Gypsy Moth Management 

in the United States:

a cooperative approach

United States 

Department of Agriculture

Forest Service

Animal and Plant Health

Inspection Service

Newtown Square, PA

NA-MR-01-08

June 2008

Draft

Supplemental Environmental 

Impact Statement

Summary

Volume I of IV
Gypsy Moth Management 

in the United States:

a cooperative approach

Draft 

Supplemental Environmental 

Impact Statement

Volume I of IV

Summary

This volume summarizes the draft supplemental environmental impact 

statement (SEIS) for the gypsy moth. The complete draft SEIS is available 

on the Web at http://na.fs.fed.us/wv/eis. Print and CD copies are available 

from the United States Department of Agriculture (USDA) Forest Service, 

Northeastern Area State and Private Forestry, 180 Canfield Street, 

Morgantown, WV 26505-3101, phone 304-285-1523.

Photo Credits

Figure 1. Caterpillar (UGA1929085) USDA Forest Service Archives, 

www.forestryimages.org; pupa (UGA2652066) USDA 

Animal and Plant Health Inspection Service, Plant Protection 

and Quarantine Archives; www.forestryimages.org; moths 

(UGA2652079) USDA Animal and Plant Health Inspection 

Service, Plant Protection and Quarantine Archives, www.

forestryimages.org; egg mass (UGA3948096) William M. Ciesla, 

Forest Health Management International, www.forestryimages.

org

Figure 2. USDA Agricultural Research Service, www.ars.usda.gov/is/kids/

suburb/story2/microscope.htm
Contents

1 Gypsy Moth in the United States............................................................................................................................1

2 Proposed Action, Purpose, and Need......................................................................................................................1

3 Programmatic Nature of the Proposed Action ........................................................................................................3

4 Alternatives Considered..........................................................................................................................................3

Alternative 1—No Action..............................................................................................................................3

Alternative 2—Add Tebufenozide.................................................................................................................3

Alternative 3—Add Tebufenozide, and Add Other New Treatments Through the Application 

of the Protocol (Preferred Alternative) ..........................................................................................................3

5 Issues Identified ......................................................................................................................................................4

6 Risk Assessments and Risk Characterization .........................................................................................................5

Overview........................................................................................................................................................5

Hazard Quotients ...........................................................................................................................................6

7 Effects of the Gypsy Moth......................................................................................................................................6

Risk to Human Health....................................................................................................................................6

Risk to Nontarget Organisms.........................................................................................................................7

8 Effects of Treatments..............................................................................................................................................7

Currently Approved Treatments.....................................................................................................................8

Bacillus thuringiensis var. kurstaki (B.t.k.) (Alternatives 1, 2, 3) ....................................................8

Risk to Human Health .........................................................................................................8

Risk to Nontarget Organisms...............................................................................................8

Diflubenzuron (Alternatives 1, 2, 3).................................................................................................8

Risk to Human Health .........................................................................................................8

Risk to Nontarget Organisms...............................................................................................8

Disparlure (as Used in Mating Disruption and Mass Trapping) (Alternatives 1, 2, 3).....................9

Risk to Human Health .........................................................................................................9

Risk to Nontarget Organisms...............................................................................................9

Dichlorvos (as Used in Mass Trapping) (Alternatives 1, 2, 3).........................................................9

Risk to Human Health .........................................................................................................9

Risk to Nontarget Organisms...............................................................................................9

Gypchek (Alternatives 1, 2, 3)........................................................................................................10

Risk to Human Health .......................................................................................................10

Risk to Nontarget Organisms.............................................................................................10

Sterile Insect Technique (Alternatives 1, 2, 3)................................................................................10

Risk to Human Health .......................................................................................................10

Risk to Nontarget Organisms.............................................................................................10

New Treatment of Tebufenozide (Alternatives 2 and 3)..............................................................................10

Risk to Human Health .......................................................................................................10

Risk to Nontarget Organisms.............................................................................................10

New Treatments That May Be Available in the Future Under Alternative 3...............................................11

9 Mitigating Adverse Effects...................................................................................................................................11

Figures and Tables

Figure 1. The gypsy moth life cycle has four stages..................................................................................................2

Figure 2. Gypsy moth hairs can cause irritation ........................................................................................................7

Table 1. Comparative Hazard Quotients (HQs) for the effects of gypsy moths and treatments on 

human health and nontarget organisms.....................................................................................................................12

Supplemental Environmental Impact Statement

Summary - Page 

1 Gypsy Moth in the United 

States.

The gypsy moth is a significant nonnative forest pest in 

the United States (Figure 1). At least 587 million acres 

(238 million hectares) of trees are susceptible to gypsy 

moth feeding and defoliation. Also at risk are countless 

urban and rural forested areas throughout the country 

where susceptible plants grow naturally or are planted.

Although both European and Asian strains exist, only 

the European strain is currently present in the United 

States. The European gypsy moth was brought to 

the United States and accidentally released in eastern 

Massachusetts around 1869. Since then, it has 

continued to spread into uninfested areas naturally and 

by artificial movement by people. The Asian strain 

occasionally has been found in the United States, 

and has been eliminated whenever it has been found. 

Unlike European female gypsy moths, which cannot 

fly, the Asian moth poses a greater risk of spread 

because females can fly and deposit egg masses miles 

from where they fed as caterpillars.

The gypsy moth continues to be a problem as it 

spreads: over the last 100 years history shows that 

gypsy moth outbreaks cause widespread defoliation, 

tree mortality, environmental and public health risks, 

and public outcry to control the outbreaks. 

2 Proposed Action, Purpose, 

and Need.

The Forest Service and the Animal and Plant Health 

Inspection Service (APHIS), as co-lead agencies, 

propose to supplement the 1996 Record of Decision 

(ROD) for the 1995 Environmental Impact Statement 

(EIS): Gypsy Moth Management in the United States: 

a cooperative approach. The Forest Service and 

APHIS are proposing to add new treatment options, 

which are described in Section 4 on the Alternatives 

Considered. New treatments that were not available 

when the 1995 EIS was written would provide gypsy 

moth managers with more flexibility in conducting 

suppression, eradication, and slow-the-spread projects. 

Making new treatments available is also expected 

to improve the gypsy moth management program, 

because each new treatment developed over the last 30 

years has proven to be safer, more cost-efficient, easier 

to use, and often more effective than older treatments. 

The supplement also provides new information on the 

gypsy moth and treatments since the 1995 EIS.

Under the 1996 decision, the three strategies of 

suppression, eradication, and slow the spread 

established a management program to address 

the full spectrum of gypsy moth populations 

found in the United States. The three strategies 

complement one another, although they differ in 

objectives and geographic locations:

Suppression reduces damage caused 

by outbreak populations of gypsy moth 

caterpillars in the generally infested area. 

Eradication prevents establishment of the 

gypsy moth in uninfested areas by eliminating 

isolated infestations that occur as a result of 

human movement of this pest. 

Slow the Spread (STS) slows the rate of spread 

of the gypsy moth from the generally infested area 

and prevents infestation of 8 million acres per year 

in the transition area, thus delaying the impacts 

and costs that occur as the gypsy moth infests new 

areas.
Summary

Summary - Page 

Gypsy Moth Life Cycle

Caterpillar

8 weeks during 

spring and early summer

Young caterpillars are black, 

developing double rows of red 

and blue spots as they mature. 

Insecticide application usually 

occurs when both the caterpillars 

and foliage are in the early stages 

of development.

Pupa

2 weeks during 

spring-summer

The female pupa is larger 

than the male; both are a dark 

reddish-brown. Caterpillars 

pupate in protected areas, 

increasing the chance of 

accidental movement of pupae 

by humans.

Adult

Several days 

during summer

Male adults are brown or gray. 

Feather-like antennae detect 

the pheromone emitted by the 

female, which is white with 

small, black markings.

Egg Mass

9 months 

summer-spring

Females lay buff-colored egg 

masses almost anywhere; 

because the egg life stage is the 

longest lasting, this stage is most 

frequently accidentally moved 

by humans.

Figure 1. The gypsy moth life cycle has four stages.
Supplemental Environmental Impact Statement

Summary - Page 

Treatments approved for use in the strategies are 

Bacillus thuringiensis var. kurstaki (B.t.k.), Gypchek, 

diflubenzuron, mating disruption (using disparlure), 

mass trapping (using disparlure alone or with 

dichlorvos), and the sterile insect technique. 

The current overall USDA gypsy moth management 

program supports an integrated pest management 

approach that includes planning, detection, evaluation, 

monitoring, defining acceptable damage, and using 

appropriate management practices to prevent or control 

gypsy moth-caused damage and losses in the United 

States. 

3 Programmatic Nature of the 

Proposed Action.

Like the 1996 ROD, the decision to be made as a 

result of this SEIS will be programmatic. It will apply 

to the overall gypsy moth management program of

suppression, eradication, and slow-the-spread projects. 

Specific decisions to undertake any treatment projects 

will be made following site-specific environmental 

analyses conducted and documented in accordance 

with agency implementing procedures for the National 

Environmental Policy Act. Project proposals will also 

be analyzed for compliance with Federal laws, such 

as the Endangered Species Act and National Historic 

Preservation Act, as well as any applicable State laws.

4 Alternatives Considered.

To learn the concerns of interested and affected people 

the preparers of the draft SEIS invited public comments 

for 45 days via a notice in the Federal Register (69 Fed. 

Reg. 23492, April 29, 2004) soliciting feedback. The 

result of the public involvement process and internal 

scoping within the agencies was the identification of 

three alternatives.

Alternative 1—No Action.

Alternative 1 would maintain the 1996 decision and 

the current gypsy moth management program; no new 

treatments would be added to the approved treatments.

Alternative 2—Add Tebufenozide.

Alternative 2 would add the insecticide tebufenozide to 

currently approved treatments.

Alternative 3—Add Tebufenozide, and 

Add Other New Treatments Through 

the Application of the Protocol 

(Preferred Alternative).

Alternative 3 would add the insecticide tebufenozide 

and add other treatment(s) that may become available 

in the future for managing gypsy moths, to currently 

approved treatments. A new treatment would be 

available for use upon the agencies’ finding that the 

treatment poses no greater risks to human health 

and nontarget organisms than are disclosed in this 

draft SEIS for the currently approved treatments and 

tebufenozide.

The protocol for making the necessary finding that a 

treatment is authorized by this Alternative is as follows: 

1. Conduct a human health and ecological risk 

assessment (HHERA). In this risk assessment 

review all scientific studies available 

for toxicological and environmental fate 

information relevant to effects on human health 

and nontarget organisms. Use this information 

to estimate risk to human health and nontarget 

organisms. Include these four elements in the 

HHERA: (a) hazard evaluation, (b) exposure 

assessment, (c) dose-response assessment, and 

(d) risk characterization. The HHERA will do 

the following: 
Summary

Summary - Page 

• Identify potential use patterns, including 

formulation, application methods, 

application rate, and anticipated frequency 

of application.

• Review chemical hazards relevant to the 

human health risk assessment, including 

systemic and reproductive effects, skin 

and eye irritation, dermal absorption, 

allergic hypersensitivity, carcinogenicity, 

neurotoxicity, immunotoxicity, and 

endocrine disruption.

• Estimate exposure of workers applying the 

chemical.

• Estimate exposure of members of the public.

• Characterize environmental fate and 

transport, including drift, leaching to 

groundwater, and runoff to surface streams 

and ponds.

• Review available ecotoxicity data including 

hazards to mammals, birds, reptiles, 

amphibians, fish, and aquatic invertebrates.

• Estimate exposure of terrestrial and aquatic 

wildlife species.

• Characterize risk to human health and 

wildlife.

2. Conduct a risk comparison of the human 

health and ecological risks of a new treatment 

with the risks identified for the currently 

authorized treatments and tebufenozide. This 

risk comparison will evaluate quantitative 

expressions of risk (such as hazard quotients) 

and qualitative expressions of risk that put the 

overall risk characterizations into perspective. 

Qualitative factors include scope, severity, and 

intensity of potential effects, as well as temporal 

relationships such as reversibility and recovery.

3. If the risks posed by a new treatment fall 

within or below the range of risks posed by the 

currently approved treatments and tebufenozide, 

publish a notice in the Federal Register of the 

agencies’ preliminary findings that the treatment 

meets the requirements of Alternative 3. The 

notice must provide a 30-day review and 

comment period and must advise the public 

that the HHERA and the risk comparison are 

available upon request.

4. If consideration of public comment leads to 

the conclusion that the preliminary finding is 

correct, publish a notice in the Federal Register 

that the treatment meets the requirements of 

Alternative 3 and, therefore, is authorized by 

that Alternative for use in the USDA gypsy moth 

management program. The Forest Service and 

APHIS will make available to anyone, upon 

request, a copy of the comments received and 

the agencies’ responses.

Like the 1996 Record of Decision, the decision 

to be made as a result of this draft SEIS will be 

programmatic. Decisions to use specific treatments 

in projects, including new treatments authorized 

under the protocol in Alternative 3, will be made after 

site-specific environmental analyses are conducted 

and documented in accordance with agency NEPA 

implementing procedures.

5 Issues Identified.

Two issues were derived from public involvement for 

this draft SEIS:

Issue 1—Risk to human health

Issue 2—Risk to nontarget organisms

The effects of each of the treatments on the identified 

issues are summarized in Section 8.
Supplemental Environmental Impact Statement

Summary - Page 

6 Risk Assessments and Risk 

Characterization.

Overview.

The consequences of the treatments in each alternative 

were determined by risk assessment for each treatment 

as well as for gypsy moth (no treatment) and by a risk 

comparison among the treatments and gypsy moth.

A risk assessment provides a logical process for 

evaluating data and analyzing potential effects of the 

gypsy moth and treatments. Risk assessments take into 

account the manner in which treatments are used in 

gypsy moth projects, including how treatment agents 

are applied, the amount applied, and the types of areas 

that receive treatment. 

Standard steps in the risk assessment process were 

followed:

• Hazard identification--gathers known 

information from laboratory and field studies on 

toxicity of the gypsy moth and treatment agents.

• Exposure assessment--describes the nature and 

magnitude of contact with the gypsy moth and 

with treatment agents as they are used in gypsy 

moth treatment projects.

• Dose-response assessment--determines 

how much exposure to the gypsy moth and 

to treatment agents is needed to produce the 

response (effect) described in the hazard 

identification. 

• Risk characterization--combines information 

from previous steps to describe the plausibility of 

observing certain effects of the gypsy moth and of 

treatments.

Each step in a risk assessment is accompanied by 

uncertainties, caused by limitations either in the 

available data or in the ability to relate the data to 

scenarios of concern. To compensate for uncertainties, 

risk assessment results tend to be conservative, 

meaning they are more likely to overestimate risks than 

to underestimate them.

Human Health and Ecological Risk Assessments 

(HHERA) were prepared by risk assessment experts 

(Syracuse Environmental Research Associates, Inc. 

[SERA]), using the best available data. The HHERAs 

also underwent independent technical review by 

other recognized experts in risk assessment methods, 

toxicology, and other applicable fields (consultants 

retained by SERA, and toxicologists and program 

specialists from APHIS and the Forest Service). The 

HHERAs cover the issues raised in scoping for this 

SEIS for both human health (human health assessment 

portion of HHERA) and nontarget organisms 

(ecological risk assessment portion of HHERA). 

Many uncertainties are inherent in conducting and 

interpreting risk assessments; however, the data 

available on the agents covered by the risk assessments, 

modeling, equations and statistics all taken together 

with the understanding of uncertainties provide 

adequate information to characterize the relative 

hazards associated with the agents evaluated. To 

compensate for missing data and any uncertainties in 

the data, numerical uncertainty factors are used in the 

dose-response assessments for potential human health 

effects, and conservative assumptions are used in both 

human health and ecological risk assessments. In 

addition, it is virtually impossible to precisely calculate 

an exposure value for every situation that may arise. 

Therefore, models, equations, and statistical techniques 

were used to quantify both plausible and extreme 

exposures and to use ranges of toxicity values to reflect 

ranges of sensitivity. These ranges for exposure and 

toxicity are then used to numerically characterize risk 

with hazard quotients that are typically expressed as 

central estimates with upper and lower bounds.
Summary

Summary - Page 

HHERAs were prepared for each of the treatments 

in the alternatives and for the gypsy moth itself. The 

relative risks of the insecticides and treatments were 

illustrated in a risk comparison evaluation.

Hazard Quotients. 

Risks to human health and to nontarget organisms 

can be estimated numerically using hazard quotients 

(HQs). HQs can be calculated only for effects on 

populations of biotic (living) organisms. The HQ is 

a screening tool commonly used in risk assessments. 

The HQ is a ratio of the exposure estimate for a 

particular and defined situation (labeled or prescribed 

conditions) for a representative population (human or 

nontarget species), divided by an effect level (dose or 

concentration level). The HQ takes into account the 

inherent toxicity of a substance, as well as its ability to 

produce specific effects on an organism (or population 

of organisms), and the degree of exposure. Table 1

provides the HQs for all of the treatments and for the 

gypsy moth.

As an example, refer to the upper bound of the HQ 

for B.t.k. for nontarget aquatic species--0.5, in Table 

1. This HQ was derived from an exposure estimate 

of 0.24 mg/L, which is calculated as the peak 

concentration of the B.t.k. formulation in water after 

a direct spray. This exposure estimate serves as the 

numerator for the HQ. The toxicity value of 0.45 

mg/L is the NOEC (no observed effect concentration) 

from a reproduction study in Daphnia magna, an 

aquatic invertebrate. This toxicity value serves as the 

denominator for the HQ. Thus, the HQ is calculated as 

follows:

HQ = exposure estimate/toxicity value 

= 0.24 mg/L / 0.45 mg/L 

= 0.533... ≈ 0.5.

Note that the HQ in the above example is rounded to 

one significant place. This is a common practice in 

presenting HQ values except for those in which the 

level of concern is marginally exceeded, i.e., an HQ of 

1.45 would be rounded to 1.4 but not to 1.0.

In risk management, the HQ must be used in 

conjunction with other factors and characteristics 

of a substance, such as the quality and quantity of 

substantiating evidence (published scientific literature, 

data, models, and risk assessments done by others such 

as industry and universities), the severity of potential 

adverse effects, and the nature of the affected species 

and populations. 

In some cases numerical expressions of risk (HQs) do 

not adequately convey the potential for hazard. For 

example, a high HQ for a mild effect, such as skin rash, 

is probably more acceptable than a much lower HQ for 

a more serious effect like neurotoxicity. Therefore, the 

use of HQ as an expression of risk and “acceptability” 

requires that a qualitative perspective also be injected 

into the analysis. Ecological risk assessments often 

involve considerations of many different species 

of plants and animals, and abiotic factors, and their 

interrelationships and interactions. Invariably, few 

data sets are available, and field studies provide only 

an overview of the complex interrelationships and 

secondary effects among species. Human health risk 

assessments and ecological risk assessments cannot 

offer a guarantee of safety. Both risk assessments offer 

a way to estimate the adverse effects and the severity of 

those adverse effects.

7 Effects of the Gypsy Moth.

Risk to Human Health.

Following exposure to gypsy moth caterpillar hairs 

(Figure 2) during gypsy moth outbreaks, children and 

others who spend time outside may develop rashes 

or irritation of the eyes or respiratory tract. Some 

individuals may develop an allergy to the gypsy moth 

following repeated exposures over 1 or more years.
Supplemental Environmental Impact Statement

Summary - Page 

Risk to Nontarget Organisms.

Environmental effects due to the gypsy moth vary, 

depending on population levels, the amount of 

defoliation, and the duration of an outbreak. The 

most pronounced effects occur when the gypsy moth 

causes heavy defoliation.. After a single year of 

heavy defoliation, tree condition suffers and mortality 

increases. Production of both wood and hard mast 

(nuts) temporarily declines, and the growth rate of 

many shrubs and herbaceous plants may increase.

Two years of heavy defoliation greatly reduce the 

production of wood, hard mast (nuts), and soft mast 

(berries). Shoestring fungus and two-lined chestnut 

borer, which attack and kill trees weakened by 

defoliation, become more abundant. Mortality is likely 

within 5 years, both among oaks and among species 

less favored by the gypsy moth. 

Three years of heavy defoliation cause high mortality 

in oaks and less-favored hosts; wood growth is 

drastically reduced, and production of hard mast will 

likely cease for at least 5 years. Regeneration of young 

forest to a mature forest requires decades, and in areas 

where trees less-favored by the gypsy moth remain, 

stands are dominated by species such as red maple, 

yellow poplar, black cherry, and yellow or black birch. 

Decomposition of leaf fragments and caterpillar 

droppings reduce oxygen levels in water and result in 

dramatic increases in algae, reducing acid-neutralizing 

capacity and increasing watershed yields. 

Increased exposure to sunlight caused by defoliation 

results in seasonal elevations in the temperature of 

soil and leaf litter, which may temporarily reduce soil 

moisture content. These factors can lead to shortlived increases in the rates of soil decomposition, 

mineralization, and plant productivity.

Heavy defoliation can affect animals, fish, and birds. 

Populations of small mammals (such as the gray 

squirrel) decline as do some bird species, although 

woodpecker populations may increase. Trout may 

decrease in number or disappear from small streams, 

along with small crayfish and snails. Forest-feeding 

moths and butterflies, particularly those that feed on 

oak also are likely to decline, as may other forestdwelling invertebrates. 

8 Effects of Treatments.

All of the treatments described in this SEIS may 

indirectly help to maintain existing forest conditions 

and environmental quality by delaying increases 

in gypsy moth populations, thereby protecting tree 

foliage.

Figure 2. Gypsy moth hairs can cause irritation.
Summary

Summary - Page 

Currently Approved Treatments.

Bacillus thuringiensis var. kurstaki (B.t.k.) 

(Alternatives 1, 2, 3).

Risk to Human Health.

Direct exposure to B.t.k. applications may cause some 

individuals (most likely project workers) to develop 

minor and transient irritation of the skin, eyes, or 

respiratory tract. Mating disruption with disparlure 

will most likely be the only treatment used in the same 

project area as B.t.k. These treatments have different 

modes of action and are applied weeks apart; therefore, 

no cumulative effects are expected between B.t.k. and 

disparlure treatments. 

Risk to Nontarget Organisms.

Permanent changes in non-target caterpillar populations 

are not likely following treatment projects that use 

B.t.k. An exception might occur in certain habitat types 

that support small, isolated populations of moths and 

butterflies that are highly susceptible to B.t.k. Repeated 

treatment of areas with B.t.k. could potentially affect 

some species of spring-feeding butterfly and moth 

caterpillars. Since B.t.k. is not used in the same 

locations as other treatments that could affect nontargets, there is no cumulative effect of different 

treatments with B.t.k. on spring-feeding caterpillars

Diflubenzuron (Alternatives 1, 2, 3).

Risk to Human Health. 

No human health effects are likely from exposure 

to diflubenzuron as used in the USDA gypsy moth 

management program. Exposure to very high levels 

of diflubenzuron may produce a detectable increase 

in methemoglobin, an abnormal blood pigment 

that reduces the oxygen-carrying capacity of blood. 

Exposure to other methemoglobinemia-inducing 

compounds in the environment may contribute to a 

cumulative effect. Individuals exposed to combustion 

smoke or carbon monoxide (agents that also cause 

oxidative damage to blood) may be at increased risk 

of developing methemoglobinemia. Individuals 

exposed to high levels of nitrates, either in air or in 

water, demonstrate increased levels of methemoglobin 

and may be at increased risk with exposure to 

compounds such as diflubenzuron. Some individuals 

have congenital methemoglobinemia and may be at 

increased risk of adverse effects from compounds that 

induce methemoglobinemia. Diflubenzuron rapidly 

dissipates from vegetation and is broken down by 

sunlight; in the environment the compound degrades 

to 4-chloroaniline, which the EPA considers a potential 

carcinogen. This is the only identified potential 

carcinogen associated with any of the agents to control 

gypsy moth. The compound is not expected to be 

present in significant amounts during application since 

4-chloroaniline does not form during application. The 

scenario of greatest concern involving 4-chloroaniline 

is a cancer risk from drinking contaminated water. 

Risk to Nontarget Organisms.

Moth and butterfly caterpillars, grasshoppers, parasitic 

wasps, some beetles, spiders, sawflies, aquatic 

insects, and bottom-dwelling and immature freefloating crustaceans may be affected by application 

of diflubenzuron. Diflubenzuron treatments are 

applied to the top of the tree canopy and the amount 

of diflubenzuron residue begins to diminish soon after 

the application. The population reduction is greater 

for those species that feed in the upper canopy as 

compared with those in the mid and lower canopy. 

Diflubenzuron may cause a reduction in some aquatic 

invertebrate populations. Diflubenzuron reduces 

numbers of stream invertebrates that process detritus; 

however, field studies have shown no decline in 

detrital decomposition rates. The populations of 

some invertebrates that feed on algae are reduced by 

diflubenzuron. An increase in algae could occur after 

the loss of algal herbivores, however, this has not been 

observed in field studies. 

Birds are not directly affected by exposure to 

diflubenzuron. Some insectivorous species may show 

subtle changes, such as a switch in diet, reduced fat 
Supplemental Environmental Impact Statement

Summary - Page 

loads, and expanded foraging territories. Similar 

changes may occur in bats that feed primarily on moths 

and butterflies.

Diflubenzuron is generally not used in conjunction 

with other treatments; however, diflubenzuron might 

be applied to the same area in multiple years for 

eradication projects. In that case, diflubenzuron might 

have a cumulative effect on nontarget invertebrates, 

such as caterpillars of moths and butterflies, 

grasshoppers, parasitic wasps, aquatic insects, bottom 

dwelling crustaceans, and immature free-floating 

crustaceans. Diflubenzuron applications as used in 

USDA treatment projects will otherwise have no 

cumulative effects.

Disparlure (as Used in Mating Disruption and 

Mass Trapping) (Alternatives 1, 2, 3).

Mating disruption entails the aerial application of 

tiny plastic flakes containing disparlure, the synthetic 

version of the gypsy moth sex attractant. This 

treatment confuses male moths and prevents them from 

locating and mating with females.

Risk to Human Health.

After direct contact with disparlure, a person (most 

commonly, a project worker) may attract male gypsy 

moths. Although this attraction could last for years, 

and could be annoying, there are no data to show it 

poses any health risk. The general public is not likely 

to be exposed to sufficient amounts of disparlure to 

experience this effect. Since disparlure seems to persist 

in humans, repeated exposures of disparlure will attract 

the gypsy moth. No information is available on the 

interaction of disparlure with other control agents or 

other chemicals usually found in the environment.

Risk to Nontarget Organisms.

Disparlure has low toxicity to vertebrates and is 

specific to the gypsy moth in North America. As used 

in mating disruption (and as an attractant in mass 

trapping), disparlure is not likely to affect populations 

of non-target organisms. Since disparlure attracts 

only the gypsy moth in North America, no cumulative 

effects are expected on nontarget organisms. 

Dichlorvos (as Used in Mass Trapping) 

(Alternatives 1, 2, 3).

Two types of traps are used in mass trapping; both 

contain disparlure to attract male moths. The smaller 

delta trap captures moths with a sticky inside surface 

but contains no dichlorvos. The large milk carton trap 

contains a pest strip impregnated with the insecticide 

dichlorvos. Both traps are also commonly used for 

survey purposes.

Risk to Human Health.

Dichlorvos as used in milk carton traps would 

pose a health risk to humans only if the individual 

were to disassemble the trap and tamper with the 

dichlorvos-impregnated strip. Skin contact with the 

strip or eating the strip could inhibit the production 

of acetylcholinesterase. This enzyme prevents the 

accumulation of acetylcholine, the buildup of which 

can impair the function of the nervous system. 

Obvious signs of toxicity to the nervous system are 

possible but unlikely. Exposure to other substances 

that inhibit acetylcholinesterase, including similar 

insecticides, could have a cumulative effect with 

dichlorvos. The carcinogenic potential of diclorovos 

has been classified as “suggestive” under the 1999 

Environmental Protection Agency Cancer Guidelines.

Risk to Nontarget Organisms. 

Invertebrates that inadvertently enter delta or milk 

carton traps are likely to die. Invertebrates that come 

into contact with a dichlorvos strip that has accidently 

fallen on the ground, on vegetation, or in water might 

also be adversely affected. The potential for adverse 

effects decreases over time as dichlorvos dissipates 

from the strip. Large animals, such as bears, that may 

tamper with traps are not likely to be affected by the 

dichlorvos strips. Experience with traps used in mass 
Summary

Summary - Page 10

trapping and survey programs shows that there are no 

cumulative effects on nontarget organisms even over 

years of use. 

Gypchek (Alternatives 1, 2, 3).

Risk to Human Health.

Irritation of the eyes, skin and respiratory tract is 

possible from exposure to Gypchek. Gypchek contains 

gypsy moth parts and may cause irritant effects 

similar to those caused by the gypsy moth caterpillars. 

Consequently, exposure to both the gypsy moth 

caterpillars and Gypchek could be cumulative, although 

there are no data showing this occurs.

Risk to Nontarget Organisms.

Since Gypchek is specific to the gypsy moth, no effects 

or cumulative effects on nontarget organisms are 

expected.

Sterile Insect Technique (Alternatives 1, 2, 3).

The release of sterile insects adds large numbers of 

sterile gypsy moths to an area, resulting in population 

reduction and eventual elimination of the infestation. 

Risk to Human Health.

The sterile insect technique temporarily increases 

the number of gypsy moths in the treatment area, 

increasing both the chance of effects due to the gypsy 

moth and contact with gypsy moth caterpillars.

Risk to Nontarget Organisms.

No effects or cumulative effects on non-target species 

are expected.

New Treatment of Tebufenozide (Alternatives 2 

and 3).

Risk to Human Health.

Exposure to very high levels of tebufenozide may 

increase detectable levels of methemoglobin, an 

abnormal blood pigment that reduces the oxygencarrying capacity of the blood. These exposure levels 

far exceed those exposures expected to occur in project 

workers and the general public from the USDA gypsy 

moth management program. The presence of other 

compounds that raise levels of methemoglobin, such as 

cigarette or other combustion smoke, carbon monoxide, 

and nitrates in air or water, may be cumulative. 

Tebufenozide does not appear to be carcinogenic and 

does not appear to cause birth defects. No human 

health effects are likely from exposure to tebufenozide 

as used in gypsy moth projects. 

Tebufenozide and diflubenzuron could have a 

cumulative effect on methemoglobin but USDA 

gypsy moth management programs would not use 

these two chemicals together in the same area at the 

same time. However, tebufenozide might be applied 

to the same area in multiple years for eradication 

projects. These multiple applications of tebufenozide 

over a period of time may increase the potential 

risk of methemoglobinemia. Exposure to other 

methemoglobinemia-inducing compounds in the 

environment may contribute to a cumulative effect. 

For example, individuals exposed to combustion smoke 

or carbon monoxide (agents causing oxidative damage 

to blood) in addition to exposure to tebufenozide may 

be at increased risk of developing methemoglobinemia. 

Individuals exposed to high levels of nitrates, either 

in air or in water, demonstrate increased levels of 

methemoglobin and may be at increased risk with 

exposure to compounds such as tebufenozide.

Risk to Nontarget Organisms.

Tebufenozide may affect other Lepidoptera, especially 

spring-feeding moths and butterflies. No adverse 

effects on birds, mammals, or aquatic species are likely 

to occur from exposure to tebufenozide.

Tebufenozide generally would not be used in 

conjunction with other treatments. Multiple year 

applications of tebufenozide might occur for 
Supplemental Environmental Impact Statement

Summary - Page 11

eradication projects in the same area, but generally 

these areas are small. Tebufenozide might have a 

cumulative effect on non-target caterpillars of moths 

and butterflies, but will not affect other aquatic and 

terrestrial species. 

New Treatments That May Be Available in the 

Future Under Alternative 3.

Treatments that might become available in the future 

for managing the gypsy moth cannot be predicted. 

Given the parameters and protocol built into 

Alternative 3, any new treatment would pose no greater 

risk to human health and nontarget organisms than 

are disclosed in this SEIS for the currently approved 

treatments and for tebufenozide.

9 Mitigating Adverse Effects.

Given the variety of places and circumstances where 

gypsy moth projects could be implemented, it will be 

necessary to develop and implement specific mitigation 

measures for each project. Mitigation measures will be 

developed and implemented on a site-specific basis for 

each project based on local conditions and concerns. 
Summary

Summary - Page 12

Table 1. Comparative Hazard Quotients (HQs) for the effects of gypsy moths and treatments on human health and nontarget 

organisms. Wherever a 0 appears in the table, the hazard quotient value is less than 0.01.

Population Gypsy Moth 

HQ

B.t.k. HQ Dichlorvos 

HQ

Diflubenzuron 

HQ

Disparlure HQ Gypchek 

HQ

Tebufenozide 

HQ

Human 

health

1.6 to 625

Upper range 

is based 

on major 

outbreaks

0 to 0.04

Unlikely 

effects

0 to 380

Upper range 

based on child 

tampering with 

strip.

0.05 to 0.5 

– workers,

0.09 to 0.1 

– public

Upper range for 

workers based 

on ground spray 

operations.

0 

No potential 

risk can be 

identified

0 to 0.02

No risks are 

plausible

0.03 to 1.5

Highest HQ 

based on 

long-term 

consumption 

of 

contaminated 

fruit 

following two 

applications 

at the highest 

application 

rate.

Nontarget 

terrestrial 

species

0.25 to 400

Upper range 

based on 

gypsy moth 

outbreak in 

sensitive 

stands

0.36 to 9.4 

Upper 

range 

based on 

sensitive 

caterpillars 

of moths 

and 

butterflies

0 

Effects not 

likely

0.18 to 32 

Upper range 

based on 

sensitive species 

of invertebrates

0 

No potential 

hazard 

identified

0 

Effects not 

likely

0 to 4

Upper range 

based on the 

consumption 

of 

contaminated 

vegetation 

by a large 

mammal

Nontarget 

aquatic 

species

0 

No adverse 

effects

0 to 0.5

Upper level 

based on 

sensitive 

species

0 

No risks 

plausible in 

normal use. 

HQ for aquatic 

invertebrates 

could reach 

up to 8 in 

accidental 

exposures

0 to 5

Upper range 

based on acute 

exposure 

to aquatic 

invertebrates 

(Daphnia)

0 to 0.4

Upper range 

based on acute 

exposures 

to sensitive 

aquatic 

invertebrates 

(Daphnia)

0 

No adverse 

effects

0 to 0.4

Upper range 

based on 

longer term 

toxicity in 

sensitive 

aquatic 

invertebrates
Pesticide Precautionary Statement

This publication reports research involving pesticides. It does not contain recommendations for their 

use, nor does it imply that the uses discussed here have been registered. All uses of pesticides must 

be registered by appropriate State and/or Federal agencies before they can be recommended.

CAUTION: Pesticides can be injurious to humans, domestic animals, desirable plants, and fish or 

other wildlife--if they are not handled or applied properly. Use all pesticides selectively and carefully. 

Follow recommended practices for the disposal of surplus pesticides and pesticide containers.

The use of trade or firm names in this publication is for reader information and does not imply endorsement 

by the U.S. Department of Agriculture.

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race, color, national origin, age, disability, and where applicable, sex, marital status, familial status, parental status, 

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Printed on recycled paper with soy-based ink.
Gypsy Moth Management 

in the United States: 

a cooperative approach 

Draft 

Supplemental Environmental 

Impact Statement 

Risk Assessments 

Volume III of IV 

Appendixes F-I 

United States 

Department of Agriculture 

Forest Service 

Animal and Plant Health 

Inspection Service 

Newtown Square, PA 

NA-MR-01-08 

June 2008 
Gypsy Moth Management 

in the United States:

a cooperative approach

United States 

Department of Agriculture

Forest Service

Animal and Plant Health

Inspection Service

Newtown Square, PA

NA-MR-01-08

June 2008

Draft

Supplemental Environmental 

Impact Statement

Volume II of IV

Chapters 1-8 and Appendixes A-E

Contents 

Volume II

Chapter 1. Purpose of and Need for Action

Chapter 2. Alternatives Including the Preferred Alternative

Chapter 3. Affected Environment

Chapter 4. Environmental Consequences

Chapter 5. Preparers and Contributors

Chapter 6. Mailing List

Chapter 7. Glossary

Chapter 8. References

Appendix A. Gypsy Moth Treatments and Application Technology

Appendix B. Gypsy Moth Management Program

Appendix C. Public Involvement and Issues

Appendix D. Plant List

Appendix E. Biology, History, and Control Efforts for the Gypsy Moth
Gypsy Moth Management in the United States: a cooperative approach

Draft Supplemental Environmental Impact Statement 

The complete Draft Supplemental Environmental Impact Statement, Gypsy 

Moth Management in the United States: a cooperative approach, consists of 

four volumes:

Volume I Summary

Volume II Chapters 1-8 and Appendixes A, B, C, D, E

Volume III Appendixes F, G, H, I

Volume IV Appendixes J, K, L, M

Lead Agency: Forest Service, U.S. Department of Agriculture 

 (USDA)

Responsible Official: James R. Hubbard, Deputy Chief for State and 

 Private Forestry

 Sidney R. Yates Federal Building

 201 14th Street, SW

 Washington, DC 20250

For More Information: Bill Oldland Team Leader 

180 Canfield Street

 Morgantown, WV 26505

Joint Lead Agency: Animal and Plant Health Inspection Service 

 (APHIS), USDA

Responsible Official: Richard L. Dunkle, Deputy Administrator for Plant 

 Protection and Quarantine

 4700 River Road

 Riverdale, MD 20737-1236

For More Information: Weyman P. Fussell, Ph.D., Gypsy Moth Program 

 Coordinator

 Emergency and Domestic Programs, APHIS

 4700 River Road, Unit 134

 Riverdale, MD 20737-1236
Abstract: The USDA Forest Service and Animal and Plant Health Inspection Service are proposing an addition 

to the gypsy moth management program that was described in the 1995 Environmental Impact Statement--Gypsy 

Moth Management in the United States: a cooperative approach--and chosen in the 1996 Record of Decision. The 

agencies are proposing these new treatment options: adding the insecticide tebufenozide, or adding the insecticide 

tebufenozide and other new treatment(s) that may become available in the future to manage gypsy moths, provided 

that the other treatment(s) poses no greater risk to human health and nontarget organisms than are disclosed in this 

Draft SEIS for the currently approved treatments and tebufenozide.

Commenting on this Draft Supplemental Environmental Impact Statement: Reviewers should provide the 

Forest Service with their comments during the review period of this draft supplemental environmental impact 

statement. Timely comments will enable the Forest Service to analyze and respond to all of the comments at one 

time and to use information acquired in the preparation of the final supplemental environmental impact statement, 

thus avoiding undue delay in the decision making process. Furthermore, the more specific and substantive 

the comments, the better for reviewers and the agencies alike. Reviewers have an obligation to structure their 

participation in the National Environmental Policy Act process so that it is meaningful and alerts the agency to 

the reviewer’s position and contentions (Vermont Yankee Nuclear Power Corp. v. NRDC, 435 U.S. 519, 553, 

1978). Environmental objections that could have been raised at the draft stage may therefore be forfeited, if not 

raised until after completion of the final environmental impact statement (Department of Transportation v. Public 

Citizen, 541 U.S. 752, 764 (2004). Comments on this draft supplemental environmental impact statement should 

be specific and should address the adequacy of the statement and the merits of the alternatives discussed (40 CFR 

1503.3).

Web Site for Draft SEIS: The Draft SEIS is available for viewing at www.na.fs.fed.us/wv/eis

Send Comments to: Bill Oldland Team Leader 

 USDA Forest Service

180 Canfield Street

 Morgantown, WV 26505

 304-285-1585

Date Comments Must Be Received: See cover letter for the date that comments are due in Morgantown, WV.
Photo Credits

Figure 1-1. (UGA1398104) USDA Forest Service, Northeastern Area Archives, www.forestryimages.org

Figure 1-2. (UGA1929085) USDA Forest Service Archives, www.forestryimages.org

Figure 1-3. (UGA0488025) John H. Ghent, USDA Forest Service, www.forestryimages.org

Figure 1-4. Left (UGA1241014) and Right (UGA1241013) John H. Ghent, USDA Forest Service, 

www.forestryimages.org

Figure 1-5 (UGA3948096) William M. Ciesla, Forest Health Management International, www.forestryimages.org

Figure 2-1. (UGA1275077) USDA Forest Service Archives, www.forestryimages.org

Figure 3-1. (UGA1275033) USDA Forest Service Archives; www.forestryimages.org

Figure 3-3. USDA Agricultural Research Service, www.ars.usda.gov/is/kids/suburb/story2/microscope.htm

Figure 4-1. (UGA1275042) USDA Forest Service Archives, www.forestryimages.org

Figure 5-1. (UGA1275050) USDA Forest Service Archives, www.forestryimages.org

Figure 6-1. (UGA1275044) USDA Forest Service Archives, www.forestryimages.org

Figure 7-1. (UGA1275010) USDA Forest Service Archives, www.forestryimages.org

Figure 8-1. (UGA1275053) USDA Forest Service Archives, www.forestryimages.org

Figure A-1. (UGA1275013) USDA Forest Service Archives, www.forestryimages.org

Figure A-2. Derek Handley 

Figure A-3. (UGA1301021) Joseph O’Brien, USDA Forest Service, www.forestryimages.org

Figure A-4. (UGA2652048) USDA Animal and Plant Health Inspection Service, Plant Protection and Quarantine 

Archives, www.forestryimages.org

Figure A-5. (UGA2652042) USDA Animal and Plant Health Inspection Service, Plant Protection and Quarantine 

Archives, www.forestryimages.org

Figure A-6. (UGA1335028) John H. Ghent, USDA Forest Service, www.forestryimages.org

Figure A-7. (UGA2253091) Bill Antrobius, USDA Forest Service, www.forestryimages.org

Figure A-8. (UGA5022085) Pennsylvania Department of Conservation and Natural Resources Forestry Archives, 

www.insectimages.org

Figure B-1. (UGA1275058) USDA Forest Service Archives, www.forestryimages.org

Figure C-1. (UGA1275037) USDA Forest Service Archives, www.forestryimages.org

Figure D-1. (UGA1275020) USDA Forest Service Archives, www.forestryimages.org

Figure E-1. (UGA1275016) USDA Forest Service Archives, www.forestryimages.org

Figure E-3. (UGA1929072) USDA Forest Service Archives, www.insectimages.org 

Figure E-4. (UGA0886002) Tim Tigner, Virginia Department of Forestry, www.insectimages.org 

Figure E-5. (UGA2652066) USDA Animal and Plant Health Inspection Service, Plant Protection and Quarantine 

Archives, www.forestryimages.org

Figure E-6. (UGA2652079) USDA Animal and Plant Health Inspection Service, Plant Protection and Quarantine 

Archives, www.forestryimages.org
Chapter 1

Purpose of and

Need for Action

Figure 1-1. In 1892, workers attempted to control gypsy moth by hand picking egg 

masses.
Chapter 1 Purpose of and Need for Action

Contents

1.1 Proposed Action.......................................................................................1

1.2 Public Involvement and Issues................................................................1

1.3 Background.............................................................................................2

1.4 Purpose of and Need for Action...............................................................3

1.5 Decision Framework................................................................................4

1.6 Scope of This Document and NEPA Requirements.................................5

1.7 Consultations...........................................................................................6

Figures and Tables

Figure 1-1. In 1892, workers attempted to control gypsy moths 

by hand picking egg masses..................................................................Cover

Figure 1-2. Feeding by gypsy moth caterpillars (larvae) 

causes defoliation..........................................................................................2

Figure 1-3. European gypsy moths (male on left and female on right) 

are found in the United States.......................................................................2

Figure 1-4. This Asian gypsy moth male (left) and female (right) 

are from Mongolia. As of this writing, the Asian gypsy moth is not 

found in the United States.............................................................................3

Figure 1-5. People unknowingly spread gypsy moths by moving 

objects on which egg masses were deposited ...............................................3

Figure 1-6. In 2006, the European gypsy moth was established in 

all or part of 19 states and the District of Columbia .....................................4

Table 1-1. Acres treated in suppression projects, by treatment, 

2002-2006 .....................................................................................................5

Table 1-2. Acres treated in eradication projects, by treatment, 

2002-2006 .....................................................................................................5

Table 1-3. Acres treated in slow-the-spread projects by treatment, 

2002-2006 .....................................................................................................5
Purpose of and Need for Action

Chapter 1- Page 

The management of gypsy moth in the United States 

takes an integrated pest management approach to 

protecting the forests and trees of the United States 

from the adverse effects caused by the gypsy moth. 

This chapter gives brief background on the gypsy moth 

and the current gypsy moth management program. The 

chapter also states the proposed changes, rationale, 

and related issues. It explains the purpose of this draft 

supplemental environmental impact statement (SEIS) 

and how it is to be used.

1.1 Proposed Action.

The United States Department of Agriculture (USDA) 

is responsible for management activities related to the 

gypsy moth (Lymantria dispar Linnaeus [L.]), for the 

Federal government. Two USDA agencies, the Forest 

Service and the Animal and Plant Health Inspection 

Service (APHIS) share this responsibility. Agency 

authorities are found in these USDA Delegations 

of Authority: 7 Code of Federal Regulations (CFR) 

2.60(a)(38) by the Under Secretary for Natural 

Resources and Environment, for the Forest Service; 

and 7 CFR 2.80(a)(36) by the Under Secretary for 

Marketing and Regulatory Programs, for APHIS.

The Forest Service and APHIS are proposing an 

addition to the gypsy moth management program 

described in the 1995 Environmental Impact Statement 

(EIS) and chosen in the 1996 Record of Decision 

(USDA 1995, 1996). The agencies are proposing to 

add new treatment options: the insecticide tebufenozide 

and the option of adding other treatments that may 

become available in the future to manage gypsy 

moths, provided such treatments pose no greater risks 

to human health and nontarget organisms than are 

disclosed in this draft SEIS for currently approved 

treatments and tebufenozide.

This draft SEIS discloses the method of use, 

effectiveness, and effects of tebufenozide, and outlines 

the protocol that would be followed in order to add 

other treatments. Appendix A provides detailed 

information about the use and effectiveness of 

tebufenozide and other treatments that are effective 

for eradicating, suppressing, or slowing the spread 

of the gypsy moth as represented in this draft SEIS. 

Information about treatments and natural control agents 

that are not used in the USDA gypsy moth management 

program is also presented in Appendix A for the benefit 

of the reader. Appendix B provides an overview 

of the USDA gypsy moth management program. 

This draft SEIS also updates effects of currently 

approved treatments and of the gypsy moth, with new 

information that has become available since the 1995 

EIS, and about the slow-the-spread strategy which is 

now an operational component of the USDA gypsy 

moth management program.

1.2 Public Involvement and 

Issues.

On April 29, 2004, the Forest Service and APHIS 

published a Notice of Intent (NOI) to Prepare 

a Supplement to the Final EIS for Gypsy Moth 

Management in the United States: a Cooperative 

Approach (69 Federal Register (FR) 23492-93, April 

29, 2004). The public was invited to comment on the 

proposed supplement. Fourteen comment letters were 

received from the public on the SEIS. Other NOIs 

were published on March 13, 2006 (71 FR 12674-75) 

and on February 7, 2007 (72 FR 5675), revising the 

dates for filing the draft and final SEIS.

The interdisciplinary team preparing this draft 

SEIS, joined by public affairs specialists and forest 

pest managers throughout the Forest Service and 

APHIS (listed in Chapter 5) actively sought public 

involvement. Two issues were derived from the scoping 

effort: Issue 1—risk to human health, and Issue 2—risk 

to nontarget organisms. These issues are described in 

Chapters 3 and 4 of this SEIS. See Appendix C for 

details of scoping efforts.
Chapter 1

Chapter 1- Page 

1.3 Background.

The gypsy moth is a significant nonnative forest pest in 

the United States. The gypsy moth caterpillar—one of 

four distinct developmental stages (Figure 1-2)—alters

ecosystems and disrupts people’s lives as it feeds on 

the foliage of trees, shrubs, and other plants. Excessive 

feeding causes defoliation, which weakens trees 

(increasing their vulnerability to other insects and 

diseases that may kill them), alters wildlife habitat, 

changes water quality, reduces property and aesthetic 

values of public and private woodlands, and reduces 

the recreation value of forested areas. When present 

in large numbers, gypsy moth caterpillars can pose 

a nuisance, as well as a hazard to health and safety. 

Effects due to the gypsy moth are described in 

Chapter 4.

At least 587 million acres (238 million hectares) of 

trees susceptible to gypsy moth feeding (Appendix 

D), are at risk in the United States (Powell and others 

1993). Also at risk are countless urban and rural 

forested areas throughout the country where susceptible 

plants grow naturally or are planted.

Although both European and Asian strains exist, only 

the European strain is currently present in the United 

States (Figure 1-3). The European gypsy moth was 

brought to the United States and accidentally released 

in eastern Massachusetts around 1869. Since then, 

it has continued to spread into uninfested areas. The 

Asian strain occasionally has been found in this 

country, but it has been eliminated whenever it has 

been found (Figure 1-4). Unlike European female 

gypsy moths, which cannot fly, the Asian moth poses 

a greater risk of spread because females can fly and 

deposit egg masses miles from where they fed as 

caterpillars (Figure 1-5).

Despite many early attempts to halt its spread, by 2006 

the European gypsy moth became established in the 

District of Columbia and in all or parts of the following 

States: Connecticut, Delaware, Illinois, Indiana, 

Figure 1-2. Feeding by gypsy moth caterpillars (larvae) 

causes defoliation.

Figure 1-3. European gypsy moths (male on left, female on 

right) are found in the United States.

Maine, Maryland, Massachusetts, Michigan, New 

Hampshire, New Jersey, New York, North Carolina, 

Ohio, Pennsylvania, Rhode Island, Vermont, Virginia, 

West Virginia, and Wisconsin (Figure 1-6). Spread 

continues into uninfested areas because of natural and 

artificial movement.

The gypsy moth continues to be a problem as it 

spreads. Historical documentation over the last 100 

years reveals gypsy moth outbreaks cause widespread 

defoliation, tree mortality, environmental and public 

health risks, and public outcry to control the outbreaks 

(Williams and Liebhold, 1995a). For more information 

about the biology, history, and control efforts for the 

gypsy moth, see Appendix E.
Purpose of and Need for Action

Chapter 1- Page 

Figure 1-4. This Asian gypsy moth male (left) and female (right) are from Mongolia. As of this writing, the Asian gypsy moth 

is not found in the United States.

Figure 1-5. People unknowingly spread gypsy moths by 

moving objects on which egg masses were deposited.

1.4 Purpose of and Need for 

Action.

In this draft SEIS the Forest Service and APHIS 

propose to add additional treatments for use in the 

gypsy moth management program. The proposed 

treatments are new and were not available when the 

1995 EIS was written. Additional treatments would 

provide gypsy moth managers with more flexibility 

in conducting suppression, eradication, and slow-thespread projects. Making new treatments available is 

also expected to improve the gypsy moth management 

program, because each new treatment developed over 

the last 30 years has proven safer to human health and 

the environment, more cost efficient, easier to apply, 

and often more effective than older treatments.

This draft SEIS also presents new information about 

currently used treatments. It...

• Introduces hazard quotients for nontarget organisms

• Reinforces that the gypsy moth poses a significant 

risk hazard to both human health and forest condition

• Confirms that spring feeding nontarget caterpillars 

are more at risk from B.t.k. applications than are 

caterpillars that come out later in the year

• Determines that disparlure formulations used for 

mating disruption are of low toxicity to daphnids

• Makes available additional epidemiological studies 

for human health effects associated with B.t.k.

• Provides data showing that slow the spread is very 

effective in slowing the natural and artificial spread 

of the gypsy moth
Chapter 1

Chapter 1- Page 

Figure 1-6. In 2006, the European gypsy moth was established in all or part of 19 states and the District of Columbia (shaded 

in dark grey).

1.5 Decision Framework.

The 1995 EIS analyzed six alternatives for managing 

gypsy moth infestations (USDA 1995). With the 

1996 Record of Decision (USDA 1996), the agencies 

selected an integrated pest management (IPM) 

approach comprised of suppression, eradication, and 

slow-the-spread strategies to manage the gypsy moth 

in the United States. The adopted alternative also 

provides delivery of technical advice and support to 

State, Tribal, and Federal cooperators by the Forest 

Service and APHIS. The USDA has carried out its 

gypsy moth responsibilities under that Record of 

Decision since 1996.

The 1996 decision provides for the use of several 

insecticides and other treatments in suppression 

(Table 1-1), eradication (Table 1-2), and slowthe-spread projects (Table 1-3). These include 

Bacillus thuringiensis var. kurstaki (B.t.k.), the 

insect growth regulator diflubenzuron, the gypsy 

moth nucleopolyhedrosis virus product Gypchek, 

a pheromone attractant disparlure used in mating 

disruption and mass trapping, the killing agent 

dichlorvos used in large-capacity pheromone traps, 

and the sterile insect technique. Human health and 

ecological risk assessments (HHERA) were prepared 

for each of these insecticides and for the proposed 

insecticide tebufenozide, and can be found in 

Appendixes F-K of this SEIS.
Purpose of and Need for Action

Chapter 1- Page 

Table 1-1. Acres treated in suppression projects, by treatment, 2002-2006.

Year B.t.k. Diflubenzuron Gypchek Total

2002 149,772 131,601 4,794 286,167

2003 67,895 25,124 10,015 103,034

2004 73,493 0 6,078 79,571

2005 7,292 0 0 7,292

2006 145,053 18,000 602 163,655

Total 443,505 174,725 21,489 639,719

Table 1-2. Acres treated in eradication projects, by 

treatment, 2002-2006.

Year B.t.k. Gypchek Mating

Disruption

Total

2002 9,961 0 650 10,611

2003 16,540 0 0 16,540

2004 10,855 0 250 11,105

2005 36,778 0 0 36,778

2006 19,960 0 0 19,960

Total 94,094 0 900 94,994

Table 1-3. Acres treated in slow-the-spread projects, by 

treatment, 2002-2006.

Year B.t.k. Gypchek

DiflubenzuronMating 

Disruption

Total

2002 28,705 0 3,938 542,600 575,243

2003 70,470 6,819 0 647,618 720,907

2004 131,282 8,230 0 588,256 727,728

2005 108,611 17,075 790 287,890 414,366

2006 95,860 7,003 12,292 426,138 541,293

Total 434,928 39,127 17,020 2,488,502 2,979,577

Like the 1996 Record of Decision, the decision to 

be made will be programmatic. No site-specific 

suppression, eradication, or slow-the-spread projects 

will be implemented as a direct result of the decision 

that will follow this SEIS. The decision to implement 

any treatment project will be made after site-specific 

environmental analyses are conducted and documented 

in accordance with agency NEPA implementing 

procedures. Analyses will address unique local issues, 

beyond the scope of this document, for site-specific 

management projects for the gypsy moth. Site-specific 

environmental analyses are more detailed and precise 

as to geographical locations, individual treatments to be 

used, and timing of treatments.

The decision on this draft SEIS will serve as the 

primary guide for management of the gypsy moth on 

Forest Service lands; treatments and strategies allowed 

by the 1996 decision will continue to be available for 

use. The USDA is not reconsidering the suppression, 

eradication, and slow-the-spread strategies, or the 

treatments made available by the 1996 Record of 

Decision. The decision whether to plan and implement 

a gypsy moth project on National Forest System lands 

rests with the responsible official in that particular 

forest.

1.6 Scope of This Document 

and NEPA Requirements.

This SEIS concerns only the USDA gypsy moth 

management program carried out by the Forest 

Service or APHIS, directly or in conjunction with 

others (States, other Federal agencies, and Tribal 

governments). Actions of other Federal or local 

agencies or private citizens to manage the gypsy moth 

on their own, are not affected or in any way constrained 

by the USDA program. Such actions are affected or 

constrained only by applicable Federal and State laws, 

local ordinances, insecticide label instructions, and any 

self-imposed constraints.
Chapter 1

Chapter 1- Page 

The information and analysis contained in this SEIS 

can be incorporated by reference, into environmental 

documents prepared for proposed gypsy moth 

management projects, in accordance with the National 

Environmental Policy Act (NEPA) (42 United States 

Code (U.S.C.) 4332) and agency NEPA procedures. 

Future environmental documentation for specific 

projects would tier to the final SEIS and to the 1995 

EIS (40 CFR 1508.28). Proposed treatment projects 

will be evaluated on an individual basis to determine 

if they are biologically sound, environmentally 

acceptable, and economically efficient.

Some gypsy moth related activities, such as treatment 

of regulated articles infested with gypsy moths, the 

boarding and inspection of ships entering U.S. seaports, 

and research and methods-development activities, 

are outside the scope of this document and were not 

examined. More information about these activities can 

be found in Appendix B.

1.7 Consultations.

As they had done on the 1995 EIS, the Forest Service 

and APHIS will informally consult on the proposed 

action (Alternative 3) under the Endangered Species 

Act. In addition, the Forest Service and APHIS will 

ensure that site-specific consultations will be done as 

necessary at the project level under the Endangered 

Species Act, the National Historic Preservation Act, 

and any other laws, regulations, executive orders, and 

agency policies that apply to site-specific projects.
Abstract: The USDA Forest Service and Animal and Plant Health Inspection Service are proposing an addition 

to the gypsy moth management program that was described in the 1995 Environmental Impact Statement--Gypsy 

Moth Management in the United States: a cooperative approach--and chosen in the 1996 Record of Decision. The 

agencies are proposing these new treatment options: adding the insecticide tebufenozide, or adding the insecticide 

tebufenozide and other new treatment(s) that may become available in the future to manage gypsy moths, provided 

that the other treatment(s) poses no greater risk to human health and nontarget organisms than are disclosed in this 

Draft SEIS for the currently approved treatments and tebufenozide.

Commenting on this Draft Supplemental Environmental Impact Statement: Reviewers should provide the 

Forest Service with their comments during the review period of this draft supplemental environmental impact 

statement. Timely comments will enable the Forest Service to analyze and respond to all of the comments at one 

time and to use information acquired in the preparation of the final supplemental environmental impact statement, 

thus avoiding undue delay in the decision making process. Furthermore, the more specific and substantive 

the comments, the better for reviewers and the agencies alike. Reviewers have an obligation to structure their 

participation in the National Environmental Policy Act process so that it is meaningful and alerts the agency to 

the reviewer’s position and contentions (Vermont Yankee Nuclear Power Corp. v. NRDC, 435 U.S. 519, 553, 

1978). Environmental objections that could have been raised at the draft stage may therefore be forfeited, if not 

raised until after completion of the final environmental impact statement (Department of Transportation v. Public 

Citizen, 541 U.S. 752, 764 (2004). Comments on this draft supplemental environmental impact statement should 

be specific and should address the adequacy of the statement and the merits of the alternatives discussed (40 CFR 

1503.3).

Web Site for Draft SEIS: The Draft SEIS is available for viewing at www.na.fs.fed.us/wv/eis

Send Comments to: Bill Oldland Team Leader 

 USDA Forest Service

180 Canfield Street

 Morgantown, WV 26505

 304-285-1585

Date Comments Must Be Received: See cover letter for the date that comments are due in Morgantown, WV.
Photo Credits

Figure 1-1. (UGA1398104) USDA Forest Service, Northeastern Area Archives, www.forestryimages.org

Figure 1-2. (UGA1929085) USDA Forest Service Archives, www.forestryimages.org

Figure 1-3. (UGA0488025) John H. Ghent, USDA Forest Service, www.forestryimages.org

Figure 1-4. Left (UGA1241014) and Right (UGA1241013) John H. Ghent, USDA Forest Service, 

www.forestryimages.org

Figure 1-5 (UGA3948096) William M. Ciesla, Forest Health Management International, www.forestryimages.org

Figure 2-1. (UGA1275077) USDA Forest Service Archives, www.forestryimages.org

Figure 3-1. (UGA1275033) USDA Forest Service Archives; www.forestryimages.org

Figure 3-3. USDA Agricultural Research Service, www.ars.usda.gov/is/kids/suburb/story2/microscope.htm

Figure 4-1. (UGA1275042) USDA Forest Service Archives, www.forestryimages.org

Figure 5-1. (UGA1275050) USDA Forest Service Archives, www.forestryimages.org

Figure 6-1. (UGA1275044) USDA Forest Service Archives, www.forestryimages.org

Figure 7-1. (UGA1275010) USDA Forest Service Archives, www.forestryimages.org

Figure 8-1. (UGA1275053) USDA Forest Service Archives, www.forestryimages.org

Figure A-1. (UGA1275013) USDA Forest Service Archives, www.forestryimages.org

Figure A-2. Derek Handley 

Figure A-3. (UGA1301021) Joseph O’Brien, USDA Forest Service, www.forestryimages.org

Figure A-4. (UGA2652048) USDA Animal and Plant Health Inspection Service, Plant Protection and Quarantine 

Archives, www.forestryimages.org

Figure A-5. (UGA2652042) USDA Animal and Plant Health Inspection Service, Plant Protection and Quarantine 

Archives, www.forestryimages.org

Figure A-6. (UGA1335028) John H. Ghent, USDA Forest Service, www.forestryimages.org

Figure A-7. (UGA2253091) Bill Antrobius, USDA Forest Service, www.forestryimages.org

Figure A-8. (UGA5022085) Pennsylvania Department of Conservation and Natural Resources Forestry Archives, 

www.insectimages.org

Figure B-1. (UGA1275058) USDA Forest Service Archives, www.forestryimages.org

Figure C-1. (UGA1275037) USDA Forest Service Archives, www.forestryimages.org

Figure D-1. (UGA1275020) USDA Forest Service Archives, www.forestryimages.org

Figure E-1. (UGA1275016) USDA Forest Service Archives, www.forestryimages.org

Figure E-3. (UGA1929072) USDA Forest Service Archives, www.insectimages.org 

Figure E-4. (UGA0886002) Tim Tigner, Virginia Department of Forestry, www.insectimages.org 

Figure E-5. (UGA2652066) USDA Animal and Plant Health Inspection Service, Plant Protection and Quarantine 

Archives, www.forestryimages.org

Figure E-6. (UGA2652079) USDA Animal and Plant Health Inspection Service, Plant Protection and Quarantine 

Archives, www.forestryimages.org
Figure 2-1. Early spray operations for gypsy moths used horse-drawn equipment. 

Chapter 2

Alternatives Including 

the Preferred Alternative
Chapter 2 Alternatives Including the 

Preferred Alternative

Contents

2.1 Background............................................................................................ 1

2.2 Alternative Chosen From the 1995 Gypsy Moth EIS ............................ 1

2.3 Alternatives in This SEIS....................................................................... 1

Alternative 1—No Action................................................................ 2

Alternative 2—Add Tebufenozide................................................... 2

Alternative 3—Add Tebufenozide, and Add Other 

New Treatments Through the Application of the Protocol 

(Preferred Alternative)..................................................................... 2

2.4 Evaluation and Comparison of Alternatives.......................................... 3

2.5 Mitigation Measures .............................................................................. 4

Human Health.................................................................................. 4

Nontarget Organisms....................................................................... 4

Mitigation Efficacy .......................................................................... 5

Figures and Tables

Figure 2-1. Early spray operations for gypsy moths used 

horse-drawn equipment..........................................................................Cover

Table 2-1. Treatments that have been approved for use in gypsy moth 

projects since the 1995 gypsy moth EIS........................................................2

Table 2-2. Treatments available under each alternative in this SEIS.............3

Table 2-3. Effects of treatments approved and proposed for use, 

by alternatives and identified issues...............................................................6
Alternatives Including the Preferred Alternative

Chapter 2 - Page 

This chapter defines the three alternatives that are being 

considered. It compares the alternatives based on their 

ability to provide flexibility for managing gypsy moth 

populations and their relation to the identified issues. 

The preferred alternative is identified. This chapter 

also describes mitigation measures that can be used to 

protect human health and nontarget organisms.

2.1 Background.

The gypsy moth is destructive to vegetative resources, 

and the human health and environmental effects 

from exposure to the pest are substantial (Chapter 

4 and Appendix L). The strategies of suppression, 

eradication, and slow the spread and the currently 

approved treatments (Table 2-1) have proven successful 

in reducing damage caused by gypsy moth outbreaks 

in the generally infested area, eliminating new isolated 

infestations of the gypsy moth introduced outside 

the generally infested area, and slowing the shortrange natural and artificial spread of this insect. For 

a description of the strategies, see Section B-5 in 

Appendix B.

These strategies form the basis for the alternatives that 

were considered in the 1995 Environmental Impact 

statement (EIS) and for the alternatives in this draft 

supplemental EIS (SEIS).

2.2 Alternative Chosen From 

the 1995 Gypsy Moth EIS.

A program consisting of the strategies of suppression, 

eradication, and slow the spread--the preferred 

alternative in the 1995 EIS--was chosen in the 1996 

Record of Decision. The following insecticide and 

noninsecticide treatments were approved for use in the 

strategies:

• Bacillus thuringiensis var. kurstaki (B.t.k.) (a 

microbial insecticide)

• Diflubenzuron (an insect growth regulator)

• Gypchek (gypsy moth nucleopolyhedrosis virus 

product)

• Mass trapping (using traps baited with the 

gypsy moth attracting pheromone disparlure and 

sometimes containing the killing agent dichlorvos)

• Mating disruption (aerially dispensed medium 

impregnated with the gypsy moth attractant 

disparlure)

• Sterile insect technique (release of sterile or partly 

sterile gypsy moth pupae or eggs)

Table 2-1 shows which treatments may be used in each 

strategy.

This alternative was adopted because it fully met the 

USDA goal of reducing the adverse effects of the gypsy 

moth on the Nation’s forests and trees. The alternative 

addresses the major issues associated with the gypsy 

moth and treatments while incorporating flexible 

options for managing ecosystems affected by the gypsy 

moth. The issues influencing the discussion in the 1995 

Gypsy Moth EIS focused on the effects of the gypsy 

moth and gypsy moth treatments on human health, 

nontarget organisms, and forest conditions.

2.3 Alternatives in This SEIS.

Like the 1996 Record of Decision, the decision to be 

made as a result of this SEIS will be programmatic. 

No site-specific suppression, eradication, or slow-thespread projects will be implemented as a direct result of 

the decision on this SEIS. The decision to implement 

any treatment project will be made after site-specific 

environmental analyses are conducted and documented 

in accordance with agency NEPA implementing 

procedures.

The following three alternatives were identified during 

scoping for the draft SEIS:

Alternative 1—No action

Alternative 2—Add tebufenozide

Alternative 3—Add tebufenozide, and add 

other new treatments through the application of the 

protocol (preferred alternative).
Chapter 2

Chapter 2 - Page 

Alternative 1—No Action.

Alternative 1 is the same as the alternative selected in 

the 1996 Record of Decision. It is the current gypsy 

moth management program of suppression, eradication, 

and slow the spread, using currently approved 

treatments. Alternative 1 would make no change to 

the 1996 Record of Decision, and it would add no 

treatment options to those approved by that decision.

Alternative 2—Add Tebufenozide.

Alternative 2 would add the insecticide tebufenozide 

to currently approved treatments. Information on the 

use and effectiveness of tebufenozide is provided in 

Appendix A. The human health and ecological risk 

assessments for tebufenozide are in Appendix J.

Alternative 3—Add Tebufenozide, and 

Add Other New Treatments Through 

the Application of the Protocol 

(Preferred Alternative).

Alternative 3 would add the insecticide tebufenozide 

and add other treatment(s) that may become available 

in the future for managing gypsy moths, to currently 

approved treatments. A new treatment would be 

available for use upon the agencies’ finding that the 

treatment poses no greater risks to human health 

and nontarget organisms than are disclosed in this 

draft SEIS for the currently approved treatments and 

tebufenozide.

The protocol for making the necessary finding that a 

treatment is authorized by this Alternative is as follows:

1. Conduct a human health and ecological risk 

assessment (HHERA). In this risk assessment 

review all scientific studies available for 

toxicological and environmental fate information 

relevant to effects on human health and nontarget 

organisms. Use this information to estimate risk 

to human health and nontarget organisms. Include 

these four elements in the HHERA: (a) hazard 

evaluation, (b) exposure assessment, (c) doseresponse assessment, and (d) risk characterization. 

The HHERA will do the following:

• Identify potential use patterns, including 

formulation, application methods, application 

rate, and anticipated frequency of application.

• Review chemical hazards relevant to the human 

health risk assessment, including systemic and 

reproductive effects, skin and eye irritation, 

dermal absorption, allergic hypersensitivity, 

carcinogenicity, neurotoxicity, immunotoxicity, 

and endocrine disruption.

• Estimate exposure of workers applying the 

chemical.

• Estimate exposure of members of the public.

• Characterize environmental fate and transport, 

including drift, leaching to groundwater, and 

runoff to surface streams and ponds.

• Review available ecotoxicity data including 

Table 2-1. Treatments that have been approved for use in gypsy moth projects since the 1995 gypsy moth EIS.

Strategy B.t.k. Diflubenzuron Gypchek

Mass Trapping 

(Dichlorvos plus 

disparlure) 

Mating 

Disruption 

(Disparlure)

Sterile Insect 

Technique

Suppression   

Eradication      

Slow the Spread      
Alternatives Including the Preferred Alternative

Chapter 2 - Page 

hazards to mammals, birds, reptiles, amphibians, 

fish, and aquatic invertebrates.

• Estimate exposure of terrestrial and aquatic 

wildlife species.

• Characterize risk to human health and wildlife.

2. Conduct a risk comparison of the human health 

and ecological risks of a new treatment with 

the risks identified for the currently authorized 

treatments and tebufenozide. This risk comparison 

will evaluate quantitative expressions of risk (such 

as hazard quotients) and qualitative expressions of 

risk that put the overall risk characterizations into 

perspective. Qualitative factors include scope, 

severity, and intensity of potential effects, as well 

as temporal relationships such as reversibility and 

recovery.

3. If the risks posed by a new treatment fall within 

or below the range of risks posed by the currently 

approved treatments and tebufenozide, publish 

a notice in the Federal Register of the agencies’ 

preliminary findings that the treatment meets the 

requirements of Alternative 3. The notice must 

provide a 30-day review and comment period and 

must advise the public that the HHERA and the 

risk comparison are available upon request.

4. If consideration of public comment leads to the 

conclusion that the preliminary finding is correct, 

publish a notice in the Federal Register that the 

treatment meets the requirements of Alternative 

3 and, therefore, is authorized by that Alternative 

for use in the USDA gypsy moth management 

program. The Forest Service and APHIS will make 

available to anyone, upon request, a copy of the 

comments received and the agencies’ responses.

Like the 1996 Record of Decision, the decision 

to be made as a result of this draft SEIS will be 

programmatic. Decisions to use specific treatments 

in projects, including new treatments authorized 

under the protocol in Alternative 3, will be made after 

site-specific environmental analyses are conducted 

and documented in accordance with agency NEPA 

implementing procedures.

2.4 Evaluation and Comparison 

of Alternatives.

Different treatments could be used under the different 

alternatives, as shown in Table 2-2. The more 

treatments that are available, the more flexibility the 

program manager has in choosing the right treatment 

for a given set of specific conditions and the greater 

likelihood of meeting the project objectives. The 

Alternatives provide increasing flexibility from 

Alternative 1 to Alternative 3. With the addition of 

tebufenozide and other treatments that may become 

available, Alternative 3—the preferred alternative—

would provide the program manager the greatest 

flexibility. This flexibility for Alternative 3 includes 

reducing the cost, streamlining the process, and 

greater efficiency in adding new treatments for gypsy 

moth management. Cost, availability, efficacy, and 

site-specific environmental effects are examples of 

considerations regarding which treatment to use for a 

specific project.

The effects of the different treatments are summarized 

by the issues in Table 2-3.

Table 2-2. Treatments available under each alternative in 

this SEIS

Alternative

Currently 

approved 

treatments*

Tebufenozide

Other 

treatments that 

may become 

available

1 

2  

3   

*Currently approved treatments:

Bacillus thuringiensis var. kurstaki

Diflubenzuron

Mass trapping (dichlorvos and disparlure)

Mating disruption (disparlure)

Gypchek

Sterile insect technique
Chapter 2

Chapter 2 - Page 

2.5 Mitigation Measures.

Given the variety of places and circumstances where 

gypsy moth projects could be implemented, it will be 

necessary to develop and implement specific mitigation 

measures for each project. Mitigation measures will be 

developed and implemented on a site-specific basis for 

each project based on local conditions and concerns.

The site-specific mitigation measures developed and 

employed in gypsy moth projects since the 1996 

Record of Decision have shown to be effective in 

addressing human health and safety concerns, adverse 

effects on nontarget organisms and potential impacts on 

economic resources such as organic farms. At the same 

time the objectives of gypsy moth projects have been 

met. Site-specific mitigation measures will continue 

to be developed and implemented. The following are 

examples of project level mitigation measures that have 

been employed in the past and could be implemented 

for future projects.

Human Health.

● Ensure workers handling insecticides wear 

appropriate personal protective gear and protective 

clothing.

● Prepare a project safety plan, disseminate it to project 

workers, and conduct safety briefings.

● Ensure workers handling dichlorvos insecticide strips 

wear gloves and assemble the gypsy moth traps 

outdoors, preferably at the trap site, and transport 

traps and trapping supplies in an air-tight plastic bag.

● Use gypsy moth traps that do not contain dichlorvos, 

when possible, in residential areas.

● Encourage public involvement to identify human 

health issues, including concerns of people sensitive 

to insecticides. Public notification is an important 

part of the program, enabling those living in 

treatment areas to plan their activities and avoid 

exposure.

● Consider social and cultural factors. Take steps 

to ensure all groups of the affected population 

understand the project and are invited to provide 

input during project development, such as the 

distribution of information pamphlets in languages 

relevant to the affected population. 

● Give notice to hospitals, schools, public health 

facilities and local law-enforcement agencies of 

treatments, the types of insecticides used and risks to 

humans.

● Give notice of pesticide treatment projects to 

organizations, groups and agencies that consist of, or 

work with, people who are chemically sensitive.

● Give notice to the public when treatments are 

scheduled, including the insecticides planned for use, 

potential health effects and other characteristics of 

the project, such as the use of low-flying aircraft.

● Give notice of treatments to people living in the 

project area sufficiently in advance to allow them to 

plan their activities and avoid exposure.

● Establish safety and protection measures for workers 

known to be sensitive to insecticides.

● Establish buffer zones as needed (for example, 

tebufenozide would not be sprayed over water or 

areas where surface water is present, and buffers 

will be maintained around these areas). Certain 

actions like using the latest advances in application 

technology as outlined in section A.5 of Appendix A 

would minimize the risk of insecticides drifting into 

bodies of water or sites such as organic farms.

● Mix, load, and unload insecticides in areas where an 

accidental spill will not enter and contaminate bodies 

of water.

Nontarget Organisms.

● Use public involvement to identify any site-specific 

issues with potential for effects on nontarget 

organisms (including threatened and endangered 

species), and to design appropriate means to mitigate 

these effects.

● Select treatments taking into consideration maximum 

project efficiency, potential effects on nontarget 

organisms (including threatened and endangered 

species), and the potential for these organisms to 

recolonize areas if they are displaced or die after 

treatment.
Alternatives Including the Preferred Alternative

Chapter 2 - Page 

● Establish buffer zones where necessary to minimize 

or eliminate insecticide drift to areas of special 

concern, such as wilderness areas or sensitive 

species habitats (for example, tebufenozide would 

not be sprayed over water or areas where surface 

water is present, and buffers will be maintained 

around these areas).

● Review maps and conduct ground inspections or 

other actions as part of the site-specific analysis to 

identify small brooks, wetlands, estuarine waters, 

areas where threatened and endangered species are 

found, bat caves and other roosts or other sensitive 

areas, and to determine actions needed to minimize 

adverse outcomes.

● Mix, load, and unload insecticides in areas where 

an accidental spill will not enter and contaminate 

bodies of water.

Mitigation Efficacy.

The mitigation measures developed and employed in 

site-specific gypsy moth projects have proven to be 

effective in protecting human health and non-target 

organisms. At the same time, the objectives of gypsy 

moth suppression, eradication, and slow-the-spread 

projects have been successfully met since 1996.
Chapter 2

Chapter 2 - Page 

Treatments and alternatives Issue 1. 

Risk to Human Health

Issue 2. 

Risk to Nontarget Organisms

B.t.k. 

Alternatives 1, 2, 3

See Appendix F for Human 

Health and Ecological Risk 

Assessment (HHERA)

May irritate the eyes, skin, and 

respiratory tract.

Reduces human health effects 

caused by gypsy moth hairs.

May reduce populations of some 

spring feeding caterpillars.

Reduces effects of gypsy moths on 

nontarget organisms.

Diflubenzuron 

Alternatives 1, 2, 3

See Appendix I for HHERA

May slightly increase 

methemoglobin in sensitive 

individuals.

Reduces human health effects 

caused by gypsy moth hairs.

Potentially affects arthropod 

species that produce chitin (hard 

exoskeleton) and are immature at 

time of treatment.

Can temporarily increase algae 

due to reduction of algae-feeding 

aquatic invertebrates. (This has not 

been observed in the field.)

Reduces effects of gypsy moths on 

nontarget organisms.

Gypchek 

Alternatives 1, 2, 3

See Appendix G for HHERA

Not likely to affect human health.

Reduces human health effects 

caused by gypsy moth hairs.

Has no effect on nontarget 

organisms.

Reduces effects of gypsy moths on 

nontarget organisms.

Dichlorvos plus disparlure (Mass 

Trapping)

Alternatives 1, 2, 3

See Appendixes H and K for 

HHERA

Used in intact traps, not likely to 

affect human health. Could impair 

the nervous system if someone 

disassembles a milk carton trap 

and tampers with the dichlorvosimpregnated strip, resulting in skin 

contact or ingestion.

Not likely to affect nontarget 

organisms.

(continued)

Table 2-3. Effects of treatments approved and proposed for use, by alternatives and identified issues. (Unless otherwise 

noted, the effects are based on the maximum registered usage rate allowed by the insecticide label.)
Alternatives Including the Preferred Alternative

Chapter 2 - Page 

Table 2-3 (continued).

Treatments and alternatives Issue 1. 

Risk to Human Health

Issue 2. 

Risk to Nontarget Organisms

Disparlure (Mating Disruption)

Alternatives 1, 2, 3

See Appendix H for HHERA

Not likely to affect human health. Has no effect on nontarget 

organisms.

Sterile Insect Technique

Alternatives 1, 2, 3

Has no effect on human health Has no effect on nontarget 

organisms.

Tebufenozide

Alternatives 2, 3

See Appendix J for HHERA

May slightly increase 

methemoglobin in sensitive 

individuals.

Reduces human health effects 

caused by gypsy moth hairs.

May affect some Lepidoptera 

species.

Reduces effects of gypsy moths on 

nontarget organisms.

Other treatment

Alternative 3

Has effects no more severe than 

those described in this SEIS for 

currently approved treatments and 

tebufenozide.

Reduces human health effects 

caused by gypsy moth hairs.

Has effects no more severe than 

those described in this SEIS for 

currently approved treatments and 

tebufenozide

Reduces effects of gypsy moths on 

nontarget organisms.

Chapter 3

Affected Environment

Figure 3-1. Undated historical image of workers involved in a gypsy moth 

management program.
Chapter 3 Affected Environment

Figures and Tables

Figure 3-1. Undated historical image of workers involved in a 

gypsy moth management program.........................................................Cover

Figure 3-2. Forest stands with 20 percent basal area or more of 

gypsy moth host trees are at the greatest risk of defoliation..........................1

Figure 3-3. Gypsy moth hairs can cause irritation.........................................4

Table 3-1. Top 20 tree species in the United States preferred by 

gypsy moths, ranked by total basal area (BA)...............................................2

Contents

3.1 General Affected Environment................................................................1

3.2 Affected Forest.........................................................................................1

Affected Plants..................................................................................1

Affected Areas...................................................................................2

 Uninhabited Forest...............................................................2

 Forest Recreation Areas........................................................3

 Forest Residential Areas.......................................................3

 Developed Areas...................................................................3

Forest Condition................................................................................3

Water Quality.....................................................................................3

Microclimate.....................................................................................3

Soil....................................................................................................3

3.3 Affected Human Populations...................................................................4

3.4 Affected Nontarget Organisms................................................................4

General..............................................................................................4

Threatened and Endangered Species................................................5
Affected Environment

Chapter 3 - Page 

This chapter describes the environment that is or could 

be affected by the gypsy moth and the USDA gypsy 

moth management program.

3.1 General Affected 

Environment.

Because this is a programmatic document, the 

description of the affected environment contained in 

this chapter is, by necessity, general. The potentially 

affected environment in the United States is anywhere 

vegetation susceptible to gypsy moth feeding is found. 

Given the known worldwide distribution of the gypsy 

moth, it is probably capable of surviving anywhere 

in the United States where suitable host plants are 

available (McFadden and McManus 1991).

3.2 Affected Forest.

Affected Plants.

Field and laboratory studies of numerous tree species 

enabled determination of the gypsy moth’s feeding 

preferences (Liebhold and others 1995; and see 

Appendix D for a list of susceptible plants). Forest 

trees grow either in pure stands comprised of a 

single species or in mixed stands as an aggregation 

of different species. Plant species composition is 

an important factor in determining the degree of 

susceptibility of a forest to the gypsy moth (McFadden 

and McManus 1991). Other factors include total 

density (basal area per acre) of preferred tree species 

and proportion of area covered by susceptible stands 

(Figure 3-2). Stands with basal area of preferred 

species greater than 20 percent are particularly at risk 

(Liebhold and others 1997).

Table 3-1 lists the total basal area of the 20 most 

common and important gypsy moth hosts in the 

United States. The more hardwoods, particularly 

oaks, in a forest, the more vulnerable it is to the gypsy 

moth. Higher numbers of susceptible species result 

in increased intensity, duration, and frequency of 

defoliation episodes (Davidson and others 1999).

The Forest Service classifies forested areas by 

combining forest cover types into “forest type groups” 

for inventory, mapping, and other purposes. Although 

forest cover types are based on and named after the 

Figure 3-2. Forest stands with 

20 percent basal area or more 

of gypsy moth host trees are at 

the greatest risk of defoliation. 

(Shading on the map represents 

the following basal areas of 

preferred hosts: white – less 

than 2%; light gray – 2-20%; 

medium gray – 21-39%; dark 

gray – 40-79%)
Chapter 3

Chapter 3 - Page 

tree species dominating the stand, other tree species 

may be present. These associated tree species may be 

susceptible to the gypsy moth as well.

Oak-hickory is the largest and most diverse susceptible 

forest type group, extending from the Great Plains to 

the eastern seaboard. Oak-pine types are found in the 

South. Oak-gum-cypress types are bottomland forests, 

typically found in the South and Southeast, especially 

within the Mississippi Delta and Piedmont. Aspenbirch forests are located in the North Central States. 

All of these forest types are susceptible to the gypsy 

moth.

Much of south-central and southeastern Alaska has 

climate and trees (paper birch, willow and alders) 

suitable for the gypsy moth. Aspen types are the most 

abundant hardwood in the intermountain area, while 

oak types predominate in California and red alder in the 

Pacific Northwest.

Compared with the European strain, the Asian strain 

of the gypsy moth feeds on more plants (USDA 1992). 

In addition to feeding on the same plant species as the 

European strain, the Asian strain of the gypsy moth 

will feed on larch and tamarack (Larix spp.) in Siberia, 

eastern Asia, and Japan (USDA 1992), and on both 

eastern (L. laricina) and western larch (L. occidentalis) 

in the United States.

Affected Areas.

Uninhabited Forest. 

Land use in uninhabited forest areas is dependent on 

the individual landowner’s management objectives 

(e.g., timber, wildlife, esthetics, recreation). This 

classification of forest has no or few residences and few 

if any paved roads. Uninhabited forest areas exhibit 

nearly complete forest canopy coverage, typically with 

three layers composed of subcanopy vegetation, ground 

layer vegetation, and a layer of organic debris at the 

Table 3-1. Top 20 tree species in the United States preferred 

by gypsy moths, ranked by total basal area (BA).

Common 

Name Species

Total BA 

(100,000,000 ft2

)

White oak Quercus alba 14.30

Sweetgum Liquidambar styraciflua 11.60

Quaking 

aspen Populus tremuloides 10.10

Northern 

red oak Quercus rubra 9.62

Black oak Quercus velutina 7.31

Chestnut oak Quercus prinus 6.84

Post oak Quercus stellata 5.47

Water oak Quercus nigra 4.34

Paper birch Betula papyrifera 3.81

Southern 

red oak Quercus falcata 3.75

Scarlet oak Quercus coccinea 3.31

American 

basswood Tilia americana 2.41

Western larch Larix occidentalis 2.40

Laurel oak Quercus laurifolia 1.94

Bigtooth 

aspen Populus grandidentata 1.90

Tan oak Lithocarpus densiflorus 1.64

Willow oak Quercus phellos 1.49

California 

red oak Quercus kelloggii 1.45

Eastern 

hophornbeam Ostrya virginiana 1.26

Canyon 

live oak Quercus chrysolepis 1.14
Affected Environment

Chapter 3 - Page 

soil level. The layers of vegetation serve to reduce 

the impact of raindrops and the subsequent chance of 

erosion due to overland runoff.

Forest Recreation Areas.

Recreation sites typical of rural settings include 

municipal, county, and state parks, national parks, 

monuments, forests and grasslands, public and private 

campgrounds, hiking trails, winter sports complexes, 

vacation cabins, forest lands for backpacking, and lakes 

and rivers used for hunting, fishing, and boating. Rural 

roads and scenic vistas provide attractive and tranquil 

settings, drawing many visiting tourists from populous, 

developed areas. All of these areas may be subject to 

gypsy moth outbreaks.

Forest Residential Areas.

Suppression projects are often conducted in areas 

where forests and people meet. Examples are forested 

residential areas that contain single- and multiplefamily housing, parks, cemeteries, schools, churches, 

and small businesses; and woodlots in farm areas 

that offer the potential for gypsy moth movement. 

These areas are typically occupied year-round, with 

landowners directly experiencing the impact of gypsy 

moth defoliation. Homeowners generally place a 

high value on their trees for shade, esthetics, privacy, 

investment, and wildlife habitat, and are consequently 

concerned when this resource is threatened. Several 

studies reveal that trees increase property values 5 to 

15 percent (Dwyer and others 1992). The presence of 

defoliated, dying or dead trees can decrease property 

value and marketability. The cost to remove a dead 

tree and stump is potentially hundreds of dollars.

Developed Areas.

Natural plant communities in developed areas tend to 

be fragmented and small, as native plants are frequently 

replaced with nonnative species.

Forest Condition.

Indicators of forest condition include tree mortality 

rates, tree growth rates, degree of insect damage 

(defoliation by gypsy moths), and species composition 

in the understory and canopy. Gypsy moth defoliation 

can not only cause mortality of trees, but can also affect 

the composition of forest communities.

The gypsy moth is not the only introduced pest that can 

adversely affect the Nation’s forest resources. Chestnut 

blight and Dutch elm disease in the past, and more 

recently beech bark disease, dogwood anthracnose, 

emerald ash borer, hemlock woolly adelgid, Asian 

longhorned beetle, Sirex woodwasp, butternut canker, 

and others threaten both natural and urban forests. 

As the gypsy moth and other introduced insects and 

pathogens spread, they all add stress to forest areas. 

This stress may be responsible, in part, for documented 

cases of widespread mortality where no single agent 

appears to be responsible (Weiss and Rizzo 1987).

Water Quality.

Lakes, streams, rivers and other surface waters in 

areas with plants susceptible to feeding by gypsy moth 

caterpillars may be part of the affected environment. 

Indicators of water quality include flow rate and water 

chemistry. 

Microclimate.

Microclimates created by moisture and temperature 

conditions found in forests vary by the amount of 

annual precipitation, elevation, and forest type group. 

Microclimates may potentially be affected in areas with 

trees susceptible to gypsy moth feeding. 

Soil.

Soil types capable of supporting vegetation susceptible 

to gypsy moth feeding are potentially part of the 

affected environment. Soil supports a great diversity 

of organisms, such as earthworms, arthropods, and 

microorganisms, which may live in the surface layer, 

beneath leaf litter, or throughout several soil layers.
Chapter 3

Chapter 3 - Page 

Soil structural differences support a wide range of soildependent organisms; for example, ground-dwelling 

arthropods in urban settings are less diverse than those 

commonly found in undeveloped areas (Gilbert 1989). 

Impervious surfaces in developed areas prevent air and 

water from penetrating the soil, which is often more 

disturbed and compacted than in undeveloped areas. 

These conditions contribute to a general reduction 

of plant vigor, root penetration, nitrogen fixation by 

legumes, and invertebrates to consume and recycle 

organic matter.

3.3 Affected Human 

Populations.

Many factors influence the health of people including 

these: diet, climate, airborne diseases, cultural 

traditions, emotional well-being, income, access 

to medical facilities, and contaminants in soil, air, 

and water. People living in or near areas with trees 

could be exposed to the gypsy moth and treatments. 

Particularly susceptible people include those with 

allergic reactions to gypsy moth hairs (Figure 3-3), 

respiratory ailments, chemical sensitivities, pregnant 

women, children, and the elderly. Those who work in 

the woods or with trees, mix or apply insecticides, or 

work in laboratories with gypsy moths could frequently 

be exposed to gypsy moths and treatments.

Perceptions and behaviors of individuals vary, 

depending upon their familiarity with the presence 

of gypsy moth caterpillars and the use of treatments. 

Reactions to the gypsy moth are usually strongest 

where outbreaks occur for the first time; people become 

alarmed when huge numbers of gypsy moth caterpillars 

suddenly appear. Perceptions and behaviors in 

response to the presence of gypsy moth caterpillars 

and gypsy moth treatment projects may also vary by 

location. Because urban dwellers are less likely to be 

exposed to the caterpillars and may never encounter the 

gypsy moth, they generally do not perceive the moths 

as being a problem unless the trees in their own yard 

are directly affected.

Suburban and rural area residents are more likely to be 

alarmed by large populations of gypsy moth caterpillars 

and treatment efforts. Inhabitants of rural agricultural 

areas tend to be less concerned about spraying to 

control gypsy moth populations due to their familiarity 

with spraying of agricultural crops.

3.4 Affected Nontarget 

Organisms.

General.

Virtually all wildlife in the United States that require 

trees as a part of their environment are within 

range of the gypsy moth. Mammals, birds, fish, 

and butterflies, for example, live in environments 

potentially affected by the gypsy moth or gypsy moth 

treatments. Detrimental effects of gypsy moths on 

native Lepidoptera were noted in a West Virginia study 

(Sample and others 1996).

Figure 3-3. Gypsy moth hairs can cause irritation. 
Affected Environment

Chapter 3 - Page 

Animal diversity is generally lower in developed 

areas, where native animal communities tend to be 

fragmented and small. Animals that do well in urban 

or fringe areas usually reproduce rapidly, and exhibit 

flexible behavior patterns, enabling them to exploit 

diverse food sources (Gill and Bonnet 1973). Species 

in urban areas (squirrels and birds like starlings, robins, 

and crows), which adapt to high human population 

density, are often found in greater numbers. Domestic 

animals and pets also comprise a sector of the animal 

life in areas with high concentrations of people. In 

contrast, forested areas sustain various populations, 

including birds (such as warblers, vireos, thrushes, 

flycatchers, and raptors), as well as large and small 

mammals such as bobcats and other predators.

Opossum, skunk, raccoon, and squirrel do well in both 

developed and undeveloped areas, and may be found in 

areas providing sufficient green space for cover. Larger 

mammals, such as bear, moose, and wolf, that are 

sensitive to human disturbances, require larger home 

ranges and tend to inhabit undeveloped regions.

The diversity of birds is lower in urban settings than in 

undeveloped areas (Gill and Bonnett 1973). Most bird 

species in urban areas are year-round residents or shortdistance migrants rather than neotropical migrants, 

which are more common to undeveloped areas.

Reptiles and amphibians do not fare well in developed 

areas where native vegetation, breeding sites and cover 

have been disturbed. Loss of habitat, travel barriers 

and pollution are reasons for fewer numbers of reptiles 

and amphibians in developed areas than in more natural 

areas (Campbell 1974a).

Threatened and Endangered Species.

Any species that is listed or proposed for listing as a 

threatened or endangered species and found in or near 

forested habitats could potentially be affected by the 

gypsy moth or gypsy moth treatments. Federally listed 

species of moths, butterflies, and insect-eating birds are 

of particular concern.

Chapter 4

Environmental 

Consequences

Figure 4-1. This experiment station and insectary in Malden, Massachusetts, was 

used for some of the earliest research on the gypsy moth.
Contents

4.1 Alternatives and Treatments...................................................................1

4.2 Risk Assessments and Risk Characterization ........................................1

Overview..........................................................................................1

Hazard Quotients .............................................................................2

4.3 Consequences of the Gypsy Moth .........................................................4

General Effects of the Gypsy Moth .................................................4

Forest Condition–Effects of Defoliation on Vegetation......4

Forest Condition—Tree Mortality ......................................5

Forest Condition—Seed and Mast Production ...................6

Water Quality......................................................................6

Soil Condition.....................................................................7

Microclimate.......................................................................7

Risk to Human Health (Issue 1).......................................................7

General................................................................................7

Groups at Special Risk........................................................7

Risk to Nontarget Organisms (Issue 2)............................................7

Mammals ............................................................................7

Birds....................................................................................8

Terrestrial Invertebrates......................................................9

Fish .....................................................................................9

Aquatic Invertebrates..........................................................9

Cumulative Effects of the Gypsy Moth ...........................................9

Risk to Human Health (Issue 1)..........................................9

Risk to Nontarget Organisms (Issue 2)...............................9

Chapter 4 Environmental Consequences
4.4 Consequences of Bacillus thuringiensis var. kurstaki (B.t.k.) (Alternatives 1, 2, 3)..........................................10

General Effects of B.t.k................................................................................................................................10

Risk to Human Health (Issue 1)...................................................................................................................10

 General...........................................................................................................................................10

 Groups at Special Risk....................................................................................................................10

Risk to Nontarget Organisms (Issue 2)........................................................................................................11

 Mammals.........................................................................................................................................11

 Birds................................................................................................................................................11

 Terrestrial Invertebrates..................................................................................................................11

 Fish.................................................................................................................................................12

 Aquatic Invertebrates......................................................................................................................12

Cumulative Effects of B.t.k..........................................................................................................................13

 Risk to Human Health (Issue 1)......................................................................................................13

 Risk to Nontarget Organisms (Issue 2)...........................................................................................13

4.5 Consequences of Diflubenzuron (Alternatives 1, 2, 3)......................................................................................13

General Effects of Diflubenzuron................................................................................................................13

Risk to Human Health (Issue 1)...................................................................................................................13

 General...........................................................................................................................................13

 Groups at Special Risk...................................................................................................................14

Risk to Nontarget Organisms (Issue 2)........................................................................................................14

 Mammals........................................................................................................................................14

 Birds................................................................................................................................................14

 Terrestrial Invertebrates..................................................................................................................15

 Fish.................................................................................................................................................15

 Aquatic Invertebrates......................................................................................................................16

Cumulative Effects of Diflubenzuron..........................................................................................................16

 Risk to Human Health (Issue 1).....................................................................................................16

 Risk to Nontarget Organisms (Issue 2)...........................................................................................17

4.6 Consequences of Disparlure (as Used in Mating Disruption and Mass Trapping) (Alternatives 1, 2, 3)..........17

General Effects of Disparlure.......................................................................................................................17

Risk to Human Health (Issue 1)...................................................................................................................17

 General............................................................................................................................................17

 Groups at Special Risk...................................................................................................................18

Risk to Nontarget Organisms (Issue 2)........................................................................................................18

 Mammals.........................................................................................................................................18

 Birds................................................................................................................................................18

 Terrestrial Invertebrates..................................................................................................................18

 Fish.................................................................................................................................................18

 Aquatic Invertebrates......................................................................................................................18

Cumulative Effects of Disparlure................................................................................................................18

 Risk to Human Health (Issue 1)......................................................................................................18

 Risk to Nontarget Organisms (Issue 2)...........................................................................................19
4.7 Consequences of Dichlorvos (as Used in Mass Trapping) (Alternatives 1, 2, 3)..............................................19

General Effects of Dichlorvos.....................................................................................................................19

Risk to Human Health (Issue 1)..................................................................................................................19

 General...........................................................................................................................................19

 Groups at Special Risk...................................................................................................................20

Risk to Nontarget Organisms (Issue 2).......................................................................................................20

 Mammals........................................................................................................................................20

 Birds................................................................................................................................................20

 Terrestrial Invertebrates..................................................................................................................20

 Fish.................................................................................................................................................20

 Aquatic Invertebrates.....................................................................................................................20

Cumulative Effects of Dichlorvos...............................................................................................................20

 Risk to Human Health (Issue 1)......................................................................................................20

 Risk to Nontarget Organisms (Issue 2)...........................................................................................21

4.8 Consequences of Gypchek (Alternatives 1, 2, 3)...............................................................................................21

General Effects of Gypchek.........................................................................................................................21

Risk to Human Health (Issue 1)...................................................................................................................21

 General............................................................................................................................................21

 Groups at Special Risk....................................................................................................................21

Risk to Nontarget Organisms (Issue 2)........................................................................................................21

 Mammals.........................................................................................................................................21

 Birds................................................................................................................................................21

 Terrestrial Invertebrates..................................................................................................................21

 Fish..................................................................................................................................................22

 Aquatic Invertebrates......................................................................................................................22

Cumulative Effects of Gypchek...................................................................................................................22

 Risk to Human Health (Issue 1).....................................................................................................22

 Risk to Nontarget Organisms (Issue 2)...........................................................................................22

4.9 Consequences of Tebufenozide (Alternatives 2 and 3)......................................................................................22

General Effects of Tebufenozide.................................................................................................................22

Risk to Human Health (Issue 1)..................................................................................................................22

 General...........................................................................................................................................22

 Groups at Special Risk...................................................................................................................23

Risk to Nontarget Organisms (Issue 2)........................................................................................................23

 Mammals........................................................................................................................................23

 Birds...............................................................................................................................................23

 Terrestrial Invertebrates..................................................................................................................24

 Fish.................................................................................................................................................25

 Aquatic Invertebrates.....................................................................................................................25

Cumulative Effects of Tebufenozide...........................................................................................................26

 Risk to Human Health (Issue 1).....................................................................................................26

 Risk to Nontarget Organisms (Issue 2)..........................................................................................26
4.10 Consequences of Adding a New Treatment Under Alternative 3....................................................................26

4.11 Summary of Effects Including Cumulative Effects.........................................................................................26

Risk to Human Health (Issue 1)...................................................................................................................26

 General............................................................................................................................................26

 Cumulative Effects..........................................................................................................................27

Risk to Nontarget Organisms (Issue 2)........................................................................................................27

 General............................................................................................................................................27

 Cumulative Effects..........................................................................................................................28

4.12 Operational Flexibility of Treatments..............................................................................................................28

4.13 Unavoidable Adverse Effects...........................................................................................................................28

4.14 Short-Term Uses and Long-Term Productivity................................................................................................28

4.15 Measures to Mitigate Adverse Environmental Impacts...................................................................................28

4.16 Urban Quality, Historic and Cultural Resources, and Design of the Built Environment................................29

4.17 Energy Requirements and Conservation Potential of Various Alternatives......................................................29

4.18 Natural or Depleted Resource Requirements and Conservation Potential of Various Alternatives.................29

4.19 Irreversible and Irretrievable Commitments of Resources..............................................................................29

4.20 Other Required Disclosures.............................................................................................................................29

Figures and Tables

Figure 4-1. This experiment station and insectary in Malden, Massachusetts, was used for some 

of the earliest research on the gypsy moth ........................................................................................................Cover

Table 4-1. Treatments available for use, by alternative.............................................................................................1

Table 4-2. Comparative Hazard Quotients (HQs) for the effects of gypsy moths and treatments 

on human health and nontarget organisms ................................................................................................................3

Environmental Consequences

Chapter 4 - Page 

This chapter examines, on a national scale, the 

environmental consequences of the alternatives as 

they relate to the issues of human health and nontarget 

organisms associated with the treatments that could be 

used. It updates the general background information 

presented in the 1995 EIS, and the human health and 

ecological risk information for the gypsy moth and 

for currently approved treatments. This chapter also 

presents human health and ecological risk information 

for tebufenozide (Alternatives 2 and 3) and other 

new treatments that may be available in the future 

(Alternative 3). Any information with a reference 

date between 1995 and 2006 is new since the 1995 

environmental impact statement. All of the information 

on tebufenozide is new.

4.1 Alternatives and 

Treatments.

Chapter 2 states the three alternatives. Table 4-1 lists 

the treatments that would be available under each 

alternative.

4.2 Risk Assessments and Risk 

Characterization.

Overview.

The consequences of the treatments in each alternative 

were determined by risk assessment for each treatment 

as well as for gypsy moth (no treatment) and a risk 

comparison among the treatments and gypsy moth (see 

Appendixes F-L for the risk assessments, and Appendix 

M for the risk comparison).

