Screening Assessment for the Challenge (2024)

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Urea,N'-(3,4-dichlorophenyl)-N,N-dimethyl-
(Diuron)

Chemical Abstracts Service Registry Number
330-54-1

Environment Canada
Health Canada

January 2011

(PDF Version - 589 KB)

Table of Contents

  • Synopsis
  • Introduction
  • Substance Identity
  • Physical and Chemical Properties
  • Sources
  • Uses
  • Summary of Information Used as Basis for Environmental Screening Assessment
  • Environmental Fate
  • Persistence and Bioaccumulation Potential
  • Potential to Cause Ecological Harm
  • Summary of Information Used as Basis for Human Health Risk Characterization
  • Conclusion
  • References
  • Appendix 1: Robust Study Summaries
  • Appendix 2: Persistence Bioaccumulation and Toxicity Model Inputs Summary Table
  • Appendix 3: Summary of Health Effects Information for CASRN 330-54-1: Urea,N′-(3,4-dichlorophenyl)-N,N-dimethyl- (Diuron)
  • Appendix 4: Summary of Available Reference Values for Diuron

Synopsis

Pursuant to section 74 of the Canadian Environmental Protection Act, 1999 (CEPA 1999), the Ministers of the Environment and of Health have conducted a screening assessment of Urea,N'-(3,4-dichlorophenyl)-N,N-dimethyl- Chemical Abstracts Service Registry Number (CAS[*] RN) 330-54-1. This substance will be referred to by its common name, diuron, in this screening assessment. Diuron was identified in the categorization of the Domestic Substances List as a high priority for action in the Challenge initiative under the Chemicals Management Plan . Diuron was identified as a high priority as it was considered to pose an intermediate potential for exposure of individuals in Canada and is classified by other agencies on the basis of carcinogenicity. The substance met the ecological categorization criteria for persistence and for inherent toxicity to aquatic organisms.

Diuron is an organic substance and according to information submitted under section 71 of CEPA 1999, it was imported into Canada for pesticidal and non-pesticidal uses above the reporting threshold of 100 kg in 2006.

Diuron is a registered active ingredient in pest control products and these uses are regulated by Health Canada’s Pest Management Regulatory Agency (PMRA) under the Pest Control Products Act (PCPA). In 2007, Health Canada’s PMRA re-evaluated pesticidal uses of diuron under Re-evaluation Program 1 and concluded that it is acceptable for continued registration. Information from the re-evaluation of pesticidal uses of diuron was taken into consideration in the current assessment of non-pesticidal uses under CEPA 1999.

This screening assessment considers the potential effects of diuron on human health and the environment as a result of non-pesticidal uses of the substance. Diuron was imported in non-pesticidal products, which were considered industrial in nature, including a hardening agent in epoxy resins and a curing agent in epoxy adhesive for bonding of metal parts. Exposure to the general population from non-pesticidal, industrial uses of diuron is considered to be negligible.

As diuron was classified on the basis of carcinogenicity by other national and international agencies, carcinogenicity was a key focus for the human health portion of this screening assessment. Exposure to diuron, via the diet, increased the incidence of urinary bladder carcinomas in both male and female Wistar rats, while mammary gland adenocarcinomas were noted in NMRI mice. The increases in tumour incidences were noted at high dose levels, which approached maximally tolerated doses. Results of genotoxicity testing, performed both in vitro and in vivo, were predominantly negative. Consideration of the available information regarding genotoxicity, and assessments of other agencies, indicate that diuron is not likely to be genotoxic. Accordingly, although the mode-of-induction of tumours is not fully elucidated, the tumours observed are not considered to have resulted from direct interaction with genetic material.

Consistent non-neoplastic effects were observed in rat, mouse and dog. These included erythrocyte damage with compensatory haematopoiesis. Urinary bladder wall thickening and swelling, together with epithelial focal hyperplasia, were observed in the high-dose groups in rat and mouse bioassays.

As general population exposure to diuron as a result of non-pesticidal, industrial uses is considered to be negligible, it is concluded that diuron is a substance that is not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health.

The ecological components of this assessment build upon the assessment work conducted by Environment Canada following the categorization of substances under CEPA 1999 and also include applicable information from the US EPA’s and PMRA’s assessments (i.e., Re-registration Eligibility Decision for Diuron, dated 2003, and Re-evaluation Decision Diuron, dated 2007). In addition, other information from the European Commission Joint Research Centre and recent scientific literature available in the public domain were considered. Finally, discussion of major metabolites of diuron, and in particular 3,4-DCA, is provided to assist in understanding the overall environmental impacts of the substance.

Although diuron is expected to be persistent in water and aerobic soil and sediment, it has a low potential to bioaccumulate. Diuron also exhibits relatively high potential for toxicity to sensitive aquatic organisms. A risk quotient analysis, integrating a conservative predicted environmental concentration with a predicted no-effect concentration, indicated that the current concentrations of diuron in water are unlikely to cause ecological harm in Canada. Due mainly to the estimated low releases of this substance to water as a result of current non-pesticidal uses, it is concluded that diuron is not entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity, or that constitute or may constitute a danger to the environment upon which life depends. Diuron meets the criteria for persistence, but does not meet criteria for bioaccumulation as set out in the Persistence and Bioaccumulation Regulations.

It is therefore concluded that diuron does not meet any of the criteria set out in section 64 of CEPA 1999.

This substance will be considered for inclusion in the Domestic Substances List inventory update initiative. In addition and where relevant, research and monitoring will support verification of assumptions used during the screening assessment.

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Introduction

The Canadian Environmental Protection Act, 1999(CEPA 1999) (Canada 1999) requires the Minister of the Environment and the Minister of Health to conduct screening assessments of substances that have met the categorization criteria set out in the Act to determine whether these substances present or may present a risk to the environment or to human health.

Based on the information obtained through the categorization process, the Ministers identified a number of substances as high priorities for action. These include substances that:

  • met all of the ecological categorization criteria, including persistence (P), bioaccumulation potential (B) and inherent toxicity to aquatic organisms (iT), and were believed to be in commerce in Canada; and/or
  • met the categorization criteria for greatest potential for exposure (GPE) or presented an intermediate potential for exposure (IPE) and had been identified as posing a high hazard to human health based on classifications by other national or international agencies for carcinogenicity, genotoxicity, developmental toxicity or reproductive toxicity.

The Ministers therefore published a notice of intent in theCanada Gazette, Part I, on December 9, 2006 (Canada 2006) that challenged industry and other interested stakeholders to submit, within specified timelines, specific information that may be used to inform risk assessment and to develop and benchmark best practices for the risk management and product stewardship of those substances identified as high priorities.

The substance Urea,N'-(3,4-dichlorophenyl)-N,N-dimethyl- was identified as a high priority for assessment of human health risk because it was considered to present IPE and had been classified by the European Union on the basis of carcinogenicity (European Commission 2000, 2001). Furthermore, the US EPA’s carcinogenicity peer review committee has classified diuron as a “known/likely” carcinogen (US EPA 1997). Diuron also met the ecological categorization criterion for persistence and inherent toxicity to aquatic organisms. The Challenge for this substance was published in the Canada Gazette on June 20, 2009 (Canada 2009). A substance profile was released at the same time. The substance profile presented the technical information available before December 2005 that formed the basis for categorization of this substance. As a result of the Challenge, submissions of information pertaining to the substance were received.

For a substance to be imported, manufactured or used in Canada, it must be listed on CEPA 1999’s Domestic Substances List (DSL) or regulated under another federal government act that is scheduled under CEPA 1999. The Pest Control Products Act (PCPA), which is administered by Health Canada’s Pest Management Regulatory Agency (PMRA), is scheduled under CEPA 1999. Pesticides must undergo a pre-market environmental and human health risk assessment by PMRA.

Diuron is a registered active ingredient in pest control products under the PCPA. Health Canada’s PMRA recently re-evaluated the pesticidal uses of diuron and concluded that it is acceptable for continued registration with the implementation of specific mitigation measures as specified in the Re-evaluation Decision Document (Health Canada 2007). Currently six end-use products containing the active ingredient, diuron, are registered in Canada (PMRA 2010).

For the screening assessment under CEPA 1999 of a registered pesticide on the DSL, the approach of Environment Canada and Health Canada is to conduct an entry characterization of the substance in Canada focusing on any non-pesticidal releases and sources.

Screening assessments focus on information critical to determining whether a substance meets the criteria as set out in section 64 of CEPA 1999. Screening assessments examine scientific information and develop conclusions by incorporating a weight-of-evidence approach and precaution[a].

This screening assessment includes consideration of information on chemical properties, hazards, uses and exposure, including the additional information submitted under the Challenge. Data relevant to the entry characterization of this substance were identified in original literature, review and assessment documents and stakeholder research reports. Recent literature searches were performed up to February 2010 for the human exposure and human health effects sections of the document, and up to April 2010 for the ecological sections of the document. Key studies were critically evaluated; modelling results may have been used to reach conclusions.

Evaluation of risk to human health involves consideration of data relevant to estimation of exposure (non-occupational) of the general population to the non-pesticidal uses, as well as information on health hazards (based principally on the weight-of-evidence assessments of other agencies that were used for prioritization of the substance). Decisions for human health are based on the nature of the critical effect and/or margins between conservative effect levels and estimates of exposure, taking into account confidence in the completeness of the identified databases on both exposure and effects, within a screening context. The final screening assessment does not represent an exhaustive or critical review of all available data. Rather, it presents a summary of the critical information upon which the conclusion is based.

This final screening assessment was prepared by staff in the Existing Substances Programs at Health Canada and Environment Canada and incorporates input from other programs within these departments.

The ecological and human health portions of this assessment have undergone external written peer review and consultation. Comments on the technical portions relevant to human health were received from scientific experts selected and directed by Toxicology Excellence for Risk Assessment (TERA) and included comments by Dr. Bernard Gadagbui (TERA), Dr. Michael Jayjock (The Lifeline Group) and Dr. Chris Bevans (CJB Consulting).

The critical information and considerations upon which the final assessment is based are summarized below.

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Substance Identity

Substance Name

For the purposes of this document, this substance will be referred to as diuron, derived from the DSL, EINECS, ENCS, SWISS and PICCS inventories (see Table 1 for definitions of abbreviations).

Table 1. Substance identity for diuron

Chemical Abstracts Service Registry Number (CAS RN) 330-54-1
DSL name Urea, N'-(3,4-dichlorophenyl)-N,N-dimethyl-
National Chemical Inventories (NCI) namesa

Urea, N'-(3,4-dichlorophenyl)-N,N-dimethyl- (TSCA, DSL, ENCS, PICCS, ASIA-PAC, NZIoC); Diuron (English, French, German, Spanish) (DSL, EINECS, ENCS, SWISS, PICCS);

3-(3,4-Dichlorophenyl)-1,1-dimethylurea (ENCS);DCMU (ENCS); Urea, N'-(3,4-dichlorophenyl)-N,N-dimethyl- (AICS);

N'-(3,4-Dichlorophenyl)-N,N-dimethyl urea (ECL);

Urea, N'-(3,4-dichlorophenyl)-N,N-dimethyl-(SWISS);

N'-3,4-DICHLOROPHENYL N,N-DIMETHYLUREA (PICCS);

3-(3,4-DICHLOROPHENYL)-1,1-DIMETHYL UREA (PICCS);

(3-(3,4-DICHLOROPHENYL)-1,1-DIMETHYL UREA (PICCS);

1,1-DIMETHYL-3-(3,4-DICHLOROPHENYL)UREA (PICCS)

Other names

1-(3,4-Dichlorophenyl)-3,3-dimethylurea; Dairon; DCMC; DCMU 9; DCMU 99; Direx; Dironet; Dironzol;

Diuron Nortox; DMU; DP Hardener 95; Duran;

Dyhard UR 200; Herbatox; HRT Dinron; HW 920; Karmax;

Karmex; Karmex D; Karmex Diuron Herbicide;

Karmex DW; Lucenit; Marmer; N'-(3,4-Dichlorophenyl)-N,N-dimethylurea; N,N-Dimethyl-N'-(3,4-dichlorophenyl)urea;

N-(3,4-Dichlorophenyl)-N',N'-dimethylurea; NSC 8950; Preventol A 6; Telvar Diuron Weed Killer;

Urea, 3-(3,4-dichlorophenyl)-1,1-dimethyl-

Chemical group

(DSL Stream)

Discrete organics
Major chemical class or use Amines
Major chemical sub-class Tertiary amines, aliphatic amines, anilines, secondary aromatic amines
Chemical formula C9H10Cl2N2O
Chemical structure Screening Assessment for the Challenge (1)
SMILESb O=C(N(C)C)Nc(ccc(c1Cl)Cl)c1

Molecular mass

(g/mol)

233.1

a National Chemical Inventories (NCI) 2007: AICS, Australian Inventory of Chemical Substances; ASIA-PAC, Asia-Pacific Substances Lists; DSL, Canadian Domestic Substances List; ECL,Korean Existing Chemicals List; EINECS, European Inventory of Existing Commercial Chemical Substances; ENCS, Japanese Existing and New Chemical Substances; NZIoC, New Zealand Inventory of Chemicals; PICCS, Philippine Inventory of Chemicals and Chemical Substances; SWISS, Inventory of Notified New Substances, List of Toxic Substances 1; TSCA, Toxic Substances Control Act Chemical Substance Inventory.
b Simplified Molecular Input Line Entry System.

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Physical and Chemical Properties

The physical and chemical properties of diuron relevant to its environmental fate are described in Table 2. The models based on quantitative structure–activity relationships (QSAR) were used to generate data for some of the physical and chemical properties of diuron. These models are mainly based on fragment addition methods, i.e., they rely on the structure of a chemical. Since these models only accept the neutral form of a chemical as input (in SMILES form), the modelled values shown in Table 2 are for the neutral form of diuron. The neutral form of the substance is expected to be predominant at environmentally relevant pH.

