn l sc nl" \dlk.'M\cs and Scalauls I'M"., of I'Y)
5X
:_s4
r.\i)(isuiv
r
Inhalation - Mist
What-If
2.3E+1
250
Dermal - Liquid
High End
5.3E+2
250
Ucloiiso
k*i/sile/ซl:i>
(l.i\/\ r
Water or Incineration or Landfill
High End
1.9E+U
00
Water or Incineration or Landfill
Conservative
3.3E-1
250
Air
Output 2
1.2E+0
250
Landfill
Output 2
1.1E+1
250
120
-------
Appendix F EXPOSURE SCENARIO ESTIMATES
(E-FAST Model Run. SEE APPENDIX G FOR REFERENCE TO FULL REPORTS
UNDER SEPARATE COVER)
INEOS: P12-0433 Exposure 1
Exposure
Scenario 1
Water
Landfil
1
Stack Air
Fugitive Air
Drinking Water
Fish Ingestion
ADR
LADD
ADR
LADD
7Q104
COC =
1
PDM
Days
Exceede
d
LADD
ADR
LADD
ADR
LADD
Release
activity(ies)
2
exposure
calculation(
s)3
mg/kg/day
mg/kg/da
y
mg/kg/da
y
mg/kg/da
y
ng/i
# Days
mg/kg/da
y
mg/kg/da
y
mg/kg/da
y
mg/kg/da
y
mg/kg/da
y
Proc: Max
ADR,
PDM,
LADD
8.99E-03
2.08E-
05
6.49E-
02
8.93E-
05
184.28
27
1.12E-04
1.43E-
02
3.39E-
05
USE: Max
ADR,
PDM1
3.90E-02
0.17
829.55
139
2.76E-
02
USE:
PDM2
784.09
247
USE: Max
LADD
4.75E-
04
2.04E-
03
9.71E-
04
1.26E-
03
5.99E-
04
Exposure scenario titles consist of release activity followed by exposure calculation abbreviation.
2Release activities are from engineering report's Manufacturing (Mfg), Processing (Proc) and Use release activity
labels. Multiple release activities are combined in one exposure scenario if their releases occur at same location.
3Exposure calculations are Acute Dose Rate (ADR), Lifetime Average Daily Dose (LADD), and Probabilistic
Dilution Model (PDM). There may be one, two, or all three exposure calculations per exposure scenario. COC is
the aquatic concentration of concern.
4This column displays concentration values for the 7Q10 streamflow, which is defined as the average streamflow
of the 7 consecutive days of lowest flow within a 10 year period.
121
-------
INEOS: P12-0453 Exposure 1
Exposure
Scenario 1
Water
Landfill
Stack Air
Fugitive Air
Drinking Water
Fish Ingestion
ADR
LADD
ADR
LADD
7Q104
COC
= 1
PDM
Days
Exceeded
LADD
ADR
LADD
ADR
LADD
Release
activity(ies)2;
exposure
calculation^)^
mg/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
ng/i
# Days
mg/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
PROC1: Max
ADR, PDM,
LADD
1.17E-02
8.43E-05
0.24
1.03E-03
239.69
84
4.58E-04
3.31E-02
5.97E-04
4.78E-03
3.52E-
05
USE1: Max
ADR, PDM1
6.08E-02
0.74
1291.67
218
3.68E-02
4.60E-02
USE1: PDM2
1246.21
247
USE1: Max
LADD
7.55E-04
9.24E-03
1.55E-03
2.01E-03
9.78E-
04
PROC2: Max
ADR, PDM,
LADD
1.35E-
02
1.46E-04
0.28
1.79E-03
277.67
126
3.42E-04
2.76E-02
7.35E-04
1.08E-04
1.19E-
06
USE2: Max
ADR, PDM,
LADD
7.24E-
03
1.43E-04
0.14
1.74E-03
148.14
248
7.35E-04
7.94E-03
4.35E-04
4.61E-03
1.01E-
04
PROC3: Max
ADR, Max
acute eco
0.14
2.85
2835.05
0.27
PROC3: PDM
257.73
200
PROC3: Max
LADD
2.59E-04
3.17E-03
1.00E-03
1.30E-03
USE3: Max
ADR, PDM1
9.61E-
03
0.10
210.38
64
2.76E-03
2.94E-02
USE3: PDM2
31.13
186
USE3: Max
LADD
4.00E-05
4.89E-04
6.16E-04
5.64E-05
6.60E-
04
Exposure scenario titles consist of release activity followed by exposure calculation abbreviation.
2Release activities are from engineering report's Manufacturing (Mfg), Processing (Proc) and Use release activity labels.
Multiple release activities are combined in one exposure scenario if their releases occur at same location.
3Exposure calculations are Acute Dose Rate (ADR), Lifetime Average Daily Dose (LADD), and Probabilistic Dilution
Model (PDM). There may be one, two, or all three exposure calculations per exposure scenario. COC is the aquatic
concentration of concern.
4This column displays concentration values for the 7Q10 streamflow, which is defined as the average streamflow of the
7 consecutive days of lowest flow within a 10 year period.
122
-------
Appendix G SUPPLEMENTAL INFORMATION
SUPPLEMENTAL DATA FOR APPENDIX E (ChemSteer Engineering Reports):
pl20433.ceb - 23 page pdf file
pl20453.ceb - 60 page pdf file
SUPPLEMENTAL DATA FOR APPENDIX F (E-FAST Exposure Reports):
P-12-043 3. exp 1 Draft Final_REVISED_022013 - 9-page pdf file
P-12-0453.expl_Draft Final_102512 - 21-page pdf file
123
------- |