A risk assessment provides a logical process for 

evaluating data and analyzing potential effects of the 

gypsy moth and treatments. Risk assessments take into 

account the manner in which treatments are used in 

gypsy moth projects, including how treatment agents 

are applied, the amount applied, and the types of areas 

that receive treatment.

Standard steps in the risk assessment process were 

followed:

• Hazard identification—gathers known information 

from laboratory and field studies on toxicity of the 

gypsy moth and treatment agents.

Table 4-1. Treatments available for use, by alternative

Treatment

Alternative 1

No action

Alternative 2

Add tebufenozide

Alternative 3

Add tebufenozide and 

other treatments

B.t.k.*   

Diflubenzuron*   

Gypchek*   

Mass Trapping (Disparlure, or disparlure and dichlorvos)*   

Mating disruption (Disparlure)*   

Sterile insect technique*   

Tebufenozide  

Other treatments 

* Currently approved treatments
Chapter 4

Chapter 4 - Page 

• Exposure assessment—describes the nature and 

magnitude of contact with the gypsy moth and with 

treatment agents as they are used in gypsy moth 

treatment projects.

• Dose-response assessment—determines how much 

exposure to the gypsy moth and to treatment agents is 

needed to produce the response (effect) described in 

the hazard identification.

• Risk characterization—combines information 

from previous steps to describe the plausibility of 

observing certain effects of the gypsy moth and of

treatments.

Each step in a risk assessment is accompanied by 

uncertainties, caused by limitations either in the 

available data or in the ability to relate the data to 

scenarios of concern. To compensate for uncertainties, 

risk assessment results tend to be conservative, 

meaning they are more likely to overestimate risks than 

to underestimate them.

Human Health and Ecological Risk Assessments 

(HHERA) were prepared by risk assessment experts 

(Syracuse Environmental Research Associates, Inc. 

[SERA]), using the best available data. The HHERAs 

also underwent independent technical review by 

other recognized experts in risk assessment methods, 

toxicology, and other applicable fields (consultants 

retained by SERA, and toxicologists and program 

specialists from APHIS and the Forest Service). The 

HHERAs and this chapter cover the issues raised in 

scoping for this SEIS for both human health (human 

health assessment portion of HHERA) and nontarget 

organisms (ecological risk assessment portion of 

HHERA).

Many uncertainties are inherent in conducting and 

interpreting risk assessments; however, the data 

available on the agents covered by the risk assessments, 

modeling, equations and statistics all taken together 

with the understanding of uncertainties provide 

adequate information to characterize the relative 

hazards associated with the agents evaluated. To 

compensate for missing data and any uncertainties in 

the data, numerical uncertainty factors are used in the 

dose-response assessments for potential human health 

effects, and conservative assumptions are used in both 

human health and ecological risk assessments. In 

addition, it is virtually impossible to precisely calculate 

an exposure value for every situation that may arise. 

Therefore, models, equations, and statistical techniques 

were used to quantify both plausible and extreme 

exposures and to use ranges of toxicity values to reflect 

ranges of sensitivity. These ranges for exposure and 

toxicity are then used to numerically characterize risk 

with hazard quotients that are typically expressed as 

central estimates with upper and lower bounds.

HHERAs were prepared for each of the treatments 

in the alternatives (Appendixes F through K) and for 

the gypsy moth itself (Appendix L). Results of the 

HHERAs are summarized later in this chapter. The 

relative risks of the insecticides and treatments are 

illustrated in a risk comparison evaluation in 

Appendix M.

Hazard Quotients.

Risks to human health and to nontarget organisms 

can be estimated numerically using hazard quotients 

(HQs). HQs can be calculated only for effects on 

populations of biotic (living) organisms. The HQ is 

a screening tool commonly used in risk assessments. 

The HQ is a ratio of the exposure estimate for a 

particular and defined situation (labeled or prescribed 

conditions) for a representative population (human or 

nontarget species), divided by an effect level (dose or 

concentration level). The HQ takes into account the 

inherent toxicity of a substance, as well as its ability to 

produce specific effects on an organism (or population 

of organisms), and the degree of exposure. The HQs 

for currently approved treatments and tebufenozide are 

described in Appendix M. Table 4-2 provides the HQs 

for all of the treatments and for the gypsy moth.
Environmental Consequences

Chapter 4 - Page 

As an example, refer to the upper bound of the 

HQ for B.t.k. for nontarget aquatic species--0.5, in 

Table 4-2. This HQ was derived from an exposure 

estimate of 0.24 mg/L, which is calculated as the peak 

concentration of the B.t.k. formulation in water after 

a direct spray. This exposure estimate serves as the 

numerator for the HQ. The toxicity value of 0.45 

mg/L is the NOEC (no observed effect concentration) 

Table 4-2. Comparative Hazard Quotients (HQs) for the effects of gypsy moths and treatments on human health and 

nontarget organisms. (Wherever a 0 appears, the hazard quotient value is less than 0.01.)

Population Gypsy Moth 

HQ

B.t.k. 

HQ

Dichlorvos 

HQ

Diflubenzuron

HQ

Disparlure 

HQ

Gypchek 

HQ

Tebufenozide 

HQ

Human 

health

(See Table 

3-4 of Appendix M 

for in-depth 

comments)

1.6 to 625

Upper range is 

based on major outbreaks

0 to 0.04

Unlikely 

effects

0 to 380

Upper range 

based on 

child tampering with 

strip.

0.05 to 0.5–

workers

0.09 to 0.1– 

public

Upper range 

for workers based on 

ground spray 

operations.

0

No potential 

risk can be 

identified

0 to 0.02

No risks are 

plausible

0.03 to 1.5

Highest HQ 

based on longterm consumption of contaminated fruit 

following two 

applications at 

the highest application rate.

Nontarget 

terrestrial 

species

(See Table 

4-4 of 

Appendix M 

for in-depth 

comments)

0.25 to 400

Upper range 

based on 

gypsy moth 

outbreak in 

sensitive 

stands

0.36 to 9.4 

Upper range 

based on 

sensitive 

caterpillars 

of moths 

and 

butterflies

0 

Effects not 

likely

0.18 to 32 

Upper range 

based on 

sensitive species 

of invertebrates

0 

No potential 

hazard 

identified

0 

Effects not 

likely

0 to 4

Upper range 

based on the 

consumption 

of 

contaminated 

vegetation by a 

large mammal 

Nontarget 

aquatic species

(See Table 

4-5 of Appendix M 

for in-depth 

comments)

0 

No adverse 

effects

0 to 0.5

Upper 

level 

based on 

sensitive 

species

0 

No risks 

plausible 

in normal 

use. HQ 

for aquatic 

invertebrates 

could reach 

up to 8 in 

accidental 

exposures

0 to 5

Upper 

range based 

on acute 

exposure 

to aquatic 

invertebrates 

(Daphnia)

0 to 0.4

Upper range 

based on 

acute exposures to sensitive aquatic 

invertebrates 

(Daphnia)

0 

No adverse 

effects

0 to 0.4

Upper range 

based on 

longer term 

toxicity in sensitive aquatic 

invertebrates
Chapter 4

Chapter 4 - Page 

from a reproduction study in Daphnia magna, an 

aquatic invertebrate. This toxicity value serves as the 

denominator for the HQ. Thus, the HQ is calculated as 

follows:

HQ = exposure estimate/toxicity value 

= 0.24 mg/L / 0.45 mg/L 

= 0.533... ≈ 0.5

Note that the HQ in the above example is rounded to 

one significant place. This is a common practice in 

presenting HQ values except for those in which the 

level of concern is marginally exceeded, i.e., an HQ of 

1.45 would be rounded to 1.4 but not to 1.0.

In risk management, the HQ must be used in 

conjunction with other factors and characteristics 

of a substance, such as the quality and quantity of 

substantiating evidence (published scientific literature, 

data, models, and risk assessments done by others such 

as industry and universities), the severity of potential 

adverse effects, and the nature of the affected species 

and populations.

In some cases numerical expressions of risk (HQs) do 

not adequately convey the potential for hazard. For 

example, a high HQ for a mild effect, such as skin rash, 

is probably more acceptable than a much lower HQ for 

a more serious effect like neurotoxicity. Therefore, the 

use of HQ as an expression of risk and “acceptability” 

requires that a qualitative perspective also be injected 

into the analysis. Ecological risk assessments often 

involve considerations of many different species 

of plants and animals, and abiotic factors, and their 

interrelationships and interactions. Invariably, few 

data sets are available, and field studies provide only 

an overview of the complex interrelationships and 

secondary effects among species. Human health risk 

assessments and ecological risk assessments cannot 

offer a guarantee of safety. Both risk assessments offer 

a way to estimate the adverse effects and their severity.

4.3 Consequences of the Gypsy 

Moth.

This section provides exsisting and updated 

information on the gypsy moth. It is intended for use 

with site-specific project analysis and for general 

information for the reader. See Appendix E for 

information on the history and biology of the gypsy 

moth. See Appendixes L and M for detailed analysis of 

risks associated with gypsy moths.

General Effects of the Gypsy Moth.

Forest Condition–Effects of Defoliation on 

Vegetation.

When gypsy moth populations are low, nearly all 

feeding and defoliation occurs on favored hosts, such as 

oaks (Campbell and Sloan 1977a). During population 

outbreaks gypsy moth caterpillars feed on more than 

300 species of broad-leaved and coniferous trees and 

shrubs (Leonard 1981) (see Appendix D, Plant List). 

Trees stripped of 50 percent or more of their leaves are 

likely to refoliate the same season, although new leaves 

are fewer and smaller than the originals (Wargo 1981a). 

The impact of defoliation depends on five key factors:

(1) How much foliage is removed;

(2) The number of successive years of defoliation;

(3) When defoliation occurs in the growing season;

(4) The presence and number of secondary organisms; 

and

(5) The physiological condition of the tree (Parker 

1981).

Defoliated trees already under stress from drought 

or other factors often succumb more quickly than 

healthier trees.

After gypsy moth outbreaks red maple (Acer rubrum) 

numbers may increase and oak numbers decrease 

in Appalachian forests (Allen and Bowersox 1989, 

Gansner and others 1994, Hix and others 1991), 

because red maple is not a preferred host and oaks are 

preferred. Trends in New England and Pennsylvania 

reveal a shift in composition towards less oak, with 
Environmental Consequences

Chapter 4 - Page 

some stands having major losses and others having 

only minor changes (USDA Forest Service 1994f). 

Moderate-to-heavy defoliation accelerates forest 

succession towards more shade-tolerant (and less 

defoliation-prone) species (Campbell and Sloan 1977a, 

Clement and Nisbet 1972, Feicht and others 1993, 

Houston 1981b, Stephens and Hill 1971).

An area that is defoliated for only 1 year will have 

minimal long-term effects. However, defoliation by 

even non-epidemic levels of gypsy moth larvae could 

have a significant, negative effect on the radial growth 

of preferred trees, except possibly aspen (Muzika 

and Liebhold 1999, Naidoo and Lechowicz 2001). 

Small feeder roots die, reducing water and mineral 

uptake and slowing tree recovery (Wargo 1978b). The 

effects of a single heavy defoliation in a mixed stand 

of oaks in eastern New England were visible for 10 

years (Campbell and Sloan 1977a). Decreases in stem 

volume growth in southern New England averaged 

approximately 20 percent in any year a tree was 

defoliated compared with no defoliation the previous 

year, and growth loss was evident up to 3 years after 

defoliation (Twery 1987, Wargo 1981a). Overall stand 

volume may decrease initially (Gansner and Herrick 

1982, Herrick and Gansner 1988) and then may 

increase over time (Gansner and others 1993b).

Defoliation reduces carbohydrate (starch) production 

(Heichel and Turner 1976, Kozlowski 1969) forcing 

trees to use root starch reserves. Most trees can tolerate 

2 years of defoliation before root starch reserves 

are depleted (Wargo 1981a). Depletion of reserves 

weakens trees, making them vulnerable to secondary 

organisms that cause further decline and death. In the 

eastern United States the principal secondary organisms 

are the shoestring fungus (Armillaria mellea) and the 

two-lined chestnut borer (Agrilus bilineatus) (Houston 

1981a, Wargo 1981b).

Increased light due to defoliation causes herbaceous 

plants to rapidly expand their density and coverage 

(Gottschalk 1988). In some areas that are subject 

to intense deer browsing, defoliated trees may fail 

to regenerate, and shrubs or herbaceous plants can 

dominate (Gottschalk 1988). 

Heavy defoliation by the gypsy moth increases fire 

danger (Gottschalk 1990a). An abundance of heavy 

fuel, standing dead snags, dense understory vegetation, 

and numerous fallen trees act in combination to 

promote spot fires, impede fire line construction, and 

extend the time needed for post-fire mop-up operations 

(Tigner 1992).

Forest Condition—Tree Mortality.

Several factors interact to produce tree and stand 

mortality: severity, frequency, and distribution of 

defoliation, site and stand factors, environmental 

conditions, tree vigor, crown condition, and presence 

and abundance of secondary organisms (Campbell and 

Valentine 1971, Kulman 1971, Staley 1965, Campbell 

and Sloan 1977a, Gansner and others 1978, Wargo 

1978a, b, Campbell 1979, Herrick and Gansner 1987, 

Fosbroke and Hicks 1989, Herrick 1982, Tigner 1992, 

Feicht and others 1993, Gottschalk and MacFarlane 

1993). Oak mortality in initial outbreaks is greater than 

in later outbreaks (Davidson and others 1999). Oaks 

and other susceptible species experience more severe 

and frequent defoliation and have higher mortality than 

do non-susceptible species (Campbell and Sloan 1977a; 

Herrick and Gansner 1987; Quimby 1985, 1987).

Mortality can vary from stand-to-stand, even when 

stands have similar characteristics with mortality 80 

to 100 percent in some stands (Campbell and Sloan 

1977a, Gansner and Herrick 1984). Most mortality 

occurs during and after the initial outbreak (Twery 

1991) with severe mortality along and behind an 

advancing outbreak front as the gypsy moth invades 

new areas (Gansner and Herrick 1984, Herrick 

and Gansner 1986, Twery and Gottschalk 1988). 

Subdominant trees typically have much higher 

mortality rates than dominant trees after heavy 

defoliation (Campbell 1979, Gansner and others 1993c. 

Quimby 1993). The most common response to canopy 
Chapter 4

Chapter 4 - Page 

gaps created by tree mortality is increased growth and 

density of existing understory woody plants (Collins 

1961, Ehrenfeld 1980, Feicht and others 1993, Hix and 

others 1991, USDA Forest Service 1994f).

Drought may increase the severity of gypsy moth 

effects on trees (Bess and others 1947, Campbell and 

Sloan 1977a, Stephens and Hill 1971). Should severe 

drought occur with repeated years of defoliation, the 

cumulative impacts may increase mortality. Stress 

from disturbances, such as timber cutting or fire, and 

naturally occurring oak decline can also increase 

mortality.

Forest Condition—Seed and Mast Production.

Nuts, seeds, and fruits that serve as food for animals 

in the forest are called mast. Seed production by 

defoliated oak trees is reduced directly through 

consumption of oak flowers and young acorns by gypsy 

moth caterpillars, and indirectly by abortion of acorns 

and—in the years after defoliation–reduced initiation of 

flower buds. Significant mortality of oaks (more than 

60 percent of basal area in a stand) must occur before 

acorn production is reduced significantly (Gottschalk 

1990b). Over the long term, an increase in soft mast, 

particularly berries, replaces the loss of hard mast 

such as acorns (Gottschalk 1990a), and mammals that 

usually eat acorns may start eating this soft mast.

Water Quality.

Defoliation by the gypsy moth may affect a number 

of characteristics of nearby water bodies, including 

temperature, flow rate and yield, sediment load, acidity 

levels, oxygen availability, nutrient concentration, and 

structural habitat for aquatic organisms. Defoliated 

riparian areas receive increased exposure to the sun. 

Increases in the amount of light penetrating stream 

surfaces and changes in water temperature can affect 

both plants and animals in the stream. Various factors 

influence stream temperature at a given point, including 

flow volume, hydraulic gradient, ground water 

discharge, degree of shading, and upstream conditions. 

Actual changes to water temperature vary from site to 

site and depend in part upon the degree and duration 

of defoliation (USDA Forest Service 1994f). On a 

headwater stream under a dense tree canopy, light 

penetration increased from 5 to 18 percent to 73 percent 

after a “massive” gypsy moth outbreak in Rhode Island 

(Sheath and others 1986). Water temperature increased 

by 3.7 °C (6.7 °F) in early July, and algal growth in the 

streambed increased dramatically.

Defoliation by the gypsy moth has been shown to 

increase water yield (Corbett and Lynch 1987), in part 

due to fewer available leaves to transpire moisture from 

the soil (Twery 1991). Increased water yields from 

forested watersheds may produce beneficial results, 

such as creating more wet areas during summer, which 

might enhance habitat for amphibians. Conversely, 

increased stream discharge may have a destabilizing 

effect on herbivorous insects (Eagle 1993).

Sediment loads from forested land are usually 

low; however, increases in stream velocities due to 

increased water yield can lead to increased erosion, 

sedimentation, and turbidity. Timber cutting, exclusive 

of disturbances caused by road construction and log 

removal, usually has little if any effect on stream 

turbidity and sedimentation (Corbett and Lynch 1987). 

Therefore, gypsy moth defoliation would be unlikely to 

cause an increase in watershed erosion.

Whenever defoliation by the gypsy moth causes tree 

mortality in riparian areas, the structural habitat of 

streams may be altered by deposition of woody debris 

in affected streams. Debris dams may trap more 

organic material, lengthening the time it is available for 

ingestion by benthic invertebrates and leaf shredders, 

and allowing for more complete energy utilization. 

Large, woody materials also provide improved fisheries 

habitat (USDA Forest Service 1994f).

Defoliation by the gypsy moth may contribute to 

alterations in water chemistry and a reduction in the 

capacity to neutralize acids in some streams associated 
Environmental Consequences

Chapter 4 - Page 

with upland watersheds in the southern Appalachian 

region (USDA Forest Service 1994f). Defoliation 

temporarily produces conditions typical of winter, such 

as reduced acid-neutralizing capacity and increased 

acidity (Downey 1991). Acid-neutralizing capacity 

determines the concentrations of hydrogen and 

aluminum in solution, which at elevated levels are toxic 

to fish and other aquatic organisms. Acid-neutralizing 

capacity of streams increases seasonally, when 

deciduous leaves are present in the tree canopy.

Increased organic matter in streams from gypsy moth 

frass and leaf fragments, in combination with increased 

light penetrating the surface of the water, may lead to 

over-enrichment and result in excessive growth of algae 

and other microorganisms. This bloom could cause 

a reduction in oxygen available to other organisms 

in the stream. Large increases in fecal coliform and 

streptococci densities have been observed in streams 

where heavy gypsy moth defoliation has occurred 

(Corbett and Lynch 1987).

Defoliation is also suspected of causing increased 

nitrate mobility, which would allow nitrate to be 

lost from a site. Elevated concentrations of nitrate 

in streams have been associated with forest harvest 

(Vitousek and Melillo 1979) and defoliation by insects 

(Swank and others 1981, USDA Forest Service 1994f). 

Defoliation by the gypsy moth can accelerate the 

transfer of nutrients from vegetation to the soil surface; 

however, there is little evidence that these nutrients are 

lost from the site and enter adjacent water bodies to a 

significant degree (Eagle 1993, Grace 1986).

Soil Condition.

Gypsy moth defoliation probably increases the rate of 

decomposition of organic matter and decreases soil 

moisture content because of the greater penetration of 

sunlight increasing biological activity (Grace 1986, 

Tomblin 1994). These changes should result in shortterm increases in biological productivity.

Microclimate. 

The microclimate of defoliated areas is affected by 

rises in soil, leaf litter, and ambient air temperatures 

due to increased exposure to sunlight (Vaughan and 

Kasbohm 1993).

Risk to Human Health (Issue 1).

General.

People coming in contact with gypsy moth larvae 

may have skin irritation, resembling mosquito bites, 

with raised patches of skin approximately 0.25 to 0.5 

inches in diameter. Some people may have itching 

persisting several days to 2 weeks and sufficiently 

severe to cause them to seek medical treatment. Heavy 

infestations or extreme outbreaks potentially cause eye 

and respiratory effects in some individuals. Heavy 

infestations are often considered a public nuisance, 

causing esthetic damage to the environment through 

tree defoliation which may induce stress or anxiety in 

some individuals.

Groups at Special Risk.

Young children are potentially at greater risk of effects 

from gypsy moth exposure perhaps because they spend 

more time outdoors than adults (Tuthill and others 

1984, Aber and others 1982, Anderson and Furniss 

1983).

Risk to Nontarget Organisms 

(Issue 2).

Mammals.

Fur reduces the risk of direct contact with gypsy 

moth hairs making skin irritation unlikely. Evidence 

of irritation to the eyes and or respiratory tract in 

mammalian wildlife species after direct contact with 

the gypsy moth is not found in the literature.
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Chapter 4 - Page 

To determine the effects of a gypsy moth outbreak 

on a population of black bears (Ursus americanus), 

Vaughan and Kasbohm (1993) monitored the behavior 

of 54 radio-collared black bears in the Shenandoah 

National Park after a gypsy moth outbreak that caused 

widespread defoliation, hard mast failures, and tree 

mortality. The outbreak had no apparent effects on cub 

production or mortality rates of cubs or adults. In the 

fall, before the gypsy moth infestation, the bears ate 

mostly acorns. When acorns were no longer available 

due to defoliation, the bears switched to eating fruit, 

which had no apparent impact on the nutritional quality 

of their diets. Seventy-one percent of bear dens were 

in tree cavities, primarily in living oaks. Gypsy mothinduced mortality of den trees was high and, by the end 

of the study, 54 percent of the living oaks used as dens 

were dead. While no short-term effects were noted, 

Vaughan and Kasbohm (1993) speculated that the longterm adverse impact of defoliation on black bears may 

be a reduction in den sites, with natural replacement 

possibly requiring 50 years. Conversely, black bears 

will use upturned stumps of large dead trees as dens. 

These would be expected to increase as tree mortality 

increases.

Variations in acorn and other mast production are 

directly related to variations in populations of squirrels, 

mice, and other small mammals (Brooks and others 

1998). Acorn crop size in the fall directly affects the 

population density of mice living in oak-dominated 

forests the following spring (McShea and Rappole 

1992, McShea and Schwede 1993). A decrease in 

acorn production has been shown to decrease the 

population of white-footed mice, Peromyscus leucopus

(Elkinton and others 1996, 2002).

White-tailed deer will migrate to areas that have not 

been defoliated. Nesting failures of grouse and turkey 

may increase. Bear, turkey, and bats may migrate to 

nondefoliated areas or less defoliated areas (USDA 

1995).

Sample and others (1996) found no significant effects 

on the consumption of insects by Virginia big-eared 

bats in areas of high gypsy moth infestation and 

defoliation.

Birds.

Some species of birds appear to avoid the gypsy moth 

as a prey species (Smith 1985), perhaps because of 

larval hairs. Reported increases in nesting failures of 

various species of birds appear to be due to increased 

predation, increased weather stress, or both, which are 

associated with defoliation (Thurber and others 1994).

Gypsy moth infestations and subsequent defoliation 

may be beneficial to some species of birds, especially 

species that favor dead wood (snags) as a habitat (Bell 

and Whitmore 1997a, b; DeGraaf 1987; DeGraaf 

and Holland 1978; Showalter and Whitmore 2002). 

Available nesting and foraging resources increased for 

several bird species as a result of more snags, windfall, 

and shrub cover after defoliation, while there was no 

substantial impact from upper canopy defoliation on 

birds residing primarily in the forest canopy (Bell and 

Whitmore 1997a, b).

Cavity-nesting birds benefit indirectly from a gypsy 

moth outbreak (Showalter and Whitmore 2002). 

Bird density increased in plots with low to moderate 

defoliation (Thurber 1993). Species richness increased 

from 19 to 23 species per plot, with declines noted only 

for tree nesters and flycatchers on high-impact plots 

(Thurber 1993). Increases in low shrub and ground 

nesters, cavity nesters, low shrub and ground foragers, 

bark foragers, forest edge species, short-distance 

migrants, year-round residents, and woodpeckers were 

widespread, but most pronounced on moderate-impact 

plots. DeGraaf and Holland (1978) reported similar 

results, finding significantly fewer numbers of only 4 

out of 36 bird species examined in heavily defoliated 

areas. No substantial effects on abundance of various 

species of birds in defoliated and nondefoliated stands 

were noted in central Pennsylvania over a 2-year period 

(DeGraaf 1987).
Environmental Consequences

Chapter 4 - Page 

Terrestrial Invertebrates.

Some lepidopteran species may be adversely affected 

by gypsy moth outbreaks. Redman and Scriber (2000) 

examined the adverse effects of the gypsy moth on the 

northern tiger swallowtail butterfly, Papilio canadensi. 

Direct effects included 100 percent mortality in Papilio

larvae exposed to leaves painted with gypsy moth body 

fluids, and 84 percent mortality in Papilio larvae fed 

leaves from aspen stands infested with gypsy moth 

larvae. 

The potential adverse effects of gypsy moth outbreaks 

to Lepidoptera was also investigated in a study 

designed to compare lepidopteran populations in 50 

acre plots in mixed oak, hickory, and pine forests in 

West Virginia (Sample and others 1996). Decreases 

in abundance and richness of larvae and adults from 

the family Arctiidae (tiger moths) were apparent in 

plots infested with gypsy moth larvae, compared with 

uncontaminated plots.

The impact of the gypsy moth is negative to only 

a small proportion of the lepidopteran community, 

primarily species that feed on oak and for which the 

larval development of the affected species and gypsy 

moth presumably coincide (Work and McCullough 

2000). Although the study does not address the 

mechanism(s) by which the gypsy moths adversely 

affect the lepidopteran community, the investigators 

suggest they might include altered host plant quality, 

increases in natural enemies, or microclimate changes.

Some reports suggest that certain lepidopteran species 

respond positively to gypsy moth infestations. In 1981, 

the number of butterfly species was at a record high for 

the New Haven, Connecticut, area, despite the record 

number of acres defoliated by the gypsy moth that 

same year (Schweitzer 1988).

Fish.

Little information is available regarding the effects 

of gypsy moth infestations on fish populations. 

Defoliation by the gypsy moth can result in an 

increase in the pH and temperature of ambient water 

(Downey and others 1994, Webb and others 1995a). 

Trout, which are very sensitive to changes in pH and 

temperature, could be adversely affected by such 

changes (Downey and others 1994). No direct data are 

available on the biological effects of such changes due 

to gypsy moth defoliation (Webb and others 1995a).

Aquatic Invertebrates.

The rate of leaf breakdown in streams apparently 

increased due to gypsy moth defoliation, which might 

result in food deficits during spring for shredders, such 

as caddisflies, stoneflies, and some dipterans (Hutchens 

and Benfield 2000). The number of shredders 

collected, however, was greater in disturbed streams 

(i.e., streams in areas of gypsy moth defoliation) than 

in control streams.

Cumulative Effects of the Gypsy 

Moth.

Risk to Human Health (Issue 1).

The available data do not permit a definitive 

assessment of the effects of exposure to the gypsy 

moth over several seasons. Some individuals may 

become sensitized to the gypsy moth after repeated 

exposures over one or more seasons. Young children 

may be a group at special risk from effects of gypsy 

moth exposure but it is not clear whether children are 

more sensitive than adults to the effects of gypsy moth 

exposure or whether responses in children appear 

greater because children spend more time outdoors than 

with adults do.

Risk to Nontarget Organisms (Issue 2). 

Effects due to the gypsy moth would be cumulative in 

situations of repeated outbreaks and defoliation in the 

same area. Repeated defoliation would lead to changes 

in forest condition that are characterized by increased 

tree mortality, stand structure and composition changes, 

a shift from production of hard to soft mast, and 

increased fire danger.
Chapter 4

Chapter 4 - Page 10

Habitats of wildlife species are altered more with 

each successive outbreak of the gypsy moth. 

Recolonization of species lost or displaced due to 

changes in habitat is possible; however, large areas of 

defoliation and frequent repeated defoliation do not 

favor recolonization by species with low dispersal 

capabilities.

Economic and recreational consequences that 

accumulate with repeated multiyear outbreaks 

include these: costs associated with annual cleanup; 

maintenance and replacement of trees that die; and loss 

of value from reduced growth and mortality of trees.

4.4 Consequences of Bacillus 

thuringiensis var. kurstaki 

(B.t.k.) (Alternatives 1, 2, 3)

See Appendixes F and M for detailed analysis of risks 

associated with B.t.k.

General Effects of B.t.k.

B.t.k. may indirectly help to maintain existing 

forest conditions, water quality, microclimate, and 

soil condition by delaying increases in gypsy moth 

populations, thereby protecting tree foliage.

Risk to Human Health (Issue 1).

General.

B.t.k. and its formulations may cause irritation to the 

skin, eyes, and respiratory tract; however, serious 

adverse health effects are improbable. Overt signs of 

systemic toxicity are not likely to be observed in any 

group–ground workers, aerial workers, or members 

of the general public–that is exposed to B.t.k. as the 

result of gypsy moth management programs conducted 

by the USDA (Appendix M). Throat irritation is the 

most frequently documented effect of B.t.k. in the 

scientific literature on human health (Appendixes F and 

M). Dermal and ocular irritations are observed at the 

extreme upper levels of exposure.

There is little indication that B.t.k. is associated with 

pathogenicity in humans and no indication of endocrine 

disruption or reproductive effects. Carcinogenic 

and mutagenic effects are not likely. Neither B.t.k.

nor its commercial formulations are highly toxic or 

infectious (Appendixes F and M). Formulations of 

B.t.k. are likely to cause irritant effects to the skin, 

eyes, and respiratory tract; however, concerns about 

serious adverse health effects are not plausible. 

This risk characterization is consistent with the risk 

characterization in the previous USDA risk assessment 

(USDA 1995), as well as with more recent risk 

assessments conducted by the U.S. Environmental 

Protection Agency (EPA) and the World Health 

Organization, and the comprehensive review of B.t.

published by Glare and O’Callaghan (2000).

Pretreatment with an influenza virus substantially 

increased mortality in mice exposed to various doses of 

B.t.k (Hernandez and others 2000). These results raise 

questions about the susceptibility of individuals who 

contact influenza or other viral respiratory infections 

prior to B.t.k. applications and have viral infections 

at the time of application. The enhancement of 

bacterial infections by a virus is not uncommon, and 

the enhancement of B.t.k. toxicity by a viral infection 

is, in some respects, not surprising. The relevance of 

this observation to public health cannot be completely 

assessed at this time. Several epidemiological 

studies have been conducted on the effects of B.t.k.

on human populations, and none have reported viral 

enhancement. It is uncertain whether epidemiology 

studies would detect such an effect or whether such an 

effect is plausible under the anticipated exposure levels 

used in programs to control the gypsy moth. The viral 

enhancement of B.t.k. toxicity is likely to be an area of 

further study in the coming years.

Groups at Special Risk. 

The available toxicity data give no indication that 

subgroups of the general population are likely to be 

remarkably sensitive to B.t.k. Nonetheless, B.t.k. 

formulations are complex mixtures and there is a 
Environmental Consequences

Chapter 4 - Page 11

possibility that certain individuals may be allergic to 

one or more of the components in the formulations. 

The study by Hernandez and others (2000) also raises 

concern regarding the susceptibility of individuals with 

influenza or other viral respiratory infections to B.t.k. 

toxicity (Appendix F). See Appendixes F and M for 

detailed information.

Risk to Nontarget Organisms 

(Issue 2).

Mammals.

Adverse effects due to B.t.k. are unlikely in mammals 

(Appendixes F and M). Most inhalation studies do not 

suggest the potential for adverse effects, even at B.t.k.

concentrations much greater than those likely to be 

encountered in the environment (Appendix F). Bats 

that feed almost exclusively on lepidopterans might 

be indirectly affected through a reduction in prey, as 

suggested by a study in West Virginia (Sample and 

others 1993a, b; Sample and Whitmore 1993). A 3-

year study (1990-1992) conducted in West Virginia 

on food of the endangered big-eared bat revealed the 

greatest impact within 3 weeks of B.t.k. application due 

to reduction of prey species. Contrasting these studies, 

Sample and others (1996) showed that the moths on 

which bats feed were not affected by B.t.k. applications.

Birds.

Acute toxic effects are not likely in birds (Appendixes 

F and M). Due to the lack of toxicity of B.t.k.

formulations, as well as of other B.t. strains, the U.S. 

EPA did not require chronic or reproductive toxicity 

studies in birds (Appendix F). This apparent lack 

of toxicity is supported by numerous field studies in 

birds. B.t.k. applied at rates sufficient to decrease 

the number of caterpillars had no substantial adverse 

effects on most bird species (Rodenhouse and Holmes 

1992, Nagy and Smith 1997, Sopuck and others 

2002). However, a study showed a significant decline 

in three species of insectivorous birds (black throated 

green warbler, eastern tufted titmouse, and yellowbilled cuckoo), but they fully recovered within 3 years 

(Strazanac and Butler 2005).

A field study that included intensive searches of plots 

in sprayed and unsprayed areas revealed no differences 

in the numbers of songbird broods between the two 

areas for any of the species examined (Sopuck and 

others 2002). A reduction of lepidopteran larvae due 

to B.t.k. application appeared to have only minimal 

effects on reproduction in hooded warblers (Nagy and 

Smith 1997). The reduction in numbers of birds in an 

area observed in some species was considered indirect 

and attributed to alterations in the availability of prey 

rather than to the direct toxicity of B.t.k. (Gaddis 1987; 

Gaddis and Corkran 1986; Norton and others 2001).

Terrestrial Invertebrates.

B.t.k. is toxic to several species of target and nontarget 

Lepidoptera. The larvae of the Karner Blue Butterfly 

(a Federally listed endangered species), two species 

of swallowtail butterflies, a promethean moth, the 

cinnabar moth and various species of Nymphalidae, 

Lasiocampidae, and Saturniidae are susceptible to B.t.k.

(Glare and O’Callaghan 2000)

Permanent changes in nontarget caterpillar populations 

do not appear likely as a result of gypsy moth 

management projects. An exception might occur 

in certain habitat types that support small isolated 

populations of lepidopterans that are highly susceptible 

to B.t.k. If unaffected individuals of the same species 

are unlikely to, or physically cannot, move from the 

treated into the untreated area, then one application 

of B.t.k. will have an effect on the ability of those 

populations to recover. These effects are limited 

to spring caterpillars that are present during B.t.k.

treatments (Strazanac and Butler 2005). Full recovery 

of nontarget spring caterpillars occurred within 1 to 2 

years after the treatment (Strazanac and Butler 2005).
Chapter 4

Chapter 4 - Page 12

In Oregon, Miller (1990) observed reductions in both 

types and numbers of nontarget caterpillars after three 

applications of B.t.k. The reductions persisted for 1 

year after treatment but not for 2 years. In another 

study (Carter and others 1995), a second application 

of B.t.k. did not increase mortality of five species of 

Lepidopterans over that caused by one application. 

The species tested were moderately resistant to B.t.k.

and had mortality rates below 50 percent after the first 

application.

While some nontarget lepidopteran species appear 

to be as sensitive to B.t.k., most studies indicate that 

effects in other terrestrial insects are likely to be of 

minor significance (Appendix F). There is relatively 

little information regarding the toxicity of B.t.k. or 

B.t.k. formulations to terrestrial invertebrates other than 

insects. For some Lepidoptera, sensitivity to B.t.k. is 

highly dependent on their developmental stage. This is 

particularly evident for the cinnabar moth, where late 

instar larvae are very sensitive to B.t.k. and early instar 

larvae are very tolerant to B.t.k. (James and others 

1993).

The variability in the response of nontarget Lepidoptera 

to B.t.k. is also illustrated in a recent field study 

in which a B.t.k. formulation was applied to two 

forests (dominated by oak, hickory, and maple trees) 

over a 2-year period, at an application rate of 40 

BIU/acre (Rastall and others 2003). Researchers 

monitored nontarget lepidopteran populations in the 

2 years prior to application as well as over the 2-year 

period in which B.t.k. was applied. The response of 

nontarget Lepidoptera varied substantially among 

different species. Larvae of three lepidopteran species 

significantly decreased in treatment years: Lambdina 

fervidaria (geometrid), Heterocampa guttivitta

(notodontid), and Achatia distincta (noctuid). For 19 

other species, larval counts were significantly higher 

in treatment years as were the total number of noctuids 

combined and the total number of all nontarget 

lepidopteran species combined. The Karner Blue 

Butterfly is susceptible to B.t.k., although the larval 

generation at risk may vary from year to year (Herms 

and others 1997).

Some predators and parasitoids may be affected 

indirectly by B.t.k. because of the loss of gypsy 

moth caterpillars that they parasitize or eat. The 

more specific the parasites and predators are for 

lepidopterans affected by B.t.k., the greater the chance 

of an effect. For example, populations of parasitoid 

tachinid flies and Braconidae wasps and Pentatomidae 

stinkbugs declined after application of B.t.k. (then 

recovered by the second year), but generalist predators 

did not decline (Strazanac and Butler 2005).

Fish.

The U.S. EPA classifies B.t.k. as virtually nontoxic to 

fish (Appendix F). This assessment is consistent with 

the bulk of experimental studies reporting few adverse 

effects in fish exposed to B.t.k. concentrations that 

exceed environmental concentrations associated with 

USDA programs (Buckner and others 1975, Otvos and 

Vanderveen 1993).

Aquatic Invertebrates.

The effects of B.t.k. on aquatic invertebrates is 

examined in standard laboratory studies and in 

numerous field studies. B.t.k. may be lethal to 

certain aquatic invertebrates, like Daphnia magna, 

at concentrations high enough to cause decreases 

in dissolved oxygen or increased biological oxygen 

demand (Young 1990). Most aquatic invertebrates 

seem relatively tolerant to B.t.k. (Appendix F, Section 

4.1.3.3). This assessment is supported by several 

field studies that failed to note effects in most species 

after exposures that substantially exceed expected 

environmental concentrations (Kreutzweiser and others 

1992, 1993, 1994).
Environmental Consequences

Chapter 4 - Page 13

Cumulative Effects of B.t.k.

Risk to Human Health (Issue 1).

Given the reversible nature of the irritant effects of 

B.t.k. and the low risks for serious health effects, 

cumulative human health effects from spray programs 

conducted over several years are not expected. Mating 

disruption with disparlure will most likely be the only 

other treatment used in the same spray blocks with 

B.t.k. However, B.t.k. is used to treat gypsy moth 

larvae, and mating disruption is used against gypsy 

moth adults, and they are applied weeks apart. These 

treatments also have different modes of action, and 

there are no known cumulative effects between the 

treatments.

Workers or members of the general public who are 

exposed to aerial or ground sprays of B.t.k. are also 

exposed to the gypsy moth and may be exposed to 

other control agents for the gypsy moth. No known 

data indicate that risks posed by these other agents 

will affect the response, if any, to B.t.k. formulations. 

Similarly, exposure to other chemicals in the 

environment may impact the sensitivity of individuals 

to B.t.k. or other agents; however, the available data 

are not useful for assessing the significance of such 

interactions.

There is no known documented evidence of a subgroup 

of individuals who are more sensitive than most 

members of the general public to B.t.k. formulations 

(Appendix F).

Risk to Nontarget Organisms (Issue 2).

Many studies indicate that B.t.k. lasts about a week in 

the environment. Repeated treatments of areas with 

B.t.k. could potentially impact some species of springfeeding butterfly and moth caterpillars. Since B.t.k. is 

not used in the same spray blocks with other treatments 

that could affect nontarget organisms, there is no 

cumulative effect between different treatments and 

B.t.k. on spring-feeding caterpillars.

4.5 Consequences of 

Diflubenzuron 

(Alternatives 1, 2, 3).

See Appendixes I and M for detailed analysis of risks 

associated with diflubenzuron.

General Effects of Diflubenzuron.

Diflubenzuron may indirectly help to maintain existing 

forest conditions, water quality, microclimate, and 

soil condition by delaying increases in gypsy moth 

populations, thereby protecting tree foliage.

Risk to Human Health (Issue 1).

General.

Diflubenzuron causes the formation of methemoglobin, 

a form of hemoglobin incapable of oxygen 

transport, normally present in the blood in small 

amounts. Methemoglobinemia, the formation of 

excess methemoglobin, is the primary toxic effect 

of diflubenzuron in every species of animal tested, 

regardless of the route or duration of exposure. While 

effects on the blood are well documented, there is little 

indication that diflubenzuron causes other specific 

forms of toxicity. Diflubenzuron does not appear to 

be neurotoxic nor immunotoxic, does not appear to 

affect endocrine function in laboratory mammals, 

and is not a carcinogen. Additionally, diflubenzuron 

does not appear to cause birth defects or to affect 

reproductive processes. Numerous studies regarding 

the subchronic and chronic toxicity of diflubenzuron in 

laboratory animals indicate that methemoglobinemia 

is the most consistent clinical symptom indicative of 

toxicity. Diflubenzuron can be absorbed via the skin 

in sufficient amounts to cause hematological effects, 

that is, methemoglobinemia and sulfhemoglobinemia. 

Nonetheless, the dermal exposure concentrations 

necessary to induce these hematological effects are 

higher than the oral exposure dosage necessary to cause 

the same effects.
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Chapter 4 - Page 14

Diflubenzuron rapidly dissipates from vegetation and 

is broken down by sunlight; in the environment the 

compound degrades to 4-chloroaniline, which the 

EPA considers a potential carcinogen. This is the only 

identified potential carcinogen associated with any 

of the agents to control gypsy moth. The compound 

is not expected to be present in significant amounts 

during application since 4-chloroaniline does not form 

during application. The scenario of greatest concern 

involving 4-chloroaniline is a cancer risk from drinking 

contaminated water. This risk would be most plausible 

in areas with sandy soil and annual rainfall rates 

ranging from about 50 to 250 inches. The estimate 

of the hazard quotient for the consumption of water 

contaminated with 4-chloroaniline and based on a 

cancer risk of 1 in 1 million is 0.09, which is 10 times 

lower than the level of concern.

None of the hazard quotients for diflubenzuron reaches 

a level of concern at the highest application rate used 

in USDA programs (Appendix I). Since many of the 

exposure assessments overestimate exposure, and 

because the dose-response assessment is based on 

similarly protective assumptions, there is no basis for 

asserting that this use of diflubenzuron poses a hazard 

to human health (Appendix I).

Groups at Special Risk.

Some individuals have congenital methemoglobinemia 

and may be at increased risk of adverse effects to 

compounds that induce methemoglobinemia (Barretto 

and others 1984). Infants less than 3 months old 

have lower levels of methemoglobin (cytochrome b5) 

reductase and higher levels of methemoglobin (1.32 

percent), compared with older children or adults (Centa 

and others 1985, Khakoo and others 1993, Nilsson 

and others 1990). Some infants with an intolerance to 

cow’s milk or soy protein exhibit methemoglobinemia 

(Murray and Christie 1993, Wirth and Vogel 1988). 

These infants would be at increased risk if exposed to 

any materials contaminated with diflubenzuron or any 

compound that induces methemoglobinemia.

Individuals with poor diets might be vulnerable to some 

chemicals. Based on a study in rats, iron deficiency 

leads to anemia but does not influence methemoglobin 

reductase activity (Hagler and others 1981). Thus, 

although individuals with poor nutritional status are 

generally a group for which there is particular concern, 

the available information does not support an increased 

risk for these individuals with respect to diflubenzuron 

exposure.

Risk to Nontarget Organisms 

(Issue 2).

Mammals.

The available field studies indicate no substantial 

impacts on mammalian wildlife from applications 

of diflubenzuron. Applications of 60 to 280 g a.i./

ha (grams active ingredient per hectare) or 0.85 to 

4 oz a.i./acre (ounces of active ingredient per acre) 

had no detectable adverse effects on the abundance 

of, or reproduction in moles, field mice, and shrews 

(O’Connor and Moore 1975; Henderson and others 

1977). Small mammals increased in abundance on a 

plot receiving 280 g a.i./ha compared with a control 

plot (Henderson and others 1977). The adverse effects 

that diflubenzuron might have on bot flies, a parasite 

of small and large mammals alike, was suggested as a 

possible explanation.

A field study reported no effect on body measurements, 

weight, or fat content in populations of mice in areas 

treated with diflubenzuron (Seidel and Whitmore 

1995). Mice in the treated areas did consume less 

lepidopteran prey, but total food consumption was not 

significantly different between treated and untreated 

plots.

Birds.

The acute toxicity of diflubenzuron to birds appears 

generally low. The lack of direct effects on birds 

is supported by several field studies summarized 

in Appendix I. Effects secondary to a reduction in 

lepidopteran prey may include increased foraging 
Environmental Consequences

Chapter 4 - Page 15

range (Cooper and others 1990), relocation (Sample 

and others 1993a, b; Sample and Whitmore 1993) and 

lower body fat (Whitmore and others 1993).

Terrestrial Invertebrates.

Arthropods, a large group of invertebrates including 

insects, crustaceans, spiders, mites, and centipedes, 

are most sensitive to diflubenzuron. Most of these 

organisms use chitin as a major component of their 

exoskeleton (outer body shell). Diflubenzuron is an 

effective insecticide because it inhibits the formation 

of chitin, disrupting normal growth and development. 

Both terrestrial and aquatic arthropods are affected, 

though some substantial differences in sensitivity are 

apparent.

Invertebrates lacking exoskeletons, such as earthworms 

and snails, do not utilize chitin, and diflubenzuron 

is relatively nontoxic to these species (Appendix I). 

Species that are most sensitive to diflubenzuron include 

lepidopteran and beetle larvae, grasshoppers, and other 

chewing herbivorous insects (Berry and others 1993, 

Butler 1993, Butler and others 1997a, Elliott and Iyer 

1982, Jepson and Yemane 1991, Kumar and others 

1994, Sample and others 1993a, Sinha and others 1990, 

Redfern and others 1980). Species that are relatively 

tolerant to diflubenzuron include flies, parasitic wasps 

(on insect eggs), adult beetles, and sucking insects 

(Ables and others 1975, Broadbent and Pree 1984, 

Brown and Respicio 1981, Bull and Coleman 1985, 

De Clercq and others 1995, Delbeke and others 1997, 

Gordon and Cornect 1986, Keever and others 1977, 

Martinat and others 1988, Webb and others 1989, 

Zacarias and others 1998, Zungoli and others 1983). 

The U.S. EPA uses the honey bee as the standard test 

species to classify the toxicity of pesticides to nontarget 

terrestrial invertebrates. Based on early acute oral 

and contact toxicity studies in honey bees (Atkins and 

others 1974, Stevenson 1978), the U.S. EPA (1997) 

classifies diflubenzuron as practically nontoxic to 

honey bees. Several other laboratory toxicity studies 

also indicate diflubenzuron is not particularly toxic to 

bees (Chandel and Gupta 1992, Elliott and Iyer 1982, 

Gijswijt 1978, Kuijpers 1989, Nation and others 1986, 

Yu and others 1984). This conclusion is supported 

by several field studies conducted at application rates 

comparable to, or substantially higher than, those used 

to control the gypsy moth (Buckner and others 1975, 

Emmett and Archer 1980, Matthenius 1975, Schroeder 

1978, Schroeder and others 1980). Additionally, 

no detectable amounts of diflubenzuron were found 

in honey bees in areas treated with diflubenzuron 

(Cochran and Poling 1995).

Fish.

Based on the available information, the U.S. EPA 

(1997) classifies acute exposure to diflubenzuron as 

“practically nontoxic” to fish. The 96-hour LC50 values 

range from greater than 25 milligrams per Liter (mg/

L) (the value for yellow perch reported by Johnson 

and Finley 1980) to greater than 500 mg/L (the value 

for fathead minnow reported by Reiner and Parke 

1975). In addition, no effects were seen in longer-term 

studies at concentrations up to 100 parts per billion 

(ppb) (Cannon and Krize 1976) or in two-generation 

reproduction studies at concentrations of up to 50 ppb 

(Livingston and Koenig 1977).

Indirect effects on fish are plausible based on a 

decrease in invertebrate populations as demonstrated 

in studies in which concentrations as low as 2.5 

ppb resulted in decreased growth of fish in littoral 

enclosures (populations of fish placed in enclosures 

along the shore of a body of water and monitored) 

(Moffett 1995, Tanner and Moffett 1995). The reduced 

growth observed in these studies is attributed to a 

reduction in macroinvertebrates, a fish food source.

None of the field studies summarized in Appendix I 

note any adverse effects on fish at application rates 

comparable to or greater than those used in the control 

of the gypsy moth. A study by Colwell and Schaefer 
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Chapter 4 - Page 16

(1980) did note a shift in the diet of fish (secondary to 

changes in food availability) but no effect on growth 

rates or general condition of the fish.

Aquatic Invertebrates.

Because diflubenzuron inhibits the synthesis of chitin, 

crustaceans are the aquatic invertebrates most sensitive 

to diflubenzuron. Many bioassays, both acute and 

chronic, have been conducted on Daphnia magna

(Hansen and Garton 1982, Kuijpers 1988, Majori and 

others 1984, Surprenant 1988) as well as a related 

species, Ceriodaphnia dubia (Hall 1986). As detailed 

further in the dose-response assessment (Appendix 

I), these organisms are among the most sensitive to 

diflubenzuron (Hall 1986, Hansen and Garton 1982). 

Several other crustacean species appear to be about 

as sensitive as or only somewhat less sensitive to 

diflubenzuron than daphnids are (Appendix I). Small 

crustaceans that consume algae and serve as a food 

source for fish, such as Daphnia species, appear to be 

the most sensitive to diflubenzuron, while larger insect 

species, such as backswimmers and scavenger beetles, 

are much less sensitive. Other aquatic invertebrates, 

crustaceans, and small- to medium-sized aquatic insect 

larvae appear to have intermediate sensitivities.

Snails, aquatic worms, and bivalves were not affected 

by exposure to diflubenzuron (Hansen and Garton 

1982; Surprenant 1989).

Field studies on the effects of diflubenzuron on aquatic 

invertebrates reinforce the standard toxicity studies, 

indicating diflubenzuron will impact invertebrate 

populations. Several of these studies, however, were 

conducted at application rates substantially higher than 

those used to control the gypsy moth. Many of the 

studies in which severe adverse effects were observed 

in aquatic invertebrate populations involved multiple 

applications at rates between about 110 g/ha and 560 

g/ha (Ali and Mulla 1978a, b; Ali and others 1988; 

McAlonan 1975). Concentrations in this range are 

substantially higher than the application rate of 17.5 g/

ha that is likely to be encountered in USDA programs. 

Similarly, other field studies involve direct applications 

to open water, a treatment method that is not part 

of USDA program activities, and which resulted in 

concentrations of diflubenzuron in water in the range 

of 10 ppb (Apperson and others 1977, Boyle and others 

1996, Colwell and Schaefer 1980, Lahr and others 

2000, Sundaram and others 1991).

Diflubenzuron reduces numbers of stream invertebrates 

that process detritus; however, field studies have shown 

no decline in detrital decomposition rates (Swift and 

others 1988). The populations of some invertebrates 

that feed on algae are reduced by diflubenzuron. An 

increase in algae could occur after the loss of algal 

herbivores, however, this has not been observed in field 

studies.

Field studies using lower application rates that are more 

typical of USDA gypsy moth management programs 

noted some effects on freshwater invertebrates, 

particularly smaller crustaceans (Farlow 1976; Griffith 

and others 1996, 2000; Hurd and others 1996; Reardon 

1995). The effects were much less severe than those 

seen at higher application rates. See Section 4.4 of 

Appendix I for further discussion.

Cumulative Effects of Diflubenzuron.

Risk to Human Health (Issue 1).

Diflubenzuron is not likely to be used with other 

treatments at the same site, so no cumulative effects 

with other treatments are likely. Multiple applications 

at lower rates per application result in lower associated 

risks than with a single application at the maximum 

approved rate.

Diflubenzuron and tebufenozide could have a 

cumulative effect on methemoglobinemia. USDA 

gypsy moth management programs do not use these 

two chemicals together in the same area at the same 

time. Exposure to other methemoglobinemia-inducing 

compounds in the environment may contribute to a 
Environmental Consequences

Chapter 4 - Page 17

cumulative effect. Individuals exposed to combustion 

smoke or carbon monoxide (agents causing oxidative 

damage to blood) may be at increased risk of 

developing methemoglobinemia. Individuals exposed 

to high levels of nitrates, either in air or in water, 

demonstrate increased levels of methemoglobin and 

may be at increased risk with exposure to compounds 

such as diflubenzuron.

Some infants with congenital methemoglobinemia 

and an intolerance to cow’s milk or soy protein exhibit 

methemoglobinemia. These infants would be at 

increased risk if exposed to any materials contaminated 

with diflubenzuron.

Risk to Nontarget Organisms (Issue 2).

Diflubenzuron is generally not used in conjunction 

with other treatments; however, diflubenzuron might 

be applied to the same area in multiple years for 

eradication projects. In that case, diflubenzuron might 

have a cumulative effect on nontarget invertebrates, 

such as caterpillars of moths and butterflies, 

grasshoppers, parasitic wasps, aquatic insects, bottom 

dwelling crustaceans, and immature free-floating 

crustaceans. Diflubenzuron applications as used in 

USDA treatment projects will otherwise have no 

cumulative effects.

4.6 Consequences of 

Disparlure (as Used in Mating 

Disruption and Mass Trapping) 

(Alternatives 1, 2, 3).

See Appendixes H and M for detailed analysis of risks 

associated with disparlure as used in mating disruption 

and mass trapping.

In mating disruption, a medium is impregnated with 

disparlure for timed release and formulated for aerial 

application over the project area. The objective is 

to flood the area with pheromone, thereby impeding 

the male moth’s ability to find and mate with female 

moths. Also, in mass trapping, a solid medium is 

impregnated with disparlure, formulated for timed 

release, and deployed in small “delta” or large capacity 

“milk carton” traps. The traps are deployed across 

the treatment area to attract and capture male moths, 

thereby preventing them from finding and mating with 

female moths. The delta and milk carton traps are also 

used in detection surveys for gypsy moth.

General Effects of Disparlure.

Disparlure is specific to the gypsy moth and may 

indirectly help to maintain existing forest conditions, 

water quality, microclimate, and soil condition by 

delaying increases in gypsy moth populations, thereby 

protecting tree foliage.

Risk to Human Health (Issue 1).

General.

Insect sex pheromones are chemicals produced by 

insects for communication between the sexes of 

the same species. Insect pheromones are generally 

regarded as nontoxic to mammals and are commonly 

employed in very low concentrations. Consequently, 

the U.S. EPA requires less rigorous testing of these 

products than is required of chemical insecticides. 

Results of acute exposure studies for oral, dermal, 

ocular, and inhalation exposure to disparlure reveal 

no adverse effects. Based on the results of studies 

on disparlure itself (i.e., the active ingredient), acute 

exposure to disparlure exhibits very low toxicity to 

mammals.

No studies were identified investigating the effects 

of chronic exposure of mammals to disparlure or 

investigating the effects of disparlure on the nervous, 

immune, reproductive or endocrine systems of 

mammals. The carcinogenic potential of disparlure 

has not been assessed, though a single study focusing 

on mutagenicity revealed no indication that disparlure 

is mutagenic. No information is available regarding 

the kinetics and metabolism of disparlure in mammals; 

available literature does not document absorption 
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Chapter 4 - Page 18

of disparlure following dermal, oral, or inhalation 

exposure. A case report of an occupational exposure 

indicates that disparlure may persist in humans for 

years (Cameron 1981, 1983).

Although studies on the acute toxicity of disparlure 

have been conducted in laboratory animals, the lack of 

either subchronic or chronic toxicity data precludes a 

quantitative characterization of risk.

Groups at Special Risk.

The toxic effects of disparlure, if any, have not been 

identified. Consequently, groups at special risk cannot 

be characterized.

Risk to Nontarget Organisms 

(Issue 2).

Mammals.

Results of acute toxicity studies for oral, dermal, ocular, 

and inhalation exposure to disparlure demonstrate very 

low toxicity to mammals. Information is not available 

regarding chronic toxicity, and no field studies exist 

assessing the impact of disparlure on mammals.

Birds.

There is no evidence that birds are affected by USDA 

treatment projects using disparlure.

Terrestrial Invertebrates.

Disparlure does not attract any other insect found in 

North America.

Fish.

Limited data are available regarding the toxicity of 

disparlure to aquatic animals. A major issue in the 

interpretation of the aquatic toxicity data on disparlure 

involves the solubility of disparlure in water. While 

no measured values are available, estimates based on 

quantitative structure-activity relationships developed 

by the U.S. EPA suggest that the solubility of disparlure 

in water is in the range of 0.0019 to 0.0028 mg/L 

(Appendix H). No risks to fish can be identified under 

foreseeable circumstances in the use of disparlure 

formulations.

Aquatic Invertebrates.

As with fish, disparlure does not appear to pose a risk 

to aquatic invertebrates due to inherent toxicity. At the 

limit of the solubility of disparlure in water, there is no 

indication that toxic effects are likely in any aquatic 

species (Appendix H). Based on the variability in the 

experimental data as well as the range of application 

rates used in USDA programs, HQs would vary from 

about 0.15 to about 0.37 below the level of concern 

by factors of about 3 to 10. This risk characterization 

applies to accidental application of disparlure to a 

body of water 1 meter deep. The HQ will vary with 

the depth of the water. Since the calculations are 

based on a 1-meter-deep body of standing water, the 

HQ would be a factor of 10 lower in a 10-meter-deep 

body of standing water and a factor of 10 higher in 

a 0.1-meter-deep body of standing water. In actual 

field applications using Disrupt II flakes, water bodies 

such as lakes and rivers are never directly treated with 

flakes, and levels of exposure in moving water would 

be magnitudes lower than the calculated static level, 

providing an even greater margin between exposure 

and potential toxicity. Further, control tests using the 

untreated carrier products (small plastic flakes) showed 

no toxicity.

In summary, the application of disparlure in mating 

disruption is unlikely to affect aquatic invertebrates.

Cumulative Effects of Disparlure.

Risk to Human Health (Issue 1).

Since disparlure seems to persist in humans, repeated 

exposures of disparlure will attract the gypsy moth. No 

information is available on the interaction of disparlure 

with other control agents or other chemicals usually 

found in the environment.
Environmental Consequences

Chapter 4 - Page 19

Risk to Nontarget Organisms (Issue 2).

Since disparlure attracts only the gypsy moth in 

North America, no cumulative effects are expected on 

nontarget organisms.

4.7 Consequences of 

Dichlorvos (as Used in Mass 

Trapping) (Alternatives 1, 2, 3).

See Appendixes K and M for detailed analysis of risks 

associated with the use of dichlorvos in mass trapping. 

Appendix A provides an in-depth discussion of how 

dichlorvos is used in mass trapping. Dichlorvos is 

not a distinct treatment in the USDA gypsy moth 

management program. It is simply an insecticide 

(formulated in a vinyl strip as a killing agent) used 

in the large-capacity milk carton trap, which can be 

deployed for mass trapping of male gypsy moths in 

a project area. This same kind of milk carton traps 

(with dichlorvos) are also used in gypsy moth surveys. 

Without this insecticide in the traps, the male gypsy 

moths that are attracted to traps (by disparlure) would 

simply fly back out.

Milk carton traps with dichlorvos have not been 

used for mass trapping since 1997 and only twice 

between 1993 and 1997, where no more then 50 acres 

were treated. Each year for surveys APHIS deploys 

approximately 19,000 milk carton traps with dichlorvos 

pest strips. The Forest Service’s slow-the-spread 

strategy also uses milk carton traps for surveys.

General Effects of Dichlorvos.

Because dichlorvos is used inside traps, no effect on 

human health and nontarget organisms is expected. A 

person or animal would have to deliberately eat the 

resin strip. In the entire history of USDA use of traps 

containing dichlorvos, such an accidental or deliberate 

action has not been encountered.

Risk to Human Health (Issue 1).

General.

Dichlorvos is readily absorbed into the body of 

mammals via all routes of exposure, and is rapidly 

metabolized and eliminated. Generally, the systemic 

effects observed after oral, inhalation, or dermal 

exposure of humans or laboratory animals to dichlorvos 

result from the inhibition of acetylcholinesterase 

(AChE). The enclosed nature of milk carton traps 

containing dichlorvos minimizes the chance that people 

will come into contact with it. In a risk assessment of 

the carcinogenic and mutagenic potential of diclorvos, 

EPA decided “The carcinogenicity potential of 

Diclorovos has been classified as ‘suggestive’ under 

the 1999 Draft Agency Cancer Guidelines and no 

quantitative assessment of cancer risk is required.” 

(Section 3.1.10 of Appendix K).

Exposure of both workers and members of the general 

public should be negligible in most cases. Workers 

taking prudent steps to limit both dermal and inhalation 

exposures can minimize the likelihood of exposure to 

dichlorvos. Similarly, exposure of the general public 

to substantial amounts of dichlorvos is unlikely. The 

dichlorvos is contained within a PVC strip to ensure 

the active ingredient is released slowly over time. The 

strip, in turn, is placed within a trap and the trap is 

placed so that it will not be accessed except in the case 

of intentional tampering or trap monitoring.

The greatest risks for workers are associated with 

inhalation exposures from assembling the traps in 

enclosed and poorly ventilated spaces, or while 

transporting the traps in the passenger compartments 

of vehicles. These risks are readily avoided. Dermal 

exposures are usually at lower levels than inhalation 

exposures.

All of the exposure scenarios for members of the 

general public described in Appendix K are accidental. 

Should a child come into contact with a dichlorvos 

strip, both dermal and oral exposures (if a child ate the 
Chapter 4

Chapter 4 - Page 20

strip) could substantially exceed a level of concern. 

See Appendix K for additional dichlorvos information 

and risk assessment scenarios.

Groups at Special Risk.

Children are of primary concern as identified in the risk 

assessment (Appendix K). As noted above, imprudent 

handling of a dichlorvos-impregnated strip would 

most likely involve a child. Additionally, very young 

children (infants less than 6 months old) may be at 

special risk because of their incompletely developed 

AChE systems and immature livers (ATSDR 1993).

Several other groups may be at special risk to all 

cholinesterase-inhibiting compounds, including 

dichlorvos. A small proportion of the population 

has an atypical variant of plasma cholinesterase that 

may make them more susceptible to effects when 

exposed to dichlorvos and other AChE inhibitors. 

Other groups known to have low plasma AChE levels 

are long-distance runners, women in early stages of 

pregnancy, women using birth control pills, individuals 

with advanced liver disease, alcoholics, individuals 

with poor nutritional status, and individuals with 

skin diseases. Asthmatics may also be at special 

risk because dichlorvos may induce or exacerbate 

respiratory distress (ATSDR 1993).

Risk to Nontarget Organisms 

(Issue 2).

Exposure would be accidental since dichlorvos is used 

inside traps.

Mammals.

The principal adverse effects of dichlorvos exposure 

are directly related to inhibition of cholinesterase. In 

the USDA program for the control of the gypsy moth, 

the use of milk carton traps employing slow release of 

dichlorvos from PVC strips essentially precludes rapid 

exposures to high doses of dichlorvos.

Birds.

No published data is available concerning the acute 

toxicity of dichlorvos encased in PVC resin to birds.

Terrestrial Invertebrates.

The only terrestrial invertebrates likely to come 

into close contact with the dichlorvos strip are male 

gypsy moths attracted by the disparlure in the trap, or 

carnivorous wasps and hornets that may enter the trap 

to feed on dead and dying gypsy moths.

Fish.

There is no indication fish are likely to be adversely 

affected by dichlorvos as used in PVC strips (Section 

4.4.3.1, Appendix K). However, dichlorvos itself 

is classified as highly toxic to both freshwater and 

estuarine fish (U.S. EPA 1999a). See Appendix K for 

comprehensive information.

Aquatic Invertebrates.

Based on the same conservative exposure assessment 

used for both fish and terrestrial vertebrates, some 

sensitive aquatic invertebrates could be adversely 

affected by dichlorvos contamination of water if a 

trap is intentionally thrown into water. As in the 

other exposure assessments developed in Appendix K 

involving contaminated water, this exposure scenario 

should be regarded as an extremely rare accident rather 

than routine. Under normal circumstances, water 

contamination from dichlorvos strips is negligible and 

consistent with the conclusions reached by U.S. EPA 

(1999a).

Cumulative Effects of Dichlorvos.

Risk to Human Health (Issue 1).

The only substantial exposures to the general public 

would occur from repeated tampering with traps 

containing dichlorvos. No such incidents have been 

reported, despite the long use of dichlorvos in traps for 

the gypsy moth and other species.
Environmental Consequences

Chapter 4 - Page 21

Workers may be exposed repeatedly to dichlorvos if 

they are involved in the assembly and placement of 

traps over a period of several weeks. No data exists 

regarding the effects of exposure to dichlorvos in 

combination with exposure to the other agents used 

to control the gypsy moth or to the gypsy moth itself. 

Inhibition of AChE is the most sensitive effect of 

dichlorvos; this effect is not associated with exposure 

to the other control agents or to the gypsy moth. 

Therefore, there is no plausible basis for assuming that 

the effects of exposure to dichlorvos and any or all of 

the other control agents or the gypsy moth are additive.

Risk to Nontarget Organisms (Issue 2).

Experience with traps used in mass trapping and survey 

programs shows that there are no cumulative effects on 

nontarget organisms even over years of use.

4.8 Consequences of Gypchek 

(Alternatives 1, 2, 3).

See Appendixes G and M for detailed analysis of risks 

associated with Gypchek.

General Effects of Gypchek.

Gypchek may indirectly help to maintain existing 

forest condition, water quality, microclimate, and 

soil condition by delaying increases in gypsy moth 

populations, thereby protecting tree foliage.

Risk to Human Health (Issue 1).

General.

According to Appendix G, there is no plausible risk to 

either workers or members of the general public from 

the use of Gypchek to control the gypsy moth.

Groups at Special Risk.

No groups at special risk are identified. Some 

individuals may be allergic to gypsy moth parts found 

in Gypchek.

Risk to Nontarget Organisms 

(Issue 2).

Mammals.

Except for eye irritation, there is little indication that 

NPV or the Gypchek formulation of NPV has any 

effect in mammals, even at extremely high levels of 

exposure. One study specifically focused on wildlife 

conducted by Lautenschlager and others (1977), 

exposed mice, short-tailed shrews, and opossums to 

various forms of NPV (gypsy moth larvae infected 

with NPV, a purified formulation of NPV, and a spray 

preparation of NPV). Based on gross observations, 

as well as necropsy and microscopic examination of 

several different tissues, no effects were seen in any of 

the species.

Birds.

Few studies are available on birds, and the results 

of these studies are essentially identical to those on 

mammals. The studies indicate exposures to NPV at 

levels that are substantially higher than those likely 

to occur in the environment are not associated with 

any adverse effects (Podgwaite and Galipeau 1978, 

Lautenschlager and others 1976).

Terrestrial Invertebrates.

Barber and others (1993) found no indication that NPV 

is pathogenic to any insect species except the gypsy 

moth. No adverse effects were observed in any species 

tested. Additionally, a recent field study noted no 

effects in nontarget insects following the application 

of Gypchek (Rastall and others 2003). There is no 

indication that adverse effects are caused in nontarget 

insects at any level of exposure.
Chapter 4

Chapter 4 - Page 22

Fish.

Two studies are available on the toxicity of NPV to 

fish (Moore 1977, Kreutzweiser and others 1997). The 

results of both studies show no toxicity in rainbow 

trout, no effects on mortality, behavior, or growth rate, 

and no viable NPV detected in the stomach or intestinal 

tract.

Aquatic Invertebrates.

No effects on mortality or reproduction were observed 

over exposure periods of up to 4 weeks (Streams 1976).

Cumulative Effects of Gypchek.

Risk to Human Health (Issue 1).

Exposure to both the gypsy moth caterpillars and 

Gypchek could be additive, although there are no data 

showing this occurs and Gypchek treatments would 

eliminate the caterpillars.

Risk to Nontarget Organisms (Issue 2).

Since Gypchek is specific to the gypsy moth, no 

cumulative effects are expected for nontarget 

organisms.

4.9 Consequences of 

Tebufenozide (Alternatives 

2 and 3).

See Appendixes J and M for detailed analysis of risks 

associated with tebufenozide.

The use of tebufenozide to manage the gypsy moth 

may adversely affect nontarget Lepidoptera. There is 

little indication that humans or other wildlife species 

will be adversely affected under normal conditions of 

use, even at the highest application rate (see the full 

analysis of tebufenozide in Appendix J). Table 4-2

provides hazard quotients (HQ) for tebufenozide and 

the other treatments and gypsy moth.

General Effects of Tebufenozide.

Tebufenozide may indirectly help to maintain existing 

forest conditions, microclimate, and soil condition 

by delaying increases in gypsy moth populations, 

thereby protecting tree foliage. Although tebufenozide 

is not highly mobile in soil, it may be transported 

by percolation, sedimentation, or runoff from soil to 

ambient water. Tebufenozide would not be sprayed 

over water or areas where surface water is present, and 

buffers will be maintained around these areas. See 

Appendix J for additional information on tebufenozide 

and water quality.

Risk to Human Health (Issue 1).

General.

A relatively detailed and consistent series of studies 

in mice, rats, and dogs indicates that the primary 

mechanism of tebufenozide toxicity in mammals 

involves effects on the blood, specifically the formation 

of methemoglobin. Tebufenozide does not appear to 

be carcinogenic and does not appear to cause birth 

defects. Nonetheless, the compound is associated with 

adverse reproductive effects in experimental mammals. 

Tebufenozide itself does not seem to be irritating to the 

skin or eyes. As discussed in the exposure assessment 

in Appendix J, dermal absorption is the primary route 

of exposure for workers. Data regarding the dermal 

absorption kinetics of tebufenozide are not available 

in the published or unpublished literature. Potential 

inhalation toxicity of the compound is not of substantial 

concern in the risk assessment in Appendix J.

At the maximum application rate, two applications at 

0.12 lb (pounds) a.i./acre spaced 3 days apart, there is 

little indication that adverse effects on human health 

are likely. The risk assessment at Appendix J suggests, 

however; that two applications at 0.08 lb a.i./acre or 

more should be avoided in areas where members of the 

general public might consume contaminated fruits or 

other contaminated vegetation.
Environmental Consequences

Chapter 4 - Page 23

Groups at Special Risk.

Individuals born with a form of congenital 

methemoglobinemia may be at increased risk of 

adverse effects to compounds like tebufenozide that 

induce methemoglobinemia (Centa and others 1985, 

Das Gupta and others 1980). Some infants with 

an intolerance to cow’s milk or soy protein exhibit 

methemoglobinemia. Infants less than 3 months old 

have lower levels of methemoglobin (cytochrome b5) 

reductase and higher levels of methemoglobin (1.32 

percent) in comparison with older children or adults 

(Centa and others 1985, Smith 1996). A similar pattern 

is seen in many species of mammals (Lo and Agar 

1986).

Risk to Nontarget Organisms 

(Issue 2).

Under normal conditions of use at the highest 

anticipated application rate, no effects are expected in 

any group of organisms: vertebrates, invertebrates, or 

plants.

Mammals.

everal standard toxicity studies in experimental 

mammals were conducted as part of the registration 

process for tebufenozide. The most sensitive effect 

in several species of experimental mammals involves 

effects on the blood, specifically the formation of 

methemoglobin.

The acute toxicity of tebufenozide is relatively low, 

with an oral LD50 greater than 5,000 mg/kg. The 

subchronic and chronic toxicity studies on tebufenozide 

were conducted in dogs, mice, and rats, with the 

most sensitive effects involving changes to the blood. 

There is no apparent dose-duration relationship for 

tebufenozide; short-term exposures are likely to lead 

to changes in the blood comparable to those observed 

following longer-term exposures (Appendix J).

Birds.

Toxicity studies have been conducted on the acute 

toxicity and reproductive effects of tebufenozide in 

birds, and a field study is available on reproductive 

effects. The acute toxicity of tebufenozide is low for 

birds (Appendix J).

Reproduction studies were conducted in mallard 

ducks (Beavers and others 1993a) and bobwhite 

quail (Beavers and others 1993b, Reinert 1995a). 

Dietary concentrations less than or equal to 1,000 

ppm tebufenozide did not cause reproductive effects 

in mallard ducks. In the quail studies results are 

inconsistent. In a study by Beavers and others (1993b), 

reproductive effects included reduced number of 

eggs laid, viable embryos and 14-day-old survivors at 

dietary concentrations of 300 and 1,000 ppm, but not 

at 100 ppm. A similar study yielded no substantial 

dose-related effects in quail exposed to dietary 

concentrations of up to 615 ppm (Reinert 1995a).

A field study on the reproductive performance of 

Tennessee warblers (Vermivora peregrina) in forests 

treated with tebufenozide has been published (Holmes 

1998). In this study, tebufenozide was applied twice 

at a rate of approximately 0.06 lb a.i./acre with a 4-

day interval between applications, in a forest area in 

Ontario, Canada. Reproductive parameters assayed 

included number of eggs laid, percent hatch, and 

growth of the hatchlings as compared with an untreated 

control plot. A total of six nests were observed in 

the control plot, and five nests were treated with 

tebufenozide in the test plots, with no statistically 

significant adverse effects noted. However, there 

were decreases in both the average number of eggs 

per nest (6.3 in the control area and 5.8 in the treated 

area) as well as the percent hatch (97.4 percent in 

the control area and 89.7 percent in the treated area). 

The small sample sizes result in a low statistical 

power, and the results are “suggestive, although not 

necessarily compelling, that reproductive parameters 

were consistently lower in the treated blocks than 

in the control block” (Holmes 1998, p. 191). Some 
Chapter 4

Chapter 4 - Page 24

differences in adult behavior were observed in the 

plot treated with tebufenozide, such as an increase in 

foraging time and an associated decrease in brooding 

time. This suggests that the primary effect on the birds 

may have been a decrease in food abundance.

This field study by Holmes (1998) combined with the 

bobwhite quail assay conducted by Beavers and others 

(1993b) raise concern that tebufenozide could cause 

adverse reproductive effects in birds. This concern 

is addressed quantitatively in the risk assessment in 

Appendix J for exposures involving the consumption of 

contaminated vegetation, fish, and insects.

Terrestrial Invertebrates.

While tebufenozide will be specifically used by the 

USDA Forest Service for the control of the gypsy 

moth, tebufenozide is effective in controlling other 

pest species, including the apple bud moth (Platynota 

idaeusalis) (Biddinger and others 1998), various 

species of spruce budworm (Payne and others 1997; 

Retnakaran and others 1997a, b), and the Indian-meal 

moth (Plodia interpunctella) (Oberlander and others 

1998). A complete list of the pest species for which 

tebufenozide is specified is provided in U.S. EPA 

(1999e).

The toxicity of tebufenozide has been assayed in 

several species, and the mechanism of action of 

tebufenozide in target insects is relatively well 

understood. Tebufenozide mimics the action of the 

invertebrate hormone 20-hydroxyecdysone, which 

controls molting in insects and various terrestrial and 

aquatic invertebrates by binding to species-specific 

ecdysone receptors present in the cytoplasm of 

epidermal cells (Addison 1996, Keller 1998, Smagghe 

and Degheele 1994a, U.S. EPA 1999e).

While 20-hydroxyecdysone is a hormone common to 

many invertebrates, the effectiveness of tebufenozide in 

mimicking 20-hydroxyecdysone activity seems to vary 

among orders and species of invertebrates. Although 

the specificity of tebufenozide is not addressed in 

detail in the recent U.S. EPA (1999e) ecological 

risk assessment, it was reviewed in detail by Rohm 

and Haas (Keller 1998). That review is consistent 

with publications in the open literature relating to 

species specificity of tebufenozide (Addison 1996; 

Biddinger and Hull 1995; Biddinger and others 1998; 

Brown 1996; Butler and others 1997; Dhadialla and 

others 1998; Rumpf and others 1998; Smagghe and 

others 1996; Valentine and others 1996). In general, 

Lepidoptera are sensitive to tebufenozide, but other 

insects are much less sensitive (Smagghe and Degheele 

1994a). The differing levels of sensitivity appear to 

be related to differences in ecdysone receptor binding 

(Smagghe and others 1996) rather than differences in 

pharmacokinetics (Smagghe and Degheele 1994b).

There are four studies regarding the effects of 

tebufenozide on terrestrial invertebrates under field or 

field-simulation conditions. Three of these studies are 

published (Addison 1996, Butler and others 1997b, 

Valentine and others 1996), and one is an unpublished 

study conducted by Rohm and Haas (Walgenbach 

1995). The studies by Addison (1996) and Butler and 

others (1997b) are most directly relevant to the risk 

assessment in Appendix J, because they assayed the 

effects on nontarget invertebrates in the forest canopy 

(Butler and others 1997b) and forest soil (Addison 

1996) after the application of tebufenozide.

In the study by Addison (1996), tebufenozide was 

incorporated into forest soil at a concentration of 72.1 

ppm. Based on a typical application rate of 70 g/ha 

and the assumption that tebufenozide will remain in 

the top 2 cm of soil, Addison (1996) estimated that 

the soil concentration of 72.1 ppm is equivalent to a 

concentration that is 100 times greater than expected 

environmental concentrations. There were no adverse 

effects on one species of earthworm (Dendrobaena 

octaedra) or on four species of Collembola (Folsomia 

candida, Folsomia nivalis, Onychiurus parvicornis, 

and Hypogastrura pannosa), which are indigenous to 

forest soils in Canada and the northern United States. 
Environmental Consequences

Chapter 4 - Page 25

Consistent with results of the Addison (1996) study, 

a standard bioassay on earthworms (Eisenia foetida) 

noted no adverse effects at soil concentrations of up 

to 1,000 ppm over a 14-day exposure period (Garvey 

1992).

Butler and others (1997b) conducted a study on canopy 

arthropods in which tebufenozide was applied at 

rates of 0.03 and 0.06 lb a.i./acre to a mixed oak plot 

in Ohio. The investigators examined the efficacy of 

tebufenozide against gypsy moth larvae and its effects 

on nontarget arthropods. Population assays included 

measures of abundance and diversity in 10 arthropod 

families and 15 lepidopteran species. A decrease in 

abundance was noted in some lepidopteran species, 

while no effects on abundance or richness were noted 

in any organisms other than lepidopteran species.

The studies by Valentine and others (1996) and 

Walgenbach (1995) involve the application of 

tebufenozide formulations to apple orchards. 

Tebufenozide had no effects on species of mites, 

spiders, various beetles (Coleoptera), and true bugs 

(Hemiptera), after being applied to apple orchards at 

rates effective in controlling lepidopteran pest species 

(Valentine and others 1996). Similarly, Walgenbach 

(1995) noted no effects on beneficial insect populations. 

These two studies support the general conclusion that 

tebufenozide is likely to have an adverse impact on 

Lepidoptera, but not on nonlepidopteran species.

Fish.

Information on the toxicity of tebufenozide to fish 

is summarized in Appendix J. All of the available 

studies were conducted in support of the registration of 

tebufenozide and submitted to the U.S. EPA.

The acute toxicity of tebufenozide to fish is relatively 

low with LC50 values of 3.0 mg a.i./L in bluegill 

sunfish (Graves and Smith 1992b) and 5.7 mg a.i./L 

in rainbow trout (Graves and Smith 1992c). There 

is greater concern, however, regarding the potential 

chronic toxicity of tebufenozide to fish. The U.S. 

EPA evaluates all studies like those summarized in 

Appendix J to determine whether the conclusions are 

consistent with the data, and in many instances the U.S. 

EPA accepts the study conclusions. For tebufenozide, 

however, the U.S. EPA has disagreed with conclusions 

for a fathead minnow egg and fry study as well 

as a fathead minnow full life cycle study. This 

disagreement is discussed further in the dose-response 

assessment (section 4.3.3.1 of Appendix J).

Aquatic Invertebrates.

Unpublished studies on the toxicity of tebufenozide 

to aquatic invertebrates submitted to the U.S. EPA 

in support of the registration of tebufenozide are 

summarized in Appendix J. Some invertebrate assays 

were conducted in support of the registration of 

tebufenozide, and the summaries of these studies in 

Appendix J are based on a review of the full text copies 

of the studies submitted to the U.S. EPA. Additional 

studies published in the open literature are discussed 

below. Unlike some of the fish studies, the studies on 

aquatic invertebrates, summarized in Appendix J, were 

accepted without exception by the U.S. EPA (1999e).

In the studies submitted for registration, the acute 

toxicity of tebufenozide to Daphnia (Crustacea) and 

midges (Insecta) is on the same order as that for fish, 

with a 48-hour LC50 value of 3.8 mg/L for daphnids 

(Graves and Smith 1992a) and a 96-hour LC50 value 

of 0.3 mg/L for midge larvae (van der Kolk 1997). 

Similarly, a study published in the open literature and 

sponsored by the U.S. Geological Survey, reported 

higher LC50 values for Crustacea (17.37 mg/L for 

Daphnia and 5.53 mg/L for Artemia) than for two 

species of mosquitoes (0.92 mg/L for Aedes aegypti

and 0.15 mg/L for Aedes taeniorhynchus) (Song and 

others 1997).

Kreutzweiser and Thomas (1995) assayed the effects 

of tebufenozide on aquatic invertebrate communities in 

lake enclosures. A dose-related decrease in cladoceran 
Chapter 4

Chapter 4 - Page 26

abundance was noted, persisting for 1-2 months at the 

two lower concentrations and for 12-13 months at the 

two higher concentrations. The decrease in cladoceran 

abundance was accompanied by an increase in the 

abundance of rotifers, suggesting that the changes in 

community structure could be attributable to secondary 

or trophic effects rather than to toxicity.

Rohm and Haas summarized the results of several 

field studies or field simulation studies (Kreutzweiser 

and others 1994, 1995) regarding the effects of 

tebufenozide to aquatic invertebrates (Keller 1998).

Cumulative Effects of Tebufenozide.

Risk to Human Health (Issue 1).

Tebufenozide and diflubenzuron could have a 

cumulative effect on methemoglobinemia. USDA 

gypsy moth management programs do not use 

these two chemicals together in the same area at the 

same time; however, tebufenozide might be applied 

to the same area in multiple years for eradication 

projects. These multiple applications of tebufenozide 

over a period of time may increase the potential 

risk of methemoglobinemia. Exposure to other 

methemoglobinemia-inducing compounds in the 

environment may contribute to a cumulative effect. 

For example, individuals exposed to combustion smoke 

or carbon monoxide (agents causing oxidative damage 

to blood) in addition to exposure to tebufenozide may 

be at increased risk of developing methemoglobinemia. 

Individuals exposed to high levels of nitrates, either 

in air or in water, demonstrate increased levels of 

methemoglobin and may be at increased risk with 

exposure to compounds such as tebufenozide.

Risk to Nontarget Organisms (Issue 2).

Tebufenozide generally would not be used in 

conjunction with other treatments; however, it might 

be applied to the same area in multiple years for 

eradication projects. Generally these areas are small 

(usually no more than 5,000 acres). Tebufenozide 

might have a cumulative effect on nontarget caterpillars 

of moths and butterflies by potentially reducing 

their populations, but it will not affect other aquatic 

and terrestrial species as used in USDA gypsy moth 

treatment projects.

4.10 Consequences of Adding 

a New Treatment Under 

Alternative 3.

At this time a prediction can not be made as to what 

new treatments might become available in the future 

for the gypsy moth. Given the protocol built into Alternative 3 (see Chapter 2), the effects and cumulative 

effects associated with any treatment(s) would pose no 

greater risk to human health and nontarget organisms 

than are disclosed in this draft SEIS for the currently 

approved treatments and for tebufenozide.

4.11 Summary of Effects 

Including Cumulative Effects.

Risk to Human Health (Issue 1).

General.

During a gypsy moth outbreak, people are exposed to 

large numbers of gypsy moths and experience skin and 

eye irritation and respiratory system effects, sometimes 

to the extent that they may seek medical treatment. 

Although both B.t.k. and Gypchek may also cause these 

effects, these irritations most likely will be less intense 

than irritations from a gypsy moth outbreak. No other 

human health effects are plausible for Gypchek; for 

disparlure, no human health risks could be identified, 

the only effect being the nuisance of male moths 

attracted to people working with traps that contain the 

female gypsy moth sex pheromone, disparlure.
Environmental Consequences

Chapter 4 - Page 27

No human health effects are likely from exposure to 

diflubenzuron and tebufenozide at application rates 

used in USDA gypsy moth projects. With very high 

exposures, increases in methemoglobin, an abnormal 

blood pigment that reduces the oxygen-carrying 

capacity of the blood, might be detectable for both 

insecticides. Should high application rates (0.12 lbs/

acre in two applications 3 days apart) of tebufenozide 

be used, ingestion of tebufenozide becomes a concern 

(for example, on contaminated fruit; the upper range 

for the HQ of 1.5 is for long-term consumption of fruit; 

see Table 4-2). Applications at these high levels are not 

likely to occur for USDA projects.

The risk posed by dichlorvos is greatest for people 

who might tamper with traps and receive high levels of 

dermal exposure, or who might ingest the insecticide 

strip contained in the trap (Table 4-2). The upper range 

of the HQ of 380 depicts a child ingesting a dichlorvos 

strip; this scenario has never been encountered in 

USDA projects.

Cumulative Effects.

Repeated defoliation over successive years by gypsy 

moth caterpillars increases the potential exposure 

and subsequent skin, eye, and respiratory reactions. 

All of the treatments would reduce this risk over 

time. Diflubenzuron and tebufenozide both evoke 

the formation of methemoglobin; however, these 

treatments would not be utilized at the same time in 

the same area. Improper handling of dichlorvos poses 

a cumulative risk to workers, especially if ventilation 

is inadequate and proper handling procedures are not 

followed.

Risk to Nontarget Organisms 

(Issue 2).

General.

Other than effects on trees, current data and literature 

on the gypsy moth reveal only minor effects on other 

terrestrial and aquatic organisms; studies were in many 

cases of short duration and evaluated only a segment of 

the ecosystem or only a few species. There is a general 

lack of long-term, multi-year studies measuring over 

decades the impact of the gypsy moth on ecosystems 

and terrestrial and aquatic species and systems. 

This deficiency of extended studies may mask and 

underestimate the long-term impacts of gypsy moth 

on terrestrial and aquatic systems. Gypchek, mass 

trapping (dichlorvos), and disparlure have no long- or 

short-term effects on nontarget terrestrial species (all 

hazard quotients are less than 0.01, see Table 4-2). 

Gypchek and dichlorvos in USDA treatment projects 

do not affect aquatic nontarget organisms. The highest 

calculated disparlure hazard quotient in any aquatic 

organism is 0.37 (some small aquatic invertebrates). 

Under normal conditions of USDA gypsy moth 

management projects, disparlure is not expected to 

impact aquatic organisms.

B.t.k. applications impact certain spring-feeding 

butterflies and moths. Many lepidopteran species are 

not affected, especially those not present in the treated 

foliage and species arriving in treatment areas after the 

B.t.k. has disappeared from the foliage.

Compared with any of the other treatments, 

diflubenzuron affects a greater variety of terrestrial 

and aquatic nontarget species: moths and butterflies, 

grasshoppers, parasitic wasps, aquatic insects, bottomdwelling crustaceans, and immature free-floating 

crustaceans (Table 4-2).

Tebufenozide affects only Lepidopterans, having no 

other expected significant effect on other terrestrial 

species or aquatic invertebrates (Table 4-2). There is 

no expectation that tebufenozide would be used at the 

highest application rates in USDA treatment projects; 

as a result the hazard quotient derived from a mammal 

eating contaminated fruit is likely to be lower than 1.5 

(Table 4-2).
Chapter 4

Chapter 4 - Page 28

Cumulative Effects.

Repeated spraying with B.t.k., diflubenzuron, or 

tebufenozide is likely to decrease lepidopteran species 

populations if the same areas are sprayed over 2 or 

more years. An expected result of cumulative impact 

on sensitive lepidopteran species from repeated annual 

spraying with any of these treatments is reasonable, as 

is the expectation that repeated annual spraying with 

diflubenzuron would have a cumulative impact on 

aquatic organisms if this insecticide reached aquatic 

ecosystems.

4.12 Operational Flexibility of 

Treatments.

For example, in order to minimize possible effects 

on threatened and endangered species that may be 

present in areas proposed for treatment, Gypchek, mass 

trapping, and mating disruption (where appropriate) 

could be selected instead of using B.t.k., diflubenzuron, 

or tebufenozide.

Tebufenozide (Alternative 2) provides the USDA 

gypsy moth management program with an additional 

treatment option that may prove useful for reducing the 

threat posed by gypsy moth outbreaks. Alternative 3 

affords the greatest flexibility to the USDA gypsy moth 

management program.

4.13 Unavoidable Adverse 

Effects.

Since this draft SEIS is programmatic in nature, no 

unavoidable adverse effects were identified for any of 

the alternatives. Any adverse effects that might occur 

would be identified and addressed in environmental 

analyses at the site-specific project level.

4.14 Short-Term Uses and 

Long-Term Productivity.

The National Environmental Policy Act (NEPA) 

requires consideration of “the relationship between 

short-term uses of man’s environment and the maintenance and enhancement of long-term productivity” [42 

U.S.C. 4322 (2)(C)]. As declared by the Congress, this 

relationship includes using all practicable means and 

measures, including financial and technical assistance, 

in a manner calculated to foster and promote the general welfare, to create and maintain conditions under 

which man and nature can exist in productive harmony, 

and fulfill the social, economic, and other requirements 

of present and future generations of Americans (NEPA 

Section 101).

The gypsy moth threatens the forest resources in 

the United States both in the short and long term, as 

described in Section 4.3 and in Appendix L. Each 

alternative provides treatments to lessen and delay the 

impacts of the gypsy moth on these forest resources. 

Alternative 2 provides an additional treatment and 

increased operational flexibility for gypsy moth treatment projects. Alternative 3 provides the greatest operational flexibility for gypsy moth treatment projects. 

Although the treatments may have short-term effects as 

outlined in Sections 4.4 – 4.9 and Table 4-2, no longterm effect could be identified--except for B.t.k. where 

sensitive spring lepidopteran species may take longer to 

recover. Mitigation measures at the site-specific project level will reduce the short- and long-term impacts 

of the treatments for each of the alternatives.

4.15 Measures to Mitigate 

Adverse Environmental 

Impacts.

Given the variety of places and circumstances where 

gypsy moth projects could be implemented, it will be 

necessary to develop and implement specific mitigation 

measures for each project. Mitigation measures will be 

developed and implemented on a site-specific basis for 

each project based on local conditions and concerns. 

See Chapter 2 for mitigation measures.
Environmental Consequences

Chapter 4 - Page 29

4.16 Urban Quality, Historic 

and Cultural Resources, 

and Design of the Built 

Environment.

In-depth, site-specific project environmental analyses 

will be performed for individual projects, as this draft 

SEIS is programmatic in nature.

4.17 Energy Requirements and 

Conservation Potential

of Various Alternatives.

All of the alternatives involve energy use, primarily 

aviation fuel used by aircraft and helicopters for 

treatment application. Designing spray blocks for 

efficiency reduces flight time and conserves fuel.

4.18 Natural or Depleted 

Resource Requirements and 

Conservation Potential of 

Various Alternatives.

All alternatives reduce the impact of the gypsy moth on 

forest resources in protecting forests from gypsy moth 

outbreaks that may cause tree mortality. Other than the 

use of air space over treatment areas, with the shortterm impacts of aviation noise and limitation of public 

use during application, no inherent natural or cultural 

resource requirements exist for the three alternatives. 

Impacting factors for specific projects will be addressed 

with site-specific environmental analyses.

4.19 Irreversible and 

Irretrievable Commitments of 

Resources.

Irreversible and irretrievable commitments of resources 

due to the presence of the gypsy moth, defoliation, and 

specific treatments occur at the project level and are 

disclosed through site-specific analyses.

4.20 Other Required 

Disclosures.

NEPA at 40 CFR 1502.25(a) directs “to the fullest extent possible, agencies shall prepare draft environmental impact statements concurrently with and integrated 

with ... other environmental review laws and executive 

orders.”

Because this draft SEIS is programmatic in nature, the 

Forest Service and APHIS will ensure that site-specific 

consultations will be done as necessary at the project 

level for the Endangered Species Act (ESA), the National Historic Preservation Act (NHPA), and any other 

laws, regulations, executive orders, and agency policies 

that apply.

Chapter 5

Preparers and

Contributors

Figure 5-1. Civilian Conservation Corps workers traveled by truck to perform 

gypsy moth field work. 
Chapter 5 Preparers and Contributors

Contents

5.1 Preparers..................................................................................................1

Interdisciplinary Team......................................................................1

Human Health and Ecological Risk Assessment Contractor............2

5.2 Contributors............................................................................................2

Management Group...........................................................................2

Individuals........................................................................................2

5.3 Information Management and Analysis Staff.........................................4

Figure

Figure 5-1. Civilian Conservation Corps workers traveled by truck to 

perform gypsy moth field work..............................................................Cover
Preparers and Contributors

Chapter 5 - Page 

Individuals listed as preparers were responsible for 

developing the content of this document. Contributors 

shared information and expertise. Those named under 

Information Management assembled the document, 

posted material on the Web, and managed supporting 

information. 

5.1 Preparers.

Interdisciplinary Team

Joseph L. Cook Supervisory Entomologist and SEIS 

Team Leader, USDA Forest Service, Northeastern 

Area State and Private Forestry, Newtown Square, 

PA. 

Three years with the Forest Service and a total of 

32 years of government service to various agencies 

and military branches, including the Navy, Army 

and Marine Corps. Positions held in the fields of 

natural resources, fisheries, marine biology, forestry, 

pest management, entomology, wildlife biology, 

cultural resources management, environmental 

management and planning. Participated in 

National Environmental Policy Act document 

preparation, implementation and administration at 

the local, regional and national level in a variety of 

assignments both in the United States and overseas. 

Served in the U.S. Army (active and reserve) as a 

Medical Entomologist. Academic degrees include 

a B.S. in Natural Resources from the University 

of Michigan in 1970, M.S. in Entomology (Forest 

Entomology) from the University of Minnesota in 

1996 and M.B.A. from University of the Incarnate 

Word in 1991. 

Weyman Fussell SEIS Team Co-Leader, USDA 

APHIS Plant Protection and Quarantine, Riverdale, 

MD.

Eight years with the USDA, including 3 years with 

APHIS Pest Detection and Management Programs 

as Gypsy Moth program manager working to 

address phytosanitation issues domestically and 

internationally, focusing on Latin America. Prior 

to joining the USDA, taught at the university 

level for 5 years and spent 15 years in overseas 

programs addressing food production planning and 

implementation. Academic degrees include M.S. 

from Purdue University in Crop Genetics with a 

minor in Plant Pathology, and a Ph.D. from the 

University of Tennessee in 1983 in Agricultural 

Genetics with a minor in Economics of International 

Development.

Derek Handley Public Affairs Specialist, USDA 

Forest Service, Northeastern Area State and Private 

Forestry, Newtown Square, PA. (No longer with 

Forest Service)

Two years with the Forest Service, 11 years with the 

U.S. Navy and U.S. Navy Reserves. Responsibilities 

included community relations, speech writing, 

and media relations. Earned B.A. in English from 

Hampton University in 1994.

William Oldland Entomologist, USDA Forest Service, 

Northeastern Area State and Private Forestry, 

Newtown Square, PA. 

Three years as an Entomologist with the USDA 

Forest Service and 8 years with the Tennessee 

Valley Authority (TVA) as contract and Federal 

employee. Served as TVA’s Medical Entomologist 

for 2 of these 8 years; responsibilities included 

monitoring mosquito populations for West Nile 

Virus, malaria, several strains of encephalitis (EEE, 

SLE, LAC) and writing the vector chapters for 

the TVA EIS/Reservoir Operations Study. While 

serving as an Environmental Scientist for the TVA, 

his duties included sample collection, evaluation 

and compliance report composition for TVA Power 

Plants. While a contract Entomologist at the TVA, he 

assisted in the management of a biological control 

program for hydrilla and purple loosestrife. Bill 
Chapter 5

Chapter 5 - Page 

also spent 2 years in private industry as a Wildlife 

Biologist/Forester. He earned a B.S. in Wildlife 

Management in 1991 and M.S. in Entomology in 

1993 from West Virginia University.

Mary Ann White Writer-Editor, USDA Forest Service, 

Northeastern Area State and Private Forestry, 

Newtown Square, PA. (No longer with Forest 

Service)

Two years with the Forest Service, 6 years on 

active duty in the U.S. Navy. Received A.S. in 

Medical Laboratory Technology from The George 

Washington University in 1979 and B.A (1984) and 

M.A.(1986) degrees in history from the University 

of Texas at El Paso.

Human Health and Ecological Risk 

Assessment Contractor 

Syracuse Environmental Research Associates, Inc., 

Fayetteville, NY

5.2 Contributors.

Contributors provided information and expertise.

Management Group 

USDA Forest Service advisory group on national gypsy 

moth policy.

Robert D. Mangold USDA Forest Service, Director, 

Forest Health Protection, Washington, DC

Jerry Boughton USDA Forest Service, Assistant 

Director Forest Health and Economics Programs, 

Northeastern Area State and Private Forestry, 

Newtown Square 

Individuals 

USDA Forest Service, USDA Animal and Plant Health 

Inspection Service (APHIS) and other USDA contacts 

provided assistance to the interdisciplinary team with 

their time, materials, critical review skills and support. 

Debra Allen-Reid USDA Forest Service, Southwestern 

Region, Albuquerque, NM

John Anhold USDA Forest Service, Southwestern 

Region, Flagstaff, AZ

Hank Appleton USDA Forest Service, Forest Health 

Protection, Washington, DC

David Bakke USDA Forest Service, Pacific Southwest 

Region, Vallejo, CA

David A. Bergsten USDA APHIS, Riverdale, MD

Beverly M. Bulaon USDA Forest Service, Northern 

Region, Missoula, MT

David R. Bridgewater USDA Forest Service, Pacific 

Northwest Region, Portland, OR

Robert Cain USDA Forest Service, Rocky Mountain 

Region, Lakewood, CO

Joseph Carbone USDA Forest Service, Ecosystem 

Management Coordination, Washington, DC

William A. Carothers USDA Forest Service, Southern 

Region, Asheville, NC

Michael D. Connor USDA Forest Service, 

Northeastern Area State and Private Forestry, St. 

Paul, MN

Jesus A. Cota USDA Forest Service, Forest Health 

Protection, Washington, DC (retired)

Frank J. Cross USDA Forest Service, Rocky 

Mountain Region, Denver, CO (retired)

Meredith Dahl USDA, Office of the General Counsel, 

Washington, DC
Preparers and Contributors

Chapter 5 - Page 

John William Dale USDA Forest Service, Pacific 

Southwest Region, Vallejo, CA (retired)

Jack P. Edmundson USDA APHIS, Riverdale, MD

John H. Ghent USDA Forest Service, Southern 

Region, Asheville, NC

Kurt W. Gottschalk USDA Forest Service, Northern 

Research Station, Morgantown, WV

John W. Hazel USDA Forest Service, Northeastern 

Area State and Private Forestry, Morgantown, WV

Donna S. Leonard USDA Forest Service, Southern 

Region, Asheville, NC

Andrew M. Liebhold USDA Forest Service, Northern 

Research Station, Morgantown, WV

Jesse Logan USDA Forest Service, Rocky Mountain 

Research Station, Logan, UT

Leonard L. Lucero USDA Forest Service, 

Southwestern Region, Albuquerque, NM

Michael L. McManus USDA Forest Service, Northern 

Research Station, Hamden, CT

Victor C. Mastro USDA APHIS, Otis Plant Protection 

Center, Otis ANGB, MA 

Paul A. Mistretta USDA Forest Service, Southern 

Region, Atlanta, GA

Wesley A. Nettleton USDA Forest Service, Southern 

Region, Atlanta, GA

Amy H. Onken USDA Forest Service, Northeastern 

Area State and Private Forestry, Morgantown, WV

Doug Parker USDA Forest Service, Southwestern 

Region, Albuquerque, NM (no longer with Forest 

Service) 

John D. Podgwaite USDA Forest Service, Northern 

Research Station, Hamden, CT

Bernard J. Raimo USDA Forest Service, Northeastern 

Area State and Private Forestry, Durham, NH

Richard C. Reardon USDA Forest Service, Forest 

Health Technology Enterprise Team, Morgantown, 

WV

Leslie Rubin USDA APHIS, Riverdale, MD

Dwight Scarbrough USDA Forest Service, 

Intermountain Region, Boise, ID

Noel F. Schneeberger USDA Forest Service, 

Northeastern Area State and Private Forestry, 

Newtown Square, PA

Dave E. Schultz USDA Forest Service, Pacific 

Southwest Region, Redding, CA (deceased)

Mark E. Schultz USDA Forest Service, Alaska 

Region, Southeast Alaska Field Office, Juneau, AK

David Sire USDA Forest Service, Ecosystem 

Management Coordination, Washington, DC

James M. Slavicek USDA Forest Service, Northern 

Research Station, Delaware, OH

Rhonda R. Solomon USDA APHIS, Riverdale, MD

Dennis J. Souto USDA Forest Service, Northeastern 

Area State and Private Forestry, Durham, NH

Harold Thistle USDA Forest Service, Forest Health 

Technology Enterprise Team, Morgantown, WV
Chapter 5

Chapter 5 - Page 

Kevin Thorpe USDA Agricultural Research Service, 

Beltsville, MD

Patrick C. Tobin USDA Forest Service, Northern 

Research Station, Morgantown, WV

Kathryn Toffenetti USDA, Office of the General 

Counsel, Washington, DC

Daniel B. Twardus USDA Forest Service, 

Northeastern Area State and Private Forestry, 

Morgantown, WV

Algimantas P. Valaitis USDA Forest Service, Northern 

Research Station, Delaware, OH

Ralph Webb USDA, Agricultural Research Service, 

Beltsville, MD

5.3 Information Management 

and Analysis Staff.