Table 2. Physical and chemical properties of diuron

Property Type Valuea Temperature
(°C)
Reference
Melting point
(ºC)
Experimental 158–159 Tomlin
2005–2006
Modelled 126.39 MPBPWIN 2008
Boiling point
(ºC)
Modelled 353.86 MPBPWIN 2008
Density
(kg/m3)
Experimental ~450 Bayer AG in ESIS 1995–2009
Vapour pressure
(Pa)
Experimental 1.1 10-7
(1.1 × 10-4 mPa)
25 Tomlin
2005-–2006
9.2 10-6
(9.2 × 10-3 mPa)
DuPont 1989
in PPD 2009
2.3 10-7
(2.3 × 10-9 hPa)
20 Bayer AG in ESIS 1995–2009
Modelled 6.2 10-4 25 MPBPWIN 2008
Henry’s Law constant
(Pa·m3/mol)
Calculated 7.04 × 10-6 Tomlin 2005–2006
Calculated 5.11 10-5 25 HENRYWIN 2008
Modelled 5.40 10-5 25
Log Kow
(octanol–water partition coefficient)
(dimensionless)
Experimental 2.85 0.03 25 Tomlin
2005–2006
Experimental 2.68 Hansch et al. 1995
Log Koc
(organic carbon–water partition coefficient)
(dimensionless)
Experimental 2.4 0.2 Thomas et al. 2002
2.6 (Koc 400) Tomlin
2005–2006
Modelled
(estimated from MCIb)
2.04 PCKOCWIN 2008
Modelled
(estimated from log Kow)
2.33
Water solubility
(mg/L)
Experimental 42 25 DuPont 1989 in PPD 2009
35
(0.035 g/L)
20 Bayer AG in ESIS 1995–2009
36.4 25 Tomlin
2005–2006
Modelled 150.6 25 WSKOWWIN 2008
Solubility in acetone (mg/L)c Experimental 4.19 104
(53 g/kg)
27 Tomlin
2005–2006
Solubility in butyl stearate (mg/L)d 1.32 103
(1.4 g/kg)
Solubility in benzene (mg/L)e 1.55 103
(1.2 g/kg)
pKa (acid dissociation constant) (dimensionless) Modelled 13.55
(pKa1)
-1.09
(pKa2)
-2.48
(pKa3)
ACD/pKaDB 2005

Abbreviations: Koc, organic carbon–water partition coefficient; Kow, octanol–water partition coefficient.
a Values in parentheses represent the original values as reported by the authors or as estimated by the models.
b First-order molecular connectivity index.
c Density of acetone = 0.791 kg/L (CRC Handbook of Chemistry and Physics 1965–1966).
d Density of butyl stearate = 0.941 kg/L (CRC Handbook of Chemistry and Physics 1965–1966).
e Density of benzene = 0.879 kg/L (CRC Handbook of Chemistry and Physics 1965–1966).

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Sources

Recent Manufacture and Importation Information

The substance diuron is not naturally produced in the environment. Based on information collected through a survey conducted pursuant to section 71 of CEPA 1999 (Environment Canada 2009b), between 100 000 and 1 000 000 kg of diuron were imported into Canada in 2006 for both pesticidal and non-pesticidal industrial uses. The substance was not reported to be manufactured in Canada. Six companies identified stakeholder interest in diuron.

Previously received information from the Domestic Substances List nomination (1984–1986) showed that the total quantity of diuron reported as imported into, manufactured in or in commerce in Canada during the calendar year 1986 was 1 000 000 – 10 000 000 kg (Environment Canada 1988).

Elsewhere, diuron has been identified as a high production volume (HPV) chemical on the lists from the following organizations: the Organisation for Economic Co-operation and Development (OECD) (OECD 2004a) and the European Union (EU) (European Commission 2010).

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Uses

Diuron is a registered active ingredient in pest control products (PMRA 2010). Diuron is an herbicide used to control weeds in food crops (grapes and asparagus) and non-cropland areas (including industrial sites and irrigation and drainage ditches) and is also used to control algae in ponds and dugouts (Health Canada 2007).

Based on an industrial survey conducted under section 71 ofCEPA 1999 (Canada 2009), the majority of diuron (96%) imported into Canada in 2006 was for its use as a pesticide. The remaining 4% of diuron was imported for non-pesticidal uses including a hardening agent for epoxy resins, as a curing agent in epoxy adhesive for bonding of metal parts, and for uses in manufactured items (such as industrial items used in the transportation industry). These non-pesticidal uses of diuron in Canada are industrial in nature. One company intends to discontinue their non-pesticidal use of diuron by the end of 2010.

However, potential for environmental impacts of diuron resulting from all non-pesticidal uses identified for the year 2006 (Environment Canada 2009b) were taken into consideration in this assessment.

Outside of Canada, diuron is used primarily as an herbicide, although it may have other pesticidal uses such as a biocide for non-agricultural use (Gangolli 1999, Spectrum Laboratories 2009, European Commission 2010, Lewis 2007, SPIN 2006).

Finally, diuron may also be used in applications described by the following use categories: impregnation or proofing materials (SPIN 2006); paints and varnishes (water-based decorative and exterior protection, volatile organic thinner, active corrosion inhibitor); coatings; printing ink and mastics; lacquers; adhesives (binding agents and glues); process regulators (building materials and additives); solvents (SPIN 2006, MSDS 2009a, MSDS 2009b, MSDS 2008a,US EPA 2003); cross-linking catalyst (Gangolli 1999); curing agent for epoxy resins and adhesives (MSDS 2001, MSDS 2003); and it is found in trace amounts in dyes used for re-colouring plastic bumper bars in automobiles (MSDS 2008b).

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Summary of Information Used as Basis for Environmental Screening Assessment

Releases to and Presence in the Environment

Reports describing environmental monitoring data of diuron resulting from non-pesticidal use and releases of the substance were not identified. Therefore, environmental concentrations of diuron due specifically to releases from industrial applications are presently unknown.

However, reports of measured environmental concentrations of diuron in surface and ground waters in Canada and elsewhere from pesticidal applications were identified (Claver et al. 2006, Thomas 2001, Thomas et al. 2002, US EPA 2002, Hiebsch 1988, US EPA 2001, Health Canada 1989, Anderson 2005, Cross 2000, Giroux 1995, Giroux 1998, Tellier 2006, Environment Canada 2007a, Environment Canada 2008, Environment Canada 2009a). The environmental risk assessment of diuron resulting from pesticidal uses was addressed by PMRA through its re-evaluation program, under authority of thePCPA (Health Canada 2007).

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Environmental Fate

Based on its physical and chemical properties (Table 2), the results of Level III fugacity modelling (Table 3) suggest that diuron is expected to predominantly reside in water and soil, depending on the compartment of release.

Table 3. Results of the Level III fugacity modelling (EQC 2003)

Substance released to: Percentage of substance partitioning into each compartment
Air Water Soil Sediment
Air (100%) 0 5.9 94.1 0
Water (100%) 0 99.7 0 0.3
Soil (100%) 0 4.1 95.9 0

If released to air, a negligible amount of the substance is expected to reside in air (see Table 3). Based on the low experimental vapour pressure of 10-7 to 10-6Pa and calculated Henry's Law constant of 7.04 × 10-6Pa·m3/mol, diuron is non-volatile. Therefore, if released solely to air the major two compartments into which this substance will partition are soil (>90%) and, to a lesser extent, water (<10 %), see Table 3.

If released into water, diuron is not expected to adsorb in large amounts to suspended solids and sediment based upon its low experimental log Koc value of 2.4. Volatilization from water surfaces is expected to be negligible, based upon this compound's calculated Henry's Law constant. Thus, if water is a receiving medium, diuron is expected to mainly reside in water (>99%), and to a very small extent partition into sediment (<1%) (see Table 3).

If released to soil, diuron is expected to have low to moderate adsorptivity to soil based upon its experimental and modelled log Koc (see Table 2). Volatilization from moist soil surfaces will likely be minimal based on its relatively low Henry's Law constant. This chemical is unlikely to volatilize from dry soil surfaces based upon its low vapour pressure. Therefore, if released to soil, diuron will mostly remain in soil and to a much lesser extent partition into water, which is illustrated by the results of the Level III fugacity modelling (see Table 3).

In addition, diuron is considered to be non-ionizing, as predicted by the modelling program pKa dB (ACD/pKaDB, 2005). The acid dissociation constants of diuron in water were calculated as pKa1 = 13.55 (an acid), pKa2 = -1.09 (base) and pKa3 = -2.48 (base) (see Table 2), which indicate that diuron is not expected to ionize in water at environmentally relevant pH.

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Persistence and Bioaccumulation Potential

Environmental Persistence

Empirical biodegradation data (CHRIP c2008, Table 4a) indicate zero percent biodegradation of diuron over 28 days in a ready-biodegradation test (test method MITI-I [OECD TG 301C]). In addition, Thomas et al. (2002) demonstrated that diuron is extremely persistent in seawater where it does not biodegrade under exposure to constant light over 42 days. In addition, photodegradation of diuron in natural seawater under ultraviolet (UV) sunlight appeared to be limited (Okamura 2002). In water, degradation of diuron can also occur through an irreversible hydrolysis of the ureide group to produce 3,4- DCA (Salvestrini et al. 2002). However, this reaction is reported to be very slow in the pH range of 6–9, which is typical of Canadian surface waters. These experimental data suggest that the half-life of diuron in water is likely to be longer than 182 days (6 months) and that the substance is therefore likely to persist in that environmental compartment.

Experiments in marine sediments indicate that diuron is subject to more rapid primary degradation under anaerobic conditions. When tested over 42 days, approximately 45% of diuron biodegraded (Thomas et al. 2003), indicating a half-life of 14 days for the substance in this environmental compartment (assuming that diuron degradation follows a pseudo first-order kinetic model). The main anaerobic degradation product was reported to be 1-(3,4-dichlorophenyl)-3-methylurea. However, both field tests and laboratory experiments have shown that diuron has a low tendency to partition onto sedimentary material. Therefore, the water column is the primary compartment of interest in relation to the persistence of this substance in the aquatic environment.

Degradation of diuron has also been studied in different types of soil. It follows that the rate of loss of diuron is affected by its concentration and the experimental temperature as well as the soil type (Madhun and Freed 1987). Madhun and Freed (1987) determined, based on application of 5 and 100 µmol/kg diuron and measurement of CO 2 evolution, that the ultimate half-life of diuron in sandy loam soil ranged from about 200 to 1400 days, whereas in mucky peat soil the half-life ranged from about 800 to 4000 days. Half-lives decreased with increased experimental temperature. Enhanced or accelerated primary biodegradation of diuron was observed in soil previously treated with the substance (Rouchaud et al. 2000). The chemical structure of the herbicide is thought to influence the occurrence of accelerated biodegradation, as this is not observed with all herbicides. When diuron was tested in the top 10 cm of loam soil at an application rate of 3 kg/ha per year, a primary degradation half-life of 80.7 days was observed in non-treated soil, whereas after 12 years of consecutive annual applications of diuron, the half-life decreased to 37 days. The products of diuron degradation were not identified in the study by Rouchaud et al. (2000).

The degradation half-life of diuron in the field was reported as 100 days (WSSA Herbicide Handbook 1989) and 90 days (Wauchope et al. 1992, WSSA Herbicide Handbook 1994) although study conditions, including incubation temperature, type of soil and the degradation endpoint (primary versus ultimate) are unknown for these results.

Microbial degradation is thought to be the primary mechanism of diuron dissipation from soil (Tixier et al. 2000). In well-oxygenated brown calcareous soil studied under field conditions, degradation is believed to occur through successive demethylation of the diuron urea group, followed by hydrolysis to produce a chlorinated aniline (Goody et al. 2002). Specifically, the following major metabolites are produced: (1)N-(3,4-dichlorophenyl)-N-methylurea (DCPMU); (2) 3,4-dichlorophenylurea (DCPU); and (3) dichloroaniline (DCA) (Goody et al. 2002). Although a half-life was not estimated in this study, results indicate that after 50 days over 80% of the diuron that was initially added was still present in the soil, and that none of the chemical’s degradation products had been lost from the 54-cm-deep soil column. The metabolites of diuron are generally less mobile than the parent compound (Howard 1991).

Under anaerobic soil conditions, a dechlorinated metabolite,N-(3-chlorophenyl)-N-methylurea (mCPMU) is formed (Attaway et al. 1982).

As indicated by the Madhun and Freed (1987) study, experimental conditions affect the degradation potential of diuron in soil; therefore, greater weight was placed on studies that described these factors. In addition, greater weight was given to data for which the chemical products of degradation (e.g., CO 2or specific metabolites) were reported, and which apply to aerobic conditions where organisms are more likely to be exposed. Therefore, the half-life of diuron in soil was taken to be longer than 182 days (6 months), based particularly on the studies of Madhun and Freed (1987) and Goody et al. (2002), indicating that it is expected to persist under aerobic conditions in this environmental compartment.

It should also be noted that properties of the soil influence the degree of sorption of diuron, which in turn determines the potential for leaching and contamination of ground water. The proportion of organic matter in soil directly influences the amount of adsorbed diuron; namely, sorption of diuron increases as soil organic matter increases (Alva and Singh 1990).

Table 4a. Empirical data for degradationof diuron

Medium Fate process Degradation value Degradation endpoint / units Reference
Water Biodegradation 0 Biodegradation / %
(ready bio-degradation test)
CHRIP c2008
Seawater Biodegradation <1 Biodegradation/ % Thomas et al. 2002
Soil Biodegradation
(aerobic)
200 – 4000
(ultimate)
Half-life (days) Madhun and Freed 1987
90 Wauchope et al. 1992
37 – 80.7
(primary)
Rouchaud et al. 2000
Marine sediment Biodegradation
(anaerobic)
~45 Biodegradation / % Thomas et al. 2003

In addition to the available experimental data on the degradation of diuron, a QSAR-based weight-of-evidence approach (Environment Canada 2007b) was also applied using the degradation models shown in Table 4b to add to the weight of evidence. Given the ecological importance of the water compartment, the fact that most of the available models apply to water and the fact that diuron is expected to be released into this compartment, biodegradation in water was primarily examined.

The results of available QSAR models for degradation are summarized in Table 4b. Ready biodegradability criteria are defined by results of BIOWIN Sub-models 3 and 5 based on Bayesian analysis to ready biodegradation data (EPIsuite 2008). A substance is predicted to biodegrade readily if the result of BIOWIN Sub-model 3 is “weeks” or faster (and the BIOWIN Sub-model 5 probability result is >0.5. For diuron, BIOWIN Sub-model 3 predicts biodegradation in “months” whereas BIOWIN Sub-model 5 probability is <0.5.