The Information Management and Analysis Staff of 

the USDA Forest Service, Northeastern Area State 

and Private Forestry, provided valuable assistance in 

preparation and printing of the draft and final reports, 

Internet site management, and computer operations 

support. 

Cindy Barnett Morgantown, WV

Roberta Burzynski Newtown Square, PA

Helen Butalla Morgantown, WV

Patty Dougherty Newtown Square, PA

Victoria Evans Morgantown, WV

Sandy Fosbroke Morgantown, WV

Nancy Lough Morgantown, WV

Juliette Watts Newtown Square, PA
Chapter 6

Mailing List

Figure 6-1. Early aerial gypsy moth treatments were manually released. 
Chapter 6 Mailing List

Contents

6.1 Federal Agencies.....................................................................................1

6.2 State and Local Agencies........................................................................6

6.3 American Indian Nations, Tribes, and Related Agencies......................24

6.4 Organizations........................................................................................32

6.5 Libraries................................................................................................57

6.6 Individuals............................................................................................73

Figure

Figure 6-1. Early aerial gypsy moth treatments were 

manually released...................................................................................Cover
Mailing List

Chapter 6 - Page 

This chapter lists agencies, organizations, libraries, and 

individuals who were mailed complete copies of the 

draft supplemental environmental impact statement or 

notified of its availability on the Web.

6.1 Federal Agencies

Alabama

United States Department of Agriculture (USDA), 

Forest Service, National Forests in Alabama 

(National Forests & Ranger Districts)

USDA, Animal Plant Health & Inspection Service 

(APHIS), Plant Protection & Quarantine (PPQ)

Alaska

United States Department of Interior (USDI), Bureau 

of Indian Affairs (BIA) Field Offices

National Marine Fisheries Service

USDA, APHIS, PPQ

USDA, Forest Service, Alaska Region

USDA, Forest Service, National Forests in Alaska 

(National Forests & Ranger Districts)

Arizona

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Arizona 

(National Forests & Ranger Districts)

USDI, BIA Field Offices

Arkansas

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Arkansas 

(National Forests & Ranger Districts)

California

USDI, BIA Field Offices

USDI, Bureau of Land Management (BLM)

United States Environmental Protection Agency 

(USEPA), Region IX

Federal Aviation Administration, Western Region

National Marine Fisheries Service

U.S. Army Corps of Engineers, South Pacific Division 

(CESPD)

USDA, APHIS, PPQ

USDA, Forest Service, Region

USDA, Forest Service, National Forests in California 

(National Forests & Ranger Districts)

Colorado

USDI, BIA, Field Offices

United States Environmental Protection Agency 

(USEPA), Region VIII

USDA, Forest Service, Comanche National Grassland

USDI, National Park Service (NPS), Intermountain 

Regional Office

USDA, Forest Service, Pawnee National Grassland

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Colorado 

(National Forests & Ranger Districts)

USDA, Forest Service, Rocky Mountain Region

Connecticut

USDA, Forest Service, National Forests in Connecticut 

(National Forests & Ranger Districts)

USDA, Forest Service, Northern Research Station

Delaware

USDI, U.S. Fish and Wildlife Service (USFWS), 

Bombay Hook National Wildlife Refuge (NWR)

USDA, APHIS, PPQ

District of Columbia

Advisory Council on Historic Preservation

Department of Defense (DOD), Armed Forces Pest 

Management Board
Chapter 6

Chapter 6 - Page 

DOD, U.S. Navy, Office of Chief of Navy Operations

Department of Energy (DOE), Office of Environmental 

Compliance Rural Energy Regulatory Commission, 

Environmental Compliance Branch

Department of Housing & Urban Development (HUD), 

Office of Environment & Energy

DOI, USFWS, Division of Endangered Species

USEPA, Office of Environmental Affairs

USEPA, Office of Federal Activities

General Services Administration, Office of Planning & 

Analysis (OPA)

Rural Electrification Administration, Environmental 

Compliance

Susquehanna River Basin Commission

U.S. Army Corps of Engineers

U.S. Coast Guard, Environmental Impact Branch

U.S. Department of the Interior, Office of 

Environmental Affairs

U.S. Department of Transportation, Environmental 

Division

USDA OPA Publication Stockroom

USDA, Forest Service

USDA, Forest Service, Office of Environmental 

Coordination

Florida

USDI, BIA, Seminole Agency

National Marine Fisheries Service

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Florida 

(National Forests & Ranger Districts)

Georgia

USEPA, Region IV

U.S. Army Corps of Engineers, South Atlantic Division 

(CESAD)

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Georgia 

(National Forests & Ranger Districts)

USDA, Forest Service, Southern Region

Hawaii

U.S. Army Corps of Engineers, Pacific Ocean Division 

(CEPOD)

USDA, APHIS, PPQ

Idaho

USDA, APHIS, PPQ

USDI, NPS, Craters of the Moon National Monument

USDA, Forest Service, National Forests in Idaho 

(National Forests & Ranger Districts)

Illinois

USDI, USFWS, Crab Orchard NWR

USDI, USFWS, Cypress Creek NWR

USEPA, Region V

Federal Aviation Administration, Great Lakes Region

USDI, NPS, Lincoln Home National Historical Site

USDI, USFWS, Mark Twain NWR

U.S. Army Corps of Engineers, North Central Division 

(CENCD)

USDA, Forest Service, National Forests in Illinois 

(National Forests &Ranger Districts)

USDA, APHIS, PPQ

Indiana

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Indiana 

(National Forests & Ranger Districts)

Iowa

USDA, APHIS, PPQ

USDI, NPS, Effigy Mounds National Monument

USDI, NPS, Herbert Hoover National Historic Site

USDI, USFWS, Union Slough NWR

USDA, Forest Service, National Forests in Iowa 

(National Forests & Ranger Districts)
Mailing List

Chapter 6 - Page 

Kansas

USDA, Forest Service, Cimarron National Grasslands

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Kansas 

(National Forests & Ranger Districts)

Kentucky

Ohio River Basins Commission

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Kentucky 

(National Forests & Ranger Districts)

Louisiana

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Louisiana 

(National Forests & Ranger Districts)

Maine

USDI, NPS, Acadia National Park

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Maine 

(National Forests & Ranger Districts)

Maryland

USDI, NPS, Antietam National Battlefield

USDI, NPS, Blackwater National Wildlife Refuge

USDI, USFWS, Eastern Neck NWR

USDI, USFWS, Patuxent Wildlife Resource Center

USDA, Agriculture Research Service (ARS), Insect 

Biocontrol Laboratory

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Maryland 

(National Forests & Ranger Districts)

USDA, National Agricultural Library

Massachusetts

USEPA, Region I

Federal Aviation Administration, New England Region

National Marine Fisheries Service

U.S. Army Corps of Engineers, New England Division 

(CENED)

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in 

Massachusetts (National Forests & Ranger 

Districts)

Michigan

USDI, BIA Field Offices

USDI, NPS, Isle Royal National Park

USDI, USFWS, Shiawassee NWR

USDI, NPS, Sleeping Bear Dunes National Lakeshore

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Michigan 

(National Forests & Ranger Districts)

Minnesota

USDI, USFWS, Agassiz NWR

USDI, BIA Field Offices

USDI, NPS, Grand Portage National Monument

USDI, USFWS, Pipestone National Monument

USDI, USFWS, Rice Lake NWR

USDI, USFWS, Tamarac NWR

USDI, USFWS, Upper Mississippi River NWR

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Minnesota 

(National Forests & Ranger Districts)

USDI, NPS, Voyagers National Park

Mississippi

USDI, BIA Field Offices

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Mississippi 

(National Forests & Ranger Districts)

Missouri

USDI, USFWS, Clarence Cannon NWR

USEPA, Region VII

Federal Aviation Administration, Central Region

USDI, NPS, Harry S. Truman National Historic Site

USDI, NPS, Ozark National Scenic Riverways

USDI, USFWS, Squaw Creek NWR

USDI, USFWS, Swan Lake NWR

USDA, APHIS, PPQ
Chapter 6

Chapter 6 - Page 

USDA, Forest Service, National Forests in Missouri 

(National Forests & Ranger Districts)

Montana

USDI, BIA Field Offices

USDI, BLM

USDI, USFWS, Bowdoin NWR

USDI, NPS, Glacier National Park

USDI, NPS, National Bison Range

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Montana 

(National Forests & Ranger Districts)

Nebraska

USDI, BIA Field Offices

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Nebraska 

(National Forests & Ranger Districts)

Nevada

USDI, BIA Field Offices

USDI, BLM

USDI, NPS, Great Basin National Park

USDI, NPS, Spring Mountains National Recreational 

Area

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Nevada 

(National Forests & Ranger Districts)

New Hampshire

USDI, NPS, St. Glaudens National Historic Park

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in New 

Hampshire (National Forests & Ranger Districts)

New Jersey

Delaware River Basins Commission

USDI, NPS, Delaware Water Gap National Recreation 

Area

USDI, USFWS, Great Swamp NWR

USDI, NPS, Morristown National Historic Site

USDI, NPS, Shady Hook Gateway National Recreation 

Area

USDA, APHIS, PPQ

New Mexico

USDI, BIA Field Offices

USDI, NPS, National Park Service Intermountain 

Support Office

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in New 

Mexico (National Forests & Ranger Districts)

New York

USDI, BIA, Field Offices

USEPA, Region II

Federal Aviation Administration, Eastern Region

Federal Highway Administration, Region I

USDI, NPS, Fire Island National Seashore

USDI, USFWS, Montezuma National NWR

USDI, USFWS, Morton National Wildlife Refuge

USDI, NPS, Roosevelt-Vanderbilt National Historic 

Site

U.S. Army Corps of Engineers, North Atlantic

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in New York 

(National Forests & Ranger Districts)

North Carolina

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in NC 

(National Forests & Ranger Districts)

North Dakota

USDI, BIA Field Offices

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in North 

Dakota (National Forests & Ranger Districts)

Ohio

USDI, USFWS, Cedar Point NWR

USDI, USFWS, Ottawa NWR
Mailing List

Chapter 6 - Page 

U.S. Army Corps of Engineers, Ohio River Division 

(CEORD)

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Ohio 

(National Forests & Ranger Districts)

Oklahoma

USDI, BIA, Field Offices

USDA, Forest Service, Black Kettle National 

Grasslands

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Oklahoma 

(National Forests & Ranger Districts)

Oregon

USDI, BIA Field Offices

USDI, BLM

USDI, NPS, Columbia Gorge National Scenic Area

USDI, NPS, Columbia River National Scenic Area

USDI, NPS, Hells Canyon National Recreation Area

Northwest Power Planning Council

USDI, NPS, Oregon Dunes National Recreation Area

U.S. Army Corps of Engineers, North Pacific Division

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Oregon 

(National Forests & Ranger Districts)

Pennsylvania

USDI, NPS, Delaware Gap National Recreation Area

USEPA, Region III

USDI, USFWS, Erie NWR

USDI, USFWS, Lamar NWR

USDA Forest Service Northeastern Research Station

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in 

Pennsylvania (National Forests & Ranger Districts)

USDI, NPS, Valley Forge National Historical Site

Rhode Island

USDA, APHIS, PPQ

South Carolina

Rural Development Administration, Region IIISoutheast

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in South 

Carolina (National Forests & Ranger Districts)

South Dakota

USDI, BIA Field Offices

USDA, Forest Service, Fall River Rd/Wall Road/

Buffalo Gap National Grasslands

USDA, Forest Service, Fort Pierre National Grassland

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in South 

Dakota (National Forests & Ranger Districts)

Tennessee

USDI, NPS, Great Smokey Mountains National Park

Tennessee Valley Authority (TVA), Environmental 

Quality

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Tennessee 

(National Forests & Ranger Districts)

Texas

USDA, BIA, Field Offices

USDA, Forest Service, Caddo LBJ National Grasslands

USEPA, Region VI

Federal Aviation Administration, Southwest Region

U.S. Army Corps of Engineers, Southwestern Division

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Texas 

(National Forests & Ranger Districts)

Utah

USDI, NPS, Arches National Park

USDI, BIA Field Offices

USDI, BLM

USDI, NPS, Bryce Canyon National Park

USDI, NPS, Canyonlands National Park

USDI, NPS, Capital Reef National Park

USDI, NPS, Cedar Breaks National Monument

USDI, NPS, Flaming Gorge National Recreation Area
Chapter 6

Chapter 6 - Page 

USDI, NPS, Natural Bridges National Monument

USDI, NPS, Timpanogos Cave National Monument

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Utah 

(National Forests & Ranger Districts)

USDI, NPS, Zion National Park

Vermont

USDA, APHIS, PPQ

Virginia

USDI, NPS, George Washington Memorial Pkwy

USDI, NPS, Mount Rogers National Recreation Area

USDI, NPS, Shenandoah National Park

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Virginia 

(National Forests & Ranger Districts)

Washington

USDI, BIA Field Offices

USEPA, Region X

Federal Aviation Administration, Northwest Region

USDI, NPS, Mt. Rainier National Park

USDI, NPS, North Cascades National Park

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Washington 

(National Forests & Ranger Districts)

West Virginia

DoD-Army Blueston Lake, US Army Corps Of 

Engineers

USDI, NPS, Harpers Ferry National History Park

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in West 

Virginia (National Forests & Ranger Districts)

Wisconsin

USDI, NPS, Apostle Islands National Lakeshore

USDI, BIA, Field Offices

USDI, USFWS, Necedah NWR

USDI, NPS, St. Croix National Scenic Riverway

USDI, USFWS, Trempealeah NWR

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Wisconsin 

(National Forests & Ranger Districts)

Wyoming

USDI, BIA Field Offices

USDI, BLM

USDI, NPS, Grand Teton National Park

USDA, APHIS, PPQ

USDA, Forest Service, National Forests in Wyoming 

(National Forests & Ranger Districts)

USDI, NPS, Yellowstone National Park

Puerto Rico

USDA, Forest Service, National Forests in Puerto Rico 

(National Forests & Ranger Districts)

6.2 State and Local Agencies

Alabama

Alabama A&M University, Forestry Sciences 

Laboratory

Cooperative Extension Service 

Department of Agriculture & Industries

Forestry Commission

Alaska

Cooperative Extension Service 

Department of Environmental Conservation

Department of Fish & Game

Department of Natural Resources

Division of Forestry

Project Analyst- Forests Alaska State Agencies

Arizona

Department of Agriculture

Department of Agriculture Plant Services Division

Game & Fish Department

State Land Department

State Parks
Mailing List

Chapter 6 - Page 

Arkansas

Cooperative Extension Service 

Forestry Commission

State Plant Board

State Plant Pest Board Division of Plant Industry

California

Department of Environmental Protection

Department of Fish & Game

Department of Food & Agriculture

Department of Forestry & Fire Protection

Department of Parks & Recreation

Department of Water Resources

Redwood Sciences Laboratory

State Lands Commission

Colorado

Department of Agriculture

Department of Natural Resources

State Conservation Board

State Forest Service

Connecticut

Bureau of Natural Resources

Department of Environmental Protection

Division of Forestry

Middlesex County Cooperative Extension Service 

State Forest Tree Nursery

University of Connecticut Cooperative Extension 

Service 

University of Connecticut Storrs Agriculture 

Experiment Station

West Hartford Cooperative Extension Service 

Windham County Cooperative Extension Service 

Delaware

Delaware Department of Agriculture

Department of Natural Resources & Environment

Division of Fish & Wildlife

Forest Service

Kent County Cooperative Extension Service 

New Castle County Cooperative Extension Service 

Sussex County Cooperative Extension Service 

University of Delaware CES

District of Columbia

DC Government

Florida

Cooperative Extension Service (CES)

Department of Agriculture & Consumer Services 

Division of Plant Industry

Department of Environmental Protection

Division of Forestry

Fish & Wildlife Conservation Commission

Pike County CES 

Georgia

Department of Agriculture Plant Protection Division

Department of Natural Resources

Environmental Council

Environmental Policy Institute

Forestry Commission

Hawaii

Department of Agriculture Plant Industry Division

Department of Land & Natural Resources

Division of Forestry & Wildlife

Idaho

Department of Environmental Quality

Department of Fish & Game

Department of Lands Northern Operations

Department of Lands Southern Operations

Department of Parks & Recreation

Department of Agriculture Division of Plant Industry

Department of Lands

Idaho County Weed Supervisor

Northwest Watershed Research Center

Water Resources Department

Illinois

Adams County CES 

Boone County CES 
Chapter 6

Chapter 6 - Page 

Brown County CES 

Bureau County CES 

Caroll County CES 

Champaign CES 

Charleston CES 

Chicago Park District

Christian County CES

Clark County CES

Clay County CES

Clinton County CES

Cook County CES

Crawford County CES

Cumberland County CES

Dekalb County CES

Department of Agriculture

Department of Conservation

Department of Natural Resources

Department of Public Health

Department of Conservation Forbes State Fish & 

Wildlife Area

Division of Forest Resources

Douglas County CES

Dupage County CES

Edgar County CES

Edwards County CES

Effingham County CES

Environmental Council

Fayette County CES

Ford County CES

Forest Research Center

Franklin County CES

Fulton County CES

Gallatin County CES

Greene County CES

Grundy County CES

Hamilton County CES

Hancock County CES

Hardin County CES

Henderson County CES

Henry County CES

Iroquois County CES

Jasper County CES

Jefferson County CES

Jerseyville CES

Jo Daviess County CES

Johnson/Massac Counties CES

Jones County CES

Kane County CES

Kankakee County CES

Kendall County CES

Knox County CES

Lasalle County CES

Lawrence County CES

Lee County CES

Livingston County CES

Logan County CES

Macon County CES

Macon County, Soil & Water

Macoupin County CES

Madison County CES

Marion County CES

Marion Extension Center

Marshall/Putnam Counties CES

Mason County CES

Mcdonough County CES

Mchenry County CES

Mclean County CES

Menard County CES

Mercer County CES

Monroe County CES

Montgomery County CES

Morgan County CES

Moultrie County CES

Natural History Survey

Nature Preserves Commission

Ogle County CES

Oquawka CES

Peoria County CES

Peorial County CES

Perry County CES

Pittsfield CES

Pope/Hardin Counties CES

Randolph County CES

Region 1 CES

Region 2 CES

Region 4 CES
Mailing List

Chapter 6 - Page 

Region 6 CES

Region 7 CES

Richland County CES

Rock Island County CES

Saline County CES

Sangamon County CES

Scott County CES

Shelby County CES

St. Clair County CES

Stephenson County CES

Tazewell County CES

Union County CES

University of Illinois CES

Urbana CES

Urbana, Department of Forestry

Vermilion County CES

Wabash County CES

Warren County CES

Washington County CES

Wayne County CES

White County CES

Whiteside County CES

Will County CES

Williamson County CES

Winnebago Count CES

Woodford County CES

Indiana

Adams County CES

Allen County CES

Bartholomew County CES

Benton County CES

Boone County CES

Brown County CES

Carroll County CES

Cass County CES

Clark County CES

Clay County CES

Crawford County CES

Davless County CES

Dearborne County CES

Decator County CES

Delaware County CES

Department of Conservation

Department of Natural Resources

Department of Health

Division of Entomology & Plant Pathology

Division of Fish & Wildlife

Division of Forestry

Division of Forestry Department of Natural Resources

Dubois County CES

Elkhart County CES

Fayette County CES

Floyd County CES

Fountain County CES

Franklin County CES

Fulton County CES

Gibson County CES

Grant County CES

Green County CES

Hamilton County CES

Hancock County CES

Harrison County CES

Hendricks County CES

Henry County CES

Howard County CES

Huntington County CES

Jackson County CES

Jasper County CES

Jay County CES

Jefferson County CES

Jennings County CES

Johnson County CES

Knox County CES

Kosciusko County CES

Lagrange County CES

Lake County CES

Laporte County CES

Lawrence County CES

Madison County CES

Marion County CES

Marshall County CES

Martin County CES

Miami County CES

Monroe County CES

Montgomery County CES
Chapter 6

Chapter 6 - Page 10

Morgan County CES

Newton County CES

Noble County CES

Owen County CES

Parke County CES

Perry County CES

Porter County CES

Posey County CES

Pulaski County CES

Purdue University, State Extension Forester

Putnam County CES

Randolph County CES

Ripley County CES

Rush County CES

Scott County CES

Shelby County CES

Starke County CES

Steuben County CES

Sullivan County CES

Switzerland County CES

Tippecano County CES

Tipton County CES

Union County CES

Vanderburgh County CES

Vermillion County CES

Vigo County CES

Wabash County CES

Warren County CES

Warrick County CES

Washington County CES

Wayne County CES

Wells County CES

White County CES

Whitley County CES

Iowa

Adair County CES

Adams County CES

Allamake County CES

Audubon County CES

Bloomfield County CES

Boone County CES

Bremer County CES

Buchanan County CES

Buena Vista County Vista CES

Butler County CES

Calhoun County CES

Carroll County CES

Cedar County CES

Cerro Gordo County CES

Cherokee County CES

Chickasa County CES

Clark County CES

Clay County CES

Clayton County CES

Clinton County CES

Council Bluffs CES

Dallas CES

Decatur CES

Delaware CES

Department of Agriculture State Horticulturist

Department of Forestry

Department of Natural Resources

Department of Natural Resources State Forest Nursery

Department of Agriculture

Des Moines County CES

Dubuque County CES

Emmet County CES

Fayette County CES

Floyd County CES

Franklin County CES

Freemont County CES

Greene County CES

Grundy County CES

Guthrie County CES

Hamilton County CES

Hamilton County Conservation Board

Hancock CES

Hardin County CES

Harrison County CES

Henry County CES

Howard County CES

Humboldt County CES

Ida Grove County CES

Iowa County CES

Iowa Department of Agriculture
Mailing List

Chapter 6 - Page 11

Iowa Department of Public Health

Iowa State University CES

Jackson County CES

Jasper County CES

Jones County CES

Keokuk County CES

Kossuth County CES

Lee County CES

Loess Hills State Park

Louisa County CES

Lucas County CES

Lyon County CES

Madison County CES

Mahaska County CES

Marion County CES

Marshall County CES

McFarland Park

Melvern CES

Mitchell County CES

Monona County CES

Monroe County CES

Montgomery County CES

Muscatin County CES

Oakland CES

O’Brien County CES

Osceola County CES

Palo Alto County CES

Plymouth County CES

Pocahontas County CES

Polk County CES

Ringold County CES

Sac County CES

Scott CES

Shimek State Forest

Sigourney CES

Sioux CES

Spirit Lake CES

State Entomologist

State Forest Nursery

Stephens State Forest

Story County CES

Taylor County CES

Toledo CES

Union County CES

Urbana CES

Van Buren County CES

Vinton CES

Wapello County CES

Washington County CES

Wayne County CES

Webster County CES

Winnebago County CES

Winneshi CES

Worth County CES

Wright County CES

Yellow River State Forest

Kansas

Department of Agriculture Plant Prot. & Weed Control 

Program

Department of Wildlife & Parks

Forest Service

Kentucky

CES

Department of Fish & Wildlife Resources

Director, Kentucky Division of Forestry

Natural Resources & Environmental

Office of State Entomologist

Louisiana

Agcenter, Louisiana CES

Department of Agriculture & Forestry

Department of Natural Resources

Department of Wildlife & Fisheries

Maine

Department of Agriculture

Division of Plant Industry, Department of Agriculture

Forest Service

University of Maine CES

Maryland

Allegany County CES

Ann Arundel CES
Chapter 6

Chapter 6 - Page 12

Baltimore County CES

Calvert County CES

Caroline County CES

Carroll County CES

Charles County CES

Department of Agriculture

Department of Agriculture Forest Pest Management

Department of Agriculture, Entomology Nursery 

Inspection

Department of Natural Resources, Forest Service

Dorchester County CES

Garrett County CES

Gypsy Moth Program, Prince Georges County

Hartford County CES

Howard County CES

Kent County CES

Montgomery County Gypsy Moth Program

Prince Georges County CES

Somerset County CES

Talbot County CES

Wicomico County CES

Worcester County CES

Massachusetts

Association of Conservation Commissions

Barnstable County CES

Berrien CES

Central Massachusetts Extension Center

Pittsfield CES

Department of Conservation

Department of Environmental Management

Department of Environmental Management Region 5 

Headquarters

Department of Food & Agriculture

Division of Fisheries & Wildlife

Division of Forests & Parks

Division of Regulatory Services, Department of 

Agriculture Resources

Division State Parks & Recreation

East Massachusetts Extension Center

Middlesex County CES

Pioneer Valley Extension Center

Plymouth County CES

Southeast Massachusetts Agriculture Center

Michigan

Alabastor Township

Alger & Marquette Soil Conservation District

Alger County CES

Allendale Charter Townships

Alpena County CES

Alpena Township

Antrim County CES

Antrim County Gypsy Moth Program

Arenac County CES

Arenac County Gypsy Moth Coordinator

Barry County CES

Bay County CES

Bay County Gypsy Moth Program

Bay County Gypsy Moth Suppression Program

Benzie CES

Bloomfield Township

Branch CES

Calhoun CES

Camden Township

Cass County CES

Charlevois CES

Charter Township of Highland

Charter Township of West Bloomfield

Cheboygan CES

Cherry Grove Township

Chippewa CES

Chippewa County CES

Clare County CES

Clare County Gypsy Moth Program

Crawford County CES

Crawford County Gypsy Moth Coordinator

Delta County CES

Department of Agriculture

Department of Agriculture Pesticide & Plant Pest 

Management Division

Department of Natural Resources

Department of Natural Resources Forestry Division

Department of Public Health

Dickson Township Manistee County
Mailing List

Chapter 6 - Page 13

Eaton City CES

Emmet County CES

Fenton Township

Frankenmuth Township

Genesee County Gypsy Moth Coordinator

Genessee County CES

Gladwin City CES

Gladwin County Gypsy Moth Coordinator

Gogebic CES

Grand Traverse County CES

Gration County Gypsy Moth Coordinator

Harrisville CES

Hawes Township Trustee

Hillsdale CES

Home Township

Houghton CES

Huron City CES

Iosco County CES

Iosco County Gypsy Moth Coordinator

Iron County CES

Iron Mountain CES

Isabella County

Jackson CES

Kalamazo CES

Kalkaska County Gypsy Moth Coordinator

Kalkaska County Gypsy Moth Program

Kent County CES

Kent County Conservation District

Lake CES

Lake County Commissioners

Lake County Gypsy Moth Program

Lansing CES

Leelanau County CES

Lenawee County CES

Livingston County Gypsy Moth Coordinator

Livingston County Gypsy Moth Program

Luce County CES

Mackinac County CES

Macob City CES

Manistee CES

Manistee County

Manistee County Commission

Manistee County Planning Commission

Manistee Soil & Water Conservation District

Marion Township

Marquette CES

Mason County CES

Mecosta County Gypsy Moth Coordinator

Menomine CES

Michigan Department of Agriculture

Michigan Department of Transportation

Michigan State University Extension

Midland County CES

Midland County Gypsy Moth Program

Midland County Gypsy Moth Suppression Program

Millen Township Supervisor

Mio CES

Missaukee County Gypsy Moth Coordinator

Missaukee County CES

Monroe County CES

Montcalm County CES

Montgomery CES

Newaygo CES

Newaygo County Gypsy Moth Coordinator

Newaygo County Gypsy Moth Suppression Program

Oakland County CES

Oakland County Gypsy Moth Program

Oceana CES

Ogemaw County Gypsy Moth Coordinator

Ontonago CES

Osceola County Gypsy Moth Program

Oscoda County Gypsy Moth Coordinator

Otsego County CES

Ottawa County Gypsy Moth Coordinator

Ottawa County Gypsy Moth Program

Pesticide & Plant Pest Management Division, 

Department of Agriculture

Presque Isle CES

Region 2, Michigan Department of Natural Resources

Rochester Parks Department

Roscommon County

Roscommon County Gypsy Moth Suppression Program

Saginaw County CES

Saginaw County Gypsy Moth Coordinator

Sanilac County CES

Shiawassee County Gypsy Moth Coordinator
Chapter 6

Chapter 6 - Page 14

St. Clair County CES

St. Jose CES

Sweetwater Township

Tuscola County CES

Van Buren CES

Village of Beverly Hills

Wexford CES

Wexford County Gypsy Moth Coordinator

Minnesota

Aitkin County CES

Becker County CES

Beltrami County CES

Benton County CES

Big Stone County CES

Blue Earth County CES

Brown County CES

Brown County Historical Society

Cass County CES

Chippewa County CES

Chisago County CES

Clay County CES

Clearwater County CES

Clearwater County Land Department

Clearwater County Land Forestry Department

Cook County CES

Cottonwood County CES

Crow Wing County CES

Dakota County CES

Department of Agriculture

Department of Agriculture Agronomy & Plant 

Protection

Department of Natural Resources

Department of Natural Resources

Department of Public Health

Division of Forestry

Dodge County CES

Douglas County CES

Faribault County CES

Fergus Falls CES

Goodhue City County CES

Graceville Township

Grant County CES

Hennepin County CES

Houston County CES

Hubbard County CES

Huntsville Township

Isanti County CES

Itasca County Extension

Jackson County CES

Kanabec County CES

Kandiyohi County CES

Kittson County CES

Koochiching CES

Lac Qui Parle CES

Lake County CES

Lake of the Woods CES

Le Seur County CES

Lincoln County CES

Lyon County CES

Mahnomen County CES

Marshall County CES

Martincounty CES

McLoud County CES

Mille Lacs County CES

Morrison County CES

Mower County CES

Murray County CES

Nicollet County CES

Nobles County CES

Norman County CES

Olmsted County CES

Ore County CES

Owatonna CES

Pennington County CES

Perham CES

Pine County CES

Pipestone CES

Polk West County CES

Ramsey County CES

Red Lake City CES

Redwood County CES

Renville County CES

Rice County CES

Rock County CES

Roseau County CES
Mailing List

Chapter 6 - Page 15

St. Louis County Land Department

Sibley County CES

St. Louis CES

Stearns County CES

Stearns County Park Department

Stevens County CES

Swift County CES

Todd County CES

Traverse County CES

University of Minnesota CES

Wabasha County CES

Wadena County CES

Waseca County CES

Washington County CES

Watonwan County CES

Wilkin County CES

Winona County CES

Wright County CES

Yellow Medicine CES

Mississippi

Department of Wildlife, Fisheries

Division of Plant Industry

Forestry Commission

Missouri

Adair County CES

Andrew County CES

Atchison City CES

Audrain County CES

Audrain County CES

Barry County CES

Barton County CES

Bates County CES

Benton County CES

Boone County CES

Buchanan City CES

Butler County CES

Caldwell County CES

Callaway County CES

Camden City CES

Carroll City CES

Carter County CES

Cass County CES

Cedar County CES

Chariton County CES

Christian City CES

Clark County CES

Clinton County CES

Cole County CES

Cooper County CES

Crawford County CES

Dade County CES

Dallas County CES

Daviess County CES

Dekalb County CES

Delta Center County CES

Dent County CES

Department of Agriculture

Department of Conservation

Department of Health

Department of Natural Resources

Douglas County CES

Dunklin County CES

Gasconade County CES

Green County CES

Grundy City CES

Harrison County CES

Hickory City CES

Hickory County CES

Hold County CES

Howard County CES

Howell County CES

Jackson County CES

Jasper County CES

Knox County CES

Laclede City CES

Laclede County CES

Lafayette City CES

Lawrence County CES

Lebanon Excess Property Center

Lewis County CES

Lincoln County CES

Linn County CES

Livingston County CES
Chapter 6

Chapter 6 - Page 16

Macon County CES

Madison County CES

Maries County CES

Marion County CES

Marles County CES

Mercer County CES

Miller City CES

Mississippi County CES

Missouri CES

Moniteau City CES

Monroe County CES

Montgomery County CES

Morgan County CES

Natural History Program

Newton County CES

Nordaway City CES

Oregon City CES

Oregon County CES

Osage City CES

Ozark County CES

Pemiscot City CES

Pettis County CES

Phelps City CES

Phelps County CES

Pike County CES

Plant Industries Division

Platte County CES

Polk County CES

Pulaski CountyCES

Putnam County CES

Randolph County CES

Ralls City CES

Ray County CES

Ripley County CES

Saline County CES

Schuyler County CES

Scotland County CES

Scott County CES

Shelby City CES

St Loius CES

St. Charles County CES

St. Clair County CES

St. Louis City CES

Ste. Genevieve City CES

Stoddard County CES

Sullivan County CES

Taney County CES

Texas County CES

University of Missouri CES

Vernon County CES

Warren County CES

Washington County CES

Webster City CES

Webster County CES

Worth County CES

Wright County CES

Montana

Department of Agriculture

Department of Natural Resources & Conservation

Department of Fish & Wildlife

Forestry Division

Nebraska

Department of Agriculture Bureau of Plant Industry

Department of Environmental Quality

Game & Parks Commission

University of Nebraska CES

Nevada

Department of Agriculture

Department of Agriculture, Entomologist

Division of Forestry

Division of Plant Industry, Department of Agriculture

Division of State Lands

Division of Wildlife

University of Nevada CES

New Hampshire

Belknap County CES

Cheshire County CES

Coos County CES

Department of Agriculture

Department of Resources & Economic Development

Division of Forests & Lands
Mailing List

Chapter 6 - Page 17

Division of Plant Industry

Grafton County CES

Merrimack County CES

Natural Heritage Inventory

New Hampshire Fish & Game Department

Rockingham County CES

Strafford County CES

Sullivan County CES

University of New Hampshire CES

University of New Hampshire CES

New Jersey

Allaire State Park Nature Center

Atlantic County CES

Bergen County CES

Bergen County Park Commission

Boonton Township

Borough of Oakland

Bureau of Forest Management

Burlington County CES

Camden County CES

Cape May County CES

Cumberland County CES

Department of Agriculture

Department of Environmental Protection

Division of Fish & Wildlife

Division of Plant Industry

Essex County CES

Federation of Shade Tree Commission

Forestry Service

Glouchester County CES

Hunterdon County CES

Jersey City CES

Mercer County CES

Monmouth County CES

Monmouth City Park System

Monmouth County Shade Tree Commission

Morris County CES

Morris County Soil Conservation District

Mount Olive Township Council

New Jersey Division of Parks & Forestry

New Jersey Forest Tree Nursery

Ocean County CES

Parvi State Park

Passaic County CES

Pinelands Commission

Salem County CES

Somerset County CES

Somerset County Park Commission

Susses County CES

Union County CES

Warren County Soil Conservation District

Warren CES

New Mexico

Clauch-Pinto Soil & Water Conservation District

Department of Agriculture

Department of Agriculture, Bureau of Entomology & 

Nursery Industries

Department of Game & Fish

Energy Minerals & Natural Resource

Energy, Minerals & Natural Resource Department

Southwest New Mexico Council of Governments

State Forestry

State Land Office

New York

Albany County CES

Allehgany County CES

Broome County CES

Cayuga County CES

Chautaugua County CES

Chenango County CES

Clinton County CES

Clinton County Legislature

Columbia County CES

Cornell CES

Cortland County CES

County of Nassau Department of Recreation & Parks

Delaware County CES

Department Environmental Conservation

Department of Agriculture & Markets Division of Plant 

Industry

Department of Environmental Conservation Forest 

Resources Management
Chapter 6

Chapter 6 - Page 18

Department of Environmental Protection

Department of Health

Dutchess County CES

Essex County CES

Franklin County CES

Fulton County CES

Genesse County CES

Greene County CES

Herkimer County CES

Jefferson County CES

Lewis County CES

Livingston County CES

Monroa County CES

Montgomery County CES

Nassau County CES

New York State CES

Niagra County CES

New York Department of Environmental Conservation 

Urban Forest Coordinator

Oneida County CES

Onondaga County CES

Ontario County CES

Orange County CES

Orange County Department of Parks, Recreation, & 

Conservation

Orleans County CES

Oswego County CES

Otsego County CES

Pocono Forest & Wildlife Service

Putnam County CES

Rensselaer County CES

Rockland County CES

Saratoga County CES

Schenectady CES

Schoharie County CES

Seneca County CES

St. Lawrence County CES

State Department of Environmental Conservation

State Department of Agriculture

Steuben County CES

Suffolk County CES

Sullivan County CES

Tioga County CES

Tompkins County CES

Town of Granville

Ulster County CES

Warren County CES

Washington County CES

Wayne County CES

Westchester County CES

Wyoming County CES

Yates County CES

North Carolina

Carolina Department of Environment & Natural 

Resources

Department of Agriculture & Consumer Services

Department of Agriculture Plant Protection Section

Division of Forest Resources

North Carolina CES

Plant Industry Division

North Dakota

Department of Agriculture

Forest Service

Game & Fish Department

State Historical Society of North Dakota

Ohio

Adams County CES

Allen County CES

Ashland County CES

Ashtabula County CES

Athen County CES

Auglaize County CES

Belmont County CES

Brown County CES

Butler County CES

Champaign County CES

Clark County CES

Clermont Couny CES

Clinton County CES

Colerain Township Parks

Columbiana County CES

Coshocton County CES
Mailing List

Chapter 6 - Page 19

Crawford County CES

Cuyahoga County CES

Darke County CES

Defiance County CES

Deleware County CES

Division of Forestry

Erie County CES

Fairfield County CES

Fayette County CES

Franklin County CES

Fulton County CES

Gallia County CES

Geauga County CES

Geauga Park District

Guernsey County CES

Hamilton County CES

Hamilton County Park District

Hancock County CES

Hardin County CES

Harrison County CES

Henry County CES

Highland County CES

Hocking County CES

Holmes County CES

Huron County CES

Jackson Area Extension Center

Jackson County CES

Jefferson County CES

Knox County CES

Lake County CES

Lawrence County CES

Licking County CES

Logan County CES

Lorain County CES

Lucas County CES

Madison County CES

Mahoning County CES

Marion County CES

Medina County CES

Meigs County CES

Miami County CES

Monroe County CES

Montgomery County CES

Morgan County CES

Morrow County CES

Muskingum County CES

Noble County CES

Ohio Department of Agriculture

Ohio Department of Natural Resources

Ohio State University CES

Ohio State University Extension

Ottawa County CES

Paulding County CES

Perry County CES

Pickaway County CES

Pike County CES

Plant Pest Control Section, Department of Agriculture

Portage County CES

Preble County CES

Richland County CES

Ross County CES

Sandusky County CES

Scloto County CES

Seneca County CES

Shelby County CES

Stark/Summit CES

Trumball County CES

Tuscarawas County CES

Union County CES

Van Wert County CES

Vinton County CES

Warren County CES

Washington County CES

Williams County CES

Wood County CES

Wyandot County CES

Oklahoma

Department of Agriculture Plant Industries & 

Consumer Services Division

Department of Agriculture, Forestry Services

Department of Environmental Quality

Department of Wildlife Conservation

Department of Agriculture

Putnam County CES
Chapter 6

Chapter 6 - Page 20

Oregon

CES Umatilla County

City of Eugene Public Works Department

City of Klamath Falls

Department of Agriculture

Department of Environmental Quality

Department of Fish & Wildlife

Department of Forestry

Department of Environmental Quality

Eugene Parks & Open Space

Fern Ridge Wildlife Area

Grant County & Harney County

Grant County Education Service District

Jackson County CES

Klamath Falls Resource Area

Lane County Environmental Health

Lane County Public Works

Lincoln County Health & Human Services

OSU Agriculture Chemistry Extension

Parks & Recreation Department

Portland Parks & Recreation

Umatilla Basin Watershed Council

Umatilla County Soil & Water

Union County Extension

Wasco County Weed Control

Pennsylvania

Adams County CES

Allegheny County CES

Armstrong County CES

Asbury Woods Nature Center

Beaver County CES

Bedford County CES

Berks County CES

Blair County CES

Bradford County CES

Bucks County CES

Butler County CES

Cambria County CES

Centre County CES

Chester County CES

Clarion County CES

Clearfield County CES

Clinton County CES

Commonwealth of Pennsylvania, Department of 

Environment

Crawford County CES

Cumberland County CES

Dams County CES

Dauphin County CES

Delaware County CES

Department of Conservation & Natural Resources

Elk County CES

Fayette County CES

Forest Pest Management

Four Mills Nature Reserve

Franklin County CES

Fulton County CES

Game Commission

Green County CES

Hawk Mountain Sanctuary

Huntingdon County CES

Indiana County CES

Jefferson County CES

Juniata City CES

Juniata County CES

Lackawanna County CES

Lancaster County CES

Lawrence County CES

Lebanon County CES

Lehigh County CES

Luzerne County CES

Lycoming City CES

Lycoming County CES

Mckean County CES

Mckean County CES

Mifflin County CES

Monroe County CES

Montgomery County CES

Montour County CES

Northhampton County CES

Northhumberland County CES

Pennsylvania Forest Stewardship Program

Perry County CES

Philadelphia County CES
Mailing List

Chapter 6 - Page 21

Pike County CES

Potter County CES

Potter County CES

Schuylkill Conservation District

Schuylkill County CES

Snyder County CES

Stone Valley Recreation Area

Sullivan County CES

Susquehanna County CES

Tioga County CES

Union County CES

Venango County CES

Warren County CES

Washington County CES

Wayne County CES

Wayne County Gypsy Moth Coordinator

Westmoreland County CES

Wyoming County CES

York County CES

Rhode Island

Division of Agriculture & Resource Marketing

Division of Forest Environment

Jennings Nature Reserve

Rhode Island Department of Environmental 

Management

South Carolina

Clemson University CES

Department of Natural Resource

Department of Plant Industry

Forestry Commission

Westvaco Forest Research Center

South Dakota

Department of Agriculture

Department of Environment & Natural Resources

Department of Game Fish & Parks

Tennessee

Department of Agriculture

Department of Agriculture - Division of Forestry

Department of Environment & Conservation

Department of Forestry, Wildlife & Fisheries

Department of Plant Industry

TN Agriculture CES

University of Tennessee Extension

Wildlife Resources Agency

Texas

Department of Agriculture

Forest Service

Parks & Wildlife Department

Utah

Department of Agriculture

Department of Natural Resources

Utah Department of Agriculture & Food Division of 

Plant Industry

Utah State University CES

Vermont

Department of Agriculture

Department of Agriculture Food & Markets Plant 

Industry Section

Department of Forest & Parks

Department of Forest & Parks Forest Resource 

Protection

Guildhall CES

Natural Resources Council

Newport Extension System

North Hero Extension System

Rutland Extension System

Skakel Forest Management

University of Vermont Extension

Vermont Agency of Natural Resources

Vermont Department of Fish & Wildlife

Vermont Extension System

Virginia

Alexandria Gypsy Moth Program

Bedford County CES

Botetourte County CES

CES
Chapter 6

Chapter 6 - Page 22

Clark County CES

Clark County Gypsy Moth Coordinator

Craig County CES

Culpeper County Gypsy Moth Coordinator

Department of Agriculture & Consumer Services

Department of Conservation & Recreation

Department of Forestry

Department of Game & Inland Fisheries

Dinwiddle County CES

Essex County CES

Fairfax County Gypsy Moth Program

Fairfax County Gypsy Moth Office

Flatwoods Civilian Conservation Center

Frederick County Gypsy Moth Program

Fulvanna County CES

Glocester County CES

Greensville County CES

Gypsy Moth Program Coordinator

Hanover County Gypsy Moth Program

Highland County Gypsy Moth Program

Isle of Wright County Gypsy Moth Program

King & Queen County CES

King County CES

King William County CES

Lancaster County CES

Louisa County Gypsy Moth Program

Madison County CES

Mathews County Gypsy Moth Program

Mecklenburg County CES

New Kent County CES

Northhampton County CES

Piedmont Environmental Council

Powhatan County Gypsy Moth Program

Prince Georges County CES

Prince William County Gypsy Moth Program

Prince William Forest Park

Rappahannock County CES

Richmond County CES

Roanoke County CES

Rockingham County Gypsy Moth Program

Rotetourt County CES

Shenandoah County CES

Sky Meadows State Park

Spotsylvania County CES

Surrey County CES

Virginia Polytechnic Institute (VPI) Extension Orange 

County

Warren County Gypsy Moth Program

Westmoreland County CES

York County Gypsy Moth Program

Washington

Alpine Lakes Protect Society

Bellevue Parks/Commission Service

Department of Ecology Office of Water Resources

Department of Fish & Wildlife

Department of Natural Resources

Department of Natural Resources, Northeast

Department of Wildlife Region 1 Habitat

Department of Agriculture Entomology

Evergreen State College

Glacier Public Service Center

Gonzaga University

King County Department of Natural Resources

King/Pierce County Farm Bureau

Natural Heritage Program

Northwest Wilderness Programs

Office of Archaeology & Historic Preservation

Okanogan Conservation District

Oregon Department of Environmental Quality

Seattle City Planning Office

State Department Agriculture

State Department of Agriculture Plant Services 

Division

State Department of Ecology

State of Washington, Department Natural Resources, 

Environmental Quality

Umatilla Forest Watch

Verlot Public Service Center

Washington Department of Natural Resources 

Southeast Region

Washington State Parks

Washington State University

Washington State University Department of Natural 

Resource Sciences
Mailing List

Chapter 6 - Page 23

Washington State University, Extension

Washington Trout

Washington Wildlife Commission

West Virginia

Department of Agriculture

Department of Agriculture Plant Industries Division

Department of Energy

Department of Commerce, Labor & Environmental 

Resources

Division of Forestry

Hardy County Extension Agency

Harrison County CES

Home Health Agency of Davis Memorial Hospital

Plant Industries Program

State Lands Management

State Lands Management Division of Forestry

Upshur County

West Virginia Division of Natural Resources

Wisconsin

Adams County CES

Ashland County CES

Ashland County CES

Barron County CES

Barron County CES

Bayfield County CES

Brown County CES

Buffalo County CES

Burnett County CES

Calumet County CES

Chippewa City CES

Clark County CES

Columbia County CES

Crawford County CES

Dane County CES

Department of Agriculture

Department of Natural Resources

Department of Agriculture

Department of Agriculture & Trade

Devil’s Lake State Park

Division of Forestry

Division of Health

Dodge County CES

Douglas County CES

Dunn City CES

Dunn County CES

Eau Claire County CES

Florence County CES

Fond Du Lac County CES

Forest County Board of Supervisors

Grant CES

Green County CES

Green Lake County CES

Iowa County CES

Iron County CES

Jackson County CES

Jefferson County CES

Juneau County CES

Kenosha County CES

Kewaunee County CES

Lacrosse County CES

Lafayette County CES

Langlade County Board of Supervisors

Langlade County Forestry Department

Layfayette County Board

Lincoln CES

Lincoln County Forestry Land & Park Departmen

Manitowo CES

Manitowoc County Soil & Water Conservation

Marathon County CES

Marinett County CES

Marquette County CES

Menomine County CES

Milwaukee County CES

Mondovi Township

Monroe County CES

Oconto County CES

Oneida County CES

Oneida County Board

Outagamie County CES

Ozaukee County CES

Pepin County CES

Pierce CES

Plant Industry Bureau
Chapter 6

Chapter 6 - Page 24

Polk County CES

Port Washington CES

Portage CES

Price County CES

Racine County CES

Richland CES

Rock County CES

Rock County Park & Conservation Division

Rusk County CES

Sauk CES

Shawano CES

Sheboygan CES

St. Croix CES

Taylor County CES

Town of Brooklyn

Town of Caledonia

Town of Jump River

Town of Knox

Town of Newbold

Town of Rhine

Town of Schoepre

Town of Spring Prairie

Town of Troy

Trempealeau County CES

Trenton Township

University of Wisconsin CES

Vernon County CES

Vilas County CES

Vilas County Community Resource

Walworth CES

Washburn County CES

Washington CES

Waukesha CES

Waupaca County CES

Winnebago CES

Winnebago County CES

Wisconsin Department of Agriculture

Wisconsin Department of Natural Resources

Wood CES

Wood County

Wyoming

Department of Agriculture

Department of Agriculture Technical Services Division

Department of State Parks, & Cultural Resources

Game & Fish Department

State Forestry Division

State Lands & Investments

University of Wyoming CES

Puerto Rico

University of Puerto Rico Agriculture Experiment 

Station

Canada

BC Ministry of Agriculture & Food

Ontario Forestry Association

6.3 American Indian Nations, 

Tribes, and Related Agencies

Alabama

Poarch Creek Indians

Alaska

Atmautluak Traditional Council

Andreafski Tribal Council

Agdaagux Tribe of King Cove

Akiachak Native Community

Akiak Native Community

Akutan Traditional Council

Alatna Tribal Office

Aleut Community of St. Paul Island

Algaaciq Tribal Council

Allakaket Village

Angoon Community Association

Anvik Tribal Council

Asa’carsarmiut Tribe

Beaver Tribal Council

Birch Creek Village

Central Council Tlingit & Haida Indian Tribes

Chalkyitsik Village
Mailing List

Chapter 6 - Page 25

Chefornak Traditional Council

Chenega Council

Chevak Traditional Council

Chickaloon Village Traditional Council

Chignik Lake Traditional Village Council

Chilkat Indian Village (Klukwan)

Chilkoot Indian Association

Chinik Eskimo Community (Aka Golovin)

Chitina Traditional Village

Chuloonawick Native Village

Circle Native Community

Craig Community Association

Crooked Creek Traditional Council

Curyung Tribal Council

Dot Lake Village Council

Douglas Indian Association

Egegik Village

Ekwok Village Council

Emmonak Village

Evansville Tribal Council

Gulkana Village

Healy Lake Traditional Council

Holy Cross Tribal Council

Hoonah Indian Association

Hughes Village Council

Huslia Village Council

Hydaburg Cooperative Association

Igiugig Village

Inupiat Community of Arctic Slope

Iqurmuit Tribe (Russian Mission)

Ivanoff Bay Village Council

Kaguyak Village

Kaktovik Village

Kenaitze Indian Tribe

Ketchikan Indian Corporation

King Island Native Community

Klawock Cooperative Assn

Knik Tribe

Kobuk Traditional Council

Kokhanok Village

Kongiganak Traditional Council

Koyukuk Native Village

Kwigillingok Council

Larsen Bay Tribal Council

Lesnoi Village

Levelock Village

Lime Village

Louden Tribal Council

Manley Hot Springs Village

Manokotak Village

Mary’s Igloo Traditional Council

Mcgrath Native Village Council

Mentasta Traditional Tribal Council

Metlakatla Indian Community

Naknek Native Village Council

Native Village Nuiqsut

Native Village of Afognak

Native Village of Akhiok

Native Village of Aleknagik

Native Village of Ambler

Native Village of Atqasuk

Native Village of Barrow Inupiat Traditional

Native Village of Belkofski

Native Village of Brevig Mission

Native Village of Buckland

Native Village of Cantwell

Native Village of Chignik

Native Village of Chignik Lagoon

Native Village of Chistochina

Native Village of Chuathbaluk

Native Village of Council

Native Village of Deering

Native Village of Diomede (Ira) (Aka Inalik)

Native Village of Eagle

Native Village of Eek

Native Village of Eklutna

Native Village of Ekuk

Native Village of Elim

Native Village of Eyak

Native Village of False Pass

Native Village of Fort Yukonnative

Native Village of Gakona

Native Village of Gambell

Native Village of Georgetown

Native Village of Goodnews Bay

Native Village of Hamilton
Chapter 6

Chapter 6 - Page 26

Native Village of Hooper Bay

Native Village of Kanatak

Native Village of Karluk

Native Village of Kasaan

Native Village of Kasigluk

Native Village of Kiana

Native Village of Kipnuk

Native Village of Kivalina

Native Village of Kluti-Kaah (Aka Copper Cent)

Native Village of Kotzebue

Native Village of Koyuk

Native Village of Kwinhagak

Native Village of Marshall (Aka Fortuna Ledge)

Native Village of Minto

Native Village of Nanwalek (Aka English Bay)

Native Village of Napaimute

Native Village of Napakiak

Native Village of Napaskiak

Native Village of Nikolski

Native Village of Noatak

Native Village of Nunapitchuk

Native Village of Ouzinkie

Native Village of Paimiut

Native Village of Pauloff Harbor

Native Village of Perryville

Native Village of Pitka’s Point

Native Village of Point Hope

Native Village of Port Graham

Native Village of Port Heiden

Native Village of Port Lions

Native Village of Savoonga

Native Village of Scammon Bay

Native Village of Shaktoolik

Native Village of Sheldon Point

Native Village of Shishmaref

Native Village of Shungnak

Native Village of South Naknek

Native Village of St. Michael

Native Village of Stevens

Native Village of Tanacross

Native Village of Tanana

Native Village of Tatitlek

Native Village of Tetlin

Native Village of Tuntutuliak

Native Village of Tununak

Native Village of Tyonek

Native Village of Unalakleet

Native Village of Venetie Tribal Government

Native Village of Wales

Native Village of White Mountain

Nelson Lagoon Tribal Council

Nenana Native Association

New Koliganek Village Council

New Stuyahok Village

Newhalen Tribal Council

Newtok Traditional Council

Nightmute Traditional Council

Nikolai Village

Ninilchik Village Traditional Council

Nome Eskimo Community

Nondalton Village

Noorvik Native Community

Northway Village,

Nulato Village

Nunakauyak Traditional Council

Organized Village of Grayling

Organized Village of Kake, Petersburg Indian 

Association

Organized Village of Kwethluk

Organized Village of Saxman

Orutsararmuit Native Council

Oscarville Tribal Council

Pedro Bay Village

Pilot Point Traditional Council

Pilot Station Traditional Village

Platinum Traditional Village Council

Portage Creek Village

Qagan Tayagungin Tribe

Qawalangin Tribe of Unalaska

Rampart Village

Ruby Tribal Council

Selawik Ira Council

Seldovia Village Tribe

Shageluk Native Village

Sitka Tribe of Alaska

Skagway Village
Mailing List

Chapter 6 - Page 27

Sleetmute Traditional Council

Solomon Traditional Council

Stebbins Community Association

Takotna Tribal Council

Tazlina Village Council

Telida Native Village Council

Teller Traditional Council

Traditional Village of Togiak

Tuluksak Native Community

Twin Hills Village Council

Ugashik Traditional Village Council

Umkumiut Native Village

Unga Tribal Council

Village of Alakanuk

Village of Anaktuvuk Pass

Village of Aniak

Village of Arctic Village

Village of Clarks Point

Village of Iliamna

Village of Kalskag

Village of Kaltag

Village of Kotlik

Village of Lower Kalskag

Village of Ohogamiut

Village of Old Harbor

Village of Point Lay

Village of Red Devil

Village of Salamatof

Village of Stony River

Village of Wainwright

Wrangell Cooperative Assn

Yakutat Tlingit Tribe

Arizona

Cocopah Tribal Council

Havasupai Tribal Council

Hopi Tribal Council

Hualapai Tribal Council

Pascua Yaqui Tribal Council

San Juan Southern Paiute Council

Tohono O’odham Nation

White Mountain Apache Tribal Council

White Mountain Apache Tribe

California

Agua Caliente Band of Cahuilla Indians

Alturas Rancheria

Augustine Band of Mission Indians

Barona Band of Mission Indians

Bear River Band of Rohnerville Rancheria

Benton Paiute Reservation

Berry Creek Rancheria

Big Lagoon Rancheria

Big Pine Reservation

Big Sandy Rancheria

Big Valley Rancheria

Bishop Reservation

Blue Lake Rancheria

Bridgeport Indian Colony

Buena Vista Rancheria

Cabazon Tribal Business Committee

Cahuilla Band of Mission Indians

Campo Band of Mission Indians

Cedarville Rancheria

Chemehuevi Tribal Council

Chicken Ranch Rancheria

Cloverdale Rancheria

Cold Springs Rancheria

Colusa Rancheria

Cortina Rancheria

Cuyapaipe Band of Mission Indians

Dry Creek Rancheria

Elem Indian Colony

Elk Valley Rancheria

Enterprise Rancheria

Fort Bidwell Reservation

Fort Independence Reservation

Fort Mojave Tribal Council

Greenville Rancheria

Grindstone Rancheria

Guidiville Rancheria

Hoopa Valley Tribal Council

Hopland Reservation

Inaja-Cosmit Reservation

Ione Band of Miwok Indians

Jackson Rancheria

Jamul Indian Village
Chapter 6

Chapter 6 - Page 28

Karuk Tribe of California

La Jolla Band of Luiseno Indians

La Posta Band of Mission Indians

Laytonville Rancheria

Lone Pine Reservation

Los Coyotes Reservation

Lytton Rancheria

Manchester - Point Arena Rancheria

Manzanita Band of Mission Indians

Mechoopda Indian Tribe of the Chico Rancheria

Mesa Grande Band of Mission Indians

Mooretown Rancheria

Morongo Band of Mission Indians

Pala Band of Mission Indians

Paskenta Band of Nomlaki Indians

Pauma/Yuima Band of Mission Indians

Pechanga Band of Mission Indians

Picayune Rancheria

Pinoleville Reservation

Pit River Tribal Council

Potter Valley Rancheria

Quartz Valley Reservation

Ramona Band of Mission Indians

Redding Rancheria

Redwood Valley Reservation

Resighini Rancheria

Rincon Band of Mission Indians

Robinson Rancheria

Round Valley Reservation

Rumsey Rancheria

San Manuel Band of Mission Indians

San Pasqual Band of Diegueno Indians

Santa Rosa Band of Mission Indians

Santa Ynez Band of Mission Indians

Santa Ysabel Band of Mission Indians

Scotts Valley Rancheria

Sheep Ranch Rancheria

Sherwood Valley Rancheria

Shingle Springs Rancheria

Smith River Rancheria

Soboba Band of Mission Indians

Stewart Point Rancheria

Susanville Indian Rancheria

Sycuan Band of Mission Indians

Table Bluff Reservation

Table Mountain Rancheria

Torres-Martinez Desert Cahuilla Indians

Trinidad Rancheria

Tule River Reservation

Twenty-Nine Palms Band of Mission Indians

United Auburn Indian Community

Upper Lake Rancheria

Viejas Band of Mission Indians

Woodfords Community Council

Yurok Tribe, Middletown Rancheria

Colorado

Southern Ute Tribe

Ute Mountain Ute Tribe

Connecticut

Mashantucket Pequot Tribe

Mohegan Indian Tribe

Florida

Miccosukee Indian Tribe

Seminole Indian Tribe

Iowa

Sac & Fox Tribe of the Mississippi in Iowa

Kansas

Iowa Tribe of Kansas & Nebraska

Kickapoo Tribe in Kansas

Prairie Band Potawatomi Indians

Sac & Fox Tribes

Louisiana

Chitimacha Indian Tribe

Coushatta Indian Tribe

Jena Band of Choctaw Indians

Tunica-Biloxi Tribe
Mailing List

Chapter 6 - Page 29

Maine

Aroostook Band of Micmacs

Houlton Band of Maliseet Indians

Passamaquoddy Tribe

Penobscot Indian Nation

Penobscot Nation Department of Natural Resources

Massachusetts

The Trustee of Reservations

Wampanoag Tribe of Gay Head (Aquinnah)

Michigan

Bay Mills Indian Community of Michigan

Grand Traverse Band of Ottawa & Chippewa

Hannahville Indian Community of Michigan

Huron Potawatomi, Inc.

Keweenaw Bay Indian Community of Michigan

Lac Vieux Desert Band of Lake Superior Chippewas

Little River Band of Ottawa Indians

Little Traverse Bay Bands of Odawa Indians

Match-E-Be-Nash-Wish Band of Pottawatomi Indians

Pokagon Band of Potawatomi Indians of Michigan

Saginaw Chippewa Indian Tribe of Michigan

Sault Ste. Marie Tribe of Chippewa Indians

Minnesota

Bois Forte Reservation Business Committee

Fond Du Lac Reservation

Fond Du Lac Reservation Business Committee

Grand Portage Reservation Business Committee

Leech Lake Reservation

Leech Lake Reservation Business Committee

Lower Sioux Indian Community of Minnesota

Mille Lacs Band of Chippewa Indians

Mille Lacs Reservation Business Committee

Minnesota Chippewa Tribe

Prairie Island Indian Community of Minnesota

Red Lake Band of Chippewa Indians

Red Lake Band of Chippewa Indians of Minnesota

Shakopee Mdewakanton Sioux Community of 

Minnesota

Upper Sioux Community of Minnesota

White Earth Reservation Business Committee

Mississippi

Mississippi Band of Choctaw Indians

Montana

Blackfeet Tribal Business Council

Chippewa Cree Business Committee

Confederated Salish & Kootenai Tribes, Tribal

Crow Tribal Council

Fort Belknap Community Council

Fort Peck Tribal Executive Board

Northern Cheyenne Tribal Council

Nebraska

Omaha Tribal Council

Ponca Tribe of Nebraska

Santee Sioux Tribal Council

Winnebago Tribal Council

Nevada

Battle Mountain Band Council

Carson Community Council

Dresslerville Community Council

Duckwater Tribal Council

Elko Band Council

Ely Colony Tribal Council

Fallon Paiute Shoshone Tribal Business Council

Fort Mcdermitt Tribal Council

Las Vegas Tribal Council

Lovelock Tribal Council

Moapa Business Council

Pyramid Lake Paiute Tribal Council

Reno-Sparks Tribal Council

Shoshone-Paiute Business Council

South Fork Band Council

Stewart Community Council

Summit Lake Paiute Council

Te-Moak Tribe of Western Shoshone

Walker River Paiute Tribal Council

Washoe Tribal Council
Chapter 6

Chapter 6 - Page 30

Wells Indian Colony Band Council

Winnemucca Tribal Council

Yerington Paiute Tribe

Yomba Tribal Council

New Mexico

Jicarilla Apache Tribe

Mescalero Apache Tribe

Pueblo of Acoma

Pueblo of Cochiti

Pueblo of Isleta

Pueblo of Jemez

Pueblo of Laguna

Pueblo of Nambe

Pueblo of Picuris

Pueblo of Pojoaque

Pueblo of San Felipe

Pueblo of San Ildefonso

Pueblo of San Juan

Pueblo of Sandia

Pueblo of Santa Ana

Pueblo of Santa Clara

Pueblo of Santo Domingo

Pueblo of Taos

Pueblo of Tesuque

Pueblo of Zia

Pueblo of Zuni

Ramah Navajo Chapter

New York

Blue Mtn. Reservation

Cayuga Nation of Indians

Oneida Indian Nation

Onondaga Indian Nation

Seneca Nation of Indians

St. Regis Mohawk Tribe

Tonawanda Band of Seneca

Tuscarora Nation

North Carolina

Eastern Band of Cherokee Indians

North Dakota

Spirit Lake Tribal Council

Standing Rock Sioux Tribal Council

Three Affiliated Tribes Business Council

Turtle Mountain Band of Chippewa

Oklahoma

Absentee-Shawnee Tribe of Oklahoma

Alabama-Quassarte Tribal Town

Apache Tribe of Oklahoma

Caddo Indian Tribe of Oklahoma

Cherokee Nation of Oklahoma

Cheyenne-Arapaho Tribes of Oklahoma

Chickasaw Nation

Choctaw Nation of Oklahoma

Citizen Potawatomi Nation

Comanche Indian Tribe

Delaware Tribe of Indians

Delaware Tribe of Western Oklahoma

Fort Sill Apache Tribe of Oklahoma

Iowa Tribe of Oklahoma

Kaw Tribe of Oklahoma

Kialegee Tribal Town

Kickapoo Tribe of Oklahoma

Kiowa Tribe of Oklahoma

Miami Tribe of Oklahoma

Modoc Tribe of Oklahoma

Muscogee (Creek) Nation

Osage Tribal Council

Otoe-Missouria Tribe of Oklahoma

Ottawa Tribe of Oklahoma

Pawnee Tribal Business Council

Peoria Tribe of Oklahoma

Ponca Tribe of Oklahoma

Quapaw Tribal Business Committee

Sac & Fox Nation of Oklahoma

Seminole Nation of Oklahoma

Seneca-Cayuga Tribe of Oklahoma

Thlopthlocco Tribal Town

Tonkawa Tribe of Oklahoma

United Keetoowah Band of Cherokee Indians
Mailing List

Chapter 6 - Page 31

Wichita & Affiliated Tribes

Wyandotte Tribe of Oklahoma

Oregon

Burns Paiute Tribe, General Council

Confederated Tribes of Coos, Lower Umpqua

Confederated Tribes of the Grand Ronde Commun

Confederated Tribes of the Umatilla Indian

Confederated Tribes of the Warm Springs Reservation

Coquille Indian Tribe

Klamath General Council

Siletz Tribal Council

Rhode Island

Narragansett Indian Tribe

South Carolina

Catawba Indian Tribe

South Dakota

Cheyenne River Sioux Tribe

Crow Creek Sioux Tribal Council

Flandreau Santee Sioux Executive Committee

Lower Brule Sioux Tribal Council

Oglala Sioux Tribal Council

Rosebud Sioux Tribal Council

Sisseton-Wahpeton Sioux Tribal Council

Yankton Sioux Tribal Business & Claims Committee

Texas

Alabama-Coushatta Tribe of Texas

Kickapoo Traditional Tribe of Texas

Ysleta Del Sur Pueblo

Utah

Goshute Business Council

Paiute Indian Tribe of Utah Tribal Council

Skull Valley Band of Goshutes

Uintah & Ouray Tribal Business Committee

Washington

Colville Business Council

Confederated Tribes of the Chehalis

Hoh Tribal Business Committee

Jamestown S’klallam Tribal Council

Kalispel Business Committee

Lower Elwha Tribal Council

Lummi Indian Business Council

Makah Indian Tribal Council

Muckleshoot Tribal Council

Nez Perce Contact

Nisqually Indian Community Council

Nooksack Indian Tribal Council

NW Indian Fisheries Commission

Port Gamble S’klallam Tribe

Puyallup Tribal Council

Quileute Tribal Council

Quinault Indian Nation - Business Committee

Samish Indian Nation

Sauk-Suiattle Tribal Council

Shoalwater Bay Tribal Council

Skokomish Tribal Council

Snoqualmie Tribal Organization

Spokane Business Council

Squaxin Island Tribal Council

Suquamish Tribal Council

Swinomish Indian Tribal Community

Umatilla Forest Resource Council

Upper Skagit Tribal Council

Yakama Indian Nation

Wisconsin

Bad River Band of Lake Superior Chippewa Indians

Forest County Potawatomi

Forest County Potawatomi Community of Wisconsin

Great Lakes Intertribal Council

Ho-Chunk Nation

Lac Courte Oreilles Band of Lake Superior Chippewas

Lac Du Flambeau Band of Lake Superior Chippewas

Lac Du Flamebeau Tribal Natural Resource Depa

Menominee Indian Tribe of Wisconsin

Menominee Tribal Enterprises
Chapter 6

Chapter 6 - Page 32

Oneida Tribe of Indians of Wisconsin

Red Cliff Band of Lake Superior Chippewa Indians

Sokaogon Chippewa (Mole Lake) Community)

Sokaogon Chippewa Community

Sokaogon Chippewa Tribe

St. Croix Chippewa Indians of Wisconsin

St. Croix Chippewa Tribe

Stockbridge Munsee Community of Wisconsin

Stockbridge-Munsee Community

Stockbridge-Munsee Tribe

Wisconsin Winnebago Business

Wyoming

Arapaho Business Committee

Shoshone Business Committee

6.4 Organizations

Alabama

Alabama Nursery Association

Auburn University CES

Auburn University, Department of Entomology & Plant 

Pathology

B.A.S.S., Inc.

Alaska

Alaska Conservation Alliance

Alaska Defenders of Wildlife

Alaska Forest Association

Alaska Rainforest Campaign

Alaskan Society of Forest Dwellers

Alaskans for Responsible Resource Management

Cook Inlet Region, Inc.

Copper River Delta Institute

Friends of Berners Bay

Friends of the Earth

Greenpeace

Rural Advisor Office of the Governor

Society of American Foresters Alaska

Tongass Conservation Society

Arizona

Arizona for Wildlife Conservation

Arizona Nature Conservancy

Arizona Wilderness Coalition

Arizona Wildlife Federation

Audubon Society, Huachuca

Audubon Society, Northern Section

Audubon Society, Tucson

Center for Biological Diversity

Citizens for Protection of Prescott Arizona

Citizens of Mt. Graham Scientific Council

Cochise Conservation Council

Conservation Chair

Defenders of Wildlife

Ecological Restoration Institute

Environmental Services, Phelps Dodge Morenci, Inc.

Environmental Services

Flagstaff Activist Network

Fort Apache Timber Co.

Foundation for Biodiversity

Grand Canyon Trust

Hebbard & Webb, Inc.

Maricopa Audubon Society

Native Plant Society of Arizona

The Nature Conservancy

North Country, Inc.

Northern Arizona Loggers Association

Northwest Pine Products

Phoenix Zoo

Plateau Group

Prescott Forest Friends

Rincon Group People for the West

Rocky Mountain Elk Foundation

Roosevelt Community Association

Salt River Project

Scentry Biologicals, Inc.

Sierra Club

Sierra Club, Grand Canyon Chapter

Sky Island Alliance

Southern Arizona Environmental Council

Sonoran Biodiversity Project,

Southwest Environmental Center

Southwest Forest Alliance
Mailing List

Chapter 6 - Page 33

Southwest Forest Watch

Trout Unlimited

Verde Watershed Association

White Mountain Conservation League

The Wilderness Society, Arizona Chapter

Wildlife Society, Arizona Chapter

Arkansas

Arkansas Forestry Association

Defenders of Quachita Forest

Ozark Organic Growers Association

California

Alternatives to Toxics

American Land Conservancy

American River Conservancy

Aspen Environmental Group

Bear Engineering

California Forestry Association

California Native Plant Society

California State Polytechnic University

California State University - Sacramento

Conservation Congress

Council for Planning & Conservation

Earthjustice, Headquarters

Entrix

Five Creek Limited Partnership

Forest Landowners of California

Golden Queen Mining Co.

Greystone

Humboldt State University, Department of Forestry

Natural Resources Defense Council

Oregon Heirs Corp.

PG&E Corporation

Preservation Officer, Oregon-California Trails 

Association 

Robert Burt & Rebecca Burt Family Trust

Santa Cruz Rainforest Action Group

Sierra Club

Society for the Protection & Care of Wildlife

Talon Associates

Trinity River Lumber Co.

University of California, Department of Ecology & 

Evolution

University of San Diego, Pardee Legal Research Center

University of the Pacific

Colorado

Burns & McDonnell

Colorado Forestry Association

The Denver Gold Group

KTUN News Director

Meet the Wilderness

Native American Fish & Wildlife Society

Rocky Mountain Bird Observatory

Sierra Club, Rocky Mountain Chapter (Colorado)

Wilderness Society

Wildlife Management Institute

Connecticut

Club, Connecticut Chapter

Connwood Foresters, Inc.

Connecticut Forests & Park Association

Dubois Forestry Land Management

Forestland Associates

Hull Forestland Management

Keep America Beautiful

Sierra Conwood, Inc.

Tamarack Tree Co.

Timberline Management Co.

University of Conneticut CES

Yale School of Forestry

Delaware

Air Enterprises, Inc.

College of Agriculture & Sciences University of 

Delaware

Delaware Campground Owners Association

Delaware Nature Education Society

Delaware Nature Society

Delaware State University

Department of Agriculture & Natural Resources

Dover Post

Drake Farms
Chapter 6

Chapter 6 - Page 34

Hopeland Farms

Stafford Homeowners Association

T. S. Smith & Sons, Inc.

The Cedars Academy

University of Delaware Entomology

District of Columbia

American Forest Council

American Forest Resource Alliance

American Forests

American Lands Alliance

American Paper Institute

American Pulpwood Association

American Recreation Coalition

American Rivers

American Ski Federation

Americans for the Environment

Association of State & Territorial Health Officers

Coalition for Scenic Beauty

Council of Governors’ Policy

Defenders of Wildlife

Ecological Society of America, The

Endangered Species Coalition

Friends of the Earth

George Washington University Library

Global Leleaf

Greenpeace USA

International Association of Fish & Wildlife

National Agricultural Chemical Association

National Association of Conservation District

National Association of Counties

National Association of State Foresters

National Audubon Society

National Coalition Against the Misuse of Pesticides

National Conference of State Legislatures

National Council for Science & the Environment

National Forest Products Association

National Governors Association

The National Grange

National Parks Conservation Association

National Tree Trust

National Urban League

Natural Resources Council of America

Natural Resources Defense Council

Public Health Foundation

Public Lands Council

Save America’s Forests

Sierra Club, Washington DC Office

Sport Fishing Institute

U.S. Public Interest Research Group

United States Tourist Council

Urban Forestry Administration

Western Governors Association

Wilderness Society, The

World Resources Institute

World Wildlife Fund

Florida

Florida A&M University

Florida Forestry Association

Florida International University

Florida Native Plant Society

Florida State University

Forest Management Trust

Great Outdoors Conservancy

St. Thomas University

Tree Advisors

University of Central Florida

University of Florida, School of Forest Resources

University of South Florida, Sarasota

University of South Florida, Tampa

University of Tampa

University of West Florida

Georgia

Atlanta Audubon Society

Augusta State University

Berry College

Columbus State University

Dalton State College

Emory University

Forest Landowners Association, Inc.

Forest Watch Coordinator

Fox Forestry, Inc.

Georgia Organic Growers Association
Mailing List

Chapter 6 - Page 35

Georgia College & State University

Georgia Department of Natural Resources

Georgia Federation of Forest Owners

Georgia Forestry Association, Inc.

Georgia Institute of Technology

Georgia Southern University

Georgia Southwestern State University

Georgia State University

Kennesaw State University

Mercer University

National Forest Products Association

North Georgia College & State University

Savannah Tree Foundation

Sierra Club, Georgia Chapter

Southeast Lumber Manufacturers Association

State University of West Georgia

Toxic Commission & Assistance Project

Trees Atlanta

Union Camp Woodlands Corp.

University of Georgia

University of Georgia, Department of Entomology

Valdosta State University

Warnell School of Forest Resources

Hawaii

Brigham Young University - Hawaii

University of Hawaii - Manoa

Idaho

Albertson College of Idaho

Association Logging Contractors, Inc.

Bioanalysts

Blue Ribbon Coalition

Boise Cascade Corporation

Boise State University

Caldwell R. & E. Center

Carney Products Co. Ltd

Ceda-Pine Beneer

CH2M Hill

Citizens for a User Friendly Forest (CUFF)

Coeur D’Alene Chamber of Commerce

Coeur D’Alene Tribal Forestry

College of Forestry University of Idaho

Croman Corp

Crown Pacific Island

Dames & Moore

Economic Modeling Specialists

Flying Resort Ranches

Friends of the Clearwater

Hells Canyon Alliance

Idaho Audubon Council

Idaho Conservation League

Idaho Forest Industries, Inc.

Idaho Nursery & Landscape

Idaho Power Co.

Idaho State University

Idaho Trails Council

Idaho Water Users Association

Idaho Watersheds Project

Idaho Wildlife Federation

Idaho Women in Timber

Intermountain Forest Industry Association

JAG, Inc.

Kootenai Environmental Alliance

Lewis-Clark State College

Ling, Nielsen & Robinson

Louisiana-Pacific Corporation

Northwest Management, Inc.

Northwest Timber Workers Resourse Council

Potlatch Corporation

President, ID Chapter, Oregon California Trails 

Association (OCTA)

Regulus Stud Mills, Inc.

Resource Solutions

Ricks College

Selkirk-Priest Basin Association

Sierra Club

Society of American Foresters, Intermountain

Spokesman-Review

St. Joe Economic Development Foundation

Three Rivers Timber, Inc.

University of Idaho Department of Forest ResourcesEntomology

University of Idaho

University of Idaho Extension Forestry
Chapter 6

Chapter 6 - Page 36

University of Idaho Forest, Wildlife & Range Policy 

Analysis Group

University of Idaho, College of Forestry, Forest 

Entomology

University of Idaho Society of American Foresters

Western Forest Environmental Alternative

Western States Equipment

Wilderness Society

Illinois

Abbot Laboratories Capd

America Defender Network

American Nurseyman

Benedictine University

Bradley University

Capital Agriculture Property Service, Inc.

Chicago Region Biodiversity Council, Chicago

Chicago State University

Depaul University

Dominican University

Eastern Illinois University

Governors State University

Harold Flying Service

Hendrickson Flying Service

Illinois Nurserymen’s Association

Illinois Forest Products Co., Inc.

Illinois Native Plant Society

Illinois State Library

Illinois State University

Illinois Valley Community College

Illinois Walnut Council

Illinois Wesleyan University

International Society of Arboriculturists

John A. Logan College

Lake Forest College

Lewis University

Loyola University, Chicago

Monmouth College

Moraine Valley Community College

National Campers & Hikers Association

North American Wildlife Foundation

Northeastern Illinois University

Northern Illinois University

Northwestern University

Northwestern University School of Law

Olivet Nazarene University

Open Lands Project

Plant Illinois

Pontiac Flying Service

Race & Heartwood

Reed’s Fly-On Farming

Rocky Moutain Elk Foundation

Southeast Illinois College

Sierra Club Great Lakes Chapter

South Suburban College

Southern Illinois University Carbondale

Tri-State Forestry

University of Chicago

University of Illinois Extension, University

University of Illinois, Chicago

University of Illinois at Urbana, Champaign

Urbana Forestry Management, Inc.

Valent Biosciences Corp

Western Illinois University

Wheaton College

Wildlife Society Illinois Chapter

Indiana

Ace Pest Control

Akard Forestry Consultants

Al’s Aerial Spraying

Anderson University

Ball State University

Berg-Warner Nursery, Inc.

C.S. Bond Forest Management

Chris Leibering & Sons

Depaul University

Forest & Land Managers, Inc.

Forest Management Services

Hanover College

Hensler Nursery, Inc.

Hopwood Forestry Consultants

Hudson Forestry Co.

Huntington College

Indiana Forestry & Woodland Owners Association

Indiana State University
Mailing List

Chapter 6 - Page 37

Indiana University

Indiana University, Kokomo

Indiana University, Northwest

Indiana University, South Bend

Indiana University, Southeast

Indiana Wood Products, Inc.

Pike Lumber Co., Inc.

Purdue University Hardwood Tree Improvement & 

Regeneration Center

Purdue University

Purdue University Entomology Department

Saint Joseph College

Schuerman Forestry Service

Stambaugh Forestry & Nursery

University of Notre Dame

University of Southern Indiana

Valparaiso University

Wabash College

Walley Lumber Co.

Walnut Council

Weston Paper

Iowa

Control Services

Cornell College

Department of Entomology Iowa State University

Drake University

Drake University

Geode Resource Conservation and Development 

(RC&D)

Grinnell College

Haugen Contracting

Iowa State University

Iowa State University Department of Plant Pathology

Iowa State University Extension Forester

Iowa State University Extension Horticulture

Iowa State University of Science & Tech.

Iowa Woodland Owners Association

Izaak Walton League of America Endowment

Krambeer Forestry Services, Inc.

Lone Tree Nursery

Murphy’s Walnut Hill Nursery

National Association of County Engineers

North Iowa Area Community College

Northwest Landscaping, Inc.

Northwestern College

Pathfinders RC&D

Pella Nurseries

Peters Logging & Milling

Soil & Water Conservation Society of America

State Library of Iowa

Tiedt Nursery & Forestry Service

Trees Forever

Trees-R-Us

University of Iowa

University of Northern Iowa

Upper Iowa University

West Enterprises

Kansas

Baker University

Colby Community College

Dodge City Community College

Fort Hays State University

Kansas City Community College

Kansas State University

Kansas Wesleyan University

University of Kansas

Washburn University

Wichita State University

Kentucky

Department of Entomology, University of Kentucky

Eastern Kentucky University

Kentucky State University

Kentucky Wesleyan College

Kentucky Woodland Owners Association

Kentucky Forest Industries Association

Morehead State University

Murray State University

Northern Kentucky University

Union College

University of Kentucky

University of Kentucky Department of Entomology

University of Kentucky Department of Forestry
Chapter 6

Chapter 6 - Page 38

University of Louisville

Western Kentucky University

Louisiana

Alexandria Forestry Center

Louisiana College

Louisiana Forestry Association

Louisiana State University Department of Entomology

Louisiana State University, Baton Rouge

Louisiana State University, Eunice

Louisiana State University, Shreveport

Louisiana Tech University

Loyola University, New Orleans

Norrie Colony, Inc.

Northwestern State University

Southeastern Louisiana University

Southern Forest Products Association

Southern University A&M College

Tulane University

University of Louisiana, Lafayette

University of Louisiana, Monroe

Xavier University of Louisiana

Maine

American Pulpwood Association, Inc.

Andrews Land Service, Inc.

Bates College

Bear Paw Lumber

Bowdoin College

Champion International

Colby College

Coolong Land Surveys

Eco- Analysts

Edgewood Tree Farm

Evans Notch Visitor Center

Forest Society of Maine

Forests by Design

International Paper Co.

The Ireland Group

James River Timber

James W. Sewall Co.

Landvest, Inc.

Leonardi Associates

Llavalley Lumber Co.

M.S. Lavoie Air

Mackintosh Forest Management Services

Maine Maritime Academy

Maple Hill Forest Services

Marine Helicopters

Marty’s Logging

National Audubon Society, Maine Audubon

Natural Resources Council of Maine

Prentiss & Carlisle Co., Inc.

S.D. Warren Woodlands

Sierra Club, Maine Chapter

Small Woodland Owners Association of Maine

Southern Maine Forestry Services

Timberland Consultants

Two Trees Forestry

University of Maine, School of Law

University of Maine, College of Natural Sciences

University of Maine, Orono

University of Maine, Presque Isle

Whittling Ridge Farm

Wood Fiber Industries

Woodlot Alternatives, Inc.

Woodlot Management Services

Maryland

Allegany College of Maryland

Alliance for Maryland Forest

American Hiking Society

Associationiatoin of Consulting Foresters America,

Center for Watershed Diversity

Chesapeake Corporation

Chesapeake Forest-Land Services

East Coast Helicopter, Inc.

Entomological Society of America

Forestry Concepts

Frostburg State University

Glatfelter Pulp Wood Co.

Helicopters Applications, Inc.

International Society of Tropical Foresters

Izaak Walton League of America, Inc.

James Bailey Agrotors
Mailing List

Chapter 6 - Page 39

Kennedy Consultants

The Land & Tree Co.

Maryland Campground Owners Association

Maryland Chistmas Tree Association

Maryland Forest, Park & Wildlife Service

Maryland Forests Association

Maryland Native Plant Society

Michel Forestry Co.

National Military Fish & Wildlife Association

The Orchards Association, Inc.

Parker Forestry Services

Parkton Woodland Services

Pickering Creek Audubon Center, Audubon Maryland

Pine Top Woodland Improvement Ser.

Piney Run Nature Center

Rachel Carson Council, Inc.

Renewable Natural Resources Foundation

Salisbury State University

Society of American Foresters

Spicer Lumber Co.

University of Baltimore

University of Maryland at College Park

University of Maryland, Baltimore County

Versar, Inc.

Washington College

Western Maryland College

Wildlife Habitat Council

Wood Products, Inc.

Massachusetts

Amherst College

Appalacian Mountain Club

Beaver Tree Work

Boston Athenaeum Library

Boston College

Boston University, School of Law

Brandeis University

Conservation Law Foundation of Northeast

Earthwatch

Forest Logic

Forester - W.D. Cowle, Inc.

Gordon College

Harvard College

Holdsworth Natural Resource Center

The Land Concern, Inc.

Massachusetts Audobon Society

Massachusetts Forestry Associatione

Massachusetts Institute of Technology

National Parks Conservation Association

New England Forestry Foundation

North American Family Campers Association

Northeastern University

Sierra Club, Massachusetts Chapter

Trust for Public Land

Tufts University

Tufts University Environmental Program

Turnagain Resources

University of Massachusetts, Amherst

University of Massachusetts CES

University of Massachusetts, Dartmouth

University of Massachusetts, Department of 

Entomology

University of Massachusetts, Department of Forestry & 

Wildlife Mgmt

University of Massachusetts Extension

University of Massachusetts, Lowell

University of Massachusetts, Medical Center

University of Massachusetts, Shade Tree Lab

University of Massachusetts, Western MA Agriculture 

Center

University of Massachusetts, Department of Zoology

Wellesley College

Western New England College

Williams College

Michigan

Abitibi-Price Corporation

Albion College

Al’s Aerial Spraying

American Motorcycle Association

American Society of Agricultural Engineers

Armintrout’s Nursery

Battelle Great Lakes Environmental Center

Biewer Sawmill, Inc.

Big Creek Associates

Central Michigan University
Chapter 6

Chapter 6 - Page 40

Champion Fleet Owners Association

Citizens for Alternatives to Chemical Contamination

Consumers Power Co.

Cycle Conservation Club of Michigan

Delta College

Dow Chemical Co.

Earl’s Spray Service, Inc.

Eastern Michigan University

The Ecology Center

Federation Natural Resources

Federation of Fly Fishers

Ferris State College

Ferris State University, Department of Bio Science

Forest-Land Services, Inc.

Georgia Pacific Corporation

Global Relief of Michigan

Great Lakes Camp & Trail Association

Grosse Pointe Woods Tree Commission

Hatfields Spraying Service

Huron Audobon Club

Hydrolake Leasing & Sales

Keweenaw Land Association, Ltd.

Lake States Forestry Alliance

Lake States Lumber Association

Lake Superior State University

Mackinac Chapter - Sierra Club

Marble Institute of America

Maurice’s Flying Service

Mead - Publishing Paper Division

Metropolitan Forestry Consultants, Inc.

Michigan Association of Private Campground Owners

Michigan Association of Timbermen

Michigan Audubon Society

Michigan Bow Hunters

Michigan Chamber of Commerce

Michigan Conservation Foudation

Michigan Environmental Council

Michigan Environmental Defense League

Michigan Forest Products Industry Dev. Council

Michigan Forests Association

Michigan Independent Wood Products Association

Michigan Nature Association

Michigan Nature Association

Michigan Recreation Canoe Association

Michigan Salmon & Steelheader Association

Michigan Snowmobile Association

Michigan State University Department of Entomology

Michigan State University Extension

Michigan State University, Pesticide Research Center

Michigan Tech University, School of Forestry & 

Wildlife

Michigan Technological University

Michigan Technological University School of Forest & 

Wood Products

Michigan Trail Riders Association

Michigan Tree

Michigan Trout Unlimited

Michigan United Conservation Clubs

Michigan Wild Turkey Federation

Michigan Wilderness Prevention

Michigans Trapper’s Association

Mid-Michigan Helicopters, Inc.

Morth County Trail Association

Michigan State University, Department of Forestry

National Gypsy Moth Management Group

National Woodlands

Nature Conservancy

North Central Michigan College

North Country Trail Association

Northeast Michigan Sportsmen Club

Northern Hardwoods

Northern Michigan University

Northwestern Financial Center

Oakland University

Oscoda Sierra Club

Outdoor Access, Inc.

Packaging Corporation of America

Pere Marquette Watershed Council

Pictured Rocks National Lakeshore

Potts Tree Farm

Ruffed Grouse Society

Screamin Eagle Aviation

Sierra Club

Sierra Club, West Michigan Group

Steelhead Anglers

Steiger Lumber Co.
Mailing List

Chapter 6 - Page 41

Timberwatch

Trout Unlimited

University of Michigan, Botanical Gardens

University of Michigan, School of Natural Resources

West Michigan Environmental Action Council

West Michigan Tourist Association

Western Michigan University

Weyerhauser

Wiggins Tree Co.

Wild Turkey Federation

The Wilderness Society

Minnesota

Airborne Custom Spraying, Inc.

Cook Co.

Forest Management Specialists, Inc.

Forestry Associates

Hamline University

Heartwood Forestry

Kunde Co. Forestry

Minnesota Forestry Association

Minnesota Ornithologists’ Union

Moorhead State University

Mountland Timber, Inc.

North Hardwoods Notes

Plant Health Associates, Inc.

Potlatch Corp.

Privatelands Forestry Consulting

Rijala Timber Co.

Saint Cloud State University

Split Rock Forestry, Inc.

Sundance Silviculture

Two by Forestry

Woodland Services, Inc.

Mississippi

Alcorn State University

Delta State University

Jackson State University

Mississipe State University Department of Entomology

Mississippi College

Mississippi Department of Agriculture & Commerce

Mississippi Forestry Commission

Mississippi State University Department of Forestry

Mississippi State University, College of Forestry

Mississippi University for Women

Mississippi State University

Mississippi Forestry Commission

National Association of Professional Forestry Schools

Sidney Malone International

University of Mississippi

University of Southern Mississippi

Missouri

Audubon Society of Missouri

Central Missouri State University

Chamberlain & De James

Department of Conservation/Forestry Division

Dowler’s Lower Place

Foremost Forest Managers

G-W Lumber Co.

Hammons Products Co.

Jefferson College

Lincoln University

Lindenwood University

Loners of America

Maryville University of Saint Louis

Meers & Associates

Metropolitan Forestry Services

Missouri Forest Products Association

Missouri Native Plant Society

Missouri Southern State College

Missouri Forest Management Co.

National Timber Consultants

Northwest Missouri State University

Port Hudson Timber Management

Rockhurst University

Saint Louis University

Schnurbusch Land Services

Sierra Club, Ozark Chapter (Missouri)

Skip Kincaid & Associates

Southeast Missouri State University

Southeastern Association of Fish & Wildlife Agencies

Southwest Missouri State University

Steward Agriculture Research Services
Chapter 6

Chapter 6 - Page 42

Truman State University

University Extension, University of Missouri

University of Missouri

University of Missouri, Columbia Department of 

Entomology

University of Missouri, Columbia Department of 

Entomology

University of Missouri, Columbia

University of Missouri, Kansas City

University of Missouri, Rolla

University of Missouri, Saint Louis

University of Missouri, School of Forestry, Fisheries & 

Wildlife

Washington University

Washington Wheatley Neighborhood Association

William Jewell College

Montana

Alliance for the Wild Rockies Ecosystem Defense

American Wildlands

Belt Creek Information Center

CS & KT Forestry

Lewis & Clark Intrepretive Center

Missoula Technology Development Center

Montana Forest Owners Association

Montana State University, Billings

Montana State University, Bozeman

Montana State University, Northern

Montana Tech/University of Montana

National Audubon Society, Montana Audubon

National Forest Foundation

Rocky Mountain Elk Foundation

Scentry Biologicals., Inc.

University of Montana

Wilderness Watch

Nebraska

Creighton University

Dana College

National Arbor Day Foundation

University of Nebraska, Kearney

University of Nebraska, Lincoln

University of Nebraska, Omaha

Wayne State College

Nevada

University of Nevada, Las Vegas

University of Nevada, Reno

New Hampshire

Blue Hills Forest Products

Dartmouth College

Foreco

Gorham Land Co.

James River Corp.

Little Pro Timber Service

New Hampshire Association of Conservation Districts

New Hampshire College

New Hampshire Landowners Alliance

New Hampshire Timberland Owners Association

NH Timberlands Owners Association

Northern Forest Lands

Preserve Appalachian Wilderness

Saint Anselm College

Sierra Club, New Hampshire Chapter

Society for the Protection of New Hampshire Forests

Trust for New Hampshire Lands

University of New Hampshire

Urban Forestry Center

Wagner Woodlands, Inc.

Waterville Valley Co.

White Mountain Attractions Association

Wonalancet Outdoor Club

New Jersey

Aeolium Nature Center

Alliance for Environmental Concerns

American Littoral Society

Arbor Management

Batsto Nature Center

Bergab County Wildlife Center

Cattus Island Nature Center

Center for Environmental Studies
Mailing List

Chapter 6 - Page 43

College of Saint Elizabeth

Conservation & Environmental Center, Inc.

Consolidated Eastern Corp.

Cordoba Helicopter Enterprises, Inc.

Cornucopia Network of New Jersey, Inc.

County College of Morris

The Delicate Balance

Downstown Aero Crop Service, Inc.

Drew University

Fairleigh Dickinson University

Flat Rock Brook, Director

Forest Management Services

Forestry Section, Cook College

Galway Forestry Services

Grassroots Environmental Organization

H & S Forestry Co., Inc.

The Hope Commission

Interstate Pest Control Compact

Irvington Outdoor Education Center

Monmouth University

Montclair State University

New Jersey Association of Conservation Districts

New Jersey Beekeepers Association, Inc.

New Jersey City University

New Jersey Coalition for Alternatives to Pesticides

New Jersey Conservation Foundation

New Jersey Environmental Federation

New Jersey Forestry Association

Ocean County College

Palisades Nature Association

Paul Cowie Associates

PGE

Poricy Park Nature Center

Princeton Education Center at Blairstown

Princeton University

Public Service Electric & Gas Co.

Rdg Associates, Inc.

Rider University

Rowan University

Rutgers University, Camden

Rutgers University, New Brunswick

Rutgers University, Newark

Sandy Hook Environmental Education Center

Seton Hall University

Seton Hall University, School of Law

Shade Tree Commission

Sierra Club Chatam

South Branch Watershed Association

Spermaceti Cove Visitor Center

Stony Brook-Millstone Watershed

Tenafly Nature Center

Trailside Nature & Science Center, Director

Upper Raritan Watershed Association

Watergate Environmental Education Center

Weis Ecology Center

The Wetlands Institute

YMCA Camp Bernie

New Mexico

Audubon Society

Audubon Society, Central New Mexico

Audubon Society, Southwest New Mexico

Carson Forest Watch

Chippeway Lumber, Inc.

Coalition for Public Lands & Natural Resources

Coalition of AZ/NM Counties

Conklin Lumber Co

Department of Natural Science, WNMU

Earth First, New Mexico

Eastern New Mexico University

Forest Conservation Council

Forest Guild

Forest Trust

Forestry Association, Inc.

Gila Conservation Coalition

Gila Watch

Hansen Lumber Co., Inc.

Hawkwatch International

Izaak Walton League 

Kuykendall Lumber Co.

La Jicarita Enterprise, Inc.

Madera Forest Products Coop

MCS Task Force of NM

Moore Cash Lumber

Native Plant Society of New Mexico

The Nature Conservancy of New Mexico
Chapter 6

Chapter 6 - Page 44

New Mexico Audubon Council

New Mexico Earth First!

New Mexico Farm & Livestock Bureau

New Mexico Highlands University

New Mexico Junior College

New Mexico Public Interest Research Group

New Mexico Public Land Council

New Mexico Rural Development Response

New Mexico State University

New Mexico Wilderness Study Committee

People for the West

Public Land Users Association

The Quivira Coalition

Sanchez Timber & Mill Co.

Santa Fe Canyon Association

Santa Fe Forestry Council

Scientech

Sierra Club

Sierra Club, Albuquerque Group

Sierra Club, Pajarito Group

Sierra Club, Rio Grande Chapter

Sierra Club, Santa Fe Chapter

Sierra Club, Tularosa Basin Group

Taos Birders

Taos Nature Society

University of New Mexico

University of New Mexico, School of Law

Western Environmental Law Center

Western Network

Western New Mexico University

Wild Turkey Federation, Las Cruces Chapter

Zuni River Watershed

New York

A to Z Forestry

Adelphi University

Adirondack Conservancy

Adirondack Council

Adirondack Eco-Center

Adirondack Forestry, Inc.

Adirondack Mountain Club

Airspray, Inc.

Alley Pond Enviroment Center, Inc.

Alpine Forestry

American Birding Association, Inc.

American Forest Council

American Nature Study Society

American Whitwater Affiliation

Appalachian Forestry Consulting Services

Arbor Care Ltd.

Arthur W. Butler Memorial Sanctuary

Ashokan Field Campus

Baltimore Woods

Bayard Cutting Arboretum

Beaver Lake Nature Center

Beaversprite Nature Center

Binghamton University

Brooklyn Botanic Garden

Brooklyn College/College University of New York 

(CUNY)

Brooks Resources Management Co.

Buttermilk Falls Tree & Turf, Inc.

Camp Greenkill Environmental Education Center

Camp Owahta Outdoor Education Center

Capake Falls State Park Nature Center

Cary Arboretum

Catskill Forest Association

Catskill Mountain Forestry Service

Catskill Real Estate Appraisals

Christian Gearwar

City College/CUNY

Clarkson University

Clear Creek Consulting

Colgate University

The College of Insurance

Columbia University

Conservation Advisory Committee

Cooper Union 

Co-operating Consultant Foresters

Cornell Laboratory of Ornithology

Cornell University

Cornell University CES

Cornell University, Department of Entomology

The Cummings Nature Center

Delaware Valley Forestry Service

Downing Enviroment & Forest Consultant
Mailing List

Chapter 6 - Page 45

Dyken Pond Environment Education Center

Earth First

East-West Forestry Association

Elmira College

Empire State Forestry Service

Environmental Action Coalition

Environmental Defense, Headquarters

Erie County CES

Ferncliff Forest

Five Rivers Environmental Education Center

Fordham University

Fordham University, School of Law

Forecon, Inc.

Forest-All: Tree & Forest Care

Forsite Forestry

Fountain Forestry, Inc.

Golden Valley Outdoor Recreation Center

Great Neck Outdoor Environmental Center

Green Chimneys Farm Center

Greenburg Nature Center

Greenwood Park

Gunlocke Co.

Heldeberg Workshop

Herbert H. Lehman College/CUNY

High Rock Park Conservation Center

Hillside Outdoor Education Center

Hofstra University

Intermountain Forestry

Intermountain Forestry

International Paper Co.

Inwood Hill Park Environmental Education Center

IPM Laboratories

Kenneth L. Willimas & Association

Long Island University

Mallery Lumber Co.

Manitoga Hudson River Nature Center

Mianus River Gorge Wildlife Refuge & Botanica

Micha Tree & Landscape Consultants

Miller Forest Products

Minna Anthony Com. Nat. Center

Monroe Tree & Landscape, Inc.

Muscoot Park Interp. Farm

Museum of Hudson Highlands

Nassau County Museum Preserve

National Audubon Society

National Campers & Hikers Association

National Council of Paper Industry

Natural Resources Defense Council, Inc.

North East State Foresters Association

New York Botanical Gardens Institute of Ecosystem 

Studies

New York Forest Owners Association

New York Law Institute

New York Law School

New York University

New York Zoological Park

Northeast Timber Services

Northern Consulting

Northern States Tree Services, Inc.

Norton Timberland Management

NY Department of Environmental Cons.

Oceanside Marine National Study Area

Pace Environmental Center

Peterson Forestry, Inc.

Pioneer Forestry Service

Planting Fields Arboretum

Plattsburgh State University

Pok-O-Moonshine Outdoor Education Center

Queens Botanical Garden

Queens College/CUNY

Quogue Wildlife Refuge

Regional Plan Association

Rockland Lake Nature Center

Rudolf Steiner Farm School

Rye Nature Center

Sackhoes Region Nature Center

Saint Bonaventure University

Saint John’s University

Sapsucker Woods Bird Sanctuary

Sarah Lawrence College

Sharpe Enviroment Center

Skidmore College

South Fork, Shelter Island Chapter

St. Lawrence University

State University College of Technology

State University College of Brockport
Chapter 6

Chapter 6 - Page 46

State University College of Cortland

State University College of Geneseo

State University College of New Paltz

State University College of Oneonta

State University College of Oswego

State University College of Potsdam

State University of New York (SUNY)

SUNY, Albany

SUNY, Buffalo

SUNY, College of Environmental Sciences & Forestry

SUNY, Farmingdale

SUNY Institute

SUNY Maritime College

SUNY, Stony Brook

SUNY, Department of Natural Resources

Sylvan Forestry Services

Synecology Forest Management

Syracuse University

T. Roosevelt Memorial Bird Sanctuary

Teatown Lake Reservation

Thompson Pond Project

Thorington Forestry Service

Timberland

Town of North Salem

Town of Warrensburg

Trailside Nature Museum

Trust for Public Land

Twin Valleys Outdoor Education Center

Union College

University of Rochester

Upland Farm

Upper Delaware Council

Upper Delaware Scenic & Rec. River

Vassar College

Wave Hill Center for Environ. Study

Weinberg Nature Center

Westmoreland Sanctuary, Inc.

Winding Hills Park Nature Center

North Carolina

Abw Lumber Indiustries, Inc.

Apex Nurseries, Inc.

Appalachian State University

Associationiated Hardwoods, Inc.

Bartlett Tree Rsch Lab

Beard, E.N. Hardwood Co.

Campbell University

Catawba College

Cramer Lumber Co.

Davidson College

Duke University

Duke University School of Forestry & Environmental 

Studies

East Carolina University

Fayetteville State University

Forest History Society, Inc.

Gilkey Lumber Co., Inc.

Interforest, Inc.

Mount Olive College

National Foundation for the Chemically Hypersensitive

National Toxics Campaign

NC Christmas Tree Association

North Carolina A & T State University

North Carolina Central University

North Carolina Forestry Association

North Carolina State University

North Carolina State University College of Forest 

Resources

North Carolina State University Department of 

Entomology

North Carolina Wesleyan College

North Carolina Wildlife Resources Commission

Oaks Unlimited

Prime Lumber Co.

Queens College

Southern Appalachian Biodiversity Project

University of North Carolina, Chapel Hill

University of North Carolina, Charlotte

University of North Carolina, Greensboro

University of North Carolina, Pembroke

University of North Carolina, Wilmington

Voohees & Pitts Lumber Co., Inc.

Wake Forest University

Watson Lumber Co., Inc.

Western Carolina University
Mailing List

Chapter 6 - Page 47

North Dakota

North Dakota State University

Sully’s Hill-N-Game Preserve

University of North Dakota

Ohio

American Farm Tree Program

Ashland University

Bob Ruhe AG Service

Bowling Green State University

Capital University

Case Western Reserve University

Cleveland State University

College of Wooster

Custom Forestry

Denison University

Downing Woodland Services

Forest Resource Consultants

Hooking College

International Association of Natural Resource

John Carroll University

Johnson’s Forest Products

The Longaberger Co.

Kent State University

Kenyon College

MacArthur Lumber & Post Co.

Marietta College

Marketing Labs Co., Inc.

Meadow Woodlands

Med Woodlands

Miami University

Miami University - Middletown

Municipal Arborists & Urban Foresters Society

Muskingum College

Native Plant Society of Northeast Ohio

Oberlin College

Ohio Christmas Tree Association, Inc.

Ohio Ecological Food & Farm Association

Ohio Forestry Association, Inc.

The Ohio Lepidopterists

Ohio Integrated Pest Management Program

Ohio Northern University

Ohio Nursery Association

Ohio State University

Ohio State University Extension Lucas Co.

Ohio Tree Consulting Services

Ohio University

Ohio Wesleyan University

Olde Forester Consultants

Otterbein College

Potts Tree Farm

Prime Air

Shawnee State University

Society of Municipal Arborists

Timberland Forestry Consutants

Tree Sentry

Treevalue Forestry Service

University of Akron

University of Cincinnati

University of Dayton

University of Findlay

University of Toledo

Wright State University

Youngstown State University

Oklahoma

East Central University

National Watershed Coalition

Native Americans for a Clean Environment

Northeastern State University

Northwestern Oklahoma State University

Oklahoma Forestry Association.

Oklahoma State University

Oklahoma State University Department of Forestry

Oklahoma Woodland Owners Association

Southeastern Oklahoma State University

Southern Nazarene University

Southwestern Oklahoma State University

University of Central Oklahoma

University of Oklahoma

University of Tulsa
Chapter 6

Chapter 6 - Page 48

Oregon

Andersen Forestry Consulting

Agri-Pacific Resources, Inc.

Alice V. Wissman Living Trust

American Fisheries Society - Oregon Chapter

American Lands Alliance

Argus Observer

Associated Oregon Loggers

Atterbury Consultants, Inc.

Audubon Society of Portland

Avion Water Co., Inc.

B & S Logging, Inc.

Bark

Big Pines RV Park

Black Butte Resort

Blue Mountain Back Country Horsemen

Blue Mountain Biodiversity Project

Blue Mountain Native Forest Alliance

Blue Mountains Biodiversity Project

Blue Mountain Lumber Products

Boise Cascade Corporation

Boise Corporation of Northeast Oregon 

Brandt-Nelson, Lark & Brandt

The Bulletin

Camp Tamarack

Capital Press

Cascadia Forest Alliance

Center for Environmental Equity

Central Cascades Alliance

Central Oregon Audubon Society

Central Oregon Community College

Central Oregon Small Woodlands Associates

Central Point Lumber

Chambers Communication Corp

Churchill, Leonard, Brown, Lodine, & Hendrie

Circle De Lumber Co.

City of Eugene, Parks & Open Space

Clouston Energy Research

Cold Springs Resort

Columbia Basin Fish and Wildlife Authority

Concerned Friends ofthe Winema

Consolidated Pine

Crane Prairie Resort

Crescent Creek Cottages

Crescent Lake Lodge & Resort

Crescent Lake RV Park

Crescent Water Association

Crown Pacific Ltd.

Crown River Corp. & Crown Zellerbach

Cultus Lake Resort

D.R. Johnson Lumber (Prarie Wood Products)

David Evans & Associates

Defenders of Wildlife

Deschutes Province Advisory Committee

Double-D-Logging

Douglas Timber Operators, Inc.

Dow Agrosciences

Dr. Johnson Lumber Companies

Dunn Family Trust

Earth Share of Oregon

Eastern Oregon University

Eastern Oregon University Baker Center

Eco-Northwest

Ecola Creek Awareness Project

Ecosystem Equity Council

Emerald Chapter, Native Plant Society of Oregon

Environmental Studies Center

Eugene Burrill Lumber Co.

Eugene Parks & Open Space

Evergreen Helicopters

Forest Recovery, Inc.

Forest Resource Management, Inc.

Forest Resource Services

Forest Service Employees for Environmental Ethics

Fowler Timber Co.

Friends of the Columbia Gorge

Friends of Black Butte Ranch

Friends of Living Oregon Waters

Friends of the Greensprings

Friends of the Metolius

Friends of the Metolius-Environmental Advocates

Glide Lumber Co./Western Timber Co.

Grande Ronde Resource Council

Grande Ronde Model Watershed Program

Grant County Conservationists

Grizzly Mountain Aviation
Mailing List

Chapter 6 - Page 49

Guistina Resources

Haglund, Kirtley, Kelly & Horngren LLP

Halfway House Gallery & Gifts

Headwaters

Hells Canyon Preservation Council

Henderson Logging

High Desert Committee, Sierra Club

Hood River County Weed & Pest Director

Independent Forest Products Association

Institute for Fisheries Resources

Izaak Walton League

J. & J. Logging

J. Herbert Stone Nursery

Joseph Timber Co.

K/P Corporation

Keerins Ranch

Kinzua Resources

Klamath Siskiyou Wildlands Center

L. & D., Inc.

Lane County Audubon Society

Lapine Forestry Services, Inc.

Larch Co.

Lewis & Clark College

Lewis & Clark Law School

Linfield College

Lowell Service Center

Lowell Work Center

M. & L. Enterprises

Malheur Lumber Co.

Malheur Timber Operators, Inc.

Mason, Bruce & Girard, Inc.

Mckenzie Guardians

Metolius Meadows Property Owners Association, Inc.

Metolius Recreation Association

Metolius River Summer Homes

Mid-Columbia Native Plant Society

Monarch Magic

Mountaineers

Mt. Bachelor, Inc.

Musselman & Association, Inc.

Myrmo & Sons, Inc.

National Audubon Society

Native Plant Society

Native Plant Society of Oregon, Blue Mountain 

Chapter

Native Plant Society of Oregon, Wm Cusick Chapter

The Nature Conservancy

Neighborhood Association

Net Work Association Ecological Consulting

North American Butterfly Association

North American Wild Sheep

North Santiam Watershed Council

North Santiam Watershed Forum

Northwest Coalition for Alternatives to Pesticides

Northwest Environmental Defense Council

Northwest Forestry Association

Northwest Resource Council

Northwest Special Forest Products Association

Northwestern School of Law

The Nugget Natural Areas Association

Ochoco Lumber Co.

Odell Lake Homeowners Association

Odell Lake Resort

Odell Sportsman

Old Cascades Wilderness Com

One World Trade Center

Oregon Association of Nurserymen

Oregon Cattlemen’s Association

Oregon Eagle Foundation

Oregon Equestrian Trails

Oregon Farm Bureau

Oregon Forest Homeowners Association

Oregon Hunters Association

Oregon Natural Desert Association

Oregon Natural Resources Council

Oregon Poison Center

Oregon Rivers Council

Oregon Sierra Club Wildlands

Oregon Small Woodlands Association

Oregon Society of American Foresters

Oregon State Public Interest Research Group

Oregon State Snowmobile Association

Oregon State University 

Oregon State University CES

Oregon State University, Department of Botany & 

Plant Pathology
Chapter 6

Chapter 6 - Page 50

Oregon State University, Department of Entomology

Oregon State University, Department of Forest 

Management

Oregon State University, Department of Rangeland 

Resources

Oregon State University Integrated Plant Protection 

Center, Cordley Hall

Oregon Tilth, Inc.

Oregon Water Resources

Ouzel Outfitters

P & M Lumber & Cedar Products

Pacific Environmental Advocacy Center (PEAC)

Pacific University

Pacificorp

The Pacific Rivers Council

Pape Brothers, Inc.

Pendleton Record

People for the West

Pine Creek Logging, Inc.

Pine Point Forest

Portland State University

Portland General Electric

Prarie Wood Products

Public Forestry Foundation

Reed College

Rei Co-Op

River Conservancy

Robert E. Morris Contracting

Rogue Valley Audubon Society of Medford

Rosboro Lumber Co.

Rosebud Contracting

Ross Trust

Salmon-Drift Creek Watersheds Group

Samuel S. Johnson Foundation

Sandy River Basin Watershed Council

Save Our Wild Salmon Coalition

Save the West

Shirbeck, Inc.

Sierra Club

Sierra Club Juniper Group

Sierra Club Oregon Chapter

Sierra Club Portland

Sierra Club, Oregon Chapter

Siskiyou Regional Educational Project

Sisters Forest Planning Committee

Smith Properties

Society Advocating Natural Ecosystems

Society for Range Management

Society of American Forersters, OR State Society

South Oregon University- Biology Department

Southeast Oregon Resource Advisory Council

Southern Oregon University

Steens Mountain Packers

Sun Mountain Water Systems, Inc.

Sunriver Nature Center

Sunriver Owners Association

Sunriver Properties Oregon Ltd.

Sunriver Resort

Thompson Timber Co.

Timber Data Co

Trout Unlimited-Oregon Council

Twin Lakes Resort

Twin Rivers Logging Co.

Umatilla Basin Watershed

University of Oregon

Upper Deschutes Watershed Council

Wagon Wheel Water Co.

Walla Walla Basin Watershed Council

Wallowa Forest Products (D.R. Johnson)

Washington Forest Law Center

Water Wonderland Improvement District

Western Ancient Forest Campaign

Western Environmental Law Center

Western Farm Service

Western Forestry & Conservation Association

Western Oregon University

Western Radio Services Co, Inc.

Wilderness Trail Riders, Inc.

Wildland Resources

Wildlife Management Institute

Wildlife Society, Oregon Chapter

Willamette Industries

Willamette Pass Inn

Willamette University

Woodsman Motel

Xerces Society
Mailing List

Chapter 6 - Page 51

Yaquina Basin Planning Team

Zacharias Logging

Pennsylvania

A.D. Renninger Lumber Co.

Agrotors

Ahora Tree Service

Alder Valley Forestry Consulting

Allegheny Acres

Allegheny College

Allegheny Foresters & Consultants

Allegheny Forestry, Inc.

Allegheny Portage Railroad

American Forestry Consultants

Andrews Woodlot Consulting Service

Appalachian Forest Consutants

Arboreal Forestry Services

Bailey Lumber Co.

Bear Run Nature Reserve

Beechwood Farms Nature Reserve

Bloomsburg University of Pennsylvania

Bradford Area Chamber of Commerce

Bradford Naturalist Club

Briar Bush Nature Center

Brownlee Lumber, Inc.

Bucknell University

C.F.E., Inc.

Carl Hunsberger Sawmill, Inc.

Chas. M. Shaffer Memorial Natl. Center

Clifford B. Carts Co.

Coastal Lumber Co.

Collins Pine Kane Hardwood Division

Columbia County CES

Derwood Nature Center

Duquesne University

Dwight Lewis Lumber Co.

East Stroudsburg University

Felton Associates

Forest Land Services, Inc.

Forest Management Center

Forest Management Associates, Inc.

Forest/Woodlot Management, Inc.

Forestry & Wildlife Consulting

Franklin & Marshall College

Franklin Foresty Services

Freeman’s Forestry & Wildlife Services

Fulton Forest Products

Glatfelter Pulpwood Co.

Haverford College

Hercon Environmental

Highlands Lumber Co., Inc.

Hyma-Devore Lumber Co.

Indiana University of Pennsylvania

International Paper Co.

J.M. Wood Products, Inc.

Jay-For Logging

John J. Tyler Arboretum

Joseph W. Arnold Association

Kane Hardwood Division, Collins Pine

Kuhns Bros. Lumber Co., Inc.

La Roche College

Lamar National Wildlife Refuge

Landon Forestry Services

Lapp Lumber Co.

Lehigh University

Longwood Gardens

Mansfield University

Meiser Lumber Co.

Miller Ag-Craft, Ltd

Miller Aircraft Limited

Millersville University of Pennsylvania

Montgomery County Community College

Mooretown Mill

Morris Arboretum

Mt. Valley Farms & Lumber Products, Inc.

Muhlenberg College

Nagy & Webb Forestry & Surveying Services

National Audobon Society, Audobon Science off

National Gypsy Moth Management

National Gypsy Moth Management Group, Inc.

Northern Timber Services

Northern Timberlands

Open Land Conservancy

Penelec Manager of Forestry

Penn Forestry Co., Inc.

Penn State New Kensington
Chapter 6

Chapter 6 - Page 52

Pennsylvania Deer Association

Pennsylvania Forestry Association

Pennsylvania State University

Pennsylvania State University Department of 

Entomology

Polaris Surveying & Forestry

Proctor & Gamble Paper Co.

Pennsylvania State University, School of Forest 

Resources

Ram Forest Products

Regional Vice President National Audubon Society

Robert Labar Forestry Consultant

Robert Morris College

Robert S. Bommer, Jr., Inc.

Rohm & Haas

Rolling Rock Farms

The Ruffed Grouse Society

Saint Joseph’s University

Seneca Highlands Association

Shippensburg University of Pennsylvania

Sierra Club, Pennsylvania Chapter

Slippery Rock University

Sparty-Wood Products, Inc.

Sunderland Surveying & Forestry

Susquehanna County Historical Society & Free Library 

Association

Swarthmore College

Tallman Aerial Spraying

Temple University

Timber Mgmt Services

Tinicum National Environmental Center

Twin Ponds Sawmill

Twin Tier Systems

University of Pennsylvania

University of Pittsburgh

University of Pittsburgh-Bradford

Valent Biosciences

Villanova University.

Walter H. Weaver Co,

West Chester University Department of Biology

West Chester University of Pennsylvania

Wheeland Lumber Co., Inc.

Whites Wood Nature Center

Rhode Island

Brown University

Florence Gray Center

Group for Alternative to Spraying Pesticides

National Network of Forest Practitioners

Pesticide Public Policy Foundation

Providence College

Rhode Island College

Rhode Island Forest Conservator’s Organization

Rhode Island Organic Farmer’s Association

Southern New England Forest Consortium, Inc.

Southern Northeast Woodland Service

Turnquist Lumber Co.

University of Rhode Island

University of Rhode Island CES

University of Rhode Island, Department of Plant 

Sciences

University of Rhode Island, Department of Forestry

South Carolina

Center for Forested Wetlands

Charleston Southern University

Clemson University

Clemson University, Department of Forest Resources

Coastal Carolina University

College of Charleston

Francis Marion University

Furman University

Lander University

South Carolina State University

Southern Appalachian Botanical Society

University of South Carolina, Aiken

University of South Carolina, Columbia

University of South Carolina, Lancaster

Winthrop University

South Dakota

Black Hills State University

Northern State University

South Dakota State University

University of South Dakota
Mailing List

Chapter 6 - Page 53

Tennessee

Austin Peay State University

Cleveland State Community College

Columbia State Community College

East Tennessee State University

Fisk University

Hardwood Forest Foundation

Hardwood Research Council

King College

Lambuth University

Middle Tennessee State University

Sierra Club, Tennessee Chapter

Tennessee Forestry Association

Tennessee State University

Tennessee Technological University

United States Aviation Underwriters

University of Memphis

University of Tennessee, Department of Forestry, 

Wildlife & Fisheries

University of Tennessee, Knoxville

University of Tennessee, Martin

University of the South

Vanderbilt University

Texas

Angelo State University

Audubon Society

Bat Conservation International

Baylor University

F. Austin State University

Geo-Marine, Inc.

Hardin-Simmons University

Howard Payne University

Lamar University

Loma Linda Homeowners Association

Midwestern State University

Mitchell Energy Corp.

Navarro College

Rice University

Saint Mary’s University

Sam Houston State University

San Antonio College

Sierra Club, El Paso Regional Group

South Texas College of Law

Southern Methodist University

Southwest Texas State University

Stephen Texas Forestry Association

Tarrant County College

Texarkana College

Texas A&M International University

Texas A&M University

Texas A&M University Research & Extension Center

Texas A&M University, Commerce

Texas A&M University, Corpus Christi

Texas A&M University, Kingsville

Texas Christian University

Texas Southern University

Texas Tech University

Trinity University

University of Houston

University of Houston, Clear Lake

University of Houston, Victoria

University of North Texas

University of Texas, Arlington

University of Texas, Austin

University of Texas, Dallas

University of Texas, El Paso

University of Texas, Pan American

University of Texas, San Antonio

West Texas A&M University

Utah

Brigham Young University

Southern Utah University

University of Utah

University of Utah, Department of Biology

Utah State University

Utah Woodland Owners Council

Vermillion Services

Weber State University

Western Association of Land Users
Chapter 6

Chapter 6 - Page 54

Vermont

Harwood Forestry Services

Nature’s Light

Society of American Foresters New England

Upland Resource Group, Inc.

Vermont Law School

Virginia

American Pulpwood Association

Association of Consulting Foresters of America

Chespeake Forest Products Co.

Citizens for a Better America

College of William & Mary

Conservation Foundation

Dupont Nature Club

Dynamic Aviation

Emory & Henry College

Friendly Forest Farms

Future Farmers of America

George Mason University

Hampton University

Helicopter Association International

Highland County CES

Hollins University

James Madison University

Labot-Anderson

Madison County Library, Inc.

Mary Washington College

National Campground Owners Association

National Recreation & Parks Association

National Wildlife Federation

National Wildlife Federation, Headquarters

National Woodland Owners Association

Nature Conservancy

Old Dominion University

Old Time Orchard

Potomac Appalachian Trail Club

Roanoke College

Tetrotech

Trout Unlimited

University of Richmond

University of Virginia

University of Virginia’s College at Wise

Virginia Polytechnic Institute & State University, 

Department of Entomology

Virginia Commonwealth University

Virginia Forestry Association

Virginia Native Plant Society

Virginia Polytechnic Institute

Virginia State University

Virginia Tech University, College of Natural Resources

Virginia Tech University, Department of Entomology

Virginia Tech University, Department of Fisheries & 

Wildlife

Virginia Tech University, Department of Forestry

Virginia Wesleyan College

Virginians for Wilderness/Earth First

Washington & Lee University

Washington

49 Degrees North Ski Area

Alpine Lakes Protect Society

Alps Trustee

Alpine Lakes Protection Society Alta Crystal Resort, 

L.L.C.

American Land Rights Association

American Lands Access Association, Inc.

American Rivers

Apple Valley Broadcasting

Auble Association

Audubon Society, North Cascades

Audubon Society, Skagit Chapter

Backcountry Bicycle Trails

Backcountry Horsemen of Washington, Inc.

Backpacker Magazine

Bear Creek Tree Farms

Bellingham Mountaineers

The Bloedel Reserve

Camp Sheppard, BSA - Chief Seattle Chapter

Center for Environmental Law & Policy

Central Washington University

Central Washington University Department of Geog. & 

Land Studies

Chinook Byways

Citizens of Greenwater
Mailing List

Chapter 6 - Page 55

Coalition for a Livable Washington

Columbia Basin Nursery

Columbiana

Common Sense Resource League

Concerned Friends of Ferry County

Crown Zellerbach Corp.

Crystal Mountain Resort

David Evans & Associates

Double Shake Co.

Dunau Associates

Dupont Forestry Products

Earth Justice Legal Defense Fund

East Lake WA Audubon Society

Eastern Washington University

Ebel & Associates

Elma Truck & TLR

Environmental Outlook, University of Washington

F.O.C.U.S.

Federal Lands Advisory Committee

Ferry County Action League

Forest Recovery-Granger Co.

Forest Stewards Guild, Northwest Regional Chapter

Forestry Sciences Lab

Fort James Corporation

Foster Wheeler Environmental Corp.

Friends of the Earth

Georgia Pacific Corp.

Grassland West Co.

Greater Greenwater Gateway Committee

Guy Bennett Lumber Co.

Inland Empire Paper Co.

Inland Empire Public Lands Council

James River Corp.

Kamerrer Family Farms

Kettle Range Conservation Group

The Lands Council

Lusignan Forestry, Inc.

Mentor Law Group

Methow Valley News

Methow Valley Snowmobile Association

Mountaineers

Mountains to Sound Greenway

National Audubon Society

National Audubon Society, Seattle Audubon Society

National Campers & Hikers

National Outdoor Leadership School, Pacific

National Wildlife Federation

Natural Resources Defense Council

The Nature Conservancy

North Cascade Audubon Society

North Cascades Conservation Council

North Cascades Institute

Northeast Washington Medical Group

Northwest Coalition for Alternatives to Pesticides

Northwest Ecosystem Alliance

Northwest Fly Anglers

Northwest Forestry Association

Northwest Timber Workers Resource Council

Okanogan Highlands Alliance

Okanogan Resource Council

Olympia Forest Sciences Laboratory

The Omak County Chronicle

P.L.U.S.

Pacific Biodiversity Institute

Pacific Crest Biodiversity Project

Pacific Rivers Council

Pilchuck Audubon Society

Plum Creek Timber Co.

Ponderay Newsprint Co.

Potlatch Corporation

Public Land Users Society

Raedeke Association., Inc.

Rafter Seven Ranch

Rainier Audubon Society

Resources Northwest, Inc.

Rivers Council of Washington

Rocky Mountain Elk Foundation

Rosboro Lumber Co.

RZ Resource Consultants

Seattle Audubon Society

Seattle Snohomish Mill Co.

Seattle University

Sierra Club

Sierra Club Legal Defense Fund

Sierra Club Northern Rockies Chapter

Sierra Club, Northwest Office
Chapter 6

Chapter 6 - Page 56

Signpost Magazine

Skagit Valley College

Skyline Wheat Ranch

Society of American Foresters

Spokane Research Center

Stevens Pass Ski Area

Tahoma Audubon Society

The Trust for Public Lands

Trout Unlimited, NW Steelhead & Salmon Council

Umatilla Forest Watch

University of Puget Sound

University of Washington

University of Washington, College of Forest Resources

Upper Columbia Resource Council

Vaagen Brothers Lumber Co.

Volunteers for Outdoor Washington

Washington Native Plant Society

Washington Contract Loggers Association

Washington Environmental Council

Washington Farm Forestry Association

Washington Forest Protection Association

Washington Friends of Farms & Forests

Washington Native Plant Society

Washington State Department of Ecology

Washington State Department of Fish & Wildlife

Washington State Farm Bureau

Washington State Mineral Council

Washington State Snowmobile Association

Washington State University

Washington Wilderness Coalition

Western Land Exchange Project

Western Resource Analysis

Western Washington University

Weyerhaeuser Co.

White River Recreation Association

Whitman College

Wild Washington Campaign

Wilderness Society

Wilderness Watch

Willapa Hills Audubon Society, Conservation B

Zahn Ranch

West Virginia

Allegheny Wood Products, Inc.

Alyeska, Inc.

Appalachian Investments

Appalachian Trail Conference

Bluefield State College

Central Tie & Lumber Co.

Coastal Lumber Co.

Concord College

Davis & Elkins College

Davis & Elkins College, Department of Biology

The Garden Works

Fairmont State College

Fairmont State College, Department of Biology

Hardscrabble Enterprises, Inc.

Harmony Hill

Lapaix Farm

Marshall University

Marshall University, Department of Biological Services

Millstone Farm

Monongahela Power Co.

Mountain Aquaculture & Producers Association

The Mulch Patch

New Dawn Farm

Parsons Volunteer Fire Department

Salem-Teikyo University

Shepherd College

Sierra Club, West Virginia Chapter

Sleepy Creek Seed Co.

Tilinghast & Neely

Twin Oaks Farm & Nursery

West Virginia Bass Federation

West Virgina Forestry Association

West Virgina RC&D Association

West Virgina Sierra Club

West Virginia State College

West Virginia University

West Virginia University CES

West Virginia University Division of Plant & Soil 

Sciences

Westvaco Corp.

Wheeling Jesuit University

Woodland Owners Association of West Viginia
Mailing List

Chapter 6 - Page 57

Wisconsin

American Pulpwood Association

Beloit College

Blue Ox Forestry Service, Inc.

Boardman, Suhr, Curry & Field

Burns Forestry Consultants

Consolidated Papers, Inc.

Florence Mining News

Forestry Services Co

Fox Valley Technical College

Georgia Pacific

Johnson Timber Corp.

Lake States Forestry

Lake States Independent Loggers

Lake States Women in Timber

Lawrence University

Lodholz North Star Acres

Louisiana-Pacific Corporation

M&M Associates

Meier Natural Resources Conservation

Michigan-Wisconsin Timber Producers Association

National Association of Conservation District

Natural Resources Services & Consulting

Oakwood Forestry Consulting

Pierre & Sweeney Lawyers

Pine River Lumber Co. 

Rhinelander Daily News

Ruffed Grouse Society

Sierra Club - Midwest Region

Steigerwaldt Land Services

Tappon-Ruetz Land Services, Inc.

Thilmany Paper Co.

Tigerton Lumber Co.

Triple “T” Enterprises, Inc.

University of Wisconsin, Green Bay College of 

Environmental Science

University of Wisconsin Extension

University of Wisconsin, Stevens Point School of 

Natural Resources

University of Wisconsin, Eau Claire

University of Wisconsin, Green Bay

University of Wisconsin, La Crosse

University of Wisconsin, Madison

Vilas County News Review

Wausau Paper Mills Co.

Whitetails Unlimited, Inc.

Wisconsin Audubon Society

Wisconsin County Forests Association

Wisconsin Forest Conservation Task

Wisconsin Forest Productivity Council

Wisconsin Paper Council

Wisconsin Woodland Owners Association

Wyoming

Central Wyoming College

Northern Rockies Conservation Cooperative

Northwest College

Sierra Club, Wyoming Chapter

University of Wyoming

Canada

Bioforest Tech., Inc.

Valent Biosciences

3M Canada Co Ltd

6.5 Libraries

Alabama

Alabama A&M University, J.F. Drake Memorial 

Library

Andalusia Public Library

Attalla-Etowah County Public Library

Auburn University at Montgomery Library

Baldwin County Library Cooperative

Cherokee County Public Library

Choctaw County Public Library

Clayton-Town & County Library

De Kalb County Public Library

Fayette County Memorial Library

Gadsden-Etowah County Library

Hale County Public Library

Huntsville-Madison County Public Library

Jacksonville State University, Houston Cole Library

Lawrence County Public Library
Chapter 6

Chapter 6 - Page 58

Macon County-Tuskegee Public Library

Marengo County Public Library

Marion-Perry County Library

Monroe County Public Library

North Shelby County Library

Phenix City-Russell County Library

Saint Clair County Library

Sumter County Library System

Troy State University, Library, Wallace Hall

Tuskegee University, Hollis Burke Frissell Library

University of Alabama, Amelia Gayle Gorgas Library

University of Alabama, Huntsville, Salmon Library

University of South Alabama, University Libraries

Washington County Public Library

Wilcox County Library

Alaska

Alaska Resources Library & Information Service

Anchorage Municipal Libraries

Arlis Library

University of Alaska, Anchorage Consortium Library

University of Alaska, Fairbanks, Elmer E. Rasmuson 

Library

University of Alaska, Southeast, Ketchikan Campus 

Library

University of Alaska, Southeast, William A. Egan 

Library

Z.J. Loussac Public Library

Arizona

Arizona State Library,

Clifton-Greenlee County Public Library

Cochise County Library District

Flagstaff City-Coconino County Public Library

Gila County Library District

Huachuca City Public Library

Maricopa County Library District

Mohave County Library District

Northern Arizona University Cline Library

Phoenix Public Library

University of Arizona Main Library

Yavapai County Library District

Yuma County Library District

Yuma County Library District Main Library

Arkansas

Arkansas State Library

Arkansas State University, Jonesboro, Dean B. Ellis 

Library

Central Arkansas Library System Main Library

Cleburne County Library

Conway County Library Headquarters

Craighead County & Jonesboro Public Library

Crittenden County Library

Dallas County Library

Faulkner County Library

Garland County Library

Grant County Library

Hot Spring County Library

Izard County Library-Melbourne

Jackson County Library

Lawrence County Library

Lonoke Prairie County Regional Library

Lyon College Mabee-Simpson Library

Montgomery County Library

Ouachita Baptist University Riley-Hickingbotham 

Library

Pine Bluff & Jefferson County Library System

Poinsett County Library

Pope County Library System

Saline County Public Library

Southern Arkansas University Magale Library

University of Arkansas, Mullins Library

University of Arkansas - Little Rock, Ottenheimer 

Library

University of Arkansas, Monticello Library

University of Arkansas - Pine Bluff, Watson Memorial 

Library

University of Central Arkansas Torreyson Library

University of the Ozarks Robson Library
Mailing List

Chapter 6 - Page 59

California

Alameda County Library

Alpine County Free Library

Anaheim Public Library

Berkeley Public Library

Butte County Library

California Institute of Technology Library

California Polytechnic State University Library

California State Library

California State University, Bakersfield Library

California State University, Chico Library

California State University, Dominguez Hills Library

California State University, Fresno Library

California State University, Fullerton Library

California State University, Hayward Library

California State University, Long Beach Library

California State University, Los Angeles Library

California State University, Northridge Library

California State University, San Bernardino Library

California State University, San Marcos Library

California State University, Stanislaus Library

Carlsbad City Library

Claremont University Center Library

Colusa County Free Library

Contra Costa County Library

College of the Sequoias Library

County of Los Angeles Public Library

Daly City Public Library

Del Norte County Library District

Fresno County Free Library

Fresno County Genealogical Society Library

Fresno County Public Library

Humboldt County Library

Humboldt State University Library

Kern County Library

Kern County Library System Library

Lake County Library

Los Angeles Public Library System

Los Gatos Public Library

Madera County Library

Marin County Free Library

Mariposa County Library

Mendocino County Library

Merced County Library

Modoc County Library

Mono County Free Library System-Northern Region

Mono County Free Library System-Southern Region

Monterey County Free Libraries

Nevada County Library

Oakland Public Library

Orange County Public Library

Pepperdine University Library

Plumas County Library

Riverside Public Library

Sacramento Public Library

San Benito County Free Library

San Bernardino County Library

San Diego County Library

San Diego Public Library Library

San Diego State University Library

San Francisco State University Library

San Jose State University Library

San Leandro Public Library

San Luis Obispo City-County Library

San Mateo County Library

Santa Clara County Free Library

Santa Clara University Library

Santa Cruz City-County Library System Headqua

Shasta County Library

Solano County Library System Library

Sonoma County Public Library

South San Francisco Public Library

Stanford University Library

Stanislaus County Free Library

Sutter County Free Library

Thousand Oaks Library

Tulare County Library System

Tuolumne County Free Library

University of California, Berkeley Library

University of California, Berkeley School of Law 

Library

University of California, Davis Library

University of California, Irvine Library

University of California, Los Angeles Library

University of California, Riverside Library

University of California, San Diego Library
Chapter 6

Chapter 6 - Page 60

University of California, Santa Barbara Library

University of California, Santa Cruz Library

University of Redlands Library

University of San Francisco Library

University of Southern California Library

Ventura County Library Services Agency

Whittier College Library

Yolo County Library

Yuba County Library

Colorado

Adams State College Library

Arkansas Valley Regional Library Service System

Baca County Public Library

Colorado College Library

Colorado School of Mines Library

Colorado State Library

Colorado State University Library

Conejos County Library

Delta County Public Library

Denver Public Library

Dolores County School Public Library

East Morgan County Library District

Elbert County Library

Garfield County Public Library System

Gilpin County Public Library

Grand County Library District

Gunnison County Public Library

Jackson County Public Library

Jefferson County Public Library

Kiowa County Public Library

Lake County Public Library

Las Animas-Bent County Public Library

Mesa County Public Library District

Mesa County Public Library District

Mesa State College Library

Mineral County Regional Library

Moffat County Libraries

Park County Public Library

Pitkin County Library

Pueblo Library District

Regis University Library

San Miguel County Public Library District No

Southern Peaks Public Library

University of Colorado, Boulder Library

University of Colorado, Colorado Springs Library

University of Colorado, Denver Library

University of Denver Library

University of Northern Colorado Library

University of Southern Colorado Library

Connecticut

Central Connecticut State University Library

Connecticut College Library

Connecticut State Library

Danbury Public Library

Eastern Connecticut State University Library

Hartford Public Library

Quinnipiac University Library

Southern Connecticut State University Library

Teikyo Post University Library

Trinity College Library

Union Free Public Library

University of Connecticut Library

University of New Haven Library

Wesleyan University Library

Western Connecticut State University Library

Yale University Library

Delaware

Appoquinimink Community Library

Delaware State University Library

Delaware Technical & Community College Library

New Castle County Public Library System

Sussex County Department of Libraries

University of Delaware Library

District of Columbia

American University Library

Catholic University of America Library

District of Columbia Public Library

Georgetown University Library

Library of Congress Library

U.S. Department of the Interior Departmental Library

U.S. Senate Library
Mailing List

Chapter 6 - Page 61

Florida

Alachua County Library District Headquarters

Bradford County Public Library

Brevard County Library System

Broward County Division of Libraries

Calhoun County Public Library

Collier County Public Library

Columbia County Public Library

Flagler County Public Library

Florida Atlantic University Library

Florida Institute of Technology Library

Franklin County Public Library

Hardee County Public Library

Hendry County Library System

Hernando County Public Library System

Highlands County Library System

Holmes County Library

Indian River County Main Library

Jacksonville Public Library

Jacksonville University Library

Jefferson County Public Library

Lake County Library System

Lee County Library System

Manatee County Public Library System

Martin County Library System

Monroe County Public Library

North Indian River County Library

Osceola County Library System

Palm Beach County Genealogical Society Librar

Palm Beach County Library System

Pasco County Library System

Putnam County Library System

Saint Johns County Public Library System

Saint Lucie County Library System

Seminole County Public Library System

State Library of Florida

Stetson University Library

Tampa-Hillsborough County Public Library

Taylor County Public Library

Union County Public Library

University of Florida Library

University of Miami Library

University of North Florida Library

Volusia County Public Library

Wakulla County Public Library

Washington County Library

Georgia

Athens Clarke County Library

Atlanta-Fulton Public Library

Bartow County Library System

Brooks County Public Library

Brunswick-Glynn County Regional Library

Chattooga County Library

Clayton County Library System

Coastal Plain Regional Library

Cobb County Public Library System

Dekalb County Public Library

Dougherty County Public Library

Elbert County Library

Fitzgerald-Ben Hill County Library

Gwinnett County Public Library

Hart County Library

Houston County Public Library System

Jefferson County Library System

Newton County Library

Thomas County Public Library System

Troup-Harris-Coweta Regional Library

Hawaii

Hawaii State Library

Hawaii State Library System

Hawaii State Public Library System

Idaho

American Falls District Library

Boundary County District Library

Camas County Public Library

Cambridge Community Library

Clark County District Library

East Owyhee County Library District

Idaho Falls Public Library

Idaho State Library

Idaho State Talking Book Library

Jefferson County District Library
Chapter 6

Chapter 6 - Page 62

Latah County Free Library District

Menan County District Library

Midvale District Library

Oneida County Free Library

Illinois

Brown County Public Library District

Calumet City Public Library

Champaign Public Library

Decatur Public Library

Evansville Public Library

Henderson County District Library

La Grange Park Public Library District

Northern Illinois Library System

Putnam County Public Library District

South County Public Library District

Warren County Public Library District

Indiana

Allen County Public Library

Bartholomew County Public Library

Benton County Public Library

Crawford County Public Library

Fayette County Public Library

Fulton County Public Library

Greensburg-Decatur County Public Library

Indianapolis-Marion County Public Library

Jackson County Public Library

Jasper County Public Library

Jay County Public Library

Monroe County Public Library

Morgan County Public Library

Newton County Public Library

North Madison County Public Library System

Ohio County Public Library

Owen County Public Library

Pike County Public Library

Saint Joseph County Public Library

Scott County Public Library

Spencer County Public Library

Sullivan County Public Library

Switzerland County Public Library

Tell City-Perry County Public Library

Tippecanoe County Public Library

Tipton County Public Library

Union County Public Library

Iowa

Altoona Public Library

Ames Public Library

Cumberland Public Library

Dubuque County Library

Public Library of Des Moines

Scott County Library System

State Library of Iowa

Union Public Library

Woodbury County Rural Library

Kansas

Coffey County Library

Finney County Public Library

Graham County Public Library

Grant County Library

Hamilton County Library

Johnson County Library

Kearny County Library

Kiowa County Library

Linn County Library District Three

Linn County Library District Two

Linn County Library District One

Morton County Library

Scott County Library

Sheridan County Library

Stanton County Library

Stevens County Library

Wichita County Library

Kentucky

Allen County Public Library

Boone County Public Library

Bowling Green Public Library

Boyd County Public Library

Boyle County Public Library

Breathitt County Library
Mailing List

Chapter 6 - Page 63

Breckinridge County Public Library

Calloway County Public Library

Campbell County Public Library 

Carroll County Public Library

Casey County Public Library

Clark County Public Library

Clinton County Public Library

Crittenden County Public Library

Cumberland County Public Library

Cynthiana-Harrison County Public Library

Daviess County Public Library

Edmonson County Public Library

Estill County Public Library

Fleming County Public Library

Floyd County Public Library

Gallatin County Public Library

Garrard County Public Library

Grant County Public Library District

Graves County Library

Grayson County Public Library

Hancock County Library

Hardin County Public Library

Harlan County Public Library

Hart County Public Library

Henderson County Public Library

Henry County Library

Hopkins County-Madisonville Public Library

Jackson County Public Library

Johnson County Public Library

Kenton County Public Library

Kentucky Department for Libraries

Knott County Public Library

Knox County Public Library

Laurel County Public Library District

Lawrence County Public Library

Lee County Public Library

Leslie County Library

Lewis County Public Library

Logan County Public Library

Louisville Free Public Library

Lyon County Public Library

Madison County Public Library

Magoffin County Library

Marion County Public Library

Marshall County Public Library

Mason County Public Library

McCreary County Public Library District

Meade County Public Library

Menifee County Public Library

Mercer County Public Library

Metcalfe County Public Library

Middlesboro-Bell County Public Library

Monroe County Public Library

Nicholas County Public Library

Ohio County Public Library

Oldham County Public Library

Owen County Public Library

Pendleton County Public Library

Pike County Public Library District

Powell County Public Library

Pulaski County Public Library

Robertson County Public Library

Rockcastle County Library

Rowan County Public Library

Russell County Public Library

Scott County Public Library

Spencer County Public Library

Taylor County Public Library

Todd County Public Library

Trimble County Public Library

Union County District Library

Washington County Public Library

Wayne County Public Library

Whitley County Library

Wolfe County Library

Louisiana

Beauregard Parish Public Library

Bossier Central Library

Caldwell Parish Library

Catahoula Parish Library

Claiborne Parish Library

Concordia Parish Library

Jefferson Parish Library Department

Lafourche Parish Public Library

Saint Martin Parish Library
Chapter 6

Chapter 6 - Page 64

Saint Mary Parish Library

Southern University, New Orleans

State Library of Louisiana

Tangipahoa Parish Library

Union Parish Library

Washington Parish Library System

West Carroll Parish Library

Winn Parish Library

Maine

Maine State Library

Portland Public Library

Maryland

Baltimore County Public Library

Calvert County Public Library

Caroline County Public Library

Carroll County Public Library

Dorchester County Public Library

Frederick County Public Libraries

Harford County Public Library

Kent County Public Library

Montgomery County Department of Public Library

Prince George’s County Memorial Library System

Queen Anne’s County Free Library

Saint Mary’s County Memorial Library

Somerset County Library System

Talbot County Free Library

Washington County Free Library

Worcester County Library

Massachusetts

Boston Public Library

Cambridge Public Library

State Library of Massachusetts

Worcester Public Library

Michigan

Alcona County Library System

Alpena County Library

Ann Arbor District Library

Bay County Library System

Benton Harbor Public Library

Charlevoix Public Library

Crawford County Library

Detroit Public Library

Dickinson County Library

East Lansing Public Library

Flint Public Library

Gladwin County Library

Jackson District Library

Kalkaska County Library

Lapeer County Library

Macomb County Library

Manistee County Library

Mason County District Library

Menominee County Library

Mideastern Michigan Library Cooperative

Missaukee District Library

Monroe County Library System

Muskegon County Library

Oakland County Library

Otsego County Library

Petoskey Public Library

Shiawassee County Library

St. Clair County Library

Washtenaw County Library

Minnesota

Anoka County Library

Anoka County Library System

Carver County Library

Douglas County Library

Fergus Falls Public Library

Great River Regional Library

Hennepin County Library

Hennepin County Library System

Jackson County Library System

Marshall-Lyon County Library

Martin County Library

Milaca Community Library

Montevideo-Chippewa County Public Library

Nobles County Library & Information Center

Ramsey County Public Library

Rock County Community Library
Mailing List

Chapter 6 - Page 65

Saint Paul Public Library

Scott County Library System

Washington County Library

Watonwan County Library

Mississippi

Biloxi Public Library

Hancock County Library System

Harrison County Library System

Humphreys County Library System

Laurel-Jones County Library

Madison County-Canton Public Library

Marks-Quitman County Library

Marshall County Library System

Mississippi Library Commission

Neshoba County Public Library

Noxubee County Library System

Raleigh Public Library’

Sunflower County Library System

Tallahatchie County Library

Union County Library

Warren County-Vicksburg Public Library

Washington County Library System

Missouri

Bollinger County Library

Camden County Library District

Cass County Public Library

Christian County Library

Dallas County Library

Daviess County Library

Douglas County Public Library

Gentry County Library

Grundy County-Jewett Norris Library

Henry County Library

Jefferson County Library

Kansas City Public Library

Livingston County Library

Mcdonald County Library

Mercer County Library

Miller County Library Service Center

Mississippi County Library

Missouri State Library

Neosho Newton County Library

New Madrid County Library

Ozark County Library

Putnam County Public Library

Ralls County Library

Ray County Library

Saint Charles City County Library District

Saint Clair County Library

Saint Louis County Library

Scotland County Memorial Library

Springfield-Greene County Library

Stone County Library

Sullivan County Public Library

Washington County Library

Wright County Library

Montana

Blaine County Library

Chouteau County Library

Daniels County Free Library

Flathead County Library

Garfield County Library

Glacier County Library

Glasgow City County Library

Liberty County Library

Lincoln County Public Libraries

Livingston-Park County Library

Meagher County City Library

Mineral County Public Library

Missoula Public Library

Montana State Library

Petroleum County Community Library

Prairie County Library

Roosevelt County Library

Rosebud County Library

Sheridan County Library

Stillwater County Library

Thompson-Hickman Free County Library
Chapter 6

Chapter 6 - Page 66

Nebraska

Garfield County Library

Grant County Library

Hooker County Library

Lincoln City Libraries

Logan County Library

Omaha Public Library

Rock County Public Library

South Sioux City Public Library

Thomas County Library

Nevada

Carson City Library

Douglas County Public Library

Elko-Lander-Eureka County Library System

Humboldt County Library

Las Vegas-Clark County Library District

Lincoln County Library

Lyon County Library System

Mineral County Public Library

Pershing County Library

Storey County Public Library

Washoe County Library

White Pine County Library

New Hampshire

East Rochester Public Library

Nashua Public Library

Unity Free Public Library

New Jersey

Atlantic City Free Public Library

Atlantic County Library

Burlington County Library

Camden County Library System

Cape May County Library

Cumberland County Library

Gloucester County Library System

Hunterdon County Library

Mercer County Library

Monmouth County Library

Morris County Library

Somerset County Library

Sussex County Library System

Trenton Public Library

Warren County Library

New York

Albany Public Library

Brooklyn Public Library

Broome County Public Library

East Rochester Public Library

La Grange Association Library

Monroe County Library System

Mount Vernon Public Library

New York Public Library

Onondaga County Public Library

Queens Borough Public Library

Schenectady County Public Library

South Country Library

South New Berlin Free Library

Tompkins County Public Library

North Carolina

Alexander County Library

Asheville-Buncombe Library System

Avery County Morrison Public Library

Bladen County Public Library

Brunswick County Library

Burke County Public Library

Caldwell County Public Library

Carteret County Public Library

Catawba County Library

Columbus County Public Library

Cumberland County Public Library & Informatio

Currituck County Public Library

Dare County Library

Davidson County Public Library System

Davie County Public Library

Duplin County Library

Durham County Library

Edgecombe County Memorial Library

Forsyth County Public Library

Franklin County Library
Mailing List

Chapter 6 - Page 67

Gates County Library

Graham County Public Library

Granville County Library System

Greene County Public Library

Harnett County Public Library

Havelock-Craven County Library

Haywood County Public Library

Henderson County Public Library

Hertford County Library

Hoke County Public Library

Iredell County Public Library

Jackson County Public Library

Kinston-Lenoir County Public Library

Lee County Library

Lincoln County Public Library

Macon County Public Library

Madison County Public Library

Mcdowell County Public Library

Mitchell County Public Library

Montgomery County Public Library

New Hanover County Public Library

Onslow County Public Library

Pamlico County Library

Pender County Library

Perquimans County Library

Person County Public Library

Robeson County Public Library

Rockingham County Public Library

Scotland County Memorial Library

Stanly County Public Library

State Library of North Carolina

Transylvania County Library

Tyrrell County Public Library

Union County Public Library

Warren County Memorial Library

Washington County Library

Watauga County Public Library

Wayne County Public Library, Inc.

Wilson County Public Library

Yancey County Public Library

North Dakota

Bottineau County Public Library

Cavalier County Library

Kidder County Library

Mckenzie County Public Library

Morton County Library

Stutsman County Library

Valley City-Barnes County Public Library

Ward County Public Library

Ohio

Adams County Public Libraries

Akron-Summit County Public Library

Brown County Public Library

Carroll County District Library

Champaign County Library

Clark County Public Library

Clermont County Public Library

County District Library

Cuyahoga County Public Library

Delaware County District Library

Findlay-Hancock County Public Library

Geauga County Library System

Geauga County Public Library

Greene County Public Library

Guernsey County District Public Library

Highland County District Library

Logan County District Library

Medina County District Library

Meigs County District Public Library

Mercer County District Library

Monroe County District Library

Paulding County Carnegie Library

Perry County District Library

Pickaway County District Public Library

Pickerington Public Library

Portage County District Library

Preble County District Library

Public Library of Cincinnati

Stark County District Library

State Library of Ohio

Troy-Miami County Public Library
Chapter 6

Chapter 6 - Page 68

Tuscarawas County Public Library

Warren-Trumbull County Public Library

Washington County Public Library

Wayne County Public Library

Williams County Public Library

Oklahoma

Beaver County Pioneer Library

Cherokee-City-County Public Library

Choctaw County Library

Delaware County Library

Latimer County Public Library

Metropolitan Library System

Nowata City-County Library

Tulsa City-County Library

Woodward Public Library

Oregon

Baker County Public Library

Clackamas County Library

Corvallis-Benton County Public Library

Crook County Library

Deschutes County Library System

Douglas County Library System

Gilliam County Library

Grant County Library

Harney County Library

Hood River County Library

Jackson County Library Services

Jefferson County Library

Josephine County Library System

Klamath County Library

La Grande Public Library

Malheur County Library

Multnomah County Library

The Dalles-Wasco County Library

Tillamook County Library

University of Oregon Library

Washington County Cooperative Library Service

Pennsylvania

Adams County Library System

Allegheny County Library

Altoona Area Public Library

Bedford County Library

Blair County Library System

Bradford County Library System Headquarters

Bucks County Free Library

California University of Pennsylvania Louis L. 

Manderino Library

Cambria County Library System & District Cent

Cameron County Public Library

Carnegie Library of Pittsburgh

Centre County Library & Historical Museum

Chester County Library

Clearfield County Public Library Federation

Columbia County Traveling Library

Crawford County Federated Library System

Dauphin County Library System

Delaware County Library System

Dickinson College Waidner-Spahr Library

Dormont Public Library

Erie County Public Library

Fayette County Library System

Forest County Library

Franklin County Library System

Free Library of Philadelphia

Fulton County Library

Greene County Library System

Huntingdon County Library

Juniata County Library, Inc.

Lackawanna County Library System

Lancaster County Library

Lawrence County Law Library

Lebanon County Library System

Mifflin County Library

Montgomery County-Norristown Public Library

Perry County Law Library

Pike County Public Library

Snyder County Library

Somerset County Library

State Library of Pennsylvania

Sullivan County Library
Mailing List

Chapter 6 - Page 69

Wayne County Public Library

West Chester Public Library

York County Library System

Rhode Island

Rhode Island State Library

South Carolina

Beaufort County Library

Berkeley County Library

Calhoun County Library

Cherokee County Public Library

Chester County Library

Chesterfield County Library

Colleton County Memorial Library

Darlington County Library

Dillon County Library

Dorchester County Library

Fairfield County Library

Florence County Library

Georgetown County Library System

Greenville County Library

Harvin Clarendon County Library

Horry County Memorial Library

Kershaw County Library

Lancaster County Library

Laurens County Library

Lee County Public Library

Marion County Library

Oconee County Library

Orangeburg County Library

Pickens County Library System

Richland County Public Library

Saluda County Library

Spartanburg County Public Libraries

Spartanburg County Public Library

Sumter County Library

Union County Carnegie Library

Williamsburg County Library

York County Library

South Dakota

Bennett County Public Library

Custer County Library

Grant County Public Library

Hand County Library

Hyde County Library

Jackson County Library

Potter County Free Public Library

Sully County Library

Tripp County Library-Grossenburg Memorial

Tennessee

Benton County Library

Bledsoe County Public Library

Blount County Public Library

Bolivar-Hardeman County Public Library

Carroll County Library

Chattanooga-Hamilton County Bicentennial Library

Cheatham County Public Library

Chester County Public Library

Claiborne County Public Library

Coffee County Lannom Memorial Public Library

Coffee County-Manchester Library

Decatur County Library

Decatur-Meigs County Library

Dickson County Public Library

Fayetteville-Lincoln County Public Library

Fentress County Public Library

Franklin County Library

Gibson County Memorial Library

Giles County Public Library

Hancock County Public Library

Hardin County Library

Houston County Public Library

Humphreys County Public Library

Johnson County Public Library

Knox County Public Library System

Lawrence County Public Library

Lebanon-Wilson County Library

Lewis County Public Library

Macon County Public Library

Marshall County Memorial Library
Chapter 6

Chapter 6 - Page 70

Maury County Public Library

Memphis-Shelby County Public Library

Moore County Public Library

Mount Juliet-Wilson County Public

Overton County Public Library

Perry County Public Library

Pickett County Library

Public Library of Nashville & Davidson

Putnam County Library System

Rutherford County Library System

Scott County Public Library

Sequatchie County Public Library

Sevier County Public Library

Smith County Public Library

Stewart County Public Library

Sullivan County Public Library

Tipton County Public Library

Unicoi County Public Library

Washington County - Jonesborough Library

Wayne County Public Library

Williamson County Public Library

Texas

Alamo Area Library System

Amarillo Public Library

Arlington Public Library

Bandera County Library

Bee County Public Library

Brazoria County Library System

Brooks County Library

Calhoun County Library

Callahan County Library

Carson County Public Library

Central Texas Library System

Chambers County Library System

Coke County Library

Cooke County Library

Corsicana Public Library

Crane County Library

Crockett County Public Library

Crosby County Library

Dallas Public Library

Dawson County Library

Delta County Public Library

Dickens County-Spur Public Library

Dimmit County Public Library

East Parker County Library

Ector County Library

Floyd County Library

Ford County Library

Fort Bend County Libraries

Gaines County Library

Goliad County Library

Hansford County Library

Harris County Public Library

Haskell County Library

Hemphill County Library

Hidalgo County Library System

Hockley County Memorial Library

Hood County Public Library

Houston Public Library

Huntsville Public Library

Irion County Library

Jackson County Library

Jefferson County Library

Karnes County Library System

Kaufman County Library

Kent County Library

Kimble County Library

Kinney County Public Library

Lamb County Library

Leon Valley Public Library

Live Oak County Library

Llano County Library System

Madison County Library

Martin County Library

Memphis Public Library

Mitchell County Public Library

Montgomery County Memorial Library System

Newton County Public Library

North Texas Regional Library System

Oldham County Library

Rains County Public Library

Reagan County Library

Real County Public Library

Red River County Public Library
Mailing List

Chapter 6 - Page 71

Reeves County Library

Roberts County Library

Rockwall County Library

San Patricio County Library System

San Saba County Library

Schleicher County Public Library

Scurry County Library

Shackelford County Library

Sherman County Public Library

Somervell County Library

Starr County Public Library

Stonewall County Library

Sutton County Library

Sweetwater County-City Library

Swisher County Library

Terrell County Public Library

Texas State Law Library

Upshur County Library

Upton County Public Library

Val Verde County Library

Waco-Mclennan County Library

Waller County Library

Wilson County Library

Winkler County Library

Yoakum County Library

Yoakum County Library

Zapata County Public Library

Utah

Davis County Library

Grand County Public Library

Morgan County Library

Salt Lake County Library System

San Juan County Library

Uintah County Library

Utah State University Natural Resources Research 

Library

Wasatch County Library

Washington County Library

Weber County Library

Vermont

Aldrich Public Library

Bennington Free Library

Landgrove Public Library

Orwell Free Library

State of Vermont Department of Libraries

Thetford Town Library

Virginia

Appomattox Regional Library

Arlington County Department of Libraries

Botetourt County Library

Campbell County Public Library

Charlotte County Library

Chesterfield County Public Library

Colonial Heights Public Library

County of Henrico Public Library

Culpeper County Library

Cumberland County Public Library

Fauquier County Public Library

Fluvanna County Library

Franklin County Library

Jefferson-Madison Regional Library

Loudoun County Public Library

Norfolk Public Library System

Nottoway County Library

Orange County Library

Pittsylvania County Public Library

Powhatan County Public Library

Pulaski County Library

Rappahannock County Library

Roanoke County Public Library

Russell County Public Library

Shenandoah County Library

Tazewell County Public Library

Washington County Public Library

York County Public Library

Washington

Asotin County Library

Forest Resource Library

King County Library System
Chapter 6

Chapter 6 - Page 72

Pacific Marine Technology Library

Pierce County Library System

Spokane County Library District

Spokane Public Library

Stevens County Rural Library District

Tacoma Public Library

Walla Walla County Library

Washington State Law Library

Washington State Library

Whitman County Rural Library

West Virginia

Alpha Regional Library

Brooke County Public Library

Cabell County Public Library

Calhoun County Public Library

Clarksburg-Harrison Public Library

Doddridge County Public Library

Elkins-Randolph County Public Library

Fayette County Public Libraries

Five Rivers Public Library

Grant County Library

Greenbrier County Library

Hamlin-Lincoln County Public Library

Hampshire County Public Library

Hardy County Public Library

Jackson County Library

Kanawha County Public Library

Keyser-Mineral County Public Library

Marion County Public Library

Mason County Library System

Mingo County Library

Monroe County Public Library

Morgan County Public Library

Ohio County Public Library

Pendleton County Public Library

Pleasants County Public Library

Pocahontas County Free Library

Putnam County Library

Raleigh County Public Library

Ritchie County Public Library

Roane County Public Library

Summers County Public Library

Taylor County Public Library

Tyler County Public Library

Upshur County Public Library

West Virginia University Evansdale Library

Wyoming County Library System

Wisconsin

Cumberland Public Library

Dodge County Library Service

Door County Libraries

Florence County Library

Marathon County Public Library

Marinette County Consolidated Public Library

Menominee Tribal County Library

Milwaukee County Federated Library System

Oneida County Mailbox Library

Portage County Public Library

Western Taylor County Public Library

Wisconsin State Law Library

Wyoming

Campbell County Public Library

Carbon County Library System

Converse County Library

Crook County Library

Goshen County Library

Hot Springs County Library

Johnson County Library

Laramie County Library System

Lincoln County Library

Natrona County Public Library

Niobrara County Library

Park County Library System

Platte County Public Library

Sheridan County Fulmer Public Library

Sublette County Library

Sweetwater County Library System

Teton County Library

Uinta County Library

Western Wyoming Community College Library
Mailing List

Chapter 6 - Page 73

6.6 Individuals

Alabama

Timothy Boyce

 J. Wayne Brewer

Jim Hyland

Adam Jackson

Tomm Johnson

Bill Moore

M. N. “Corky” Pugh

Bruce Shupp

Debi Summersell

Alaska

Randy Bates

Candace Beery

Charles Bell

Joel Bennett

Tim Blust

Corrie Bosman

Roger Burnside

Andre Camara

Niles Cesar

Melanie Duchin

Hugh Durham IV

Ken Fisher

John Fox

Sylvia Geraghty

Bob Gorman

Owen Graham

Mike Holloway

Eric Hummel

Jeff Jahnke

Edgar Jenks

Christina Jewett

Jason Loos

Gary Morrison

Don Muller

Mary Pete

Frank Rue

Joseph Sebastian

Marlo Shedlock

Fred Sorenson

Dave Sturdevant

Tara Sweeney

Doug Warner

Ronald Wolfe

Arizona

Thomas Abrams

Gale & Vesta Aldrich

Peggy Alexander

Theresa L. Allison

Andrea Anderson

Fred Anderson

Kathryn Anderson

John Anhold

Scott and Toni Arena

Donald Arganbright

Don Arkin

Rachel Aschmann

Janina Austin

Jean & Trevor Avenett

Charles Babbitt

Pat Baca, Jr.

Paula Bachman-Williams

John Baker

Theodore Barbone

Ariel Barfield

Annie Barva

Beverly Bass

Charles & Mar Bast

Kay Bawden

Don Beck

Paul Beier

Cher Beilfuss

William Beloret

Fray H. Belshe

Robert Bennet

Patsy Bennett

Linda Bentley

Don & Linda Bentley

David Bertelsen

Rebecca Berton

Andy Bessler
Chapter 6

Chapter 6 - Page 74

Jean Calhoun

A. P. Camps

Stephen Canning

Lawrence Carlson

Larry Carlson

Royce Carlson

Ron Carswell

L. M. Case

Eligio & Anna Castillo

Mr. Bob Celaya

Don Chaney

Frank Chapman

Holden Chase

Alan Chatfield

Helen & Placi Chavez

Donna Chesner

Gary Christensen

Earl Christian

Paula Chronister

Ray & Sabine Cichlar

Becki Cimadevilla

Lamar & Pat Clark

Letha Cline

Christopher Cloud

Vonna Cluff

Joanne Cockerill

Mark Colby

Rose Coleman

A. Consolo

Mel Copeland

Brenda Corkin

Tania Corliss

Walter & R. Craig

Cullen Cramer

Cole Crocker-Bedford

Lauerl Crosby

Martha Crosby

Laura Cummings

Diane Cusack

Rudy Dankwort

Diana Davis

Vada L. Davis

Loren Dawn

Porter Dean

Roy & Joanne Dechant

Vic DeFrancesco

Tom DeGomez

Marc Delany

Eileen DeLauer

Shanna & Rod Denault

Charlie Denton

Sonya Diehn

Carl Dietrich

Linda Dills

Robert Dink, Jr.

John & Annie Dunn

Michael Durgain

Dane & Marlen Dyrland

Auroroa B. Eagar

Dustan & Cynthia Eagar

Roland & Ruth Eagar

Karen Earley

Barbara Edwards

David E. Eisley

Dan Ellerbroek

Dillie A. Ellis

Dock Ellis

Glen L. Ellison

Terrance Ely

Larry & Suzan Ely

Lloyd Engel

Rick Erman

Corwin Estes

Paula Fan

Christine Farney

Albert Farr

Dewey H. Farr

W. John Faust

James & Glend Finch

R. J. Finch

Brent Finley

Holly Finstrom

Heather Fitar

Bobby & Linda Fite

Chase & Lance Fite

M. Fitzgerald

Carl Beyerhelm

Jessie Bhangoo

Bettina Bickel

Dennis J. Bigelow

Robert Bigelow

Rulon & Lucie Bigelow

Evelyn Billo

Toni Bish

Michael Bissontz

Joe Blaszczak

Edward Blumer

W. Brent & Joann Bogdanski

Larry Borden

Roy Boss

Melville S. Bowers

Virginia Gail Bowers

Curtis Bradley

Richard Bradshaw

Christine & M. Brady

Susan Brandes

Clait Braun

Ruth Brawdy

Robert Breen

Sheri Brennan

Ronald Brill

Scott Brill

ArthurandAnne Britt

Bob Broscheid

Grant & Innis Brown

Jarrod Brown

Kelli Brown

Rayanne & Jam Brown

Tonda Brown

James Brown

James A. Bruder

Adam Burdick

Jeff Burgess

Brian & Sarah M. Burnett

Carolyn Burns

M. & Jackie Butler

Paul Byers

John Caid

William Calder
Mailing List

Chapter 6 - Page 75

Jenny Flynn

Chris Ford

Tyler Forman

B. & Kendie Foster

Cheryl Foster

Zee Fowler

Linda Franks

Steve Franks

Dirk Frauenfelder

Rae Frederickson

Jim Frich

Bruce Friedemann

Kevin J. Fritz

Eric Gabel

Sharon Galbreath

Charles Gallagher

Victor Gallegos

Joe Ganey

Karen R. Garley

Ted Gartner

Connie Gartner

Robert Gay

Roxanne George

Roxane George

Don Gerrard

Mary Ann Gibbons

Lee Roy Gibson

M. Gice

Curtis & Jean Gillespie

Brandon Gilliam

D. Gilliam

Joyce & Stone Gilliam

Robert & Elvia Gillies

Pamela Goldman

Kenn Goldman

Robert Gomez

Jerry Gonzales

Jose Gonzalez

Joni Goode

Donald & Evel Goodman

John & Karen Goodwin

Kenneth Gouker

Penny Govedich

Scott Graff

L. Graham

Richard Grapp

Jesse Greenberg

Michael Gregory

D. Grhoton

Tim Griffith

Anastacia Gutierrez

Brent Hall

Douglas & Michelle Hamblin

Lana Hamblin

Marion Hamblin

Cheles Hancock

Holly Hancock

Tina Hancock

Robbie Hannawacker

Byron & Roz Harding

Mr. & Mrs. Jeff Harper

N. Harper

Ryan Harper

Shae Lynn Harris

Leo Hartke

Syble & LaMar Hartley

Meredith Hartwell

Cynthia Hartzell

Ron Harvey

Michael Haseltine

C. A. & Wilma Haught

Jim & Sherri Haught

Kay Hauser

Mike & Joelle Hauser

Steven & Christine Heap

Craig Hegel

Merrill Hentz

Robert Herdliska

Terry Heslin

Tom Hicks

Cody Hill

Ron Hill

Sky Hilts

Dawn Hines

Sidney Hirsh

Orne H. Hiscox

David Hodges

Bobbie Holaday

William Holden

Jim & Karen Horton

Caroline Hotaling

Michael Houghtaling

Sally L. Hulsey

Michael Humphrey

Michael & Pat Humphrey

Kip & Alicia Hunsaker

Sean Hunsaker

T. & A. Hunsaker

Joseph & Mari Hunter

Peter Ianchiou

Mike Ingram

Wayne D. Iverson

Glen Jacobs

David Jason Jaramillo

Manuel Jaramillo

Jeff Jenness

Shane & Paula Johnson

Bonnie Johnson

Richard & Fran Jones

Mitzi Jones

T. J. Jordan

Oweta Josleyn

Cecelia Juszczak

Michelle Kaczynski

Charlie Kane

Joseph Kantauskis

Jennifer Katcher

John Keane

Dennis Kee

Bart Keehn

Charlotte Keller

Patricia Kelly

M. Keoppen

June A. & Don W. Kimble

Barbara & W. Kinman

Edward Kirsten

Larry Kivela

Henry J. Klassen

Keith Kleber
Chapter 6

Chapter 6 - Page 76

Barbara & Fred Klug

Edwin Knochel

Julie Kornmeyer

Kimberly Kuehnert

Paul Kuenstler

Gene & Barbar Laird

Jason Laird

Pierre Landau

Bob Landis

Linda Laney

G. James Langello

Kimberly Larsen

Alicia Larson

Susan Lascelles

Calvin Lash, Jr.

Arthur N. Lee

Danny Lee

Katherine Lee

Dan Leeds

Kent LeSueur

N. LeSueur

Vera LeSueur

Robin Levenworth

Lainie Levick

Jerry Levitt

Glen Lewis

H. Lewis

Timothy Lewis

Tracy Liston

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S. Petersen

Michael Plumer

B. Poliska

B. Reed

R. Reed

James Reid

Kitlyn Rescinito

Virginia Schick

Debbie Scott

Cindy Sheridan

Stan Sipp

Peter Skuba

S. Smith

Peggy Snow

Leellen F. Solter

Martha Speir

Theodore Steck

Randy Stephens

Russell Sutton

Richard Thom

Dale Thurber

Eric F. Ulaszek

Amber Urban

Carissa Van Nevel

Brad Virden

Stephen Warble

Roger Weber

Nancy Wedow
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Patrick Weicherding

Mark Willi

Jason J. Zylka

Indiana

Mrs. Parks Adams

Charles Anderson

Wilford Barlett

Joe Bruggenschmidt

Jim Brummett

Harold Bruner

Glenn Burham

James Burke

Richard Burt

Kevin Clark

Dan Cole

Louise Cummings

Joseph Davison

D. Eby

J. Ellis

Clay Faller

Burnell Fischer

C. Gallowitch

J. Golod

Donald Goodwin

Steven Goodwin

Frank Gottbrath

Anthony Grossman

Harry Halstead

William Kautz

Gregory Koontz

Norman Lamunion

Ronnie Linville

P. Marshall

Philip T. Marshall

Dan Meisberger

James Morris

G. Moughler

Brian Mueller

J. Ogle

Jim W. Pleasant

Tom Rathert

Phil Reid

Charles Rush

Dan Schmoker

Z. Smith

Zachery M. Smith

Jill Strawder

Jeffrey Swackhamer

Jim Sweeney

Steward Turner

Lloyd Vanderstreek

Bruce Wakeland

Robert D. Waltz

D. Yaninek

Iowa

Jim Ahrens

Allan Beck

Mike Brandrup

Mark Dungan

Wayne Fuhlbrugge

Merlin Glade

Jay Gold

Denny Haugen

Bill Haywood

Roger Jacob

Robert Kassmeyer

Steve Lekwa

Al Manning

Darwyn Peters

Robin D. Pruisner

Norman Riemersma

David Sell

Brian Wade

Mark Webb

Pat Wenke

Dan & Vicki Fogle

Kansas

Raymond Aslin

B. F. Barker

Pete Garfinkel

Russell A. McKinney

Kentucky

C. Thomas Bennett

Allen Hale

Carl Harper

Emily Loomis

Leah W. MacSwords

Mike Madryga

Jennifer Reed

Lynne Rieske-Kinney

Sara Sanders

Jeffrey W. Stringer

Barbara Werner

Louisiana

Esther Boykin

Richard Goyer

Louis Heaton

James Jenkins, Jr.

Peter Martinat

Edward Rivera

Bill Spitzer

Will Waring

Alvin Wells

David Wilbur

Robert Zinn

Maine

Ohn Ackerman

John Ackley

Richard Aishton

Jeff Albert

Kennard Allen

David Allen

Linda Alverson

Phillip Andrews

Richard Arbour

Mark Armstrong

Walt Armstrong

Dennis Arsenault

James Austin

William Barron

Mark Beauregard
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Michael Benjamin

Peter Beringer

Heather Bessey

Richard Blodgett

Leon Blood

Wendall Bradford

Jack Brake

Jerri Brandt

David Brann

J. Brennan

Michael Brown

Raymor Brown

Barbara Brusila

Robert Bryan

Stephen Bumps

Harold Burnette

William Calderwood

Albert Carey

Ernest Carle

Iver Carlson

Stanley Carpenter

Katherine Carter

Robert Chadbourne

Robert Chandler

Pete Chase

John Churchhill

Michael Cline

John Colannino

Phillip Conkling

Terri Coolong

Roger Coolong

Andrea Corbett

John Cote

Robb Cotiaux

Hewlette Crawford

Stephen Croman

Chester Curtis

Fernald Curtis

Mike Cyr

Brian Dangler

Debra Davidson

Theresa Davis

Christopher Deane

Mark Deschene

Herbert Dickey

Jack Dirkman

John Dirkman

Kevin Doran

David Dow

Harold Dow

Phillip Dow

Amanda Dow-Smith

William Drisko

Michael Duddy

Martin Duffany

Eugene Dumont

Geneva Duncan

Ralph Dunn

Emile Dupuis

Harold Dwyer

James Ecker

Stephen Elliot

Kirby Ellis

Leon Emery

Grant Estelle

Peter Farnsworth

Floyd Farrington

Robert Fenderson

Richard Finlay

Robert Finlay

Emma Finn

Gary Fish

Stephen Follette

Norman Forbes

Clifton Foster

Chester Gage

Tim Gammell

Robert Gammons

Arthur Garland

Stephen Gettle

Ann Gibbs

Robert Giffin

Dale Gilman

Daniel Gilmore

Walter Gooley

Douglas Gray

Glenn Gurney

Dana Hall

Frederick Hallenberg

Michael Hammond

Geoffrey Hancock

James Harris

M. E. Hartley

Martin Hartley

Patsy Hartley

Michael Hartung

David Harvey

Hugh Hastings

Bob Haynes

Thomas Henderson

Kerry Herbert

Frederick Herrick

Mindie Hesketh

Eric Holden

Dave Hopkins

Patricia Hopkins

Fred Huntress

Gary Inman

Wayne Jackson

Lonnie Jandreau

William Jarvis

Roger Johndro

Linda Johns

Marc Johnson

Albert Johnston

Alfred Johnson

Anita Johnson

Bela Johnson

Peter Johnson

Rick Jones

Mike Kankainen

Alan Kimball

Charles Kinney

Richard Kircher

Peter Klachany

David Knupp

John Kochi

Joseph Koller

Gloria Krellman
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Mitch Lansky

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John Leslie

Robert Leso

Phillip Levasseur

David Libby

Robert Liske

Robert Locke

Ronald Locke

John Long

Robert Lumppio

Michael MacDonald

Sandra MacGown

David Maddocks

Alan Magrath

Douglas Mahan

Tristan Manchester

David Manski

Douglas Marston

Mark Martin

Todd Massey

Sue May

Earl McCheeney

M. McClean

Malcolm McConnell

Izzy McKay

Jack McMullen

James McMullen

Paul Memmer

Mark Michaud

Paul Miller

Brian Milligan

Roger Milligan

Brooks Mills

John Mills

Robert Moore

Gary Morse

Keith Morse

Christopher Murdock

Kenneth Nawfel

Scott Nelson

Thomas Nelson

William Newcomb

A. Newell

Merle Parise

Scott Pease

Stephen Pelletier

D. Perkins

Christopher Polson

Neil Postlewaite

Andrew Pottle

Gerald Poulin

Eugene Putnam

Don Quellette

Kevin Raye

Steve Reynolds

Linwood Rideout

Merle Ring

Bruce Ripley

George Ritz

Hugh Roak

Jonathan Robbins

Wallace Robbins

Michael Rochester

Robert Rochester

David Rock

David Rocque

Fred Rooney

Megan Ropiak

Edwin Rosso

Paul Rudd

Daniel Russell

Michael Sackett

J. Sass

Wendall Saucier

Dave Schaible

Wilhelm Schloth

Timothy Scott

David Shaible

Randy Shaw

Charles Simpson

Carl Sjogren

Dan Smith

Robert Smith

Donald Soctomah

Brian Somers

Michael Spellman

Robert Spivey

Frank Spizuoco

Ellis Sprague

George Stadler

Susan Stetson

Timothy Stevens

Andrew Strachon

Kenneth Strickland

David Struble

Joel Swanton

Christopher Taylor

Theresa Tenney

Alan Thiebeault

Terry Thomas

Barry Tibbetts

Sam Timberlake

Irvin Tower

Everett Towle

Peter Tracy

James Trask

Joel Tripp

Theodore Tryon

Gail Tunstead

Daren Turner

Bob Umberger

Mark Vannah

Vite Vitale

Peter Volz

Carol Voss

Dave Walker

James Ward

David Warren

Dean Webster

Andrew Weegar

David Wellman

John Wenteel

Forest Weston

James Wheeler

Thomas Whitworth
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Carroll Willette

Maynard Wilson

Kenny Wing

Donald Winslow

Mark Woodbury

David Woodhead

Robert Wright

R. Alec Giffen

D. Eric Johns

James St. Pierre

Russell F. Roy

Ronald St. Saviour

Scott & Doree Olson

Maryland

Robert Albert

Lynn Alexander

Leslie Amtower

Ronald Antill

Jim Bailey

Glen Besa

Fred Bolton

Rusty Booker

Susan A. Bright

Scott Burroughs

Steward Callis

Ryan Carter

Joseph Clisham

John Cullom

John Davis

Henry Debruin

Chuck Divan

Richard J. Dolesh

Wade Dorsey

Alice Eastman

Nathan Erwin

Drew Foerster

Jeanne Frantz

Weyman P. Fussell

Jim Getter

Anne Harmeyer

Rex Harper

Harry Hartment

Joe Hautzenroder

Zoh Hieronimus

Martha Holdridge

John Houser

Rolf Hubbe

Kristin Iden

Andrea Illig

Clint Irwin

John & Linda Jacobs

John Jordan

Charles Keeley

John Kennedy

Jerry Kimmel

Steve Koehn

Steven Koehn

John Kowalski

Craig Kuhn

Margret Leary

Rodney Lipscomb

Robert Loomis

Chris Lynch

Melissa McDaniel

Todd McDonald

David Miller

Robert More

Phil Nester

Sonny Newhall

Dawn Parker

Bob Prettyman

Heidi Pringle

Bob Rabaglia

Melvin Reuber

Chuck Schneider

Larry Sharpe

Earl Sheble

Paul Shogren

Don Sisler

Richard Smucker

Mark Souterland

Warren Spencer

Marshall Stacy

Frank Stark

Henry Stasick

Jerry Stokes

Raymond Stralka

Kathleen Talman

Mark C. Taylor

Matt Taylor

Kevin W. Thorpe

Robert Tichenor

John Van Horn

Adrienne Venables

Carole Vila

Marie Walz

Kenneth Willets

Shawn Winterberg

Peter Wood

Len Wrabel

Amelia Wright

Massachusetts

Warren Archey

Suzanne Artemieff

Abigail Avery

Glenn Ayers

Kerri Belval

Norman Berberi

Charles Burnham

Susan Campbell

Paul Cole

John Deans

James Dennesen

Charles Diehl

James Dimaio

Alexandra Durbin

Tom Emerson

Burt Germond

Kenneth Gooch

Jeff Grove

Kerowyn Guillotte

Donna Hampson

Robert Hannon

Donald Harris

Jeff Hourdain

Mike Kiernan

Lisa Kroeber
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Michael Leonard

Aaron Maia

Victor C. Mastro

Lawrence May

Phyllis Michalewich

Richard Mytkowicz

Robert Perschel

John E. Rice

David Richard

David Sanderson

Jack Savastano

Michael Sikora

Grechen Smith

Edward Syrjala

Charles Thompson

Rick Tormala

Lucille Traina

Peter Tucker

Kimberly Walsh

Robert Watjen

Mary Wigmore

Michigan

Steve Alguire

John Arend

C. Arsnoe

B. Austin

Ralph Babcock

Dan Bailey

Samuel Bailor

J. Barense

Jim Barry

Andrew Barton

William Beach

Martin Blackedge

Gary Boushelle

Maurice Bracken

M. Brackenbury

Darryl Braun

Dan Braun

Peter Buehler

Christopher Burnett

Robert Caouette

Gordon Cole

Richard Cooper

James Crowfoot

Tracy Curlee

D. Dantuma

B. Dayton

Leanne Dijok

Homer Earl

Lee Eavy

M. Eddy

J. Edgerly

L. Eisbrenner

Lee Ekstrom

Leon Erbe

George Ferrar

Walter Fifelski

Marlene Fluharty

Ken Ford

Suzanne Gartz

Dennis Good

C. Gorsuch

Earl Gorsuch

Kathryn Griffith

Gerald Grossman

Shawn Hagan

Ron Hansen

John Hanson

Kevin Harsch

W. Hatfield

William Hatfield

P. Henry

Bob Heyd

J. Hill

John P. Hill

Max Holden

Bernard Hubbard

Don Ingle

Larry Jobson

Stanley Johns

Jim & Patricia Johnson

Carl Johnson

Martha Jones

Norbert Karmann

C. Keefe

Gerald Keiser

Duane Kenaga

Phil Kline

B. Kroll

Dane Lamb

R. Larowe

C. Lenchek

Frank Lendzion

Larry Lindenberg

R. Lintemuth

Jack Lockwood

J. Loncar

Rory Mattson

James Mccumber

D. Mckay

M. Meriwether

Cynthia Merrow

Kip Miller

A. Mrozinski

G. Patchen

M. Philip

Mike Philip

T. Plachta

Richard Potts

M. Quesnel

Don Quick

K. Rauscher

K. Reis

Robert Rohn

David Roose

Forrest Ruppert

J. Ryan

Eino Sainio

Ken Salo

John Santeiu

Frank Sapio

A. Schiffer

Albert Schiffer

Kathy Schiffer

M. Schiffer

Lisa Schoppmann
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Chapter 6 - Page 92

Josheph Schott

Walter Selski

Ronald Sievertson

Shaun Smakal

John Sosnowski

L. Sperry

Thomas Stadt

Edward Stulberg

Russell J. Sutton

Mike Tansy

Howard Taylor

M. Ticehurst

Jack Titus

Donald Tracey

Jack Tucker

G. Voyle

A. Wallace

Alicia H. Wallace

Richard Wallace

G. Wheeler

John Wiggins

Richard Wilcox

Gary Willis

J. Winkler

John Witter

Sylvester Wood

Dennis Worst

Christopher Wright

Ralph Zandt

Minnesota

Kenneth Anderson

Lillian Baker

David Banta

S. Burks

John Calgaro

Jennifer Callahan

Michael Carroll

Valerie J. Cervenka

Marty Christensen

M. Connor

Michael D. Connor

Mike Connor

Ronald Daigle

P. Deerwood

Robert Despot

Gene Dressely

Carlos Eberhardt

Jeff Fellows

Donald Ferguson

Geir Y. Friisoe

Thomas Gearhart

Richard Gitar

Cathryn Greene

Ralph Greilig

Keith Hanson

Dennis A. Haugen

T. Heyer

Roger Holmes

Ralph Hovind

Mark Johnson

Alan Jones

R. A. Knoll

Mert Lammi

Raymond Lopresto

Brian Lutenegger

Greg Magnuson

Susan McGuire

Wade Mier

Jim Mohler

Steve Nelson

Kevin O’Brien

Glen Olson

Ralph Olson

Arlet Phillips

Kevin Proescholdt

Arthur Reese

Kathryn Robbins

M. Roberts

Geart Searfoss

Michael Shepard

Don Small

E. Karsten Smelser

D. Solum

Lance Sorenson

George Stever

John Swanson

Michael Swift

Kimberly Thielen Cremers

Maynard Underbakke

Carl Vogt

Mississippi

John Corban

Jeffery Head

Sidney Malone

Thomas A. Monaghan

Tamara Muller

Evan Nebeker

Sam Polles

Robert Simonds

James Sledge, Jr.

Missouri

Burl Ashley

Randall Bolyard

M. Brown

Scott Brundage

Michael Collins

K. Combs

Jerry Conley

John Dwyer

Troy Gordon

Amy Grubert

Kerwin Hafner

Randall Herberg

John Keesey

Robert Krepps

Jay Law

Rob Lawrence

Gregory Linn

Ronald Lumb

James McClure

Sarah Messbarger

Robert Miley

Charles Phillips

M. Roling
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Cheryl Rosenfeld

Don Schultz

Robert Simonds

Clell Soloman

Delores Ward

Rebecca Weisser

Montana

Gary Adams

Vick Applegate

Tom Corse

Bruce Erickson

Roger Gowan

Patrick J. Graham

Cathey Hardin

Bob Harrington

Lee Harry

Bruce Higgins

Jack Holmes

Sharon Klinkhammer

Steve Kohler

Sue MacMeekan

Bill Michels

George Nickas

Jane Olson

Wes Paulson

Christopher Reichert

Jerry Sass

Steve Slaughter

Kathy Tribby

Lynn Vrlanic

Val Walker

John Weinert

M. Whalen

Michael Wood

Nebraska

Gary Hergenrader

Susan Schimmer

Ruth Wusk

Nevada

R. C. Bechtel

Valerie Buchanan

Frank R. Cervantes

Gail Durham

Gail Ferrell

Trevor & Tracy Ganske

Shelly Germann

Steven Glimp

Doug Hunt

Anna Keyzers

Robert King

P. C. Martinelli

Marian Mckenzie

Leanne Miller

Steve Robinson

Patrick Rucker

Marilyn Tomkins

Robert Vaught

Roy & Ruby Venghams

Sean & Erin Wallace

New Hampshire

Charles Baylies

George Bell

Putnam Blodgett

Jennifer Bofinger

Charles Bond

Philip Bryce

Michael Burke

Norman Charest

Richard Chase

Raymond Conley

Gibb Dodge

Tom Durkis

Kevin Evans

Rick Evena

Peter Farrell

Donna Gamache

Buhrman Garland

Walter Graff

Frank Hammond

Bruce Jacobs

Jonathan Janis

Kenneth Jordan

Keith Kidder

Charles Koch

Fred Kocher

Ted Lacey

Kyle Lombard

Quentin Mack

John Martinson

Bruce McAllister

Brooks McCandlish

Joseph McKeever

Dennis McKenney

Charles Moreno

David Noyes

David Olson

Daniel Reed

Pete Renzelman

Peter Rhoades

Jay Seavey

Bruce Sloat

Norma Sorgum

David Thompson

Robert Todd

Wayne E. Vetter

John Twichell

John Violette

Steve Walasewicz

Wayne Young

New Jersey

Diane Allen

James Barresi

Paul Barrett

Joseph Bateman

Diane Beatty

Judson Bennett

Paul Berezny

Deborah Boerner

Melani Bolyai

Ian Borden

Paul Borokhov
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William Brash

Tom Brodde

Thomas Bullock

Gene Burks

Hugh Carola

Dennis Chandler

Nancy Coleman

Paul Cowie

Joseph Dunn

David Edelman

G. Ettenger

Galen E. Ettinger

Lorens Fasano

Robert Fimber

R. Fine

David Finley

Richard Goodenough

Ted Gordon

Heather Gracie

William Grundmann

Christina Harrigan

Cora Hartshorn

Kris Hasbrouck

Curtis Helm

Neil Hendrickson

Frank Hennion

Madelyn Hoffman

Lewis Howell

Christina Hurd

Mark Hynson

Matthew Immergut

Craig Kane

Constance Katzenbach

George Paul Koeck

Mary Lamielle

John Linson

Lucine L. Lorrimer

Gary Lovalo

Linda Mack

Wendy Malmid

Timothy Matthews

Steve Maurer

John Mayyyott

Bob McDowel

William Metterhouse

Nenneth Meyers

David Moore

Robin Murphy

Steven Panter

Linda Price

Stuart Rich

Paul Rodriguez

Nicole Roskos

Barbara Sachau

Dale Schweitzer

Jack Shuart

Robert Sidor

Timothy Slavin

Sunil Somalwar

Hank Suydam

Kenneth Taaffe

Linda Tatem

Douglas Tavella

Dena Temple

Luke Templin

Bob Williams

J. Worrell

George Zimmerman

Joe Zoltowski

New Mexico

Doug Abbott

Ann Alexander

Craig Allen

Debra Allen-Reid

Josefina Alvarez

Jim Bailey

James Bailey

Randy Balice

Mary Ann Baruch

Hugh E. Bearup

Joanie Berde

Jacque E. Blackman

Butch Blazer

Doug & Penny Bogart

Jess Bowman

Billy O. Branman

Jerry Brock

Kelly L. Bryan

Ysabel Campbell Luecke

Henry Carey

Elizabeth G. Chapman

Betty Jane Curry

Leslie Davis

Nelson Denman

Rich Detry

Ellen Dietrich

George Duda

Robert & Lill Dunn

Ron Ensminger

Karen Epperson

Judith Espinosa

John Fowler

Ric Frost

Delbert L. Fulfer

Charles Fuller

Sid Goodloe

Frank E. Gorskey

Randall Gray

Carrie Green

William Gross

Wayne Gyulai

Darrol L. Harrison

Bruce Hayward

Martin T. Heinrich

David Henderson

Sam Hitt

Mark Hoak

Larry Hughes

Timothy Humphrey

Abe Jacobson

Thomas Jervis

Sarah Johnson

Jennifer Johnson

Jack T. Jordan

David Keller

Ed Kelly

Suedeen Kelly

Richard Khanlain
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Anthony Ladino

Gabe Lopez

Willie Lucero

Leonard Lucero

Richard V. Ludwig

Paul Luehrmann

Greg Magee

Anne Malone

Jerry Maracchini

Freddie Martin

Aron Martinez

Michael Massey

Pat Mathis

Fred C. & Wil May

Ann McCampbell

Mr. Jerry McCrea

Tracy McFarland

Julie McIntyre

Stephen Mergault

Mark Miller

Verna Miller

Tim Mitchusson

William H. Moore

Patrick Morandi

Alex Mueller

Nance Neskaukas

Rick Norris

James Norwick

Douglas B. Osborn

Manuel Pacheco

Doug Parker

Jeff Pierce

Gregory Pollak

Joanna Prukop

Don Rauch

Harold Reynolds

Eric Roybal

Deleen Ruebush

Richard Ryan

Calvin J. & B. Salars

Ben Sanchez

Pat D. Sanchez

Stephani Sandoval

Mike Sauber

Joe Savage

Melissa Savage

Hazel M. Shuck

Jack Stauder

Mary Steele

Carol Sutherland

Fred Swetnam

John Talberth

Dierde Tarr

Chancel Teague

Alex Thal

David Ther

Jerry Sue Thompson

Bruce Thompson

Lauro Vanegas

Arlene Walsh

Rhonda Ward

Jan Ward

Dave Wilgues

Marjorie Williams

Rex Wilson

Wade Worrell

Raleigh Zellers

New York

Scott Aldridge

Clifford Asdal

Richard Audette

Jordan Bain

Todd Baldwin

Wilford Bartlett

Richard Bell

Nadya Carolyn Bennett

Glen Berger

Bobbie Blowers

Herb Boyce

Bernard Braun

John Burton

Steve Callahan

Jerry A. Carlson

Marcia Carlson

Steve Collins

William Connally

James Danoff-Burg

Robert Davies

Richard DeBadts

Brian Dermody

Jane Difley

Dana Drake

John Earl

James Farrar

Richard Fox

I. M. Frellsen

Ronald Frisbee

Bill Galdstone

Edward Gammon

Cindy Garfield

David Gee

John Gibbs

Jody Gray

Ann E. Hajek

Glen Hampton

David Hawke

John Herrington

Ken Hotopp

Robert Howard

Sabrina Islam

Calvin John

Jack Karnig

Susan Keister

John Koshorek

David Kotzle

Ted Kozlowski

Peter Levatich

Lowell Lingo

Wendy Lochner

Lisa Maybee

Warren McKeon

John Miller

Richard Monti

Robert Moore

Bob Mungari

Aprille Nace

Gary Nelson
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Evan Nichols

Rodney Nielsen

Robert Peterson

Richard Pettus

Barlow Rhodes

Jim Roberts

Monique Roberts

William Roe

Ed Schaffer

James Schuler

David Seymour

Phillip Siarkowski

David Sinclair

Vernon St. Louis

John Stratton

Robert Synowiez

Patricia Testa

Christopher Thompson

Louis Tirtito

Wayne Tripp

Sarah Waite

Fred Wilhelm

Rodman Williams

Alex Wipf

Allyn Wright

Michael Wright

Jennifer Zeh

David Zlomek

North Carolina

Stanford Adams

Weedie Barnard

Phil Bell

Erica Blackwell

Nancy Brown

Richard Brown

Leo Bunce

Kevin Carpenter

Gene Cross

Tracy Davids

Brian Fireman

Joann Fireman

Paul Gallimore

S. Andrew Gerry

John Ghent

Lessie Mae Graves

Robert Giles

Harris Gruber

Fred P. Hain

Rick Hamilton

Edward Harrison

Brian Heath

Phillip Heatherly

James B. Jones

John Kent

Donna Leonard

M. Leonard

Olivia Lim

D. Martin

Mike Massey

A. Mustian

James Padgett

Brett Pendergrass

Ethel Pittman

Derek L. Puckett

Robert Reiman

Donald F. Rogers

Stephen P. Schmidt

Aron Sebastian

Terry Seyden

Jill Sidebottom

Walton Smith

Deborah Steward

Robert Thatcher

Ron Thigpen

Michael Thompson

Robert Trickel

Diana Underhill

Ralph Willard

James Yount

North Dakota

John Brauner

Dean Hildebrand

Dave Hirsch

Larry Kotchman

Joe Maxwell

Ohio

Dave Adkins

Daniel Balser

Joel Berry

Pamela S. Blackburn

Robert Boley

Michael J. Budzik

B. Burke

Brian Burke

Richard Cappell

Greg Crandall

John Dorka

Robert Endebrock

Judy Fink

David Fleischer

Jeff Frontz

Tammy Frye

Margaret Garwook

D. Geglein

Stephanie Glazer

Robert Hampel

Margaret Harwreak

Betty Jean Herner

Tim Humprey

A. Lacy Johnson

Lacy Johnson

Robert Lamoreaux

William Lebold

Michael Littlejohn

Frank Luppino

Steve McKee

Thomas Morban

K. Niese

Galen Oakes

Peter Oros

Gilbert Papsy

B. Ramsey

Richard Ramsey

Deb Reed

C. Richards
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Chris Richards

Delores Rogers

Amanda Schackow

James Scheetz

Siobhan Wolf Shaffer

J. Slavicek

James M. Slavicek

Mary Smallsfeed

Kathleen Smith

Jim Stafford

Thomas Stanley

A. Stone

James Suhanic

Stanley Swierz

Robert Tapeman

Algimantas P. Valaitis

Becky Violey

Joseph Vitti

Oklahoma

John Burwell

Sancho M. Dickinson

Jacob Frank

Jack Gobin

Pat Mcdowell

Bill Ross

Oregon

Hilary Abraham

Sherry L. Adams

Dennis Ades

Thomas P. Ady

John Aguirre

Steven Akehurst

Nate Alexander

Tena Alvarez

Ed Alverson

Bob Amundson

Arvid Andersen

Debbie Anderson

Don E. & Pat Anderson

Donald J. Anderson

Jim Anderson

Zach Anderson

Martin Andre

Bret & Doris Armacost

Jerome & Jane Arnold

Larry Aschenbrenner

Kathy Askren

Mary M. Atkinson

Alan Ayres

Dale Badrick

Barry B. Bai

Ric Bailey

James Bailey

Tim Bailey

Ric Bailey

James Baker

Jerry Baker

Bruce Ball

Cindy Banzer

Jamie Barbour

Stanley Barg

Lyle Barkman

John R. Barnes

Rick Barnes

Dan Barnett

Richard Barnette

Donald & June Barnum

John Barry

Edward & Bob Bartell

Conrad Bateman

John & Patricia Bates

Robert A. Batty

Byron Beach

Joy Belsky

LeRoy Bennett

Leslie Benscoter

Ken Benson

Ted Berg

Lars Bergstrom

A. Berier

Thomas Berkemeier

Daniel Berman

Donald Bettis

Ann Bettman

Gary Betts

John & Linda Biehl

Ken Bierly

Mack Birkmaier

Tom Birkmaier

Cal & Lorna Birrer

Dan Bishop

David Bishop

Neil Bjorklund

Erin Black

Scott Black

Christopher & J. Blake

Larry Bliesner

Richard Bloom

Dick Blum

Mike Blumm

Mark & Vicki Bolton

Donald & Donna Bond

Dale Bonnell

Tracy Bosen

Cassandra Botts

Gerald Bowerly

Grant Bowerman

Sue Bowers

Sandy Boyce

Jack Boyd

Jim Boyle

David Boyles

Durward L. Boyles

Albert Bradford

James & Doris Brady

Mike & Kathy Brandis

Bruce Brandt

William D. Brand

La Dora Brasel

David M. Braun

John & Lynne Breese

Lynne Breese

Eugene Brick

Dave Bridgwater

Chris Broadfoot

D. Brodie
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Chapter 7

Glossary

Figure 7-1. A respirator prevented inhalation of wing scales and fine hairy 

particles from gypsy moth life stages. 
Chapter 7 Glossary

Figure

Figure 7-1. A respirator prevented inhalation of wing scales and 

fine hairy particles from gypsy moth life stages....................................Cover
Glossary

Chapter 7 - Page 

Terms are defined as they pertain to this Supplemental 

Environmental Impact Statement (SEIS).

A

absorption — process by which the agent is able 

to pass through the body membranes and enter the 

bloodstream. The main routes by which toxic agents 

are absorbed are the gastrointestinal tract, lungs, and 

skin

acetylcholine — compound released at nerve endings, 

active in the transmission of the nerve impulse

acetylcholinesterase — enzyme that occurs in nerve 

endings and prevents accumulation of acetylcholine; 

acetylcholinesterase inhibition results in acetylcholine 

accumulation, which impairs the nervous system 

acinar-cell adenomas — type of benign tumor

actinomycete — any bacterium in the order 

Actinomycetales, which contains filamentous 

branching bacteria of the genera Actinomyces and 

Streptomyces

active ingredient — (a.i.) toxic part of an insecticide 

formulation

acute exposure — single exposure or multiple 

exposures occurring within a short time frame (24 

hours or less)

acute toxicity — potential of a substance to cause 

injury or illness in a single dose or in multiple doses 

over a period of 24 hours or less

adenoma — benign epithelial tumor; glandular

additive effect — combined effect of two chemicals 

is equal to the sum of the effect of each chemical 

alone. The effect most commonly observed when two 

chemicals are administered together is an additive 

effect

adjuvant(s) — formulation factors used to enhance the 

pharmacological or toxic effect of the active ingredient

absorption — tendency of one chemical to adhere to 

another material

adverse-effect level — (AEL) signs of toxicity 

that must be detected by invasive methods, external 

monitoring devices, or prolonged systematic 

observations. Symptoms that are not accompanied by 

grossly observable signs of toxicity 

AEL — acronym for adverse-effect level.

aerobes — organisms that require oxygen.

aesthetic damage — undesirable change in appearance

Agricultural Research Service (ARS) — USDA 

agency that develops the means to protect trees in 

forests, parks, yards, and other nonforest environments; 

conducts research to support activities against the 

gypsy moth

a.i. — abbreviation for active ingredient

alkaline — having a high pH; a basic solution, 

compared with an acidic solution

allergic reaction — situation where a pre-exposure of 

the chemical is required to produce the toxic effect via 

an antibody

alopecia — hair loss

alternative — one possible way to accomplish a 

proposed action; a way to manage the gypsy moth in 

the United States
Chapter 7

Chapter 7 - Page 

amino acids — relatively simple carbon-nitrogen 

molecules that are the subunits of proteins

amphiphod — any of the various small crustaceans 

in the order Amphipoda, with laterally compressed 

bodies found primarily in aquatic habits; examples are 

sandhoppers, beach fleas and skeleton shrimp 

anaerobes — organisms that do not require oxygen

analogy to other compounds — using data on one set 

of compounds to predict the activity of another set of 

compounds

anemia — decrease in the concentration of red blood 

cells in whole blood

Animal and Plant Health Inspection Service

— (APHIS) joint-lead agency for this environmental 

impact statement on the gypsy moth; the USDA 

Agency that enforces national quarantine, coordinates 

with States on the National Gypsy Moth Survey, 

provides assistance to States to eradicate isolated 

infestations of the gypsy moth on 640 acres or less, 

develops new methods to improve gypsy moth 

quarantine and eradication practices, and conducts 

technology transfer activities 

anthelmintic — compound used to rid an organism of 

parasitic worms

antibodies — large protein molecules that interact with 

antigens and deactivate antigens 

antigen — substance capable of inducing an immune 

response

APHIS — acronym for Animal and Plant Health 

Inspection Service

aplastic — pertaining to or characterized by aplasia–

the lack of development of an organ or tissue, or of the 

cellular products from an organ or tissue

aplastic anemia — form of anemia that is difficult to 

treat

ARS — acronym for Agricultural Research Service

arthropods — large group of invertebrate animals that 

includes insects, spiders and crustaceans

artificial spread — spread of the gypsy moth by other 

than natural means, for example, by insect life stages 

attaching to and being moved on recreational vehicles, 

automobiles, nursery stock, outdoor household articles, 

and cargo

Asian strain — refers to strains of the gypsy moth 

originating in the Far East, which have some females 

that can fly, and may have the capacity to establish in a 

broader host range, be larger, and hatch earlier than the 

European strain

assay — a test (noun); to test (verb)

atrophy — decrease in the size of a cell, tissue, or 

organ, often associated with exposure to a toxic agent

B

Bacillus thuringiensis (B.t.) — bacterium; found in 

most of the world useful in regulation and/or control 

of insect populations. This microorganism produces 

several agents (toxins) active against insects

Bacillus thuringiensis var. kurstaki (B.t.k.) — 

scientific name of a bacterium that is specifically 

pathogenic to caterpillars of many moths and 

butterflies; the active ingredient in biological 

insecticides sold under the trade names Dipel, Foray, 

and Thuricide 

basal area — cross-sectional area of a tree determined 

from the diameter of the trunk at breast height; the total 

area of ground covered by trees measured at breast 

height 

benchmarks — results of toxicological tests, such as 

LCD or EC50 values.

beneficial organism — any organism that eats, 

parasitizes, or regulates in some way populations of 

other organisms that are pests
Glossary

Chapter 7 - Page 

benign — not malignant, not recurrent, favorable for 

recovery

benthic — pertaining to the sea bed, river bed, or lake 

floor 

beta-exotoxin — proteinaceous toxin in some forms 

of B.t. that is mutagenic in mammals; this toxin is not 

present in B.t.k.

biliary — referring to bile, a substance in which many 

chemicals are eliminated from the body 

bioassay — determination of the relative strength 

of a substance (e.g., drug, insecticide) by comparing 

its effect on a test organism with that of a standard 

preparation

biodiversity — variety of life and its processes; 

includes the variety of living organisms, the genetic 

differences among them, and the communities and 

ecosystems in which they occur

biologically sensitive — term used to identify a group 

of individuals who, because of their developmental 

stage or some other biological condition, are more 

susceptible than the general population to a chemical or 

biological agent in the environment

biomass — total weight, volume, or energy equivalent 

of organisms in a given area 

biota — plants and animals

BIU — acronym for billion international units

B.t.k. — abbreviation for Bacillus thuringiensis var. 

kurstaki

C

cancer potency parameter — model-dependent 

measure of cancer potency (mg/kg/day) over lifetime 

exposure; often expressed as a q, which is the upper 95 

percent confidence limit of the first dose coefficient (q,) 

from the multistage model

canopy — uppermost layer of foliage in forest 

vegetation, formed by the crowns of trees

carcinogen — chemical capable of inducing cancer

carcinoma — malignant tumor

carrier — in commercial formulations of insecticides 

or control agents, a substance added to the formulation 

to make it easier to handle or apply

caterpillar — soft-bodied larva of the gypsy moth or 

other moth, butterfly, or sawfly 

cell-mediated response — response originating from 

materials within the cell, in contrast to a humoral 

response

cfu — acronym for colony forming units

chironomid — ecologically important group of aquatic 

insects belonging to the family Chironomidae (order 

Diptera), often occurring in high densities and diversity, 

and feeding on a great variety of organic substrates; 

important prey of most aquatic predators

chitin — hard substance made of a complex 

carbohydrate (acetyl glucosamine) similar to cellulose; 

main component in the skin (cuticle) of insects, spiders, 

and crustaceans

cholinergic — refers to nerve cells that release 

acetylcholine

cholinesterase — group of enzymes that degrade 

acetylcholine and similar compounds. Cholinesterases 

that occur in nerve tissues have a clear function. Other 

cholinesterases, such as those occurring in red blood 

cells or plasma, do not have a clear function but are 

used as indicators of insecticide exposure

chromatography — method of separating chemicals 

prior to quantitative analysis

chronic exposure — long-term exposure studies 

often used to determine the carcinogenic potential of 

chemicals; these studies are usually performed on rats, 

mice, or dogs and extend over the average lifetime of 

the species; for example, chronic exposure for a rat is 2 

years
Chapter 7

Chapter 7 - Page 

chronic toxicity — adverse biologic response, such 

as mortality or an effect on growth or reproductive 

success, resulting from repeated or long-term (equal 

to or greater than 3 months) doses (exposures) of a 

compound, usually at low concentrations

circadian rhythm — influence of the time of day on 

the rate of metabolism of foreign compounds, often 

observed in a given animal species; a variation in 

the metabolic rate often correlated with variations in 

endocrine functions, as influenced by the light-dark 

cycle to which the animal is exposed

cladoceran — small aquatic crustaceans in the order 

Cladocera; water fleas

coliforms — bacteria that indicate recent fecal 

contamination of water

colony forming unit (cfu) — index of bacterial levels 

in a medium such as air or water; a cfu represents 

a collection of a droplet or particulate from air that 

contains one or more viable spores or vegetative cells 

of B.t.k. 

common logarithm — common logarithm of a 

number, X, is defined as the number, Y, which when 

used as the exponent of 10 results in X. Thus, if X = 

101

, then the log of X is Y, which is often written using 

the notation, log(X) = Y

community — association of potentially interacting 

plants and/or animals, more or less distinguishable 

from other such associations, usually defined by the 

nature of their interaction or the place in which they 

live

compliance agreement — written agreement between 

APHIS Plant Protection and Quarantine and a person 

who grows, handles, or moves regulated articles to 

comply with APHIS regulations 

confounders — term used in discussions of studies 

regarding human populations (epidemiology studies) 

to refer to additional risk factors which, if unaccounted 

for in a study, may lead to erroneous conclusions

congenital — refers to conditions present at birth, 

regardless of their cause

conidium — asexual spore produced by fungi (pl. 

conidia)

conjugation — in metabolism, a linkage of one 

molecule with another; common step in the elimination 

of many chemicals from the body

conjunctiva — thin mucous membrane that lines the 

eyelids 

conjunctivitis — inflammation of the membrane that 

lines the eyelids

connected actions — exposure to other chemical and 

biological agents in addition to exposure to a treatment 

agent used to control gypsy moth

connective tissue — tissue that binds together and 

supports the various structures of the body

contaminants — for chemicals, impurities present in 

a chemical-grade chemical; for biological agents, other 

agents that may be present in a commercial product

control — maintain or try to maintain a population 

density of insects or other undesirable animals below 

the point where injury to man’s interests occurs 

conspecific — belonging to the same species

cooperative project — management project conducted 

by a State or Federal agency, under agreement and 

with financial and technical assistance of the U.S. 