Table 4b. Modelled data for degradation of diuron

Fate process Model
and model basis
Model result and prediction Extrapolated half-life (days)
Air
Atmospheric oxidation AOPWIN 2008a t1/2 = 0.49 days <2
Ozone reaction AOPWIN 2008a n/ab n/a
Water
Hydrolysis HYDROWIN 2008a t1/2 = >1 year n/a
Primary biodegradation
Biodegradation (aerobic) BIOWIN 2008a
Sub-model 4: expert survey
(qualitative results)
3.18c
may biodegrade fast
<182
Ultimate biodegradation
Biodegradation (aerobic) BIOWIN 2008a
Sub-model 3: expert survey
(qualitative results)
2.27c
biodegrades slowly: months
>182
Biodegradation (aerobic) BIOWIN 2008a
Sub-model 5:
MITI linear probability
0.06d
not readily biodegradable
>182
Biodegradation (aerobic) BIOWIN 2008a
Sub-model 6:
MITI non-linear probability
0.01d
not readily biodegradable
>182
Biodegradation (aerobic) TOPKAT 2004
Probability
0.00d
biodegrades slowly
>182
Biodegradation (aerobic) CPOPs 2008
% BOD
(biological oxygen demand)
% BOD = 8.30
biodegrades very slowly
>182

a EPIsuite (2008)
b n/a., not applicable. Model does not provide an estimate for this type of structure.
c Output is a numerical score from 0 to 5.
d Output is a probability score.

In air, a predicted atmospheric oxidation half-life value of 0.5 days (see Table 4b) demonstrates that diuron is likely to be quickly oxidized. The substance is not expected to react appreciably with other photo-oxidative species in the atmosphere, such as O3, nor is it likely to degrade via direct photolysis. Therefore, it is expected that reactions with hydroxyl radicals will be the most important fate process in the atmosphere for diuron. With a half-life of 0.5 days via reactions with hydroxyl radicals, diuron is considered to be not persistent in air.

In water, a predicted hydrolysis half-life that is greater than 1 year (see Table 4b) suggests that this chemical is likely to be slowly hydrolyzed. This result is consistent with the fact that diuron does not contain functional groups expected to undergo rapid hydrolysis unless demethylation of the diuron urea group has occurred. As indicated previously, experimental evidence shows that the rate of hydrolysis in circum- neutral water at 25°C is very slow (Salvestrini et al. 2002; Giacomazzi and Cachet 2004). However, under well-oxygenated conditions its degradation products, namely the second sequential metabolite DCPU, may undergo hydrolysis to produce chlorinated anilines (Goody et al. 2002).

There is also strong evidence based on the results of biodegradation models that diuron does not biodegrade readily in water (Table 4b). All of the ultimate biodegradation models suggest that diuron biodegradation is slow and that the half-life in water would be >182 days. Moreover, predictions of CPOPs (2008) and TOPKAT (2004) for diuron are in the domains of both models. Thus, those models are considered to be reliable and suggest a very slow rate of biodegradation. Model results in Table 4b are generally consistent with the previously described empirical data indicating the slow ultimate biodegradation of diuron.

The result of the BIOWIN Sub-model 4 (primary survey model) exceeds the threshold value of =3 by a narrow margin of 0.1 to 0.2 indicating a potential for relatively fast biodegradation. However, the empirical degradation data previously described indicate that this model may overestimate the rate of primary biodegradation of diuron.

Using an extrapolation ratio of 1:1:4 for a water:soil:sediment biodegradation half-life (Boethling et al. 1995), and an ultimate biodegradation half-life of >182 days in water, the half-life in oxic soil is also >182 days and the half-life in oxic sediments is >365 days. However, as presented in Table 4a, there is empirical evidence for rapid degradation of diuron in anoxic sediment. Application of the water:soil:sediment extrapolation ratio indicates that diuron is expected to be persistent in oxic soil and sediment. This extrapolation result is consistent with the available empirical evidence for degradation of diuron in oxic soil (Table 4a).

Based on the empirical and modelled data (see Tables 4a and 4b), diuron meets the persistence criteria in water as well as in oxic soil (half-lives in soil and water are =182days) and oxic sediment (half-life in sediment ³365 days), but does not meet the criteria for air (half-life in air is ³2 days) as set out in thePersistence and Bioaccumulation Regulations (Canada 2000).

Potential for Bioaccumulation

Limited experimental data for bioaccumulation of diuron exists. Furthermore, experimental data for bioaccumulation of this substance in aquatic organisms is contradictory. Studies (Call et al. 1987, Tucker and Kingsbury 2003) where fish have been exposed to diuron through treatments of water for relatively short periods of time and allowed to recover during periods of no exposure point to low potential for bioconcentration of this substance. Bioconcentration potential of diuron has been tested in rice fish (Oryzias latipes) by the Japanese Institute of Technology and Evaluation (NITE), and it was concluded that bioconcentration of diuron is low with the reported bioconcentration factor (BCF) values of <2.9–14 (CHRIP c2008). In addition, a laboratory study performed on cupped oysters (Crassostrea gigas) from the Bay of Veys, France, indicated some accumulation of diuron in the tissues as well as effects on the reproductive cycle and digestive gland tubules; however, exposure in the field led to no detectable accumulation of diuron (Buisson et al. 2008).

The experimental log kow values of durion are moderate at 2.68 and 2.85 (Table 2). Modelled bioconcentration data for fish, using the experimental Kow of 2.85, were generated and suggest that the substance does not bioaccumulate; all calculated BCF and BAF) values were much less than 5000 (BBM with Mitigating Factors 2008, BCFBAF 2008). Modelled bioconcentration data are not discussed further in this assessment.

Finally, results of a recent study performed in the Biosphere Reserve of Camargue, France, (Roche et al. 2009) suggest that diuron bioconcentrates and biomagnifies through several trophic levels, despite its moderate log Kow. However, as explained in more detail below, there are uncertainties about how the data in this study were interpreted and consequently the results of this study were given relatively little weight in this assessment.

Call et al. (1987) investigated toxicity, uptake and elimination of diuron in freshwater fish and concluded that diuron does not accumulate in fish tissue to a large extent. Briefly, fathead minnows (Pimephales promelas) were exposed to14C-labelled diuron at concentrations of approximately 3 and 30 µg/L in water over 24 days. The 14C-labelled diuron was rapidly taken up by the fish, with equilibrium between water and fish established after 24 hours. Diuron was rapidly eliminated within 24 hours after transfer of fish to clean water (84% and 76% for lower and higher exposures, respectively), with nearly 99% of the substance eliminated after 21 days. A BCF of 2 was determined for diuron.

The metabolism of diuron was also investigated in rainbow trout (Oncorhynchus mykiss). Two groups of rainbow trout were injected with 1 µCi of 14C-labelled diuron and sacrificed after 24 hours, with one group having been pre-exposed to unlabelled diuron at a sub-lethal concentration for 4 to 5 days (Call et al. 1987). Diuron was rapidly eliminated by trout with more than 90% of radioactivity present in the water after 24 hours. Parent diuron comprised 35% to 40% and at least four other metabolites were present, one of which was identified as 3,4-DCA and two others were demethylated compounds.

Tucker and Kingsbury (2003) investigated both the bioaccumulation potential of diuron and the potential for diuron residues to carry over from one season of exposure to the next in catfish (Ictalurus punctatus). Commercial catfish ponds were treated with diuron to prevent cyanobacteria outbreaks that cause undesirable off-flavours and render fish unacceptable for sale. The limited tolerance level of diuron of 2.0 mg/kg in catfish fillets was established by the US EPA (Federal Register 1999). Three ponds were stocked with three populations of catfish: large fish averaging 650 g, large fingerlings at 75 g per fish and small fingerlings at approximately 20 g per fish, to represent typical catfish culture ponds. Ponds were treated with diuron at 0.01 mg/L for nine consecutive weeks in two seasons: first in the fall and second in the following spring–summer months. Resulting waterborne concentrations of diuron were not measured in this study (Tucker and Kingsbury 2003); thus, BAFs could not be calculated.

Fish samples (six fish per sample) were collected as follows: (1) 4 days before first diuron application, (2) 1 day after last diuron application in the first treatment series, (3) 4 and 6 months after completion of the first treatment series, (4) 1 day before second treatment series and (5) 1 day after final diuron application. Diuron residue levels remained below 1 mg/kg after nine consecutive weekly treatments with diuron. However, after the first treatment, tissue levels of diuron in catfish varied significantly among ponds. This was presumably due to potential differences in diuron residue levels in the water and pond environmental conditions such as biomass and pond microflora. Diuron tissue residue levels did not vary among ponds after the second treatment, presumably due to higher water temperatures that promote faster pesticide biodegradation. Also, variation of diuron residue levels among all fish was reported. Specifically, following the first diuron treatment in the fall, the mean tissue concentration of diuron varied by nearly an order of magnitude among all fish (0.078 to 0.724 mg/kg) and as much as three-fold in fish within a pond (0.078 to 0.247 mg/kg). This variation was lower following the second treatment in the spring–summer, when the diuron residue levels varied less than four-fold in all fish (0.05 to 0.191 mg/L) and again three-fold within a pond (0.05 to 0.152 mg/kg). The overall observed variation may reflect the relative rates of uptake and elimination of diuron. Diuron levels in fillets fell below the limit of detection (0.05 mg/kg) within 4 to 6 months after the final application and residues did not carry over from one year to the next. Tucker and Kingsbury (2003) concluded that the exposure of catfish to diuron in consecutive years would not cause residues in fillets to exceed 2.0 mg/kg, the tolerance limit set by the US EPA.

Buisson et al. (2008) investigated whether exposure to pesticides acted as an additional stressor in the significant summer mortality events of cupped oysters in the Bay of Veys, Normandy, France. These summer mortality events, likely caused by a combination of several factors including elevated temperature, low dissolved oxygen, xenobiotic stress and physiological stress related to reproduction, threaten commercial oyster production in the region. Diuron was assessed in both a 4-month field study (February to May) in two bay locations, and in a 7-day laboratory exposure. Diuron concentrations were measured in the water as well as in the oyster tissues. Environmental water concentrations of diuron varied from 0.015 µg/L in early spring to 0.254 µg/L in the late spring. These levels of contamination did not lead to any detectable diuron bioaccumulation in the oyster tissues. However, it should be noted that concentrations of organochlorine pollutants in shellfish vary among sampling locations, collection seasons and species (Buisson et al. 2008).

In the laboratory study, cupped oysters were exposed to three concentrations of diuron: 0.1, 1 and 10 µg/L for 7 days. Diuron concentrations accumulated to 7.4 and 53 ng/g wet weight in oysters at exposure concentrations of 1 and 10 µg/L, respectively. Bioconcentration factors were calculated: the BCF was 7–7.5 at 1 µg/L and 5.3 at an exposure of 10 µg/L (Buisson et al. 2008). In addition, diuron bioaccumulation of 5.3 ng/g wet weight in oysters was also observed at an exposure level of 0.5 µg/L in a mixture containing four pesticides: diuron, isoprouron, deethylatrazine (DEA) and bentazon (Buisson et al. 2008). The laboratory study demonstrates that diuron has the potential to bioaccumulate at a very low level in the cupped oyster over a short exposure period in the test medium. However, the calculated BCF values of 5.3 to 7.5 are well below criteria specified in the Persistence and Bioaccumulation Regulations (Canada 2000), which state that a substance is bioaccumulative if its BAF or BCF is ≥ 5000.

Results presented by Roche et al. (2009) are in contradiction to the above mentioned studies (CHRIP c2008, Call et al. 1987, Tucker and Kingsbury 2003, Buisson et al. 2008). According to the authors, ecotoxicological studies focused on the biota of the Vaccarès Lagoon, the Biosphere Reserve in Rhone Delta, France, during the decade 1996–2006, show contamination at all levels of the food web due to direct bioconcentration of diuron from the water and food transfer. In contrast, studies discussed above address bioaccumulation of diuron in fish solely from exposure to water. The Vaccarès Lagoon is exposed to the inflow of pollutants from the Rhone River and pesticides from spraying of the bordering agrosystems. The authors reported contamination of trophic web components sampled in 2002 and 2005, during two seasons, spring and autumn.

Briefly, trophic levels of invertebrates and fish in the Vaccarès Lagoon were assessed based on δ15N enrichment, and concentrations of diuron in trophic guilds were lipid-normalized. The following trophic levels were assigned: Producers (Zostera species and sedimentary organic matter), primary consumers or Consumers 1 (phytoplanktivorous [zooplankton] and depositivorous [Sphaeroma species, co*ckles, gammarids, mysids]) and secondary consumers, specifically, Consumers II-1 (zooplanktivorous [shrimps, juvenile eel]), Consumers II-2 (benthivorous fish [goby, pipefish, sand smelt] and piscivorous fish [mullet, stickleback, juvenile eel, sole, bream, pikeperch]) and Consumers II-3 (top-consumer fish [yellow eel]). Diuron was found at every trophic level and was the most abundant contaminant in the Vaccares Lagoon in 2005. However, environmental concentrations of diuron in the waters of the Vaccarès Lagoon were not reported by the authors. Eels, the top predators, were found to be highly contaminated (1000 ng/g dry weight of tissue in the spring and 3000 ng/g dry weight in the fall). Roche et al. (2009) suggested that this high level of contamination contradicted the Tucker and Kingsbury (2003) study, in which a low diuron transfer to fish was shown after several consecutive treatments with 0.01 mg/L of diuron in the water. However, since environmental concentrations of diuron in the waters of the Vaccarès Lagoon were not reported by the authors, the levels of diuron causing such contamination of eels can only be inferred to have been high.

Diuron biomagnification factors (BMFs) greater than 1 were calculated for all trophic levels (Roche et al. 2009); for Producers versus Consumers I the BMF was 2, for Consumers I versus Consumers II-1 the BMF was 1.61, for Consumers II-1versus Consumers II-2 the BMF was 1.3 and finally for Consumers II-2 versus Consumers II-3 the BMF was 2. These results suggested that the rate of uptake of diuron in aquatic organisms of the Vaccarès Lagoon exceeded its rate of elimination and that this substance has the ability to bioaccumulate in this food web.

The evidence for the bioaccumulation potential of diuron presented was carefully considered. The empirical database of acceptable BCF and BAF values examined by Arnot and Gobas (2006) reveals that the majority of chemicals with log kowvalues less than ~3.5 are not expected to have BCF or BAF values = 5000. Indeed, all of the available BCF evidence indicates that diuron has a low bioaccumulation potential (i.e., BCF = 5000). Furthermore, this is consistent with the physical and chemical properties of diuron (i.e., chemical structure, moderate water solubility, low log Kow) as well as the high potential for tissue biotransformation in vivo (Call et al. 1987). Field evidence from the Tucker and Kingsbury (2003) study also suggests that accumulation of diuron in fish from consistent long-term exposures is mitigated by in vivobiotransformation. Diuron levels in fish fillets in this study declined to non-detectable amounts once exposures were stopped, which suggested a relatively rapid elimination in these fish. Laboratory BCF values in filter feeders (Buisson et al. 2008), which might be expected to reflect a worst-case bioconcentration potential due to high uptake rates and low metabolic rates, were also significantly less than 5000.