Department of Agriculture, to control forest diseases 

and insects such as the gypsy moth

Cooperative State Research, Education, and 

Extension Service — (CSREES) USDA agency that 

administers a research grants program, including gypsy 

moth research; plans cooperative research projects 

through the State Agriculture Experiment Station 

System and coordinates information and education 

activities

cooperator — State or Federal agency that enters into 

an agreement with the U.S. Department of Agriculture 
Glossary

Chapter 7 - Page 

to conduct a cooperative project

copepod — small marine or freshwater crustaceans in 

the class Copepoda, exhibiting great diversity in form 

and life history

corixid — insects in the family Corixidae (order 

Hemiptera); referred to as true water bugs

corneal opacity — cloudy area on the cornea

corneal ulcer — small area of damaged tissue on the 

surface of the eye 

corticosteroid — anti-inflammatory agent

corrosive effect — effect that causes visible 

destruction and alteration in tissue at the site of contact

cover type — vegetation, described in terms of its 

general form or dominant species, comprising the plant 

community in a given area

critical habitat — area determined by the U.S. Fish 

and Wildlife Service to be essential to the conservation 

of threatened or endangered species and that may 

require special management considerations or 

protection

crown condition — combination of tree crown density, 

coloration, leaf-rolling, mortality, or other factors that 

provide an indication of tree health

crustaceans — organisms such as crabs, lobsters, 

shrimp, crayfish, wood lice, pill bugs, and water fleas 

that have hard exoskeletons made of chitin, as do other 

arthropods

CSREES — acronym for Cooperative State Research, 

Education, and Extension Service

cumulative effects — effects attributable to 

exposure(s) that may last for several days to several 

months, or effects resulting from gypsy moth program 

activities that are repeated more than once during a 

year or for several consecutive years

cumulative exposure — exposure that may last for 

several days to several months or exposures resulting 

from program activities that are repeated more than 

once during a year or for several consecutive years

cytosolic — found in the cytoplasm of a cell 

D

dam(s) — female parent(s)

DDVP — abbreviation of the chemical name for 

dichlorvos—2,2 dichloroethenyl dimethyl ester 

phosphoric acid—an insecticide contained in some 

gypsy moth traps

defoliation — noticeable loss of foliage due to 

feeding by insects, such as gypsy moth caterpillars; 

light defoliation is normal background defoliation of 

less than 30 percent, moderate defoliation is 30 to 60 

percent, heavy defoliation is greater than 60 percent

defoliation survey — visually examining trees from 

the ground or the air, to detect defoliation

degradation — breakdown of a compound by physical 

and chemical or biochemical processes, into basic 

components with properties different from those of the 

original compound

degraded — broken-down or destroyed

degrees of freedom — number of data points minus 

the number of parameters in a model. For example, 

two points are required to define a straight line. In 

statistical jargon, using two points to define a straight 

line is fitting a two-parameter model with zero degrees 

of freedom

delimiting survey — using pheromone-baited traps to 

determine the approximate size of an infested area

delineation — a process used in slow the spread to 

estimate numbers and presence of gypsy moths in an 

area

delta-endotoxin — proteinaceous toxin in B.t.k. that is 

toxic to gypsy moth larvae

dermal — pertaining to the skin
Chapter 7

Chapter 7 - Page 

dermatitis — inflammation of the skin; characterized 

by redness, swelling, pain, and warmth

detection survey — using pheromone-baited traps to 

determine whether the gypsy moth is present and where 

delimiting may be necessary 

detritus — fragmented, particulate-organic matter 

resulting from the decomposition of plant and animal 

remains 

developed forest — privately owned forested 

residential areas

dichlorvos — another name for DDVP

diflubenzuron — active ingredient of chemical 

insecticide formulations sold under the trade name 

Dimilin®; acts as a growth regulator by interfering 

with chitin synthesis, preventing molting in gypsy 

moth caterpillars, some other immature insets, and 

crustaceans 

Dimilin® — trade name of diflubenzuron formulations 

registered for use against the gypsy moth 

DiPel — one of the commercial formulations of B.t.k.

dipteran — insect belonging to the order Diptera 

(meaning two wings), which includes flies and 

mosquitoes

direct effect — reaction of an organism after exposure 

to a chemical or non-chemical agent that is not 

medicated through another organism. For example, 

caterpillars that eat leaves with diflubenzuron on them 

fail to molt, and die as a result of their direct exposure 

to this insecticide; the direct effect of an unchecked 

gypsy moth infestation could be a change in species 

composition of trees

dislodgeable residues — residue of a chemical or 

biological agent on foliage as a result of aerial or 

ground spray applications, which can be removed 

readily from the foliage by washing, rubbing, or having 

some other form of direct contact with the treated 

vegetation

disparlure — synthetic version of the pheromone 

produced by female gypsy moths to attract males for 

mating

diuresis — increased urinary excretion

diurnal rhythm — normal changes in the body that 

occur during the day; most diurnal variations have been 

shown to be related to eating and sleeping habits

dominant trees — trees with crowns extending above 

the general level of the canopy and receiving full light 

from above and from the side 

dose — quantity of material that is taken into the body; 

dosage is usually expressed in amount of substance 

per unit of animal body weight, often in milligrams of 

substance per kilogram (mg/kg) of animal body weight, 

or other appropriate units; in radiology, the quantity of 

energy, or radiation absorbed

dose-response assessment — description of the 

relationship between the dose of a chemical and the 

occurrence or intensity of an effect

draft environmental impact statement — detailed, 

written statement of effects expected as a result of a 

major Federal action that is released to the public and 

other agencies for review and comment, as required 

under Section 102 (2)(c) of the National Environmental 

Policy Act 

E

EC50 — acronym for median effective concentrate

ecology — study of the interrelationships between 

living organisms and their environment

ecosystem — living organisms interacting with each 

other and with their physical environment, usually 

described as an area for which it is meaningful to 

address these interrelationships

ecosystem management — holistic approach to 

achieving productive healthy ecosystems by blending 

social, physical, economic, and biological needs and 

values
Glossary

Chapter 7 - Page 

eczema — form of dermatitis associated with swelling 

and redness of the skin

effect level — dose or concentration of a substance 

reported to have no harmful (adverse) effects on people 

or animals.

effector cell — cell stimulated by a nerve cell to 

effect a certain function. Examples include muscle and 

sensory cells

egg mass survey — visually examining an area in 

a systemic manner, either (1) outside the generally 

infested area, to obtain evidence that gypsy moths are 

present and reproducing, or (2) in an infested area, to 

assess the population density

EIS — acronym for environmental impact statement

empirical — refers to an observed, but not necessarily 

fully understood relationship; in contrast to a 

hypothesized or theoretical relationship

enantiomer — reference to molecules that are 

structurally identical except for differences in the threedimensional configuration

endangered species — Federal designation for any 

species that is in danger of extinction throughout all 

or a significant part of its range. The Federal list of 

endangered species is maintained by the Secretary of 

the Interior

endemic — something that is always present in a 

population but not always prevalent or present in high 

numbers; often applied to diseases or infestations

endospore — thick-walled body containing genetic 

material that forms inside the vegetative cell of some 

types of bacteria, including bacillus, under adverse 

conditions. When conditions improve, the endospore 

can develop into a vegetative cell

endpoints — components of an ecosystem that indicate 

its sensitivity to the type of disturbance expected 

from the gypsy moth or treatments; five endpoints 

were selected for the ecological risk assessment: 

nontarget organisms, forest condition, water quality, 

microclimate, and soil fertility and productivity

Entomophaga maimaiga — scientific name for a 

fungus that causes disease in gypsy moth caterpillars

environmental analysis — investigation of alternative 

actions and their predictable environmental effects 

through a systemic interdisciplinary approach, which 

ensures the integrated use of the natural and social 

sciences and the environmental design arts in planning 

and in decision making that may have an impact on the 

human environment

Environmental Assessment — (EA) a concise public 

document that a Federal agency prepares under the 

National Environmental Protection Act (NEPA) to 

provide sufficient analysis and evidence for either a 

finding of no significant impact or preparation of an 

environmental impact statement 

Environmental Impact Statement — (EIS) a 

detailed public document written by a Federal agency 

to disclose significant environmental impacts that 

would result from a planned action and used to make 

decisions about the action

enzyme — biological catalyst; a protein produced by 

an organism itself, which enables the splitting (as in 

digestion) or fusion of other chemicals

Ephemeroptera — order of aquatic insects including 

mayflies

epidemiology — branch of science that deals with 

the incidence, distribution, and control of disease in a 

population 

epidermis — outermost layer of the skin

epizootic — occurrence of a disease in animals that is 

widely prevalent and spreads rapidly

eradication — strategy of eliminating an isolated 

infestation of the gypsy moth

erythema — name applied to redness of the skin 

produced by congestion of the capillaries, which may 

result from a variety of causes
Chapter 7

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erythrocyte — red blood cell

European strain — strain of the gypsy moth 

historically found in Western Europe and the original 

source of the North American population, which has 

females that do not fly

evaluation — gypsy moth survey to determine the 

need for treatment or to determine the effectiveness of 

treatment

exclusion — policy pursued by APHIS to prevent 

animal and plant pests and diseases, including the 

gypsy moth, from being introduced into the United 

States

exotic — refers to all species of plants and animals 

not naturally occurring, either now or in the past, in an 

ecosystem of the United States

exposure — skin contact, inhalation, or ingestion of a 

substance that may have a harmful effect

exposure assessment — process of estimating the 

extent to which a population will come into contact 

with a chemical or biological agent

extra risk — risk in the population that can be 

attributed to exposure to the agent 

extrapolation — use of a model to make estimates 

outside of the observable range

exuviae — cast-off skins or outer coverings of insects 

and animals that shed skin

F

fecal — relating to feces (solid waste)

fibroma — benign tumor composed mainly of fibrous 

or fully developed connective tissue 

fibrosarcoma — malignant tumor derived from 

fibroblasts that produce collagen 

FIFRA — Federal Insecticide, Fungicide, and 

Rodenticide Act; establishes procedures for the 

registration, classification, and regulation of pesticides

final environmental impact statement — detailed, 

written statement of the analysis of a major Federal 

action, released to the public as required under sec. 102 

(2)(c) of the National Environmental Policy Act

financial assistance — money provided by the Forest 

Service and APHIS to Federal and State agencies 

through several pest control or management programs 

to suppress, eradicate, or slow the spread of the gypsy 

moth. On Federal lands the cost of gypsy moth 

projects are paid in full; on State and private lands cost 

may be shared with State cooperators. See technical 

assistance for other assistance provided

food chain — feeding sequence used to describe the 

flow of energy and materials through the system

food web — interconnected food chains in the 

ecosystem, representing the various paths of energy 

flow through populations in the community

Foray — one of the commercial formulations of B.t.k.

forest — land at least 10 percent occupied by forest 

trees or formerly having had such tree cover and 

not currently developed for non-forest use. Lands 

developed for non-forest use include areas for crops, 

improved pasture, residential or administrative areas, 

improved roads of any width, and adjoining roadclearing and power line clearing of any width

forest condition — species composition, tree growth 

rates and mortality rates, productivity, and degree of 

insect damage

forest cover type — description based on and named 

after the tree species that forms a plurality of the basal 

area in a stand; other tree species may also be part of 

the stand 

Forest Service — lead agency for this environmental 

impact statement; the largest USDA agency, which 

conducts research and develops the means to control 

the gypsy moth in forests; conducts surveys and 

evaluations on lands managed by other Federal 
Glossary

Chapter 7 - Page 

agencies; helps State and other Federal agencies to 

conduct detection surveys, evaluation and suppression; 

to test and transfer technology designed to improve 

gypsy moth control and reduce damage; and to conduct 

eradication on Federal or adjacent land, and on nonFederal land for infestations of more than 640 acres

forest type group — grouping of forest cover types for 

inventory, mapping, or other purposes

forestomach — front or foremost portion of the 

stomach in animals

formulation — commercial preparation of a chemical 

including any inert ingredients or contaminants

frank effects — obvious signs of toxicity

Frank Effect Level (FEL) — dose or concentration 

of a chemical or biological agent that causes gross and 

immediately observable signs of toxicity

frass — fecal excrement of gypsy moth caterpillars

fumigant — pesticide applied as a liquid or powder 

which volatilizes to gas; usually applied beneath a tarp, 

sheet, or other enclosure

fumigation — process of using a fumigant to destroy 

pests, usually applied under a cover or shelter

FWS — Fish and Wildlife Service, an agency of the 

U.S. Department of the Interior

G

gavage — placement of a toxic agent directly into the 

stomach of an animal, using a gastric tube

gene — basic unit of inheritance, by which hereditary 

characteristics are transmitted from parent to offspring. 

Genes consist of short lengths of DNA (or RNA in 

some viruses) that direct the synthesis of protein. These 

in turn influence the form and function of the organism

generally infested area — (regulated or quarantined 

area) the area in the eastern United States where the 

European strain of the gypsy moth is considered to 

be permanently established; also the area quarantined 

by APHIS and the States. All life stages are present, 

and populations are continuous. Population outbreaks 

occur, and defoliation is common. In 1994, the area 

extended from Maine to northern North Carolina and 

west to West Virginia, Ohio, and Michigan

genotoxic — causing direct damage to genetic 

material, associated with carcinogenicity

genotoxicity — specific adverse effect on the genome 

(the complement of genes contained in the haploid 

set of chromosomes) of living cells, which upon the 

duplication of the effected cells can be expressed as a 

mutagenic or carcinogenic event because of specific 

alteration of the molecular structure of the genome

geocorid — big-eyed bug

geometric mean — measure of an average value often 

applied to numbers for which a log-normal distribution 

is assumed

gestation — period between conception and birth; in 

humans, the period known as pregnancy

gram (g) — metric unit of measure for weight or mass

growth regulator — chemical that controls the rate 

of growth, or interferes with successful growth in an 

animal; diflubenzuron is a growth regulator for insects 

and other chitinous animals

guild — group of species with similar modes of 

existence

Gypchek — trade name for a biological insecticide 

containing gypsy moth nucleopolyhedrosis virus, 

which is registered and produced by the Forest Service 

and APHIS

gypsy moth — all life stages of the Asian and 

European strains of the insect with the scientific name 

Lymantria dispar (L.), previously Porthetria dispar

(L.)
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H

Haber’s Law — in toxicology, the assumption that 

the concentration or dose, multiplied by the duration 

of exposure (time) will always have the same effect. 

This relationship is true for some chemicals and some 

endpoints but not true for others. Even when true for a 

particular chemical and effect, it may be true only over 

certain ranges of exposure

habitat — place or type of site where a plant or animal 

naturally or normally lives and grows

half-life — time required for the concentration 

of a chemical to decrease by half of the original 

concentration (the longer the half-life, the more 

persistent a chemical is considered to be)

hazard — adverse effects to humans or the 

environment as a result of exposure to the gypsy moth 

or treatments; compare risk

hazard assessment — component of a risk 

assessment that consists of the review and evaluation 

of toxicological data to identify the nature of the 

hazards associated with a chemical, and to quantify the 

relationship between dose and response

hazard identification — process of identifying the 

array of potential effects that an agent may induce in an 

exposed population

hazard quotient — ratio of the estimated level of 

exposure to the risk-reference value or some other 

index of acceptable exposure; a hazard quotient greater 

than 1 raises concern 

Heinz bodies — dark-staining granules found in red 

blood cells, which are signs of oxidative damage; 

formation of Heinz bodies can lead to red cell 

dysfunction and breakdown of the cell membrane

hemangiosarcoma — malignant tumor formed by 

proliferation of endothelial and fibroblastic tissue

hematological — pertaining to the blood

hemipteran — insect belonging to the order 

Hemiptera, including the true bugs

hemoglobin — iron-containing respiratory pigment in 

red blood cells of vertebrates

herbaceous — relating to plants that have nonwoody 

stems and die down annually

herbivorous insect — insect that eats plants and plant 

material; the gypsy moth is an herbivorous insect 

because it eats leaves

HHERA — acronym for Human Health and Ecological 

Risk Assessment

histamine — naturally occurring chemical; causes 

dilation of the capillaries and muscle contraction

histopathology — signs of tissue damage that can be 

observed only by microscopic examination

homopteran — insect in the order Homoptera, which 

includes aphids, scale insects, and cicadas

host — living organism that provides subsistence or 

lodging for another organism

humoral — associated with agents dissolved in the 

blood or body fluids, in contrast to materials contained 

in cells (cell-mediated)

hydroxylation — addition of a hydrogen-oxygen 

or hydroxy (−OH) group to one of the electron rings 

of a compound. Hydroxylation increases the water 

solubility of aromatic compounds, particularly when 

followed by conjugation with other water-soluble 

compounds in the body, such as sugars or amino acids, 

hydroxylation greatly facilitates the elimination of the 

compound in the urine or bile

hymenopteran — any of highly specialized insects in 

the order Hymenoptera, usually with four membranous 

wings, the abdomen borne on a slender pedicel and 

associated with large colonies and complex social 

organization; includes bees, wasps, ants, ichneumonid 

flies, sawflies, and gall wasps

hypoactivity — less active than normal
Glossary

Chapter 7 - Page 11

I

immunocompetent — having normal immune 

function

immunocompromised — having an impaired immune 

system, such as people with HIV or AIDS

immunodeficient — organism with impaired immune 

function

in vitro — in glass; a test-tube culture; any laboratory 

test using living cells taken from an organism

in vivo — in the living organism; in vivo tests are those 

laboratory experiments carried out on whole animals or 

human volunteers

indirect effect — reaction of an organism to a change 

in the environment that is a direct result of exposure 

to a chemical or non-chemical agent. For example, 

wasps that prey on caterpillars that eat leaves with 

diflubenzuron on them could obtain diflubenzuron 

that the caterpillars ate, thus exposed indirectly to the 

chemical; the indirect effect of an unchecked gypsy 

moth infestation could be the change in woodland 

structure, a direct effect of the gypsy moth

inerts — adjuvants or additives in commercial 

formulations of gypsy moth control agents that do not 

cause mortality in the gypsy moth

inert ingredients — additives in insecticide 

formulations that do not effect the organism targeted 

but are added for a variety of reasons, such as to 

stabilize the formulation, to improve its weatherability, 

or to prevent growth of contaminating microorganisms

infestation — presence of the gypsy moth and an 

indication of a reproducing population, based on the 

results of surveys 

infested area — isolated infestation or generally 

infested area

inhalation — act of breathing

innocuous — something that produces no injury; 

harmless; inoffensive 

insecticide — pesticide that kills, debilitates, or 

controls the growth of insects

instar — stage between molts in the development of 

the gypsy moth caterpillar and other arthropods

Integrated Pest Management (IPM) — selecting 

strategies to mange pest-host systems for specific 

objectives; includes planning, detection, evaluation, 

monitoring, establishing acceptable damage thresholds, 

and use of appropriate management practices to prevent 

or control pest-caused damage and losses

intercept — in a simple linear equation, the value of 

the dependent variable when the independent variable 

is zero

interdisciplinary team — team of varied resource 

specialists with different professional backgrounds who 

conduct an environmental analysis; members of the 

interdisciplinary team who prepared this environmental 

impact statement are listed in chapter 5, Preparers and 

Contributors

interpolation — use of mathematical models within 

the range of observations 

intraperitoneal — injection into the abdominal cavity

invertebrates — animals without a spinal column, 

such as insects, spiders, and crustaceans

IPM — acronym for Integrated Pest Management

iritis — inflammation of the iris

irritant effect — reversible effect, compared with a 

corrosive effect

isolated infestation — defined area infested with the 

gypsy moth outside the generally infested area; or, a 

defined area infested with the Asian strain of the gypsy 

moth within the generally infested area

issue — public concern or significant problem that 

might occur when the gypsy moth is present or 

treatments are applied

IU — International Unit
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L

land use — type of activity occurring on the land 

surface, e.g. forestland, farmland, pastureland, etc

landscape — physical features of an area (e.g. slope, 

aspect, drainage) that affects the characteristics of the 

plant and animal communities in the ecosystem

Latin Hypercube — stratified sampling technique 

designed to sample from all portions of a distribution

larva — stage in development between hatching and 

attaining adult form 

larval survey — placing tar paper, burlap, or similar 

material around the trunks of susceptible trees, to create 

hiding places for gypsy moth caterpillars so they can be 

captured and counted

LC50 — acronym for lethal concentration50

LD1 — acronym for lethal dose1

LD50 — acronym for lethal dose50

leaf expansion — percentage of leaf growth from 0 to 

100

lentic — water bodies that do not flow (e.g., lakes, 

ponds)

lepidopteran — insects in the order Lepidopteran, 

characterized by adults with two pairs of scale-covered 

wings and coiled sucking-mouthparts, including moth 

and butterflies

lethal concentration50 (LC50) — calculated 

concentration of a toxicant in air (or water) to which 

exposure for a specific length of time is expected to 

cause death in 50 percent of a defined test animal 

population

lethal dose1

 (LD1) — dose of a chemical or biological 

agent calculated to cause death in 1 percent of a defined 

test animal population

lethal dose50 (LD50) — dose of a chemical or biological 

agent calculated to cause death in 50 percent of a 

defined test animal population

lethargy — decrease in the normal amount of activity

life stage — distinctive period in an insect’s life 

(Nichols 1989); life stages of the gypsy moth are: egg 

(in an egg mass), larva or caterpillar, pupa, and adult 

moth

lipophilic — having a tendency to dissolve or partition 

to fatty substances

LOAEL — acronym for lowest-observed-adverseeffect level

log-normally — a logarithmic function with a normal 

distribution

lotic — water bodies that flow and have running waters 

(e.g. streams, rivers)

lowest-observed-adverse-effect level (LOAEL) — 

lowest measured amount of a chemical that produces 

significant increases in frequency or severity of adverse 

effects in an exposed human population

M

macroinvertebrates — invertebrates large enough to 

be seen with the unaided eye

malignant — cancerous 

mammary gland — breast

management practice — specific act, measure, cause 

of action, or treatment 

mass trapping — using pheromone-baited traps to 

catch all or nearly all the male gypsy moths in an area 

having low gypsy moth populations

mast — fruit and seeds of trees and other forest 

vegetation eaten by wildlife; hard-mast includes nuts 

and seeds (such as acorns, walnuts, hickory nuts, maple 
Glossary

Chapter 7 - Page 13

seeds); soft-mast is fruit (such as apples, blackberries, 

wild grapes) 

mating disruption — saturating an area with gypsy 

moth pheromone to confuse male gypsy moths, thereby 

preventing them from locating and mating with females

median effective concentration (EC50) — 

concentration of a substance that results in some effect 

being exhibited by 50 percent of the test organisms

median lethal concentration — concentration of a 

toxicant necessary to kill 50 percent of the organisms 

in a population being tested; usually expressed in parts 

per million (ppm), milligrams per liter (mg/L), or 

milligrams per cubic meter (mg/m3

)

median lethal dose — dose necessary to kill 50 

percent of the test organisms; usually expressed in 

milligrams of chemical per kilogram of body weight 

(mg/kg)

metabolite — compound formed as a result of 

the metabolism or biochemical change of another 

compound

metastatic — pertaining to or of the nature of 

metastasis; the transfer of disease from one organ or 

part to another not directly connected with it; may be 

due either to the transfer of pathogenic microorganisms 

(e.g., bacilli) or to the transfer of cells, as in malignant 

tumors

methemoglobinemia — rare blood disorder in 

which there is a deficiency of the enzyme that turns 

methemoglobin into hemoglobin (methemoglobin 

differs from hemoglobin in being unable to combine 

reversibly with oxygen)

mg/cm2

 — milligrams per square centimeter

mg/kg — milligrams per kilogram

mg/m3 — milligrams per cubic meter

microclimate — climate of the immediate 

surroundings or habitat, differing from the 

macroclimate, as a result of the influences of local 

topography, vegetation and soil

microinvertebrates — invertebrates too small to be 

seen without magnification

microlepidopterans — general term for the most 

primitive families of moths whose members usually 

have the smallest body size among lepidopterans 

microorganism — organism so small that a 

microscope is necessary to see it

microsomal — pertaining to portions of cell 

preparations commonly associated with the oxidative 

metabolism of chemicals

mineralization — conversion of an organic substance 

into an inorganic substance as a result of microbial 

decomposition

minimal risk level (MRL) — route-specific (oral 

or inhalation) and duration-specific estimate of an 

exposure level that is not likely to be associated with 

adverse effects in the general population, including 

sensitive subgroups

mixture of concern — mixture on which a risk 

assessment is being conducted. See sufficient 

similarity.

molting — process of shedding an old skin and 

creating a new one, as an insect grows or changes in 

form

monitor — to observe or check that treatments are 

carried out as planned, or to determine whether effects 

of treatments are those that were predicted

Monte Carlo simulation — technique used to 

simulate systems with probabilistic elements; one 

or more variable in a Monte Carlo simulation is 

determined by drawing a random number from a 

probability distribution (such as the normal or uniform 

distribution), which describes the natural variation in 

that variable

most-sensitive effect — adverse effect observed at 

the lowest dose of a substance—an important concept 

in risk assessments; if the most-sensitive effect is 

prevented, no other effects will develop
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multiple-chemical sensitivity — syndrome that affects 

individuals who are extremely sensitive to chemicals at 

extremely low levels of exposure 

mutagenicity — ability of a substance (mutagen) 

to cause genetic damage, that is, damage to DNA or 

RNA (mutation); mutations can lead to birth defects, 

miscarriages, or cancer

N

nabid — damselbug belonging to Order Hemiptera of 

Class Insecta

NADH — acronym for nicotinamide adenine 

dinucleotide phosphate; a molecule that is common 

in all living systems and is necessary for the proper 

function of many enzymes 

nanogram (ng) — one billionth of a gram 

National Environmental Policy Act (NEPA) 

Act of 1969 (42 U.S.C. 4321) — established a 

national policy that encourages harmony between 

man and the environment; requires that Federal 

agencies proposing legislation or a major action use a 

systemic, interdisciplinary approach to planning and 

decisionmaking, and prepare a detailed statement that 

includes the following: the environmental impact of the 

proposed action, any adverse environmental effects that 

cannot be avoided, alternatives to the proposed action, 

the relationship between local short-term uses of man’s 

environment and the maintenance and enhancement 

of long-term productivity, and any irreversible and 

irretrievable commitment of resources 

National Gypsy Moth Survey — minimal detection 

survey administered by APHIS in cooperation with the 

States to detect isolated infestations of the gypsy moth 

outside the generally infested area

natural landmark — site on the National Registry 

of Natural Landmarks, administered by the National 

Park Service, U.S. Department of the Interior, 

preserved as an outstanding example of plant or animal 

communities, geological features, scenic grandeur, or 

other attribute 

natural spread — movement of gypsy moths from an 

infested area: (1) of first instar larvae by wind, (2) of 

larger larvae by crawling, (3) of adult females of the 

European strain by crawling, (4) of some adult females 

of the Asian strain by flying

necropsy — examination of a body after death, usually 

refers to a gross examination of the major organs

nematodes — elongated cylindrical worms that are 

parasitic in animals or plants or free-living in soil or 

water 

neotropical migrant — bird that nests in North 

America but migrates to the Neotropics (region of the 

New World south of the Tropic of Cancer, includes 

South America, Central America, southern Mexico, the 

West Indies, and Caribbean) during winter

NEPA — acronym for National Environmental Policy 

Act

neuropathy — damage to the peripheral nervous 

system

ng — nanogram, one billionth of a gram

NIOSH — acronym for the National Institute for 

Occupational Safety and Health

nm — nanometer, one billionth of a meter

NOAEL — acronym for non-observed-adverse-effect 

level

NOEL — acronym for no-observed-effect level

no-observed-adverse-effect level (NOAEL) — 

highest measured amount of a chemical at which no 

increase in frequency or severity of adverse effects 

is observed in an exposed human population when 

compared with a control; effects may be produced, but 

they are not considered to be adverse

no-observed-effect level (NOEL) — dose of a 

chemical or biological agent at which there are 

no biologically or statistically significant effects 

attributable to treatment
Glossary

Chapter 7 - Page 15

non-insecticidal treatments — gypsy moth treatments 

that do not involve spraying of insecticides; in this 

environmental impact statement, they include mass 

trapping, mating disruption, and the sterile insect 

technique

non-target organism — any living organism that is 

not the target of a management practice

normal distribution — theoretical frequencydistribution of variable data generally shaped in a bellshaped curve

Notice of Intent — announcement that preparation of a 

new national gypsy moth supplemental environmental 

impact statement was beginning, which appeared in 

the April 29, 2004, Federal Register (vol. 69, no. 83, p. 

23,492 – 23,493)

NPV — acronym for nucleopolyhedrosis virus

nucleopolyhedrosis virus (NPV) — category of 

naturally occurring viruses that cause a usually fatal 

disease, mainly in larvae of moths, butterflies, sawflies, 

wasps, ants, bees, and others. The nucleopolyhedrosis 

virus specific to the gypsy moth is the active ingredient 

in the insecticide Gypchek 

nymph — larvae of an insect with incomplete 

metamorphosis that differs chiefly in size and degree of 

differentiation from the final adult stage

O

OB — acronym for occlusion bodies

occlusion bodies (OB) — virus particles containing 

variable numbers of genetic material within one protein 

envelope

octanol-water partition coefficient (Kow) — 

equilibrium ratio of the concentrations of a chemical in 

n-octanol and water, in dilute solution

ocular — pertaining to the eye

odonates — insects in the order Odonata; dragonflies 

and damselflies

1-day health advisory — drinking water concentration 

(mg/L) not likely to cause adverse effects in the general 

population, including sensitive subgroups, after 1 day 

of exposure

one-storied stand — stand of trees that is 

characterized by the predominance of trees the same 

size

ophthalmic — pertaining to the eye, as an ophthalmic 

solution–a solution of medication intended to be 

applied to the eye

oral — pertaining to the mouth

oral toxicity — toxicity of a compound when given 

or taken by mouth, usually expressed as milligrams 

of chemical per kilogram of body weight of animal 

(mg/kg)

organoleptic — relating to an objectionable taste or 

smell

organophosphate — class of insecticides that are toxic 

to the nervous system

orthopteran — insects in the order Orthoptera, 

which includes grasshoppers, crickets, locusts, and 

cockroaches

osteosarcoma — malignant tumor derived from bone 

tissue

outbreak — cyclic rise in gypsy moth populations 

when feeding by caterpillars causes widespread 

moderate-to-heavy defoliation

ovicide — chemical toxic to the eggs of the target 

animal

P

parasite — organism that lives in, on, or at the expense 

of another, from which it obtains food, shelter, or other 

requirements; a parasite is usually smaller than the host 

and weakens it

parasitoid — parasite that eventually kills its host, for 
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example, insects that kill life-stages of the gypsy moth

parenteral — any form of injection

partition — in chemistry, the process by which a 

compound or mixture moves between two or more 

media

pathogen — an agent, such as a virus or bacterium, 

that causes disease

pathogenic — causing or capable of causing disease

pathway — in metabolism, a sequence of metabolic 

reactions

peroxide — molecule that contains two or more 

oxygen atoms in series, such as —O—O—; these 

molecules are often involved in the degradation of 

polymers, including proteins

persistence — characteristic of an insecticide or a 

compound to remain in the environment as an effective 

residue; persistence is related to volatility, chemical 

stability, and degradation

pesticide — substance or mixture of substances that 

kill insects, rodents, fungi, weeds, or other forms of 

plant or animal life that are considered to be pests

pH — measure of acidity and alkalinity on a scale 

from 0 to 14, of which 7 is neutral; lower numbers are 

acidic, higher numbers are alkaline; numbers vary by a 

factor of 10, i.e., pH 3 is 10 times more acidic than 

pH 4 

pharmacokinetics — quantitative study of the 

metabolic processes of absorption, distribution, 

biotransformation, and elimination of drugs

pheromone — chemical produced and emitted by 

an animal as a form of communication with other 

individuals of the same species, for example, the sex 

attractant given off by the female gypsy moth to attract 

males for mating

phytoplankton — small algal cells suspended in the 

water column of water bodies

phytotoxic — toxic or harmful to plants

piloerection — condition in which the hair stands on 

end

pituitary-adrenal axis — hormonal interaction 

between the pituitary and the adrenal glands

planktonic — suspended in the water of seas, lakes, 

rivers, or other water bodies

plasma cholinesterase — another term for pseudocholinesterase; the normal physiological role of 

this cholinesterase is not known, inhibition of this 

enzyme is considered an index of exposure to many 

oganophosphate insecticides

plasma — fluid portion of the blood in which 

particulates are suspended

plasmid — sub-cellular elements in bacteria that 

contain genetic material for relatively narrow and 

specific traits; plasmids can be transferred from one 

microorganism to another of the same species; transfer 

may also occur between two microorganisms of 

different species

Plecoptera — order of insects; includes stoneflies

polymer — generic term for a molecule composed of 

repeating units of less complex molecules; for example, 

proteins are polymers of amino acids

polyvinyl chloride — nontoxic polymer of vinyl 

chloride 

population — group of gypsy moths that occupy a 

defined area, separated to some degree from other 

groups, and are reproducing

population survey — counting egg masses in the 

generally infested area to determine if suppression 

treatments are warranted, or using pheromone traps 

in the transition area to determine if slow-the-spread 

treatments are warranted

post-treatment evaluation or survey — defoliation, 

egg mass, or larval survey conducted in a treatment 

area to evaluate treatment effectiveness
Glossary

Chapter 7 - Page 17

potentiation — action of two or more substances 

from which one or more (the potentiator) enhances the 

toxicity of another

ppb — parts per billion; the number of parts of 

chemical substance per billion parts of the substrate in 

question

ppm — parts per million; the number of parts of 

chemical substance per million parts of the substrate in 

question

predator — animal that obtains the energy it needs to 

live and grow by eating animals of other species, for 

example, some mice are predators of the gypsy moth

probit analysis — analysis technique that relates doses 

to measures of standard deviation away from the 50 

percent response level, using the cumulative normal 

distribution

programmatic — broad or general rather than site 

specific

proposed species — any species of fish, wildlife, 

or plant that is proposed in the Federal Register for 

listing as a threatened or endangered species under the 

Endangered Species Act 

proteinaceous — consisting or composed of proteins

proteolytic enzymes — enzymes that breakdown 

proteins

prototoxins — proteins that can be converted to toxins

pruritis — itching; an unpleasant skin sensation that 

provokes the desire to rub or scratch

pseudocholinesterase — term for cholinesterase 

found in the plasma; the normal physiological role 

of this cholinesterase is not known; inhibition of this 

enzyme is considered an index of exposure to many 

organophosphate insecticides

public involvement — actions taken by the Forest 

Service and APHIS to involve the various individuals, 

groups, and organizations who are interested in or may 

be affected by this environmental impact statement and 

the decision that may result

pupa — developmental stage of gypsy moth or any 

lepidoptera, between the caterpillar and adult moth 

stages, during which the insect undergoes major 

structural changes

Q

quarantine — designating an area as generally 

infested, so as to regulate the movement of articles 

(such as outdoor household articles, logs, and nursery 

stock) and prevent artificial spread of gypsy moth lifestages to uninfested areas of the United States 

R

racemic mixture — 50:50 blend of a (+) enantiomer 

and (−) enantiomer

recreational forest — publicly owned forest used 

predominantly for hiking, hunting, camping, day-use, 

and sightseeing

reference concentration — concentration in air (mg/

m3) not likely to be associated with adverse effects over 

lifetime-exposure, in the general population, including 

sensitive subgroups

reference dose (RfD) — oral dose (mg/kg/day) 

not likely to be associated with adverse effects over 

lifetime exposure in the general population, including 

sensitive subgroups

regeneration — renewal of a tree or stand of trees; 

restocking of an area

regulatory activities — activities conducted by APHIS 

and the States to prevent the artificial spread of the 

gypsy moth from the regulated area to the uninfested 

area; activities include inspection and treatment of 

regulated articles on which the gypsy moth commonly 

deposits egg masses. See quarantine

renal — pertaining to the kidneys
Chapter 7

Chapter 7 - Page 18

reproductive effects — adverse effects on the 

reproductive system that may result from exposure 

to a chemical or biological agent. The toxicity of the 

agent may be directed to the reproductive organs or the 

related endocrine system. The manifestations of these 

effects may be noted as alternatives in sexual behavior, 

fertility, pregnancy outcomes, or modification in other 

functions dependent on the integrity of the reproductive 

system

residue — quantity of insecticide and its metabolites 

remaining on and in vegetation, soil, or water

resistance — ability of a population or ecosystem 

to absorb an impact without significant change from 

normal fluctuations; for plants and animals, the ability 

to withstand adverse environmental conditions and/or 

exposure to toxic chemicals or disease

RfD — acronym for reference dose

rhinitis — inflammation of the mucous membranes of 

the nose

riparian — pertaining to, living in, or situation on, 

the banks of rivers and streams (Lincoln and Boxshall 

1987)

risk — likelihood that adverse effects will occur; 

compare hazard

risk assessment — evaluation of the likelihood 

that adverse effects may occur in humans or the 

environment as a result of exposure to one or more 

stressors, such as the gypsy moth and treatments

risk characterization — process of estimating the 

incidence of a healthy effect in a human population 

under the different conditions of exposure described in 

the exposure assessment

risk comparison — the practice of comparing one risk 

to another in order to promote a better understanding of 

the consequences of different treatment options as well 

as the consequences of no treatment

risk reference-value (RRV) — generic term used as 

an estimate of dose that is not likely to induce adverse 

health effects in humans under specific conditions of 

exposure such as duration and route

route-of-exposure — way in which a chemical or 

biological agent enters the body. Most typical routes 

include oral (eating or drinking), dermal (contact of the 

agent with the skin), and inhalation

RRV — acronym for risk reference value

S

safety factor — factor used to give a margin-oferror to the screening index in the Ecological Risk 

Assessment; safety factors are selected based on the 

amount of error likely in estimating toxicological 

benchmark values or concentrations of a toxicant in the 

environment

salvage — cutting and removing dead, dying, or 

deteriorating trees before they lose their value as timber

sarcoma — tumor made up of a substance like 

embryonic connective tissue; often highly malignant

scientific notation — the method of expressing 

quantities as the product of a number between 1 and 10, 

multiplied by 10 raised to some power. For example, in 

scientific notation, 

1 kg = 1,000 g [is expressed as] 1 kg = 1 × 103 g; 1 mg 

= 0.001 [is expressed as] 1 mg = 1 × 10 −3 g

scission — in metabolism, breaking or cleavage of part 

of a molecule

scoping — open process, including public notification 

and participation, by which an agency identifies 

significant environmental issues and determines the 

extent of analysis needed to make an informed decision 

on a proposed action

screening index — index used to determine whether 

a species exposed to a toxic agent is at risk. The 

screening index is a conservative estimate of species at 

risk. It is more likely to indicate that a species is at risk 

when it actually may not be than to miss species that 

are at risk
Glossary

Chapter 7 - Page 19

secondary organism — pathogens or insects that 

attack trees already weakened by defoliation and that 

sometimes cause death of the trees

SEIS — acronym for Supplemental Environmental 

Impact Statement

sensitive subgroup — subpopulation that is much 

more sensitive than the general public to certain agents 

in the environment

septicemia — occurrence of pathogens or pathogenic 

toxins in the blood or other body fluids 

serotype — classification of a microorganism based on 

occurrence of antigens in the cell

silviculture — practice of applying treatments to forest 

stands, to maintain and enhance them for any purpose 

(Smith 1986); silvicultural treatments may also be 

applied to forested areas in urban and suburban areas 

slow the spread — strategy being pilot-tested on a 

large-scale to determine its biological effectiveness 

and economic efficiency in slowing the gypsy 

moth’s natural spread from areas where it is already 

established or is a permanent resident by keeping lowlevel populations from increasing

species composition — assemblage of species 

inhabiting a defined area

species diversity — ecological concept that 

incorporates both the number of species in a given area 

and the number of individuals per species

species richness — number of species in a local area, 

region, or community

species-to-species extrapolation — method involving 

the use of exposure data on one species (usually an 

experimental mammal) to estimate the effects of 

exposure in another species (usually humans)

squamous-cell papillomas — type of benign tumor

stand — contiguous group of trees sufficiently uniform 

in species composition, age, and condition to be 

distinguishable as a unit

stand composition — variety of vegetation species in 

a stand

stand growth — increases in wood, dry matter, or 

biomass with a stand

stand structure — combination of species, ages, sizes, 

and numbers of trees that describe a stand

standard deviation — expression of the variability in 

a sample or population

standard-normal distribution — normal distribution 

with a mean of zero and a standard deviation of one

sterile insect technique — gypsy moth treatment 

that reduces the chance of fertile female gypsy moths 

mating with fertile males and producing fertile eggs, 

by the release of large numbers of (1) male pupae 

sterilized by radiation, (2) male pupae irradiated but not 

sterilized, or (3) eggs from mating of irradiated males 

with non-irradiated females

stewardship and stewardship incentives programs

— cooperative programs between the Forest Service 

and States, to provide financial and technical assistance 

for silvicultural planning on non-Federal forested areas 

for private landowners 

strain — group within a species that differs 

physiologically rather than in form or structure

strategy — planned actions with specific objectives; 

the strategies of eradication, suppression, and slow 

the spread make-up the alternatives examined in this 

environmental impact statement

Streptococcus (pl. Streptococci) — genus of bacteria, 

which—depending on its classification—may be 

associated with infections in humans

stressor — an agent, such as an insecticide or the 

gypsy moth, that causes stress to an ecosystem

subcanopy — cover of branches and foliage formed 

collectively by trees and other woody growth that is 

below the principal canopy
Chapter 7

Chapter 7 - Page 20

subchronic exposure — exposure studies that can last 

for different periods of time, but 90 days is the most 

common duration; the subchronic exposure study is 

usually performed in two species (rat and dog) by the 

route of intended use or exposure

subchronic reference dose — oral dose (mg/kg/day) 

not likely to be associated with adverse effects over a 

less-than-lifetime exposure, in the general population, 

including sensitive subgroups

subchronic toxicity — adverse biologic response of 

an organism, such as mortality or an effect on growth 

or reproductive success, resulting from repeated or 

short-term (3 month) doses (exposures) of a compound, 

usually at low concentrations

subconjunctival — refers to the area beneath the 

membrane that lines the eyelids and eyeball

subcutaneous — just below the skin 

subdominant trees — trees with crowns below the 

general level of the canopy and that receive little or no 

direct light from above; trees whose crowns make up 

the subcanopy (Smith 1986)

substrate — with reference to enzymes, the chemical 

that the enzyme acts upon 

succession — natural and gradual replacement of one 

community of plants by another

succinylcholine — neuromuscular blocking agent

sufficient similarity — as applied to chemical 

mixtures, whether or not the data on one or more 

samples of a complex and variable mixture can or 

should be used for dose-response assessments for all 

such mixtures

sulfhemoglobinemia — presence of abnormal 

pigments, other than methemoglobin, in red blood cells

Supplemental Environmental Impact Statement — 

a document that is written to provide a supplement to 

the original Environmental Impact Statement 

suppression — strategy of reducing outbreak 

populations of the gypsy moth in areas where it is 

already established, or is a permanent resident, to 

prevent or minimize damage to resources

survey — see defoliation survey, delimiting survey, 

detection survey, egg mass survey, larval survey, 

National Gypsy Moth Survey, population survey, posttreatment survey, and transition area survey 

susceptible plants — plants with leaves the gypsy 

moth will eat

synapse — space between two nerve cells or a nerve 

cell and an effector cell such as muscle

synergism — action of two or more substances 

to achieve an effect of which each is individually 

incapable; synergistic effects may be greater or less 

than the sum of effects of the substances in question

synergistic effect — situation in which the combined 

effects of two chemicals are much greater than the sum 

of the effect of each given agent alone

systemic — entering and then distributing throughout 

the body of an organism

systemic effects — effects that require absorption of a 

toxic agent at an entry point and distribution to a distant 

site at which effects are produced

systemic toxicity — effects that require absorption 

and distribution of a toxic agent to a site distant from 

its entry point at which point effects are produced; 

systemic effects are the obverse of local effects

T

technical assistance — any of a whole range of direct 

and indirect help that USDA provides to Federal 

and State cooperators, short of providing monetary 

funds; this assistance includes but is not limited to 

providing training, providing assistance in preparing 

environmental documents, work and safety plans, 

contracts, and monitoring plans, and providing 

assistance on site during the conduct and evaluation of 

gypsy moth projects 
Glossary

Chapter 7 - Page 21

technology transfer — disseminating research results 

and adapting innovations so government and private 

parties can use them

1-day health advisory — drinking water concentration 

(mg/L) not likely to cause adverse effects in the general 

population, including sensitive subgroups, after 1 day 

of exposure

teratogenic — relating to or causing developmental 

malformations

teratology — study of malformations induced during 

development from conception to birth

thinning from below — silvicultural technique of 

removing the subdominant trees in a forest stand, 

leaving the dominant trees more or less evenly 

distributed over the stand 

threatened species — Federal designation for any 

species that is likely to become an endangered species 

within the foreseeable future throughout all or a 

significant portion of its range (the Federal list of 

threatened species is maintained by the Secretary of the 

Interior)

threshold — maximum dose or concentration level of 

a chemical or biological agent that will not cause an 

effect in the organism

threshold-limit value — air concentration, in 

milligrams per cubic meter (mg/m3

), not likely to cause 

adverse effects in exposed workers, over a normal 

period of work

Thuricide — one of the commercial formulations of 

B.t.k.

toxic — poisonous to organisms

toxicant — poisonous substance such as the active 

ingredient in pesticide formulations that can injure or 

kill plants, animals, or microorganisms

toxicity — capacity of a poison to cause adverse effects

toxicological benchmark value (or benchmark 

value) — values determined for any of a number 

of toxicological tests, such as lethal dose 50, lethal 

concentration 50, no-observed-adverse-effect level, 

lowest-observed-adverse-effect level

toxicology — science that deals with poisons and 

their effects and problems involved (such as clinical, 

industrial, or legal) 

toxins — chemicals that may cause toxic effects, often 

used when referring to naturally occurring toxic agents, 

especially proteins

transition area — area between the uninfested 

area and generally infested area; populations are 

discontinuous, consist mostly of adult male moths, and 

occasionally other life stages; population outbreaks do 

not occur, and defoliation is uncommon

transition area survey — monitoring gypsy moths 

in the transition area to provide date that support 

the decision to quarantine an area or to take other 

management action

treatment threshold — population level reached by 

an insect pest that indicates treatment is necessary to 

prevent unacceptable damage to other resources

triangular distribution — theoretical frequencydistribution shaped like a triangle and described by a 

minimum, maximum, and likeliest values

trichopteran — insects in the order Trichoptera, in 

which the adults are terrestrial and immature life stages 

are almost exclusively aquatic in freshwater; caddisflies

trophic levels — feeding levels—for example, primary 

producer, herbivore, and first-level carnivore

U

uncertainty factor — factor used in deriving the riskreference values and similar values from experimental 

data; uncertainty factors are intended to account for 

variation in sensitivity among people, the uncertainty in 

extrapolating animal data to humans, and other sources 

of uncertainty; common uncertainty factors are 10, 100, 

and 1,000
Chapter 7

Chapter 7 - Page 22

understory — vegetation layer below the canopy of 

other plants, formed by shade-tolerant trees and low 

shrubs, grasses, and other herbaceous plants

uninfested area — area outside the generally infested 

area and ahead of the transition area; adult male moths 

are occasionally found, other life stages are rarely 

found; no populations are found, and no outbreaks 

occur

uniform distribution — theoretical frequencydistribution described by a minimum and a maximum 

value; all values in the uniform distribution have an 

equal probability of occurrence

Urban and Community Forestry Program — 

cooperative program between the USDA Forest Service 

and States to provide financial and technical assistance 

to municipalities, school districts, communities, and 

nonprofit organizations for managing trees on nonFederal lands in urban environments

urban forest — forested areas in cities, towns, and 

communities

urinalysis — testing of urine samples to determine 

whether toxic or other physical effects have occurred in 

an organism

urticaria — skin condition marked by the development 

of wheals

USDA — acronym for U.S. Department of Agriculture

U.S. EPA — acronym for U.S. Environmental 

Protection Agency

V

vehicle — substance (usually a liquid) used as a 

medium for suspending or dissolving the active 

ingredient; commonly used vehicles include water, 

acetone, and corn oil

vertebrates — animals with a spinal column, such as 

mammals, fish, birds, amphibians, and reptiles

volatile — referring to compounds or substances 

that have a tendency to vaporize; material that will 

evaporate quickly

volatility — tendency of a substance to evaporate at 

normal temperatures and pressures

vulnerability — likelihood that a tree or plant will die 

if defoliated

W

watershed — area of land with a characteristic 

drainage network that contributes to the same surface 

flow

wheal — smooth, slightly elevated area on the body 

surface, which is more red or more pale than the 

surrounding skin; often accompanied by severe itching 

and usually changing size or shape or disappearing 

within a few hours; the typical lesion of urticaria, the 

dermal evidence of an allergic reaction (allergy), and 

in sensitive persons may be provoked by mechanical 

irritation of the skin; also called a hive

X

xenobiotic — chemical that does not naturally occur 

in an organism; term is often applied generically to all 

synthetic or man-made chemicals

Z

zooplankton — animals that are dependent on movement of water or air for their position or distribution
Chapter 8

References

Figure 8-1. Civilian Conservation Corps workers scouted for gypsy moths.
Chapter 8 References

Figure

Figure 8-1. Civilian Conservation Corps workers scouted for 

gypsy moths...........................................................................................Cover
References

Chapter 8 - Page 

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A

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Chapter 8 - Page 

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Pesticide Precautionary Statement

This publication reports research involving pesticides. It does not contain recommendations for their 

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CAUTION: Pesticides can be injurious to humans, domestic animals, desirable plants, and fish or 

other wildlife--if they are not handled or applied properly. Use all pesticides selectively and carefully. 

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Printed on recycled paper with soy-based ink.
Appendix A

Gypsy Moth Treatments 

and Application 

Technology

Figure A-1. Public notices warned campers about transporting gypsy moth eggs 

and caterpillars.
Appendix A Gypsy Moth Treatments and 

Application Technology

Contents

A.1 Treatments in the 1996 Record of Decision...........................................1

Bacillus thuringiensis var. kurstaki (B.t.k.).......................................1

 B.t.k. Use..............................................................................1

 B.t.k. Effectiveness...............................................................2

Diflubenzuron...................................................................................2

 Diflubenzuron Use...............................................................2

 Diflubenzuron Effectiveness................................................3

Gypchek (Nucleopolyhedrosis Virus)...............................................3

 Gypchek Use........................................................................4

 Gypchek Effectiveness.........................................................4

Mass Trapping (Disparlure Only, or Disparlure and Dichlorvos).....4

 Mass Trapping Use..............................................................5

 Mass Trapping Effectiveness...............................................5

Mating Disruption (Disparlure).........................................................5

 Mating Disruption Use........................................................6

 Mating Disruption Effectiveness.........................................6

Sterile Insect Technique....................................................................7

 Sterile Insect Technique Use................................................7

 Sterile and Substerile Male Pupae..........................7

 Inherited Sterility....................................................8

 Sterile Insect Technique Effectiveness...................8

A.2 The New Proposed Treatment of Tebufenozide.....................................8

 Tebufenozide Use.................................................................9

 Tebufenozide Effectiveness..................................................9

A.3 Treatments That Include Natural Control Agents..................................9

Fungal Pathogens..............................................................................9

Parasitoids.......................................................................................11

Predators.........................................................................................12

Nematodes.......................................................................................14

Microsporidia..................................................................................14

A.4 Miscellaneous Means of Gypsy Moth Management............................15

Removing and Destroying Egg Masses...........................................15

Tree Trunk Bands and Barriers.......................................................15

Broad-Spectrum Insecticides..........................................................15

Silviculture......................................................................................16

A.5 Advances in Application Technology...................................................16
Figures

Figure A-1. Public notices warned campers about transporting gypsy moth eggs and caterpillars................... Cover

Figure A-2. Researchers hope to be able to manufacture Gypchek in bioreactors such as this................................. 4

Figure A-3. Gypsy moth larvae killed by the nucleopolyhedrosis virus typically hang in an inverted V................. 4

Figure A-4. Sterile gypsy moths are reared on artificial diet in a climate-controlled environmental chamber......... 8

Figure A-5. A parasitic wasp lays eggs on gypsy moth pupal case; eggs hatch into wasp larvae, which feed 

on and kill the host................................................................................................................................................... 11

Figure A-6. Tachinid flies will parasitize gypsy moth caterpillars (photo was taken in Mongolia)........................ 12

Figure A-7. White-footed mice feed on gypsy moth larvae..................................................................................... 13

Figure A-8. The Calosoma beetle is a gypsy moth predator introduced from Europe ............................................ 13

Gypsy Moth Treatments and Application Technology

Appendix A - Page 

This appendix describes treatments used and proposed 

for use in managing the gypsy moth. These treatments 

vary in effectiveness in different situations. Some are 

not effective in meeting the objectives of eradication, 

suppression, or slow-the-spread projects; but they are 

presented in order to provide the reader with a fuller 

understanding of the range of control and natural agents 

that regulate gypsy moth populations.

The treatments are divided into four categories. 

The first category includes those treatments in the 

1996 Record of Decision: Bacillus thuringiensis 

var. kurstaki (B.t.k), diflubenzuron, the gypsy 

moth nucleopolyhedrosis virus product (Gypchek), 

mating disruption, mass trapping, and sterile insect 

technique. The second treatment category consists 

of the new proposed treatment of tebufenozide. The 

environmental and human health risks associated with 

the use of treatments in these first two categories are 

analyzed and presented in Appendixes F-K of this 

supplemental environmental impact statement (SEIS). 

The environmental effects are summarized in 

Chapter 4.

The third category contains some natural control 

agents that help regulate gypsy moth populations in 

North America and in other places around the world 

where gypsy moth exists. These natural control agents 

include fungal pathogens, parasitoids, predators, 

nematodes, and microsporidia. Unfortunately, cost 

effective technology does not yet exist to develop and 

propagate these agents for use within the USDA gypsy 

moth management program.

The fourth category contains the miscellaneous 

treatments of removing and destroying egg masses, tree 

trunk bands and barriers, broad-spectrum insecticides, 

and silviculture. These treatment methods do not 

meet the objectives of eradication, suppression, and 

slow-the-spread projects. Some of the treatments may 

have value, however, for protecting individual trees in 

homeowner’s yards or other landscape situations, rather 

than in a forest setting or in a large treatment area.

A.1 Treatments in the 1996 

Record of Decision.

Bacillus thuringiensis var. kurstaki 

(B.t.k.). 

Bacillus thuringiensis, commonly called B.t., is a 

bacterium that moves by using whip-like appendages 

called flagella and forms a resting spore. B.t. occurs 

naturally in soils throughout the world. Unique to this 

species is formation of a protein crystal next to the 

spore at the time of sporulation.

B.t. commercial formulations used for managing 

defoliating forest caterpillars in North America are 

preparations of the HD-1 strain of B.t. variety kurstaki

(B.t.k.). B.t.k. spores and crystals are ingested by the 

gypsy moth caterpillar along with foliage. Enzymes in 

the mid-gut of the caterpillar dissolve the crystals and 

release delta-endotoxins, which are insecticidal crystal 

proteins. The proteins bind to specific receptors on 

the cellular lining of the midgut and penetrate the cell 

membrane. The insect stops feeding and dies within a 

few hours or days.

Natural epizootics caused by B.t.k. have not been 

observed as a control factor for the gypsy moth 

(Reardon and others 1994). B.t.k. is not expected to 

infect more than the current year generation of gypsy 

moths present when it is applied (Dubois and others 

1988).

B.t.k. Use.

A number of commercial preparations of B.t.k. are 

registered for aerial and ground application to gypsy 

moth populations. The typical application rate used 

in USDA cooperative suppression projects is one 

application at 24 to 38 BIU per acre (60-95 BIU/ha). 

For eradication treatments, the typical dose rate is 24 to 

25 BIU per acre (60-63 BIU/ha), applied one to three 

times with application times being from a few days to 

over a week apart.
Appendix A

Appendix A - Page 

The United States Department of Agriculture (USDA) 

first used B.t.k. in cooperative suppression projects 

for the gypsy moth in 1980. Between 1995 and 2003, 

B.t.k. was used in more than 68 percent of the total 

acreage treated in cooperative suppression projects, 

more than 1.4 million acres (0.5 million ha) in nine 

States (U.S. Department of Agriculture, Forest Service 

2003).

The timing of B.t.k. application in gypsy moth projects 

is generally dictated by foliage and insect development 

(Dubois 1991). The optimal timing of application is 

when most of the insects are in the second instar, and 

not delayed beyond early third instar. To be effective, 

B.t.k. must be consumed by the caterpillar. The 

timing of B.t.k. application is a subjective judgment 

considering foliage expansion, larval stage, population 

density and predicted level of defoliation (Reardon and 

others 1994).

Phenology models such as BIOSIM may be used to 

help predict insect development in eradication projects 

where gypsy moth numbers are so low that egg masses 

and larvae cannot be located and monitored. This 

allows application of B.t.k. at the most opportune time. 

Caged egg masses are sometimes deployed in treatment 

areas and monitored for egg hatch so that the optimal 

timing of B.t.k. application can be estimated.

B.t.k. Effectiveness.

The effectiveness of B.t.k. in cooperative suppression 

projects from 2000 to 2003 varied from a low of 

84 percent to a high of 100 percent; the average 

success rate of suppression projects was 95 percent in 

reducing gypsy moth populations (U.S. Department of 

Agriculture, Forest Service 2003). Greater reductions 

in gypsy moth populations generally occurred with 

higher dose rates (24 and 38 BIU per acre: 60 and 

95 BIU/ha) (U.S. Department of Agriculture, Forest 

Service 2003).

Many factors affect B.t.k. efficacy, including the timing 

of the application with regard to insect and foliage 

development, weather conditions during and after 

application, and the quality of the application, that is, 

good pilot skills and properly functioning equipment. 

Most important is application timing and delivering a 

dose sufficient to kill the insects. The species of host 

plant may also affect the effectiveness of B.t.k. (Farrar 

and others 1996).

During eradication applications in or near areas that 

contain rare, endangered, or desirable moths and 

butterflies, extra effort should be taken to minimize 

drift (see Advances in Application Technology later in 

this appendix). See Appendix F for the risk assessment 

on B.t.k.

Diflubenzuron.

Diflubenzuron belongs to a group of compounds 

called insect growth regulators. When ingested by 

gypsy moth caterpillars, diflubenzuron disrupts the 

formation of a new cuticle (outer skin) during molting. 

The caterpillar cannot complete the molting process, 

its body wall ruptures from internal pressure, and the 

insect dies. Ingestion of diflubenzuron is lethal to the 

gypsy moth caterpillar.

Diflubenzuron Use.

Diflubenzuron is registered for aerial or ground 

application for gypsy moth. The label prohibits 

application directly to water, to areas where surface 

water is present, or to intertidal areas below the high 

water mark—except under the forest canopy when 

aerially applied. Typically, diflubenzuron is aerially 

applied at the rate of 0.5 ounces active ingredient in 

0.75 to 1.00 gallon spray volume-per-acre, twice in 

eradication projects and once in suppression projects. 

Diflubenzuron application in suppression projects may 

be at a much lower dosage than the commonly used 

0.5 ounce active ingredient per acre and still achieve 

project objectives (McLane 1993).
Gypsy Moth Treatments and Application Technology

Appendix A - Page 

Diflubenzuron Effectiveness. 

Diflubenzuron effectively reduces gypsy moth 

populations and protects foliage, both key objectives of 

suppression projects. Data collected from 2000 to 2003 

from areas treated with diflubenzuron in cooperative 

suppression projects with States reveal diflubenzuron 

has a 95 to 98 percent success rate in meeting 

foliage protection objectives. From 1995 to 2003, 

diflubenzuron was used on about 30 percent of the total 

acres treated in cooperative suppression projects (GM 

Digest 2004). See Appendix I for the risk assessment 

on diflubenzuron.

Gypchek (Nucleopolyhedrosis Virus).

The gypsy moth nucleopolyhedrosis virus (NPV) is 

one of several natural agents found in eastern North 

America that infect gypsy moth (Podgwaite and 

Campbell 1972). The virus is a member of the genus 

Baculovirus and is unrelated to arthropod-borne viruses 

and other viruses that infect man (Mazzone and others 

1976). The disease caused by the gypsy moth virus is 

commonly referred to as “wilt disease” because of the 

limp appearance of infected caterpillars.

The disease can reach outbreak levels naturally 

as gypsy moth populations increase. Epizootics 

caused by the gypsy moth virus are thought to be 

density dependent, and display one or more waves of 

mortality; intensity is proportional to larval density 

and viral inoculum (Doane 1970, Woods and others 

1990). Outbreaks of this type result from increased 

transmission rates of the virus within and between 

generations of the gypsy moth. Small gypsy moth 

caterpillars become infected and die on leaves in 

the tree crowns, the cadavers disintegrate, and the 

viral particles disperse, infecting other gypsy moth 

caterpillars.

The virus appears to spread rather easily when egg 

masses are laid on virus-contaminated surfaces. Birds, 

mammals, gypsy moth parasitoids, and invertebrate 

predators may also play a role in spreading the virus, 

although they themselves are not affected. The virus 

may kill up to 90 percent of the caterpillars in dense 

gypsy moth populations, reducing populations to levels 

that cause only minimal defoliation the following year 

(Reardon and Podgwaite 1992, Reardon and others 

1996).

USDA began investigating the feasibility of developing 

gypsy moth virus as an alternative to chemical 

insecticides in the late 1950s. The viral product 

Gypchek was registered with the U.S. Environmental 

Protection Agency (U.S. EPA) in 1978 as a general use 

insecticide for ground and aerial application (Reardon 

and Podgwaite 1992, Reardon and others 1996). 

Gypchek must be used under the supervision of the 

Forest Service.

Gypchek is specific to the gypsy moth and does not 

affect other caterpillar species or any other nontarget 

organisms that might be present in treatment areas 

(Barber and others 1993, Rastall and others 2003). 

This fact renders Gypchek a desirable insecticide for 

use where threatened or endangered species might 

be found or in other environmentally sensitive areas; 

however, the availability of Gypchek is limited.

Gypchek is produced by the Forest Service and APHIS 

in quantities sufficient to treat about 8,000 acres 

(3,240 ha) each year. Production involves raising 

large numbers of gypsy moth caterpillars, inoculating 

and then processing the infected caterpillars at the 

appropriate time. Anywhere from 500 to 1,000 infected 

caterpillars are required to produce enough Gypchek 

to treat 1 acre with two applications. Widespread 

operational use of Gypchek hinges on availability and 

cost (Reardon and Podgwaite 1992,Reardon and others 

1996). Gypchek can be applied with aerial or ground 

techniques, and when applied properly can achieve 

suppression rates similar to B.t.k. (Thorpe and others 

1998).

On-going research may result in the future ability to 

manufacture Gypchek in bioreactors, avoiding the 

higher costs and difficulty of rearing caterpillars to 
Appendix A

Appendix A - Page 

produce the virus (Figure A-2). Research also seeks 

to produce a strain of Gypchek that is more effective 

against the gypsy moth.

Gypchek must be ingested by the gypsy moth 

caterpillar. The rod-shaped virus particles, or virions, 

are liberated in the gut of the insect. The virions invade 

the gut wall and attack the internal organs and tissues, 

causing infection. The virus multiplies rapidly in 

cells of the insect and eventually causes breakdown of 

internal tissue and death. The entire process takes from 

10 to 14 days, depending on the size of the caterpillar, 

viral dose, and ambient temperature. First and second 

instar caterpillars are most susceptible to Gypchek. 

Dead caterpillars typically hang in an inverted “V” 

from foliage and branches and often rupture, releasing 

more virus that can infect other gypsy moths (Reardon 

and Podgwaite 1992, Reardon and others 1996) (Figure 

A-3).

Gypchek Use. 

Gypchek must be formulated at the mixing and 

loading site before aerial application. The standard 

tank mix consists of water (pH 5.0-8.0), ultravioletlight sunscreen and a sticking agent (to aid adhesion 

to leaf surfaces). During the years 1995 to 2003, 

Gypchek was used on an average of 5,014 acres per 

year for suppression, eradication, and slow-the-spread 

treatments (U.S. Department of Agriculture, Forest 

Service 2003). Gypchek is usually applied against first 

or second instars of the gypsy moth.

Gypchek Effectiveness.

Gypchek is preferably used against moderate-to-high 

gypsy moth populations (300-5,000 egg masses/acre 

[741-12,355 egg masses/ha]). Gypchek does not 

adversely affect nontarget species (Rastall and others 

2003). See Appendix G for the risk assessment on 

Gypchek.

Mass Trapping (Disparlure Only, or 

Disparlure and Dichlorvos).

Mass trapping uses disparlure (synthetic sex 

pheromone) to attract male moths to traps placed in a 

grid pattern across a target area, with the objective of 

capturing male gypsy moths before they are able to 

locate and mate with female moths. Two types of traps 

are used, depending on the expected number of moths 

to be caught: the smaller delta trap and larger milk 

carton trap.

Figure A-2. Researchers hope to be able to manufacture 

Gypchek in bioreactors such as this. (Forest Service 

laboratory, Delaware, OH)

Figure A-3. Gypsy moth larvae killed by the 

nucleopolyhedrosis virus typically hang in an inverted V.
Gypsy Moth Treatments and Application Technology

Appendix A - Page 

The standard “delta” trap is a small-capacity trap, 

approximately 8 inches (20 cm) long, 4 inches (10 cm) 

high, and triangular in cross section. A tiny plastic strip 

impregnated with disparlure or a string impregnated 

with disparlure is stapled to the inside of the trap to 

attract male gypsy moths. The inside surface of the 

trap is coated with a sticky substance to capture the 

moths and prevent their escape from the traps.

The second type of trap is called the “milk carton” 

trap because it resembles a half-gallon cardboard 

milk container. This type of trap is used in areas 

where large numbers of male moths are expected to 

be caught and would quickly overwhelm the sticky 

surface of the smaller delta trap. As in the delta trap, a 

small pheromone wick containing disparlure is placed 

inside the milk carton trap to attract moths. Unlike 

the delta trap, the milk carton trap also contains a 

1-inch by 4-inch (2.5 cm by 10 cm) laminated plastic 

strip impregnated with the insecticide dichlorvos 

(2,3 dichloroethenyl dimethyl ester phosphoric acid 

[DDVP]) to kill the moths and prevent their escape 

from the traps. Dichlorvos, registered with the 

U.S. EPA, is manufactured by AMVAC Chemical 

Corporation (City of Commerce, CA). When used 

in milk carton traps, dichlorvos is formulated and 

registered as Vaportape II (Hercon Environmental 

Company, Emigsville, PA). A risk analysis for 

dichlorvos is found in Appendix K of this SEIS.

Mass Trapping Use.

Both types of traps are used for detecting and 

monitoring gypsy moth populations. Delta traps 

are most commonly used in the uninfested area 

of the United States to detect and delimit isolated 

infestations of the gypsy moth. Milk carton traps are 

more commonly used in areas where large numbers of 

male moths are likely to be caught. The delta trap is 

primarily used for mass trapping, though milk carton 

traps might be considered if the expected catch per trap 

is greater than 15 moths, which would overwhelm the 

sticky surface in the smaller delta trap. When used for 

mass trapping, delta or milk carton traps are deployed 

in an intensive grid pattern across an infested area and 

an adjacent buffer area, at the rate of at least 9 traps per 

acre (25 traps/ha).

Mass Trapping Effectiveness.

The success of mass trapping depends on the density 

of the gypsy moth population in the treatment area, 

because the tactic relies on luring all male moths into 

the traps before they can mate with females. The 

higher the population density, the greater the risk that 

a male will find and mate with a female before being 

lured into a trap. Therefore, the treatment is best used 

where there are less than 10 egg masses per acre (25 

egg masses/ha) (USDA Forest Service 1989).

Mass trapping is a labor-intensive treatment, especially 

over large areas; it is typically used on small 

infestations of less than 100 acres (40.4 ha). Nontarget 

organisms are unaffected, except those that accidentally 

find their way into the traps (primarily flying insects).

See Appendix H for the risk assessment on disparlure, 

and Appendix K for the risk assessment on dichlorvos.

Mating Disruption (Disparlure).

Disparlure is a gypsy moth pheromone that attracts 

male moths to female moths for mating. Synthetically 

produced disparlure can be used to disrupt the mating 

of gypsy moths. Mating disruption relies on the use of 

the gypsy moth pheromone disparlure (cis-7,8-epoxy2-methyloctadecane [racemic disparlure]) as the active 

ingredient; however, a 50:50 mixture of the plus (+) 

and minus (−) enantiomers of synthetic disparlure are 

used rather than only the plus (+) enantiomer used 

in trap lures. This 50:50 mixture of enantiomers, 

called racemic disparlure, lacks the highly attractive 

characteristics of plus disparlure. Instead of luring 

adult male gypsy moths away from females, application 

of racemic disparlure interferes, or “disrupts,” the male 

moths’ normal mate search behavior, which prevents 

them from finding and mating with the females.
Appendix A

Appendix A - Page 

Mating Disruption Use. 

Mating disruption was first used operationally in a 

USDA cooperative eradication project in Virginia in 

1983. Widespread use of this treatment did not begin 

until initiation of the slow-the-spread pilot project 

in 1993. Research and technology development 

accelerated during the pilot project (1993 to 1999). By 

the time slow the spread transitioned to operational 

status in 2000, mating disruption evolved into the 

treatment of choice. Between 2000 and 2004, mating 

disruption accounted for more than 84 percent of the 

total acres treated in association with slow-the-spread 

projects. During that period, a total of 1.7 million 

acres in the slow-the-spread area received treatment, 

with 1.4 million acres receiving mating disruption 

treatments. Only 15 uses of mating disruption as the 

primary treatment are recorded for eradication projects 

between 1983 and 2003. This represents less than 

2 percent of the total acreage treated for eradication 

during that period (U.S. Department of Agriculture, 

Forest Service 2003).

Mating disruption is accomplished either by ground 

or aerial application of controlled-release dispensers, 

formulated to slowly exude their load of active 

ingredient (racemic disparlure) into the environment. 

The formulation used for ground application consists 

of a laminated polymeric dispenser or tape impregnated 

with the pheromone for gradual release into the 

environment (USDA Forest Service 1989). The tape is 

manually attached to trees in a grid pattern, making this 

method labor intensive, especially in large treatment 

areas. An evaluation of this method concluded that 

additional research is needed before considering it for 

operational use (Kolodny-Hirsch and others 1990). 

The tape is no longer produced, but it is still registered 

with the U.S. EPA and could be made available in the 

future if requested from the manufacturer (Hercon 

Environmental Co, Emigsville, PA).

Although numerous controlled-release dispensers 

have been evaluated for use in aerial gypsy moth 

mating disruption projects, only Hercon’s Disrupt II 

is registered with EPA and available for commercial 

use. This formulation consists of a layer of resin 

impregnated with racemic disparlure sandwiched 

between two outer layers of plastic laminate. The 

laminate is chopped into small flakes; thus the 

commonly used term “pheromone flakes” or simply 

“flakes” when referring to Disrupt II treatments. 

Aircraft using custom-designed application equipment 

apply the flakes, which slowly release their load of 

pheromone into the environment over 3 to 4 months. 

Other promising formulations, such as microcapsules, 

microtubes, or emulsified concentrates, will continue to 

be evaluated for use in gypsy moth mating disruption 

projects.

Mating Disruption Effectiveness.

The effectiveness of mating disruption varies with the 

population density of gypsy moths in the treatment and 

surrounding areas. Mating disruption is only effective 

when used against very low-density populations of 

the gypsy moth; in higher-density populations where 

dozens of moths of both sexes may emerge on the same 

tree bole, the chance of male moths locating females 

is high. Therefore, mating disruption is best suited 

for areas that contain less than 10 egg masses per acre 

(25 egg masses/ha) (USDA Forest Service 1989). 

Population densities of this sort are typically found in 

the STS or eradication area, but not in the suppression 

area.

The trend in recent years has been toward lower doses 

for gypsy moth slow-the-spread projects. Until 1999, 

the standard dose used in slow-the-spread mating 

disruption projects was 30 grams active ingredient per 

acre (30 g a.i. per acre). Reduction of the standard 

dose by 50 percent started in 2000, to 15 g a.i. per 

acre. Further research confirms doses as low as 6 g a.i. 

per acre can effectively disrupt mating in low-density 

populations of the gypsy moth.

Mating disruption may be used alone or in conjunction 

with other treatments. Typically, it is used alone, 

but in some situations large infestations contain core 
Gypsy Moth Treatments and Application Technology

Appendix A - Page 

population(s) that already exceed the threshold at 

which mating disruption can be effective. In these 

cases, a small area treated with B.t.k., diflubenzuron, or 

Gypchek might be embedded within the boundaries of 

the larger mating disruption block.

Treatments using mating disruption as part of 

cooperative slow-the-spread projects between 1993 

and 2001, were at least as effective as treatments using 

B.t.k. Further, the frequency of repeated treatments 

was higher after using B.t.k. than mating disruption. 

This information must be further evaluated, considering 

that mating disruption is typically used on the lower 

population densities, whereas B.t.k. is used on both 

low and high population densities (Sharov and others 

2002a).

The use of disparlure as a mating disruption agent 

is desirable because the pheromone does not affect 

nontarget organisms. Once the pheromone dissipates, 

however, the plastic dispensers may remain in the 

environment for some time before disintegrating. 

Nonetheless, the use of this treatment will continue to 

be critical to STS projects where rare, threatened, or 

endangered species arecommonly encountered.

See Appendix H for the risk assessment on disparlure.

Sterile Insect Technique.

The sterile insect technique has not been used in 

recent years, but is available as a treatment tool for 

gypsy moth control. The objective of the sterile insect 

technique is to reduce the chance that female moths 

will mate with fertile males. Its success is more likely 

with the release of large numbers of sterile males in 

consecutive years. The resultant progressive reduction 

of fertile egg mass production leads to the eventual 

elimination of the population.

Sterile insect technique is ideally suited for application 

to gypsy moth populations with one generation per 

year. Male moths may mate several times; female 

moths usually mate only once and lay an egg mass that 

may contain up to 1,600 eggs (Reardon and Mastro 

1993). Recognition of the potential of this approach for 

managing low-density and isolated infestations of the 

gypsy moth took place in the mid-1950s. Treatment 

was not practical, however, until the development of 

methodologies for rearing large quantities of quality 

insects and quantifying the impact of the releases 

(Mastro and others 1981).

Sterile Insect Technique Use. 

One of three different approaches is selected (Reardon 

and Mastro 1993): (1) deploying male pupae sterilized 

by irradiation; (2) deploying male pupae irradiated, but 

not fully sterilized (substerile); or (3) broadcasting eggs 

from a female mated with an irradiated male (inherited 

sterility). None of these approaches is without 

biological or logistical limitations, which hamper 

operational use.

Sterile and Substerile Male Pupae. 

Initially, the sterile insect technique focused on 

deploying male pupae treated with a sterilizing 

dose of radiation. Pilot projects in Maryland, 

Michigan, and South Carolina during the 1970s and 

1980s demonstrated the efficacy of this technique. 

Nevertheless, the limited time period during which 

pupae must be released and the need to synchronize 

rearing of mass quantities of pupae for that release 

(treated pupae cannot be stockpiled) are obstacles to an 

operational program (Figure A-4) (Reardon and Mastro 

1993). A major logistical difficulty is the necessity of 

repeatedly releasing the treated insects over the 4-week 

flight period because male moths live only 2 to 3 days.

Deploying substerile insects is the preferred of the 

two techniques that release male pupae, because (1) 

the substerile insects suffer less tissue damage and are 

therefore more competitive than sterile males; (2) the 

progeny of substerile males and wild females develop 

in the field and are, in theory, hardy and in synchrony 
Appendix A

Appendix A - Page 

with the native population; and (3) the suppressive 

effect on the native population spans at least two life 

cycles (Knipling 1979, Snow and others 1971).

Inherited Sterility.

n induced inherited sterility (or F1 sterility), males 

are irradiated but not sterilized before they mate with 

non-irradiated females in the laboratory. More of the 

resulting progeny are sterile than in the treated parental 

generation, and the sex-ratio of the progeny is skewed 

in favor of males (LaChance 1985, North 1975). 

Release of F1 sterile eggs has advantages over the 

other two techniques: only a single release of treated 

gypsy moth eggs is required before wild eggs hatch, 

the production window is wider because eggs can be 

stockpiled, and the logistics of shipment and release are 

simpler.

Sterile Insect Technique Effectiveness. 

Between 1988 and 1992, eight isolated infestations of 

the gypsy moth were treated by releasing F1 sterile 

eggs, with favorable results; but numerous problems 

were identified (Reardon and Mastro 1993): (1) how to 

predict when wild eggs hatch, and how to synchronize 

release and hatching of eggs produced in the 

laboratory; (2) how to reduce mortality that occurs in 

early F1 instars; (3) dispersal of F1 young caterpillars 

and adult males; and (4) the relative competitiveness of 

caterpillars.

When evaluated against low-level gypsy moth 

populations in Virginia (Reardon 1991), results with 

substerile pupae generally proved more favorable than 

with F1 sterile eggs. Of the three approaches, the 

deployment of sterile male pupae is the least desirable. 

Release of F1 sterile eggs is preferred; however, the 

obstacles described are major impediments to more 

general use of this technique (Reardon and Mastro 

1993). The deployment of substerile pupae, in spite of 

its disadvantages, appears closest to operational use, 

although availability of substerile insects is limited.

Recent advances in insect engineering for use in sterile 

insect technique programs show promise for increasing 

the effectiveness and efficiency of the program. At 

this writing, however, no operational program has been 

developed for the gypsy moth.

A.2 The New Proposed 

Treatment of Tebufenozide.

Tebufenozide, like diflubenzuron, belongs to a 

group of compounds called insect growth regulators. 

Tebufenozide, which induces premature molts by 

direct stimulation of the ecdysteroid receptors (whereas 

diflubenzuron affects chitin synthesis at the regularly 

scheduled molt), mimics the action of a natural insect 

Figure A-4. Sterile gypsy moths are reared on artificial diet 

in a climate-controlled environmental chamber.
Gypsy Moth Treatments and Application Technology

Appendix A - Page 

hormone. Upon ingestion of tebufenozide, larvae stop 

feeding and undergo an early, incomplete and lethal 

molt.

Tebufenozide Use. 

Label instructions permit ground or aerial applications 

of tebufenozide. The labeled application rates for 

tebufenozide range from 0.06 lbs a.i. per acre to 0.12 

lbs a.i. per acre. Tebufenozide is applied to early first 

to third instar larvae.

Tebufenozide Effectiveness. 

Tebufenozide has not been used operationally by 

the USDA in suppression, eradication, or slow-thespread projects. Forest Service tests of tebufenozide 

at 0.06 lbs a.i. per acre generally found the product 

to be effective with 95 to 99 percent control achieved 

(Reardon and others 2000). See Appendix J for the risk 

assessment on tebufenozide.

A.3 Treatments That Include 

Natural Control Agents.

Few natural control agents accompanied the accidental 

introduction of the gypsy moth to this country. Some 

of those agents, as well as agents native to the United 

States, can play an important role in regulating gypsy 

moth populations throughout the generally infested 

area.

Fungal Pathogens. 

Fungal products labeled for use against the gypsy moth 

are not available at this writing. A fungus capable of 

infecting the gypsy moth is Entomophaga maimaiga

Humber, Shimazu, and Soper. This fungus, commonly 

found in Japan (Soper and others 1988), was brought 

to the United States in the early 1900s and released, 

but was not recovered until 1989 (Hajek and others 

1996). E. maimaiga is known to infect only the gypsy 

moth and other closely related caterpillars that spend 

significant periods of time on the soil surface (Reardon 

and Hajek 1993, Hajek and others 2000).

A field survey, conducted from 1989 to 1995 of 

lepidopteran cadavers infected with E. maimaiga, 

found three species of lymantriids, from the genus 

Dasychira, infected with E. maimaiga (Hajek and 

others 1996b). The field survey method was chosen 

because entomopathogens can infect hosts in the 

laboratory that are never found infected in the field. 

During a similar study in 1994, Hajek and others tested 

and found two species infected with E. maimaiga. 

Under laboratory conditions, conidia (infective spores) 

produced by an alternate host were determined to be 

ineffective (Hajek and others 1995b).

Epizootics of E. maimaiga in gypsy moth identified 

in the northeastern United States in 1989 represent 

the first reported occurrence of this fungus in North 

American gypsy moth populations (Andreadis and 

Weseloh 1990, Hajek and others 1990). Unlike the 

gypsy moth nucleopolyhedrosis virus, associated with 

high gypsy moth population densities, the fungus 

appears capable of causing dramatic mortality to 

middle-and late-stage gypsy moth caterpillars at low 

densities (Shimazu and Soper 1986). Since the fungus 

tends to cause mortality earlier than the virus, tree 

defoliation may not be as severe. Prediction of longterm impacts of E. maimaiga is inconclusive (Valenti 

1998).

A gypsy moth larvae infected with E. maimaiga

produces one or both types of spores—resting spores 

(azygospores) and conidia spores (infectious spores).

The age of the larvae is the primary factor in 

determining which type of spore is produced. Second 

instar larvae rarely contain resting spores, while fifth 

instar larvae produce resting spores when temperatures 

are increasing (Hajek and Shimazu 1996). The resting 

spore of the fungus overwinters on the bark of trees, in 

leaf litter, and in soil (Shimazu and others 1986). The 

resting spore germinates in the spring and produces a 

single conidium, which is released into the environment 

and may be carried in the air. Once on a susceptible 

caterpillar, the conidium germinates, penetrates the 
Appendix A

Appendix A - Page 10

insect’s skin, spreads throughout the caterpillar and 

kills it. Infected dying caterpillars typically hang with 

head down, in a stretched-out position on the stems of 

infested trees (Hajek and Roberts 1992). The fungal 

spores may remain alive in the soil for up to 10 years 

(Weseloh and Andreadis 2002).

With favorable conditions, high humidity and 

temperatures between 13 °C to 19 °C, the fungus grows 

out of the caterpillar through the skin and produces 

and releases more conidia, which may subsequently 

infect other caterpillars (Hajek and others 1996c). 

This secondary infection cycle is a major contributor 

to the dramatic epizootics observed in gypsy moth 

populations. When conditions are unfavorable, or midto-late June as the end of the feeding period of gypsy 

moth caterpillars approaches, E. maimaiga begins to 

produce resting spores inside the dead caterpillars, 

which slowly disintegrate and scatter the resting 

spores into the environment, with most accumulating 

in the soil. Laboratory determinations indicate that 

spores buried at least 1 cm below the surface are 

unable to infect gypsy moth larvae (Hajek and others 

1998a). These resting spores will not germinate for 

approximately 9 months after production (Hajek and 

Humber 1997). Diet of the gypsy moth larvae could 

also influence the development of E. maimaiga (Hajek 

and others 1995b).

Since 1989, the fungus has spread across a large portion 

of the generally infested area, apparently by spore 

movement on the wind and intentional introduction 

(Elkinton and others 1991, Smitley and others 1995), 

infecting gypsy moth throughout its range (Hajek 

and others 1999). Epizootics of E. maimaiga have 

occurred in New England and some Middle Atlantic 

States, and its distribution continues into areas more 

recently colonized by the gypsy moth. The fungus is 

so widespread in parts of Michigan and Virginia that 

it is difficult to determine whether the presence of the 

fungus at an individual location resulted from natural 

migration or spread from a release (inoculation) site 

(Reardon and Hajek 1995).

It is not clear why the fungus suddenly appeared 

almost 80 years after its initial introduction into the 

United States. Among the hypotheses offered, the 

most plausible may be these two: (1) a more aggressive 

strain of E. maimaiga arose through natural selection 

some time after its release in 1910-1911; or (2) more 

of the fungus was accidentally introduced (Hajek and 

others 1995a, Weseloh 1998b).

Numerous constraints limit the development of E. 

maimaiga for use as an insecticide (Reardon and Hajek 

1993). Fungi are often short-lived in storage and 

relatively expensive to produce, and foliar applications 

of fungi are sensitive to heat, humidity, sunlight, and 

rainfall. Formulation and application of dried fungal 

preparations also present the unique challenges of their 

adherence to leaf surfaces and protecting them from 

adverse environmental conditions.

The release of E. maimaiga into uninfested areas on 

a large scale is problematic as well. Because of its 

natural rate of spread, it is probably not necessary to 

physically introduce it into new areas. Intentional 

introduction of the fungus by moving soil or other 

inoculation into the soil would require registration 

and labeling of a product for this purpose with the 

Environmental Protection Agency (Podgwaite, John, 

Microbiologist, USDA Forest Service [Conversation 

with Joseph L. Cook]. 28 July 2004). Though E. 

maimaiga is a virulent pathogen of the gypsy moth, 

known to cause extensive epizootics in Japan (Shimazu 

and Soper 1986), it poses no known health risks to 

humans or pets.

E. maimaiga may eventually contribute to the longterm control of the gypsy moth. However, studies 

have only begun to identify the information about 

host-pathogen interactions that are vital to developing 

the fungus for effective biological control of the gypsy 

moth. Computer models can assist in management 

decisions by predicting short-term gypsy mothfungus interactions and the effectiveness of the fungus 

(Weseloh 2003b).
Gypsy Moth Treatments and Application Technology

Appendix A - Page 11

Parasitoids. 

Parasitoids live in or on another organism and benefit 

from the relationship, at a cost to the host, which 

often dies (Figure A-5). Two approaches used to 

introduce parasitoids into the gypsy moth population 

in North America are classic biological control and 

augmentation. The discovery, importation, release, and 

attempted establishment of exotic natural enemies of 

the gypsy moth are all part of classic biological control 

(Reardon 1981). Manipulation to initiate or increase 

effective biological control through established 

parasites is termed augmentation (Blumenthal and 

others 1981).

Parasitoids, in conjunction with other natural enemies 

(predators and pathogens), help regulate populations of 

the European strain of the gypsy moth by reducing their 

numbers. Most researchers do not believe that they 

play a major role in regulating gypsy moth populations 

(Elkinton and Liebhold 1990).

The rate of parasitism by a particular parasitoid 

species varies from site-to-site and from year-to-year, 

depending on such factors as the number of gypsy moth 

caterpillars, the number of alternative hosts, and the 

weather. Parasitoids are thought to help maintain lowdensity populations of the European strain, but do not 

prevent the buildup of already increasing populations 

(Campbell 1974b). The tachinid flies, Compsilura 

concinnata (Meigen) and Parasetigena silvestris

(Robineau-Devoidy), may play a role in suppressing 

incipient outbreak populations, but such population 

declines may go unnoticed (Elkinton and Liebhold 

1990).

The State of Massachusetts and the (then) Federal 

Bureau of Entomology initiated foreign exploration 

for gypsy moth parasitoids in 1904, and the effort 

continues today by the USDA. Over 250,000 

parasitoids of more than 85 species have been sent 

to the United States from collection areas around the 

world. Ten of these imported species were released 

and became established in the United States (Elkinton 

and Liebhold 1990). Additionally, several parasitoids 

native to the United States have become opportunistic 

parasitoids of the gypsy moth.

The principal egg parasitoids in North America 

are Ooencyrtus kuvanae (Howard) (Hymenoptera: 

Encyrtidae) and, to a much lesser degree, Anasatus 

disparis (Ruschka [Hymenoptera; Eupelmidae] ). O. 

kuvanae typically attacks 10 to 40 percent of the eggs 

in an egg mass (Brown 1984). The rate of parasitism 

is greater in the smaller egg masses typical of highdensity declining gypsy moth populations (Bellinger 

and others 1988, Brown and Cameron 1979).

Cotesia (Apanteles) melanoscelus (Ratzeburg) is a 

small braconid wasp that parasitizes early instar gypsy 

moth caterpillars, and has two generations per year. 

Hyperparasitoids, which prey on other parasitoids, 

severely reduce the numbers of C. melanoscelus

that overwinter (Weseloh 1983). Also limiting the 

wasps’ effectiveness is the poor synchronization of the 

parasitoid’s second generation with its host (Weseloh 

1976). Higher parasitism rates, however, reportedly 

occur when early gypsy moth instars are prolonged, as 

when they ingest sublethal doses of B.t.k. (Weseloh and 

Andreadis 1982).

Figure A-5. A parasitic wasp lays eggs on gypsy moth pupal 

case; eggs hatch into wasp larvae, which feed on and kill the 

host.
Appendix A

Appendix A - Page 12

Parasetigena silvestris (Diptera: Tachinidae) is a 

tachinid fly that lays an egg on the outer skin of the 

gypsy moth caterpillar, and has a single generation 

per year. Most active during daylight, the fly often 

causes more mortality than any other parasitoid. Peak 

parasitism tends to occur after gypsy moth populations 

decline from high densities (Elkinton and Liebhold 

1990). In Europe, parasitism by P. silvestris sometimes 

exceeds 95 percent (Bogenschutz and others 1989).

The tachinid fly, Blepharipa pratensis (Meigen) 

(Diptera: Tachinidae), is a major source of mortality in 

intermediate-density gypsy moth populations (Ticehurst 

and others 1978) (Figure A-6). It lays small eggs on 

foliage being fed upon by gypsy moth caterpillars. The 

eggs hatch after being ingested by caterpillars.

Brachymeria intermedia (Nees) (Hymenoptera: 

Chalcididae) is a small wasp that attacks gypsy moth 

pupae and other hosts. Introduced in 1908 but not 

recovered until 1942, it was abundant by 1971 (Doane 

1971). The parasitoid was observed causing high 

mortality of gypsy moths in Pennsylvania (Ticehurst 

and others 1978) and on Cape Cod (Elkinton and others 

1989). B. intermedia tends to be scarce in low-density 

gypsy moth populations (Elkinton and Liebhold 1990).

Lastly, the tachinid fly Comsilura concinnata (Diptera: 

Tachinidae) has many hosts and several generations 

per year; it can remain abundant when gypsy moth 

populations are low. This fly often causes higher 

mortality than other parasitoids in low-density gypsy 

moth populations (Elkinton and Liebhold 1990).

Augmentation of these established parasitoids has not 

proven to be an effective means to control gypsy moth 

populations (Blumenthal and others 1981). Classic 

biological control efforts continue to be an important 

avenue for study, and the search for and importation of 

gypsy-moth-specific natural enemies from Europe and 

Asia remains promising.

Predators. 

Many species of animals eat the gypsy moth as well 

as other forest-defoliating insects. Some predators 

feed on only one life stage of the gypsy moth, while 

others consume two or more life stages (Smith 

1985). Predation can help maintain sparse, stable 

gypsy moth populations indefinitely, though periods 

of low predatory pressure do not necessarily lead 

to an outbreak. Once an outbreak starts, as well as 

during subsequent outbreak decline, predation has no 

significant effect on population densities (Smith and 

Lautenschlager 1981).

The gypsy moth predator community is complex and 

includes approximately 50 species of birds, 20 species 

of mammals, some amphibians, reptiles, fish, insects, 

and spiders. Only a few of these predators are known 

to affect gypsy moth population dynamics (Elkinton 

and Liebhold 1990, Smith and Lautenschlager 1981). 

The predators are all opportunistic feeders, meaning 

that their taste for the gypsy moth depends upon the 

scarcity of preferred food. Robins, for example, may 

eat gypsy moth caterpillars when earthworms become 

scarce.

Bess and others (1947) first suggested that predation by 

small mammals is important to gypsy moth population 

dynamics in North America. Vertebrate predators, 

Figure A-6. Tachinid flies will parasitize gypsy moth 

caterpillars. (Mongolia).
Gypsy Moth Treatments and Application Technology

Appendix A - Page 13

especially the white-footed mouse (Peromyscus 

leucopus) (Figure A-7), are major sources of latelarval and pupal mortality in low-density gypsy moth 

populations (Campbell and Sloan 1977b, c, Campbell 

and others 1977), but not at higher gypsy moth 

densities (Campbell and others 1975, 1977). Small 

mammals help to maintain low-density gypsy moth 

populations (Elkinton and Liebhold 1990).

The earliest study of predation by birds, conducted by 

Forbush and Fernald in 1896, listed 38 bird species 

seen eating one or more life stages of the gypsy moth. 

Studies of bird predation tend to show that gypsy moth 

is not a major food item of most species (Cooper 1988). 

In feeding preference studies birds favored hairless 

caterpillars over gypsy moth caterpillars (Whelan and 

others 1989). Predation by birds is frequently cited in 

European literature as an important influence on gypsy 

moth population dynamics, but few studies exist to 

support that claim (Elkinton and Liebhold 1990).

The impact of invertebrate predators, such as ground 

beetles and ants, on gypsy moth pupae is less than that 

of vertebrates (Campbell and Sloan 1976, Elkinton 

and others 1989). Most predation by invertebrates 

occurs in leaf litter; little predation occurs in the tree 

canopy (Weseloh 1988). Adult and immature stages of 

Calosoma sycophanta (L.), a large, predaceous ground 

beetle introduced into North America from Europe, 

feed on gypsy moth caterpillars and pupae (Figure 

A-8). C. sycophanta populations increase in response 

to high-density gypsy moth populations and tend to lag 

1 to 3 years behind the onset of gypsy moth outbreaks 

(Weseloh 1985a, Smith and Lautenschlager 1978). The 

impact of C. sycophanta on low-density gypsy moth 

populations is thought to be minor (Weseloh 1985b, 

Smith and Lautenschlager 1978). Gypsy moth hairs 

defend the moth from spiders (Bardwell and Averill 

1996).

Predators can be encouraged by maintaining habitat 

diversity. People unknowingly destroy good habitat for 

predators by removing brush in an effort to “clean up” 

yards and woodlots. Such cleanup efforts significantly 

decrease the survival of small mammals and increase 

the survival of gypsy moths. For example, leaving 

dead “snag” trees increases populations of cavity 

nesting birds such as woodpeckers, which eat gypsy 

moths. Placing nesting boxes to supplement snags 

may also encourage cavity-nesting birds. Leaving 

piles of brush might encourage populations of small 

mammals, such as mice and shrews, which eat gypsy 

moths. Forest type may also affect predation (Liebhold 

and others 1998). Forest thinning does not affect 

predation, but it was found that invertebrates are the 

Figure A-7. White-footed mice feed on gypsy moth larvae.

Figure A-8. The Calosoma beetle is a gypsy moth predator 

introduced from Europe.
Appendix A

Appendix A - Page 14

main predators on larvae, and small mammals the 

major predators on pupae (Grushecky and others 1998). 

As vertebrate densities increase, invertebrate predation 

may decrease (Cook and others 1995, Hastings and 

others 2002).

Nematodes. 

Nematode results against defoliators such as the 

gypsy moth are inconsistent (Gaugler 1981, Kaya and 

Reardon 1982, Kaya and others 1981). Depending on 

the species, nematodes may actively search out their 

hosts and enter their body openings. In one study, 

two species of commercially available nematodes, 

Steinernema carpocapsae (Weiser) and S. feltiae

(Filipjev), were applied to cloth-lined burlap and plastic 

bands around the tree boles to infect resting gypsy 

moth caterpillars. The results were highly variable 

between trees, primarily due to the nematode’s need 

for a humid environment (Reardon and others 1986). 

Because nematodes may have potential for use against 

the gypsy moth, research continues.

Microsporidia. 

The gypsy moth was introduced into North America 

without the normal complement of natural enemies that 

help to regulate populations in Europe. There are many 

groups of entomopathogens (organisms that infect 

insects), viruses, fungi, and protozoans found in gypsy 

moth populations in Europe (especially during outbreak 

years) that are not found in this country.

Microsporidia (protozoa) are a diverse group of 

obligate intracellular parasites that use most animals 

(including insects) and humans as hosts and are 

relatively host specific. According to Maddox and 

others (1999), six species of microsporidia described 

from gypsy moth populations in Europe and several 

isolates that have not been described or identified are 

recorded in the literature; microsporidia have never 

been reported from gypsy moth populations in North 

America.

The significance of microsporidian pathogens 

as mortality agents of gypsy moths is frequently 

overlooked. Among the pathogens commonly found in 

European gypsy moth populations, microsporidia are 

prevalent during the gradation period prior to outbreaks 

and then persist at low levels among gypsy moth 

populations in the years between outbreaks. Different 

microsporidian species that infect the gypsy moth target 

various tissues within their host, including the silk 

glands, midgut and associated muscle tissue, body fat, 

nerve tissue, and reproductive organs. Several authors 

from Europe report that microsporidia caused over 80 

percent mortality of late-stage gypsy moth larvae in 

the Balkans and Ukraine, and caused high mortality in 

overwintering egg masses.

Because this strong evidence suggests that 

microsporidia are significant mortality factors in the 

dynamics of gypsy moth populations in central Europe, 

the USDA Forest Service initiated a foreign exploration 

program in 1993 to search for microsporidia in gypsy 

moth populations in several European countries. The 

program compares these isolates with previously 

described species and evaluates isolates that might 

be candidates for introduction as classical biological 

control agents, to enhance the natural control of this 

pest in North America.

Significant progress has been made in accumulating 

basic knowledge of the biology and life history of 

select isolates, which is necessary to resolve safety and 

regulatory issues of concern prior to consideration for 

possible introduction into the United States (McManus 

and Solter 2003). An extensive series of laboratory 

studies assesses the host specificity of select isolates 

against 49 species of non-target Lepidoptera known 

to occur in U.S. oak forests. Multi-year studies in 

Bulgaria and Slovakia evaluated the susceptibility of 

non-target forest Lepidoptera found with the gypsy 

moth in those countries. The development of molecular 

techniques aids in clarifying the taxonomy of European 

isolates and in fingerprinting individual isolates.
Gypsy Moth Treatments and Application Technology

Appendix A - Page 15

Preliminary discussions are ongoing with the EPA and 

APHIS to review the biological and ecological data 

accumulated on these gypsy moth pathogens during the 

past 10 years and to assess the feasibility of introducing 

them as classical biological control agents in small 

controlled experiments.

A.4 Miscellaneous Means of 

Gypsy Moth Management.

Removing and Destroying Egg 

Masses. 

One of the first gypsy moth treatments involved 

removing and destroying egg masses. Broad 

application of this technique to control the gypsy moth 

reached its zenith in the 1930s with the employment of 

Civilian Conservation Corps workers in New England 

during the fall, winter, and early spring, to seek out 

and destroy egg masses in towns and woodlands. The 

technique is labor- and time-intensive and impractical 

for large areas. Experience has shown that in a 

forested area, many more egg masses are present 

than are actually seen and disposed of, though the 

technique may be helpful in urban or suburban areas 

on accessible trees or ornamental plantings. Careful 

searching, removal, and destruction of egg masses may 

help reduce the potential for damage due to the gypsy 

moth in these situations.