The biomagnification field study from the Vaccarès Lagoon (Roche et al. 2009) suggests that diuron has the ability to bioconcentrate and biomagnify in the studied food web, with calculated BMF values ranging from 1.3 to 2. However, because of uncertainties related to the approach to data analysis and interpretation in this study, the reliability of the reported BMFs may be questioned.

Details on how data were combined for the trophic and residue level comparisons were not provided by Roche et al. (2009). There were substantial seasonal as well as year-to-year variations in chemical residue levels; these variations were not always in the same direction for season, year and particular species. It is not clear whether the variations were adequately accounted for; however, it appears that data were combined together without regard for year and season. As such, there is potential for erroneous trophic and residue level comparisons, which could possibly lead to unreliable calculated BMF values. Moreover, BMF values provided for some highly studied chemicals, such as polychlorinated biphenyls (PCBs), are not always in agreement with the well-known environmental behaviour of these chemicals.

Due to the large variability of ambient concentrations and biological processes in aquatic ecosystems, as well as potential measurement errors associated with lipid and contaminant analysis, the BMFs reported for diuron by Roche et al. (2009) are likely not large enough to be significant. Therefore, the BMFs reported by Roche et al. (2009) are not considered to indicate a high bioaccumulation potential for this substance.

Considering all of the above evidence on the bioaccumulation potential of diuron, including consideration that the most recent field study (Roche et al. 2009) suggests potential for biomagnification in contradiction to previous studies as well as the chemical property information that indicate a low potential for bioaccumulation, it is concluded that there is a stronger body of bioaccumulation evidence suggesting that diuron does not meet the bioaccumulation criteria of BAF or BCF = 5000 as set out in thePersistence and Bioaccumulation Regulations (Canada 2000).

In addition, empirical data on bioaccumulation of the biodegradation product of diuron, 3,4-DCA, indicate a relatively low bioaccumulation potential in fish (European Commission, 2006). Ensenbach et al. (1996), for example, reported BCF values of 20 and 38 for rainbow trout and zebrafish, respectively.

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Potential to Cause Ecological Harm

Ecological Effects Assessment

A – In the Aquatic Compartment

There is experimental evidence that shows that diuron causes harm to aquatic organisms following short-term (acute) as well as long-term (chronic) exposure at relatively low concentrations. Although modelled predictions for aquatic toxicity were performed for this substance, given the numerous experimental data available (see Table 5), they are not included in this report. Some of the key studies in which experimental data were reported were critically reviewed for validity. These reviews (Robust Study Summary) are provided in Appendix 1.

The Ecological Fate and Effects Division of the US EPA Office of Pesticide Programs (OPP) has reviewed a large number of diuron toxicity studies performed by the US EPA, US Department of Agriculture and US Fish and Wildlife Service laboratories on several aquatic and terrestrial species. Summaries of these studies are available from the US EPA Office of Pesticide Programs (OPP) Ecotoxicity Database (OPP Pesticide Ecotoxicity Database 2008). These studies have been reviewed by the US EPA biologists and judged acceptable for use in the ecological risk assessment process (OPP Pesticide Ecotoxicity Database 2008). As such, they are considered adequate for consideration in this assessment, even though experimental conditions or details are not available and some studies are unpublished. Results reported in the OPP ecotoxicity database and from other published sources for aquatic organisms are summarized and presented as ranges in Table 5. Fish study reports for which the age of experimental organisms was not reported were disregarded. Results for terrestrial organisms are presented in Table 6.

Table 5. Empirical data for aquatic toxicity

Test organism Type of test Endpoint Value (mg/L) Reference
Algae
Algae
(Scenedesmus subspicatus)
Chronic
(72 hours)
EC50a growth inhibition 0.036 Schäfer et al. 1994
Green algae
(Chlorella pyrenoidosa)
Chronic
(96 hours)
0.0013 Ma et al. 2001
Freshwater green algae species Chronic
(72, 96, 240 hours)
0.0024–0.037
(2.4–37 ppb)
(n = 7)b
OPP Pesticide Ecotoxicity Database 2008
Marine algae species Chronic
(72 hours)
0.01–0.05
(10–50 ppb)
(n = 8)b
Crustaceans
Brine shrimp
(Artemia franciscana)
Acute (24 hours) LC50c 12.5 Koutsaftis and Aoyama 2008
Freshwater shrimp
(Paratya australiensis)
Acute (96 hours) 8.8
(8800 µg/L)
Kumar et al., 2009
Brown shrimp (Penaeus aztecus) Acute (48 hours) 1.0
(1.0 ppm)
OPP Pesticide Ecotoxicity Database 2008
Mysid
(Americamysis bahia)
Acute (96 hours) 1.1
(1.1 ppm)
Scud
(Gammarus fasciatus)
Acute (96 hours) 0.16
(0.16 ppm)
Water flea
(Daphnia magna, D. pulex, Simocephalusspecies)
Acute (48 hours) EC50 growth inhibition 1.4–8.4
(1.4–8.4 ppm)
(n = 3)b
Water flea
(Daphnia magna)
Chronic (21 or 28 days) LOECd 0.113–0.2
(0.113–0.2 ppm)
(n = 2)b
Fish
Fathead minnow
(Pimephales promelas)
Acute
(24, 48, 96, 192 hours)
LC50 7.7–23.3
(n = 4)b
Call et al. 1987
Early life-stage chronic toxicity (60 day) MATCe 0.033–0.078
(33.4-–78.0 µg/L)
Bluegill sunfish
(Lepomis macrochirus)
Acute (96 hours) LC50 2.8–84
(2.8–84 ppm)
(n = 2)b
OPP Pesticide Ecotoxicity Database 2008
Cutthroat trout
(Oncorhynchus clarkii)
1.4
(1.4 ppm)
Lake trout
(Salvelinus namaycush)
1.2
(1.2 ppm)
Rainbow trout
(Oncorhynchus mykiss)
Chronic (28 days) 0.23 Okamura et al. 2002
Acute (96 hours) 1.95
(1.95 ppm)
OPP Pesticide Ecotoxicity Database 2008
Striped mullet
(Mugil cephalus)
Acute (48 hours) 6.3
(6.3 ppm)
Fathead minnow
(Pimephales promelas)
Early life-stage chronic toxicity (60 day) LOEC 0.062
(61.8 ppb)
Sheepshead minnow
(Cyprinodon variegates)
Early life-stage chronic toxicity (38 day) 0.44
(0.44 ppm)
Amphibians
Pacific treefrog (Pseudacris regilla) Chronic (14 day) LC50 15.2 Schuytema and Nebeker 1997
Bullfrog (Rana catesbeiana) Chronic (21 day) 12.7
Red-legged frog (Rana aurora) Chronic (14 day) 22.2
African clawed frog (Xenopus laevis) Chronic (14 day) 11.3

aEC50- The concentration of a substance estimated to cause some effect on 50% of the test organisms.
b The value in parentheses represents the number of values included in the range.
c LC50 – The concentration of a substance estimated to be lethal to 50% of the test organisms.
d LOEC – The Low Observed Effect Concentration is the lowest concentration in a toxicity test that caused a statistically significant effect in comparison to the controls.
e MATC - The maximum allowable toxicant concentration, generally presented as the range between the NOEC(L) and LOEC(L) or as the geometric mean of the two measures. NOEC – The No Observed Effect Concentration is the highest concentration in a toxicity test not causing a statistically significant effect in comparison to the controls.

Diuron may pose an increased hazard to phytoplankton and aquatic plants because it acts as a specific inhibitor of photosynthesis (Taiz and Zeiger 2006). Briefly, diuron prevents electron transfer at the quinone acceptors of photosystem II by competing for the binding site of plastoquinone. Thus, binding of diuron effectively blocks electron flow and inhibits photosynthesis. In addition, electron flow in photosynthetic bacteria that have quinone-type electron acceptor complexes may also be affected by diuron. Environmentally realistic diuron concentrations produced responses in both target organisms (algae) and non-target organisms (bacteria) (Ricart et al. 2009). Growth inhibition following exposure to diuron was studied in green algae in static test systems; the resulting EC50 after 72 hours of exposure was 0.036 mg/L for the species Scenedesmus subspicatus(Schäfer et al. 1994). The EC50 for speciesChlorella pyrenoidosa following 96 hours of exposure was 0.0013 mg/L (Ma et al. 2001). Also, growth inhibition in several species of fresh and marine water algae was observed in chronic exposure studies with EC50 values ranging from 0.0024 to 0.05 mg/L (OPP Pesticide Ecotoxicity Database 2008).

Acute toxicity of diuron to three tropical seagrass species,Halophilia ovalis, Cymodocea serrulata andZostera capricorni, was investigated at environmentally relevant concentrations that ranged from 0 to 100 µg/L (Haynes et al. 2000). The effect of diuron on photosynthesis was assessed by measuring change in chlorophyll fluorescence following a 5-day exposure period and a 5-day recovery period. Over the 5-day exposure period, concentrations of 10 and 100 µg/L of diuron reduced effective quantum yield in the three seagrass species by 50% to 75%, whereas lower diuron concentrations of 0.1 and 1 µg/L reduced effective quantum yield by 10% and 30% in H. ovalis and Z. capricorni, respectively, and no changes were observed in C. serrulata. Recovery of photosynthetic ability following return to clean seawater was initially rapid in the three seagrass species; however, all species exhibited fluctuations in effective quantum yield over the 5-day recovery period.

Diuron has been shown to be highly toxic to crustaceans. Acute LC50 values for brown shrimp, scud and mysid were 1, 1.1 and 0.16 mg/L, respectively (OPP Pesticide Ecotoxicity Database 2008). Also, moderate acute diuron toxicity was observed in two shrimp species, Artemia franciscana and Paratya australiensis, with LC50 values at 12.5 mg/L (Koutsaftis and Aoyama 2008) and 8.8 mg/L (Kumar et al. 2009), respectively. The water flea (Daphnia magna) was exposed to diuron in both acute and chronic toxicity tests; LC50values from acute (48 hour) diuron exposure ranged from 1.4–8.4 mg/L while chronic LOECs for D. magna ranged from 0.113 to 0.2 mg/L (OPP Pesticide Ecotoxicity Database 2008).

Toxicity as well as effects of temperature and salinity from diuron exposure were studied in the brine shrimp by Koutsaftis and Aoyama (2008). The LC50 value following an acute 24-hour exposure to diuron was determined to be 12.5 mg/L. Furthermore, it was established that the mortality of brine shrimp following exposure to diuron decreased with decreasing temperature and salinity.

Laboratory exposure to low levels of diuron produced sublethal effects in the cupped oyster in laboratory experiments over the period of 7 days (Buisson et al. 2008). At an exposure concentration of 1 µg/L of diuron, both accelerated spawning events in oysters and atrophy of the digestive tubule epithelium were noted. A stimulation of spawning in aquatic organisms often occurs when they are transferred to experimental conditions; however, exposure to diuron appeared to increase this phenomenon. Incidences of atrophy of digestive tubule epithelium were significantly more frequent in oysters exposed to 1 µg/L of diuron than in the control group.

Technical grade diuron was tested on freshwater fish. In fathead minnows, acute tests indicated a high to moderate toxicity to diuron at 7.7 to 23.3 mg/L for exposure times ranging from 196 to 24 hours, respectively (Call et al. 1987). NOEC and LOEC values were also determined in early life-stage toxicity tests where egg hatching success, survival rate and fish growth (wet weight and length) endpoints were measured during 60 days of diuron exposure. NOEC and LOEC values were determined to be 0.03 mg/L and 0.078 mg/L, respectively (Call et al. 1987). The MATC is therefore expected to fall in this range. LOECs of 0.44 and 0.062 mg/L were observed for sheepshead and fathead minnows, respectively, when fish were exposed to diuron in early life-stage chronic toxicity tests (OPP Pesticide Ecotoxicity Database 2008). Okamura et al. (2002) determined both acute and chronic LC50 values in juvenile rainbow trout following exposure to diuron. LC50 values for 7-day, 14-day, 21-day and 28-day exposures were 74, 15, 5.9 and 0.23 mg/L, respectively (28-day exposure value reported in Table 5). Finally, high to moderate toxicity was observed in several freshwater fish species including bluegill sunfish, striped mullet, as well as lake trout, rainbow trout and cutthroat trout following acute 96 hour exposure to diuron. The resulting LC50 values ranged from 1.2 mg/L for lake trout to 84 mg/L for bluegill (OPP Pesticide Ecotoxicity Database 2008).

Toxicity of diuron was investigated in the early life stages of the pink snapper (Pagrus auratus), an important commercial and recreational fish in Australia (Gagnon and Rawson 2009). Fertilized fish eggs were exposed to diuron for 36 hours at concentrations of 50, 5, 0.5 and 0.05 µg/L, which covered the range of estimated and measured concentrations in the marine environment. At the highest diuron exposure concentration of 50 µg/L, there was a significant increase in the rate of spinal deformities in hatched pink snapper and a decrease in the proportion of eggs that hatched and had normal development up to 36 hours after spawning. Although EC50 values for diuron exposure were not established from this study, Gagnon and Rawson (2009) suggest a LOEC of 50 µg/L and a NOEC of 5 µg/L.

The toxicity of diuron was studied in different species of frogs: African clawed frog (Xenopus laevis), bullfrog (Rana catesbeiana), red-legged frog (Rana aurora) and Pacific treefrog (Pseudacris regilla) (Schuytema and Nebeker 1997). Tadpoles were exposed to diuron in water during static-renewal tests for 14 days, and up to 21 days, to estimate chronic effects. The resulting mean LC50 values were moderately toxic at 15.2 mg/L for the Pacific treefrog, 12.7 mg/L (21 days) for bullfrog, 22.2 mg/L for red-legged frog and 11.3 mg/L for the African clawed frog. In addition, both NOEC and LOEC values were determined in tadpole tests based on growth rate; the lowest NOEC values calculated were 14.5 mg/L for the Pacific treefrog and 7.6 mg/L for bullfrog, red-legged frog and African clawed frog; LOEC values were 21.1 mg/L for the Pacific treefrog and 14.5 mg/L for bullfrog, red-legged frog and African clawed frog.

There is also preliminary in vitro evidence that suggests that diuron triggers effects on certain endocrine endpoints and pathways. Briefly, effects of diuron were investigated on endocrine endpoints in two in vitro assays at environmentally relevant concentrations: estrogenic/androgenic and anti-estrogenic/anti-androgenic receptor-mediated activity was tested in a recombinant yeast assay, and disruption of the steroidogenic pathway was tested using cultured Xenopusoocytes (Orton et al. 2009). Diuron did not show estrogenic or androgenic activity in the yeast screen assays; however, both anti-estrogenic and anti-androgenic activities were observed following exposure to diuron concentrations in the range of 15.6–0.008 µM. Furthermore, diuron was more potent as an anti-estrogen than as an anti-androgen. Effects in the ovulatory response and hormone production in Xenopus oocytes were noted following exposure to diuron at 62.5 µg/L; namely, there was a depression in production of testosterone accompanied by a decrease in ovulation.

Finally, with regard to the toxicity of diuron degradation products, Tixier et al. (2001) showed by means of Microtox assays that the main metabolites of diuron, obtained from the process of demethylation under aerobic conditions, have higher toxicity to the marine bacterium Vibrio fischeri than the parent substance. The Microtox test consists of determining the concentration of a toxic compound that inhibits 50% of the natural bioluminescence (EC50); the EC50 value for diuron was 68 mg/L. Of six metabolites tested, two hydroxylated compounds showed toxicity similar to that of diuron, with EC50 values of about 72 mg/L; three compounds had EC50 values in the range of 7.3–18 mg/L, and 3,4-DCA had the greatest toxicity with an EC50 of 0.48 mg/L.

As noted previously, the metabolite 3,4-DCA was identified as a substance of toxicological concern to human health (US EPA 2003). This metabolite may act as a non-specific membrane irritant or metabolic inhibitor (Scheil et al2009). It does not appear to have herbicidal effects (US EPA 2001).

Some empirical ecotoxicity data for 3,4-DCA are available. A 72-hour EC50 for 3,4-DCA was established in a static test system for the green alga, Scenedesmus subspicatus, and this substances appeared to be moderately toxic at a concentration of 15 mg/L (Schäfer et al. 1994). In juvenile freshwater fish, 3,4-DCA appears to be moderately toxic, with LC50 values ranging from 2.7 mg/L for the rainbow trout to 9 mg/L for the guppy (Ensenbach et al. 1996). However, high toxicity was observed in zebrafish larvae after an 11-day exposure to 3,4-DCA, with the LC50 value reported at 0.388 mg/L (Scheil et al. 2009). The aquatic toxicity of 3,4-DCA was reviewed by the European Commission in the European Union Risk Assessment Report (European Commission 2006). It was determined that daphnids were the most sensitive species in short-term tests with reported 48-hour LC50 for Daphnia magna of 0.23 mg/l and a 96-hour LC50 of 0.16 mg/L. The most sensitive species in long-term tests were zebrafish and guppy with 42-day NOECs of 2 µg/L.

Effects of immigration on the recovery of aquatic macroinvertebrate communities following exposure to 3,4-DCA were investigated in a recent study using outdoor pond microcosms (Maund et al. 2009). Recovery rates of aquatic macroinvertebrates were compared between experimental groups, such that groups where only autochthonous processes occurred (such as growth, reproduction or hatching of resting stages) were compared to those groups in which organisms were added to simulate immigration. 3,4-DCA exposure concentration of 10 mg/L corresponding to the median acute toxicity value for several aquatic invertebrates, was chosen with the intent to cause substantial effects on the macroinvertebrate communities, rather than to represent an environmentally realistic concentration. The difference in recovery rates between the two treatments showed that simulated immigration had a substantial influence on the recovery potential of macroinvertebrates at both population and community levels. In treated systems where no organisms were added, recovery was slow or did not occur. The results of the Maund et al. (2009) study suggested that the process of immigration has a potentially strong influence on community recovery after xenobiotic chemical effects. Moreover, the speed and extent of recovery can be influenced by toxicological and ecological factors, and by chemical dynamics, i.e., species sensitivities, reproduction and dispersal rates as well as dissipation rates of chemicals. The slow dissipation of 3,4-DCA clearly had an influence on the recovery profile of the test systems.

The weight of evidence based on available experimental data indicates that diuron, as well as its metabolite 3,4-DCA, has the potential to cause acute harm to sensitive aquatic organisms at low concentrations (i.e., acute LC50 values are =1.0 mg/L).

B – In Other Environmental Compartments

Empirical toxicity data of diuron for terrestrial organisms were identified in the US EPA OPP Pesticide Ecotoxicity Database (2008). Experimental values for birds, insects and mammals are presented in Table 6.

Toxicity studies on several bird species including bobwhite (Colinus virginianus), Japanese quail (Coturnix japonica), mallard duck (Anas platyrhynchos), and ringed-neck pheasant (Phasianus colchicus) were summarized. Diuron was administered either in the diet or water (in ppm), or as oral gavage or capsule (mg/kg body weight) for an exposure period of 8 days up to 14 days. The reported LC50 values ranged from 1730 ppm for 9-day-old bobwhite quails to 5000 ppm for 12-day-old Japanese quails and 15-day-old ring-necked pheasants. The LC50 values for 3-month-old mallard ducks were 2000 mg/kg body weight following 14 days of diuron exposure, and 5000 ppm for 10-day-old birds following 8 days of exposure. These studies were performed in the 1970s and reviewed by the US EPA in 1982.

A summary of a toxicity study performed in 1980 for the terrestrial insect, the honey bee (Apis mellifera), was described in the OPP Pesticide Ecotoxicity Database (2008). Adult honey bees were exposed to diuron in an acute contact study from which the LD50 was established as 145.3 µg diuron per bee. This result indicates that diuron is relatively non-toxic to bees on an acute contact basis (US EPA 2001).

An oral acute toxicity study with laboratory rats was also described in the US EPA’s Environmental Risk Assessment for the Re-registration of Diuron (US EPA 2001). Based on this study, the US EPA concluded that diuron was in Toxicity Category III, i.e., slightly toxic or irritating to small mammals on acute oral basis (US EPA 2001).

Table 6. Empirical data for terrestrial toxicity

Test organism Type of test Endpoint Value Reference
Birds
Bobwhite quail (Colinus virginianus), mallard duck (Anas platyrhynchos),
Japanese quail (Coturnix japonica), ring-necked
pheasant (Phasianus colchicus)
Acute
(8–14 days; food, water or oral gavage/
capsule)
LC50 1730–5000 ppm
(n = 4)a
OPP Pesticide Ecotoxicity Database 2008
Insects
Honey bee
(Apis mellifera)
Acute (48 hours; contact) LD50b 145.3 µg per bee OPP Pesticide Ecotoxicity Database 2008
Mammals
Laboratory rat Acute LD50 5000 mg/kg (males)
1000 mg/kg
(females)
US EPA 2001, 2003

a The value in parentheses represents the number of values that are included in the range.
b LD50 – The dose of a substance that is estimated to be lethal to 50% of the test organisms.

Studies of the toxicity of diuron to plants were not identified in the available published literature. However, several unpublished diuron plant toxicity studies for non-target species were conducted by registrants as a requirement for herbicide registration. These studies were analyzed by the US EPA and compiled in the Registration Eligibility Decision document (US EPA 2003). Briefly, Tier II terrestrial plant seedling emergence and vegetative vigour toxicity studies were conducted with four species of monocotyledonous plants (corn, onion, sorghum and wheat) and six species of dicotyledonous plants (soybean, pea, rape [canola], cucumber, sugar beet and tomato). The most sensitive species for seedling emergence were onion and tomato, with reported EC25 values at 0.099 and 0.08 lbs active ingredient/A, respectively. For plant vegetative vigour studies, wheat and tomato were most sensitive, with reported EC25 values at 0.021 and 0.002 lbs active ingredient/A, respectively.

Ecological Exposure Assessment

A – Industrial Release

When diuron is used industrially releases are mainly expected to be to water. Diuron is also expected to be found at waste management sites, due to the eventual disposal of manufactured items containing the substance. However, releases of diuron from such waste disposal sites are expected to be negligible.

Conservative industrial release scenarios were used to estimate the aquatic concentrations of diuron. Two aquatic concentrations of diuron were estimated based on the known uses of diuron as a hardening agent in epoxy resins and a curing agent in epoxy adhesive. The scenarios were made conservative by assuming that the total quantity of the substance is used by an industrial facility at a single specific site in Canada. Input parameters of each exposure scenario were adjusted to reflect some realistic scenario assumptions based on the available information regarding the substance use pattern codes and location of industrial sites. The losses to sewer were estimated at 0.015% of the total quantity based on the Emission Scenario Document on Plastic Additives (OECD 2004b), resulting from the cleaning of chemical containers and process equipment. The scenarios also assumed that the releases occurred 250 days per year, typical for small and medium-sized industrial facilities, and were sent to a local sewage treatment plants (STPs) with estimated removal rates for the substance of either 31.2% (primary and secondary removal rate estimated using SimpleTreat 1997) or 4.8% (primary removal rate estimated by STP 2001). Upon combining with the STP effluent, the receiving waters were estimated to have an actual flow of either 15 000 m3 per day or 140 400 m3 per day, derived from the specific sites chosen for the scenario. The dilution factor of the receiving waters was limited to a maximum of 10. Based on the above assumptions, an aquatic concentration of 9.8x 10-6mg/L was estimated for this substance as a result of its use as a hardening agent in epoxy resins, and 2.5x10-7 mg/L as a result of its use a curing agent in epoxy adhesive. (Environment Canada 2010a).

B – Consumer Release

Releases of diuron by consumers are expected to be limited. Current known non-pesticidal applications of the substance do not indicate dispersed or widespread consumer uses. Therefore, it is expected that end-use consumer releases of diuron into environment are not likely to be significant.

Characterization of Ecological Risk

The aim of the present characterization of ecological risk is to assess the impacts of diuron releases as a result of non-pesticidal industrial uses of the substance. Information from ecological risk assessments of the pesticidal uses of diuron developed by the US EPA and PMRA was used where applicable. The approach taken in this assessment was to examine the available scientific information and develop conclusions based on a weight-of-evidence approach and using precaution as required under CEPA 1999. Lines of evidence considered include results from conservative risk quotient calculations, as well as information on the persistence, bioaccumulation, toxicity, sources and fate of the substance.

As described previously, diuron is expected to have relatively long half-lives in aerobic, sediment, soil and water. There are significant uncertainties associated with a recent claim (Roche et al. 2009) that diuron has the potential to bioaccumulate in aquatic organisms and be efficiently transferred through successive trophic levels. With the exception of the study by Roche et al. (2009), all of the available data suggest that diuron has a low potential to bioaccumulate. Available empirical toxicity data suggest that diuron is highly to moderately hazardous to both aquatic and terrestrial organisms.

There is also preliminary in vitro evidence based on studies in cultured Xenopus oocytes suggesting that diuron may exert effects on certain endocrine endpoints and pathways (Orton et al. 2009).

In 2006, approximately 4% of the total amount of diuron imported to Canada was used for non-pesticidal purposes (Environment Canada 2009b). These included industrial applications as a hardening agent in epoxy resins, a curing agent in epoxy adhesive as well as contained in manufactured items. As a result of non-pesticidal industrial uses of diuron, as identified by section 71 survey (Environment Canada 2009b), diuron may be released into the environment with water being the primary receiving medium. Also, a quantity of diuron is expected to be found in waste management sites (landfills or incinerators), due to the eventual disposal of manufactured items containing it. . Releases of diuron from such waste disposal sites are uncertain, but expected to be negligible.

Scenarios were developed to estimate conservative concentrations of the substance in water resulting from industrial discharges. This yielded predicted environmental concentrations (PECs) of 9.8x 10-6 mg/L and 2.5x10-7 mg/L as result of applications of diuron as a hardening agent in epoxy resins and a curing agent in epoxy adhesive, respectively. Details regarding the inputs used to estimate this concentration and the output of the model are described in Environment Canada (2010a).

A conservative predicted no-effect concentration (PNEC) was derived from the lowest toxicity value identified for an organism relevant to the Canadian environment; this was a chronic EC50 of 0.0013 mg/L for green algae (Ma et al. 2001, see Table 5). A Robust Study Summary was completed for the Ma et al. (2001) study and the study was determined to be reliable with a high level of confidence (see Appendix 1). In addition, the EC50 of 0.0013 mg/L for green algae was divided by an assessment factor of 100 to account for the extrapolation from laboratory to field conditions, and inter- and intra- species variation in sensitivity. The resulting value of 0.000013 mg/L was used as the PNEC.

The conservative risk quotients (PEC/PNEC) of 0.75 and 0.02, calculated for diuron as a result of its industrial uses as a hardening agent in epoxy resins and a curing agent in epoxy adhesive, respectively, indicate that exposure values are unlikely to be high enough to cause harm to aquatic organisms. Since the majority of releases of this substance are likely into water at industrial manufacturing sites, and results of fugacity modeling indicate that most of the substance discharged into water will remain in that compartment, significant exposure of organisms at other types of locations or in media other than water are unlikely. While one of the risk quotients is approaching 1, this was the result of a conservative exposure scenario in which it was assumed that the total imported amount of the substance for non-pesticidal uses was released from an industrial facility a t a single site. This scenario is expected to be an overestimate of the actual risk associated with its non-pesticidal uses.

In addition, a conservative risk quotient for the diuron metabolite 3,4-DCA was calculated using scenarios and assumptions similar to those described above. Details regarding the inputs used to estimate this concentration and the output of the model are described in Environment Canada (2010b). Briefly, the quantity of 3,4-DCA was determined based on the total imported amount of diuron for the year 2006 for non-pesticidal uses and on the information presented in the US EPA 2003 RED document (US EPA 2003) in which 3,4-DCA was said to be formed through hydrolysis at <1% of applied diuron. The predicted environmental concentrations (PEC) for 3,4-DCA were 1 × 10-7 mg/L and 2 × 10-9mg/L as result of applications of diuron as a hardening agent in epoxy resins and as a curing agent in epoxy adhesive, respectively.