Tree Trunk Bands and Barriers. 

As with removal and destruction of egg masses, 

removal and destruction of gypsy moth caterpillars may 

be useful in localized urban and suburban situations 

where small numbers of trees are at risk. The habit of 

caterpillars to move down from the crown and rest in 

protected areas during the day can be used to collect 

them. Bands, commonly of burlap, are placed around 

the trunks of susceptible trees to serve as resting areas 

for caterpillars seeking shelter. During an outbreak, 

the bands must be checked and the larvae need to be 

scraped off and killed. However, caterpillars may 

remain in the canopy and feed night and day during 

an outbreak, thus reducing the effectiveness of this 

method. Except as a survey tool, use of this technique 

in a forest situation is impractical.

A variety of trunk barriers is commercially available. 

Shown to be the most effective are barriers that include 

a sticky surface (Webb and Boyd 1983). An effective 

sticky barrier can be fashioned by wrapping the trunk 

of a susceptible tree with duct tape (to protect the bark 

and provide a smooth surface) and applying a thin 

layer of Tanglefoot. Gaps between the tape and the 

tree surface can be filled with fabric, polyester pillow 

stuffing, or any other suitable material. Trunk barriers 

should be placed just before gypsy moth eggs hatch, 

usually in March or April, depending on location. 

Insecticides can be combined with trunk barriers. A 

product that combines an insecticidal latex coating and 

burlap trunk barriers (White and others 1997) caused 

significant larval mortality for 30 days after application 

and reduced the need for manual removal of larvae.

While properly maintained sticky barriers are 

extremely effective at preventing caterpillars from 

climbing trees, they have no effect on caterpillars 

already in the canopy. For this reason, the impact of 

the barriers is usually limited to a 20- to 30-percent 

reduction in caterpillar numbers in treated trees 

over the season (Thorpe and Ridgway 1994, Thorpe 

and others 1993). The expected degree of foliage 

protection is even more variable, but usually averages 

20 to 30 percent as well. Therefore, while trunk 

barriers provide some benefit, they should never be 

relied upon as the sole method to protect foliage.

Broad-Spectrum Insecticides. 

A number of insecticides other than B.t.k., 

diflubenzuron, tebufenozide, and Gypchek are 

registered by the U.S. EPA for gypsy moth control. 

These include carbaryl, which was used in the past by 

USDA for gypsy moth management programs. Some 

insecticides are registered either for gypsy moth control 

or for control of pests where gypsy moth is likely to be 
Appendix A

Appendix A - Page 16

present; for example, in areas with susceptible shade 

or ornamental trees. Insecticides such as carbaryl and 

diazinon may be available for homeowners. All of 

these insecticides are excluded from this SEIS because 

they affect a wider range of non-target organisms 

than do B.t.k., diflubenzuron, and tebufenozide and, 

therefore, are not part of the USDA program. However 

any of these registered insecticides may be used by 

private applicators outside of the USDA program.

Silviculture. 

Silviculture is the practice of applying treatments to 

forest stands to maintain and enhance their utility for 

any purpose (Smith 1986). Silvicultural guidelines 

are designed to minimize the effects of the gypsy moth 

on forest stands and trees and are being evaluated for 

effectiveness. The guidelines recommend application 

of treatments to minimize gypsy moth impacts before, 

during, and after outbreaks (Gottschalk 1993).

The greatest number of silvicultural options for 

gypsy moth control are available before the insect 

becomes established in an area. Before outbreaks, 

silvicultural treatments may reduce stand susceptibility 

and vulnerability. Treatments might include these: 

increasing stand and tree vigor, removing trees most 

likely to die, reducing gypsy moth habitat (trees with 

large numbers of dead branches with rough peeling 

bark), reducing preferred gypsy moth food sources, 

improving predator and parasite habitats, regenerating 

stands that are close to maturity or understocked, and 

encouraging regeneration of nonpreferred gypsy moth 

food sources. Silvicultural considerations in urban 

and suburban areas include planting trees that are less 

susceptible to the gypsy moth. These silvicultural 

treatments should be conducted at least 2 years before 

gypsy moth arrives in an area to allow the remaining 

or newly planted trees to recover from the stress of 

treatment.

Silvicultural techniques vary according to the condition 

of the site, as in converting a stand to nonpreferred 

species. A thinning of healthy and vigorous sites 

performed 1 year before or after a gypsy moth outbreak 

enhances the vigor of the residual stand (Brooks 

and Hall 1997). When considering regeneration of 

preferred gypsy moth species, stump sprouts should 

be thinned to one stem per stump to improve vigor and 

resistance.

Another silviculture technique is the use of prescribed 

burns. This technique is occasionally used for oak 

regeneration. When a prescribed burn is properly used, 

it does not enhance susceptibility to gypsy moth.

Once the gypsy moth becomes established, or 

outbreaks occur or are imminent, silvicultural options 

are reduced. During outbreaks, silvicultural guidelines 

help prioritize stands that are candidates for receiving 

treatments and help determine if stands can be 

regenerated. Performing thinning during these times 

may reduce the density of egg masses.

Following gypsy moth outbreaks, silvicultural 

treatments focus on the efficient salvage of dead trees 

and the regeneration of stands that suffered heavy 

mortality or are close to maturity.

One advantage of silvicultural treatments is that action 

can be taken long before the gypsy moth arrives. 

Years might be required to treat large areas, as other 

resource considerations may limit the amount of cutting 

in an area. These treatments are prohibited in select 

areas, such as designated wilderness. Silviculture 

techniques are not quick-fixes for protection against 

gypsy moth or for gypsy moth suppression, but can be 

useful given proper planning.

A.5 Advances in Application 

Technology.

Advances in aerial application technology since 1990 

have dramatically improved the (pilot’s) applicator’s 

ability to control aerially applied sprays, thus reducing 

drift and minimizing unintended environmental 

consequences.
Gypsy Moth Treatments and Application Technology

Appendix A - Page 17

The primary advance came in the early 1990s with 

the introduction of Differential Global Positioning 

System (DGPS) navigation technology for use with 

aerial applicator apparatus. Prior to availability of this 

technology, the pilot used visual markers on the ground 

as guidance to direct spraying. Prior to this it was 

difficult to know the true location of an aircraft at any 

instant, and standard errors in absolute position were 

typically around 300 meters, with no truly accurate 

methodology to precisely log flight paths.

DGPS navigation systems revolutionized aerial 

application by providing knowledge of the aircraft’s 

absolute position within 1m (as of 2004) and the ability 

to log that information at 1Hz (once per second).

Use of the technology begins on the ground, marking 

block corners (defining the area to be treated) using 

DGPS and then moving this electronic file into the 

cockpit DGPS or by marking the desired block into a 

GIS system and loading that file. Most of the modern 

systems allow the pilot to view either a full map of the 

block in the cockpit or an idealized block outline. This 

map can be used to mark aviation hazards, landing 

pads, home base and other items. Current capabilities 

allow the pilot to automate flow control on and off 

functions, enabling the system to automatically turn on 

the spray at the block edge.

This technology is coupled with high-speed flow 

control, allowing precise application based on accurate 

aircraft position information. As powerful as the 

guidance functions are, the logging capabilities allow 

an operational manager to see exactly where the plane 

flew as well as evaluate the amount of material released 

throughout the entire operation.

Current state-of-the-art technology facilitates the use 

of highly accurate meteorological data and affords 

the ability to log release height and even provide 

predictions of spray movement after leaving the 

aircraft. These technologies greatly improve the ability 

of the pilot to execute an accurate, neat application—

reducing costs, drift, and unintended environmental 

effects (Thistle 2004).

Figure B-1. Early efforts to treat the gypsy moth followed a piecemeal approach 

that focused on roadsides and towns. 

Appendix B

Gypsy Moth 

Management Program
Figures

Figure B-1. Early efforts to treat the gypsy moth followed a 

piecemeal approach that focused on roadsides and towns.....................Cover

Figure B-2. As of 2006 the gypsy moth quarantine area covered 

all or parts of 19 States and the District of Columbia................................... 2

Figure B-3. Egg mass survey plots typically consist of 1/40-acre 

fixed radius plots (18.6 feet) throughout a sample area. The total 

sample is based on management goals of the site and distribution 

of host species............................................................................................... 4 

Figure B-4. Three strategies have proven successful against the 

gypsy moth: suppression in the generally infested area, slow the 

spread in the transition area, and eradication in the uninfested area ............ 5

Figure B-5. Slow-the-spread treatments are planned in a systematic 

step-wise fashion........................................................................................... 7

Appendix B Gypsy Moth Management 

Program

Contents

B.1 General...................................................................................................1

B.2 Prevention...............................................................................................3

Port-of-Entry Activities.....................................................................3

Regulatory Activities.........................................................................3

B.3 Survey.....................................................................................................3

Population Survey Within the Quarantine and Transition Areas......3

Larval Survey....................................................................................3

Detection Survey Outside the Quarantine Area................................3

Delimiting Survey.............................................................................3

B.4 Public Involvement and Notification......................................................4

B.5 Treatment Projects..................................................................................5

Suppression.......................................................................................5

Eradication........................................................................................6

Slow the Spread................................................................................6

B.6 Monitoring and Evaluation.....................................................................6

B.7 Assistance in Planning for Forests and Trees........................................6

B.8 Methods Development, Technology Transfer, and Research.................7

B.9 Information and Education.....................................................................7
Gypsy Moth Management Program

Appendix B - Page 

This appendix describes the activities that make up 

the Federal gypsy moth management program, which 

is conducted by agencies of the U.S. Department of 

Agriculture under authority of public law.

B.1 General. 

The gypsy moth (Lymantria dispar) is a nonnative 

invasive species. It was intentionally imported 

into North America in the late 1800’s by a 

private researcher. In 1868 or 1869, near Boston, 

Massachusetts, it escaped. The gypsy moth has spread 

steadily since that time. Various strains of the gypsy 

moth are defoliators of forest and shade trees on four 

continents (Asia, Africa, Europe, and North America). 

The gypsy moth can cause profound changes in forest 

ecosystems (Work and McCullough 2000) and, in the 

case of severe population outbreaks, adverse human 

health effects (Anderson and Furniss 1983; Tuthill and 

others 1984).

The United States Department of Agriculture (USDA) 

has played a role in gypsy moth management since 

1906, when Connecticut and Massachusetts first 

requested aid from the Federal government. The 

USDA Forest Service, the USDA Cooperative 

State Research, Education, and Extension Service 

(CSREES), Agricultural Research Service (ARS), and 

Animal and Plant Health Inspection Service (APHIS) 

all play a role. APHIS maintains a quarantine of the 

generally infested area and enforces regulations to 

prevent human-assisted spread of the pest (Figure B-2). 

In collaboration with State departments of agriculture, 

APHIS also implements an intense program for early 

detection and eradication of the moth when it is found 

outside the quarantine area. This monitoring program 

includes the deployment of approximately 225,000 

pheromone traps nationwide, outside the quarantine 

area. 

A memorandum of understanding between the Forest 

Service and APHIS identifies the roles and responsibilities in eradicating the European strain of the gypsy 

moth (USDA 1989). APHIS is responsible for conducting eradication projects on non-Federal lands when 

infestations cover less than 640 acres (259 ha). The 

Forest Service conducts eradication on National Forest 

System lands and cooperates with other agencies in 

projects on other Federal lands. The Forest Service also 

conducts eradication projects in cooperation with States 

on non-Federal land, when infestations cover 640 or 

more contiguous acres. 

The USDA Forest Service carries out activities to 

suppress gypsy moth populations on Federal lands 

within the quarantine area. The Forest Service also 

conducts research on gypsy moth and develops tools 

for forest managers and others to use to help manage 

the insect. Slow the spread, as the name implies, is a 

program implemented by the Forest Service and APHIS 

to reduce the natural and short range artificial rate of 

spread of gypsy moth populations from quarantine 

areas to adjacent non-infested areas. 

The CSREES provides technical information to 

businesses and landowners for management and 

eradication of gypsy moth on private property. ARS 

conducts research and evaluations on gypsy moth and 

development of tools to help manage the insect. ARS 

also conducts research and evaluations in support of the 

slow-the-spread strategy.

The USDA assigned responsibilities to these agencies, 

defined their roles to avoid duplication, and established 

the following policy by Departmental Regulation 

(USDA 1990): 

• Provide a comprehensive program of gypsy moth 

management activities coordinated by a designated 

lead agency (Forest Service) 

• Prevent establishment of gypsy moth outside the 

quarantine area;

• Develop and implement effective gypsy moth 

eradication and suppression measures;

• Conduct gypsy moth detection surveys and population 

assessments in cooperation with the States;
Appendix B

Appendix B - Page 

European Gypsy Moth (Lymantria dispar) Quarantine

• Protect Federal lands and assist States in protecting 

non-Federal lands from gypsy moth damage;

• Plan and conduct research on the gypsy moth in 

partnership with the agricultural experimental stations 

and other cooperators, to support Federal and State 

gypsy moth programs;

• Prevent further introduction of the gypsy moth from 

abroad;

• Coordinate research planning and cooperation within 

USDA and other Federal and State and private 

agencies;

• Emphasize research deemed necessary by Federal and 

State cooperators from the research, extension, and 

action communities;

• Follow an integrated pest management approach 

(USDA 1993)

USDA performs its duty as defined under authority 

provided by several statutes: 

• The Plant Protection Act (7 U.S.C. section 7701-

7759)—Prevent the introduction of pests into the 

United States, and prevent the movement of pests 

across state lines. 

• Cooperation with State Agencies in Administration 

and Enforcement of Certain Federal Laws (7 

U.S.C. section 450)—Enter into cooperative 

agreements with States to avoid duplication of 

functions, facilities, and personnel and to attain 

closer coordination and greater effectiveness in 

administering Federal and State laws and regulations 

to control or eradicate plant pests.

• The Cooperative Forestry Assistance Act of 1978 

(16 U.S.C. section 2101), as amended by the Forest 

Figure B-2. As of 2006 the gypsy moth quarantine area covered all or parts of 19 States and the District of Columbia.
Gypsy Moth Management Program

Appendix B - Page 

Stewardship Act of 1990 (16 U.S.C. section 2101)—

Assist in controlling forest insects and diseases 

directly on National Forest System lands, and in 

cooperation with other Federal Departments and 

States for control of pests on other Federal land and 

non-Federal lands of all ownerships. 

B.2 Prevention.

Port-of-Entry Activities.

APHIS is responsible for developing policies and 

operational guidelines to prevent the introduction of 

harmful, exotic agricultural quarantine organisms 

from entering at air, sea, and land border ports of 

entry. Vessels and cargo are inspected for gypsy 

moth contamination by the Department of Homeland 

Security, Customs and Border Protection.

Regulatory Activities. 

The Secretary of Agriculture is authorized to quarantine 

States or portions of States generally infested by the 

gypsy moth when necessary to prevent human assisted 

spread of gypsy moth. Regulated articles, such as 

nursery stock, trees without roots, outdoor household 

articles, mobile homes, logs, firewood, and pulpwood, 

are inspected for the presence of gypsy moth life 

stages. Articles found to be infested are treated or 

cleaned of the gypsy moth life stages before movement 

of the articles is permitted. Public information 

campaigns serve to increase awareness and compliance 

with regulatory efforts to prevent the spread of gypsy 

moth.

B.3 Survey.

Surveys are conducted to monitor gypsy moth 

populations and to determine the extent of infestations.

Population Survey Within the 

Quarantine and Transition Areas. 

The Forest Service monitors gypsy moth populations 

within the generally infested area to determine when 

suppression activities are warranted. The Forest 

Service also tracks incipient gypsy moth populations 

within the transition area to guide STS activities. The 

Forest Service is responsible for conducting surveys 

within the National Forests, on other Federal lands 

in cooperation with Federal agencies, and on nonFederal lands in cooperation with States. Surveys are 

accomplished in the generally infested area primarily 

by visual examination for egg masses (Figure B-3). 

Surveys are conducted in the transition area using 

specially designed traps baited with a manufactured 

version of the pheromone produced by the female 

gypsy moth to attract male moths. 

Larval Survey.

Larval surveys may be conducted to assess 

development of gypsy moth caterpillars to determine 

the proper timing for insecticide applications. Larval 

surveys use sticky bands, burlap, or similar material 

placed around the trunks of trees to capture the larvae 

(USDA APHIS 1990). 

Detection Survey Outside the 

Quarantine Area. 

APHIS and the Forest Service conduct detection 

surveys with pheromone traps to locate new 

infestations and monitor treated areas. APHIS is 

responsible for conducting detection surveys for gypsy 

moth on all lands outside the generally infested area 

(USDA 1989). State agencies cooperate on non-Federal 

lands, and Federal agencies cooperate on Federal lands. 

Detection surveys are conducted from late spring to 

late summer. 

Delimiting Survey. 

When adult male gypsy moths are caught, a delimiting 

survey using pheromone-baited traps may be used to 

confirm the presence of a reproducing population, the 
Appendix B

Appendix B - Page 

approximate size of the population, and the geographic 

range of the infestation. The information from the 

delimiting survey is used to design the appropriate 

eradication treatment. Delimiting surveys are 

conducted in cooperation with the respective State 

governments.

B.4 Public Involvement and 

Notification. 

The Forest Service and APHIS actively seek public 

participation at the local level for planned treatment 

projects. Before suppression, eradication, or slowthe-spread projects are carried out, public outreach 

is carried out and generally includes the following 

actions:

• Convening of public meetings facilitated by USDA;

• Identification of Federal officials who may be 

contacted to answer questions;

• Notification about planned treatment activities in 

local newspapers, and through newsletters and other 

media such as radio or television;

• Provision of the SEIS, environmental assessments, 

and related documents to agencies, groups, and 

individuals who are interested in the proposed action;

• Announcement of treatment dates and times to make 

it possible for those with concerns about insecticide 

application to avoid exposure. 

Public meetings facilitated by USDA may include these 

elements:

18.6 feet

18.6 feet

Figure B-3. Egg mass survey plots typically consist of 1/40-acre fixed radius plots (18.6 feet) throughout a sample area. The 

total sample is based on management goals of the site and distrution of host species.
Gypsy Moth Management Program

Appendix B - Page 

• Presentation of the reason for the treatment project 

and its objective;

• Discussion of the recommended treatment and 

various alternatives, and their consequences;

• Soliciting of public input to identify local issues that 

should be addressed in the design and deployment of 

the project;

• Review of the details of the implementation 

procedure and the timing of activities.

B.5 Treatment Projects. 

Any of the treatments authorized under the USDA 

gypsy moth management program may be conducted 

under any one of the strategic objectives of 

suppression, eradication, or slow the spread (Figure 

B-4). These strategies include planning, detection, 

evaluation, monitoring, and using appropriate 

methods to prevent establishment of new infestations, 

reduce damage caused by outbreaks, and slow the 

natural and short range spread of the gypsy moth. 

A project authorized under the program must be 

developed in compliance with Federal statutes, such 

as the Endangered Species Act and the National 

Environmental Policy Act. 

Suppression. 

The objective of suppression is to reduce damage 

caused by outbreak populations of gypsy moth in 

the generally infested area, thus minimizing severe 

defoliation of trees. Suppression does not attempt to 

eliminate the gypsy moth from the generally infested 

area, but reduces damage to ecosystems and effects on 

people. 

Participation of State and Tribal governments or other 

Federal agencies in cooperative suppression projects 

is voluntary; private landowners may participate by 

coordinating with State and local agencies. In some 

communities, however, local nuisance ordinances 

or other orders may not permit private landowners 

to voluntary withdraw from treatments. Within the 

generally infested area, USDA provides assistance to 

Federal, Tribal, and State agencies for suppression 

projects wherever gypsy moth outbreaks cause 

unacceptable levels of defoliation, by conducting 

projects in residential and recreational areas, forests, 

and special use areas, such as scenic byways and 

watersheds.

Site-specific environmental analyses are prepared by 

Federal resource managers on Federal lands, by forest 

supervisors on National Forests, and by Forest Service 

regional foresters or the Northeastern Area director 

on State and private lands. Gypsy moth populations 

are suppressed directly by the Forest Service on 

National Forest System lands, in cooperation with State 

agencies on non-Federal lands, and in cooperation with 

responsible officials on other Federal and Tribal lands. 

Proposed suppression projects must meet these criteria 

to be considered by the U.S. Department of Agriculture 

for funding (USDA Forest Service 1990b):

• Show strong potential for effective control

• Be supported by a biological evaluation that 

substantiates the need for the project

• Be environmentally acceptable, having met 

requirements of the National Environmental Policy 

Act

• Be supported by economic analysis, and a project 

work and safety plan.

Figure B-4. Three strategies have proven succesful against 

the gypsy moth: suppression in the generally infested area, 

slow the spread in the transition area, and eradication in the 

uninfested area.

Suppression

(Generally Infested Area)

Eradication 

(Uninfested Area)

Slow-the-Spread

(Transition Area)
Appendix B

Appendix B - Page 

Eradication. 

The objective of eradication projects is to eliminate 

infestations detected in the uninfested area of the 

United States. The most common cause of isolated 

infestations is people moving egg masses or pupae 

on outdoor household articles, recreational vehicles, 

and boats, from the generally infested area to the 

uninfested area. Locations most likely to have isolated 

infestations in the future are wooded residential areas 

with high rates of relocation by people, as well as 

sawmills, plant nurseries, mobile home parks, and 

tourist attractions such as campgrounds and State and 

National parks.

Participation in eradication projects is governed 

by State law and by policies and regulations of the 

cooperating State agency. In some states, participation 

of land owners in eradication projects may be 

mandatory; if it is determined that State actions are 

inadequate, the U.S. Secretary of Agriculture can 

declare an emergency and conduct an eradication 

project.

Eradication activities may also target the Asian strain 

of the gypsy moth in the area generally infested by the 

European gypsy moth, as well as in the uninfested area. 

Eradication projects are conducted in cooperation with 

Federal and State agencies and based on the availability 

of Federal funds, a mutually agreed-upon plan of 

work, and the results of site-specific environmental 

analyses conducted in accordance with the National 

Environmental Policy Act (CEQ 1992). 

Slow the Spread. 

The objectives of slow the spread are to slow the 

natural and short-range artificial spread of the European 

strain of the gypsy moth from the generally infested 

area to uninfested areas, and to delay the adverse 

effects associated with infestations of new areas. 

Mating disruption and B.t.k. are used most frequently 

in slow-the-spread projects. Mating disruption 

is accomplished by the application of tiny flakes 

containing disparlure to confuse male gypsy moths and 

disrupt their normal mate-search behavior, preventing 

them from finding and mating with females. Slowthe-spread treatments have reduced the historic rate of 

spread of 13 miles per year (20.9 kilometers/year) to 

less than 6.25 miles per year (10.1 kilometers/year), as 

the gypsy moth moves into previously uninfested areas 

(Sharov and others 2002b).

Slow-the-spread treatments are applied in the 

transition area (also called the slow-the-spread action 

zone). When detected, gypsy moth populations 

are further delineated, then treated to eliminate the 

moths and retard their spread (Figure B-5) caused by 

“leapfrogging,” which occurs when recently established 

populations (beyond the expanding population front) 

grow and coalesce, contributing to the movement of the 

population front (Sharov and Liebhold 1998). A more 

detailed description of slow the spread and how the 

program works can be found on page 32 of Sharov and 

others (2002b).

Slow the spread includes conducting intensive surveys 

with pheromone-baited traps to detect low-level gypsy 

moth populations in the transition area. Populations 

meeting specific criteria (based on counts of male 

moths, or other life stages, or both) are treated. 

B.6 Monitoring and Evaluation. 

The Forest Service and APHIS monitor treatment 

projects, with particular attention to those in 

environmentally sensitive areas, to ensure treatments 

are executed as prescribed. Environmental monitoring 

determines treatment effects and evaluates treated areas 

to assess project effectiveness.

B.7 Assistance in Planning for 

Forests and Trees.

The Stewardship Program, led by the Forest Service 

in cooperation with the States, provides technical and 

financial assistance for forest management planning. 
Gypsy Moth Management Program

Appendix B - Page 

These programs furnish an opportunity to assess 

potential damage from the gypsy moth and to develop 

contingency management plans.

The Urban and Community Forestry Program, also led 

by the Forest Service, encourages replacing susceptible 

tree species with resistant or less susceptible species 

(USDA Forest Service 1993). In keeping with the 

Forest Service’s philosophy of ecosystem management, 

long-range tree care plans and continued inventories 

need to emphasize species that are less preferred by 

gypsy moth caterpillars. Financial and technical 

assistance, with the gypsy moth as a major management 

consideration, are available to municipalities, school 

districts, communities, and nonprofit organizations (but 

not individual landowners) for managing individual 

trees or groups of trees on non-Federal lands in urban 

environments.

Figure B-5. Slow-the-spread treatments are planned in a 

systematic step-wise fashion.

B.8 Methods Development, 

Technology Transfer, and 

Research.

The Forest Service, APHIS, and Agricultural Research 

Service (ARS) research and identify new or improved 

methods of dealing with the gypsy moth. The Forest 

Service and APHIS also implement new technology 

required to support gypsy moth management activities 

(USDA 1990). Forest Service research develops ways 

to manage the gypsy moth where forests and wildlands 

meet urban areas, emphasizing safe and cost-effective 

practices that prevent populations from increasing 

above harmless levels and that suppress outbreaks. 

ARS develops the means to protect high-value trees 

for yards, communities, parks, and other nonforest 

environments and technology to support the activities 

of the Forest Service and APHIS. The APHIS Methods 

Development Center emphasizes development of gypsy 

moth trapping technology, pheromones, and rearing and 

monitoring techniques.

B.9 Information and Education.

USDA agencies participating in the Department’s gypsy 

moth program conduct information and education 

activities to support their specific management 

responsibilities. Activities include these: developing, 

printing, and distributing technical publications, 

research reports, and briefs on the gypsy moth and 

gypsy moth management, preparing and distributing 

slide programs and videos for use in public information 

and education activities, developing computer software 

programs and geographic information systems to assist 

in gypsy moth management, making presentations 

and participating in gypsy moth workshops, and 

participating in public meetings and hearings. 

2-km

trap grid

Isolated

Colony

Detected

Delimitation

Survey

Treatment

Transition Zone populations are 

recently established, still at low 

levels, and discontinuous from 

one another.

STS uses grids of pheromone 

traps to locate and delimit 

isolated populations.

Once these populations have 

been delineated, they are 

eradicated before they grow too 

large.

STS locates and retards isolated 

populations in the transition zone 

to prevent them from growing 

and coalescing, thereby slowing 

the spread of gypsy moth.

STS Treatment Grid

Appendix C

Public Involvement

and Issues

Figure C-1. This undated photo shows woodland defoliation caused by gypsy 

moths in Princeton, Massachusetts. 
Figure

Figure C-1. This undated photo shows woodland defoliation 

caused by gypsy moths in Princeton, Massachusetts.............................Cover 

Appendix C Public Involvement and Issues

Contents

C.1 Public Involvement Activities.................................................................1

Informational Bulletins......................................................................1

Periodic Press Releases......................................................................1

C.2 Outcome and Analysis of Public Involvement Activities.......................2

Issue 1—Risk to Human Health........................................................2

Issue 2—Risk to Nontarget Organisms.............................................2

C.3 Draft SEIS Public Involvement and Comment Analysis.......................2
Public Involvement and Issues

Appendix C - Page 

This appendix describes the public involvement 

activities that were planned and carried out for this 

SEIS. It states three issues that arose from the initial 

comments. It also summarizes comments received on 

the draft SEIS.

C.1 Public Involvement 

Activities. 

Planned activities inform the public and create a 

process to enable comment by individuals and groups 

with concerns, suggestions, and ideas for shaping the 

content of the gypsy moth supplemental environmental 

impact statement (SEIS). To identify and reach the 

interested and affected public across the United States, 

the interdisciplinary team joined with public affairs and 

forest pest management contacts throughout the Forest 

Service and the Animal and Plant Health Inspection 

Service (APHIS) (see Chapter 5 for names of those 

who contributed to this document). This network also 

provided technical review and guidance to ensure this 

SEIS serves all areas of the United States. A public 

outreach plan was developed and implemented in 

June 2004. A national mailing list was compiled, and 

informational materials prepared about the SEIS project 

and the gypsy moth. 

On April 29, 2004, the Forest Service and APHIS 

published a Notice of Intent (NOI) to Prepare 

a Supplement to the Final EIS for Gypsy Moth 

Management in the United States: a Cooperative 

Approach (69 Federal Register (FR) 23492-93, April 

29, 2004). The public was invited to comment on the 

proposed supplement. Other NOIs were published on 

March 13, 2006 (71 FR 12674-75) and on February 

7, 2007 (72 FR 5675), revising the dates for filing the 

draft and final SEIS. 

Using a mailing list developed for the SEIS, an 

informational bulletin asking for comments was mailed 

to nearly 13,000 individuals and organizations in May 

2004, including scientists, members of conservation 

and environmental groups, persons working in forestry 

and related industries, homeowners, landowners, over 

2,000 libraries, and Federal, State, and local officials. 

A distribution of letters to personnel within the Forest 

Service and APHIS solicited their input.

Team members personally met with Forest Service 

officials in Regions 1, 2, 3, 4, 5, 6, 8 and 9, 

Northeastern Area State and Private Forestry field 

offices, and APHIS representatives from the same 

areas. Team members conferred with agencies of 

25 different States interested in the SEIS and gave 

presentations at several meetings and conferences on 

the gypsy moth. Attendees represented (at least) an 

additional 21 States, APHIS personnel from across 

the country, and additional Forest Service and USDA 

personnel with gypsy moth management and research 

duties from across the country. Because gypsy moth 

management occurs on Department of Defense lands, 

the team delivered a presentation, by invitation, to 

representatives from the Marine Corps, Air Force, 

Army, and Navy, at the 2004 Department of Defense 

Pest Management Workshop and Entomology Meeting.

Informational Bulletins.

Informational bulletins were developed and mailed 

throughout the development of the SEIS. Three 

informational bulletins have been mailed to date 

(March 15, 2007): the first conveyed information about 

the April 29, 2004, Notice of Intent to prepare an SEIS; 

the second bulletin provided information about the 

biology, host preferences, and current distribution of 

the moth; the third covered gypsy moth management; 

and the fourth covered gypsy moth research. A future 

informational bulletin mailing is planned on the APHIS 

gypsy moth program.

Periodic Press Releases.

Media releases are planned to advise of the availability 

of both the draft and final SEIS to generate public 

involvement.
Appendix C

Appendix C - Page 

C.2 Outcome and Analysis of 

Public Involvement Activities.

The initial comment period concluded in June 2004; all 

comments were acknowledged by postcard. Comments 

and suggestions identified from these letters and 

various meetings and conferences were grouped under 

two significant issues. Significant issues were defined 

as those directly or indirectly caused by implementing 

the proposed action. No nonsignificant issues were 

identified. Nonsignificant issues would have included 

those ... (1) outside the scope of the proposed action; 

(2) already decided by law, regulation, Forest Plan, 

or other higher level decision; (3) irrelevant to the 

decision to be made; or (4) conjectural and not 

supported by scientific or factual evidence.

The Forest Service identified the following significant 

issues during scoping:

Issue 1— Risk to human health. 

Issue 2— Risk to nontarget organisms. 

Issue 1—Risk to Human Health.

The issue of human health includes the potential 

effects from contact with the gypsy moth and from 

exposure to treatments. Effects are measured by risk 

assessments (RAs) done for the gypsy moth and each 

of the treatments to include hazard identification, 

exposure assessment, dose-response assessment, and 

risk characterization. Included are the potential effects 

on project workers, the general public, and groups 

of people who may be at special or increased risk. 

The potential high risk group includes those who are 

sensitive to specific chemicals and those with multiple 

chemical sensitivity. Mitigation measures that can be 

implemented to lessen or remediate effects on human 

health are identified in Chapter 2. 

Issue 2—Risk to Nontarget 

Organisms.

The issue of nontarget organisms includes potential 

effects due to the gypsy moth and the treatments on 

mammals, birds, terrestrial invertebrates, fish, and 

aquatic invertebrates. These effects are measured by 

risk assessments (RAs) done for the gypsy moth and 

each of the treatments, to include hazard identification, 

exposure assessment, dose response assessment, and 

risk characterization. Mitigation measures that can 

be implemented to remediate effects on nontarget 

organisms are identified in Chapter 2. 

C.3 Draft SEIS Public 

Involvement and Comment 

Analysis.

THIS SECTION WILL BE COMPLETED WHEN 

LETTERS ARE RECEIVED FOR THE DRAFT SEIS 

AND MADE AVAILABLE WITH THE FINAL SEIS
Appendix D

Plant List

Figure D-1. White oak is one of the gypsy moth caterpillar’s preferred foods. 
Appendix D Plant List

Figure

Figure D-1. White oak is one of the gypsy moth caterpillar’s 

preferred foods.......................................................................................Cover
Plant List

Appendix D - Page 

Genus and species Common name Susceptibility 

index

Abelia grandiflora glossy abelia 3

Abies amabilis Pacific silver fir; silver fir; lovely fir; amabilis fir 2

Abies balsamea balsam fir; Canada balsam; eastern fir 3

Abies balsamea 

var. phanerolepis

balsam fir; bracted balsam fir 3

Abies bifolia Rocky Mountain subalpine fir 3

Abies bracteata bristlecone fir; Santa Lucia fir; silver fir 2

Abies chinensis 

 var. grandiflora 

Glossy abelia 3

Abies concolor white fir; concolor fir; silver fir 2

Abies fraseri Fraser fir; southern balsam fir; southern fir 3

Abies grandis grand fir; lowland white fir; lowland fir; balsam fir 2

Abies holophylla needle fir; Manchurian fir 2

Abies lasiocarpa subalpine fir; alpine fir; balsam fir; white balsam fir; Rocky 

Mountain fir

2

Abies lasiocarpa 

var. arizonica

corkbark fir 2

Abies lowiana California white fir; white fir; Sierra white fir 2

Abies magnifica California red fir; red fir; silvertip; golden fir 2

Abies procera noble fir; red fir; white fir 2

Acacia baileyana Bailey acacia; cootamundra wattel 2

Acacia farnesiana huisache; sweet acacia; Texas huisache; cassie 2

Acacia greggii Gregg catclaw; catclaw acacia; Texas catclaw; devilsclaw; 

long-flowered catclaw

2

This appendix lists the susceptibility of plant species 

to feeding by gypsy moth caterpillars (Liebhold and 

others 1995). The susceptibility index, based on 

preference and weight gain of both European and Asian 

strains of the gypsy moth, takes into account preference 

variances between strains. The index numbers provide 

a general ranking:

1 – Susceptible (these are plants the gypsy moth 

prefers to eat)

2 – Resistant (although not preferred by the gypsy 

moth, it will eat these plants) 

3 – Immune (these species of plants are not eaten 

under any circumstances) 

The index terms, suggested by Montgomery (1991), 

indicate the likelihood of plant defoliation. Plant 

names were selected from several sources (Dirr 1990, 

Little 1979, Rehder 1951, Taylor 1961, Van Dersal 

1938, Viertel 1979).
Appendix D

Appendix D - Page 

Genus and species Common name Susceptibility 

index

Acacia longifolia golden wattle; Sydney golden wattle 2

Acacia spp. acacia 2

Acacia tortuosa huisachillo; catclaw; twisted acacia; Rio Grande acacia 2

Acacia wrightii Wright catclaw; Texas catclaw; Wright acacia 2

Acer barbatum Florida maple; sugar maple; hammock maple 2

Acer campestre hedge maple; English field maple 2

Acer circinatum vine maple 2

Acer dasycarpum silver maple; cut-leaf maple 2

Acer ginnala amur maple 3

Acer glabrum Rocky Mountain maple; dwarf maple; mountain maple; Sierra 

maple

2

Acer grandidentatum canyon maple; bigtooth maple; sugar maple; Uvalde bigtooth 

maple

2

Acer japonicum fullmoon maple 2

Acer leucoderme chalk maple; white-bark maple 2

Acer macrophyllum bigleaf maple; Oregon maple; broadleaf maple 2

Acer negundo boxelder; ash-leaved maple; boxelder maple; Manitoba maple 2

Acer nigrum black maple; black sugar maple; hard maple; rock maple 2

Acer palmatum Japanese maple 2

Acer pensylvanicum striped maple; moosewood 3

Acer platanoides Norway maple 2

Acer pseudoplatanus planetree maple; sycamore maple 2

Acer rubrum red maple; scarlet maple; swamp maple; soft maple 2

Acer saccharinum silver maple; soft maple; river maple; silverleaf maple 3

Acer saccharum sugar maple; hard maple; rock maple 2

Acer spiatum mountain maple; moose maple 3

Acer tartaricum tartarian maple; Tartar maple 2

Achras emarginata wild-dilly 2

Acoelorrhaphe wrightii paurotis palm 3

Adonica merrillii Manila palm 3

Aesculus californica California buckeye 3

Aesculus glabra Ohio buckeye; fetid buckeye; stinking buckeye; American 

horsechestnut

2
Plant List

Appendix D - Page 

Genus and species Common name Susceptibility 

index

Aesculus 

hippocastanum

horsechestnut; common horsechestnut 3

Aesculus octandra yellow buckeye; sweet buckeye; big buckeye 3

Aesculus sylvatica painted buckeye; dwarf buckeye; Georgia buckeye 2

Ailanthus altissima ailanthus; tree of heaven; Chinese tree-of-heaven; copaltree 2

Albizia julibrissin silktree; mimosa; mimosa-tree; powderpuff-tree 3

Aleurites fordii tung-oil-tree; tungtree 2

Alnus maritima seaside alder 1

Alnus oblongifolia Arizona alder; Mexican alder; New Mexican alder 1

Alnus rhombifolia white alder; Sierra alder 2

Alnus rubra red alder, Oregon alder, western alder, Pacific Coast alder 1

Alnus rugosa speckled alder; smooth alder; tag alder; gray alder; hoary 

alder; hazel alder

1

Alnus serrulata hazel alder; smooth alder; common alder; tag alder; black 

alder

2

Alnus sinuata Sitka alder; mountain alder, wavyleaf alder 2

Alnus tenuifolia mountain alder; thinleaf alder; river alder 1

Alvaradoa amorphoides Mexican alvaradoa 2

Amelanchier alnifolia western serviceberry; saskatoon serviceberry; serviceberry; 

juneberry; western shadbush

2

Amelanchier arborea downy serviceberry; Allegheny serviceberry; shadblow; apple 

shadbush

2

Amelanchier 

canadensis

thicket serviceberry; oblongleaf juneberry 2

Amelanchier laevis Allegheny serviceberry; downy serviceberry; smooth 

serviceberry

2

Amelanchier spp. serviceberry 2

Amphitecna latifolia black calabash 3

Amyris elemifera torchwood; candlewood; sea amyris 2

Annona glabra pond-apple; alligator-apple 2

Aralia spinosa devils-walkingstick; Hercules-club; prickly-ash; angelica-tree 3

Arbutus arizonica Arizona madrone; madrona; Arizona madrono 2

Arbutus menziesii Pacific madrone; madrone; madrona 1

Arbutus texana Texas madrone; madrona 2
Appendix D

Appendix D - Page 

Genus and species Common name Susceptibility 

index

Arbutus unedo strawberry madrone; strawberrytree 2

Ardisia escallonioides marlberry; marbleberry 2

Ardisia japonica Japanese ardisia; marlberry 3

Arecastrum 

romanzoffianum 

queen palm 3

Asimina triloba pawpaw; common pawpaw; pawpaw apple; false-banana 2

Avicennia nitida black-mangrove; blackwood 2

Betula alba European white birch; white-barked canoe birch; cut-leaved 

birch

2

Betula alleghaniensis yellow birch; gray birch; silver birch; swamp birch 3

Betula caerulea blueleaf birch 1

Betula eastwoodiae Yukon birch 1

Betula lenta sweet birch; black birch; cherry birch 2

Betula nigra river birch; red birch; black birch; water birch 1

Betula occidentalis water birch; red birch; black birch; spring birch; paper birch 2

Betula papyrifera paper birch; canoe birch; white birch; silver birch 1

Betula pendula European birch; European white birch; cut-leaf weeping 

birch; blueleaf birch

1

Betula populifolia gray birch; grey birch; white birch; wire birch; fire birch; 

oldfield birch

1

Betula pumila swamp birch; bog birch 1

Betula verrucosa European white birch 1

Bourreria ovata Bahama strongback; Bahama strongbark; strongback 3

Broussonetia papyrifera paper mulberry; common paper mulberry 3

Bumelia lanuginosa gum bumelia; woolly buckthorn; chittamwood; swiftwiggum; gum elastic; buckthorn

2

Bursera simaruba gumbo-limbo; West-Indian-birch; gum-elemi 2

Callitris glaucophylla white cypress-pine 3

Calocedrus decurrens incense-cedar 3

Calycanthus floridus common sweetshrub; Carolina allspice; hairy (Caroline) 

allspice

3

Calyptranthes pallens pale lidflower; spicewood; white spicewood 2

Calyptranthes zuzygium myrtle-of-the-river, spicewood 2

Canella winterana canella; cinnamonbark; wild-cinnamon 2
Plant List

Appendix D - Page 

Genus and species Common name Susceptibility 

index

Canotia holacantha canotia; Mohave thorn; crucifixion-thorn 2

Capparis 

cynophallophora

Jamaica caper; capertree; Jamaica capertree 2

Caragana arborescens peatree; peashrub; Siberian peashrub; Siberian pea tree 2

Carica papaya papaya; pawpaw 2

Carpinus caroliniana American hornbean 2

Carya aquatica water hickory; bitter pecan; swamp hickory; bitter water 

hickory

2

Carya cordiformis bitternut hickory; bitternut; swamp hickory; pignut; pignut 

hickory

2

Carya floridana scrub hickory; Florida hickory 2

Carya glabra pignut hickory; pignut 2

Carya illinoensis pecan; sweet pecan 2

Carya laciniosa shellbark hickory; big shellbark hickory; king nut hickory; big 

shagbark hickory

2

Carya leiodermis pignut hickory; swamp hickory 2

Carya myristiciformis nutmeg hickory; swamp hickory; bitter water hickory 2

Carya ovalis red hickory; small pignut; sweet pignut 2

Carya ovata shagbark hickory; shellbark hickory; upland hickory; 

scalybark hickory

3

Carya pallida sand hickory; pignut hickory; pale hickory; pallid hickory 2

Carya spp. hickory 2

Carya texana black hickory; bitter pecan; Buckley hickory; pignut hickory 2

Carya tomentosa mockernut hickory; mockernut; white hickory; whiteheart 

hickory

2

Caryota urens toddy palm; white palm; fishtail palm; wine palm 3

Castanea dentata American chestnut; chestnut 2

Castanea ozarkensis Ozark chinkapin; Ozark chestnut 2

Castanea pumila Allegheny chinkapin 2

Castanopsis 

chrysophylla

giant chinkapin; golden chinkapin; giant evergreen chinkapin 1

Casuarina equisetifolia horsetail casuarina; beefwood; Australian pine; horsetail-tree 2

Casuarina stricta coast beefwood 2

Catalpa bignonioides southern catalpa; common catalpa; catawba; Indian-bean; 

cigartree

3
Appendix D

Appendix D - Page 

Genus and species Common name Susceptibility 

index

Catalpa speciosa northern catalpa; hardy catalpa; western catalpa; catawba 3

Catalpa spp. catalpa; hardy catalpa 3

Ceanothus arboreus feltleaf ceanothus; island myrtle; Catalina ceanothus 3

Ceanothus integerrimus deer brush 3

Ceanothus maritimus ceanotus 2

Ceanothus spp. ceanothus 3

Ceanothus thysiflorus blueblossom; blue-myrtle; blue-brush; blueblossom ceanothus 3

Cedrus atlantica atlas cedar 2

Cedrus deodara deodar cedar 2

Cedrus libani Cedar of Lebanon 2

Celtis laevigata sugarberry; southern hackberry; Mississippi hackberry; Texas 

sugarberry

3

Celtis occidentalis hackberry; northern hackberry; sugarberry; nettletree 3

Celtis tenuifolia Georgia hackberry; dwarf hackberry; upland hackberry 3

Cephalanthus 

occidentalis

buttonbush; buttonball bush; honey-balls; globeflowers 2

Cercidium floridum blue paloverde; Texas paloverde; paloverde 2

Cercidium 

microphyllum

yellow paloverde; littleleaf hornbeam; foothill paloverde; 

littleleaf paloverde

2

Cercis canadensis eastern redbud; redbud; Judas tree 3

Cercis occidentalis California redbud; western redbud; Arizona redbud 3

Cercocarpus betuloides birchleaf cercocarpus; birchleaf mountain-mahogany; 

alderleaf cercocarpus 

2

Cercocarpus breviflorus hairy cercocarpus; Wright mountain-mahogany; hairy 

mountain-mahogany 

2

Cercocarpus intricatus little leaf mountain-mahogany 2

Cercocarpus ledifolius curlleaf cercocarpus; mountain-mahogany; curlleaf mountainmahogany

2

Cercocarpus montanus alderleaf cercocarpus; alderleaf mountain-mahogany; 

mountain-mahogany; true mountain-mahogany

2

Cereus giganteus saguaro; giant cactus; pitahaya 2

Chamaecyparis 

lawsoniana

Port-Orford-cedar; Port-Orford white-cedar; Oregon-cedar; 

Lawson cypress

3

Chamaecyparis 

nootkatensis

Alaska-cedar; Nootka cypress; Alaska yellow-cedar; Sitka 

cypress

3
Plant List

Appendix D - Page 

Genus and species Common name Susceptibility 

index

Chamaecyparis 

thyoides

Atlantic white-cedar; Atlantic cedar; white-cedar; southern 

white-cedar

3

Chilopsis linearis desert-willow; desert catalpa 3

Chionanthus virginicus fringtree; fringe tree; old-mans-beard 2

Chrysobalus icaco cocoplum 2

Chrysophyllum 

oliviforme

satinleaf 2

Cinnamonum camphora camphor-tree 1

Citharexylum 

fruticosum

fiddlewood; Florida fiddlewood 2

Citrus aurantifolia lime; key lime 2

Citrus limon lemon 3

Citrus sinensis orange; navel orange; sweet orange 2

Cladrastis lutea yellow-wood 2

Clethra alnifolia sweet pepperbush; summersweet clethra 3

Clethra spp. clethra; pepperbush 3

Cliftonia monophylla buckwheat-tree; titi; black titi 2

Coccoloba diversifolia pigeon-plum; doveplum; tie-tongue 2

Coccoloba uvifera seagrape; grape-tree 2

Coccothrinax argentata Florida silverpalm; Biscayne-palm; brittle thatch; thatchpalm 3

Cocos nucifera coconut; coconut palm 3

Colubrina reclinata soldierwood 2

Conocarpus erectus button-mangrove; buttonwood; silver buttonwood 2

Cordia sebestena geiger-tree 3

Cornus alternifolia alternate-leaf dogwood; blue cornel 3

Cornus drummondii roughleaf dogwood 3

Cornus florida flowering dogwood; dogwood; cornel; boxwood 2

Cornus nuttallii Pacific dogwood; flowering dogwood; mountain dogwood 3

Cornus racemosa gray dogwood 3

Cornus rugosa roundleaf dogwood; roundleafed cornel 3

Cornus spp. dogwood; cornel 3

Cornus stolonifera red-osier dogwood; American dogwood; redstem dogwood; 

kinnikinnik

3

Corylus americana American hazelnut; American filbert; wild hazelnut 1
Appendix D

Appendix D - Page 

Genus and species Common name Susceptibility 

index

Corylus avellana European hazelnut; European filbert 1

Corylus avena 1

Corylus cornuta beaked hazelnut; beaked filbert; western hazelnut 2

Corylus rostrata beaked hazelnut 1

Cotinus obovatus American smoketree; smoketree; chittamwood; yellowwood 1

Cotoneaster pyracantha firethorn; everlasting thorn 1

Cowania mexicana cliffrose; Stansbury cliffrose; quininebush 2

Crataegus berberifolia barberry hawthorn; bigtree hawthorn; barberryleaf hawthorn 1

Crataegus boyntonii Biltmore hawthorn; Boynton hawthorn 1

Crataegus 

brachycantha

blueberry hawthorn; blue haw; pommette blue 1

Crataegus coccinea scarlet hawthorn; scarlet haw 1

Crataegus crus-galli cockspur hawthorn; hog-apple; cockspur-thorn; Newcastle 

thorn

1

Crataegus douglasii black hawthorn; Douglas hawthorn; river hawthorn 1

Crataegus induta downy hawthorn; turkey hawthorn 1

Crataegus intricata Biltmore hawthorne 1

Crataegus marshallii parsley hawthorn; parsley-leaf hawthorn 1

Crataegus mollis downy hawthorn 1

Crataegus monogyna oneseed hawthorn; singleseed hawthorn; English hawthorn; 

European hawthorn

2

Crataegus opaca riverflat hawthorn; English hawthorn; May hawthorn; May 

haw; apple haw

1

Crataegus oxyacantha English hawthorn 1

Crataegus pedicellata scarlet hawthorn 1

Crataegus pruinosa frosted hawthorn; waxy-fruit thorn 1

Crataegus pyracantha firethorn; white thorn 1

Crataegus saligna willow hawthorn 1

Crataegus spathulata littlehip hawthorn; small-fruit hawthorn; pasture hawthorn 1

Crataegus spp. hawthorn 1

Cunninghamia 

lanceolata

China fir; blue Chinese fir 3

Cupressocyparis 

leylandii

Leyland cypress 3

Cupressus arizonica Arizona cypress 3
Plant List

Appendix D - Page 

Genus and species Common name Susceptibility 

index

Cupressus bakeri Baker cypress; Siskiyou cypress; Modoc or MacNab cypress 3

Cupressus goveniana Gowen cypress 3

Cupressus 

guadalupensis

Guadalupe cypress; Forbes’ cypress; Tecate cypress 3

Cupressus macrocarpa Monterey cypress 3

Cupressus sargentii Sargent cypress 3

Cydonia japonica common flowering quince; dwarf Japanese quince; Japan 

quince

2

Cydonia vulgaris quince 2

Cyrilla racemiflora swamp cyrilla; swamp ironwood; leatherwood 2

Dalea spinosa smokethorn; smoketree; indigobush 2

Diospyros texana Texas persimmon; black persimmon; Mexican persimmon 3

Diospyros virginiana persimmon; common persimmon; eastern persimmon; 

possumwood

3

Dipholis salicifolia willow bustic; bustic; willow-leaf bustic; cassada 2

Drypetes lateriflora Guiana-plum 3

Elaeagnus angustifolia Russian-olive; oleaster 3

Elaeagnus hortensis oleaster 2

Elliottia racemosa elliottia; southern plume 2

Enallagma latifolia black-calabash 3

Eriobotrya japonica loquat; loquat tree 2

Erythrina herbacea southeastern coralbean; eastern coralbean; Cherokee-bean 2

Ethretia anacua anaqua 3

Eucalyptus botryiodes bastard mahogany; bangalay 2

Eucalyptus 

camaldulensis

longbeak eucalyptus; camal eucalyptus; redgum 2

Eucalyptus camphora eucalyptus 3

Eucalyptus cinerea silver dollar eucalyptus 1

Eucalyptus diversifolia eucalyptus 3

Eucalyptus globulus bluegum eucalyptus; Tasmanian bluegum; bluegum 2

Eucalyptus gunnii cider gumtree 1

Eucalyptus leucoxylon white ironbark 2

Eucalyptus 

polyanthemos

redbox eucalyptus; redbox-gum; Australian beech; silver 

dollar gum

2
Appendix D

Appendix D - Page 10

Genus and species Common name Susceptibility 

index

Eucalyptus pulchella white peppermint 2

Eucalyptus rudis desert gum 2

Eucalyptus sideroxylon red ironbark 2

Eucalyptus spp. eucalyptus; gum-tree 2

Eucalyptus tereticornis horncap eucalyptus 2

Euonymus 

atropurpureus

eastern burningbush; burningbush; eastern wahoo; 

strawberry-bush

2

Euonymus europaeus European spindletree; European enonymus 2

Euonymus japonicus Japanese euonymus; evergreen euonymus 2

Euonymus occidentalis western burningbush; wahoo; western wahoo 2

Euonymus verrucosa spindle tree 2

Exostema caribaeum princewood; Caribbean princewood 2

Exothea paniculata inkwood; butterbough 2

Fagus grandifolia American beech; beech 2

Fagus sylvatica European beech 2

Fatsia japonica Japanese fatsia; Japanese aralia 3

Ficus aurea Florida strangler fig; golden fig; strangler fig; wild fig 2

Ficus benjamina Java fig; Java willow; Benjamin fig 3

Ficus carica fig; common fig 2

Ficus elastica India-rubber fig; rubber plant; India rubber tree 2

Ficus lyrata fiddle-leaf fig 2

Firmiana platanifolia Chinese parasoltree 2

Forestiera acuminata swamp-privet; forestiera; common adelia; whitewood 3

Fraxinus americana white ash; Biltmore ash; Biltmore white ash 3

Fraxinus anomala singleleaf ash; dwarf ash 3

Fraxinus caroliniana Carolina ash; water ash; Florida ash; pop ash; swamp ash 3

Fraxinus cuspidata fragrant ash; flowering ash 3

Fraxinus excelsior European ash 2

Fraxinus greggii Gregg ash; littleleaf ash; dogleg ash 3

Fraxinus latifolia Oregon ash 3

Fraxinus nigra black ash; swamp ash; basket ash; brown ash; hoop ash; water 

ash

3

Fraxinus pennsylvanica green ash; red ash; Darlington ash; white ash; swamp ash; 

water ash

3
Plant List

Appendix D - Page 11

Genus and species Common name Susceptibility 

index

Fraxinus profunda pumpkin ash; red ash 3

Fraxinus 

quadrangulata

blue ash 3

Fraxinus spp. ash 3

Fraxinus texensis Texas ash 3

Fraxinus velutina velvet ash; Arizona ash; desert ash; Modesto ash; leatherleaf 

ash; smooth ash; Toumey ash

3

Garrya fremontii Fremont silktassel; silk-tassel 3

Gaultheria shallon salal; shallon 2

Ginkgo biloba ginkgo; maidenhair tree 3

Gleditsia aquatica waterlocust 3

Gleditsia texana honeylocust; Texas honeylocust 3

Gleditsia triacanthos honeylocust; sweet-locust; thorny-locust 3

Gordonia lasianthus loblolly-bay; tan bay; gordonia; bay; holly-bay 2

Grevillea ‘noellii’ grevillea 3

Grevillea robusta silk-oak; silky oak 3

Guaiacum sanctum roughbark lignumvitae; holywood lignumvitae; lignumvitae 2

Guettarda elliptica elliptic-leaf velvetseed; Everglades velvetseed; velvetseed 2

Guettarda scabra roughleaf velvetseed 2

Gyminda latifolia falsebox; false boxwood; West Indies falsebox 2

Gymnanthes lucida oysterwood; crabwood 3

Gymnocladus dioicus Kentucky coffeetree; coffeetree 3

Hakae spp. 2

Halesia carolina Carolina silverbell; silver bell; snowdrop-tree; opossum-wood 3

Hamamelis virginiana witch-hazel; common witch-hazel; southern witch-hazel 1

Heteromeles arbutifolia toyon; Christmas berry; California-holly; hollyberry 2

Hibiscus rosa-sinensis Chinese hibiscus 2

Hibiscus tiliaceus sea hibiscus; mahoe; tree hibiscus 2

Hippomane mancinella manchineel 3

Ilex aquifolium English holly 3

Ilex cassine dahoon; dahoon holly; Alabama dahoon; Christmas-berry 3

Ilex coriacea large gallberry; tall inkberry; gallberry; bay-gallbush 3

Ilex decidua possumhaw; deciduous holly; winterberry 3

Ilex glabra inkberry; gallberry 3
Appendix D

Appendix D - Page 12

Genus and species Common name Susceptibility 

index

Ilex krugiana tawnyberry holly; Krug holly; southern holly 3

Ilex montana mountain winterberry; mountain holly 3

Ilex opaca American holly; holly; white holly 3

Ilex verticillata common winterberry; black-alder; winterberry 3

Jasminum nudiflorum winter jasmine 3

Juglans californica southern California walnut; California walnut; California 

black walnut

2

Juglans cinerea butternut; white walnut; oilnut 2

Juglans hindsii northern California walnut; Hinds walnut; California black 

walnut

2

Juglans major Arizona walnut; Arizona black walnut 2

Juglans microcarpa little walnut; Texas walnut; Texas black walnut; river walnut 2

Juglans nigra black walnut; eastern black walnut; American walnut 2

Juniperus ashei Ashe juniper; mountain-cedar; rock-cedar; post-cedar; 

Mexican juniper

3

Juniperus californica California juniper 3

Juniperus coahuilensis redberry juniper; roseberry 3

Juniperus communis common juniper; dwarf juniper; prostrate juniper 3

Juniperus deppeana alligator juniper; checker-bark juniper; western juniper 3

Juniperus erythrocarpa redberry juniper; red-fruited juniper 3

Juniperus flaccida drooping juniper; weeping juniper; Mexican drooping juniper 3

Juniperus monosperma oneseed juniper; cherrystone juniper; West Texas juniper 3

Juniperus occidentalis western juniper, Sierra juniper 3

Juniperus osteosperma Utah juniper; bigberry juniper 3

Juniperus pinchotii Pinchot juniper; redberry juniper 3

Juniperus scopulorum Rocky Mountain juniper; Rocky Mountain cedar; redcedar; 

Colorado redcedar

3

Juniperus silicicola southern redcedar; redcedar; sand-cedar; coast juniper 3

Juniperus virginiana eastern redcedar; redcedar; red juniper; savin 3

Krugiodendron ferreum leadwood; black-ironwood 2

Laguncularia racemosa white-mangrove; white buttonwood; buttonwood 2

Larix decidua European larch 1

Larix laricina tamarack; eastern larch; American larch; Alaska larch; 

hackmatack

1
Plant List

Appendix D - Page 13

Genus and species Common name Susceptibility 

index

Larix lyallii subalpine larch; alpine larch; timberline larch; tamarack 1

Larix occidentalis western larch; hackmatack; Montana larch; mountain larch 1

Leitneria floridana corkwood 2

Lindera benzoin spicebush 3

Liriodendron tulipifera yellow-poplar; tuliptree; tulip-poplar; white-poplar 3

Lithocarpus densiflorus tanoak; tan oak; tanbark-oak 1

Lyonia ferruginea tree lyonia; staggerbush; titi; rusty lyonia 2

Lyonothamnus 

floribundus

Lyontree; Catalina-ironwood; lyonothamnus; Santa-Cruzironwood2

Lysiloma bahamensis Bahama lysiloma 2

Maclura pomifera Osage-orange; bodark; bodock; bowwood; hedge-apple; 

horse-apple

3

Magnolia acuminata cucumbertree; cucumber magnolia; mountain magnolia 3

Magnolia ashei Ashe magnolia; sandhill magnolia 3

Magnolia fraseri Fraser magnolia; mountain magnolia; earleaf cucumbertree 3

Magnolia grandiflora southern magnolia; evergreen magnolia; bull-bay; big-laurel 3

Magnolia macrophylla bigleaf magnolia; umbrella-tree; large-leaf cucumbertree 3

Magnolia pyramidata pyramid magnolia; southern cucumbertree; mountain 

magnolia

3

Magnolia soulangeana saucer magnolia; rustica rubra 3

Magnolia tripetala umbrella magnolia; umbrella-tree; elkwood 3

Magnolia virginiana sweetbay; swampbay; southern sweetbay; laurel magnolia 3

Malus angustifolia southern crab apple; narrowleaf crab apple; wild crab apple 1

Malus coronaria sweet crab apple; American crab apple; wild crab 1

Malus diversifolia Oregon crab apple; Pacific crab apple; western crab apple; 

wild crab apple

1

Malus glabrata sweet crab apple; Biltmore crab apple; wild crab 1

Malus ioensis prairie crab apple; wild crab apple; Iowa crab 1

Malus spp. apple 1

Melaleuca decussata lilac melaleuca 1

Melaleuca 

quinquenervia

cajeput-tree; punktree; bottlebrush 2

Melia azedarach chinaberry; umbrella chinaberry; chinatree; pride-of-India 2

Mespilus germanica medlar; showy mespilus; European medlar 2
Appendix D

Appendix D - Page 14

Genus and species Common name Susceptibility 

index

Metasequoia 

glyptostroboides

dawn redwood 2

Metopium toxiferum Florida poisontree; poisonwood; West Indies poisontree 1

Morus alba white mulberry; silkworm mulberry; weeping mulberry 3

Morus alba var. 

tatarica

Russian mulberry 3

Morus nigra black mulberry 3

Morus rubra red mulberry; moral 3

Morus tartarica Tartarian mulberry 2

Mustichodendro 

foetidissimum

false mastic 2

Myrica californica Pacific bayberry; California bayberry; Pacific waxmyrtle; 

western waxmyrtle; California waxmyrtle

2

Myrica cerifera southern bayberry; southern waxmyrtle; bayberry; 

candleberry

2

Nyssa aquatica water tupelo; tupelo-gum; cotton-gum; sourgum 3

Nyssa ogeche Ogeechee tupelo; sour tupelo-gum; Ogeechee-lime; sour 

tupelo

3

Nyssa sylvatica black tupelo; blackgum; sourgum; pepperidge; tupelo 3

Nyssa sylvatica var.

biflora

swamp tupelo; blackgum; swamp blackgum 3

Olea europaea olive; common olive 3

Olneya tesota tesota; desert ironwood; Arizona-ironwood 2

Osmanthus americana devilwood; wild-olive 3

Ostrya knowltonii Knowlton hophornbeam; western hophornbeam; wolf 

hophornbeam

2

Ostrya virginiana eastern hophornbeam; hophornbeam; American 

hophornbeam; hornbeam; leverwood

1

Oxydendrum arboreum sourwood; sorrel-tree; lily-of-the-valley-tree 2

Parkinsonia aculeate Jerusalem-thorn; horsebean; Mexican paloverde 2

Paulownia tomentosa royal paulownia; empress-tree; princess-tree; paulownia 3

Paurotis wrightii paurotis-palm; paurotis 3

Persea americana avocado; zutano avocado; alligator-pear 2

Persea borbonia redbay; shorebay 2

Photinia arbutifolia toyon; Christmas berry 2
Plant List

Appendix D - Page 15

Genus and species Common name Susceptibility 

index

Photinia glabra Japanese photinia 2

Photinia serrulata Chinese photinia: Chinese medlar 2

Photinia spp. toyon; photinia 3

Picea abies Norway spruce 2

Picea breweriana Brewer spruce; weeping spruce 2

Picea engelmannii Engelmann spruce; Columbian spruce; mountain spruce; 

silver spruce; white spruce

2

Picea glauca white spruce; skunk spruce; Canadian spruce; cat spruce 2

Picea mariana black spruce; bog spruce; swamp spruce; shortleaf black 

spruce

2

Picea polita tigertail spruce 2

Picea pungens blue spruce; Colorado blue spruce; Colorado spruce; silver 

spruce

2

Picea rubens red spruce; yellow spruce; West Virginia spruce; eastern 

spruce

2

Picea sitchensis Sitka spruce; coast spruce; tideland spruce; yellow spruce 2

Picea spp. spruce 2

Picramnia pentandra bitterbush; Florida bitterbush 2

Pinckneya pubens pinckneya; fevertree; Georgia-bark; fever-bark 2

Pinus albicaulis whitebark pine; scrub pine; white pine 2

Pinus aristata bristlecone pine; hickory pine; foxtail pine 2

Pinus attenuata knobcone pine 2

Pinus balfouriana foxtail pine 2

Pinus banksiana jack pine; scrub pine; gray pine; black pine; Banksian pine 2

Pinus cembroides Mexican pinyon; nut pine; Mexican stone pine 2

Pinus clausa sand pine; scrub pine; spruce pine 2

Pinus contorta lodgepole pine; shore pine; beach pine 2

Pinus coulteri Coulter pine; bigcone pine; pitch pine 2

Pinus discolor border pinyon 2

Pinus echinata shortleaf pine; shortleaf yellow pine; yellow pine 2

Pinus edulis pinyon; two-leaf pinyon; two-needle pinyon 2

Pinus elliottii slash pine; yellow slash pine; swamp pine; pitch pine 2

Pinus engelmannii Apache pine; Arizona longleaf pine 2

Pinus flexilis limber pine; white pine; Rocky Mountain white pine 2
Appendix D

Appendix D - Page 16

Genus and species Common name Susceptibility 

index

Pinus glabra spruce pine; cedar pine; Walter pine; bottom white pine 2

Pinus halepensis Aleppo pine 2

Pinus jeffreyi Jeffrey pine; western yellow pine; bull pine; black pine; 

ponderosa pine

2

Pinus lambertiana sugar pine; California sugar pine 2

Pinus leiophylla var.

chihuahuana

Chihuahua pine; yellow pine 2

Pinus longaeva intermountain bristlecone pine 2

Pinus monticola western white pine; mountain white pine; Idaho white pine; 

silver pine

2

Pinus mugo mugo pine; mountain pine; Swiss mountain pine 2

Pinus muricata bishop pine; prickle-cone pine; Santa Cruz Island pine 2

Pinus nigra Austrian pine; European black pine 2

Pinus palustris longleaf pine; swamp pine; longleaf yellow pine; southern 

yellow pine

3

Pinus pinea Italian stone pine 3

Pinus ponderosa ponderosa pine; western yellow pine; yellow pine 2

Pinus ponderosa var.

arizonica

Arizona pine; Arizona ponderosa pine; yellow pine 2

Pinus pungens Table Mountain pine; mountain pine; hickory pine 2

Pinus quadrifolia Parry pinyon; four-needle pinyon; nut pine 2

Pinus radiata Monterey pine; insignis pine 2

Pinus resinosa red pine; Norway pine 2

Pinus rigida pitch pine 3

Pinus sabiniana Digger pine; bull pine; gray pine 2

Pinus serotina pond pine; marsh pine; pocosin pine 2

Pinus spp. Pine 2

Pinus strobiformis southwestern white pine; Mexican white pine; border white 

pine

2

Pinus strobus eastern white pine; northern white pine; white pine 2

Pinus sylvestris Scotch pine; Scots pine 2

Pinus taeda loblolly pine; oldfield pine; shortleaf pine 2

Pinus thunbergiana Japanese black pine 3
Plant List

Appendix D - Page 17

Genus and species Common name Susceptibility 

index

Pinus torreyana Torrey pine; Del Mar pine; Soledad pine 2

Pinus virginiana Virginia pine; Virginia scrub pine; spruce pine; Jersey pine; 

scrub pine; poverty pine

2

Pinus washoensis Washoe pine 2

Piscidia piscipula Florida fishpoison-tree; Jamaica-dogwood; Florida 

fishfuddletree

2

Pistacia texana Texas pistache; American pistachio; wild pistachio 1

Pistacia vera pistachio 1

Planera aquatica water-elm; planertree 2

Platanus orientalis Oriental planetree 2

Platanus racemosa California sycamore; western sycamore; California planetree 3

Platanus wrightii Arizona sycamore; Arizona planetree 3

Populus alba white poplar; silver poplar 2

Populus angustifolia narrowleaf cottonwood; black cottonwood; mountain 

cottonwood; narrowleaf poplar

1

Populus balsamifera balsam poplar; balm; balm-of-Gilead; bam; tacamahac 1

Populus deltoides eastern cottonwood; eastern poplar; southern cottonwood 2

Populus fremontii Fremont cottonwood; cottonwood 2

Populus grandidentata bigtooth aspen; largetoothed aspen; aspen; poplar; popple 1

Populus heterophylla swamp cottonwood; black cottonwood; river cottonwood 1

Populus nigra var.

italica

Lombardy poplar 1

Populus palmeri eastern cottonwood; eastern poplar; Palmer cottonwood 1

Populus sargentii plains cottonwood; great plains cottonwood; sargent 

cottonwood 

1

Populus spp. cottonwood; poplar 1

Populus tremuloides quaking aspen; trembling aspen; golden aspen 1

Populus trichocarpa black cottonwood; western balsam poplar; cottonwood; 

balsam cottonwood

1

Populus wislizenii Rio Grande cottonwood; valley cottonwood 1

Prosopis juliflora honeylocust; mesquite; algaroba 2

Prosopis pubescens screwbean mesquite; screwbean 2

Prunus alleghaniensis Allegheny plum; sloe plum; sloe; Allegheny sloe; northern 

sloe

2
Appendix D

Appendix D - Page 18

Genus and species Common name Susceptibility 

index

Prunus americana American plum; wild plum; red plum; river plum; yellow 

plum

2

Prunus angustifolia Chickasaw plum; sand plum 2

Prunus avium mazzard; common sweet cherry; English cherry 2

Prunus caroliniana Carolina laurelcherry; laurel cherry; cherry-laurel 2

Prunus domestica garden plum; plum; Damson plum 2

Prunus emarginata bitter cherry; quinine cherry; wild cherry 2

Prunus fremontii desert apricot 2

Prunus glandulosa flowering almond; dwarf flowing almond; almond cherry; 

wild peach

2

Prunus hortulana Hortulan plum 2

Prunus japonica Japanese plum 2

Prunus laurocerasus cherry laurel; English laurel 2

Prunus lyonii Catalina cherry 2

Prunus maritima beach plum 2

Prunus mexicana Mexican plum; bigtree plum; inch plum 2

Prunus munsoniana wildgoose plum; Munson plum 2

Prunus myrtifolia West Indies cherry; myrtle laurel cherry; laurelcherry 2

Prunus nigra Canada plum; red plum; horse plum; wild plum 2

Prunus padus European bird-cherry; black serviceberry 2

Prunus pensylvanica pin cherry; wild red cherry; fire cherry; northern pin cherry; 

pigeon cherry; bird cherry

3

Prunus persica peach; nectarine; heavenly white nectarine; Tilton apricot 2

Prunus pissardi purple-leaved prune 2

Prunus pumila sand cherry 2

Prunus serotina black cherry; wild black cherry; rum cherry; mountain black 

cherry

2

Prunus spinosa sloe; blackthorn 2

Prunus spp. cherry; plum 2

Prunus subcordata Klamath plum; Sierra plum; Pacific plum; western plum; wild 

plum

2

Prunus umbellata flatwoods plum; black sloe; hog plum; sloe 2

Prunus virginiana chokecherry; common chokecherry; black chokecherry; 

California chokecherry

2
Plant List

Appendix D - Page 19

Genus and species Common name Susceptibility 

index

Pseudophoenix 

sargentii

buccaneer-palm; Florida cherrypalm; Sargent cherrypalm 3

Pseudotsuga 

macrocarpa

bigcone Douglas-fir; bigcone-spruce; hemlock 2

Pseudotsuga menziesii Douglas-fir; red-fir; Oregon-pine; Douglas-spruce 2

Psidium guajava guava; common guava; guayaba 2

Ptelea trifoliata hoptree; common hoptree; wafer-ash 2

Punica granatum pomegranate 2

Pyracantha coccinea scarlet firethorn; everlasting thorn; fire thorn 2

Pyrus angustifolia narrowleaf crab apple 1

Pyrus arbutifolia red chokecherry; red chokeberry; chokeberry 2

Pyrus communis pear 2

Pyrus fusca Oregon crab apple 1

Pyrus malus wild apple; common apple 1

Quercus agrifolia coast live oak; California live oak 1

Quercus alba white oak; stave oak 1

Quercus arizonica Arizona white oak; Arizona oak 1

Quercus austrina Durand oak; Durand white oak; bluff oak 1

Quercus bicolor swamp white oak 1

Quercus chapmanii Chapman oak; Chapman white oak; scrub oak 1

Quercus chrysolepis canyon live oak; California live oak; canyon oak; goldcup 

oak; live oak; maul oak

1

Quercus cinerea bluejack oak 1

Quercus coccinea scarlet oak; black oak; Spanish oak 1

Quercus douglasii blue oak; California blue oak; iron oak; mountain white oak; 

mountain oak

1

Quercus durandii Durand oak; Durand white oak; bluff oak; white oak 1

Quercus ellipsoidalis northern pin oak; jack oak; black oak; Hill oak 1

Quercus emoryi Emory oak; black oak; blackjack oak 1

Quercus engelmannii Engelmann oak; evergreen white oak; mesa oak; Engelmann 

spruce

1

Quercus falcata southern red oak; Spanish oak; water oak; red oak 1

Quercus gambelii Gambel oak; Rocky Mountain white oak; Utah white oak; 

white oak

1
Appendix D

Appendix D - Page 20

Genus and species Common name Susceptibility 

index

Quercus garryana Oregon white oak; Oregon oak; Garry oak; post oak; white 

oak; Brewer oak; shin oak

1

Quercus grisea gray oak; Arizona gray oak 1

Quercus hemisphaerica laurel oak; Darlington oak 1

Quercus hypoleucoides silverleaf oak; white-leaf oak 1

Quercus ilicifolia bear oak; scrub oak 1

Quercus imbricaria shingle oak; laurel oak 1

Quercus incana bluejack oak; cinnamon oak; sandjack; bluejack; shin oak; 

turkey oak

1

Quercus kelloggii California black oak; black oak; Kellogg oak 1

Quercus laevis turkey oak; Catesby oak; scrub oak 1

Quercus laurifolia laurel oak; Darlington oak; diamond-leaf oak; swamp laurel 

oak

1

Quercus lobata valley oak; California white oak; valley white oak; water oak 1

Quercus lyrata overcup oak 1

Quercus macrocarpa bur oak; mossy cup oak; blue oak; mossy-overcup oak; scrub 

oak

1

Quercus margaretta sand post oak; small post oak; dwarf post oak; post oak 1

Quercus marilandica blackjack oak; blackjack; barren oak; black oak; jack oak 1

Quercus michauxii swamp chestnut oak; basket oak; cow oak 1

Quercus muehlenbergii chinkapin oak; yellow chestnut oak; chestnut oak; rock 

chestnut oak

1

Quercus myrtifolia myrtle oak; scrub oak 1

Quercus nigra water oak; possum oak; spotted oak 1

Quercus nuttallii Nuttall oak; red oak; Red River oak; pin oak 1

Quercus oblongifolia Mexican blue oak 1

Quercus oglethorpensis Oglethorpe oak 1

Quercus pagoda cherrybark oak; swam red oak; bottomland red oak 1

Quercus palustris pin oak; swamp oak; water oak; swamp Spanish oak; Spanish 

oak

1

Quercus phellos willow oak; pin oak; peach oak; swamp willow oak 1

Quercus prinus chestnut oak; basket oak; rock chestnut oak; rock oak; tanbark 

oak

1

Quercus rubra northern red oak; red oak; common red oak; gray oak; eastern 

red oak; mountain red oak

1
Plant List

Appendix D - Page 21

Genus and species Common name Susceptibility 

index

Quercus shumardii Shumard oak; Shumard red oak; spotted oak; Schneck oak; 

Schneck red oak; southern red oak

1

Quercus spp. oak 1

Quercus stellata post oak; iron oak 1

Quercus suber cork oak 1

Quercus undulata Rocky Mountain shin oak; wavyleaf oak 1

Quercus velutina black oak; yellow oak; quercitron oak; yellow-bark oak; 

smooth-bark oak

1

Quercus virginiana live oak; Virginia live oak 1

Quercus wislizenii interior live oak; highland live oak; Sierra live oak 1

Rapanea guianensis Guiana rapanea 2

Reynosia 

septentrionalis

darling-plum; red-ironwood 2

Rhamnus caroliniana Carolina buckthorn; Indian-cherry; yellow buckthorn; tree 

buckthorn; yellowwood

3

Rhamnus cathartica European buckthorn; common buckthorn; European waythorn 3

Rhamnus frangula glossy buckthorn; alder buckthorn 3

Rhamnus purshiana cascara buckthorn; cascara; cascara sagrada; bearberry; 

chittam; coffeetree

2

Rhizophora mangle mangrove; red mangrove 2

Rhus copallina shining sumac; dwarf sumac; winged sumac; wing-rib sumac; 

flameleaf sumac

2

Rhus corallina mountain sumac 1

Rhus cotinus smoketree; common smoketree 2

Rhus glabra smooth sumac; scarlet sumac; common sumac; Rocky 

Mountain sumac; red sumac

1

Rhus integrifolia lemonade sumac; sourberry; lemonade-berry; mahogany 

sumac

2

Rhus typhina staghorn sumac; velvet sumac 1

Ribes uva-crispa English gooseberry 2

Robinia neomexicana New Mexico locust; New Mexican locust; southwestern 

locust

3

Robinia pseudoacacia black locust; common locust; yellow locust; white locust 3

Robinia spp. locust 2
Appendix D

Appendix D - Page 22

Genus and species Common name Susceptibility 

index

Robinia viscosa clammy locust 3

Rosa bracteata Macartney rose 2

Rosa eglanteria sweetbriar; sweetbriar rose 2

Rosa setigera prairie rose; climbing prairie rose 2

Rosa spp. rose 1

Roystonea elata Florida royalpalm; Cuban royalpalm; royalpalm 3

Sabal palmetto cabbage palmetto; common palmetto; Carolina palmetto; 

palmetto; cabbage-palm

3

Salix alaxensis feltleaf willow 1

Salix alba white willow; European white willow 1

Salix alba var. tristis golden weeping willow 1

Salix amygdaloides peachleaf willow; peachleaved willow; almond willow; peach 

willow; southwestern peach willow

1

Salix babylonica weeping willow; Babylon weeping willow; Napolean willow 2

Salix bonplandiana Bonpland willow; Toumey willow; red willow; polished 

willow

1

Salix caroliniana Coastal Plain willow; Ward willow; southern willow; 

Harbison willow

1

Salix cordata heartleaf willow; heart-leaved willow 1

Salix discolor pussy willow; glaucous willow; silvery pussy willow 1

Salix eriocephala pussy willow 1

Salix fragilis crack willow; brittle willow; snap willow 1

Salix hookerana Hooker willow; coast willow; Yakutat willow; bigleaf willow 1

Salix interior sandbar willow; coyote willow; acequia willow; basket 

willow; gray willow; sandbar willow

1

Salix laevigata Bondpland willow; red willow; Toumey willow; polished 

willow

1

Salix lasiandra Pacific willow; whiplash willow; black willow; red willow; 

western black willow; yellow willow

1

Salix lasiolepis arroyo willow; white willow 1

Salix lucida shining willow; shiny willow 1

Salix mackenzieana Mackenzie willow 1

Salix nigra black willow; swamp willow; Goodding willow; western 

black willow; Dudley willow

1

Salix pentandra laurel willow; bay willow; bayleaf willow 2
Plant List

Appendix D - Page 23

Genus and species Common name Susceptibility 

index

Salix scouleriana Scouler willow; fire willow; black willow; mountain willow; 

Nuttall willow

1

Salix spp. willow 1

Salix taxifolia yewleaf willow; yew willow 1

Salix viminalis basket willow; osier; common osier; silky osier 1

Sambucus callicarpa Pacific red elder; Pacific elder; coast red elder; redberry elder; 

red elderberry

2

Sambucus canadensis American elder; common elderberry; common elder; 

blackberry elder

3

Sapindus drummondii western soapberry; wild chinatree; cherioni 2

Sapindus marginatus wingleaf soapberry; Florida soapberry 2

Sapindus saponaria wingleaf soapberry; Florida soapberry; southern soapberry; 

Mexican soapberry; wild chinatree

2

Sapium sebiferum tallowtree; Chinese tallowtree 3

Sassafras albidum sassafras; white sassagras 2

Schinus molle California peppertree 1

Sequoia sempervirens redwood; coast redwood; California redwood 2

Sequoiadendron 

giganteum

giant sequoia; sequoia; bigtree; Sierra redwood 2

Sideroxylon 

foetidissimum

false-mastic; mastic; wild-mastic; wild-olive 2

Simarouba glauca paradise-tree; bitterwood 2

Sophora affinis Texas sophora; coralbean; pink sophora; Eves-necklace 3

Sophora japonica Japanese pagoda-tree 3

Sophora secundiflora mescalbean; frigolito; coralbean; Texas-mountain-laurel 2

Sorbus americana American mountain-ash; mountain-ash; roundwood 1

Sorbus aucuparia European mountain-ash; Rowan-tree 1

Spiraea bumalda Bumalda spirea; spirea 3

Stewartia koreana Korean stewartia; stewartia 3

Stewartia ovata mountain stewartia; mountain-camellia; angel-fruit stewartia 2

Swietenia mahagoni West Indies mahogany; mahogany 2

Symphoricarpos albus snowberry; waxberry; common snowberry 3

Symplocos tinctoria sweetleaf; horse-sugar; common sweetleaf; yellowwood 2

Tamarix parviflora small-flower tamarisk 2

Taxodium distichum baldcypress 3
Appendix D

Appendix D - Page 24

Genus and species Common name Susceptibility 

index

Taxodium mucronatum Montezuma baldcypress; Mexican cypress 3

Taxus brevifolia Pacific yew; western yew 3

Taxus floridana Florida yew 3

Thrinax microcarpa key thatchpalm; silvertop palmetto; prickly thatch; brittle 

thatch; brittle thatch palm

3

Thrinax parviflora Jamaica thatchpalm 3

Thuja occidentalis northern white-cedar; white-cedar; eastern arborvitae; 

American arborvitae; eastern white-cedar

3

Thuja orientalis oriental arborvitae; Chinese arborvitae 3

Thuja plicata western redcedar; giant western arborvitae; Pacific redcedar; 

giant-cedar; arborvitae; canoe-cedar

3

Tilia americana American basswood; American linden; basswood 1

Tilia caroliniana Carolina basswood; Florida basswood; basswood; Carolina 

linden; Florida linden

1

Tilia cordata littleleaf linden; small-leaved linden; small-leaved European 

linden

1

Tilia europaea European linden 1

Tilia floridana Florida basswood; Carolina basswood 1

Tilia heterophylla white basswood; beetree; linden; beetree linden 1

Torreya californica California torreya; California-nutmeg 3

Torreya taxifolia Florida torreya; stinking-cedar 3

Torrubia longifolia longleaf blolly; Brace blolly roundleaf blolly; beeftree; 

beefwood

2

Toxicodendron vernix poison-sumac; poison-dogwood; poison-elder; thunderwood 1

Trema micrantha Florida trema 2

Tsuga canadensis eastern hemlock; Canadian hemlock; Canada hemlock; 

hemlock spruce; common hemlock

2

Tsuga caroliniana Carolina hemlock 2

Tsuga heterophylla western hemlock; Pacific hemlock; west coast hemlock 2

Tsuga mertensiana mountain hemlock; black hemlock; alpine hemlock; hemlock 

spruce

2

Ulmus alata winged elm; wahoo elm; cork elm; wahoo 2

Ulmus americana American elm; white elm; water elm; soft elm; Florida elm 2

Ulmus campestris English elm; European elm 2

Ulmus crassifolia cedar elm; basket elm; red elm; southern rock elm 2
Plant List

Appendix D - Page 25

Genus and species Common name Susceptibility 

index

Ulmus glabra Scotch elm; wych elm 2

Ulmus montana Scotch elm 2

Ulmus parvifolia Chinese elm; lacebark 2

Ulmus pumila Siberian elm; Asiatic elm; dwarf Asiatic elm; Pekin elm 2

Ulmus racemosa rock elm; cork elm 2

Ulmus rubra slippery elm; red elm; gray elm; soft elm 3

Ulmus serotina September elm; red elm 2

Ulmus spp. elm 2

Ulmus thomasii rock elm; cork elm 2

Umbellularia 

californica

California-laurel; California-bay; Oregon-myrtle; Pacificmyrtle; pepperwood; spice-tree3

Vauquelinia californica Torrey vauquelinia; Arizona-rosewood 2

Veitchia merrillii Manila palm 3

Viburnum acerifolium mapleleaf viburnum; dockmackie; maple-leaved arrowwood 3

Viburnum ellipticum western blackhaw; oval-leafed virburnum 2

Viburnum lantana wayfaringtree 2

Viburnum opulus European cranberrybush; highbush cranberry; cranberry tree 3

Viburnum prunifolium blackhaw; stagbush; sweethaw 2

Viburnum pubescens downy viburnum; hairy nannyberry; downy arrowwood 2

Viburnum 

rhytidophyllum

leatherleaf viburnum 2

Viburnum spp. viburnum; wayfaringtree 3

Viburnum tomentosum doublefile viburnum 3

Washingtonia filifera California washingtonia; California-palm; fanpalm; 

California fanpalm; desert-palm

3

Ximenia americana tallowwood; hogplum 3

Zanthoxylum 

americanum

common prickly-ash; toothache-tree; northern prickly-ash; 

prickly ash

2

Zanthoxylum clavaherculis

Hercules-club; pepperbark; southern prickly-ash; toothachetree; tingle-tongue2

Zanthoxylum fagara lime prickly-ash; wild-lime-tree; wild-lime 2

Zanthoxylum flavum West Indies satinwood; yellowheart; satinwood; yellowwood 2

Appendix E

Biology, History, 

and Control Efforts 

for the Gypsy Moth

Figure E-1. Small hand sprayers were used to apply DDT in 1945 (Gill, MA). 
Figures and Tables

Figure E-1. Small hand sprayers were used to apply DDT in 

1945 (Gill, MA) .....................................................................................Cover

Figure E-2. A historic county quarantine map shows the spread 

of the gypsy moth from 1909 to 2007........................................................... 1

Figure E-3. Female gypsy moths add hairs from their abdomens 

to their egg masses........................................................................................ 2 

Figure E-4. The gypsy moth caterpillar (larva) develops pairs 

of distinctive red and blue spots as it grows................................................. 2

Figure E-5. The gypsy moth pupa lasts for about 2 weeks........................... 2

Figure E-6. The gypsy moth adult male (left) and female (right) 

are visibly different....................................................................................... 2

Table E-1. Differences between the European and Asian strains 

of the gypsy moth, by life stage and cause of mortality ............................... 4

Appendix E Biology, History, and 

Control Efforts for the Gypsy Moth

Contents

E.1 About the Gypsy Moth............................................................................1

Life Cycle..........................................................................................1

European and Asian Strains..............................................................3

Population Phases..............................................................................5

 Innocuous Phase...................................................................5

 Release Phase.......................................................................5

 Outbreak Phase....................................................................5

 Decline Phase.......................................................................5

Host Plants........................................................................................5

E.2 1869 to 1910: Biological Controls Fail..................................................5

E.3 1911 to 1939: Chemical Insecticides Gain Favor...................................6

E.4 1940 to 1957: DDT Gets Widespread Use..............................................7

E.5 1958 to the Mid-1980s: Safer Treatments Needed.................................8

E.6 Mid-1980s to the Present: Adoption of Integrated Pest Management....9

E.7 1991 to the Present: Asian Strain Creates Additional Concern............10
Biology, History, and Control Efforts for the Gypsy Moth

Appendix E - Page 

This appendix describes the progression of control 

efforts that paralleled the spread of the gypsy moth 

from 1869 to 2005 in the United States (Figure E-2). 

Biological information includes coverage of its life 

cycle, differences between the European and Asian 

strains, and the four population phases and host plants.

E.1 About the Gypsy Moth.

The following information is provided to facilitate 

better understanding of the insect, the problems it 

creates, and treatments. 

Life Cycle.

Producing one generation per year, the gypsy moth 

goes through four life stages: egg, larva, pupa, and 

adult moth (Figures E-3, E-4, E-5, E-6).

Figure E-2. A historic county quaratine map shows the spread of the gypsy moth from 1909 to 2007.

After mating, the female gypsy moth deposits eggs in 

a well-defined mass, containing from a few hundred to 

a thousand eggs, typically in a protected area such as 

bark crevices, on the underside of branches, and in leaf 

litter. She coats the eggs with hairs from her abdomen, 

giving the egg mass a furry appearance and buff color. 

Though the embryos within the eggs develop into 

caterpillars in 4 to 6 weeks, the caterpillars remain in 

the eggs during winter. Survival and hatching success 

depend on a combination of time and temperature 

requirements. A prolonged period of chilling and 

sufficient time for subsequent incubation are necessary 

for egg hatch the following spring (Giese and 

Casagrande 1981). 

1994

1989

1999

1974

1969

1949

1914

1909 1938

2007
Appendix E

Appendix E - Page 

Figure E-3. Female gypsy moths add hairs from their 

abdomens to their egg masses.

Figure E-4. The gypsy moth catepillar (larva) develops 

pairs of distintive red and blue spots as it grows.

Figure E-5. The gypsy moth pupa lasts for about 2 weeks.

Figure E-6. The gypsy moth adult male (left) and female 

(right) are visibly different.

Coinciding with the appearance of spring leaves, eggs 

laid the previous year hatch (during April and May 

in the Middle Atlantic States), and caterpillars climb 

multiple varieties of trees, bushes, and other objects, 

spinning a thread of silk from which they hang freely. 

A phenomenon termed “ballooning,” by which the 

wind carries them to new locations, relocates most 

caterpillars before they begin feeding. Caterpillars 

may balloon several times before they settle and begin 

feeding on foliage (Nichols 1980). 

The small caterpillars move into the tree canopy 

where they feed on leaves for the next 6 to 8 weeks. 

Caterpillars grow from one-tenth of an inch (3 mm) 

to as large as 31⁄2 inches (90 mm) by going through a 
Biology, History, and Control Efforts for the Gypsy Moth

Appendix E - Page 

series of growth stages called instars. A molt (shedding 

of the outer layer) precedes each instar stage; the 

discarded “skins” can cause respiratory problems for 

some people. Male caterpillars grow through five 

instars, females through six; an additional instar is not 

uncommon (Doane and McManus 1981). 

The caterpillars develop distinctive markings on their 

ash-colored bodies as they grow--a yellow stripe 

down the back, with rows of five blue spots followed 

by six red spots on both sides of the yellow stripe. 

Their excrement, called frass, can create health risks; 

large populations of caterpillars excrete so much frass 

it sounds like rain falling through the leaves. Frass 

in runoff water can also pollute lakes and streams, 

threatening fish (Sharpe 1982). 

Caterpillars typically feed at night to avoid predators, 

though feeding may occur at any time of the day when 

caterpillar populations explode and competition for 

food increases. The feeding caterpillar is the life stage 

targeted in most gypsy moth treatment projects because 

of the potential for defoliation.

When population levels are low, caterpillars move 

down the tree during the day and rest in protected areas 

under tree bark and in crevices, returning to the tree 

canopy to feed at night. When populations are elevated 

and competition for foliage high, caterpillars remain in 

the tree canopy and feed night and day. After stripping 

the foliage of the host tree, the caterpillars descend, 

crawling in search of new food sources (McManus and 

others 1989). 

Following the last instar (June and July in the Middle 

Atlantic States), caterpillars find any available 

protected spot in trees, on buildings, and even on the 

ground, entering their pupal stage over the next 2 days. 

Approximately 2 weeks later, adult moths emerge. 

Male gypsy moths appear first, followed several days 

later by the females. The egg-laden females emit a 

pheromone, attracting males for mating. The female 

moths then deposit their egg masses, beginning the 

cycle anew the following spring.

European and Asian Strains.

The European strain of the gypsy moth became 

established in North America from a single introduction 

of closely related individuals, and genetic studies have 

shown little variation within or between populations 

(Wallner 1992). In North America, the European strain 

is also called the North American strain.

The common reference to “the Asian strain” of the 

gypsy moth is actually several strains, which display 

considerable variability. The most notable variances 

are the female’s flying abilities (some females of 

the Asian strain are strong fliers, capable of flights 

exceeding 18 miles [28.9 km]) and the capacity to 

establish in a broad range of hosts (Wallner 1992). 

The European and Asian strains of the gypsy moth are 

similar in appearance; however, behavioral differences 

between them are significant, particularly the inability 

of the European strain female to fly (Wallner 1992). 

Females of the Asian strain are attracted to light and 

more likely to deposit their eggs near light sources, 

thus potentially increasing the social “nuisance” factor 

usually associated with the gypsy moth (Hofacker 

1994). 

The Asian strain feeds on some hosts that are only 

marginally acceptable to the European strain, 

increasing their potential to establish themselves 

and cause even more extensive defoliation than their 

European cousins (USDA APHIS 1992).

Other differences between the European and Asian 

strains are minor (Table E-1). The most reliable 

method for distinguishing between the strains, other 

than the flight of the female, is genetic testing .

Prior to the first known introductions of the Asian 

strain in 1991, eradication actions were singularly 

focused against the European strain. Efforts against 
Appendix E

Appendix E - Page 

Table E-1. Differences between the European and Asian 

strains of the gypsy moth, by life stage and cause of mortality 

(adapted from Wallner 1992, p. 2).

Life Stage European Strain

(North America)

Asian Strain

(Siberia, Russia, 

Far East)

Caterpillars First instars 

disperse

Color uniform

Main hosts: oak, 

birch, poplar, 

willow, alder

Early instars feed 

in the canopy at 

night and move 

to resting sites 

during the day.

First and second 

instars disperse

Color highly 

variable

Main hosts: oak, 

larch, birch, 

willow

Early instars feed 

in the canopy at 

night and remain 

on the host during 

the day.

Pupae Pupates in 

protected spots in 

bark crevices, in 

leaf litter

Pupates on 

foliage

Adult 

Females

Flightless Strong flier, 

attracted to light

Egg 

Masses

On tree trunks, 

rocks, leaf litter

On foliage, tree 

trunks, rocks, 

objects near lights

Cause of 

Mortality

Virus, B.t., fungus, 

parasites, various 

predators

Virus, B.t., 

fungus, 

microsporidia, 

parasites and 

predators

the European strain, then and now, are conducted 

outside the generally infested area. Because of the 

flight capabilities of the Asian strain and the expanded 

potential host range, USDA policy is to eradicate 

moths exhibiting characteristic traits or genetic markers 

consistent with the Asian strain wherever feasible–even 

inside the generally infested area.

Knowledge of the time, location, and extent of an 

introduction is required to trigger eradication of the 

Asian strain in the generally infested area. In cases 

where deductive, circumstantial, or investigative 

information can be developed about an introduction of 

uncertain origin, eradication may also be conducted. 

The goal is to eradicate gypsy moths that exhibit traits 

characteristic of the Asian strain in a specific area 

wherever it may occur (within or outside the generally 

infested area). 

Treatments available are the same for both strains, but 

the timing of application differs. Eradication of the 

European strain begins with a detection survey that 

locates isolated infestations, followed by a delimiting 

survey confirming the presence of established 

populations and determining the approximate size and 

geographic extent of the infestation (see Appendix 

B for survey descriptions). Treatment ensues at 

the conclusion of the delimiting survey; time from 

detection to initial treatment is 1 to 2 years. 

Treatment for the Asian strain begins the year after 

detection. Time is not taken to conduct a delimiting 

survey, as an isolated infestation of the Asian strain 

could spread significantly because of the female’s 

flight capability, resulting in the need for an even 

larger eradication project. During the year following 

detection, the treatment area is determined using the 

best information available; the area extends beyond 

where male moths are collected, compensating for the 

distance females might fly. Delimiting surveys follow 

treatment, conducted throughout and significantly 

beyond the treated area. 
Biology, History, and Control Efforts for the Gypsy Moth

Appendix E - Page 

Population Phases.

Populations of the gypsy moth periodically build to 

high levels for one or more years, then collapse and 

remain at low levels for varying periods of time before 

increasing again. These changes in population levels 

pass through four phases (Doane and McManus 1981, 

USDA Forest Service 1989):

Innocuous Phase.

Populations are low and stable. Predation by small 

mammals and birds and parasitism by other insects 

appear to keep populations low (Campbell 1976, 

Elkinton and Leibhold 1990). This phase was 

undoubtedly the major contributing factor in the 1900 

decision to cancel the eradication program.

Release Phase.

Populations build rapidly. While not fully understood, 

mild winters followed by warm, dry springs and 

summers may increase survival and lead to population 

expansion and increase (Campbell and Sloan 1977c). 

Outbreak Phase.

Populations reach high levels, and feeding causes 

widespread moderate-to-heavy defoliation of 

susceptible hosts. Although predation and parasitism 

of caterpillars continue, the impact on gypsy moth 

populations is minor. As the outbreak progresses, 

the gypsy moth virus--a naturally occurring 

nucleopolyhedrosis virus--or a fungus (Entomophaga 

maimaiga) may begin to build in the population and 

contribute to its collapse (Campbell and Sloan 1977c). 

Decline Phase.

Populations collapse from overpopulation, starvation, 

infection by the virus or fungus and decreased 

reproduction. Males frequently outnumber females 

in these populations; the other phases exhibit 

approximately equal numbers of males and females. 

Host Plants.

Caterpillars of the European strain eat foliage from 

a wide variety of trees and shrubs. They prefer oaks, 

apple, sweetgum, speckled alder, basswood, gray and 

white birch, poplar, willow and hawthorn (McManus 

and others 1989). All instars feed on these species; 

later instars feed on some additional tree species 

shunned by early instars, such as cottonwood, hemlock, 

southern white cedar, and the pines and spruces in the 

eastern United States. The gypsy moth usually does 

not feed on some plants, including rhododendron, 

laurel, dogwood, and yellow poplar, although during an 

outbreak gypsy moth caterpillars will feed on almost all 

vegetation (McManus and others 1989). Appendix D 

provides the gypsy moth’s feeding preferences for over 

700 plant species. 

The Asian strain exhibits a broader range of preferred 

hosts than does the European strain (USDA APHIS 

1992). Studies show the Asian strain thrives with 

greater vigor than the European strain on many of the 

hosts species present in the United States, with the 

largest variability in growth rate observed on conifers 

(Wallner 1994).

E.2 1869 to 1910: Biological 

Controls Fail.

The European strain of the gypsy moth was considered 

a curiosity when it first escaped around 1869 from 

an insectary in Medford, Massachusetts. Public 

perception of the moth as a problem developed two 

decades later as the gypsy moth population exploded; 

citizens soon realized the consequences of allowing the 

moths to remain uncontrolled:

In the summer of 1889 it [the gypsy 

moth] threatened to overrun Medford, 

Massachusetts. The startled townspeople 

discovered caterpillars in astounding 

numbers, swarming through trees, eating 

leaves, and coating the ground below with 
Appendix E

Appendix E - Page 

droppings. People swept insects from their 

sidewalks, porches, and clothes; carried 

umbrellas to ward off droppings and falling 

caterpillars; and even wore face nets. The 

town, unable to deal with the situation, 

appealed to the state for aid. The striking 

nature of the infestation and its occurrence 

in an urban area brought quick response 

from the commonwealth, and an ambitious 

effort to deal with the pest was begun. 

(Dunlap 1980, p. 118)

Gypsy moth, initially assumed native, was first 

identified as an exotic pest in 1889 (Weseloh 2003c). 

Control methods included destroying egg masses, 

burning infested trees and shrubs, banding trees to 

trap caterpillars, and spraying insecticides. Paris 

green (copper aceto-arsenate), the first gypsy moth 

insecticide, was replaced with lead arsenate in 1893 

(McManus and McIntyre 1981).

Massachusetts discontinued efforts to eradicate the 

gypsy moth in 1900, mistakenly considering the project 

fully successful. The actual reason for the diminished 

presence of the pest was its entry into the innocuous 

phase, one of four periodic population phases. A 

second outbreak in 1906 prompted the Federal 

government to take action to eliminate the non-native 

insect. Eradication proved impossible; gypsy moth was 

already widespread. 

Entomologists with the (then) USDA Bureau of 

Entomology initiated studies to determine the life 

cycle of the insect and identify natural enemies from 

Europe for use against the pest. Introduction of 

identified natural enemies failed to stop the moths, and 

these biological control efforts were deemed failures. 

Funding reductions affected even basic research 

activities: 

Biological control proved to be much 

more difficult than either the scientists or 

the public had anticipated. Importing and 

establishing the moth’s natural enemies 

was neither simple nor inexpensive. 

Some of the parasites immediately died 

in the new environment, others refused 

to breed, and still others vanished 

without a trace when released. Some 

survivors were found to be preying on 

the moth, but with no noticeable effect 

on its population. (Dunlap 1980, p. 

121).

E.3 1911 to 1939: Chemical 

Insecticides Gain Favor.

The Bureau of Entomology issued a report in 1911, 

stating the parameters of effective use of biological 

controls against the gypsy moth in the United States:

. . . all fifty of the moth’s known 

European predators [would have to be 

imported and established], which would 

require long-term studies of the ecology 

of the moth and its enemies. (Dunlap 

1980, p. 121) 

The public, as well as scientists and politicians, quickly 

realized the successful use of biological controls would 

require extensive research and funding: 

With the end of hope that natural 

enemies would control the moth, both 

state and federal workers fell back on 

a piecemeal approach. They sought 

to reduce damage in highly visible 

and economically important areas—

roadsides and towns. (Dunlap 1980, 

p. 123)
Biology, History, and Control Efforts for the Gypsy Moth

Appendix E - Page 

The gypsy moth spread throughout New England. 