A conservative predicted no-effect concentration (PNEC) was also derived from the acute toxicity value identified for Daphnia magna, which was an LC50 value of 0.16 mg/L (European Commission 2006). The metabolite 3,4-DCA was observed to be more toxic to daphnids than the parent compound diuron (Sinclaire and Boxall 2003). Additionally, the LC50 of 0.16 mg/L for D. magna was divided by an application factor of 1000 to account for extrapolation from laboratory to field conditions, inter- and intra- species variation in sensitivity and extrapolation from acute to chronic exposure. The resulting value of 0.00016 mg/L was used as the PNEC. The conservative risk quotients (PEC/PNEC) of 0.0006 and 0.00002 were obtained for diuron metabolite 3,4-DCA as a result of industrial applications of diuron as a hardening agent in epoxy resins and as a curing agent in epoxy adhesive, respectively. This indicates that exposure to the degradation product 3,4-DCA associated with non-pesticidal industrial uses of diuron is unlikely to be high enough to be causing harm to aquatic organisms.

Diuron, as well as its metabolite 3,4-DCA, are therefore unlikely to cause ecological harm in Canada as a result of non-pesticidal uses of the substance.

Uncertainties in Evaluation of Ecological Risk

A number of experimental physico-chemical properties as well as aquatic and terrestrial toxicity studies were identified for diuron in existing databases or handbooks. Even though these sources are considered reliable and experimental procedures are believed to have been reviewed, in some cases experimental details and conditions were not available and could not be re-evaluated for this assessment. When available, experimental values for physico-chemical properties are used in QSAR models such as EPIsuite (EPIsuite 2008). Nevertheless, modelled data may be based on experimental values for which some degree of uncertainty exists since only summaries of studies were accessible for review.

Uncertainty with respect to the bioaccumulation potential of diuron was raised by a field study of the saltwater food web in the Vaccarès Lagoon, France. Data from the Vaccarès Lagoon study are not consistent with previous aquatic studies as well as the physico-chemical properties of the substance. However, following a careful analysis of the experimental details and the method by which diuron and other chemical residue data were gathered and analyzed, and the manner in which data were pooled before calculating trophic biomagnification potentials, the results of this study were considered to be uncertain and, therefore, of limited use in this assessment. The weight of evidence from all other available empirical and QSAR modelled bioaccumulation data indicate that diuron does not meet the bioaccumulation criteria set out in the Persistence and Bioaccumulation Regulations(Canada 2000). Studies addressing bioaccumulation potential of diuron, including bioconcentration and biomagnification factors, are not available for the Canadian environment.

Finally, there are uncertainties and limitations associated with the use of in vitro anti-estrogen/anti-androgen yeast screen assays to infer the endocrine disruption potential of diuron. These types of assays provide valuable insight into molecular mechanisms of action and their ability to alter normal endocrine function. However, they are restricted in their capacity to mimic whole-animal metabolism and the complexity of regulatory processes in animals and, as such, provide only preliminary evidence for endocrine disruption potential of the substance.

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Summary of Information Used as Basis for Human Health Risk Characterization

Health Canada has recently published a human health risk assessment for pesticidal uses of diuron[b], and the following text briefly summarizes the database.

In Canada, diuron is used primarily as a herbicide. All non-pesticidal uses of diuron are industrial in nature (Canada 2009, Canada 1988, Environment Canada 1988, Health Canada 2010). Therefore, exposure to the general population of Canada from non-pesticidal uses of diuron is considered to be negligible.

A summary of the health effects database and relevant reference values for diuron can be found in Appendices 3 and 4, respectively.

Diuron is considered to be of low acute toxicity when administered by the oral, dermal or inhalation routes. It was not irritating to the eyes or skin of rabbits and was not sensitizing in guinea pigs (US EPA 2003).

The toxicity review of diuron developed by the US EPA and adopted by the PMRA of Health Canada (2007) identified the haematopoietic system, urinary bladder and kidney as the primary target sites of subchronic and chronic exposure to diuron. Erythrocyte damage and compensatory haematopoiesis were observed in rats, mice and dogs. Urinary bladder wall thickening and swelling, together with epithelial focal hyperplasia, were observed at the high doses in rats and mice (US EPA 2003). Based on evidence of haemolytic anaemia and compensatory haematopoiesis observed in a combined chronic toxicity/carcinogenicity study in rats, the US EPA identified a chronic dietary LOAEL of 1.0 mg/kg-bw/day (US EPA 2003).

Effects on developing foetuses were not observed at doses below those that clearly induced maternal toxicity; hence, developmental effects are probably secondary to maternal toxicity. No effects on reproductive endpoints were observed in a two-generation reproductive toxicity study carried out in rats (US EPA 2002). A male reproductive toxicity study did not find significant differences in testosterone levels, sperm counts or sperm morphology. However, there were reductions in the weights of uteruses containing foetuses, and in the number of foetuses in females mated to males of the high-dose group when compared with those mated to control males (Fernandes et al. 2007).

Diuron has been classified by other agencies on the basis of carcinogenicity (European Commission 2000, 2001; US EPA 2002). Exposure to diuron via the diet increased the incidence of urinary bladder carcinoma in both male and female Wistar rats at the highest exposure concentration tested (Schmidt 1985). A non-significant increase in the incidence of kidney carcinoma was observed in the male rat, also observed only at the highest exposure. In a feeding study performed in NMRI mice, mammary gland adenocarcinomas were more prevalent in mice exposed to the highest concentration than in controls (Eiben 1983).

Additionally, results indicated that diuron may act as a promoter of urinary bladder carcinogenesis, but not of mammary carcinogenesis in Swiss mice. Diuron did not initiate or promote liver carcinogenesis in a rat liver bioassay (Grassi et al. 2007, de Moura et al. 2010). Genotoxicity testing performed both in vitro and in vivo provided predominantly negative results. However, in the absence of a fully elucidated mode of action framework for tumour development, and based on the chronic/carcinogenicity study in male Wistar rats, the US EPA conservatively established an oral cancer slope factor

(Q1*) = 1.91 x 10-2(mg/kg-bw/day)-1. Consideration of the available information regarding genotoxicity, and assessments of other agencies, indicates that diuron is not likely to be genotoxic. Accordingly, although the mode-of-induction of tumours is not fully elucidated, the tumours observed are not considered to have resulted from direct interaction with genetic material.

Diuron released into the environment is recognized to undergo metabolism that results in the formation of metabolites that are hydrolysable to 3,4-DCA (US EPA 2003). Metabolic studies have been carried out in both rats and dogs. In both models, the exposure concentration was from 25 to 2500 ppm in the diet for 9 months to 2 years. These studies examined a variety of tissues as well as urine and feces for residues of diuron. It was concluded that diuron accumulation or sequestration did not occur, and that tissue levels of diuron and its residues were proportional to the dose received. Diuron and its metabolites were excreted in both the urine and the feces with the predominant metabolite beingN-(3,4-dichlorophenyl)-urea (DCPU). Additional metabolites were found in small amounts and includedN-(3,4-dichlorophenyl)-methylurea (DCPMU), 3,4-DCA, 3,4-dichlorophenol and unmetabolized diuron (Hodge et al. 1967). Extensive metabolism was observed in rats exposed to14C-labelled diuron with only 2% of the parent compound recovered in the feces. Diuron was biotransformed byN-demethylation, ring hydroxylation and conjugation. In this study eight metabolites were detected with the predominant metabolites being identical to those listed previously (Wu 1996). Exposure to diuron and its primary metabolites as a result of pesticidal uses has been assessed by PMRA under the PCPA(Health Canada 2007). Since exposure to diuron from non-pesticidal uses is expected to be negligible, any potential health effects relating to exposure to these metabolites are not considered relevant to this assessment.

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Conclusion

Based on the information presented in this final screening assessment, it is concluded that diuron is not entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity or that constitute or may constitute a danger to the environment upon which life depends. Additionally, diuron meets the criteria for persistence, but does not meet the criteria for bioaccumulation potential as set out in the Persistence and Bioaccumulation Regulations (Canada 2000).

As exposure to the general population from non-pesticidal uses of diuron is expected to be negligible, it is concluded that diuron is not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health.

It is therefore concluded that diuron does not meet any of the criteria set out in section 64 of CEPA 1999.

This substance will be considered for inclusion in the Domestic Substances List inventory update initiative. In addition and where relevant, research and monitoring will support verification of assumptions used during the screening assessment.

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Appendix 1: Robust Study Summaries

Robust Study Summary: Aquatic iT

No. Item Weight Yes/No Specify
1 Reference: Ma et al. 2001. Acute toxicity of 33 herbicides to the green alga Chlorella pyrenoidisa. Bull Environ Contam Toxicol 66:536–541.
2 Substance identity: CAS RN n.a.a 330541
3 Substance identity: chemical name(s) n.a. diuron
4 Chemical composition of the substance 2 Y 50%
5 Chemical purity 1 Y Formulation was provided: 50% wettable powder
6 Persistence and/or stability of test substance in aquatic solution reported? 1 N
Method
7 Reference 1 Y
8 OECD, EU, national or other standard method? 3 Y Li SH. 1959. Mass culture of unicellular algae. Acta Hydrobiol Sin 4:121–129
9 Justification of the method–protocol if a standard method was not used 2 n.a.
10 Good Laboratory Practice (GLP) 3 N
Test organism
11 Organism identity: name n.a. Chlorella pyrenoidisa
12 Scientific or both scientific and common names reported? 1 Y green alga
13 Life cycle age and/or stage of test organism 1 n.a.
14 Length and/or weight 1 n.a.
15 Sex 1 n.a.
16 Number of organisms per replicate 1 initial concentration, 6 × 105 cells/mL
17 Organism loading rate 1 N
18 Food type and feeding periods during the acclimation period 1 n.a.
Test design / conditions
19 Test type (acute or chronic) n.a. Y Chronic, 96-hours, but authors describe it as acute
20 Experiment type (laboratory or field) n.a. Y lab
21 Exposure pathways (food, water, both) n.a. Y water
22 Exposure duration n.a. Y 96 hours
23 Negative or positive controls (specify) 1 Y negative (water only)
24 Number of replicates (including controls) 1 Y 3
25 Nominal concentrations reported? 1 Y 0–50 mg/L
26 Measured concentrations reported? 3 N
27 Food type and feeding periods during the long-term tests 1 n.a.
28 Were concentrations measured periodically (especially in the chronic test)? 1 n.a.
29 Were the exposure media conditions relevant to the particular chemical reported (e.g., for the metal toxicity: pH, DOC/TOC, water hardness, temperature)? 3 Y
30 Photoperiod and light intensity 1 Y
31 Stock and test solution preparation 1 Y
32 Was solubilizer or emulsifier used if the chemical was poorly soluble or unstable? 1 n.a.
33 If solubilizer or emulsifier was used, was its concentration reported? 1 n.a.
34 If solubilizer or emulsifier was used, was its ecotoxicity reported? 1 n.a.
35 Monitoring intervals (including observations and water quality variables) reported? 1 N
36 Statistical methods used 1 Y
Information relevant to the data quality
37 Was the endpoint directly caused by the chemical's toxicity, not by the organism’s health (e.g., when mortality in the control > 10%) or physical effects occurred (e.g., 'shading effect')? n.a. Y
38 Was the test organism relevant to the Canadian environment? 3 Y
39 Were the test conditions (pH, temperature, DO, other) typical for the test organism? 1 Y
40 Does system type and design (static, semi-static, flow-through, sealed or open, other) correspond to the substance's properties and organism's nature and habits? 2 Y
41 Was pH of the test water within the range typical for the Canadian environment (6 to 9)? 1 pH was not reported
42 Was temperature of the test water within the range typical for the Canadian environment (5 to 27°C)? 1 Y 25°C
43 Was toxicity value below the chemical’s water solubility? 3 Y
Results
44 Toxicity values (specify endpoint and value) n.a. n.a. EC50
45 Other endpoints reported; e.g., BCF/BAF, LOEC/NOEC (specify)? n.a. N
46 Other adverse effects (e.g., carcinogenicity, mutagenicity) reported? n.a. N
47 Score 81.8%
48 EC reliability code 1
49 Reliability category (high, satisfactory, low) High confidence

a n.a., not applicable or not available

Robust Study Summary: Aquatic iT

No Item Weight Yes/No Specify
1 Reference: Schafer et al. 1994. Biotests using unicellular algae and ciliates for predicting long term effects of toxicants. Ecotoxicol Environ Safety 27:64–81.
2 Substance identity: CAS RN n.a.a 330541
3 Substance identity: chemical name(s) n.a. diuron
4 Chemical composition of the substance 2 N
5 Chemical purity 1 n.a
6 Persistence and/or stability of test substance in aquatic solution reported? 1 n.a
Method
7 Reference 1 Y OECD
8 OECD, EU, national or other standard method? 3 Y OECD Guideline 201
9 Justification of the method– protocol if a standard method was not used 2 N n.a.
10 Good Laboratory Practice (GLP) 3 N n.a.
Test organism
11 Organism identity: name n.a. algae
12 Scientific or both scientific and common names reported? 1 Y Scenedesmus subspicatus
13 Life cycle age and/or stage of test organism 1 Y 3 days in culture and based on the cell count
14 Length and/or weight 1 N n.a.
15 Sex 1 N n.a.
16 Number of organisms per replicate 1 Y 6 × 104 cells/mL in triplicate
17 Organism loading rate 1 n.a
18 Food type and feeding periods during the acclimation period 1 N n.a.
Test design / conditions
19 Test type (acute or chronic) n.a. chronic, 72 hours
20 Experiment type (laboratory or field) n.a. lab
21 Exposure pathways (food, water, both) n.a. water
22 Exposure duration n.a. 72 hours
23 Negative or positive controls (specify) 1 Y negative control
24 Number of replicates (including controls) 1 3
25 Nominal concentrations reported? 1 Y not for diuron, atrazine is shown as an example
26 Measured concentrations reported? 3 N
27 Food type and feeding periods during the long-term tests 1 n.a.
28 Were concentrations measured periodically (especially in the chronic test)? 1 Y
29 Were the exposure media conditions relevant to the particular chemical reported (e.g., for the metal toxicity: pH, DOC/TOC, water hardness, temperature)? 3 Y
30 Photoperiod and light intensity 1 Y
31 Stock and test solution preparation 1 Y
32 Was solubilizer or emulsifier used, if the chemical was poorly soluble or unstable? 1 N n.a.
33 If solubilizer or emulsifier was used, was its concentration reported? 1 N n.a.
34 If solubilizer or emulsifier was used, was its ecotoxicity reported? 1 N n.a.
35 Monitoring intervals (including observations and water quality variables) reported? 1 Y
36 Statistical methods used 1 Y
Information relevant to the data quality
37 Was the endpoint directly caused by the chemical's toxicity, not by the organism’s health (e.g., when mortality in the control > 10%) or physical effects occurred (e.g. 'shading effect')? n.a. Y
38 Was the test organism relevant to the Canadian environment? 3 Y found in fresh water plankton
39 Were the test conditions (pH, temperature, DO,other) typical for the test organism? 1 Y
40 Does system type and design (static, semi-static, flow-through, sealed or open, other) correspond to the substance's properties and organism's nature and habits? 2 Y static flow
41 Was pH of the test water within the range typical for the Canadian environment (6 to 9)? 1 Y 7.1–7.2
42 Was temperature of the test water within the range typical for the Canadian environment (5 to 27°C)? 1 Y 20°C
43 Was toxicity value below the chemical’s water solubility? 3 Y
Results
44 Toxicity values (specify endpoint and value) n.a. EC50
45 Other endpoints reported; e.g., BCF/BAF, LOEC/NOEC (specify)? n.a. NOEC
46 Other adverse effects (e.g., carcinogenicity, mutagenicity) reported? n.a. no
47 Score 84.4%
48 EC reliability code 1
49 Reliability category (high, satisfactory, low) High confidence

a n.a., not applicable or not available

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Appendix 2: Persistence Bioaccumulation and Toxicity Model Inputs Summary Table

Phys-Chem/Fate Fate Fate Fate Fate Fate Fate PBT profiling Ecotoxicity
Model EPIWIN Suite

(all models, including: AOPWIN, KOCWIN, BCFWIN BIOWIN and ECOSAR)

STP (1)

ASTreat (2)

SimpleTreat (3)

(required inputs are different depending on model)

EQC (required inputs are different if Type I vs. Type II chemical) TaPL3 (required inputs are different if Type I vs. Type II chemical) OECD POPs Tool Arnot–
Gobas BCF/BAF
Model
Gobas Wolf BMF Model Canadian–POPs

(including: Catabol, BCF Mitigating Factors Model, OASIS Toxicity Model)

Artificial Intelligence

Expert System (AIES)/
TOPKAT/
ASTER

SMILES Code O=C(N(C)C)Nc(ccc(c1Cl)Cl)c1 x x
Molecular weight (g/mol) x (1, 2, 3) 233.1 x (I, II) x
Melting point (ºC) *158 158 x (I)
Boiling point (ºC) *353.86
Data temperature (ºC) 20 x (I, II)
Density (kg/m3) x (2)
Vapour pressure (Pa) *9.2 × 10-6 x (1, 3) 2.3 × 10-7 x (I)
Henry’s Law constant (Pa·m3/mol) *5.4 × 10-5 x (3)
Log Kaw
(Air–water partition coefficient, dimensionless)
x (2) x (II) x (II) x
Log Kow
(Octanol–water partition coefficient, dimensionless)
*2.68 x (1) x (I) x (I) x x x
Kow
(Octanol–water partition coefficient, dimensionless)
x (2, 3) 2.68
Log Koc
(organic carbon–water partition coefficient, L/kg)
Water solubility (mg/L) *42 x (1, 3) 35 x
Log Koa
(octanol–air partition coefficient, dimensionless)
x
Soil–water partition coefficient (L/kg)a x (II) x (II)
Sediment–water partition coefficient (L/kg)a x (II) x (II)
Suspended particles–water partition coefficient (L/kg)a x (2) x (II) x (II)
Fish–water partition coefficient (L/kg)b x (II) x (II)
Aerosol–water partition coefficient (dimensionless)c x (II) x (II)
Vegetation–water partition coefficient (dimensionless)a x (II)
Enthalpy (Kow) -20 (3)
Enthalpy (Kaw) 55 (3)
Half-life in air (days) x (I, II) x (I, II) x
Half-life in water (days) x (I, II) x (I, II) x
Half-life in sediment (days) x (I, II) x (I, II)
Half-life in soil (days) x (I, II) x (I, II) x
Half-life in vegetation (days)d x (I, II)
Metabolic rate constant (1/days) * *
Biodegradation rate constant (1/days) or (1/hour); specify x (3, 1/hour) (2, 1/days)
Biodegradation half-life in primary clarifier (t1/2-p) (hours) x (1)
Biodegradation half-life in aeration vessel (t1/2-s) (hours) x (1)
Biodegradation half-life in settling tank (t1/2-s) (hours) x (1)

a Derived from log Koc
b Derived from BCF data
c Default value
d Derived from half-life in water

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Appendix 3: Summary of Health Effects Informationa for CASRN 330-54-1: Urea,N'-(3,4-dichlorophenyl)-N,N-dimethyl- (Diuron)

Endpoint Lowest effect levelsb: results
Laboratory animals and in vitro
Acute toxicity

Lowest oral LD50(rat) = 4721 mg/kg for males, >5000 mg/kg for females (Rosenfeld 1985a).

Lowest inhalation LC50 (rat)>7100 mg/m3 (Kinney 1987).

Lowest dermal LD50 (rat) >2000 mg/kg (Rosenfeld 1985b).

Other Studies: Boyd and Krupa 1970; Kimmerle 1972; Thyssen 1974, 1975; Heimann 1981, 1984; Mihail 1981; Heimann and Thyssen 1983; Gaines and Linder 1986; Wandrag 1993

Short-term repeated- dose toxicity

LOEL (inhalation, rat): 47.6 mg/m3based on increased spleen weights in females. Wistar rats (10/sex/group) were exposed (head only) to 0, 6.6, 47.6, or 311 mg/m3 diuron in the form of respirable aerosols for 6 h/day, 5 days/week for 3 weeks. 92% of the relative mass had a mass median aerodynamic diameter of <5 µm. All rats survived the treatment period. At the high dose, all rats were affected in appearance and displayed piloerection after dosing from day 17 for males and day 15 for females. Piloerection disappeared before the start of each new dose. Body weight gain was low in all groups, including controls, and was attributed to exposure-induced stress due to exposure tubes. The number of Heinz bodies and reticulocytes increased in the mid- and high-dose females and high-dose males while erythrocyte counts were slightly decreased in correlation with concentration. An increase in the erythrocyte volume was observed in high-dose males and females and indicated reduced stability leading to reduced erythrocyte life. T3 and T4 levels were slightly reduced while TBK was increased. These findings suggest a slight decrease in thyroid function in the high-dose rats, which was more pronounced in males. Enzyme induction in the livers of high-dose males and females was evidenced by increased O-demethylase activity. Urinalysis was normal. Necropsy revealed dark, swollen spleens with increased incidence in animals exposed to the high dose, and in females exposed to the mid dose (Pauluhn 1986a).

LOEL (inhalation, rat): 37.4 mg/m3based on significantly increased reticulocyte counts and increased presence of Heinz bodies in females, which also had dark and enlarged spleens. A follow-up study to the study summarized above exposed Wistar rats (head only) to respirable aerosols at mean concentrations of 0, 4.1, 37.4, or 268.1 mg/m3 for 6 h/day, 5 days/week, for 8 weeks. Findings were very similar to those of the 3-week study (Pauluhn 1986b).

NOEL (dermal, rabbit): 1200 mg/kg bw/day. Only slight or mild oedema was noted in one male and one female of the high dose group. New Zealand white rabbits (5/sex/group) were exposed topically to 0, 50, 500, or 1200 mg/ kg bw/day for 21 days. Exposure was on the backs or rabbits whose hair was clipped. No treatment-related effects were observed on clinical signs (body weight, body weight gain, food consumption, haematology, organ weights or histopathology). U.S EPA considered this study to have a NOAEL of 1200 mg/kg bw/day for systemic toxicity (Mackenzie 1992).

A similar study exposed New Zealand white rabbits (6/sex/group) to 0, 50, or 250 mg/kg for 6 h/day, 5 days/week, for 3 weeks. In each group, 3 males and 3 females had their skin abraded with sandpaper before test substance application. No adverse effects were observed. Skin irritation was scored by the Draize system and skin fold thickness. Body weight monitoring, hematology and urinalysis did not identify any differences between control and exposure groups. The only finding considered noteworthy was a decrease in alkaline phosphatase levels in all groups. One high-dose male rabbit had a congested spleen as shown by histopathological analysis (Mihail and Schilde 1984).

Subchronic toxicity

LOEL (oral, rat): 1.6–1.8 mg/kg bw/day based on pigment (iron) deposition in the spleens of both males and females. Rats (BOR:WISW[SPF Cpb]) (10/sex/group) were exposed to 0, 4, 10, or 25 ppm in the diet (average doses: 0, 0.3, 0.7, or 1.6 mg/kg bw/day for males, and 0, 0.3, 0.8, and 1.8 mg/ kg bw/day for females) for 6 months. An increase in reticulocytes was observed in the high-dose females at all intervals and was significant (P < 0.05) at 12 and 26 weeks. Average haemoglobin concentrations were slightly depressed (non-significant, within 5% of controls) in the high-dose females at the 12- and 26-week intervals. Urinary bladder lesions were observed in all treated groups of both sexes but did not appear to be dose related. From control to high exposure, incidences of dilated blood vessels were 0/10 (i.e., 0 out of 10), 3/10, 1/10, and 3/10 in females and 0/10, 1/10, 2/10, and 3/10 in males; increased firmness occurred in 0/10, 5/10, 2/10, and 3/10 females, and in 0/10, 1/10, 2/10, and 3/10 males; reduced transparency was noted in 0/10, 1/10, 1/10, and 2/10 females. Hyperplasias of the urinary bladder were noted in 1 low dose male, and in 1 control, 2 low-dose, and 2 high-dose females. Thickening of the urinary epithelium, due to hypertrophy, was observed in 2 low-dose males and 1 high-dose male. Measurements were reported for the thickness of the female urinary epithelia. They were (from control to high dose): 437, 492, 448, and 486 µm. These results were considered equivocal by the US EPA. Pigment deposition, indicative of iron accumulation, was observed in the spleens of all rats, control and treated, but was increased in severity and extent in the high-dose groups (Schmidt and Karbe 1986).

LOEL (oral, rat): 6.43 mg/kg bw/day based on a non-significant increase in the occurrence of hyperplasia of the urinary epithelium in male Wistar rats. Male Wistar rats (15/group) were exposed to 0, 125, 500, or 2500 ppm (converted to 6.43, 12.9, or 64.5 mg/kg bw/day using Health Canada reference values) in the diet for 20 weeks. A significant (P < 0.0001) reduction in the mean body weight of rats exposed to the high dose was observed beginning at the 2nd week. Increased urinary pH was observed at the 4th and 13th weeks in the 6.43 mg/kg bw/day groups and in the 9th week in the 12.9 and 64.5 mg/kg bw/day groups. Macroscopic lesions were not observed in the livers, kidneys or urinary bladders of any rats nor were any microscopic lesions observed in their livers. Simple hyperplasia or the urinary epithelium was evident in 2/10, 2/10, and 7/10 (P < 0.005) rats in the 6.43, 12.9, and 64.5 mg/kg bw/day groups, respectively. Simple hyperplasia was also observed in the renal pelvis in rats of the mid- and high-dose groups (8/10 [P < 0.005] and 6/10 [P < 0.05], respectively). PCNA staining indicated that rats exposed to the high dose had significantly increased indices of cell proliferation in the urinary bladder mucosa. At the high dose, areas of necrosis and exfoliation were observed over the entire mucosa with hyperplasia and small round cells having uniform or pleomorphic microvilli present (Nascimento 2006).

Other studies considered: Wandrag 1996.

Chronic toxicity/ carcinogenicity

Non-carcinogenic effects

Non-cancer LOEL (oral, rat): 1.0 and 1.7 mg/kg bw/day for males and females, respectively, based on haemolysis coupled with haematopoiesis. Wistar rats (60/sex/group) were exposed to 0, 25, 250, or 2500 ppm (0, 1.0, 10, or 111 mg/kg bw/day for males, and 0, 1.7, 17, or 203 mg/kg-bw/day for females) for 24 months. At 12 months, 10 animals/sex/group were sacrificed for interim analysis. Survival was not affected. Discoloured (red) or bloody urine was observed in some high-dose males. Body weights were significantly (P < 0.01) decreased in males (12–15%) and females (6–14%) receiving the high dose. Weight gain was also depressed in these groups. At the mid dose, body weight gain was slightly decreased in males (4–6%) and was statistically significant (P < 0.05), but not considered biologically significant. Food efficiency was decreased for high-dose males and females. Haemolytic anaemia with compensatory haematopoiesis was observed (significant decreases in red blood cell [RBC] counts, haemoglobin levels, and hematocrit. Increased mean cell volume [MCV], mean cell haemoglobin [MCH], abnormal RBC form, reticulocyte counts and leukocyte counts) in mid- and high-dose males and females and in low-dose females. High-dose rats had increased plasma bilirubin levels (39–50%). In mid- and high-dose males and females, increased incidence of swollen, dark and discoloured spleens was noted. A dose-related increase in relative spleen weight (18–220%) was evident for all groups at both 12 and 24 months with increased severity evident in females. Spleen fibrosis was increased in incidence in high-dose males and females. Bone marrow activation was evident from increased hematopoietic bone marrow in mid- and high-dose rats at 12 and 24 months along with a decrease in fat marrow at 12 months. Effects on the urinary tract included bladder wall thickening in low- and high-dose males and in high-dose females. Epithelial focal hyperplasia in the urinary tract and renal pelvis was increased in incidence in high-dose males at 12 months, high-dose females at 12 and/or 24 months, and mid-dose females at 24 months. Liver changes included increased weight, swelling, discolouration, vacuolar cell degeneration, round cell infiltration and hyperaemia, which were observed at the low and mid doses in males and the low and high doses in females (Schmidt 1985).