By 1914, the generally infested area included the 

southern half of New Hampshire, Rhode Island, eastern 

Connecticut, southern Vermont, and the eastern half of 

Massachusetts (McManus and McIntyre 1981). The 

use of chemical insecticides evolved as the favored 

form of control: 

Their popularity was due in part to ... 

the public’s desire for an immediate 

[visible] solution to [gypsy moth] 

problems and its reluctance to invest 

in long-term research that did not 

promise a certain or immediate return. 

Chemicals...gave immediate and 

gratifying visible results. Best of 

all, they could be used by individual 

landowners or towns without regard 

to coordination with other people or 

jurisdictions...

In forest spraying, however, chemicals 

proved ineffective. Better equipment 

and sprays now made roadside and urban 

spraying practical...[while] skyrockets 

and aerial bombs proved interesting 

but impractical. Spraying from 

planes or autogiros [early helicopters] 

seemed promising...[but] the hazards 

of tall trees, crosswinds, and irregular 

terrain made spraying difficult, but the 

most important factor was economic: 

American forests had too low a return 

per acre to justify the expense and 

repeated sprayings that were necessary 

to control the moth. 

The same economic calculations also 

doomed another, ecological, control 

method [that of silvicultural...] 

replacing stands of susceptible or 

favored food species with those that 

were most resistant to the moth’s attacks 

or less palatable. Unfortunately, this 

approach, like extensive forest spraying, 

presupposed a relatively high return per 

acre, and nothing came of it. (Dunlap 

1980, p. 123-124)

The United States Department of Agriculture (USDA), 

in cooperation with the infested States and Canada, 

established a barrier zone in 1923, extending from 

the Canadian border along the Hudson River and 

Champlain Valleys to Long Island. Gypsy moth 

infestations east of this barrier were designated for 

treatment by the States; infestations to the west for 

eradication. The first major infestation west of the 

barrier zone occurred in Pennsylvania in 1932. Six 

years later, the New England hurricane of September 

21, 1938, spread the gypsy moth hundreds of miles into 

new territory. In the following year the barrier zone 

had become generally infested. 

E.4 1940 to 1957: DDT Gets 

Widespread Use.

Consideration and experimentation of new insecticide 

controls occurred both before and during World War 

II. Experimental use of cryolite as a gypsy moth 

insecticide in Pennsylvania in the 1940s proved 

ineffective; the most promising new insecticide was 

a synthetic organic chemical, dichloro-diphenyltrichlorethane (DDT): 

Even before the end of World War II, 

American and Canadian scientists 

were using experimental lots of the 

new chemical for aerial spraying on 

northern forests to test DDT against the 

gypsy moth and the spruce budworm. 

The results were astounding. Less 

than a pound of DDT per acre killed 

almost all the caterpillars, but it did 

not, apparently, cause any significant 

damage to wildlife. (Dunlap 1980, p. 

124) 
Appendix E

Appendix E - Page 

Experimental use of DDT in Pennsylvania proved 

more effective than cryolite, leading to the erroneous 

conclusion that successful eradication of gypsy moth 

in the State occurred by 1948. Undetected infestations, 

however, led to further outbreaks and continued spread 

(Nichols 1961). 

Gypsy moth infestations proliferated in the 1950s, 

and another barrier zone was set up through the 

Adirondack plateau in an attempt to prevent spread to 

the south and west. However, detection of the insect in 

previously uninfested areas occurred by the mid-1950s. 

New Jersey, New York, Pennsylvania, and Michigan 

reported populations, initiating a major Federal effort to 

eradicate the gypsy moth: 

The first phase, to begin in the spring 

of 1957, involved aerial spraying to 

eliminate outlying populations of the 

moth in New York, Pennsylvania, and 

Michigan. If these were successful, a 

second phase would follow, wiping 

out the main body in New England... 

The moth’s periodic outbreaks caused 

serious but local damage, and there was 

no urgent demand to quell the latest 

one. The only clear rationale was the 

availability of DDT...

Spraying began in April 1957 and 

lasted until June, covering more than 

3 million acres in the Northeast with 

DDT. It brought a storm of criticism 

from the populace, from scientists, and 

from local and state officials. Some 

objected to the nuisance; cars dotted 

with scum or pools covered with layers 

of oil [from the carrier used to spray the 

DDT]. Other effects were more serious: 

dairy farmers complained that DDT fell 

on their pastures and passed into the 

milk, contaminating it. Organic farmers 

on Long Island also protested, for the 

sprays rendered their crops unsuitable 

for the special markets... The program 

also met legal challenge, the first serious 

environmental litigation against a pest 

control program...it [proved to be] too 

controversial for officials and bureaus 

whose budgets depended on public 

goodwill. (Dunlap 1980, p. 124-125) 

E.5 1958 to the Mid-1980s: 

Safer Treatments Needed.

During its use, DDT application for gypsy moth control 

totaled over 12 million acres (4.9 million ha) of forest 

in nine northeastern States and Michigan (U.S. EPA 

1975). Questions concerning the non-target effects of 

DDT led to its replacement by the carbamate, carbaryl, 

in the late 1950s. DDT use came to an end soon after 

publication of Rachel Carson’s book Silent Spring, in 

1962. The Forest Service stopped using DDT in its 

Eastern Region (Paananen and others 1987). Although 

considered safer than DDT, in certain formulations 

carbaryl demonstrated toxicity to honeybees (USDA 

1985).

Between 1970 and 1981, suppression of gypsy moth 

outbreaks was accomplished with aerial applications 

of broad-spectrum insecticides, including carbaryl 

and the organophosphate trichlorfon and, to a lesser 

degree, acephate. These broad-spectrum, nerve-poison 

insecticides killed not only gypsy moth caterpillars, but 

many other immature and adult insects in treated areas.

The initiation of research efforts to find effective means 

of gypsy moth control began in the 1970s, including the 

use of the gypsy moth nucleopolyhedrosis virus (NPV) 

and Bacillus thuringiensis var. kurstaki (B.t.k.) as 

biological control agents. Gypchek, registered in 1978, 

is an insecticide made from NPV. The insect growthregulator, diflubenzuron, also registered in 1978, 

offered an attractive alternative with fewer effects on 

non-target organisms than other chemical insecticides. 
Biology, History, and Control Efforts for the Gypsy Moth

Appendix E - Page 

The USDA increased exploration for foreign parasites 

and predators of the gypsy moth in 1971, also funding 

research on a synthetic pheromone (disparlure) and 

gypsy moth population dynamics and environmental 

effects (McManus and McIntyre 1981). Results of 

this and other research led to the development of noninsecticidal methods, such as mass trapping, mating 

disruption and the sterile insect technique for use in 

gypsy moth projects. 

Attempts to eliminate the gypsy moth from the United 

States were abandoned in the 1970s and a two-phase 

management approach adopted: suppression of 

outbreaks in the generally infested area and eradication 

of isolated infestations resulting from inadvertent 

transport of the insect by people into the uninfested 

area. 

Diflubenzuron and B.t.k. largely replaced carbaryl and 

trichlorfon as the insecticides of choice in cooperative 

gypsy moth suppression projects by the mid-1980s. 

Cooperative suppression projects last used trichlorfon 

in 1984 and carbaryl in 1987 (USDA Forest Service 

1994d). Use of broad-spectrum chemical insecticides 

in cooperative eradication projects ceased in 1989 

(USDA APHIS 1992).

E.6 Mid-1980s to the Present: 

Adoption of Integrated Pest 

Management.

Integrated pest management (IPM) became the standard 

approach to gypsy moth suppression and eradication in 

the 1980s, and continues to this writing. This approach 

employs the use of various management practices, 

including the application of chemical and biological 

insecticides and utilization of non-insecticidal methods.

Up to this point, controls against high-density 

populations of the gypsy moth were employed in 

relatively small treatment blocks. Three successive 

studies began attempts to keep low-density populations 

from expanding over geographic areas of increasing 

size.

The Forest Service led Federal, State, and county 

agencies in an IPM study of a five-county area in 

Maryland from 1983 to 1987 (Reardon and others 

1993). Using geographic information system (GIS) 

computer technology to collect and store data, this 

first study accomplished advances in the operational 

use of controls specific to gypsy moth. An improved 

formulation of the nucleopolyhedrosis virus (Gypchek) 

resulted, as well as the first release of sterile eggs. 

The second study, conducted in 38 counties along the 

Appalachian Mountains in Virginia and West Virginia, 

began in 1987 and concluded in 1992 (USDA Forest 

Service 1989). Researchers successfully minimized 

damage in the project area, reducing adverse 

environmental effects using gypsy-moth-specific 

treatments in an IPM approach, demonstrating the 

technical feasibility of slowing the spread of the gypsy 

moth (USDA Forest Service 1994e). 

The third study, a 5-year pilot project started in 1992, 

utilized the same concepts and methodologies in four 

States (Virginia, West Virginia, North Carolina, and 

Michigan), to determine the operational and economic 

feasibility of a nationwide program to slow the spread 

of the gypsy moth.

Building upon these three field studies, development 

and improvement of methods for IPM continued on 

a national scale for the gypsy moth and for all major 

forest pests. Participants included the Forest Service 

National Center for Forest Health Management and 

other units of the Forest Service, APHIS, Agricultural 

Research Service and Cooperative State Research, 

Education, and Extension Service.
Appendix E

Appendix E - Page 10

Following the issuance of the 1995 Final 

Environmental Impact Statement “Gypsy Moth 

Management in the United States: a cooperative 

approach” and the subsequent signing of the 1996 

Record of Decision, USDA implemented a program 

that included eradication, suppression, and slow-thespread (STS) projects to control the gypsy moth. An 

STS pilot project concluded in 1999, leading to the 

first full-scale projects using STS methodologies 

operationally in 2000. Eradication, suppression, and 

STS projects continue to this writing.

E.7 1991 to the Present: Asian 

Strain Creates Additional 

Concern.

The Asian strain of the gypsy moth, found for the first 

time in the United States in 1991, is of concern because 

females have the ability to fly. This mobility poses 

the possibility of the Asian strain spreading at an even 

faster rate than does the European strain. Eradication 

projects for the Asian strain were conducted in Oregon 

and Washington in 1992 and 1993; in North Carolina 

in 1994 and 1995; and in Portland, Oregon, in 2000. 

A single Asian male found in a survey trap in the 

Port of Long Beach, California, in 2003 led to the 

implementation of preventive control projects there in 

2004.

Isolated infestations of the European strain continue 

to be a problem outside of the generally infested area, 

usually resulting from inadvertent movement of gypsy 

moth life stages on articles such as cars, campers, 

outdoor furniture, and nursery stock. Port-of-entry 

activities to prevent all gypsy moth strains from 

entering the United States are ongoing (see Appendix 

B). Surveys using pheromone traps continue nationally 

to detect introduction and determine if eradication is 

necessary.
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Gypsy Moth Management in the United States: a cooperative approach 

Draft Supplemental Environmental Impact Statement 

The complete Draft Supplemental Environmental Impact Statement, Gypsy 

Moth Management in the United States: a cooperative approach, consists of 

four volumes: 

Volume I Summary 

Volume II Chapters 1-8 and Appendixes A, B, C, D, E 

Volume III Appendixes F, G, H, I 

Volume IV Appendixes J, K, L, M 

Lead Agency:	 Forest Service, U.S. Department of Agriculture 

(USDA) 

Responsible Official:		 James R. Hubbard, Deputy Chief for State and 

Private Forestry 

Sidney R. Yates Federal Building 

201 14th Street, SW 

Washington, DC 20250 

For More Information: Bill Oldland, SEIS Team Leader 

180 Canfield Street 

Morgantown, WV 26505 

Joint Lead Agency:	 Animal and Plant Health Inspection Service 

(APHIS), USDA 

Responsible Official:		 Richard L. Dunkle, Deputy Administrator for Plant 

Protection and Quarantine 

4700 River Road 

Riverdale, MD 20737-1236 

For More Information:	 Weyman P. Fussell, Ph.D., Gypsy Moth Program 

Coordinator 

Emergency and Domestic Programs, APHIS 

4700 River Road, Unit 134 

Riverdale, MD 20737-1236 
 

 

 

 

Abstract: The USDA Forest Service and Animal and Plant Health Inspection Service are proposing an addition 

to the gypsy moth management program that was described in the 1995 Environmental Impact Statement--Gypsy 

Moth Management in the United States: a cooperative approach--and chosen in the 1996 Record of Decision. 

The agencies are proposing these new treatment options: adding the insecticide tebufenozide, or adding the 

insecticide tebufenozide and other new treatment(s) that may become available in the future to manage gypsy 

moths, provided that the other treatment(s) poses no greater risk to human health and nontarget organisms than are 

disclosed in this Draft SEIS for the currently approved treatments and tebufenozide. 

Commenting on this Draft Supplemental Environmental Impact Statement: Reviewers should provide the 

Forest Service with their comments during the review period of this draft supplemental environmental impact 

statement. Timely comments will enable the Forest Service to analyze and respond to all of the comments at one 

time and to use information acquired in the preparation of the final supplemental environmental impact statement, 

thus avoiding undue delay in the decision making process. Furthermore, the more specific and substantive 

the comments, the better for reviewers and the agencies alike. Reviewers have an obligation to structure their 

participation in the National Environmental Policy Act process so that it is meaningful and alerts the agency to 

the reviewer’s position and contentions (Vermont Yankee Nuclear Power Corp. v. NRDC, 435 U.S. 519, 553, 

1978). Environmental objections that could have been raised at the draft stage may therefore be forfeited, if not 

raised until after completion of the final environmental impact statement (Department of Transportation v. Public 

Citizen, 541 U.S. 752, 764 (2004). Comments on this draft supplemental environmental impact statement should 

be specific and should address the adequacy of the statement and the merits of the alternatives discussed (40 CFR 

1503.3). 

Web Site for Draft SEIS: The Draft SEIS is available for viewing at www.na.fs.fed.us/wv/eis 

Send Comments to: Bill Oldland, SEIS Team Leader 

USDA Forest Service 

180 Canfield Street 

Morgantown, WV 26505
 

304-285-1585
 

Date Comments Must Be Received: See cover letter for the date that comments are due in Morgantown, WV. 
 

 

 

 

 

 

 

 

Contents 

Appendix F. Bacillus thuringiensis kurstaki (B.t.k.) Risk Assessment 

Appendix G. Gypchek (Nucleopolyhedrosis Virus) Risk Assessment 

Appendix H. Disparlure Risk Assessment 

Appendix I. Diflubenzuron Risk Assessment 

Figures 

Figure F-1. This insecticide mist blower, designed and constructed in 1946 by Quincy Forestry Department, 

Massachusetts, was used to spray trees in residential areas. 

Figure G-1 Creosote was used in 1895 to treat gypsy moth egg masses. 

Figure H-1. Female gypsy moth pupae were gathered in Massachusetts in 1948 in order to obtain sex 

attractant for trapping programs. 

Figure I-1. The first power spraying apparatus was used in gypsy moth control operations before 1900. 

Photo Credits 

Figure F-1. (UGA1275008) USDA Forest Service Archives, www.forestryimages.org 

Figure G-1. (UGA1275948) USDA Forest Service Archives, www.forestryimages.org 

Figure H-1. (UGA1275009) USDA Forest Service Archives, www.forestryimages.org 

Figure I-1. (UGA1275014) USDA Forest Service Archives, www.forestryimages.org 
Appendix F 

Bacillus thuringiensis 

kurstaki (B.t.k.) 

Risk Assessment 

Figure F-1. This insecticide mist blower, designed and constructed in 1946 by 

Quincy Forestry Department, Massachusetts, was used to spray trees in residential 

areas. 

SERA TR 03-43-05-02d 

Control/Eradication Agents for the


Gypsy Moth ­

Human Health and Ecological Risk Assessment for


Bacillus thuringiensis var. kurstaki (B.t.k.)


FINAL REPORT
 

Prepared for: 

USDA, Forest Service


Forest Health Protection 

GSA Contract No. GS-10F-0082F
 

USDA Forest Service BPA: WO-01-3187-0150
 

Requisition No.: 43-3187-1-0269
 

Task No. 5
 

Submitted to:
 

Dave Thomas, COTR
 

Forest Health Protection Staff
 

USDA Forest Service
 

Rosslyn Plaza Building C, Room 7129C


1601 North Kent Street
 

Arlington, VA 22209
 

Submitted by:


Patrick R. Durkin
 

Syracuse Environmental Research Associates, Inc.

5100 Highbridge St., 42C


Fayetteville, New York 13066-0950
 

Telephone: (315) 637-9560


Fax: (315) 637-0445


E-Mail: SERA_INC@msn.com
 

Home Page: www.sera-inc.com
 

June 8, 2004 (Risk Assessment)


July 20, 2007 (Correction of Pagination)
 
 

TABLE OF CONTENTS
 

TABLE OF CONTENTS ...................................................... ii
 

LIST OF TABLES .......................................................... iv
 

LIST OF FIGURES .......................................................... v
 

LIST OF APPENDICES ...................................................... v
 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . vi
 

COMMON UNIT CONVERSIONS AND ABBREVIATIONS . . . . . . . . . . . . . . . . . . . . . . viii
 

CONVERSION OF SCIENTIFIC NOTATION ................................... ix
 

EXECUTIVE SUMMARY .................................................... x
 

1. INTRODUCTION ....................................................... 1-1
 

2. PROGRAM DESCRIPTION .............................................. 2-1
 

2.1. Overview ...................................................... 2-1
 

2.2. Chemical Description and Commercial Formulations . . . . . . . . . . . . . . . . . . . . 2-1
 

2.3. Use Statistics ................................................... 2-3
 

3. HUMAN HEALTH RISK ASSESSMENT ................................... 3-1
 

3.1. HAZARD IDENTIFICATION ..................................... 3-1
 

3.1.1. Overview. .................................................... 3-1
 

3.1.2. Epidemiology Studies. .......................................... 3-1
 

3.1.3. Mechanism of Action............................................ 3-7
 

3.1.4. Acute Oral Toxicity. ............................................ 3-8
 

3.1.5. Subchronic or Chronic Systemic Toxic Effects. . . . . . . . . . . . . . . . . . . . . . . . 3-9
 

3.1.6. Effects on Nervous System. ...................................... 3-9
 

3.1.7. Effects on Immune System. ...................................... 3-9
 

3.1.8. Effects on Endocrine System. .................................... 3-11
 

3.1.9. Reproductive and Teratogenic Effects. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-11
 

3.1.10. Carcinogenicity and Mutagenicity. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-11
 

3.1.11. Irritation (Effects on the Skin and Eyes). . . . . . . . . . . . . . . . . . . . . . . . . . . 3-11
 

3.1.12. Systemic Toxic Effects from Parenteral Exposure. . . . . . . . . . . . . . . . . . . 3-13
 

3.1.13. Inhalation Exposure. .......................................... 3-13
 

3.1.14. Impurities. .................................................. 3-13
 

3.1.15. Inerts....................................................... 3-14
 

3.2. EXPOSURE ASSESSMENT ........................................... 3-16
 

3.2.1. Overview. ................................................... 3-16
 

3.2.2. General Issues. ............................................... 3-16
 

3.2.3. Workers. .................................................... 3-17
 

3.2.4. Member s of General Public. ............. .................. . 3-18
 

3.3. DOSE-RESPONSE ASSESSMENT ...................................... 3-19
 

3.3.1. Overview. ................................................... 3-19
 

3.3.2. Existing Guidelines. ........................................... 3-19
 

3.3.3. Human Data. ................................................. 3-20
 

3.3.4. Animal Data.................................................. 3-23
 

3.3.5. Values Used in Risk Characterization. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-25
 

3.4. RISK CHARACTERIZATION .......................................... 3-27
 

3.4.1. Overview. ................................................... 3-27
 

3.4.2. Irritant Effects. ............................................... 3-28
 

ii 
 

TABLE OF CONTENTS (continued) 

3.4.3. Serious Adverse Effects......................................... 3-29
 

3.4.4. Groups at Special Risk.......................................... 3-31
 

3.4.5. Cumulative Effects and Connected Actions. . . . . . . . . . . . . . . . . . . . . . . . . 3-32
 

4. ECOLOGICAL RISK ASSESSMENT ....................................... 4-1
 

4.1. HAZARD IDENTIFICATION ..................................... 4-1
 

4.1.1. Overview. .................................................... 4-1
 

4.1.2. Toxicity to Terrestrial Organisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-2
 

4.1.2.1. Mammals ............................................. 4-2
 

4.1.2.2. Birds ................................................. 4-3
 

4.1.2.3. Terrestrial Invertebrates .................................. 4-3
 

4.1.2.4. Terrestrial Plants (Macrophytes) . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-8
 

4.1.2.5. Terrestrial Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-8
 

4.1.3. Aquatic Organisms.............................................. 4-8
 

4.1.3.1. Fish .................................................. 4-8
 

4.1.3.2. Amphibians ........................................... 4-9
 

4.1.3.3. Aquatic Invertebrates .................................... 4-9
 

4.1.3.4. Aquatic Plants ......................................... 4-9
 

4.2. EXPOSURE ASSESSMENT ........................................... 4-10
 

4.2.1. Overview. ................................................... 4-10
 

4.2.2. Terrestrial Animals. ........................................... 4-10
 

4.2.2.1. Terrestrial Vertebrates .................................. 4-10
 

4.2.2.2. Terrestrial Invertebrates ................................. 4-11
 

4.2.2.3. Other Terrestrial Species ................................ 4-12
 

4.2.3. Aquatic Organisms............................................. 4-12
 

4.3. DOSE-RESPONSE ASSESSMENT ...................................... 4-14
 

4.3.1. Overview. ................................................... 4-14
 

4.3.2. Toxicity to Terrestrial Organisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-14
 

4.3.2.1. Terrestrial Vertebrates .................................. 4-14
 

4.3.2.2. Terrestrial Invertebrates ................................. 4-15
 

4.3.3. Aquatic Organisms............................................. 4-17
 

4.3.3.1. Fish ................................................. 4-17
 

4.3.3.2. Invertebrates .......................................... 4-18
 

4.4. RISK CHARACTERIZATION .......................................... 4-20
 4.4.1. Overview. ................................................... 4-20
 4.4.2. Terrestrial Organisms........................................... 4-20
 4.4.2.1. Terrestrial Vertebrates .................................. 4-20
 4.4.2.2. Terrestrial Invertebrates ................................. 4-21
 4.4.3. Aquatic Organisms............................................. 4-22
 

5. List of Studies Consulted. . . . . . . . . . . . . . ................... 5-1
 

iii 
TABLE OF CONTENTS (continued)
 

LIST OF TABLES
 

Table 2-1: Commercial formulations of B.t.k. that may be used in USDA Gypsy Moth 

Programs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-1 

Table 2-2: Use of B.t.k. from 1995 to 2001 for Suppression, Eradication, and Slow 

the Spread . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-2 

Table 3-1: Epidemiology Studies on B.t.k. Formulations . . . . . . . . . . . . . . . . . . . . . . . Tables-3 

Table 3-2: Publically available information on inerts used in B.t.k. formulations . . . . . Tables-4 

Table 3-3: Overview of exposure data for workers and members of the general public Tables-5 

Table 3-4: Post-spray symptoms reported by ground-spray workers and controls . . . . Tables-6 

Table 3-5: Summary of the number of symptoms per worker in 120 ground-spray workers

segregated by exposure groups and use of protective masks . . . . . . . . . . . . . . . Tables-7 

Table 3-6: Self-reported symptoms in individuals before and after the aerial application 

of B.t.k. ........................................................ Tables-8 

Table 3-7: Exposure conversions for mice and humans with effects noted in mice after 

intranasal instillations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-9 

Table 3-8: Risk characterization for serious health effects from exposure to B.t.k. .. Tables-10 

Table 4-1: Mortality in species subject to foliage treated with Foray 48B at 

89 BIU/ha . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-11 

Table 4-2: Mortality in species subject to foliage treated with Dipel 8AF at 

99 BIU/ha . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-13 

Table 4-3: Summary of exposures used in ecological risk assessment . . . . . . . . . . . . Tables-14 

Table 4-4: Summary of toxicity values used in ecological risk assessment . . . . . . . . Tables-15 

Table 4-5: Data used in dose-response assessment for non-target insects . . . . . . . . . . Tables-16 

Table 4-6: Risk characterization for ecological risk assessment of B.t.k. .......... Tables-17 

iv 
TABLE OF CONTENTS (continued)
 

LIST OF FIGURES
 

Figure 3-1: Number of symptoms per worker based on total exposure to B.t.k. 

(millions of cfu hours) and the use of protective masks . . . . . . . . . . . . . . . . . Figures-1 

Figure 3-2: Dose-response relationships in mice after intranasal administration of B.t.k. 

With or without previous challenge with influenza virus at 4% of the LD50 . . Figures-2 

Figure 4-1: Dose-response assessment for non-target terrestrial invertebrates . . . . . . Figures-3 

Figure 4-2: Distribution of sensitivity in various non-target lepidoptera . . . . . . . . . . Figures-4 

LIST OF APPENDICES 

Appendix 1: Toxicity in Mammals.........................................................................Appendix 1-1
 

Appendix 2: Toxicity in Birds..................................................................................Appendix 2-1

Appendix 3: Toxicity in Non-target Lepidoptera.....................................................Appendix 3-1

Appendix 4: Toxicity in Non-Target Terrestrial Insects Other Than Lepidoptera...Appendix 4-1
 

Appendix 5: Toxicity of B.t.k. and B.t.k. Formulations to Fish...............................Appendix 5-1
 

Appendix 6: Toxicity of B.t.k. and B.t.k. Formulations to Aquatic Invertebrates....Appendix 6-1
 

v 
GENERAL ACRONYMS, ABBREVIATIONS, AND SYMBOLS
 

a.i. 

AEL 

APHIS 

ARS 

ATSDR 

BCF 

B.t.k. 

BIU 

bw 

cfu 

cm 

DFB 

EC50 

EC100 

EEC 

EIS 

F 

F1 

FH 

FS 

FTU 

g

GC 

GRAS 

HQ 

IARC 

IRIS 

i.p.

IU 

kg

Ko/c 

Kow

Kp

L 

LdNPV 

lb 

LC50 

LD50 

LD95 

LOAEL 

m 

M 

mg

mg/kg/day 

mL 

MSDS 

MW 

NCI 

NOAEL 

NOEC 

active ingredient 

adverse-effect level 

Animal and Plant Health Inspection Service 

Agricultural Research Station 

Agency for Toxic Substances and Disease Registry 

bioconcentration factor 

Bacillus thuringiensis var. kurstaki 

Billions of international units 

body weight 

colony forming units 

centimeter 

diflubenzuron 

concentration causing 50% inhibition of a process 

concentration causing complete inhibition of a process 

expected environmental concentration 

environmental impact statement 

female 

first filial generation 

Forest Health 

Forest Service 

forestry toxic units 

gram 

gas chromatography 

generally recognized as safe 

hazard quotient

International Agency for Research on Cancer 

Integrated Risk Information System 

intraperitoneal 

international units 

kilogram 

organic carbon partition coefficient 

octanol-water partition coefficient 

skin permeability coefficient 

liter 

gypsy moth (Lymantria dispar) nucleopolyhedrosis virus 

pound

lethal concentration, 50% mortality 

lethal dose, 50% mortality 

lethal dose, 95% mortality 

lowest-observed-adverse-effect level 

meter 

male 

milligram 

milligrams of agent per kilogram of body weight per day 

milliliter 

material safety data sheet 

molecular weight 

National Cancer Institute 

no-observed-adverse-effect level 

no-observed-effect concentration 

vi 
ACRONYMS, ABBREVIATIONS, AND SYMBOLS (continued) 

NOEL no-observed-effect level 

NRC National Research Council 

OPP Office of Pesticide Programs 

ORD Office of Research and Development 

OTS Office of Toxic Substances 

ppm parts per million 

RBC red blood cells 

RfD reference dose 

UF uncertainty factor 

U.S. United States 

U.S. EPA U.S. Environmental Protection Agency 

USDA United States Department of Agriculture 

> greater than 

$ greater than or equal to 

< less than 

# less than or equal to 

= equal to 

� approximately equal to 

vii 
COMMON UNIT CONVERSIONS AND ABBREVIATIONS
 

To convert ... Into ... Multiply by ... acres 

acres 

hectares (ha) 

square meters (m ) 2 

0.4047 

4,047 

atmospheres millimeters of mercury 760 

centigrade Fahrenheit 1.8CE+32 

centimeters inches 0.3937 

cubic meters (m ) 3 liters (L) 1,000 

Fahrenheit centigrade 0.556FE-17.8 

feet per second (ft/sec) miles/hour (mi/hr) 0.6818 

gallons (gal) liters (L) 3.785 

gallons per acre (gal/acre) liters per hectare (L/ha) 9.34 

grams (g) ounces, (oz) 0.03527 

grams (g) pounds, (oz) 0.002205 

hectares (ha) acres 2.471 

hectares (ha) square meters 10,000 

kilograms (kg) ounces, (oz) 35.274 

kilograms (kg) pounds, (lb) 2.2046 

kilograms per hectare (hg/ha) pounds per acre (lb/acre) 0.892 

kilometers (km) 

liters (L) 

miles (mi) 

cubic centimeters (cm ) 3 

0.6214 

1,000 

liters (L) gallons (gal) 0.2642 

liters (L) ounces, fluid (oz) 33.814 

miles (mi) kilometers (km) 1.609 

miles per hour (mi/hr) cm/sec 44.70 

milligrams (mg) ounces (oz) 0.000035 

meters (m) feet 3.281 

ounces (oz) grams (g) 28.3495 

ounces per acre (oz/acre) grams per hectare (g/ha) 70.1 

ounces per acre (oz/acre) 

ounces fluid 

kilograms per hectare (kg/ha) 

cubic centimeters (cm ) 3 

0.0701 

29.5735 

pounds (lb) grams (g) 453.6 

pounds (lb) kilograms (kg) 0.4536 

pounds per acre (lb/acre) 

pounds per acre (lb/acre) 

pounds per acre (lb/acre) 

kilograms per hectare (kg/ha) 

mg/square meter (mg/m ) 2 

μg/square centimeter (μg/cm ) 2 

1.121 

112.1 

11.21 

pounds per gallon (lb/gal) 

square centimeters (cm ) 2 

square centimeters (cm ) 2 

square meters (m ) 2 

grams per liter (g/L) 

square inches (in ) 2 

square meters (m ) 2 

square centimeters (cm ) 2 

119.8 

0.155 

0.0001 

10,000 

yards meters 0.9144 

Note: All references to pounds and ounces refer to avoirdupois weights unless otherwise specified. 

viii 
CONVERSION OF SCIENTIFIC NOTATION
 

Scientific Decimal Verbal 

Notation Equivalent Expression 

1 @ 10-10 0.0000000001 One in ten billion 

1 @ 10-9 0.000000001 One in one billion 

1 @ 10-8 0.00000001 One in one hundred million 

1 @ 10-7 0.0000001 One in ten million 

1 @ 10-6 0.000001 One in one million 

1 @ 10-5 0.00001 One in one hundred thousand 

1 @ 10-4 0.0001 One in ten thousand 

1 @ 10-3 0.001 One in one thousand 

1 @ 10-2 0.01 One in one hundred 

1 @ 10-1 0.1 One in ten 

1 @ 100 1 One 

1 @ 101 10 Ten 

1 @ 102 100 One hundred 

1 @ 103 1,000 One thousand 

1 @ 104 10,000 Ten thousand 

1 @ 105 100,000 One hundred thousand 

1 @ 106 1,000,000 One million 

1 @ 107 10,000,000 Ten million 

1 @ 108 100,000,000 One hundred million 

1 @ 109 1,000,000,000 One billion 

1 @ 1010 10,000,000,000 Ten billion 

ix 
EXECUTIVE SUMMARY
 

This document updates the human health and ecological risk assessments on Bacillus 

thuringiensis var. kurstaki (B.t.k.) prepared in 1995 in support of the Final Environmental

Impact Statement (FEIS) for the Cooperative Gypsy Moth Management Program sponsored by

the USDA Forest Service and APHIS. B.t.k. is used in USDA Forest Service and APHIS 

programs to control or eradicate the gypsy moth (Lymantria dispar). The updated risk 

assessments define the environmental consequences of using B.t.k. in these programs. 

This is a technical support document and it addresses some specialized technical areas. Thus,

parts of this document may contain information that is difficult for some readers to understand. 

These technical discussions are necessary to support the review of the document by individuals

with specialized training. Nevertheless, an effort is made to ensure that the conclusions 

reached in the document and the bases for these conclusions can be understood by individuals

who do not have specialized training in the chemical and biological sciences. In addition to 

this executive summary, each major section of the document starts with an overview section

that is intended to summarize the technical discussion in a manner that most individuals will 

understand. 

Sensitive terrestrial insects are the only organisms likely to be seriously affected by exposure to

B.t.k. or its formulations. All sensitive terrestrial insects are lepidoptera and include some

species of butterfly, like the endangered Karner blue and some swallowtail butterflies and

promethea moths. At the application rates used to control gypsy moth populations, mortality

rates among sensitive terrestrial insects are likely to range from approximately 80% to 94% or 

more. The risk characterization for other wildlife species is unambiguous: under foreseeable

conditions of exposure, adverse effects are unlikely to be observed. 

In terms of potential human health effects, formulations of B.t.k. are likely to cause irritation to

the skin, eyes, and respiratory tract; however, serious adverse health effects are implausible. 

For members of the general public, exposure levels are estimated to be below the functional

human NOAEL for serious adverse effects by factors of about 28,000 to 4,000,000 [4 million]. 

At the extreme upper range of exposure in ground workers, exposure levels are estimated to be

below the functional human NOAEL for serious effects by a factor of 25. This assessment is 

based on reasonably good monitoring data, conservative exposure assumptions, and an 

aggressive and protective use of the available toxicity data. 

PROGRAM DESCRIPTION 

Bacillus thuringiensis (B.t.) is a bacteria that is found in most of the world. Various strains of 

B.t., including B.t.k., are commonly found in soil, foliage, wildlife, water, and air. All 

commercial formulations of B.t.k. used by the USDA contain the HD-1 strain. Ten 

formulations of B.t.k. are used in USDA programs and all are supplied by Valent USA Corp or 

subsidiaries. Historically, each of the producers of B.t.k. formulations maintained separate 

stock strains and it appears that B.t.k. strain HD-1 may actually be a set of related strains or 

sub-strains. 

B.t.k. formulations are complex chemical mixtures. B.t.k. is cultured or grown in a media

containing water and nutrients including sugars, starches, proteins, and amino acids. These 

nutrients are themselves chemically complex and variable biological materials such as animal

foodstuffs, a variety of flours, yeasts, and molasses. Relatively small quantities of essential

elements, minerals, or salts also may be added to create optimal growth conditions. Other 

materials may also be used at various stages of production to enhance growth or facilitate the

recovery of B.t.k. from the growth media. The other components of the formulation are mostly 

x 
water and a complex mixture of culture media and metabolites. The composition of the growth

media used by a manufacturer may change over time, as different sources of nutrient material

are used. 

Application rates are expressed in billions of international units (BIU), which is a measure of

the activity or potency of the formulation rather than an expression of mass. Typical 

application rates for B.t.k. range from 24 BIU/acre to more than 36 BIU/acre. The range of

application rates used in the current risk assessment is 20 to 40 BIU/acre, which is equivalent

to about 49 to 99 BIU/ha. Any preparation of bacteria carries the potential for contamination

with other possibly pathogenic microorganisms, which must be addressed by proper quality

control procedures. U.S. EPA requires that spore preparations of B.t. are produced by pure

culture fermentation procedures with adequate quality control measures to detect either

contamination with other microorganisms or changes from the characteristics of the parent B.t. 

strain. Although B.t.k. formulations may be applied by aerial spray or by ground spray, the

number of aerial applications far exceeds the number of ground applications. More than 1 

million pounds of B.t.k. are applied annually in the United States to control the gypsy moth. A 

total of 2,743,816 acres were treated with B.t.k. formulations between 1995 and 2002, for an 

average annual treatment rate of approximately 343,000 acres per year. 

HUMAN HEALTH RISK ASSESSMENT 

Hazard Identification – Most risk assessments for chemical and biological agents are based

on relatively standard toxicity studies in experimental mammals. B.t.k., however, is different in 

that several epidemiology studies – i.e., studies on populations of humans who have been

exposed to B.t.k. – provide useful information regarding the plausibility of observing human

health effects after B.t.k. applications that are identical or closely related to applications used in

USDA programs to control the gypsy moth. The results of standard toxicity studies on B.t.k. 

and its formulations are used in this risk assessment to supplement information provided by

epidemiology studies. 

Irritation of the eyes, skin, and respiratory tract might be associated with exposures to B.t.k. and 

commercial formulations of B.t.k. Irritant effects are noted in experimental animal studies as

well as in epidemiology studies and case reports. Other more serious signs of toxicity are not 

likely to occur as a result of human exposure to B.t.k. Specifically, there is little indication that 

B.t.k. is associated with pathogenicity in humans and no indication of endocrine disruption or

reproductive effects in humans after exposure to B.t.k. formulations. In addition, carcinogenic 

and mutagenic effects are not likely to results from exposure to B.t.k. or its formulations. The 

potential for allergenicity of B.t.k. is somewhat more difficult to assess. There are reported

incidents of potential skin sensitization and antibody induction in some individuals after 

exposure to B.t.k. formulations. 

Exposure Assessment – Exposure assessments usually estimate the amount or concentration 

of an agent to which an individual or population might be exposed via ingestion, dermal

contact, or inhalation. The exposure assessments are then compared with toxicity studies based

on similar types of exposure—i.e., the dose-response assessment—and then the risk is

quantified. The human health risk assessment for B.t.k. is unusual in two respects. First, the 

most directly relevant data used to characterize risk are based on actual applications of B.t.k. 

formulations where exposure is best characterized as an application rate. Second, the apparent 

lack of a specific mechanism of toxicity for B.t.k. makes selecting the most appropriate 

measure of exposure somewhat arbitrary. 

xi 
Dose-Response Assessment – Based on conclusions reached by the U.S. EPA and World

Health Organization that irritation of the skin, eyes, or respiratory tract are most likely the only

human health effects to be expected from exposure to B.t.k., the dose-response assessment is 

relatively simple. Moreover, there is no information from epidemiology studies or studies in

experimental mammals that B.t.k. is likely to cause severe adverse health effects in humans

under any set of plausible exposure conditions. Notwithstanding these assertions, a recent

epidemiology study suggests that the irritant effects of B.t.k. may occur with notable frequency

at exposure levels that are typical of those used in programs to control the gypsy moth. By

comparison, a study in workers demonstrates that the frequency of the irritant effects does not

increase substantially even at very high exposure levels. This lack of a strong dose-response

relationship is somewhat unusual but is consistent with experimental data in mammals. 

Based on recent experimental studies which are not typically used in a quantitative doseresponse assessment, it is possible to define very high exposure levels for B.t.k. which might

pose a serious health hazard and it is possible to define a NOAEL for such effects that is

consistent with the available human data. The exposure data are expressed in units of colony 3 forming units (cfu). Specifically, cumulative exposures of up to 1.4×1010 cfu/m × hour are not

likely to result in adverse effects. 

The same study that can be used to derive this NOAEL also suggests that pre-exposure to viral

infections of the respiratory tract may increase the risk of serious adverse effects, including

mortality in experimental mammals. While the dose-response relationship can be defined for a

specific exposure scenario—i.e., exposure of mice to 4% of the LD50 of an influenza 

virus—these data are not directly or quantitatively applicable to the human health risk 

assessment. 

Risk Characterization – The risk characterization regarding exposure to B.t.k. and its 

formulations is generally consistent with that of the previous USDA risk assessment as well as

more recent risk assessments conducted by the U.S. EPA and the World Health Organization:

B.t.k. and its formulations are likely to cause irritation to the skin, eyes, and respiratory tract;

however, serious adverse health effects are implausible. Nonetheless, more recent information 

alters the approach taken to quantifying the risk of exposure-related irritant effects and more

serious health effects, thereby affecting the risk characterization. Unlike the previous USDA

risk assessment, there is no attempt to quantify the risk of irritant effects. This approach is 

taken because the threshold for these effects cannot be determined. At application rates

similar to those conducted by USDA in programs to control or eradicate the gypsy moth, some

members of the general public as well as workers are likely to experience throat irritation,

which is the best documented effect in the B.t.k. literature on human health effects. 

Nonetheless, dermal and ocular irritation are also likely effects, although perhaps only at the

extreme upper levels of exposure. 

B.t.k. applications to control or eradicate the gypsy moth are not expected to cause serious

adverse health effects in humans. At the extreme upper range of exposure in ground workers,

exposure levels are estimated to be below the functional human NOAEL for serious effects by

a factor of 25. For members of the general public, exposure levels are estimated to be below

the functional human NOAEL by factors of about 28,000 to 4,000,000 [4 million]. This 

assessment is based on reasonably good monitoring data, conservative exposure assumptions,

and an aggressive and protective use of the available toxicity data. Based on these data, it is 

not likely that overt signs of toxicity will be observed in any group— ground workers, aerial

workers, or members of the general public—exposed to B.t.k. as the result of gypsy moth 

control and eradication programs conducted by the USDA. 

xii 
There is no documented evidence of a subgroup of individuals who are more sensitive than

most members of the general public to B.t.k. formulations. According to a recent epidemiology

study, asthmatics are not likely to be adversely affected by aerial applications of B.t.k. The 

literature on B.t.k. includes one anecdotal claim of a severe allergy to a carbohydrate in a B.t.k. 

formulation; however, neither the claim nor observations of similar effects are substantiated in 

the available published epidemiology studies. On the other hand, B.t.k. formulations are 

complex mixtures, and the possibility that individuals may be allergic to some of the

components in the formulations is acknowledged by a state health service. 

Pre-treatment with an influenza virus substantially increased morality in mice exposed to

various doses of B.t.k. This effect raises concern about the susceptibility of individuals who

have influenza or other viral respiratory infections to severe adverse responses to B.t.k. 

exposure. The viral enhancement of bacterial infections is not uncommon and the 

enhancement of B.t.k. toxicity by a viral infection is, in some respects, not surprising. The 

relevance of this observation to public health cannot be assessed well at this time. No such 

effects are reported in the epidemiology studies conducted to date. It is, however, not clear that 

the epidemiology studies would detect such an effect or that such an effect is plausible under

the anticipated exposure levels (typical or extreme) used in programs to control the gypsy

moth. The viral enhancement of B.t.k. toxicity is likely to be an area of further study in the 

coming years. 

ECOLOGICAL RISK ASSESSMENT 

Hazard Identification – The hazard identification for mammals is closely related to the hazard

identification for the human health risk assessment in that both are based, in part, on numerous

standard toxicity studies in experimental mammals. Although B.t.k. may persistent in

mammals for several weeks after exposure, there is little indication that oral or dermal

exposure leads to any serious adverse effects. Most inhalation studies do not suggest a 

potential for adverse effects even at B.t.k. concentrations much greater than those likely to be 

encountered in the environment. The lack of a positive hazard identification is supported by

field studies which demonstrate a lack of adverse effects in populations of mammals after

applications of B.t.k. 

Toxicity studies in birds are limited to standard acute exposures required by U.S. EPA for

product registration. The studies all involve either single-dose gavage administration or five

daily dose gavage administrations, and none of the studies reports signs of toxicity or

pathogenicity at single oral doses up to 3333 mg formulation/kg bw or at multiple oral doses up

to 2857 mg formulation/kg bw. Due to the lack of toxicity of B.t.k. formulations as well as 

other B.t. strains, the U.S. EPA did not require chronic or reproductive toxicity studies in birds. 

This apparent lack of the toxicity is supported by numerous field studies in birds. In one field 

study, a transient decrease in abundance was noted in one species, the spotted towhee (Pipilo 

maculatus). This observation is inconsistent with other field studies on B.t.k., and, according to 

the investigators, may be an artifact of the study design. 

The mechanism of action of B.t.k. in lepidoptera is relatively well characterized. B.t.k. 

vegetative cells produce spores and crystals. After the insect consumes the crystals, toxins are 

formed that attach to the lining of the mid-gut of the insect and rupture the cell walls. The 

B.t.k. spores germinating in the intestinal tract enter the body cavity through the perforations

made by the crystal toxins and replicate causing septicemia and eventually death. While 

various strains of B.t. are often characterized as selective pesticides, B.t.k. is toxic to several 

species of target and non-target lepidoptera. Sensitive non-target lepidoptera include larvae of 

xiii 
the Karner blue butterfly, two species of swallowtail butterflies, a promethea moth, the

cinnabar moth, and various species of Nymphalidae, Lasiocampidae, and Saturniidae. 

While some non-target lepidopteran species appear to be as sensitive as target species to B.t.k., 

most studies indicate that effects in other terrestrial insects are likely to be of minor

significance. There is relatively little information regarding the toxicity of B.t.k. or B.t.k. 

formulations to terrestrial invertebrates other than insects. Some oil-based B.t.k. formulations 

may be toxic to some soil invertebrates; however, the toxicity is attributable to the oil in the

formulation and not to B.t.k. There is no indication that B.t.k. adversely affects terrestrial 

plants or soil microorganisms. 

The U.S. EPA classifies B.t.k. as virtually non-toxic to fish, and this assessment is consistent

with the bulk of experimental studies reporting few adverse effects in fish exposed B.t.k. 

concentrations that exceed environmental concentrations associated with the use of B.t.k. in 

USDA programs. Although there are no data regarding the toxicity of B.t.k. or its formulations 

to amphibians, other strains of B.t. appear to have low toxicity to amphibians. The effects of 

B.t.k. on aquatic invertebrates is examined in standard laboratory studies and in numerous field

studies. At concentrations high enough to cause decreases in dissolved oxygen or increased

biological oxygen demand, B.t.k. may be lethal to certain aquatic invertebrates, like Daphnia 

magna. Most aquatic invertebrates, however, seem relatively tolerant to B.t.k. This assessment 

is supported by several field studies that have failed to note remarkable effects in most species

after exposures that substantially exceed expected environmental concentrations. As with 

effects on terrestrial plants, the toxicity of B.t.k. to aquatic plants has not be tested. 

The U.S. EPA (1998) has raised concerns that some batches of B.t. may contain heat labile 

exotoxins that are toxic to Daphnia. The production of these toxins is an atypical event

thought to be associated with abnormal or poorly controlled production process. The U.S. EPA 

requires manufacturers to submit a daphnid study on each new manufacturing process to

demonstrate that heat labile exotoxin levels are controlled. 

Exposure Assessment – Based on the hazard identification, exposure assessments are

presented for three groups: small mammals, terrestrial insects, and aquatic species. While a 

number of different exposure scenarios could be developed for terrestrial mammals, the only

positive hazard identification for B.t.k. involves inhalation exposures. As in the human health 

risk assessment, inhalation exposures of 100 to 5000 cfu/m3

 are used to assess potential risks of 

serious adverse effects in terrestrial vertebrates. These concentrations are applied to a 20 g

mouse and correspond to inhaled doses of 0.00336 to 0.168 cfu/mouse. While there is no basis 

for asserting that any oral and/or dermal exposures are likely to cause adverse effects in

terrestrial vertebrates, an extremely conservative exposure assessment is developed for

combined oral (water and vegetation) and dermal (direct spray) exposures that yields an

estimated maximum dose of about 184 mg/kg body weight. For terrestrial insects, the toxicity

values used to assess the consequences of observing effects is given in units of BIU/ha. 

Consequently, the exposure assessment for this group is simply the range of application rates

used in USDA programs —i.e., about 49 to 99 BIU/ha. For aquatic organisms, toxicity data are

expressed in several different units such as mg formulation/L, IU/L, and cfu/L. Based on 

application rates used in USDA programs and conservative assumptions concerning the depth

of water over which B.t.k. might be sprayed, concentrations in water would be expected to be at

or below 0.24 mg formulation/L. As discussed in the hazard identification, there is no basis for 

asserting that adverse effects in birds, plants, soil microorganisms, or soil invertebrates other

than insects are of plausible concern. Consequently, explicit exposure assessments are not 

conducted for those groups. 

xiv 
Dose-Response Assessment – The dose-response assessment parallels the exposure 

assessment. Specific dose-response assessments are presented for three groups: small

mammals, terrestrial insects, and aquatic animals. For small mammals, dose-response

assessments are given for inhalation and oral exposure. The risk assessment for inhalation 

exposure is based a mouse study in which mortality increased significantly after intranasal

instillations of B.t.k. A dose of 107 cfu/mouse is taken as the NOAEL and 108 cfu/mouse is 

taken as a frank effect level —a dose associated with 80% mortality. The risk assessment for 

oral exposure, on the other hand, is based on a free-standing NOAEL, which is to say that there

is no evidence that oral exposure levels, however high, will cause adverse effects in mammals

or birds. For this risk assessment, the dose of 8400 mg/kg/day is used as the NOAEL. For 

terrestrial invertebrates, sufficient data are available to estimate dose-response relationships for

sensitive species as well as for relatively tolerant species. Sensitive species, which consist 

entirely of lepidoptera, have an LD50 value of about 21 BIU/ha. Tolerant species, which consist

of some lepidoptera and other kinds of terrestrial insects, have an LD50 of about 590 BIU/ha, 

which is about 28 times greater than the LD50 value for sensitive species. For both sensitive 

and tolerant species, dose-response curves are developed which permit mortality estimates for

any application rate. As with terrestrial insects, dose-response assessments are provided for

tolerant and sensitive species of fish and aquatic invertebrates. Fish appear to be somewhat 

less sensitive than invertebrates to B.t.k.. For tolerant species of fish, the NOEC is taken as 

1000 mg/L, which corresponds to 2.5×1010 cfu/L, and is taken from a study in mosquito fish. 

For sensitive species of fish, the LOEC is based on a trout study in which marginally 7 significant mortality was observed at 1.4 mg/L or about 2.87×10 cfu/L. The most sensitive 

invertebrate species appears to be Daphnia magna, with a chronic NOEC of 0.45 mg/L or 8 6.24×10 cfu/L for reproductive effects and mortality. The NOEC for tolerant species is taken 

as 36 mg/L based on bioassays in mayflies and caddisflies. 

Risk Characterization – Terrestrial insects are the only organisms likely to be adversely 

affected by exposure to B.t.k. or its formulations. Separate dose-response curves can be

generated for both sensitive and tolerant terrestrial insects. At the application rates used to

control gypsy moth populations, mortality rates among sensitive terrestrial insects are likely to

range from approximately 80% to 94% or more. All sensitive terrestrial insects are lepidoptera

and include some species of butterfly, like the endangered Karner blue and some swallowtail

butterflies and promethea moths. For some lepidoptera, sensitivity to B.t.k. is highly dependent 

on developmental stage. This is particularly evident for the cinnabar moth, where late instar

larvae are very sensitive to B.t.k. and early instar larvae are very tolerant to B.t.k. Given the 

mode of action of B.t.k.—i.e., it must be ingested to be highly toxic to the organism— effects

on even the most sensitive species will occur only if exposure coincides with a sensitive larval

stage of development. In tolerant species, including non-lepidopteran insects and certain larval

stages of some lepidoptera, the anticipated mortality rates are much lower (on the order of less

than 1% to about 4%). The risk characterization for terrestrial mammals is unambiguous:

under foreseeable conditions of exposure, adverse effects are unlikely to be observed. 

Similarly, based on a very conservative exposure assessment for aquatic species, effects in fish

and aquatic invertebrates appear to be unlikely. As discussed in the hazard identification, 

effects in birds, plants, soil microorganisms, or soil invertebrates other than insects are not of

plausible concern. Thus, quantitative risk characterizations for these groups are not conducted. 

For oil-based formulations of B.t.k. (or any other pesticide), effects in some soil

invertebrates—i.e., Collembola or earthworms—are plausible. 

xv 
1. INTRODUCTION
 

This document updates the human health and ecological risk assessments on Bacillus 

thuringiensis var. kurstaki (B.t.k.) prepared in 1995 in support of the Final Environmental

Impact Statement (FEIS) for the Cooperative Gypsy Moth Management Program (Durkin et al.

1994; USDA 1995) sponsored by the USDA Forest Service and APHIS. B.t.k. is used in 

USDA Forest Service and APHIS programs to control or eradicate the gypsy moth (Lymantria 

dispar). The updated risk assessments define the environmental consequences of using B.t.k. in 

these programs. 

This is a technical support document and it addresses some specialized technical areas. Thus,

parts of this document may contain information that is difficult for some readers to understand. 

These technical discussions are necessary to support the review of the document by individuals

with specialized training. Nevertheless, an effort is made to ensure that the conclusions 

reached in the document and the bases for these conclusions can be understood by individuals

who do not have specialized training in the chemical and biological sciences. Each major

section of the document starts with an overview section that is intended to summarize the 

technical discussion in a manner that most individuals will understand. In addition, certain 

technical concepts, methods, and terms common to all parts of the risk assessment are

described in plain language in a separate document (SERA 2001). Some of the more 

complicated terms and concepts are defined, as necessary, in the text. 

In the preparation of this risk assessment, literature searches of B.t.k were conducted in the 

open literature using PubMed, TOXLINE, AGRICOLA, as well as the U.S. EPA CBI files. 

The body of literature regarding the environmental fate and toxicology of B.t.k is expansive. 

In addition to the previously prepared risk assessments (Durkin 1994; USDA 1995), there are

several books (Entwistle et al. 1993; Hickle and Fitch 1990; Glare and O’Callaghan 2000) and

a relatively comprehensive review by the World Health Organization (WHO 1999) concerning

the toxicology, environmental fate, and other issues associated with the use of B.t., including

B.t.k. Several other reviews of various topics involving B.t. are published in the open literature

(e.g., Addison 1995; Auckland District Health Board 2002; Drobniewski 1994; McClintock et

al. 1995b; Meadows 1993; Siegel 2001; Swadener 1994). 

Also, numerous studies were submitted to the U.S. EPA/OPP in support of the reregistration of

B.t., and most of these studies are reviewed in U.S. EPA (1998), which summarizes the product

chemistry, mammalian toxicology, and ecotoxicology studies submitted by industry. The U.S. 

EPA Office of Pesticide Programs kindly provided the full text copies of most of these studies

(n=222). The CBI studies were reviewed during the preparation of this risk assessment, and

synopses of the information that can be disclosed from these studies are included in this

document. 

Genetic material from B.t.k. is incorporated into some food crops. In its evaluation of the 

process, the U.S. EPA concluded that although the endotoxin is not toxic to mammals or other

vertebrates, it may be toxic to lepidopteran species (U.S. EPA 2000a) . For the most part, this 

risk assessment does not address the use of B.t.k. toxins in food crops (e.g., Raps et al. 2001;

Wraight et al. 2000); however, certain studies involving transgenic food crops (Fares and ElSayed 1998; Yu et al. 1997) are considered because they are relevant to the hazard

identification for humans and non-target mammalian species. 

While this document discusses the studies used to support the risk assessments, it makes no

attempt to summarize all of the information cited in the existing reviews. This is a general 

1-1
 
approach in all Forest Service risk assessments. For B.t.k. in particular, an attempt to

summarize all of the available data would tend to obscure the key studies which should and do

have an impact on the risk assessment. 

The Forest Service updates their risk assessments periodically and welcomes input from the

general public regarding the selection of studies included in the risk assessment. This input is

helpful, however, only if recommendations for including additional studies specify why the

new or not previously included information is likely to alter the conclusions reached in the risk 

assessments. 

The risk assessment methods used in this document are similar to those used in risk 

assessments previously conducted for the Forest Service as well as risk assessments conducted

by other government agencies. Details regarding the specific methods used to prepare the

human health risk assessment are provided in SERA (2001). This document has four chapters,

including the introduction, program description, risk assessment for human health effects, and

risk assessment for ecological effects or effects on wildlife species. Each of the two risk 

assessment chapters has four major sections, including an identification of the hazards

associated with B.t.k. and its commercial formulations, an assessment of potential exposure to

the product, an assessment of the dose-response relationships, and a characterization of the

risks associated with plausible levels of exposure. These are the basic steps recommended by

the National Research Council of the National Academy of Sciences (NRC 1983) for

conducting and organizing risk assessments. 

Variability can be a dominant factor in any risk assessment. The current risk assessment 

addresses variability as appropriate. Within the context of this risk assessment, variability has

a minimal impact on the human health risk assessment. As discussed in Section 3, the human 

experience with B.t.k. applications allows for a relatively unambiguous assessment of risk. In 

the ecological risk assessment (Section 4), the major source of variability involves differences

among and within groups of organisms. For terrestrial insects which comprise the basic group 

most likely to be affected directly by B.t.k. applications, data are adequate to derive separate

dose-response curves for sensitive and tolerant species and to suggest possible distributions of

tolerance for species with intermediate sensitivity. For other groups, the data are less detailed

but some attempt is made to express differences within groups when appropriate. 

1-2
 
2. PROGRAM DESCRIPTION
 

2.1. Overview 

Bacillus thuringiensis (B.t.) are naturally occurring bacteria that can be found in soil, foliage, 

wildlife, water, and air. All commercial formulations of B.t.k. used by the USDA contain the 

HD-1 strain. Historically, each of the producers of B.t.k. formulations maintained separate 

stock strains. Based on an analysis of cellular fatty acids in various commercial and standard 

cultures of B.t.k., it appears that B.t.k. strain HD-1 may actually be a set of related strains or 

sub-strains. Ten different formulations of B.t.k. are used in USDA programs and all are 

supplied by Valent USA Corp or subsidiaries. Typical application rates for B.t.k. range from

24 BIU/acre to more than 36 BIU/acre. The range of application rates used in this risk

assessment is 20 to 40 BIU/acre, which corresponds to approximately 49 to 99 BIU/ha. Since 

any preparation of bacteria has the potential for contamination with other possibly pathogenic

microorganisms, U.S. EPA requires that spore preparations of B.t. are produced by pure culture

fermentation procedures with adequate quality control measures to detect either contamination

with other microorganisms or changes from the characteristics of the parent B.t. strain. 

Although B.t.k. formulations may be applied by aerial spray or by ground spray, the number of

aerial applications far exceeds the number of ground applications. More than 1 million pounds

of B.t.k. are applied annually in the United States to control the gypsy moth. A total of 

2,743,816 acres were treated with B.t.k. formulations between 1995 and 2002, for an average

annual treatment rate of about 343,000 acres per year. 

2.2. Chemical Description and Commercial Formulations

Bacillus thuringiensis (B.t.) are rod-shaped, gram-positive, spore-forming aerobic bacteria

found in most of the world (Cheon et al. 1997). B.t. was first isolated from diseased silk 

worms in Japan in 1901. In 1915, Berliner isolated B.t. from diseased flour moths. Depending 

on the classification systems used, between 1600 and 40,000 strains of B.t. have been isolated 

(Addison 1995). The vegetative cells are 1 ìm wide, 5 ìm long, and have flagellae, which are 

short hair-like structures used for locomotion. Various strains of B.t., including B.t.k. , are 

ubiquitous in the environment and can be isolated from soil, foliage, wildlife, water, and air

(Damgaard et al. 1997b; Iriarte et al. 1998; Maeda et al. 2000; Martin 1994; Swiecicka et al.

2002). 

B.t.k. was first isolated in France by Kurstak in 1962. A new strain of B.t.k. was identified in 

the pink bollworm and named the HD-1 strain by Dulmage et al. (1971). All commercial 

formulations of B.t.k. used by the USDA contain the HD-1 strain (U.S. Department of

Agriculture, Forest Service 1994a). The HD-1 strain produces the Cry1Ac, Cyr1Aa, Cry2Aa,

and Cyr2Ab delta-endotoxins (Saxena et al. 2002) as well as chitinase (Wiwat et al. 2000). 

Different serotypes of B.t.k. , in addition to HD-1, have been identified (Lee et al. 2001; Li et 

al. 2002). 

Some strains of B.t. contain the beta-exotoxin, which is mutagenic in mammals (Meretoja et al. 

1977). Such strains are not permitted commercial formulations of B.t.k. that are sold in 

Canada or the United States (British Columbia Ministry of Health 1992, U.S. EPA 1988b). 

Batches of commercial B.t.k. are assayed for beta-toxins to ensure that the commercial batches 

do not contain the beta-exotoxin (Chen et al. 1990k; Chen et al. 1990l; Isaacson 1991b). 

Historically, each of the producers of B.t.k. formulations maintained separate stock strains (e.g.,

Smith and Regan 1990k; Smith and Regan 1990m; Smith and Regan 1990n). The U.S. EPA 

(1998, pp. 3-4) RED on B.t. designates eight different strains of B.t.k. The identity of

commercial strains is based on flagella antigen serotyping (Chen and Macuga 1990o; Chen and

Macuga 1990p; Chen and Macuga 1990q), endotoxin characteristics (Chen and Macuga 1990r; 

2-1
 
Chen and Macuga 1990s; Chen and Macuga 1990t; Fitch et al. 1990; Swysen and Hoogkamer

1991) and differential sensitivity to antibiotics (Smith and Regan 1989d; Smith and Regan

1989e; Smith and Regan 1989f). 

Analysis of cellular fatty acids in various commercial and standard cultures of B.t.k., suggests 

that B.t.k. strain HD-1 may actually be a set of related strains or sub-strains (Siegel et al. 

2000). The U.S. EPA (1998) discontinued the grouping of isolates under subspecies names

because the genetic material for delta endotoxins resides in plasmids that can be transferred

from one isolate to another. 

As discussed in Section 4, there is concern that heat stable toxins may occur in some batches of

B.t.k. Most B.t.k. toxins are heat labile—i.e., the insecticidal/toxic activity of the toxins are

destroyed by autoclaving (e.g., Chen et al. 1990h; Chen et al. 1990i; Chen et al. 1990j). 

Table 2-1 provides a list of the specific B.t.k. formulations registered for control of the gypsy 

moth in forestry applications. Typically, the potency of commercial formulations of B.t.k. is 

expressed as BIU/gallon of formulated product or BIU/pound of formulated product. The term 

BIU is an acronym for billions of international units. This potency is measured in a bioassay 

using the cabbage looper (Dulmage et al. 1971). During production and formulation, each 

commercial batch of B.t.k. is used in the bioassay to determine the LC50 for the test insect, 

expressed as mg product/kg diet. The potency of the batch is then adjusted to the nominal

requirement, as specified for the various formulations listed in Table 2-1. Hence, the use of 

BIU/acre to express an application rate is meaningful in terms of insecticidal efficacy,

assuming that toxic potency to the gypsy moth is related to the toxic potency of B.t.k. to the test 

species used in the bioassay of the formulation. The potency of B.t.k. formulations varies from 

about 14 to about 48 BIU/lb formulated product. The label for Foray 48F specifies potency in

units of Forestry Toxic Equivalents [FTUs]. FTU is a measure of potency similar to BIU

except that the bioassay is based on the gypsy moth rather than the cabbage looper. This 

approach is taken because some formulations such as Foray 48F contain different ratios of

crystals that are more effective against forestry pests (i.e., the gypsy moth and tussock moth)

rather than agricultural pests (e.g., the cabbage looper). Typical application rates for B.t.k. 

expressed in units of BIU range from 24 to more than 36 BIU/acre (USDA Forest Service.

1999). The range of application rates used in this risk assessment is 20 to 40 BIU/acre, which

is equivalent to about 49 to 99 BIU/ha [i.e., 2.471 acres per hectare]. 

As indicated in Table 2-1, the commercial formulations of B.t.k. contain between 3.5% and 

10.3% protein toxins—i.e., the delta-endotoxin. The remainder of the formulations consists of 

materials that are classified as inerts. The inerts in B.t.k. formulations are discussed in Section 

3.1.15 of this risk assessment. 

The chemical and biological variability of B.t.k. formulations is not well characterized. One 

index of variability, however, is the number of viable spores in the formulation. Because the 

viable spores, together with the crystalline toxins, are agents that exert a toxic effect on the

gypsy moth, there are some data regarding the number of spores in various formulations. For 

Foray 48B, microbial analyses of individual batches over a 2-year period indicate that the

number of spores per unit of weight of the formulation can vary by a factor of 50 (Overholt

1994). 

Any preparation of bacteria has a potential for contamination with other possibly pathogenic

microorganisms, and this concern must be addressed by proper quality control procedures

(Bernhard and Utz 1993). Between 1985 and 1987, random samples of B.t.k. purchased by the

various states or provinces were found to contain various bacterial contaminants, although none 

2-2
 
were considered pathogenic. In response to the concerns raised by this contamination,

manufacturers took steps in 1988 to ensure that each batch of B.t.k. is free of detectable levels 

of contaminants. Since 1988, no substantial levels of bacterial or yeast contaminants were 

found in B.t.k. samples (Reardon et al. 1994). As part of an epidemiology study conducted by

Noble et al. (1992), Foray 48B samples were tested and found to contain no other bacteria. 

U.S. EPA (1988b) requires that spore preparations of B.t. are produced by pure culture

fermentation procedures with adequate quality control measures to detect either contamination

with other microorganisms or changes from the characteristics of the parent B.t. strain. In 

addition, prior to final formulation, each lot must be tested by subcutaneous injection of at least

1 million spores into at least five mice. 

2.3. Use Statistics 

Although B.t.k. formulations may be applied by aerial spray or by ground spray, the number of

aerial applications far exceeds the number of ground applications. More than 1 million pounds

of B.t.k. are applied annually in the United States to control the gypsy moth (Green et al. 1990). 

 As indicated in Table 2-2, a total of 2,743,816 acres were treated with B.t.k. formulations 

between 1995 and 2002, for an average annual treatment rate of about 343,000 acres per year. 

In order to minimize the ecological effects and human health effects of gypsy moth

infestations, the USDA adopted various intervention strategies that are roughly categorized as

suppression, eradication, and slow the spread (Liebhold and McManus 1999). Suppression

efforts are conducted by the USDA Forest Service in areas of well established gypsy moth

infestations to combat or interdict periodic gypsy moth population outbreaks. Eradication 

efforts are conducted by USDA/APHIS to completely eliminate gypsy moth populations in

areas where new populations of the gypsy moth are found. Slow the spread, as the name

implies, is a program to reduce the expansion of gypsy moth populations from areas of

established populations to adjacent non-infested areas. 

2-3
 
3. Human Health Risk Assessment 

3.1. HAZARD IDENTIFICATION 

3.1.1. Overview 

Most risk assessments for chemical and biological agents are based on relatively standard

toxicity studies in experimental mammals. B.t.k., however, is different in that several 

epidemiology studies provide useful information regarding the plausibility of observing human

health effects after B.t.k. applications that are identical or closely related to applications used in

USDA programs to control the gypsy moth. The results of standard toxicity studies on B.t.k. 

and its formulations are used to supplement information provided by epidemiology studies. 

In humans, irritation of the eyes, skin, and respiratory tract are effects that might be associated

with exposure to B.t.k. and its commercial formulations. These irritant effects are reported in

experimental animal studies as well as in epidemiology studies and case reports. The 

plausibility of such effects resulting from the use of B.t.k in USDA programs is considered 

further in the risk characterization (Section 3.4). Other more serious signs of toxicity are not 

likely to occur as a result of human exposure to B.t.k. Specifically, there is little indication that 

B.t.k. will be associated with pathogenic effects in humans and essentially no indication of

endocrine disruption or reproductive effects in humans after exposure to B.t.k. Carcinogenic 

and mutagenic effects are not likely to be associated with exposure to B.t.k. or B.t.k. 

formulations. The potential for allergenicity is somewhat more difficult to assess in light of the

reported incidents of potential skin and systemic sensitization and antibody induction in some

individuals after exposure to B.t.k. formulations. 

3.1.2. Epidemiology Studies

Epidemiology studies involve observations on human populations to assess whether or not a

particular agent or exposure is associated with one or more effects. Case studies are different 

from epidemiology studies in that they generally involve reports of adverse effects in one or

more individuals associated with a specific incident. Although case reports are discussed in the

various subsections below, this section is restricted to the available epidemiology studies for

which an overview is presented in Table 3-1. Most of the studies discussed compare the 

responses of populations exposed to aerial applications of B.t.k. formulations with responses of

populations in unsprayed areas (e.g., Elliott et al. 1988; Noble et al. 1992; Aer'aqua Medicine 

Ltd. 2001). In one study, responses in a population are compared before and after application 

of a B.t.k. formulation (Petrie et al. 2003). A recent study in British Columbia (Pearce et al.

2002; Valadares de Amorim et al. 2001) concerns individuals in treated and untreated areas but

focuses specifically on children with a history of asthma. Two studies involve workers, either 

individuals applying a B.t.k. formulation (Cook 1994; Noble et al. 1992) or workers harvesting 

crops that were treated with B.t.k. (Bernstein et al. 1999). This section focuses on a description 

of the individual studies. In the following subsections, this information is used in conjunction

with the case studies and toxicology data in mammals to document the assessment of plausible

effects. 

The first substantial epidemiology study of B.t.k. applications was conducted in Oregon as part

of a program to control a gypsy moth infestation (Elliott 1986; Elliott et al. 1988; Green et al.

1990). In the Oregon program, spray operations were conducted in April, May, and June of

1985 and 1986. B.t.k. was applied to more than 250,000 acres in 1985 and 270,000 acres in 

1986. The B.t.k. was sprayed from helicopters in three separate applications (approximately 7

to10 days apart) over forest, rural, and urban areas. All spraying was conducted between

daybreak and approximately 10:00 a.m. (Elliott et al. 1988). None of the publications on the 

Oregon Program reports the nominal application rate. According to the Oregon Department of 

3-1
 
Agriculture, the application rate was 16 BIU/acre of a Dipel formulation. The health 

surveillance activities that accompanied the Oregon spray program are reported by Green et al.

(1990). The total population of Lane County at the time of the study was 260,000. The 1985 

spray covered an area with a population of approximately 80,000; the 1986 spray covered an

area with a population of approximately 40,000. A surveillance program was established

involving the four largest clinical laboratories in the area, three of which were associated with

hospitals and one of which was an outpatient facility. All clinical cultures that were positive 

for any Bacillus species were subcultured, and the presence of B.t.k. in the subcultures was 

determined. As a control, the same procedure was followed for an unsprayed community

approximately 60 miles from the spray area. No B.t.k. positive samples (n=7) were identified 

from the unsprayed community. In the samples from Lane County, a total of 55 B.t.k. positive

cultures were found over the 2-year study period, 52 of which were associated with incidental

contamination. Two of the three remaining samples may have been the result of 

contamination. The third sample was from an abscess in an IV drug user and “..., B.t. could 

have been responsible for this localized infection, but it could also have been a skin or wound

contaminant, or it could have colonized an abscess caused by another organism.” (Green et al. 

1990, p. 851). 

Another relatively large epidemiology study involving applications of B.t.k. formulations to 

control gypsy moth populations was conducted somewhat later in British Columbia (Bell 1994;

Cook 1994; Noble et al. 1992). The aerial applications were conducted over a period of

approximately 10 weeks, April 18 to June 30, 1992, at a rate of 50 BIU/ha or 20.2 BIU/acre (50

BIU/hectare ÷ 2.471 acres/hectare). According to records kept by a selected group of family

practice physicians, there were no detectable effects of exposure among members of the general

public (Noble et al. 1992). The records of 1140 physicians' office visits were reviewed. Of 

these, 675 were classified as clearly unrelated to symptoms that might be associated with the

spraying. The remaining records involved reports of allergies, asthma, rhinitis, conjunctivitis,

infections of the ear, sinus, or respiratory tract, and skin rashes. Although the available data 

did not permit an assessment of each individual's exposure to B.t.k., available information on 

postal zones for each individual's residence suggested that the numbers of these complaints

were evenly divided between individuals living inside and outside of the spray area. In 

addition, 3500 records of admissions to hospital emergency departments were reviewed. In no 

case was B.t.k. implicated as an agent causing any disease or clinical complaint. 

An analysis of all Bacillus isolates from all the hospitals and laboratories in the study area

indicated that many people were exposed to B.t.k.; however, in all cases, chromatography of 

cellular fatty acids indicated that the B.t.k. recovered from these sources was different from that 

used in the aerial spray (Noble 1994). Of 10 different vegetable samples assayed for B.t.k., five 

were positive during the spray period. As with the B.t.k. recovered from human samples, the 

B.t.k. in the vegetable samples was different from the B.t.k. used in the aerial spray. This 

indicates that oral exposure to B.t.k. was common in this area but that this exposure was not 

attributable to the aerial spraying. As discussed in the program description (see Section 2), 

B.t.k. is commonly found in nature, and widespread incidental exposure to B.t.k. is to be 

expected. In no case was B.t.k. the agent causing an infection (Noble et al. 1992). When B.t.k. 

was recovered in stool samples, the medical histories did not suggest that the B.t.k. was 

associated with signs or symptoms of food poisoning or a disease with watery diarrhea similar

to or suggestive of Bacillus cereus. 

Some ground workers from the British Columbia study involved in the application of B.t.k. 

remained culture positive for long periods of time. Of 115 workers exposed to B.t.k. and 

available for follow-up studies, 15 yielded positive B.t.k. cultures from nose swabs 30 to 60 

days after exposure. Five were positive at 120 days after exposure. No positive cultures were 

3-2
 
identified after 140 days from the termination of exposure. Signs of respiratory or nasal 

infections and other health effects attributed to B.t.k. were not observed in any of the workers at 

any time (Cook 1994). 

Similar results are reported by Bernstien et al. (1999) who studied various groups of workers

involved in harvesting crops treated with Javelin, an agricultural formulation of B.t.k. that is 

not used in USDA programs. In this study, various crops (i.e., celery, parsley, cabbage, kale,

spinach, and strawberries) were treated with the B.t.k. formulation at an unspecified application 

rate. The product label for Javlin (www.greenbook.net), indicates that the formulation is 

typically applied at a rate of about 0.12 to 1.5 lbs/acre. Since Javelin contains 17 BIU/lb, the

likely rate used in these studies ranges from 2 to 25.5 BIU/acre. 

The Berstien et al. (1999) study consisted of a longitudinal, follow-up investigation of 48

(46M, 2F) workers who were involved in picking Bt-sprayed crops (celery, parsley, cabbage,

kale, spinach, strawberries) and who were tested during 4 visits: Visit 1(N=48, baseline 1,

classified as Low for exposure), visit 2 (N=32, baseline 2, just prior to Bt-spraying, classified 

as Low for exposure), visit 3 (N=32, one month after Bt-spraying, classified as High for 

exposure) and visit 4 (N=20, 4 months after Bt-spraying, classified as High for exposure). Two 

additional groups were included: Group 2, Low (N=44) who handled a crop (onions) not Btsprayed and located 3 miles away from Bt-sprayed fields; and a Group 3 Medium (N=34), who

washed and packed Bt-sprayed vegetables. Tests included a clinical evaluation for the presence 

of allergy or atopy, skin-prick tests to B.t.k. and non-B.t.k. (control) extracts, blood testing for

IgE and IgG antibodies specific to a) Javelin water-soluble pesticide extracts (J-WS); b)

Javelin-mercaptoethanol-sodium dodecyl sulfate (J-ME-SDS); Javelin proteinase K spore

extracts (J-PK); and Javelin-associated pro-delta-endotoxin (J-PROTOX), and nasal and mouth

lavages for bacterial counts. As is the case with the study by Cook (1994), nasal cultures were 

positive for B.t.k. in 66% of the high exposure workers 1 month after exposure. Positive B.t.k. 

nasal cultures were also noted in other groups and a statistically significant (p<0.05)

association was noted with respect to the qualitative exposure groups. While the atopic status

was similar across all groups of workers, Bernstien et al. (1999) classify 3 of 9 workers who

handled B.t.k.-treated vegetables (parsley, spinach or celery) reporting clinically defined skin

manifestations due to irritant/contact dermatitis of the forearms after contact at work with the

vegetables. It is not clear, however, whether these were incidences of contact dermatitis due to 

B.t.k. exposure or whether they reflect skin contact sensitivities to the vegetables alone. 

Thirteen of the 32 Group 1workers (~40%) who were tested on two occasions (baseline and 1

month after spraying) converted from skin-prick negative (baseline) to skin-prick positive

while 3 of 4 workers who were positive at baseline remained positive. Similarly, of the 20

workers who were serially (longitudinal study) tested on all three visits (baseline, and at 1 and

4 months after spraying), 13 (65%) converted from negative to positive reactions, whereas skin

test conversions from positive to negative occurred in two workers. Thus, the number of 

positive skin-prick tests to both J-WS and J-ME-SDS extracts but not to J-PK and J-PROTOX

increased 1 month after exposure and persisted for 4 months after exposure to Javelin spray.

Taken together these studies indicate that while a small number of workers were sensitized to

B.t.k. prior exposure, de novo sensitization occurred in a significant number of workers

following exposure to an aerial spray of B.t.k. formulations. 

Data on the development of IgE and IgG antibodies specific to various B.t.k.-related antigens

are less clear since these data suffer from a significant non-random loss of sera which were not

available for testing at various points of the study. This is especially true for Group 1, visit 3 at

4 months after spraying in which the number of sera tested dropped from 22 to 8 for IgE and to

6 for IgG. Therefore, the results presented in Bernstien et al. (1999, Table 5, page 579) should

be interpreted with caution. It is evident that in the longitudinal study of Group 1, the number 

3-3
 
of IgE-positive sera to J-WS increased significantly after exposure compared to baseline values

(p<0.05). The cross-sectional study in which Group 1is compared to Groups 2 and 3, indicated

that the incidence of IgE-positive sera in Group 1 was significantly higher from that in Groups

2 and 3 for both the J-WS and J-ME-SDS antigens while results with BtkVeg and BtaVeg

antigens were not significantly different among the 3 Groups. Of significance to this review is

the observation that the sera of 10 workers tested at pre-exposure and at 4 months after

exposure showed a significant increase in IgE-specific titres (prior exposure OD, 0.08 ±0.01

SEM; post-exposure: mean OD, 0.22 ±0.07 SEM, compared to 14 non-exposed urban controls;

mean OD 0.12 ±0.01 SEM). This clearly reflects an anamnestic response – i.e., a late response

to antigen. In contrast, data on the IgG response indicated that the incidence of IgG-positive

sera from Group 1 workers was high at baseline and remained high in all subsequent visits. In

the cross-sectional study of all exposure groups the incidence of IgG-positive titres specific for

J-WS was significantly higher compared to Group 2 (control) whereas the incidence of IgGpositive titres specific for J-ME-SDS was significantly higher compared to Groups 2 and 3.

These data suggest that workers in Group 1 may have been exposed previously to B.t.k. which 

resulted in a substantial number of these producing IgG antibodies to a variety of B.t.k 

components and that a further increase in antigen-specific IgG antibodies upon re-exposure 

was minimal. Thus, it is clear from this study that exposure to B.t.k. may result in sensitization

of workers as indicated by the increase in IgE titres following exposure. It is less clear, 

however, whether the presence of IgE antibodies would result in clinical manifestations of

allergy. From the data presented in the Bernstein et al. (1999) study it is evident that an

increase in IgE titers from 0.08 to 0.22 occurred in pre- to post-exposure workers without any

clinically defined exposure-associated manifestations of allergy. The possibility exists that

levels of IgE antibodies may increase upon repeated exposures. 

However, as has been observed in the Laferriere et al. (1987) study, antibody titres are reduced

rapidly after exposure has ceased and the probability that this would result in clinically defined

allergenicity in these workers would be low. This study included workers who took part in the

Quebec Ministry of Energy and Resources (M.E.R.) spraying program which lasted for two

years (May 1994 – June 1995). Sera from 112 workers (manual/technical laborers) were tested 

for antibody to B.t.k. vegetative cells or to spores or to a spore-crystals mixture. This study’s

results should be interpreted with caution since several sera are missing throughout the testing

period, and the class of B.t.k-antibodies – i.e. reaginic (IgE) or IgG – is not reported. A small 

number (5/112 or 5%) of workers who were tested in May 1994 (start of the spraying) and in

June 1994 (middle of the activity) were reported to be positive for antibodies to vegetative cells

by June 1994. Of the 5 positive subjects, the titre in worker #12 in June was the same as that in

May, in workers #23 and #29 doubled in June over that in May, and in workers #16 and 24

titers in June were 1/80 and 1/160 respectively but for these workers titres were not available

for May. Weak titres of 1/20 to spores and spores-crystals mixture were recorded only in

worker #29 by June but sera were not analyzed in May for this subject. Three of these workers 

(#12, 16 and 23) were followed up during the next year’s activity (sera were collected in May,

July and September 1995). Workers # 12 and 23 showed an increase in titres to vegetative

cells by July, while the titre to vegetative cells in worker #16 was higher in May compared to

July. The titres in all three workers decreased by September. Worker #16 who was negative in

June 1984 to spores-crystals antigens became weakly positive to the same antigens by July

1985 and remained positive in September 1985. Worker #19, who was not tested in 1984, had a

titre of 1/320 by May 1985 and was reduced by September 1985. Serum for July 1985 was not

available. Five additional workers (technicians) who were tested in 1985 were negative for

antibodies to vegetative cells and spores. These, however, were weakly positive (titre of 1/20)

in May to the spores-crystals mixture. In June 1986 (approximately 1 year after exposure), sera

from three manual laborers who had strongly reacted in the 1985, were re-tested and found to 

3-4
 
be negative for all three antigens. This study did not report any exposure-related clinical 

manifestations in these workers. Collectively, these data suggest that a small number of 

workers become sensitized to B.t.k. constituents and that upon re-exposure the antibody levels

increase transiently, decrease within a month, and are undetectable after one year. 

An epidemiology study specifically designed to assess potential effects of B.t.k. exposure on

children with asthma was conducted in Vancouver Island, British Columbia (Pearce et al. 

2002). In this study, 29 children with asthma were identified in the area to be treated and were

matched to 29 children with asthma outside of the spray area. Endpoints examined included

recorded symptoms and peak expiratory flow rates. The spray zone and no spray zone were 

separated by 1 kilometer. Exposures were assessed by Kromecote cards, air concentrations of 

B.t., and nasal swabs. The treated area received three sprays of Foray 48B at a rate of 4 L/ha. 

This is equivalent to approximately 8.452 pints per 2.471 acres or 3.4 pints/acre, in the midrange of the application rate used in Forest Service programs—i.e., 1.3 to 6.7 pints/acre (Table

2-1). Three separate applications were made at 10-day intervals. There were no apparent

differences between the children in treated and untreated areas with regard to asthma symptoms

or peak respiratory flow rates. It is noteworthy that children in the “non-treated” areas did

receive some level of exposure to B.t.k. based on Kromecote cards (78% positive in treated area

and 9% positive in untreated area) as well as positive cultures from nasal swabs. It is also 

interesting that five nasal swabs were positive for B.t.k. prior to any spray. The average 

3 concentration of B.t.k. in the spray zone was 739 cfu/m during spraying. Monitoring data 

regarding B.t.k. concentrations in air are reported also by Teschke et al. (2001). Although it

appears that both groups of children were exposed to B.t.k., there was an apparent lack of

increased symptoms in either group. Consequently, the study by Pearce et al. (2002) seems to

demonstrate that adverse effects were not associated with the B.t.k. spray. 

Another large epidemiology study conducted in New Zealand (Aer’aqua Medicine Ltd. 2001).

This study involves a program in which Foray 48B was sprayed for the control of the whitespotted tussock moth in two regions of New Zealand during 1996 and 1997. The total exposed 

population was comprised of approximately 88,000 individuals. During the spray program,

self-reports of adverse reactions were recorded and sentinel physicians were actively used to

assess changes in disease pattern. After the spray program, records of reported diseases were

reviewed and the incidence of birth outcomes were analyzed. No effects were noted based on 

reported cases of anaphylaxis from sentinel physicians, incidences of birth defects or changes

in birth weight, the incidence of meningococcal disease, or reported infections with B.t.k. 

Among 375 self-reported incidents of potential adverse effects, the only notable response was

an increase in respiratory, dermal, and ocular irritation. All applications appear to have been 

made at the rate of 5 L/ha of Foray 48B (Aer’aqua Medicine Ltd. 2001, Appendix 6,

Appendices p. 10), which is equivalent to about 10.6 pints (2.113 pints/L) per 2.471 acres or

4.3 pints Foray 48B per acre. As indicated in Table 2-1, this application rate is within the

upper range of application rates typically used to control gypsy moth infestations—i.e., 1.3 to

6.7 pints/acre. 

Petrie et al. (2003) conducted another epidemiology in New Zealand, which is somewhat

smaller than the study by Aer’aqua Medicine Ltd. (2001) and involves only self-reporting

surveys of symptoms. A major difference in the Petrie et al. (2003) study, however, is that the

investigators surveyed the same individuals both before (n=292) and after (n=181) the

application of Foray 48B. Several of the 25 endpoints surveyed by Petrie et al. (2003) are

classified as statistically significant—i.e., sleep problems, stomach discomfort, irritated throat,

itchy nose, dizziness, diarrhoea, “gas discomfort”, extra heart beats, and difficulty

concentrating. The investigators categorize these effects into three general classes: irritant

effects, gastrointestinal effects, and effects characterized as neuropsychiatric—i.e., sleep 

3-5
 
disorder, difficulty in concentrating, and dizziness. A significant increase was noted in 

participants with a history of hay fever (p=0.02) after spraying compared with those

participants not previously diagnosed with hay fever. There was no significant increase in the 

number of participants with a history of asthma (p=0.14) or other allergies (p=0.22) when

compared with participants without these diagnoses (Petrie et al. 2003, page 4). The increase 

in hay fever could be incidental, since the pollen season in Aukland is from October to

February and this may have influenced upper airway and hay fever symptoms reported by the

participating workers. 

Petrie et al. (2003) recommend caution when interpreting this kind of self-reporting survey

because only about 62% of the individuals in the pre-application survey responded to the postapplication survey, and, in self-reporting studies such as this, individuals who feel they were

adversely affected by exposure are more likely to respond in the post-application survey. Petrie 

et al. (2003) note also that there was no significant change in the frequency of visits to health

care providers after the spray program. In other words, while the subjective reports suggest an

increase in frequency of undesirable effects, the severity of the effects were not sufficient to

cause the individuals to seek medical care. This pattern was also noted in the study by

Aer’aqua Medicine Ltd. (2001) in which most of the individuals reporting adverse effects did

not seek medical attention. 

Although Petrie et al. (2003) do not specify the application rate for Foray 48B , they indicate

that the spray program in Auckland involved the control of the painted apple moth. The risk 

assessment for this program is available from the Auckland District Health Board (2002) and

specifies an application of 5 L per hectare, identical to that used in the white-spotted tussock

moth program in New Zealand (Aer’aqua Medicine Ltd. 2001). The Auckland District Health 

Board (2002) also specifies that the application rate corresponds to 500 mg Foray 48B per m2 

and that as many as 15 applications can be made to a single property, which brings the total

application rate to as much as 75 L per hectare or 7.5 g Foray 48B per m . 2 Petrie et al. (2003) 

do specify that their survey was conducted after three aerial sprays. While it is possilbe that

other pesticides were applied in some areas over the course of this study, no information on

such applications is discussed in Petrie et al. (2003). This study is discussed further in the 

dose-response assessment (Section 3.3.3). 

Blackmore (2003) also compiled a self-reported series of incidents associated with effects in

individuals living in the area studied by Petrie et al. (2003). This compilation appears to be an

advocacy document from an organization called the “Society Targeting Overuse of Pesticides

NZ” and does not attempt to provide any analysis or draw any conclusions on causality. 

Nonetheless, the information presented by Blackmore (2003) is generally consistent with the

analysis presented by Petrie et al. (2003). 

Other epidemiology reports involving exposure to B.t.k. are much less detailed, but they 

generally support those described above. In a study in which B.t.k. 3a3b was applied at a rate 

6 6 of 22 @ 10 to 25 @ 10 IU per hectare to control the spruce budworm, no medical problems were 

detected in a survey conducted among B.t.k. workers, 80 volunteers living in the treated area,

and 80 controls living in an untreated area (Valero and Letarte 1989). Industrial reports also 

indicate that B.t.k. can be cultured from various superficial sites on exposed humans and that 

antibodies to B.t.k. are greater in individuals in areas sprayed with B.t.k. than in individuals in 

untreated areas (Abbott Labs 1992). No illnesses or infections attributed to B.t.k. were noted. 

The medical records of workers exposed to B.t.k. contained no references to ocular infection, 

soft tissue infection, or chronic respiratory infection attributable to B.t.k. (Abbott Labs 1992). 

3-6
 
3.1.3. Mechanism of Action (Persistence and Pathogenicity)


While the mechanism of action of B.t.k. and other strains of B.t. is understood relatively well in
 

target species (Section 4.1), there is little indication that B.t.k. or several other insecticidal
 

strains of B.t. have any specific mechanism of action in humans or other vertebrate species


(Addison 1995; Drobniewski 1994; McClintock et al. 1995b; Meadows 1993; Siegel et al.


1987; Siegel 2001). 


Persistence refers to the ability of the organism to survive rather than multiply within a host. 

Several studies indicate that B.t.k. can be recovered from exposed mammals but that recovery

decreases over time after exposure is terminated. B.t.k. and other strains of B.t. can be detected 

in experimental mammals several weeks after exposure (Oshodi and Macnaughtan 1990a,b,c;

Siegel and Shadduck 1990; Tsai et al. 1995). Similarly, several of the epidemiology studies

discussed in Section 3.1.2 (Cook 1994; Noble et al. 1992; Valadares de Amorim et al. 2001)

report the recovery of B.t.k. from nasal swabs for up to several months after exposure—e.g., up

to 120 days after workers applied B.t.k. (Cook 1994; Noble et al. 1992). 

By definition, a pathogen will actively multiply in the host and cause damage. Various 

Bacillus species are clearly pathogenic to mammals (Drobniewski 1994). B.t.k. is clearly

pathogenic to some insects including the gypsy moth but there is very little information

suggesting that B.t.k. is pathogenic in other species. 

Nonetheless, B.t.k. can cause toxicity in mammalian cell cultures in vitro. Tayabali and Seligy

(2000) conducted numerous studies regarding the effects of a commercial formulation of B.t.k. 

(identified as F48B and presumably referring to Foray 48B) and subfractions of the formulation

on human cell cultures. The cell culture endpoints examined were non-specific indices of

cytotoxicity, including loss in bioreduction, morphological changes, changes in cell proteins,

and cell breakdown (cytolysis). In addition, the cytotoxic effects of B.t.k. were compared to B. 

cereus. In general, the cytotoxic effects of B.t.k. were similar to those of B. cereus and could 

be blocked by antibiotics. In terms of the potential adverse human health effects in vivo, the 

authors note that “... a sustained infection would be needed to generate sufficient amounts of

vegetative cells and their cytolytic exoproducts”. 

The suggestion that B.t.k. may be pathogenic to humans (or other vertebrates) is limited to only 

one published study. Samples and Buettner (1983a,b) report that a farmer splashed a

commercial formulation of B.t.k. (DiPel solution) in his right eye, causing eye irritation. 

Irrigation of the eye and application of an antibiotic ointment were ineffective in relieving the 

symptoms. Four days after the accident, the farmer was treated with 0.1% ophthalmic solution

of dexamethasone, a corticosteroid given to relieve the irritation. A corneal ulcer was observed 

10 days after the accident. The farmer was then treated with subconjunctival injections of 

antibiotics. B.t.k. was isolated and cultured from the ulcer. The farmer recovered with no 

permanent eye damage. Although this incident might be interpreted as evidence of an eye 

infected with B.t.k., it can also be interpreted as severe eye irritation accompanied by the

recovery of incidental, viable B.t.k. known to have been accidentally introduced into the 

farmer's eye (U.S. EPA 1986b). Other case reports of B.t. pathogenicity in humans involve 

strains other than B.t.k. (Siegel 2001). 

Two studies have suggested that B.t.k. may contain diarrheal enterotoxins similar or identical to 

those in B. cereus (Damgaard 1995; Bishop et al. 1999). Damgaard (1995) used enzyme-linked

immunosorbent analysis (ELISA), a very sensitive analytical method, and did detect

enterotoxigenic activity in B.t.k. strain HD-1 as well as B.t.k. isolated from DiPel, Foray, and 

other formulations. The level of enterotoxigenic activity, however, was substantially less than 

3-7
 
that of B. cereus (positive control): HD-1 11%, Dipel 0.8%, and Foray 3.4% [Damgaard 1995 

Table 1, p. 247]. Also using an immunoassay, Bishop et al. (1999) detected diarrheal 

enterotoxins in B.t.k.. On the other hand, clinical signs of toxicity were not observed in rats at 12 6 oral doses of 10 spores per rat or subcutaneous doses of 10 spores per rat. Fares and ElSayed (1998) report that “B.t.k. HD-14” affects the gastrointestinal tract of mice. As discussed 

by Siegel (2001), however, the identification of HD-14 as B.t.k. may be incorrect. In any event, 

HD-14 is not present in commercial formulations of B.t.k. used in USDA programs to control 

the gypsy moth. 

Some strains of B.t. produce a heat-stable substance commonly referred to as thuringiensin 

(U.S. EPA 1998). The beta-exotoxin is toxic to mammals and other non-target species

(Section 4) and the mode of action involves the inhibition of RNA-polymerase (McClintock et

al. 1995b). B.t.k. and other insecticidal strains of B.t. used in the United States do not contain a 

beta-exotoxin. Other strains of B.t. may contain a heat-labile alpha-exotoxin that causes effects 

similar to B. cereus (McClintock et al. 1995b). 

Strains of B.t. are genetically similar to Bacillus cereus, a known human pathogen (Helgason et 

al. 2000). B. cereus was involved in cases of food-poisoning, causing both diarrhea and

vomiting (Notermans and Batt 1998). Some strains of B.t., not identified as B.t.k. , were 

implicated in episodes of gastroenteritis (Jackson et al. 1995). Furthermore, Vazquez-Padron 

et al. (2000) demonstrated that the Cry1Ac protoxin in B.t.k. strain HD-73 can bind to the 

gastrointestinal tract of mice, while Honda et al. (1991) demonstrated that the hemolysin in

B.t.k. HD-1 is identical to the hemolysin produced by B. cereus. Hemolysin also was identified 

in several other strains of B.t. (Yang et al. 2003). Although Wencheng and Gaixin (1998) did 

not detect hemolysin in B.t.k. HD-1 or HD-73, hemolysin was detected in several other strains 

of B.t. 

There is concern that different strains of B.t. may produce or acquire the capability to produce 

enterotoxins similar to those of B. cereus. Plasmid transfer between different species of B.t. 

under environmentally relevant conditions was demonstrated by Thomas et al. (2000). As 

discussed in the U.S. EPA (1998) RED for B.t. formulations, the transfer of diarrhoeal 

enterotoxins from B. cereus to various strains of B.t. is possible. Because of the relatively low 

incidence of food poisoning associated with B. cereus (i.e., about 0.64% of all cases of food

poisoning), the lack of fatalities in cases of food poisoning associated with B. cereus, and the 

normal measures routinely taken to prevent all causes of food poisoning, the U.S. EPA (1998)

does not consider the potential transfer to diarrhoeal enterotoxins from B. cereus to commercial 

strains of B.t. to be a substantial human health hazard. 

Overall, the evidence for pathogenicity of B.t.k. is extremely limited. While the in vitro studies 

by Tayabali and Seligy (2000) clearly suggest that B.t.k. may damage cells in culture, the only 

in vivo study suggesting a infection in humans (Samples and Buettner 1983a,b) may reflect the

persistence of B.t.k. rather than an infection. The human experience with B.t.k. is substantial, 

and, as summarized in Table 3-1 and discussed in Section 3.1.2, several epidemiology studies

have looked for but failed to find evidence of B.t.k. pathogenicity in humans. 

3.1.4. Acute Oral Toxicity

The U.S. EPA requires standard acute oral toxicity studies for the registration of most

pesticides, including B.t.k. For microbial pesticides, an additional requirement includes assays 

for pathogenicity. The standard assays involving B.t.k. or its formulations are summarized in 

Appendix 1. The interpretation of these studies is reasonably unequivocal, suggesting that

acute oral doses of B.t.k. or its formulations are essentially non-toxic and non-pathogenic (U.S. 

3-8
 
EPA/OPP 1998). The same conclusion was reached by the World Health Organization (WHO 

1999). 

There is one controlled study in humans involving oral exposure to B.t.k.. Fisher and Rosner 

(1959) summarize a study in which 18 volunteers ingested a Thuricide formulation at a rate of

1000 mg per day for 5 days and were exposed to an inhalation dose of 100 mg per day (as a

powder using an inhaler) for 5 days. No signs or symptoms of toxicity were reported and no

changes in standard clinical tests of blood and urine were noted. 

3.1.5. Subchronic or Chronic Systemic Toxic Effects

There are no recent studies regarding the subchronic or chronic toxicity of B.t.k. A standard 

90-day subchronic feeding study and a 2-year chronic rat feeding study were conducted on an

early commercial formulation of B.t.k. at a dose of 8400 mg/kg/day. No effects were seen in 

the 90-day study and the only effect noted in the 2-year study was a decrease in weight gain in

female rats (McClintock et al. 1995b). Hadley et al. (1987) fed sheep (n=6 per group) two 

commercial formulations of B.t.k., a Dipel formulation and Thuricide HP, for 5 months at a

concentration of 500 mg per kg per day (corresponding to approximately 1012 spores per day). 

Loose stool or diarrhea was noted in some of the sheep consuming B.t.k. diets. This effect was 

not observed in untreated or vehicle controls. No other remarkable signs of toxicity were 

apparent. B.t.k. was detected in the rumen, blood, and some tissues of treated sheep. 

3.1.6. Effects on Nervous System 

A neurotoxicant is a chemical that disrupts nerve function, either by interacting with nerves

directly or by interacting with supporting cells in the nervous system (Durkin and Diamond

2002). This definition of neurotoxicant is critical because it distinguishes agents that act 

directly on the nervous system (direct neurotoxicants) from those agents that might produce

neurological effects that are secondary to other forms of toxicity (indirect neurotoxicants).

Virtually any agent (microbial or chemical) will cause signs of neurotoxicity in severely

poisoned animals, and, therefore, can be classified as an indirect neurotoxicant. 

Studies designed specifically to detect impairments in motor, sensory, or cognitive functions in

animals or humans exposed B.t.k. or other strains of B.t. are not reported in the open literature

or in the list of studies submitted to the U.S. EPA to support the registration and re-registration

of B.t. Specifically, the U.S. EPA/OPTS (2003) has standard protocols for several types of 

neurotoxicity studies including a neurotoxicity screening battery (Guideline 870.6200), acute

and 28-day delayed neurotoxicity of organophosphorus substances (Guideline 870.6100). 

Neither of these types of studies was conducted on any strain of B.t. Further, the RED for B.t. 

(U.S. EPA 1998) does not specifically discuss the potential for neurological effects. 

As discussed in Section 3.1.2, a variety of effects characterized as neuropsychiatric—i.e., sleep

disorder, difficulty in concentrating, and dizziness —are reported in the epidemiology study by

Petrie et al. (2003). Consistent with the discussion presented by Petrie et al. (2003), these

effects are most likely to reflect either anxiety or nuisance caused by aerial applications in

general. Consequently, there is no indication that B.t.k. or other strains of B.t. are specific 

neurotoxins in humans or other mammalian species. 

3.1.7. Effects on Immune System

Immunotoxicants are chemical agents that disrupt the function of the immune system. Two 

general types of effects, suppression and enhancement, may be seen and both of these effects

are generally regarded as adverse. Agents that impair immune responses (immune suppression)

enhance susceptibility to infectious diseases or cancer. Enhancement or hyperreactivity can 

give rise to allergy or hypersensitivity, in which the immune system of genetically predisposed 

3-9
 
individuals inappropriately responds to chemical or biological agents (e.g., plant pollen, cat

dander, flour gluten) that pose no threat to other individuals or autoimmunity, in which the 

immune system produces antibodies to self components leading to destruction of the organ or 

tissue involved. 

Neither the published literature nor CBI files provide any clear indication that B.t.k. will cause 

immune suppression. This is consistent with the assessment of the U.S. EPA (1998, p. 13): No 

known toxins or metabolites of Bacillus thuringiensis have been identified to act as endocrine

disrupters or immunotoxicants. Based on studies of B.t.i. (Bacillus thuringiensis israelensis) in

immune suppressed mice, WHO (1999) concluded that individuals with compromised immune

systems are not at special risk from exposure to commercial formulations of B.t. (Section

6.1.7.2 of WHO 1999). 

More recently, Hernandez et al. (2000) noted that a strain of B.t. was associated with increased 

mortality in mice treated with B.t. as well as an influenza virus. The strain of B.t. used by

Hernandez et al. (2000) is identified as serotype 3a3b from Abbott Labs, identical to the active

ingredient in an unspecified pesticide formulation. Serotype 3a3b3c is B.t.k. (Glare and 

O’Callaghan 2000, Table 2.1, p.2.1). Serotype 3a3b has been used to designate B.t.k., but it 

can be applied to HD-1 or HD-73 (Hofte and Whiteley 1989, Table 4, p. 245). Thus, it is 

unclear whether the report from Hernandez et al. (2000) applies to B.t.k. HD-1. Moreover, it is 

not clear whether the mechanism of the increased mortality reflected immune suppression or a

simple addition of stress to the animal. Nonetheless, the increase in mortality was dose-related 

in terms of the B.t. exposure combined with the influenza virus at 4% of the LD50 —i.e., 4 of

2 4 7 20 mice at 10 spores/mouse, 8 of 20 mice at 10 spores/mouse, and 14 of 20 mice at 10

spores/mouse with no mortality observed in the control group (0 of 20 mice) when mice were

treated only with the influenza virus at 4% of the LD50 with no B.t. exposure. In addition, 

weight loss was observed in mice treated with influenza virus at 2% of the LD50 and this 

correlated well with the dose of B.t. 3a3b used to infect the mice suggesting that a low 

innoculum of B.t. was able to complicate an influenza virus respiratory tract infection in mice. 