Non-cancer LOEL (oral, mouse): 640.13 mg/kg bw/day in males, and 867 mg/kg bw/day in females based on haemolytic anaemia and liver toxicity in both sexes, and urinary bladder toxicity in females. NMRI(SPF HAN) mice (60/sex/group) were administered 0, 25, 250, or 2500 ppm (0, 5.4, 40.8, or 640.13 mg/kg-bw/day for males and 0, 7.5, 77.5, or 867.0 mg/kg bw/day in females) in the diet for 24 months. Ten mice/sex/group were sacrificed after 12 months for interim observations. Body weights were significantly reduced in high-dose males after 78 weeks of treatment. Overall food intake was increased in both males and females receiving the highest dose while food efficiency was decreased in males and females of the high-dose group (21–22% compared with controls). Small but significant increases in group MCV and MCH occurred in all treated mice at various times throughout the study. Increased reticulocyte counts were seen in males and females of the high-dose group. Changes in haematology were accompanied by increases in absolute and relative spleen weights, serum bilirubin and iron deposits in the spleen of the high-dose mice. Leukocyte counts were increased by 48 and 51% (P < 0.01) in male and females receiving the high dose at 18 months. After 2 years, high-dose females had a 98% increase in leukocyte counts. Serum alanine aminotransferase (ALT) activity was increased by 95% in males after 24 months and by 66% in females after 6 months at the high dose compared with controls. Absolute and relative liver weights were increased in males receiving the high dose after 24 months. Liver toxicity was indicated by microscopic examination, which identified increased mitosis, centrilobular hypertrophy in males, Kupffer cell clusters in males, enlarged and degenerative liver cells in females and single cell necroses in females. Effects on the urinary bladder included oedema, thickened mucosa and epithelial hyperplasia in high-dose females after 24 months of treatment. Hyperplasia was also evident in the high-dose females at 12 months. Uterine horn diameters were increased in females of the high-dose group (Eiben 1983).

Non-cancer LOEL (oral, dog): 18.8 mg/kg bw/day for males based on decreased body weight gains, and 93.8 mg/kg bw/day for females based on anaemia and body weight loss. Three dogs/sex/group were administered the test substance (80% purity) in feed at 0, 1.8, 9.4, 18.8, or 93.8 mg/kg-bw/day (converted from: 0, 25, 125, 250, or 1250/2500 ppm [2500 ppm for 2 weeks, basal diet for 3 weeks, 1250 ppm for the remainder of the 2-year study]) for 24 months. No treatment related deaths or clinical signs. Body weight was decreased over the entire study in the 2500/1250 ppm group in both males and females. There was also a decrease in body weight gain in males receiving 18.8 mg/kg bw/day during the first year of the study. High-dose dogs had normocytic to macrocytic normochromic anaemia between days 225 and 720 in males, and from week 2 on in females. Brown pigment in the Kupffer cells of all high-dose dogs was suggestive of haemolysis. High-dose animals had increased numbers of erythroid precursors, and a moderate reduction in marrow fat as assessed by histological preparations. At the high dose, absolute and relative liver weights and liver-to-brain weights were increased in males and females (Hodge and Downs 1964).

Carcinogenic effects (see non-carcinogenic effects for study protocols)

Oral (in diet), rat: Urinary bladder carcinoma incidence was increased in both males and females at the highest dose (33/49 [33 out of 49] males, 11/50 females) versus controls (1/50 males, 0/50 females). Most lesions were classified as transitional epithelial carcinomas. A non-significant increase in the incidence of urinary bladder papillomas and three neoplastic lesions in the renal pelvis in high-dose males were considered as treatment related. These findings were considered conclusive evidence of carcinogenicity by the US EPA (Schmidt 1985).

Oral (in diet), mouse: Incidences of mammary adenocarcinomas and ovarian luteomas were significantly (P < 0.05 in both instances) increased in the high-dose females compared with controls. Notably, the mammary adenocarcinoma incidence observed in the high-dose group was at or near the high range of incidences seen in historic controls. When all ovarian sex cord-stromal tumours were looked at together, there was not a significant increase in ovarian tumours (Eiben 1983).

Developmental toxicity

Oral developmental LOEL (rat): 400 mg/kg bw/day based on whole litter resorption, reduced foetal body weights and delayed ossification of the vertebrae and sternebrae. Developmental effects only occurred at levels exceeding those that were maternally toxic. Twenty-five pregnant Crl:COBS®CD®(SD)BR rats per group were exposed, by gavage, to 0, 16, 80, or 400 mg/kg bw/day on Gestational day (GD) 6–15. Dams were sacrificed on GD 20 and examined. Maternal effects: absolute body weights were significantly (P < 0.01) decreased from GD 6 to 11. From GD 6 to 9, mid- and high-dose dams had a net weight loss compared with a net gain by controls. High-dose dams continued to have weight loss from GD 9 to GD 12. Food consumption was significantly reduced in the mid- and high-dose groups (P < 0.01). After exposure was terminated, weight gain was significantly (P < 0.01) greater in mid- and high-dose groups. Developmental effects: two high-dose dams had total litter resorptions. Mean foetal body weights were significantly (P < 0.01) reduced in the high-dose groups compared with controls. No treatment related external or visceral malformations were observed in any dosing group. Delayed ossification of the vertebrae and sternebrae was evident (P < 0.05) in foetuses of the high-dose group. This study was considered “unacceptable/guideline” by the US EPA due to deficiencies in the variability of the test-article concentrations in the mid- and high-dose groups; however, the study was determined to be acceptable for the assessment of rat susceptibility (Dearlove 1986a).

Oral developmental NOEL (rabbit): 50 mg/kg bw/day. Groups of 24–25 artificially inseminated New Zealand white rabbits per group were exposed to 0, 2, 10, or 50 mg/kg bw/day by gavage on GD 7–19. All rabbits were sacrificed on GD 29 and examined. Maternal effects: Absolute body weights were significantly (P < 0.01) reduced in high-dose female rabbits compared with controls. Mean body weight gains were reduced in the same group compared with controls (P < 0.05 or P < 0.01) during gestational periods GD 10–13, 13–16 and 7–20. Weight gain was significantly (P < 0.01) greater in high-dose females compared with controls during the recovery period (GD 19–29). Food consumption was lower (P < 0.01) compared with that in controls in the high-dose group during the gestational intervals GD 13–16, 16–20 and 7–10. Developmental effects: A thorough examination of developmental endpoints identified no treatment-related effects (Dearlove 1986b).

Reproductive toxicity

Oral reproductive NOEL (rat): 1750 ppm (101–157 mg/kg bw/day for males and females of the F0and F1 genertions). Rats (30/sex/group) were fed 0, 10, 250, or 1750 ppm (0, 0.58, 14.8, and 101 mg/kg-bw/day for F0 males, and 0, 0.71, 18.5, and 131 mg/kg-bw/day for F0 females). One litter was produced by each generation. Dosing in the F–1 generation was 0.77, 18.9, and 139 mg/kg bw/day and 0.8, 22.1, and 157 mg/kg bw/day in males and females, respectively. F0 and F1 animals received either control or test substance containing diets for 73 and 105 days, respectively, prior to mating, and throughout mating, gestation, lactation, and until sacrifice. Low- and mid-dose groups of both exposed generations experienced occasional significant differences in body weights, body weight gains, food consumption and food efficiencies. These effects were not considered treatment related. At the high dose, F0 males and females had significantly (P < 0.05) reduced body weights, and body weight gains, compared with controls. Food consumption was significantly lower (P < 0.05) in both males and females of the F0group at several different time intervals. F0 males and females both had significantly reduced (P < 0.05) food efficiencies. The F1– generation experienced similar effects. F1 and –F2 pup body weights were significantly (P < 0.05) reduced in litters of rats exposed to the high dose compared with controls. No effects were noted on fertility indices, gestation length, pup survival, pup clinical observations, or pup anomalies (Cook 1990).

Male Wistar rats (18-20/group) were exposed for 30 days to 0, 125, or 250 mg/kg bw/day by gavage. Half of the control and treated males were housed with females for 15 days beginning 1 day after exposure was terminated. Terminable body weight was reduced by approximately 5% in the high-dose group. When the low and high dose were compared with each other, relative testis weight was decreased (P = 0.03) and absolute prostate and seminal vesicle weight (P = 0.01) was increased in the low-dose group. No differences were observed in the weights of reproductive organs compared with those of the control group. Absolute and relative liver weights were increased in both dosage groups compared with controls (P = 0.007 and P = 0.0007, respectively). No differences were observed in plasma testosterone concentrations in any group. Sperm counts in the testes and epididymis were unaffected. Sperm morphology was normal in all groups. In females mated to the 125 mg/kg bw/day exposure-group rats, reductions in maternal weight, weight of uterus with foetuses and number of foetuses were observed. Discreet centrilobular hypertrophy was observed in 80% of rats receiving the high dose (Fernandes et al. 2007).

Genotoxicity and related endpoints: in vivo

Chromosomal aberrations/micronucleus formation

Negative: Sprague Dawley rats were exposed once to 0, 50, 500, or 5000 mg/kg bw by gavage. Overt toxicity was noted at the high dose (mortality, body weight loss, ocular discharge, depression, laboured respiration, diarrhoea and tremors). The test was negative in male rats; however, a significant (P < 0.05) increase in the incidence of abnormal cells and aberrations per cell was seen when the data for high- and mid-dose males and females were combined. Combining the data for males and females also indicated a significant positive linear trend for abnormal cells and aberrations per cell. However, all values fell within the range of historical controls for this strain of rat (Ullman 1985; Cox 1997).

Negative: Female Hsd/Win:NMRI mice were exposed to a single intraperitoneal injection of 0 or 700 mg/kg. No evidence of clastogenicity was observed. Signs of toxicity included: apathy, roughened fur, staggering gait, sternal recumbency, spasm, twitching, difficulty breathing and eyelids being stuck together (Herbold 1998).

Positive: Swiss mice (6/group) were exposed intraperitoneally (IP) to 85, 170 or 340 (Maximum tolerated dose) mg/kg once. Bone marrow cells were examined for micronuclei 30, 48 and 72 h after injection. Exposure to 170 or 340 mg/kg significantly increased micronuclei at 30 and 48 h but not at 72 h. 85 mg/kg had no effect on the number of micronuclei observed (Agrawal et al. 1996).

Dominant lethal assay

Positive: Groups of 20 male mice were exposed IP to acute doses of 170 or 340 mg/kg bw, or via feed to concentrations of 3200 ppm (427 mg/kg bw/day converted using Health Canada standard values for mice) for 8 weeks. Male mice were co-habitated with two females 24 hours after exposure. An increase in pre-implantation loss and the number of dead implants was observed at all doses (Agrawal et al. 1997). However, the study was poorly reported and presented no statistical analyses.

Comet assay

Negative: Male Wistar rats (5/group) were exposed to 0, 125, 500 or 2500 ppm (converted to 6.43, 12.9, or 64.5 mg/kg bw/day using Health Canada reference values) in the diet for 20 weeks. Positive controls were exposed toN-methy-N-nitrosourea (NMU) 2h prior to testing. Peripheral blood cells and cells of the urinary bladder mucosa were analyzed by the alkaline elution comet assay. No evidence of increased DNA damage was found (Nascimento 2006).

Genotoxicity and related endpoints: in vitro

Bacterial reverse mutation

Negative: In independent trials, Salmonella typhimurium tester strains TA1535, TA97, TA98 and TA100 were negative for reverse mutation up to the highest dose tested (10 µg/plate +S9, 250 µg/plate -S9) with and without metabolic activation (Arce 1984; Herbold 1984; Poet 1985).

Mammalian cell gene mutation

Negative: In independent trials, Chinese hamster ovary (CHO)/HGPRT cells did not have increased incidences of forward gene mutation at the hprgt locus with or without activation up to doses that were cytotoxic (117 µg/mL) (Rickard 1985).

Unscheduled DNA synthesis (UDS)

Negative: No increased incidence of UDS was observed in rat primary hepatocytes treated with concentrations up to 76 µg/mL (cytotoxic dose) (Arce 1985).

Comet assay

Negative: Diuron did not cause induce DNA cross-links in Chinese hamster ovary cells as assessed by a modified comet assay (da Rocha et al. 2010).

Irritation/sensitization

Negative: All irritation cleared within 48 h of rabbits being exposed in the eye with the test substance. The substance is not considered a rabbit eye irritant (Rosenfeld 1985c).

Negative: All irritation cleared within 72 h of rabbits being exposed topically. The substance is not considered a rabbit skin irritant (Rosenfeld 1985d).

Negative: The substance was not sensitizing in guinea pigs (Rosenfeld 1985e).

Humans
Immunotoxicity No data identified.
Other studies In two human case studies, 39-year-old women ingested 20 or 39 mg/kg bw, respectively, along with aminonitroazole. No ill effects or signs of intoxication were observed (Gosselin et al. 1984; Hayes and Laws 1991).

a Health effects data dated before 2003 were obtained from the US EPA’s Registration Eligibility Decision Toxicology Chapter for Diuron, 2003 unless stated otherwise.
b LC50, median lethal concentration; LD50, median lethal dose; LOEC, lowest observed-effect concentration; LOEL, lowest observed-effect level.

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Appendix 4: Summary of Available Reference Values for Diuron

Reference value type Reference
Oral cancer slope factor
1.91 × 10-2 (mg/kg bw/day)-1

Chronic oral reference dose
2 × 10-3 mg/kg bw/day

Acceptable daily intake (food)
1.56 × 10-2 mg/kg bw/day

Maximum acceptable concentration (drinking water)
0.15 mg/L

US EPA 1998
IRIS 2010
Health Canada 1987
Health Canada 1987

Footnotes

[*] CAS RN= Chemical Abstracts Service Registry Number. The Chemical Abstracts Service (CAS) information is the property of the American Chemical Society and any use or redistribution, except as required in supporting regulatory requirements and/or for reports to the Government of Canada when the information and the reports are required by law or administrative policy, is not permitted without the prior written permission of the American Chemical Society.
[a] A determination of whether one or more of the criteria of section 64 are met is based upon an assessment of potential risks to the environment and/or to human health associated with exposures in the general environment. For humans, this includes, but is not limited to, exposures from ambient and indoor air, drinking water, foodstuffs, and the use of consumer products. A conclusion under CEPA 1999 on the substances in the Chemicals Management Plan (CMP) Challenge Batches 1-12 is not relevant to, nor does it preclude, an assessment against the hazard criteria specified in the Controlled Products Regulations, which is part of regulatory framework for the Workplace Hazardous Materials Information System [WHMIS] for products intended for workplace use. Similarly, a conclusion under CEPA 1999 does not preclude actions being undertaken under the Environmental Emergency Regulations based on consideration of the hazard and related properties of the substance.
[b] To obtain an electronic copy of the document, Proposed Acceptability for Continuing Registration: Re-evaluation of Diuron, please contact publications@hc-sc.gc.ca.

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