No mortality was observed in any of the mice but there was a statistically significant decrease 4 7 2 in body weight at 10 spores/mouse and 10 spores/mouse but not at 10 spores/mouse. Also, 

the observed partial protection to mice after use of a thuringolysin-specific monoclonal

antibody suggests that additional B.t.-produced toxins such as phospholipase C and

sphingomyelinase could be involved. Since treatment of mice with the influenza-virus 

infection inhibitor, amantadine, demonstrated that B.t. alone was not pathogenic, the authors

speculated that the influenza virus may have transiently altered the function of the non-specific

defense mechanisms of the respiratory tract – i.e., macrophages and other leukocytes – thus 

rendering the host susceptible to a pulmonary infection by a very low innoculum of B.t. 

As detailed in Section 3.1.2, there is evidence that some workers may become sensitized to

B.t.k (Bernstein et al. 1999; Laferriere et al. 1987). In addition to the possible development of 

sensitivity to B.t.k., Swadener (1994) reports the following incident: 

...during the 1992 Asian gypsy moth spray program in Oregon, a

woman who was exposed to Foray 48B had a preexisting allergy

to a carbohydrate that was present as an inert ingredient.

Within 45 minutes of exposure, the woman suffered from joint

pain and neurological symptoms. (Swadener 1994, p. 16) 

The description of this incident is attributed to a letter, dated August 12, 1992, from the Oregon

Department of Human Resources to Martin Edwards of Novo Nordisk. In itself, this report 

3-10
 
does not provide sufficient information to assess the credibility that the effect was associated

with Foray 48B or to assess the seriousness of the reported effect. Although the Oregon Health 

Services (2003) B.t.k. fact sheet discusses the possibility that individuals may be allergic to

components of the bacterial growth media in B.t.k. formulations, the incident summarized by

Swadener (1994) is not mentioned. 

3.1.8. Effects on Endocrine System

In terms of functional effects that have important public health implications, effects on

endocrine function would be expressed as diminished or abnormal reproductive performance. 

This issue is addressed specifically in the following section (Section 3.1.9). Mechanistic assays

are generally used to assess the potential for direct action on the endocrine system (Durkin and

Diamond 2002). Neither B.t.k. nor any other strain of B.t. was tested for activity as an agonist

or antagonist of the major hormone systems (e.g., estrogen, androgen, thyroid hormone). 

Accordingly, all inferences concerning the potential effect of B.t. on endocrine function must 

be based on inferences from standard toxicity studies. As noted in the previous section, U.S.

EPA (1998) concludes that there is no basis for asserting that strains of B.t. are likely to have 

an impact on the endocrine system. 

3.1.9. Reproductive and Teratogenic Effects

Specific tests regarding the effects of B.t.k. and other strains of B.t. on reproduction and

development were not conducted and effects of that nature are not addressed specifically in the

existing reviews or compendia on B.t.—e.g., Glare and O’Callaghan (2000), U.S. EPA (1998),

WHO (1999). As with effects on the nervous, immune, and endocrine systems, there is no

credible concern that B.t.k. or other strains of B.t. are to cause adverse effects on reproduction 

or developement in humans or other mammals. 

As noted in Section 3.1.3.3, Petrie et al. (2003) surveyed birth outcomes before and after a

Foray 48B spray program and noted no adverse effects. As discussed further in Section 4.1, the 

lack of adverse reproductive effects in mammals is supported in field studies conducted in

areas treated with B.t.k. 

3.1.10. Carcinogenicity and Mutagenicity

While the cancer risks of exposures to chemical carcinogens are relatively well characterized,

carcinogenic and mutagenic effects are not typically associated with bacteria. As reviewed by 

McClintock et al. (1995b), B.t.k. was subject to a 2-year chronic dietary study in rats in which

no effects were noted other than a decrease in weight gain among treated females. This is the 

kind of study typically conducted as an assay for potential carcinogenicity in mammals. 

A formulation of B.t.k. (HD-1) from China was shown to cause a dose-related increase in

chromatid and chromosome breaks in spermatogonia when injected into the abdomen of 5th 

instar grasshoppers (Oxya chinensis) (Ren et al. 2002). As discussed by Ren et al. (2002), this 

study may suggest a mechanism of action in insects. This study, however, does not suggest a 

potential human health risk. 

3.1.11. Irritation (Effects on the Skin and Eyes)

As with acute oral toxicity, the U.S. EPA requires standard assays for dermal and eye irritation,

and these studies are summarized in Appendix 1. While most studies indicate that B.t.k. is not 

a strong irritant to either the eyes or the skin, the study by Bassett and Watson (1999b) is

somewhat unusual in that the erythema appears to be more pronounced than in most of the

other studies. Moreover, in at least one animal, the erythema appears to have progressed rather

than reversed over the 14-day post-observation period. Mild eye irritation is consistently seen 

3-11
 
in studies involving exposure to Dipel (Kuhn 1999b) or Foray (Berg 1991a,b; Berg and Kiehr

1991). 

As discussed further in the dose-response assessment, throat irritation in humans appears to be

a plausible effect based on the epidemiology studies by Cook (1994) and Petrie et al. (2003). 

Furthermore, local inflammatory responses were observed in mice after intranasal instillations

of B.t.k. (Hernandez et al. 2000). 

The epidemiology study by Cook (1994) includes workers involved in both ground and aerial

applications of B.t.k. During the ground application, the commercial formulation of B.t.k., 

diluted with water, was delivered as a high pressure spray from high-lift units. Dilutions 

ranged from an initial 200:1 to 75:1. The decrease in the dilution rate was associated with the 

use of a finer spray. In the last spray cycle, a jet turbine aerosol generator (Rotomister)

mounted on a trailer was used. Two contractor teams, designated A and B, were involved in 

the ground applications. A separate group of workers was involved in monitoring the

effectiveness of the aerial application by the placement of cards used to measure droplet

deposition. These individuals were generally exposed to air-delivered aerosol during the aerial

application and for 2 hours or more after the application. In general, the workers did not wear 

protective equipment (e.g., goggles or face masks). Worker exposure was monitored by 

microbiological air sampling. Symptoms, including transient irritation of the eyes, nose, and

throat, dry skin, and chapped lips, developed in approximately 63% of the workers, but in only

38% of the control group. No days of work loss were attributable to B.t.k. exposure. These 

data are discussed further in the dose-response assessment (Section 3.3). 

Two other incidents involving eye irritation in humans after exposure to B.t.k. were reported in 

the literature (Green et al. 1990; Samples and Buettner 1983). The studies by Samples and

Buettner (1983a,b) regarding the pathogenicity and persistence of B.t.k. is discussed in detail in 

Section 3.1.3. The report by Green et al. (1990) describes an incident in which a worker

involved in the application of B.t.k. splashed the B.t.k. mixture in his face and eyes. The 

worker developed dermatitis, pruritus, burning, swelling, and erythema, with conjunctival

irritation. A culture of the conjunctiva was positive for B.t.k. The worker was treated 

effectively with steroid cream applications to the eyelid and skin. 

Ocular exposure to B.t.k. does not always result in serious eye irritation. Noble (1992) briefly

summarizes an incident in which two individuals on bicycles were accidently sprayed in the

face by ground spray workers. The face and eyes were washed immediately after the incident,

and no residual eye irritation developed in either individual over a 21-day follow-up period. In 

a separate incident, two workers on the ground spray team in the British Columbia study were

accidently sprayed in the face with the B.t.k. formulation. These workers experienced only

slight redness of the eyes for several hours after exposure (Cook 1994). The ground spray

workers in this study reported a higher rate of eye irritation, compared with the control

population (Cook 1994). 

In terms of the weight-of-evidence assessment, there seems to be little doubt that exposures to

B.t.k. can result in irritation of the skin, eyes, and respiratory tract, all of which are

demonstrated in animals studies as well as in epidemiology studies and case reports. Thus, all 

three irritant effects are rated with the highest possible score—i.e., I.A.1.a. As discussed 

further in the dose-response assessment and risk characterization, irritant effects are the most

likely effects to result from general applications of B.t.k. over widespread areas. 

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3.1.12. Systemic Toxic Effects from Parenteral Exposure

Parenteral exposures involve injecting a substance into an animal, usually into a vein (i.v.) or

into the abdominal cavity (i.p.). Several such studies were conducted on B.t.k. or B.t.k. 

formulations and these studies are summarized in Appendix 1. As discussed by McClintock et

al. (1995b), these studies are used primarily as qualitative screening tools to assess

pathogenicity and infectivity. In addition, these studies may be used to assess variations in

toxicity among different commercial batches of B.t.k. formulations (e.g., Vlachos 1991) as well

as differences in toxicity associated with different culture conditions (Siegel 2001). According

to Siegel (2001), these tests may be most relevant to risk characterization in terms of

comparing the toxicity of the microbial agent to known pathogens such as B. anthracis, which 

has an LD50 in mice of about 2.64 spores by intraperitoneal injection. As noted in Appendix 1, 8 little or no mortality was observed in mice at intraperitoneal B.t.k. doses of up to 10 [one

hundred million] cfu. Thus, relative to highly pathogenic bacteria, the apparent acute lethal 

potency of B.t.k. is extremely low. 

3.1.13. Inhalation Exposure

Most of the studies summarized in Appendix 1 are reasonably consistent with the general

assessment regarding the toxicology of B.t.k. formulations: irritant effects but no systemic 

toxic effects or infectivity. Two studies, however, are inconsistent with the other available 

information. In one of these studies, inhalation exposure of rats to very high levels of B.t.k. 

caused piloerection (an atypical condition in which the hair stands erect), lethargy, and frequent

urination during exposure (Holbert 1991). Alopecia (hair loss) was observed in the rats several 

days after exposure. This study involved whole body exposures over a 4-hour period to a level 

of B.t.k. formulation (3.22 mg/L Foray 76B) that caused the rats to become coated with the test

material. The investigators indicated that the hair loss was probably related to B.t.k. exposure.

While the implications for human risk assessment, if any, are unclear, this is an unusual

finding. The reason for the hair loss cannot be determined, and this effect is inconsistent with 

other studies on B.t.k. 

Only two studies (David 1990c; Hernandez et al. 2000) have reported mortality after exposure 

to B.t.k. and both of these studies, while related to inhalation toxicity, involve atypical routes of 

exposure. Intratracheal instillations of bacteria are analogous to inhalation exposures in that

the bacteria is essentially inserted into the lungs. One such study (David 1990c) was conducted 

on a B.t.k. Dipel formulation. As detailed in Appendix 1, toxic responses including death were

observed in treated animals and the time-to-clearance (estimated from linear regression) was

prolonged. Also, Hernandez et al. (2000) assayed the toxicity of B.t.k. after intranasal 

instillations in mice. This method of dosing is also analogous to inhalation exposures in that

the material is deposited in nasal passages and the B.t.k. is gradually transported to the lungs by 2 4 6 8 inhalation. Doses of 10 , 10 , and 10 cfu/mouse caused only local inflamation. A dose of 10 

cfu/mouse resulted in 80% lethality. The relevance of these two studies to the human health 

risk assessment is discussed further in Section 3.3 (Dose-Response Assessment). 

3.1.14. Impurities

Any preparation of bacteria has the potential for contamination with other possibly pathogenic

microorganisms, which presupposes the need for proper quality control procedures (Bernhard

and Utz 1993). Between 1985 and 1987, random samples of B.t.k. purchased by the various

states or provinces were found to contain various bacterial contaminants, although none was

considered pathogenic. In response to the concerns raised by this contamination, manufacturers

took steps in 1988 to ensure that each batch of B.t.k. is free of detectable levels of 

contaminants. Since 1988, no substantial levels of bacterial or yeast contaminants were found 

in B.t.k. samples (Reardon et al. 1994). As part of an epidemiology study conducted by Noble

et al. (1992), Foray 48B samples were tested and found to contain no other bacteria. 

3-13
 
U.S. EPA (1998) requires that spore preparations of B.t. are produced by pure culture

fermentation procedures with adequate quality control measures to detect either contamination

with other microorganisms or changes from the characteristics of the parent B.t. strain. 

3.1.15. Inerts 

Inerts are defined as compounds that do not have a direct toxic effect on the target species.

Nonetheless, some inerts may be toxic to non-target species, including humans. For some 

chemicals, the presence of toxic inerts may be a substantial issue in a risk assessment. The 

minimal testing requirements for compounds that have been used as inerts or adjuvants for

many years is a general problem in many pesticide risk assessments. For new inerts, the U.S. 

EPA does require more extensive testing (Levine 1996). U.S. EPA (2001) proposes to 

discontinue the use of the term inerts for the following reason: 

Many consumers are mislead by the term "inert ingredient",

believing it to mean "harmless." Since neither the federal law 

nor the regulations define the term "inert" on the basis of

toxicity, hazard or risk to humans, non-target species, or the

environment, it should not be assumed that all inert ingredients

are non-toxic. (U.S. EPA 2001). 

Nonetheless, the term inerts, as defined above, is used widely in the literature regarding

pesticides, including the current risk assessment. U.S. EPA (2001) classifies inerts into four

lists: toxic inerts (List 1), potentially toxic inerts (List 2), inerts that cannot be classified

because of limitations in the available data (List 3), and inerts that are nontoxic or generally

recognized as safe (List 4). 

The identity of some inerts in some formulations of B.t.k. are reported in the open literature, 

and this information is summarized in Table 3-2. As indicated in Table 3-2, most inerts 

identified in the open literature are classified as GRAS (generally recognized as safe)

compounds and are approved for use as food additives (Clydesdale 1997). Two of the 

compounds listed in Table 3-2, methyl paraben and polyacrylic acid, are not approved as food

additives and are classified as List 3 inerts in U.S. EPA (2001). Swadener (1994) raises 

concerns about many of the additives in Foray 48B, a B.t.k. formulation used in USDA 

programs, including those approved as food additives, and similar concerns are expressed by

groups opposed to the use of B.t.k. formulations (e.g., http://www.vcn.bc.ca/stop/preface.html).

For example, Swadener (1994) correctly notes that concentrated sodium hydroxide is a severe

corrosive and can be extremely hazardous. This, however, is not germane to the hazard 

identification of Foray 48B or any other B.t.k. formulations. In these formulations, sodium 

hydroxide is used in relatively low concentrations. While the specific amount and function of

sodium hydroxide cannot be publically disclosed, Clydesdale (1997) notes that sodium

hydroxide is commonly used as a pH control agent. In this and other approved uses of sodium

hydroxide as a food additive, sodium hydroxide is not likely to pose any risk whatsoever. In an 

aqueous solution such as a formulation of B.t.k., sodium hydroxide (NaOH) will dissociate to + ­ the sodium cation (Na ) and the hydroxide anion ( OH), both of which are natural and essential - components of all living organisms. Furthermore, Na+

 and OH concentrations are highly

regulated by normal biological processes. 

Much more detailed information regarding the inerts in B.t.k. formulations and the 

manufacturing processes was obtained from the U.S. EPA in the preparation of this risk

assessment (e.g., Berg et al. 1991; Birkhold 1999; Coddens 1990a; Coddens and Copper 1990;

Eyal 1999; Jensen et al. 1990a,b,c,d,e; Hargrove 1990a,b,c; Knoll 1990a; Newton 1999;

Rowell 2000; Sorensen et al. 1990a,b). These studies, which include details regarding the 

3-14
 
product chemistry and manufacturing processes, are protected under FIFRA Section

12(a)(2)(D), therefore, cannot be released to the general public or summarized in any

significant detail. 

As noted in Table 2-1, Valent USA Corporation holds the current registrations for B.t.k. 

formulations. Nonetheless, some information is available in the open literature from previous

registrants—i.e., Novo Nordisk (1993) and Abbott Labs (1992)—and this information remains

relevant to the current risk assessments and can be disclosed. Novo Nordisk (1993) published

a brief summary of the issues associated with the use of inerts in Foray 48B and the proprietary

nature of inerts. Foray 48B is a mixture of B.t.k. and fermentation materials, which comprise 

almost 90% of the product. The added inerts (that is, those other than incidental fermentation

products) include materials to inhibit the growth of bacterial or fungal contaminants. These 

additives are approved for use in foods in the United States and Canada. All of the Novo 

Nordisk inerts are on U.S. EPA List 3 or 4. No volatile solvents are used in Foray 48B. The 

Oregon Department of Human Resources reviewed the complete formulation in Foray 48B and

determined that "... exposure to the ingredients in the Foray 48B formulation are unlikely to

pose a public health threat to populations exposed to the spray in eradication programs"

(Fleming 1993 p.1). More recently, Van Netten et al. (2000) analyzed the volatile components

in Foray 48B and identified numerous organic compounds that are present in trace amounts. 

Many of these compounds are on the U.S. EPA List 3 or List 4. It is unclear which of these 

compounds are specifically added to the formulation (i.e., as inerts) and which compounds are

by-products of the fermentation process used to produce Foray 48B. 

Some additional information is also publically available regarding the manufacturing process

for B.t.k. formulations. B.t.k. formulations are complex chemical mixtures. B.t.k. is cultured in 

large vats that contain, for the most part, water and nutrients. The nutrients consist primarily of 

sugars, starches, proteins, or amino acids. These nutrients are not added as pure and defined

compounds but rather as chemically complex and variable biological materials such as animal

foodstuffs, a variety of flours, yeasts, and molasses. Relatively small quantities of essential

elements, minerals, or salts also may be added to create optimal growth conditions. Adjuvants,

such as antifoaming agents, may also be used at various stages of production to enhance growth

or facilitate the recovery of B.t.k. from the growth media. The other components of the

formulation are mostly water and a complex mixture of culture media and metabolites. The 

composition used by a manufacturer may change over time, as different sources of nutrient

material are used (Bernhard and Utz 1993). 

As detailed further in the dose-response assessments for B.t.k., the presence and identity of 

inerts, adjuvants, and contaminants in B.t.k. formulations has little impact on the dose-response

assessment for potential human health effects (Section 3.3) or ecological effects (Section 4.3). 

In both cases, the available data are much better suited to a “whole mixture” risk assessment 

than a component based risk assessment. Thus, a component based assessment of each inert 

was not conducted because component based assessments for highly complex mixtures

generally are not useful given that the uncertainty of a component based risk assessment

increases as the number of components in a mixture increases (Mumtaz et al. 1994, U.S.

EPA/ORD 2000). As recommended by U.S. EPA/ORD (2000), the risk assessment is based on

the mixtures of concern, which, in this case, are the commercial formulations of B.t.k. The 

limitations and benefits of this approach are discussed further in the risk characterization

(Section 4). 

3-15
 
3.2. EXPOSURE ASSESSMENT 

3.2.1. Overview 

Exposure assessments usually estimate the amount or concentration of an agent to which an

individual or population might be exposed via ingestion, dermal contact, or inhalation. The 

exposure assessments are then compared with toxicity studies based on similar types of

exposures—i.e., the dose-response assessment—and then the risk is quantified. The human 

health risk assessment for B.t.k. is unusual in two respects. First, as discussed in Section 3.1 

(Hazard Identification) and discussed further in Section 3.3 (Dose-Response Assessment), the

most directly relevant data used to characterize risk are based on actual applications of B.t.k. 

formulations where exposure is best characterized as an application rate. Second, the apparent 

lack of a specific mechanism of toxicity for B.t.k. makes selecting the most appropriate 

measure of exposure somewhat arbitrary. 

3.2.2. General Issues 

As discussed in Section 2 and considered further in Section 4.1, the potency of B.t.k. is often 

expressed as BIU or FTU and exposures or application rates are expressed in units of BIU or

FTU per acre. Although these units may be meaningful expressions of exposure for the gypsy

moth, they are not necessarily or even likely to be a meaningful measures of human exposure. 

Toxicity to sensitive insects like the gypsy moth is generally attributed to a combination of the

delta-endotoxin and the spore coat. These two factors probably account for the potency of the

commercial formulations in the bioassays used to determine the BIU/mg of commercial

product. Unlike the gut of the gypsy moth, which has a high pH (that is, the gut is alkaline or

basic) the stomach of most mammals, including humans, has a low pH (that is, the stomach

contents are acidic). Thus, the delta-endotoxin is not toxicologically significant for humans. 

Another commonly used measure of exposure to B.t.k. formulations is colony forming units or 

cfu. When B.t.k. formulations are applied, either by aerial spray or ground spray, one or more

viable spores contained in droplets or particulates is suspended in the air and deposited on

sprayed surfaces. These droplets may be collected, either by air sampling or direct deposition,

onto various types of filters. The filters are then cultured in a nutrient medium under 

conditions conducive to bacterial growth. As the bacteria grow, visible masses of bacteria,

referred to as colonies, appear on the media. In the case of monitoring B.t.k. formulations, 

some of the colonies will be B.t.k. and some colonies will be other endogenous bacteria. 

Microscopic examination, differential culturing, or other methods may be used to determine the

number of colonies that are B.t.k. By this general method, the number of cfu per unit of surface

area or volume of air, depending on the sampling method, may be determined. Each cfu can be 

formed from a droplet or particulate that contains one or more viable spores. Thus, the number 

of cfu per unit of surface area or volume of air does not correspond directly to the number of

viable spores per unit of surface area or volume of air. Dilution methods can be used to 

determine the number of viable spores (Palmgren et al. 1986). 

The significance of cfu as a measure of human exposure is limited. As discussed in Section 

3.1.3, there is little indication that B.t.k. is a human pathogen. Consequently, the number of

viable spores, albeit an important measure of exposure for the gypsy moth, does not appear to

be toxicologically significant to humans. In this respect, cfu like BIU are of limited 

significance. Nonetheless, at least for short-term exposures, cfu can be used as a practical

measure of relative exposure to a B.t.k. formulation. 

For example, assume that an aerial application of a B.t.k. formulation is made and that two air 

samples are taken, one immediately at the spray site and one upwind from the spray site. 

Droplets containing viable spores as well as other components in the B.t.k. formulation are 

3-16
 
sampled at both sites for a fixed period of time. If the sample taken at the spray site yields 200

cfu and the sample upwind yields 20 cfu, it seems clear that the level of human exposure to the

B.t.k. formulation at the upwind site is 10% of that directly beneath the spray. This is, 

however, only a conclusion regarding relative exposure to B.t.k. and implies nothing about its 

toxic potency. Accordingly, the number of cfu is used as a surrogate for exposure to the B.t.k. 

formulation. 

As discussed below in Section 3.2.3 for workers and in Section 3.2.4 for members of the 

3 general public), data are available regarding cfu per volume of air (cfu/m ) during application

and for intervals up to several days after application. For such measurements, it is not 

reasonable to assume that cultured colonies represent exposure to the formulation. Some 

components in the formulation, like water or other volatile materials, will have evaporated,

whereas other nonvolatile materials, like starches, sugars, minerals, proteins, and amino acids,

will have degraded or partitioned from the viable spores. Thus, measurements of cfu taken 

long after the spray application can be interpreted as viable B.t.k. spores that probably adsorbed 

to particulates and were re-suspended. 

Some of the available toxicity studies (Appendix 1) express exposure in units of mg of

formulation per unit of body weight or volume of air, depending on the route of exposure. As 

with cfu, these measures may be applicable to the risk assessment in so far as the anticipated

exposures involve the entire commercial formulation. Exposures of this nature usually occur 

during or immediately after application. 

3.2.3. Workers 

Studies that quantify exposures to workers (and members of the general public) are

summarized in Table 3-3. No new worker exposure studies became available since the 1995 

risk assessment. The two worker studies summarized in Table 3-3, Cook (1994) and Elliott et

al. (1988), are identical to the studies used in the 1995 risk assessment. 

In the study by Elliott et al. (1988), portable sampling pumps with 37-mm (0.8 micron pore

size) cellulose ester membrane filters were used for personal and area air monitoring. Flow 

rates on the sampling pumps ranged from 0.1 to 2.0 L per minute, and the duration of sampling

ranged from 0.25 to 4 hours. All personal monitoring done during 1986 was conducted with a 

flow rate of 0.1 L per minute. Microbial culture and microscopic examinations were used to 

assay for B.t. on the filter media. Initially, all plates (inoculated with membrane filters from the

monitoring pumps) were incubated and inverted for 24 hours at 30EC, after which time 

colonies were counted. The plates were then incubated for 5 more days at room temperature. 

Colonies resembling B.t. were examined microscopically. B.t. was identified by the presence 

of diamond-shaped toxin crystals (Elliott et al. 1988). Measurements made during 1985 could 3 not be expressed as cfu/m because of the extreme numbers of colonies obtained on the culture

plates. The results presented in Table 3-3 are based on 1986 monitoring of personal air. 

Much higher exposure levels are reported in the study by Cook (1994). The substantial 

difference in exposure concentrations may be related to work practices and application

methods,, which include ground applications in the study by Cook (1994) and aerial

applications in the study by Elliott et al. (1988). In general, ground applicators are exposed to

much higher concentrations of pesticides, compared with aerial applicators. 

3-17
 
3.2.4. Members of the General Public 

As noted in Section 2, B.t.k. as well as other strains of B.t. are naturally occurring bacteria. 

B.t.k. HD-1, the same strain used as a pesticide against the gypsy moth, is found in food as well

as other environmental media (Damgaard et al. 1996; Damgaard et al. 1997b; Glare and 

O’Callaghan 2000). 

In terms of exposure levels that can be meaningfully related to USDA program activities, the

most appropriate measure of exposure with respect to workers is summarized in Table 3-3 in

terms of cfu/m .3 The consistency among the various studies is noteworthy. During spray, 

members of the general public may be exposed to concentrations in the range of about 200 to 3 4000 cfu/m , which is about 2 to 3 times lower than of the range of exposure levels for workers 3 involved in aerial applications— i.e., about 400 to 11,000 cfu/m — but very far below the

exposure levels that Cook (1994) observed in ground workers (Table 3-3). 

After spray, B.t.k. and the formulation products will disperse depending on wind speed and 

deposition. Teschke et al. (2001) note that concentrations in outdoor air may decrease by a

factor of about 10 within 5 to 6 hours after spraying but that concentrations in indoor air may

remain higher than those in outdoor air, probably due to decreased dissipation. 

3-18
 
3.3. DOSE-RESPONSE ASSESSMENT 

3.3.1. Overview 

In some respects, the dose-response assessment of B.t.k. is relatively simple. There is no 

information from epidemiology studies or studies in experimental mammals to indicate that

B.t.k. will cause severe adverse health effects in humans under any set of plausible exposure

conditions. This is also the conclusion reached by the U.S. EPA and the World Health

Organization. The only human health effects likely to be observed after exposure to B.t.k. 

involve irritation of the skin, eyes, or respiratory tract. 

Nonetheless, a recent epidemiology study suggests that the irritant effects of B.t.k. may occur

with notable frequency at exposure levels typical of those used in programs to control the gypsy

moth. On the other hand, a worker study indicates that the frequency of observing these irritant

effects does not appear to increase substantially even at extremely high levels of exposure. The 

lack of a strong dose-response relationship is somewhat unusual but is consistent with

experimental data in mammals. 

From recent experimental studies not typically used in a quantitative dose-response assessment,

it is possible to define extremely high exposures for B.t.k. that might pose a serious health

hazard and it is possible to define a NOAEL for such effects that is consistent with the3 available human studies. Specifically, cumulative exposures of up to 1.4×1010 cfu/m × hour

are not likely to result in adverse effects. 

The same study that can be used to derive this NOAEL also suggests that pre-exposure to viral

infections of the respiratory tract may substantially increase the risk of serious adverse effects,

including mortality in experimental mammals. While the dose-response relationship can be

defined for a very specific situation —i.e., exposure of mice to 4% of the LD50 of an influenza 

virus—these data cannot be applied directly and quantitatively to the human health risk 

assessment. 

3.3.2. Existing Guidelines

Dose-response assessments for the systemic toxic effects of most pesticides are based on an

RfD, an estimate of a dose or exposure that is not likely to induce substantial adverse effects in

humans. The RfD, in turn, is typically based on a NOAEL (no observed adverse effect level)

divided by an uncertainty factor. Risk is then characterized as a hazard quotient (HQ) which is

the estimated level of exposure divided by the RfD. If the HQ is below unity—i.e., the

exposure is less than the RfD —there is no credible risk. If the HQ is above unity, risk is

characterized based on dose-response or dose-severity relationships. 

This approach, however, was not taken by the U.S. EPA in the re-registration eligibility

decision (RED) document (U.S. EPA 1998) for B.t. Similarly, the World Health Organization

declined to derive an acceptable daily intake (ADI) value, an estimate that is analogous to the

RfD, for B.t. (WHO 1999). In both cases, the decision not to quantify the dose-response

relationship appears to be based on the very low mammalian toxicity of B.t. and its 

formulations as well as the human experience with B.t. considered in these documents. 

Specifically, the U.S. EPA states: 

...no known mammalian health effects have been demonstrated in 

any infectivity/pathogenicity study .... The sum total of all

toxicology data submitted to the Agency complete with the lack of

any reports of significant human health hazards of the various

Bacillus thuringiensis strains allow the conclusion that all 

3-19
 
infectivity/pathogenicity studies normally required ... be waived

in the future as long as product identity and manufacturing

process testing data indicate there is no mammalian toxicity

associated with the strain (U.S. EPA, 1998, p. 11). 

The application methods suggest that the potential for eye,

dermal and inhalation exposure to mixers, loaders and

applicators does exist. ... However, because of a lack of

mammalian toxicity, the risk from occupational exposure is

minimal ... the health risk [to the general public] is expected to be

negligible due to: (1) The lack of toxicological concerns

associated with Bacillus thuringiensis, and (2) Bacillus

thuringiensis has been used as a pesticide for approximately 50

years with no known adverse effects (U.S. EPA, 1998, p. 14). 

The World Health Organization reaches a similar conclusion: 

Owing to their specific mode of action, Bt products are unlikely

to pose any hazard to humans or other vertebrates or to the great

majority of non-target invertebrates provided that they are free

from non-Bt microorganisms and biologically active products

other than the ICPs [insecticidal crystal proteins]. Bt products

may be safely used for the control of insect pests of agricultural

and horticultural crops as well as forests (WHO 1999, Section

1.7, not paginated). 

In terms of the standard risk assessment paradigm—hazard identification, exposure assessment,

dose-response assessment, and risk characterization— U.S. EPA (1998) and WHO (1999)

reach essentially the same functional conclusion: since no hazard identification can be made for

a clearly adverse effect, a formal dose-response assessment is not necessary. 

The current risk assessment does not substantially disagree with the assessment in U.S. EPA

(1998) and WHO (1999). The available data do not indicate that any serious adverse effects

are likely to occur under plausible conditions of exposure. Notwithstanding this assertion, the 

failure to quantify risk has limitations. First, as noted in the Introduction (Section 1), this risk 

assessment of B.t.k. is accompanied by risk assessments on other agents used against the gypsy

moth and the failure to quantify risk prevents an explicit comparison of risks that may be useful

in risk management decisions. Second, additional studies were published since the risk

assessments presented by U.S. EPA (1998) and the WHO (1999) which are potentially useful

for expanding on the dose-response assessment. Last, substantial public concern is often 

expressed over widespread aerial applications of B.t.k. and these concerns may be more fully 

addressed with an aggressive interpretation of the data. 

3.3.3. Human Data 

The quantitative dose-response assessment in the previous USDA risk assessment of B.t.k. 

(Durkin 1994; USDA 1995) is based largely on the worker study by Cook (1994), and this

study remains the most complete assessment of the effects of B.t.k. in workers. Cook (1994)

provides data on the overall incidence of various health effects in workers, compared with a

control group of individuals not involved in the application of B.t.k. These data are 

summarized in Table 3-4. Based on a comparison between the control group and the workers,

the data demonstrate (using the Fisher exact test and a p-value of 0.05) a statistically significant 

increase in the incidence of irritant effects in workers. The significantly increased effects 

3-20
 
include generalized dermal irritation (dry or itchy skin and chapped lips), irritation to the

throat, and respiratory irritation (cough or tightness). Moreover, the overall incidence of all 

symptoms combined was increased significantly among the workers, compared with the

controls . 

In dealing with multiple comparisons, however, the use of the standard p-value of 0.05 may 

overestimate the number of significant associations. For example, if 100 sets of comparisons

are made within the same population—i.e., there are by definition no differences because there

is only one population—some comparisons may appear to be statistically significant only

because of random differences in the sampling. To address this issue, one standard approach is

to divide the pre-determined significance level, typically taken as 0.05, by the number of

comparisons being made. This is referred to as Bonferroni’s correction (e.g., Curtin and Schulz 

1998). Thus, in the study by Cook (1994), the seven effects (excluding all effects combined)

would lead to an acceptance level for statistical significance of about 0.007 [p-value of 0.05 ÷ 7

= 0.00714]. 

While it is beyond the scope of this risk assessment to discuss Bonferroni’s correction in detail,

it should be noted that Bonferroni’s correction is conservative—i.e., it will reduce the number 

of false positive associations. In terms of a risk assessment, Bonferroni’s correction may be

viewed as anti-conservative in that the presence of a large number of trivial comparisons could

obscure statistically and biologically significant results for a subset of important comparisons. 

Thus, as discussed by Perneger (1998), judgement and an assessment of biological plausibility

must be exercised in the application of Bonferroni’s correction. Specifically for this risk 

assessment of B.t.k., these judgements are discussed further in Section 3.2.5). When 

Bonferroni’s correction is applied to the data from Cook (1994) in Table 3-4, none of the

effects are statistically significant at p<0.007; however, skin irritation (p.0.0077) and throat 

irritation (p.0.0079) are marginally significant. 

Confidence in the biological and statistical significance of these effects would be enhanced if

dose-related or at least exposure-related trends were demonstrated. Cook (1994) does not 

provide incidence data segregated by exposure levels. Nevertheless, as summarized in Table 3­

5 and illustrated in Figure 3-1, Cook (1994) provides data on the number of symptoms per

worker segregated into three exposure groups as well as categories based on the use of 3 protective masks. The exposure groups are based on cumulative cfu/m × hours over three

ranges: <1 to 100, 100 to 300, and >300. The use of masks is simply characterized as none, 

occasional, or regular. If the B.t.k. exposure levels are related to the symptoms considered by

Cook (1994) as specified in Table 3-4, one might expect to see a positive association with

exposure and fewer symptoms in workers wearing protective masks. As illustrated in Figure 3­

1, such associations are few within or among the variables. Cook (1994) does not provide

information about the control group in terms of average number of symptoms per worker and

this lack of information may obscure an association. On the other hand, based on the results 

presented in Table 3-4, which include the incidence of various effects in the control group, it is

not clear that combining all effects as a measure of response is meaningful. In other words, if 

only dermal irritation and irritation to the throat are statistically significant effects, the lack of

clear exposure-response pattens for all effects combined (significant effects as well as random

effects) might be expected. 

At least one of the more recent epidemiology studies may be useful in further assessing the

report by Cook (1994). Since the publication of the previous risk assessment, a number of

epidemiology studies were published (Table 3-1), most of which fail to note remarkable or

statistically significant effects, like the epidemiology studies considered in the 1995 risk

assessment (i.e., Elliott et al. 1988; Elliott 1986; Green et al. 1990; Noble et al. 1992). 

3-21
 
Although some of the more recent studies are discussed further in the risk characterization

(Section 3.4), the study by Petrie et al. (2003) is the only recent study that reports statistically

significant effects. 

As discussed (see Section 3.1.2), Petrie et al. (2003) surveys a group of individuals prior to a

B.t.k. spray (n=292) and a subset of the group after a B.t.k. spray (n=181) recording their 

responses for 25 different endpoints. Based on the per cent responses reported in Table 1 of the

study, Table 3-6 presents the number of responders with each effect before and after the spray

operation. The statistical significance, using the Fisher Exact test is provided in the last

column of Table 3-6. 

The Petrie et al. (2003) study, like the Cook (1994) study, involves multiple comparisons. 

When the Bonferroni correction is applied to 25 comparisons, the adjusted p-value

corresponding to 0.05 for a single comparison is 0.002 [0.05/25]. Based on this correction, 

only one endpoint, throat irritation, with a pair-wise p-value of 0.000048, is regarded as

statistically significant. The interpretation of the respiratory effects observed in the study by

Petrie et al. (2003) is less than straightforward because the effect could be due to or influenced

by pollen count. As noted in the discussion by Petrie et al. (2003), pollen counts in Auckland

peak from October to February. The pre-exposure survey was conducted at the end of October

over a 10-week period prior to spraying, which started in January. The post-exposure survey

was conducted at the end of March, about 12 weeks after the start of spraying. Consequently,

portions of the pre-exposure and post-exposure periods and all of the spray period occurred

during the pollen season. Since portions of the pre-spray and post-spray periods were

concomitant with the pollen season, it is not clear whether this factor introduces a serious bias. 

Nonetheless, both Cook (1994) and Petrie et al. (2003) report throat irritation as an effect in

workers involved in the spray application of B.t.k. The effect is of marginal significance in

Cook (1994) and of clear statistical significance in Petrie et al. (2003), using a statistically 

conservative correction for multiple comparison. This consistency combined with the animal

data indicating that irritation of the mucus membranes of the throat and respiratory tract is a

biologically plausible effect (see Section 3.1.13) suggests that these effects should be attributed 

to B.t.k. exposure. 

As indicated in the exposure assessment (Table 3-3), workers in the study by Cook (1994) were 6 3 3 exposed to concentrations of B.t.k. of up to 15.8 × 10 cfu/m —i.e., about 16 million cfu/m . 

As indicated in Table 3-4, throat irritation was noted in 7% of the control group and 29% of

workers applying B.t.k. Under the assumption of independence, the response associated with 

B.t.k. can be calculated using Abbott’s correction: 

P = (P* - C) ÷ (1 - C) 

where P* is the observed proportion responding, P is the proportion responding that can be 

attributed to exposure (in this case to B.t.k.) and C is the proportion responding in the control 

group (Finney 1972, p. 125). Using this correction, the estimated proportion of workers

evidencing throat irritation attributable to B.t.k. exposure is about 0.24 [(0.29 - 0.07) ÷ (1 ­

0.07) = 0.2366 ] or 24%. 

Petrie et al. (2003) did not monitor B.t.k. concentrations in air. Based on monitoring data from 

similar applications (Table 3-3), members of the general public may be exposed to air 3 concentrations ranging from approximately 100 to 4000 cfu/m during or shortly after aerial

applications of B.t.k. similar to those conducted in the study by Petrie et al. (2003). This range 

3-22
 
6 3 is a factor of 3950 to 158,000 less than the 15.8 × 10 cfu/m from the study by Cook (1994). 

In terms of the quantitative response for throat irritation, Petrie et al. (2003) report rates of

47÷292 (16%) in the pre-spray population and 58÷181 (32%) in the post-spray population. 

Again applying Abbott’s correction, the estimated proportion of the population evidencing

throat irritation attributable to B.t.k. exposure is about 0.19 [(0.32 - 0.16) ÷ (1 - 0.16) = 0.1904 

] or 19%. In that way, as with the number of symptoms per individual summarized in Table 3­

5 and Figure 3-1 from the study by Cook (1994), there appears to be no dose-response

relationship for throat irritation. 

Two factors in the Petrie et al. (2003) study may obscure any underlying dose-response

relationship. First, as noted above, the study was conducted during a period that overlapped

with high pollen counts. Since the high pollen season encompassed the pre-spray and postspray surveys, the extent of bias may not be substantial. The only way to have assessed this

further would have been to include a non-exposed control population, which was not done in

the Petrie et al. (2003) study. The other factor is the possible bias associated with the postspray population. Only 181 of 292 (about 62%) of the individuals responding to the pre-spray

survey responded in the post-spray survey. As noted by Petrie et al. (2003), it is reasonable to

presume that individuals who felt that they were affected by the spray would be more likely to

respond in the post-spray survey, compared with individuals who felt that they were not

affected. This possible source of bias could be further assessed by considering the pre-spray

survey results only for those individuals responding to the post-spray survey. This information, 

however, is not provided in the Petrie et al. (2003) publication. 

3.3.4. Animal Data 

As noted in Section 3.1.13 and summarized in Appendix 1, there is essentially no information

indicating that inhalation exposure to B.t.k. will cause serious adverse health effects. 

Extremely severe inhalation exposures that coat the test species with commercial formulations

of B.t.k. are associated with decreased activity, discolored lungs, and other effects but not

mortality. Although the animal data are consistent with data regarding human exposure B.t.k., 

the animal studies are all based on single concentrations and cannot be used in a meaningful

dose-response assessment. 

The only study that provides a clear dose-response relationship for exposure to B.t.k. involves 

intranasal instillations (Hernandez et al. 2000). In the Hernandez et al. (2000) study, groups of 2 4 7 20 mice were dosed at rates of 10 , 10 , and 10 cfu/mouse with or without doses of influenza 

virus at 4% of the LD50. In mice not exposed to the influenza virus, the only effect noted was 

local inflamation. Hernandez et al. (2000) do not discuss dose-severity or dose-response

patterns for the inflammation. In an earlier study, mortality increased to 80% after 24 hours in 

mice dosed at 108 cfu/mouse evidenced 80% mortality (Hernandez et al. 1999). No mortality 

was observed In mice exposed to the influenza virus alone at 4% of the LD50 or in mice 

2 4 7 exposed to B.t.k. alone at doses of 10 , 10 , and 10 cfu/mouse. In mice exposed to both the 2 4 7 influenza virus at 4% of the LD50 along with B.t.k. at doses of 10 , 10 , and 10 cfu/mouse,

mortality was 4 of 20, 8 of 20, and 14 of 20 (Hernandez et al. 2000). 

The data from the Hernandez et al. (1999, 2000) studies are illustrated in Figure 3-2, where,

mortality is plotted on the Y-axis and log10 dose of B.t.k. (cfu/mouse) is plotted on the X-axis. 

The solid circles represent mortality data from mice treated with influenza and B.t.k. The solid 

line represents the fit of the mortality data to the the probit model using the U.S. EPA

Benchmark Dose Software (http://www.epa.gov/ncea/bmds_training/software/overp.htm). The 

curved dashed line represents the 95% upper limit on risk. The probit model satisfactorily fits

the data (p<0.0001), and the lower limit on the benchmark dose, based on an extra risk of 0.1,

is estimated as 30 cfu/mouse. Because only one dose for the mice not treated with influenza 

3-23
 
virus yielded partial mortality, no formal statistical analyses of these data are conducted. These 

data are simply illustrated in Figure 3-2 and a straight line is drawn from the highest dose at 8 which no mortality occurred to the 80% mortality rate at a dose of 10 cfu/mouse.

In terms of the human health risk assessment, these data are not directly useful. Furthermore, 

the route of exposure (intranasal instillation) makes any use of these data somewhat tenuous. 

Concern with the use of this atypical route of exposure in a dose-response assessment is

exacerbated because the Hernandez et al. (2000) study does not specify whether or not the

instillations were adjusted to a constant volume. If the installations were not adjusted to a

constant volume, it is possible that could be observed in animals with a compromised

respiratory tract (i.e., because of viral infection) because of volumetric bronchial obstruction or

a combination of bronchial obstruction and B.t.k. 

Notwithstanding these reservations, the Hernandez et al. (1999, 2000) studies provide the best

dose-response data available in experimental mammals. Table 3-7 provides dose conversions

that may be valuable in further exploring the useful of these data. In Table 3-7, the first 

column indicates the cfu/mouse from the studies by Hernandez et al. (1999, 2000)and the

second column provides the estimated concentration of B.t.k. required to achieve the cfu/mouse 

dose in a 1-hour exposure. This value is calculated as cfu/mouse divided by the estimated 3 breathing rate (m /hour) of a 20 g mouse.

The calculated concentrations in air from cfu/mouse may be extremely conservative in the

assumption that all of the inhaled B.t.k. will be retained. Nonetheless, the study by Holbert

(1991) noted no mortality but some signs of toxicity in mice after 4-hour inhalation exposures

to Foray 76B at a concentration of 3.13×109 cfu per L. This concentration is equivalent to 12 3 13 3.13×10 cfu/m . Adjusting for the 4-hour exposure, the concentration is about 1.3×10 3 12 3 cfu/m × hours [3.13×10 cfu/m × 4 hours], which is approximately 5.5 times less than the 

concentration associated with 80% lethality in mice exposed to B.t.k. via intranasal 

installation (Hernandez et al. 1999) and approximately 1.8 times greater than the highest

concentration associated with inflamation. While this cannot be overly interpreted, the signs of

toxicity but lack of mortality observed in the Holbert (1991) inhalation study do appear to be 3 reasonably consistent with the conversion of cfu/mouse to cfu/m × hours presented in Table 3­

7. 

The best approach for extrapolating from mice to humans is uncertain. Following the

suggestion by Siegel (2001), dose in units of cfu/mouse are converted to an equivalent cfu per

human by adjusting body weight—i.e., 70 kg÷0.02 kg. These values are given in the third 

column of Table 3-7. The equivalent concentration in air is then calculated as the cfu per 3 human divided by the breathing rate (m /hour) of a human engaging in moderate physical

activity, presented in the fourth column of Table 3-7. 

As noted in Section 3.2.3, exposures over a wide range of B.t.k. concentrations in air are 

associated with respiratory irritation in humans. At the lower end of the exposure range,

concentrations probably in the range of 100 to 4000 cfu/m3 are associated with an increased 

incidence of throat irritation in members of the general population based on the epidemiology

study by Petrie et al. (2003). Monitoring data reported by Teschke et al. (2001) suggest that

concentrations in outdoor air after 5 to 6 hours would be about 10-fold lower but that 

3 concentrations in indoor air could be approximately 250 cfu/m (see Table 3-3). At the upper 6 3 range of exposure, B.t.k. concentrations of up to 15.8 × 10 cfu/m are associated with throat 

irritation in workers (Cook 1994). Both studies report similar response rates: about 19% in the

lower exposure for the general public and about 24% in the occupational exposures. 

3-24
 
According, there is no clear or strong exposure-response relationship. Severe adverse effects 

are not reported in either study. 

This pattern is consistent with the available toxicity data in mice. Over a broad range of

intranasal doses—i.e., 100 to 100-million cfu/mouse— the only effects reported by Hernandez 3 et al. (2000) involve inflammation. Based on the estimates of human equivalent cfu/m × hour 5 presented in Table 3-7, exposures ranging from approximately 100,000 (1×10 ) to 3 approximately 10,000,000,000 (1×1010 or 10 billion) cfu/m × hours are likely to result in local

inflamation but not mortality. 

The mouse studies were conducted at doses that are not likely to be encountered by members of

the general public exposed to B.t.k. Consequently, the mouse data cannot be used directly to

support the responses reported by Petrie et al. (2003). Nonetheless, the weight-of-evidence

suggests that some members of the general public could experience respiratory irritation at

B.t.k. concentrations ranging from 100 to 4000 cfu/m . 3 The apparent lack of a strong doseresponse relationship in humans is consistent with the wide dose range leading to local

inflamation in mice. 

Finally, the failure to note any severe adverse effects in humans exposed to B.t.k. 

6 3 7 3 concentrations of up to 15.8 × 10 cfu/m (1.58 × 10 cfu/m ) reported by Cook (1994) is also 

consistent with the available animal data suggesting that no mortality would be expected at 10 3 concentration of up to 1.4 ×10 cfu/m × hours. In other words, a worker would need to be 

exposed to 1.58 × 107 cfu/m for about 37 days to reach a cumulative dose of 1.4 ×10 3 10 cfu/m 3 

10 3 7 3 × hours [(1.4 ×10 cfu/m × hours) ÷ 1.58 × 10 cfu/m = 886 hours or about 37 days]. The 8 3 highest cumulative exposure reported by Cook (1994) is >3×10 cfu/m × hours, a factor of 3 about 50 below the highest estimated non-lethal exposure of 1.4 ×1010 cfu/m × hours base on

the available data in experimental animals. 

3.3.5. Values Used for Risk Characterization 

In some respects, the dose-response assessment for B.t.k. is not much different from that of the 

previous risk assessment (Durkin 1994; USDA 1995). Under plausible conditions of exposure, 

there is no indication that B.t.k. will cause severe adverse effects and the most plausible effects 

are likely to involve irritation. 

The current dose-response assessment can be elaborated in two ways. First, based on a 

consideration of the study by Hernandez et al. (2000) and the estimates of equivalent human

exposures given in Table 3-7, it seems plausible that cumulative exposures up to 1.4×1010 

3 7 cfu/m × hour will not cause adverse effects. This assumption is based on the 1×10 cfu/mouse 

dose group in the study by Hernandez et al. (2000) in which local inflammation was the only 3 adverse effect observed. Further support is drawn from the NOAEL of 3×108

 cfu/m × hours 

for adverse health effects in humans reported in the Cook (1994) study in which the only

effects of marginal significance are throat irritation and skin irritation. The potential need for 3 an uncertainty factor on the 1.4×1010 cfu/m × hour is questionable given the reasonable

consistency of the human data with the animal data. This issue is discussed further in Section 

3.4 (Risk Characterization). 

While a human NOAEL for serious signs of toxicity can be estimated, the NOAEL for irritant

effects cannot be estimated. The data suggest that at low and plausible concentrations

associated with the normal application of B.t.k., irritant effects may be reported by a substantial

number of individuals—i.e., about 20% of the population. Irritant effects will also be reported

at much higher concentrations, although the incidence of the effects may not be substantially 

greater. 

3-25
 
Another major difference between the previous dose-response assessment for B.t.k. (Durkin

1994; USDA 1995) and the current risk assessment is the identification in the current risk

assessment of a potential concern for individuals with respiratory diseases such as influenza. 

As illustrated in Figure 3-2, the study by Hernandez et al. (2000) clearly suggests that otherwise

non-lethal doses of B.t.k. can be associated with pronounced lethality in mice infected with

otherwise non-lethal doses of influenza virus. Based on the probit model, a benchmark dose of 

30 cfu/mouse can be calculated. 

Concern for the report by Hernandez et al. (2000) is somewhat enhanced by an earlier study by

Berg (1990) in which rats were given an intravenous dose of 1 mL Foray 48B. 

Histopathological findings in the liver and the reticuloendothelial system were attributed to a

background infection. The pathology results, however, were more severe in the exposed group

compared with the controls. This could suggest that the B.t.k. may have aggravated this disease 

condition. Most of the histopathological findings, however, appear to have been due to

extensive removal of bacteria by the reticuloendothelial system, including Kupffer cells in the

liver, spleen, and lymph nodes. Thus, this study may simply suggest that B.t.k. organisms can

survive and reproduce in a mammalian host (i.e., persistence) rather than suggest any

underlying pathogenicity. 

It is unclear whether or not the data on mice exposed to both B.t.k. and an influenza virus can 

or should be applied directly and quantitatively to the human health risk assessment. One very

significant problem in the quantitative use of these data is in the interpretation of 4% of the

LD50 for mice relative to possible disease conditions in human populations. This issue is 

discussed further in the risk characterization. 

3-26
 
3.4. RISK CHARACTERIZATION 

3.4.1. Overview 

The risk characterization for B.t.k. and its formulations is consistent with the risk 

characterization in the previous USDA risk assessment as well as more recent risk assessments

conducted by the U.S. EPA and the World Health Organization: B.t.k. and its formulations are 

likely to cause irritant effects to the skin, eyes, and respiratory tract; however, serious adverse

health effects are not of plausible concern. Nevertheless, the approach used to quantify risk for 

irritant effects and more serious health effects is different, based on recent information 

regarding B.t.k. exposure. 

Unlike the previous USDA risk assessment on B.t.k., this document does not attempt to

quantify the risk of irritant effects since there is no clear threshold for those effects. When 

B.t.k. is applied under conditions similar to those used in USDA programs to control or

eradicate the gypsy moth, irritant effects are likely to occur in some members of the general

public as well as in some workers. Throat irritation is the best documented health effect in 

humans after exposure to B.t.k.; however, skin irritation and eye irritation are also likely to

occur, although perhaps at the upper extremes of exposure. 

Although serious adverse health effects in humans are not likely to result from B.t.k. 

applications, this risk assessment, unlike the previous USDA risk assessment and the risk

assessments conducted by the U.S. EPA and the World Health Organization, considers the

possibility that serious adverse effects may result from exposure to B.t.k. and quantifies the 

risk. The bases for this approach are the recent in vitro studies suggesting that cellular damage 

is a plausible effect of B.t.k. exposure and the in vivo studies indicating that serious effects,

including mortality, are possible at extremely high exposure levels. There is however, no 

reason to assume, given the reasonably good monitoring data, conservative exposure

assumptions, and highly aggressive and conservative use of the available toxicity data, that any

human population—ground workers, aerial workers, or members of the general public—are

likely to experience overtly toxic effects from the normal use of B.t.k. in programs like those 

conducted by the USDA. At the extreme upper range for ground workers, exposure levels are

estimated to 25 times lower than the functional human NOAEL. For members of the general

public, exposurelevels are estimated to be approximately 28,000 to 4,000,000 [4 million] times

lower than the functional human NOAEL. 

The available toxicity data give no indication that subgroups of the general population are

likely to be remarkably sensitive to B.t.k.. Two recent epidemiology studies have found that

asthmatics are not likely to be adversely affected by aerial applications of B.t.k. On the other 

hand, there is one essentially anecdotal reference involving a severe allergy to a carbohydrate in 

a B.t.k. formulation which is not supported, however, in any of the published epidemiology

studies. Nonetheless, B.t.k. formulations are complex mixtures and there is a possibility that

certain individuals may be allergic to one or more of the components in the formulations, as

acknowledged by a state health service. 

An incidence in which mortality increased substantially in mice pre-treated with an influenza

virus and exposed to various doses of B.t.k. raises concern regarding the susceptibility of

individuals with influenza or other viral respiratory infections to B.t.k. toxicity. The viral 

enhancement of bacterial infections is not uncommon, and the enhancement of B.t.k. toxicity by 

a viral infection is not altogether surprising. Nonetheless, the relevance of this observation to 

public health cannot be assessed well at this time. Although the concurrence of viral 

enhancement and B.t.k. exposure are not reported in the available epidemiology studies, it is

not clear that the studies would detect such an event or that the effect is of plausible concern at 

3-27
 
the typical or even extreme exposure levels anticipated in gypsy moth control programs. The 

viral enhancement of B.t.k. toxicity is likely to be an area of further study in the coming years. 

3.4.2. Irritant Effects 

As discussed in the Hazard Identification (Section 3.1), B.t.k. formulations can be irritating to 

the skin, eyes, and respiratory tract. This conclusion is consistent with previous risk 

assessments of B.t.k. and other strains of B.t. (U.S. EPA 1998; WHO 1999). Moreover, most 

of the material safety data sheets for B.t.k. include warnings about dermal, ocular, and 

respiratory tract irritation. 

The extent to which these irritant effects are classified as adverse is largely semantic. Based on 

the available epidemiology studies (Table 3-2), these effects are not severe enough to compel

the general public to seek medical attention or to cause individuals involved in the application

of B.t.k. to lose time from work. Even so, among the adverse human health effects associated 

with B.t.k. exposure, irritant effects are the most common. 

The principal issue in quantifying the risk for irritant effects in humans exposed to B.t.k. is the 

lack of a clearly defined threshold. As discussed in the dose-response assessment (see Section

3.3), throat irritation was reported by members of the general public after aerial applications of

B.t.k. at rates typical of those used in USDA programs (Petrie et al. 2003). While a number of 

other adverse or at least undesirable effects also are noted by Petrie et al. (2003), the

association of these effects with exposure to B.t.k. is less clear. For throat irritation, however, 

the association seems compelling (Table 3-6). In addition, workers reported throat irritation 

after exposure to higher levels of B.t.k. There does not appear to be a remarkable doseresponse relationship for the incidence of throat irritation—i.e., about 19% in members of the

general public at presumably low exposure levels and about 24% in workers at much higher

concentrations. 

The lack of a dose-response relationship raises questions concerning the biological significance

of this effect, particularly at low exposure levels. As discussed by Petrie et al. (2003), there

may be biases in an epidemiology study involving self-reporting that reflect anxiety rather than

physical damage. Furthermore, as Petrie et al. (2003) indicate, their study was conducted

during a period of high pollen counts, which may explain the apparent increase in throat

irritation, assuming that the effect was confounded by allergies. Although a full study using a 

control population not exposed to B.t.k. might help to address the issue, both the pre-exposure

and post-exposure periods covered by the study did partially encompass the pollen season. 

Supported by data on human exposure and the experimental studies in other mammals (see

Section 3.1.11), the weight-of-evidence suggests that throat irritation reported by Petrie et al.

(2003) may be biologically as well as statistically significant. 

The inability to define a clear threshold for irritant effects and the lack of an apparent doseresponse or dose-severity relationship substantially impairs the quantitative expression of risk

based on the standard hazard quotient approach. For example, one approach to defining a

pseudo-human NOAEL might be to assert that responders in the Petrie et al. (2003) study were 3 probably exposed to higher concentrations of—i.e., greater than1000 cfu/m —and to propose 3 that the lower range of plausible exposure —e.g., 100 cfu/m —might be used as a functional

NOAEL for deriving hazard quotients. An approach analogous to this is taken in the previous 

USDA risk assessment of B.t.k. (Durkin 1994; USDA 1995). 

The proposed approach is not taken in the current risk assessment because, in addition to the

obvious problems with the logic of the approach and lack of data to support the presumed

NOAEL, the resulting hazard quotients would be meaningless in terms of expressing risk. For 

3-28
 
3 example, individuals exposed to 1000 cfu/m would have a hazard quotient of 10 [1000 ÷ 100 3 6 3 cfu/m ] and workers exposed to 15.8 × 10 cfu/m (i.e., workers in the study by Cook 1994) 3 would have a hazard quotient of 158,000 [15,800,000 ÷ 100 cfu/m ], leading to the conclusion,

based on the hazard quotients, that workers exposed to B.t.k. are at much greater risk than the

general public to irritant effects, which is not the case, as noted in Section 3.3.3. Moreover, 

there is no evidence that a hazard quotient of 10 has any greater effect than hazard quotients of

10,000 or 100,000 or any lesser effect than a hazard quotient of 2. 

Accordingly, the potential risks for irritation are not quantified in this risk assessment, and are

addressed only qualitatively. As discussed in Section 3.3.3 (Dose-Response Assessment,

Human Data), the studies by Cook (1994) and Petrie et al. (2003) provide credible evidence

that some members of the general population and some workers may experience throat

irritation after exposure to B.t.k. from aerial or ground applications. Irritation to the skin and 

eyes is also plausible, although less well supported by the available data in humans except

under extreme exposure conditions. 

Eye irritation may result when small amounts of commercial formulations of B.t.k. are splashed 

into the eyes. The probabilities of this event occurring under various exposure scenarios (that

is, number of hours worked) cannot be estimated from available data. Nonetheless, there are 

reports of eye irritation resulting from direct splashing of B.t.k. formulations in the eye (i.e., 

Samples and Buettner 1983; Green et al. 1990). Thus, the probability of such an event seems

sufficiently high to justify precautions when handling concentrated formulations in such a way

that splashing into the eyes is not a potential risk. Also, workers exposed to B.t.k. may be at

risk of skin irritation, and the study by Bernstien et al. (1999) suggests that skin sensitization is

a plausible effect of exposure. 

3.4.3. Serious Adverse Effects 

The previous risk assessments on B.t.k., including the previous risk assessment conducted for

the USDA, accept the general premise that B.t.k. is essentially incapable of causing serious

adverse health effects under any conditions (Durkin 1994; U.S. EPA 1998; USDA 1995; WHO

1999). More recent studies on B.t.k., however, suggest that adverse effects are possible, albeit

under extreme exposure conditions that are not representative of field applications of B.t.k. 

formulations. Tayabali and Seligy (2000) demonstrated that B.t.k. causes cytotoxicity in vitro. Also, as discussed in the dose-response assessment (see Section 3.3.4), the studies by

Hernandez et al. (1999, 2000) allow for an estimate of lethal doses as well as doses in which no

adverse effects, other than local inflamation, were noted. 

The use of these data quantitatively in a risk assessment is admittedly tenuous. Nonetheless, 

as discussed in Section 3.3.4, these are the best data available. Although intranasal instillation

is not a directly relevant route of exposure, the estimates of non-lethal and lethal concentrations

are consistent with the in vivo inhalation study by Holbert (1991), and the estimated human

NOAEL is consistent with the worker data from Cook (1994). 

Based on the calculations summarized in Table 3-7, equivalent human exposure concentrations 3 of 1×1010 cfu/m × hour could be adopted directly as a NOAEL with a 10-fold higher dose 11 3 [1×10 cfu/m × hour] as a LOAEL. As noted in Section 3.3, a case could be made for 

applying an uncertainty factor to the NOAEL. Typically, an uncertainty factor of 100 is used to

account for species-to-species extrapolation or sensitive individuals. As detailed in Table 3-7, 

however, the very conservative approach used to the estimate the equivalent human

concentration in air is less than that of the equivalent concentration for the mouse by a factor of 3 more than 500. Thus, no additional uncertainty factor for the NOAEL of 1×1010 cfu/m × hour

3-29
 
is used in this risk assessment. The potential for effects on sensitive individuals is discussed 

further in Section 3.4.3). 

3 Using an approximated NOAEL of 1×1010 cfu/m × hour for human exposure, the risk

characterization for serious toxic effects is summarized in Table 3-8. As indicated in the first 

column, three groups of individuals are considered: members of the general public, workers

involved in aerial applications of B.t.k., and workers involved in ground applications of B.t.k. 

A plausible range of concentrations for each group is based on published studies detailed in

Table 3-3. For members of the general public, the concentration ranges from 100 to 5000 

cfu/m . 3 The lower end of this range is somewhat higher than outdoor concentrations 

anticipated 5 to 6 hours after spraying (Teschke et al. 2001). The upper range is set to 3 encompass the highest reported concentration—i.e., 4200 cfu/m from Elliott et al. (1988). The 

concentrations for aerial workers are based on the study by Elliott et al. (1988), and the

concentrations for ground workers are based on the study by Cook (1994). For members of the 

general public, the duration of exposure is taken as 24 hours. Based on the monitoring data by

Teschke et al. (2001), this duration is likely to be extremely conservative but is intended to

encompass the possibly higher concentrations of B.t.k. measured in indoor air relative to 

outdoor air 5 to 6 hours after application (Teschke et al. 2001). For workers, the duration of 

exposure is taken as 8 hours to account for a regular work day. Since workers are not likely to 

spend 8 hours applying B.t.k. due to other job requirements, this exposure duration is probably 

somewhat conservative. An additional ground worker group, labeled as extreme range, is 

added to account for the report in Cook (1994) that some ground workers may have been 3 exposed to B.t.k. concentrations greater than 300 million cfu/m × hour. The cumulative 

exposure is then calculated in the fourth column of Table 3-8 as the product of the

concentration and duration of exposure—i.e., hours × cfu/m . 3 The hazard quotient is given in 

the last column as the cumulative exposure divided by the estimated human NOAEL of 1×1010 

3 cfu/m × hour.

The interpretation of the hazard quotients is simple and unambiguous. Given the reasonably

good monitoring data, conservative exposure assumptions, and aggressive and conservative use

of the available toxicity data, there is no reason to assume that any member of the human

population—ground workers, aerial workers, or members of the general public —are likely to

experience overtly toxic effects from the normal use of B.t.k. in programs like those conducted 

by the USDA. The extreme upper range of exposure levels for ground workers are estimated to

be below the functional human NOAEL by a factor of 25. For members of the general public,

exposures are estimated to be below the functional human NOAEL by factors of about 28,000

to 4,000,000 [4 million]. 

These or any other numerical expressions of risk must be interpreted with some caution. In the 

recent review of the toxicity of several strains of B.t.k. to mammals, Siegel (2001) quotes an

earlier assessment by Burges (1981) concerning general testing needs for microbial pesticides,

and this quotation bears repeating: 

... a “no risk” situation does not exist, certainly not with

chemical pesticides and even with biological agents one cannot

absolutely prove a negative. Registration of a chemical is

essentially a statement of usage in which the risks are

acceptable. The same must apply to biological agents. – Burges 

(1981, pp. 738-739). 

Within this definition of safety or acceptable risk, there remains no basis for asserting that the

use of B.t.k. to control the gypsy moth is likely to have adverse toxic effects on any group. 

3-30
 
A major and extremely important uncertainty in this risk characterization concerns the use of a

toxicity study involving nasal instillation and the attendant uncertainties in extrapolating this

type of study to inhalation exposures in humans. An inhalation study similar in general design

to the study by Hernandez et al. (2000) – i.e., using mice challenged with an influenza virus as

well as appropriate controls – would be necessary for assessing more fully and improving the

quality of the risk characterization. 

3.4.4. Groups at Special Risk 

The previous USDA risk assessment (Durkin 1994; USDA 1995) notes a weakly positive

relationship in the incidence of irritant effects in ground workers with and without a history of

asthma, seasonal allergies, or eczema (Cook 1994). Swadener (1994) also notes that some 

formulations of B.t.k. contain sodium sulfite, which may cause adverse effects in asthmatics

taking steroid treatments. As discussed in Section 3.1.2, Pearce et al. (2002) conducted an

epidemiology study designed specifically to address the potential increased risk for young

asthmatics exposed to B.t.k.. The results of the study indicate that there were no significant

differences among individuals present inside or outside the treated area. The study, which

involved subjective reports of health as well as clinical measurements of peak expiratory flow

rates has limitations. Specifically, the treated and control areas were close to one another, and 

the monitoring data indicate that individuals in the treated and control areas were exposed to

B.t.k. Nonetheless, there was no detectable adverse effects in either population (Pearce et al.

2002). 

Swadener (1994) summarizes an incident in which a carbohydrate inert in Foray 48B may have

caused an allergic response in one woman. As discussed in Section 3.1.7, the incident is not 

well documented and the interpretation remains uncertain. Commercial formulations of B.t.k. 

are complex mixtures of many different carbohydrates and other materials to which certain

members of the general population may be allergic (Oregon Health Services 2003). There is, 

however, no documented case of a severe allergic response in the epidemiology studies

conducted on B.t.k. (Table 3-1). 

Hernandez et al. (2000) demonstrate a substantial increase in mortality in mice pre-treated with

an influenza virus and exposed to various doses of B.t.k. The study raises concern regarding

the susceptibility of individuals with influenza or other viral respiratory infections to the

toxicity of B.t.k.. As illustrated in Figure 3-2, increased mortality was observed at a very low

dose—i.e., 100 cfu/mouse —which is one-million times lower than the lethal dose in non-viral

8 treated mice—i.e., 1×10 cfu/mice. Based on an extra risk of 0.1, the estimated lower limit on 

the benchmark dose is 30 cfu/mouse (see Section 3.3.4). Following the conversion approach

used in Table 3-7, this value corresponds to a human exposure level of 42,000 cfu/m . 3 The use 

of the LD10 is not to suggest that such a risk is acceptable but rather to illustrate an exposure

level for which the response rate would be readily detected in most epidemiology studies. 

The potential significance of the Hernandez et al. (2000) study to public health is difficult to 

assess. As noted in Table 3-3, most human exposure levels are well below 42,000 cfu/m . 3 On 

the other hand, cumulative exposure levels for the general public, based on the conservative 3 estimates used for this risk assessment, could range up to 360,000 cfu/m × hours. More 

plausible estimates, based on only a 2-hour rather than a 24-hour duration, range from 1200 to 3 30,000 hours × cfu/m for members of the general public. Consequently, it is not clear whether 

the human experience with B.t.k.—i.e., the epidemiology studies summarized in Table 3­

3—can be used as evidence to preclude the possible association between viral infections and

the enhanced toxicity of B.t.k. or to establish that the viral enhancement of B.t.k. toxicity is not 

of plausible concern regarding human exposure. Such effects were not observed in ground 

workers, who clearly are exposed to B.t.k. concentrations far greater than 42,000 

3-31
 
3 cfu/m × hours. Nonetheless, the viral enhancement of bacterial infections is not uncommon 

and the enhancement of B.t.k. toxicity by a viral infection seems plausible. This issue is likely

to the subject of further study in the coming years and should be monitored by groups involved

in the use of B.t.k. 

3.4.5. Cumulative Effects and Connected Actions 

The cumulative effects associated with the application of B.t.k. formulations must consider the 

normal background exposure to B.t.k., residual exposure to B.t.k. and formulation products

after a single application, and the effects of multiple applications in a single season and over

several years. Since the dose-response assessment is based on measures of cumulative 3 exposure —i.e., hours × cfu/m —and is supported by epidemiology studies, this type of

cumulative effect is implicitly considered in the dose-response assessment. Given the 

reversible nature of the irritant effects of B.t.k. and the low risks for serious health effects, 

cumulative effects from spray programs conducted over several years are not expected. 

Workers or members of the general public who are exposed to aerial or ground sprays of B.t.k. 

also will be exposed to the gypsy moth and may be exposed to other control agents. There are 

no data indicating that risks posed by these other agents will affect the response, if any, to B.t.k. 

formulations. Similarly, exposure to other chemicals in the environment may impact the

sensitivity of individuals to B.t.k. or other agents; however, the available data are not useful for

assessing the significance of such interactions. 

3-32
 
4. ECOLOGICAL RISK ASSESSMENT 

4.1. HAZARD IDENTIFICATION 

4.1.1. Overview. 

The hazard identification for mammals is closely related to the hazard identification for the

human health risk assessment in that both are based, in part, on numerous standard toxicity

studies in experimental mammals. Although B.t.k. may persistent in mammals for several

weeks after exposure, there is little indication that oral or dermal exposure leads to any serious

adverse effects. Most inhalation studies do not suggest a potential for adverse effects even at 

B.t.k. concentrations much greater than those likely to be encountered in the environment. The 

lack of a positive hazard identification is supported by field studies which demonstrate a lack

of adverse effects in populations of mammals exposed to applications of B.t.k. Nonetheless, 

there are data to suggest that extremely high concentrations of B.t.k. in air might pose a hazard. 

Toxicity studies in birds are limited to standard acute exposures required by U.S. EPA for

product registration. The studies all involve either single-dose gavage administration or five

daily-dose gavage administrations, and none of the studies reports signs of toxicity or

pathogenicity at single oral doses up to 3333 mg formulation per kg bw or at multiple oral

doses up to 2857 mg formulation per kg bw. Due to the lack of toxicity of B.t.k. formulations 

as well as other B.t. strains, the U.S. EPA did not require chronic or reproductive toxicity 

studies in birds. The apparent lack of B.t.k. toxicity is supported by numerous field studies in 

birds. In one field study, a transient decrease in abundance was noted in the spotted towhee 

(Pipilo maculatus). This observation is inconsistent with other field studies on B.t.k., and, 

according to the investigators, may be an artifact of the study design. 

The mechanism of action of B.t.k. in lepidoptera is relatively well characterized. B.t.k. 

vegetative cells produce spores and crystals. After the insect consumes the crystals, toxins are

formed that attach to the lining of the mid-gut of the insect and rupture the cell walls. The 

B.t.k. spores germinate in the intestinal tract and enter the body cavity through the perforations

made by the crystal toxins. The bacteria replicate in the body cavity, causing septicemia and 

eventual death. While various strains of B.t. are often characterized as selective pesticides, 

B.t.k. is toxic to several species of target and non-target lepidoptera. Sensitive non-target

lepidoptera include larvae of the Karner blue butterfly, two species of swallowtail butterflies, a

promethea moth, the cinnabar moth, and various species of Nymphalidae, Lasiocampidae, and

Saturniidae. 

While some non-target lepidopteran species appear to be as sensitive as target species to B.t.k., most studies indicate that effects in other terrestrial insects are likely to be of minor

significance. There is relatively little information regarding the toxicity of B.t.k. or B.t.k. 

formulations to terrestrial invertebrates other than insects. Some oil-based B.t.k. formulations 

may be toxic to some soil invertebrates; however, the toxicity is attributable to the oil in the

formulation and not to B.t.k. There is no indication that B.t.k. adversely affects terrestrial 

plants or soil microorganisms. 

The U.S. EPA classifies B.t.k. as virtually non-toxic to fish, and this assessment is consistent

with the bulk of experimental studies reporting few adverse effects in fish exposed B.t.k. 

concentrations that exceed environmental concentrations associated with the use of B.t.k. in 

USDA programs. Although there are no data regarding the toxicity of B.t.k. or its formulations 

to amphibians, other strains of B.t. appear to have low toxicity to amphibians. The effects of 

B.t.k. on aquatic invertebrates is examined in standard laboratory studies and in numerous field

studies. At concentrations high enough to cause decreases in dissolved oxygen or increased 

4-1
 
biological oxygen demand, B.t.k. may be lethal to certain aquatic invertebrates, like Daphnia 

magna. Most aquatic invertebrates, however, seem relatively tolerant to B.t.k. This assessment 

is supported by several field studies that have failed to note remarkable effects in most species

after exposures that substantially exceed expected environmental concentrations. As with 

effects on terrestrial plants, the toxicity of B.t.k. to aquatic plants has not been tested. 

U.S. EPA (1998) raises concerns that some batches of B.t. may contain heat labile exotoxins 

that are toxic to Daphnia. The production of these toxins is an atypical event thought to be

associated with abnormal or poorly controlled production process. The U.S. EPA requires

manufacturers to submit a daphnid study on each new manufacturing process to demonstrate

that heat labile exotoxin levels are controlled. 

4.1.2. Toxicity to Terrestrial Organisms. 

4.1.2.1. Mammals –The hazard identification for mammals is closely related to the hazard

identification for the human health risk assessment (see Section 3.1) in that both are based, in

part, on numerous standard toxicity studies in experimental mammals (Appendix 1). As 

discussed in Section 3.1 and summarized inAppendix 1, B.t.k. may persistent—i.e., may

survive and be recovered—in mammals for several weeks after exposure; however, there is

little indication that oral or dermal exposure leads to serious adverse health effects. Most 

inhalation studies do not suggest a potential for adverse effects even at B.t.k. concentrations 

much greater than those likely to be encountered in the environment. The lack of a positive

hazard identification is supported by field studies in which no adverse effects were observed in

populations of mammals exposed to B.t.k. applications of (Belloq et al. 1992; Innes and 

Bendell 1989). Nonetheless, as discussed in the human health risk assessment (see Section

3.3.4), there are data to suggest that extremely high air concentrations of B.t.k. in air might pose 

a hazard. 

Acute oral doses of up to approximately 5000 mg per bw of B.t.k. formulations do not cause 

adverse effects in rodents (Bassett and Watson 1999a; Kuhn 1998b; Cuthbert and Jackson

1991; Kuhn 1991). Other acute oral toxicity studies report exposure levels in units of cfu per

rat and indicate that doses of up to 108 cfu per rat are not associated with signs of toxicity 

(David 1990b; Harde 1990b). Similarly, in longer-term studies, B.t.k. doses of up to 8400

mg/kg/day were not associated with adverse effects in rats over a 2-year period (McClintock et

al. 1995b) and doses of up to 500 mg/kg/day B.t.k. (corresponding to approximately 1012 spores

per day) were not associated with adverse effects in sheep over a 5-month exposure period

(Hadley et al. 1987). The only suggestion of an adverse effect is the death of one of four male

Sprague-Dawley rats 1 day after a gavage dose of 5050 mg DiPel technical powder per kg. 

This effect, however, was attributed to a gavage dosing error that resulted in the accidental

aspiration of the test material —i.e., inadvertently transporting the material into the lungs

(Bassett and Watson 1999a). Thus, as in the human health risk assessment, the hazard 

identification for the oral route of exposure is essentially negative—i.e., there is no indication

that adverse effects will result from oral exposure to B.t.k. or B.t.k. formulations at 

concentrations far higher than exposure levels which might be anticipated in the environment. 

Although the available studies report very high NOAELs, no LOAELs are reported. 

Similarly, there is no indication that dermal exposures will result in adverse systemic effects. 

As summarized in Appendix 1, dermal applications of undiluted B.t.k. formulations will lead to 

irritant effects in rats and rabbits; however, no signs of systemic toxicity—i.e., effects other

than those at the site of application—are reported in the literature (Kuhn 1998b; Kuhn 1999a;

Meher et al. 2002; Bassett and Watson 1999b; Jacobsen 1993; Berg et al. 1991; Kiehr 1991a). 

4-2
 
Unlike oral or dermal exposure to B.t.k., there is probable concern that extreme inhalation

exposures may pose a risk of adverse health effects. As discussed in Section 3.1.13, this 

assessment is based on the studies by David (1990c) and Hernandez et al. (2000) indicating that

intratracheal instillations and intranasal instillations, respectively, may lead to mortality in rats. 

Concern regarding the possible risk posed by inhalation exposure to B.t.k. is enhanced by

reports of less severe adverse effects in rats (Holbert 1991, Appendix 1) as well as the report by

Bassett and Watson (1999a), discussed above, indicating that accidental aspiration of a B.t.k. 

powder might have caused death in a rat. As discussed further in the dose-response assessment

(Section 4.3) and risk characterization (Section 4.4), this information leads to the same

assessment of risk as for oral and dermal exposures—i.e., the risk at environmentally plausible

concentrations is very low. Unlike the case with either oral or dermal exposures, however, a

LOAEL for serious toxic effects can be approximated for inhalation exposures. 

4.1.2.2. Birds – Toxicity studies in birds are limited to standard acute exposures required by

U.S. EPA for product registration. The studies all involve either single-dose gavage

administration (Beavers et al. 1988a) or five daily-dose gavage administrations (Beavers

1991b; Lattin et al. 1990a,b,c,d,e,f,g), and none of the studies reports signs of toxicity or

pathogenicity at single oral doses up to 3333 mg formulation/kg bw or at multiple oral doses up

to 2857 mg formulation/kg bw (Appendix 2). Due to the lack of evidence regarding acute 

toxicity in birds exposed to B.t.k. formulations or other B.t. strains, the U.S. EPA did not 

require chronic or reproductive toxicity studies in birds. 

The apparent lack of B.t.k. toxicity to birds is supported by several field studies summarized in 

Appendix 2. B.t.k. applied at rates sufficient to decrease the number of caterpillars had no

substantial adverse effects on most bird species (Rodenhouse and Holmes 1992; Nagy and

Smith 1997; Sopuck et al. 2002). The relatively minor effects observed in some species were

considered indirect and attributed to alterations in the availability of prey rather than to the 

direct toxicity of B.t.k. (Gaddis 1987; Gaddis and Corkran 1986; Norton et al. 2001). 

Sopuck et al. (2002) report an unusual observation regarding effects in songbirds exposed to

B.t.k. As summarized in Appendix 2, these investigators conducted population surveys of 42

species of songbirds in areas treated with three applications of Foray 48B at a rate of 50 BIU/ha

(approximately 20 BIU/acre). Significant effects were noted in only one species, the spotted 

towhee (Pipilo maculatus); however, the effect (a decrease in abundance) was noted only

during the spray year and not 1year after treatment. As discussed by Sopuck et al. (2002), the

reason(s) for this decrease are not apparent; however, the time course of the effect was not

related to a decrease in caterpillar abundance. The authors suggest that the effect might be an

artifact of using only a single pre-application survey. Generally, this study is consistent with

other field studies indicating no substantial effects on bird populations exposed to B.t.k. 

4.1.2.3. Terrestrial Invertebrates 

4.1.2.3.1. Lepidoptera – The mechanism of action of B.t.k. in lepidoptera is relatively 

well characterized. B.t.k. vegetative cells produce spores and crystals. The crystals are

repeating protein subunits composed of proteinaceous toxins, enzymes, and other proteins. 

B.t.k. must be eaten in order to be effective as an insecticide. The crystals dissolve in insect

gastrointestinal tracts that have a high pH—i.e., they are alkaline or basic. Proteolytic enzymes

in the insect gut and in the crystals themselves break down the crystals (prototoxins) into active

toxic subunits. The toxins attach to the lining of the mid-gut of the insect and rupture the cell

walls, which allows the alkaline contents of the gut to spill into the body cavity (Drobniewski

1994). The B.t.k. spores germinate in the intestinal tract and enter the body cavity through the

perforations made by the crystal toxins, replicate, and cause septicemia. The body tissues of 

4-3
 
the insect are consumed by B.t.k. The infected insect usually stops feeding within 1 hour 

(Abbott Labs 1992). 

While strains of B.t. are often characterized as selective pesticides (e.g., Paulus et al. 1999),

various strains of B.t. are active in a large number of lepidopterans (e.g., Peacock et al. 1998)

and are used to control of a variety of lepidopteran pests: spruce budworm (Choristoneura 

fumiferana), eastern hemlock looper (Lambdina fiscellaria), the diamondback moth (Perez et

al. 1997a,b) et al. (Addison and Holmes 1996; Cooke and Regniere 1999; Gloriana et al. 2001;

Masse et al. 2000). The insecticidal potency of B.t. varies depending on the strain of bacteria

and type of insect (Frankenhuyszen et al. 1992, Navon 1993; Peacock et al. 1998). 

Appendix 3 summarizes studies regarding the effects of B.t.k. on lepidopteran species. This 

appendix represents a subset of the most relevant available literature and is not comprehensive. 

As reviewed by Glare and O’Callaghan (2000), there are approximately 1500 reports that assay

the effect of B.t.k. in different lepidopteran species. Some studies, like Miller (1990b) assay

effects as changes in species abundance in non-target lepidoptera after applications of B.t.k. to 

control a pest species. In terms of the ability to characterize risk, however, this risk assessment

focuses on studies that are useful for quantifying effects on non-target lepidoptera as well as

differences in sensitivity among various species of non-target lepidoptera. 

Herms et al. (1997) demonstrate the only dose-response relationships after applications of B.t.k. 

to both target and non-target lepidoptera. In this study, the toxicity of Foray 48B was assayed

in larvae of both the gypsy moth and the Karner blue butterfly, an endangered species of

butterfly indigenous to the northern United States (Minnesota to New Hampshire). Bioassays

in both species involved applications of Foray 48B to vegetation (wild lupine leaves for the

Karner blue and white oak leaves for the gypsy moth) at treatment levels equivalent to either 30

to 37 BIU/ha per ha (low dose) or 90 BIU/ha (high dose). A negative control consisted of st nd nd untreated vegetation. The insect larvae (either 1 or 2 instar for the Karner blue and 2 instar 

for the gypsy moth) were placed on the vegetation 7 to 8 hours after treatment and allowed to

feed for 7 days. Survival rates for Karner blue larvae were: 100% for controls, 27% at the 30 to 

37 BIU/ha treatment rate, and 14% at the 90 BIU treatment rate. Survival rates for gypsy moth

larvae were: 80% for controls; 33% for low-dose treatment, and 5% for high-dose treatment. 

As detailed further in the dose-response assessment (Section 4.3), the differences between the

gypsy moth and Karner blue do not appear to be substantial and the Karner appears to be as

sensitive as the target species to B.t.k. 

The sensitivities of larvae of two species of swallowtail butterflies (Papilio glaucus and 

Papilio canadensis) and the promethea moth (Callosamia promethea) also appear to be similar 

to that of the gypsy moth (Johnson et al. 1995). In the study by Johnson et al. (1995), several

different types of trees (amalanchier, balsam poplar, black cherry, quaking aspen, and white

ash) at several locations were treated with Foray 48B by backpack at a rate of 40 BIU/ha. On st nd the day of treatment or 1 day after treatment, 1 and 2 instar larvae of the test species were 

placed on foliage of the treated trees or untreated trees and mortality was monitored daily for 7

to 8 days. Given this experimental design, mortality could have occurred due to B.t.k. spray, 

natural causes, or predation. No significant differences were observed in mortality among the

different types of vegetation but mortality was significantly and consistently greater on B.t.k. 

treated trees compared with untreated trees. Overall, survival after 8 days was about 30% to

40% in untreated trees and only 6% to 11% in treated trees (Johnson et al. 1995, Table 1, p.

292). Consistent with many other studies —see the review by Glare and O’Callaghan (2000)—

mortality rates tended to be greater in shaded vegetation because of the longer persistence of

B.t.k. In a separate series of studies with Papilio glaucus, significant mortality was noted when

the larvae were placed on shaded vegetation for up to 30 days after the application of B.t.k. As 

4-4
 
discussed by Johnson et al. (1995, p. 292), this is an unusual finding. In most other studies, the 

residual activity of B.t.k. ranges from about 2 to 10 days. One explanation for this effect

offered by Johnson et al. (1995) is that the application by backpack may have resulted in

coverage of both the top and bottom surfaces of the leaves thus increasing the functional

persistence of B.t.k. on vegetation. Johnson et al. (1995, p. 294) also cite preliminary

unpublished bioassay data from their laboratory indicating that swallowtail caterpillars may be

over 100 times more sensitive than the gypsy moth to B.t.k. than the gypsy moth. In the 

absence of detailed data, this statement is difficult to evaluate. As discussed further in the 

dose-response assessment (Section 4.3), the survival rates reported by Johnson et al. (1995) are

consistent with those in the gypsy moth and Karner blue from the study by study by Herms et

al. (1997). 

As noted above, Johnson et al. (1995) detected no significant differences in the toxicity of B.t.k. 

among different types of vegetation. In the forest tent caterpillar (Malacosoma disstria), a

remarkably different pattern is observed with the target species apparently 100 times more

sensitive to B.t.k. contaminated leaves from a secondary host, the sugar maple, compared with 

B.t.k. contaminated leaves from their primary host in north-eastern American, the quaking

aspen (Kouassi et al. 2001). 

James et al. (1993) assayed the toxicity of (Dipel-HG) to both the cinnabar moth (Tyria st th jacobaeae) larvae (1 to 5 instar), a non-target beneficial species, and the cabbage looper 

(Trichoplusia ni), a target species (1st instars). This study involves the treatment of tansy 

ragwort, a pest weed that is consumed by the cinnabar moth, with various concentrations of

B.t.k. equivalent to application rates of 2 to 250 BIU/ha. As summarized in Appendix 2 and

discussed further in the dose-response assessment (Section 4.3), substantial differences were

noted in sensitivity, with early instars of the cinnabar moth being relatively tolerant (LC50 

values of 427 to 575 BIU/ha) and later instars being extremely sensitive (LC50 values of 19 and 

26 BIU/ha). The sensitive instars are about as sensitive to the B.t.k. formulations as the target 

species (LC50 of 16 BIU/ha). 

Not all non-target lepidoptera are as sensitive as the gypsy moth to B.t.k.. By far the most 

complete study regarding the toxicity of B.t.k. to non-target lepidoptera is the publication by 

Peacock et al. (1998). This investigators in this study used two formulations of B.t.k., Foray

48B at a rate equivalent to 89 BIU/ha and Dipel 8AF at a rate equivalent to 99 BIU/ha. Foray

48B was assayed in 42 species from 7 families of lepidoptera and Dipel 8AF in 14 species from

4 families of lepidoptera. Various instars of larvae from each species were exposed to either

control/untreated vegetation or vegetation treated with one of the formulations. Different 

bioassays used either Carya ovata (Shellbark hickory), Juniperus virginiana (Eastern red 

cedar), or Quercus alba (White oak). Larvae were placed on the treated vegetation, and

mortality rates were observed for 5 to 7 days. Some bioassays using Foray were repeated in

different years to assess variability in the potency of different batches of the formulation. The 

results of this study are summarized in Tables 4-1 (Foray formulation) and 4-2 (Dipel

formulation). For both Foray and Dipel formulations, substantial differences in sensitivity

among species and in some cases among families were noted. All species of Nymphalidae

(n=3), Lasiocampidae (n=2), and Saturniidae (n=3) exhibited significant mortality in response

to Foray. As in the study by Johnson et al. (1995), significant mortality was also observed in

Papilo glaucus (Papilionidae). The largest number of species tested were from the Noctuidae

(n=15), and significant mortality was established in only five species. Remarkably similar

results were noted in all of the eight species tested with Foray using the same instar—i.e., the

results were highly reproducible with little indication of substantial variability in the potency of

different batches. The results with Dipel 8AF (Table 4-2) were similar to those with Foray 48B

for nine species and different for only one species, Eupsilia vinulenta. This species appeared to 

4-5
 
be sensitive to Foray 48B in two separate assays but insensitive to Dipel 8AF in one assay. 

This difference is noted by Peacock et al. (1998) but no explanation is offered. The only

apparent difference in the two sets of bioassays is that the Foray assays were conducted on n1/n–2 instars whereas the Dipel assay was conducted only on n–2 instars. Although the use of

only one dose level for each formulation in the study by Peacock et al. (1998) precludes a direct

dose-response assessment, these data can be used to bracket plausible ranges of sensitivity

among non-target lepidoptera, as discussed further in Section 4.3. 

The variability in the response of nontarget lepidoptera to B.t.k. is also illustrated in the recent 

field study by Rastall et al. (2003). In this study, a B.t.k. formulation (Foray 48F) was applied

to two forests (dominated by oak, hickory, and maple trees) over a two year period at an

application rate of 40 BIU/acre. This application rate is equivalent to about 99 BIU/ha,

identical to the upper range of the application rate used in the bioassay study by Peacock et al.

(1998). Rastall et al. (2003) monitored nontarget lepidopteran populations in the two years

prior to application as well as over the two year period in which B.t.k. was applied. The 

response of nontarget lepidoptera varied substantially among different species. Larvae of three 

lepidopteran species were significantly decreased in treatment years: Lambdina fervidaria 

[geometrid], Heterocampa 

guttivitta [notodontid], and Achatia distincta [noctuid]. For 19 other species, larval counts

were significantly higher in treatment years as were the total number of noctuids combined and

the total number of all nontarget lepidopteran species combined. 

4.1.2.3.2. Other Terrestrial Insects – Some non-target lepidopteran species may be as 

sensitive as target species to B.t.k.; however, most studies indicate that effects in other 

terrestrial insects are likely to be minor. As with the non-target lepidopteran species, there is a

large body of literature available on other non-target insects. Most of the open literature is

reviewed in Glare and O’Callaghan (2000), and much of the unpublished literature is reviewed

in U.S. EPA (1998) and Abbott Labs (1992). This risk assessment focuses on those studies 

that suggest some plausible basis for concern in at least some species as well as those studies

that can be used to quantitatively assess sensitivity relative to both target and non-target

lepidoptera (Appendix 4). 

There are no recent published or unpublished studies—i.e., since the preparation of the

previous risk assessment for the USDA gypsy moth program (USDA 1995)—that report

substantial effects in non-target insects, other than lepidoptera, exposed to B.t.k.. Wang et al.

(2000) conducted a field study with Foray 47F on ants and noted no substantial effects on

abundance and species richness, composition, or diversity over a 3-year post-application

period. A slight decrease in abundance was noted in the third year of this study but was

attributed to over-trapping. A substantial and significant decrease in collembolan populations

was noted after the application of Dipel 8L that resulted in soil concentrations 1000 times

greater than expected environmental concentrations (Addison and Holmes 1995). Dipel 4L is

an oil-based formulation and the decrease in collembolan populations was also seen with the

oil blank—i.e., the formulation inerts without B.t.k. Since the effect was not seen with Dipel 8

AF (which does not contain oil) or with unformulated B.t.k., the effect on collembolan 

populations was attributed to the oil carrier rather than B.t.k. It should be noted that Dipel 4L 

is not used in USDA programs. As indicated in Section 2 (Program Description), only one oilbased formulation is used, Dipel ES, and no data regarding the toxicity of this formulation was

encountered in the literature. As indicated in the risk characterization (Section 4.4), however,

it is likely that any oil-based formulation could pose an increased risk to non-target species. 

Other recent studies on B.t.k. either report no effects in non-target species (e.g., Mohaghegh et 

4-6
 
al. 2000) or are studies designed to assess the efficacy of B.t.k. in other pest species (Robacker 

et al. 1996). 

One of the very few studies to report dose-related adverse effects in a non-target species is the

early study by Haverty (1982). In this study, direct spray of lady beetles (Hippodamia 

convergens) and green lacewing (Chrysopa carnea) adults or larvae at rates equivalent to 79

and 158 BIU/ha resulted in slight but significant increases in mortality. Although this study

also involved the use of Dipel 4L, mortality was not attributable solely to the oil carrier

(Haverty 1982). As discussed further in the dose-response assessment, the rates of mortality

observed in these species are consistent with those of B.t.k. in relatively tolerant non-target 

lepidoptera. 

Honey bees are an important non-target insect for any pesticide, and bioassays on honey bees

are required of all pesticides during the registration process. As noted by U.S. EPA (1998), the

bioassays in honey bees submitted in support of the registration of B.t.k. suggest: “minimal 

toxicity for B. thuringiensis subspecies kurstaki” (U.S. EPA 1998, p. 21). This conclusion is 

also consistent with numerous laboratory bioassays and field studies concerning the effects of

B.t.k. (Glare and O’Callaghan 2000; WHO 1999). 

The current risk assessment does not substantially dispute these conclusions. Nonetheless, one 

of the studies cited by U.S. EPA (1998— i.e., Atkins 1991a cited as MRID 419835-01 on p. 19

of the EPA document) suggests that bees may be somewhat more sensitive than some nontarget lepidoptera to B.t.k. exposure. In the study by Atkins (1991a), adult worker honey bees 

(Apis mellifera) were exposed to a dry flowable powder formulation of B.t.k. (14.52 BIU/lb) at

deposition rates of 0 (control), 7.735, 15.470, and 23.205 μg/bee and these rates were

equivalent to 0, 0.70, 1.4, and 2.1 lbs/acre. These application rates correspond to 0, 1.73, 3.45,

or 5.19 lb/ha [1 acre = 0.4047 ha]. Given the potency of 14.52 BIU/lb, these application rates

correspond to 25, 50, and 75 BIU/ha. As indicated in Appendix 4, these exposures resulted in

mortality rates of 7.17 % (control), 18.96% (low exposure), 25% (mid exposure), and 24.91%

(high exposure). As discussed in the dose-response assessment, these response rates are greater

than the responses rates expected in relatively tolerant non-target lepidoptera. 

4.1.2.3.3. Other Terrestrial Invertebrates – There is relatively little information 

regarding the toxicity of B.t.k. or its formulations to other terrestrial invertebrates. An early

report by Benz and Altweg (1975) found no statistically significant effects (compared with

water treated plots) on mixed earthworm populations over a period of about 8 weeks (May 5 to

July 7) after the application of an older Dipel formulation (not otherwise specified) and a

"Bactospeine" formulation of B.t.k. after soil applications equivalent to 1X, 10X, and 100X of 

the recommended application rates. Both Dipel 8AF (water-based formulation) and Dipel 8L

(oil-based formulation) were tested at 1000X the expected environmental concentration 3 (EEC)— i.e., 1.2 L/cm in soil—by Addison and Holmes (1996) in a microcosm study using

earthworms (Dendrobaena octaedra). Dipel 8AF caused no effect on earthworm populations

over a 10-week observation period; however, Dipel 8L and the oil blank (i.e., the formulation

without B.t.k.) caused decreased growth, greater than 50% mortality of the worms, and a

decrease in the number of viable cocoons by week 6. Based on these results, Addison and 

Holmes (1996) further assayed Dipel 8L at 1X, 10X, 100X, and 1000X EEC. A significant 

reduction in survival, growth, and cocoon production was noted at 1000X EEC but no 

significant adverse effects on survival, growth, or reproduction were noted at 10X or 100X

EEC. As discussed in Section 4.1.2.3.2 regarding effects on collembolan populations, the

toxicity of Dipel 8L appeared to be related to the oil used in the formulation rather than to B.t.k. 

4-7
 
 

4.1.2.4. Terrestrial Plants (Macrophytes) – As indicated in the re-registration eligibility 

document on B.t. (U.S. EPA 1998) , toxicity testing in non-target plant species was not

required to support the re-registration of products containing B.t. because “...a review of the 

literature on B. thuringiensis and its byproducts indicate no known detrimental effects on plant

life...”(U.S. EPA, 1998, p. 25). No information was found in the more recent literature 

regarding the toxicity of B.t.k. or its formulations to plants, suggesting that effects on plants are

not likely and that the phytotoxicity of B.t.k. has not generated substantial interest. As 

reviewed by Glare and O’Callaghan (2000, p. 52), some lepidopteran species are used as

biological control agents for weeds—such as the cinnabar moth (Tyria jacobaeae) to control

ragweed. As discussed in Section 4.1.2.3.1 and detailed further in the dose-response

assessment (Section 4.3), late instars of this species appear to be sensitive to B.t.k. and the use 

of B.t.k. could have secondary effects on the control of some weed species. It is likely, 

however, that the main impact of B.t.k. when used to control the gypsy moth will be in

minimizing damage to terrestrial plants that would otherwise be damaged by gypsy moth

infestations. 

4.1.2.5. Terrestrial Microorganisms – There are relatively few studies regarding the effects of 

B.t.k. applications on terrestrial microorganisms. At exposure levels equivalent to 100X of the 

typical application rate for B.t.k. strain A20, Bernier et al. (1990) noted no effect on other soil 

microorganisms. At the recommended the rate, Dipel 176 (another oil-based formulation of 

B.t.k.) caused no effects on cellulose degradation, microbial biomass, or microbial respiration. 

At 1000X of the normal application rate, nitrite and ammonia metabolism were reduced and

microbial biomass and respiration were increased after 8 weeks. As noted by Glare and 

O’Callaghan (2000), these effects could have been due either to B.t.k. germination or the effect 

of the oil in the formulation. 

4.1.3. Aquatic Organisms.

4.1.3.1. Fish – As summarized in the previous USDA (1995) risk assessment on B.t.k., field 

studies (Buckner et al., 1974; Otvos and Vanderveen 1993; Surgeoner and Farkas 1990) report

no apparent fish kills or other adverse effects resulting from the use of B.t.k. Similarly, U.S. 

EPA (1998) classifies B.t.k. as virtually non-toxic to fish, based on an assessment of several

acute toxicity studies in trout and one study in bluegills. These conclusions are consistent with 

a relatively large number of experimental studies that report very few if any effects in fish at

much higher concentrations than would be encountered in the environment after the use of

B.t.k. (Appendix 5). Acute exposure to B.t.k. formulations at concentrations up to 1000 mg/L

are not associated with fish mortality (e.g., Meher et al. 2002), and longer-term studies of

formulated B.t.k. in bluegills (Christensen 1990c), sheepshead minnow (Christensen 1991e)

and trout (Christensen 1990d,i) report only decreased growth at concentrations up to 40,000X

expected environmental concentrations. 

The only suggestion of an adverse effect in fish is from the study by Martin et al. (1997). 

These investigators report an unexplained fish kill in Maryland after the application of B.t.k. 

In addition, these investigators conducted bioassays on Koi carp (Cyprinus carpio) at 1X and 

10X ECC via food and water in experimental tanks for 32 days. The only adverse effects 

reported were changes in fish weight and plasma protein values. The Martin et al. (1997)

report, however, is only an abstract and a full publication of this study was not found in the

literature. Given the sparse detail in the abstract, it is difficult to interpret the significance of 

this study. No further information found regarding the fish kill purportedly associated with 

B.t.k., and the information summarized in Appendix 5 as well as the information reported by

Martin et al. (1997) do not support the contention that fish would be killed following the

application of B.t.k. 

4-8
 
4.1.3.2. Amphibians – There is available information regarding the toxicity of B.t.k. or B.t.k. 

formulations to amphibians. Other strains of B.t., specifically B.t. israelensis and B.t. 

tenebrions, appear to have a very low toxicity to amphibians (Glare and O’Callaghan 2000;

WHO 1999). 

4.1.3.3. Aquatic Invertebrates – As summarized in Appendix 6, the effects of B.t.k. on aquatic

invertebrates was investigated in both standard laboratory studies as well as a number of field

studies. At concentrations sufficiently high to cause a decrease in dissolved oxygen or an

increase in biological oxygen demand, B.t.k. exposure may be lethal to some aquatic 

invertebrates such as Daphnia magna (e.g., Christensen 1991d; Young 1990). Most organisms, 

however, seem relatively tolerant even to concentrations of B.t.k. in water that are up to

200,000 times higher than expected environmental concentrations (Christensen 1991f). Black 

fly larvae may be somewhat more sensitive than most other aquatic invertebrates to B.t.k. (Eidt

1985). Nevertheless, as discussed by Glare and O’Callaghan (2000), the different studies are

difficult to compare with one another and some are difficult to relate to plausible

environmental exposures because of different units in which exposures are expressed. 

Several field studies (e.g. Kreutzweiser et al. 1992, 1993, 1994; Richardson and Perrin 1994)

do not report remarkable effects in most species exposed to B.t.k. at levels that exceed expected 

environmental concentrations (EEC) by factors of up to 100. Possible exceptions may be 

stonefly larvae and mayfly larvae. Kreutzweiser et al. (1993, 1994) did note increased drift in

decreased populations of stonefly larvae (Leuctra tenuis) at application rates equivalent to 10X 

EEC. After applications of B.t.k. at rates of 50 to 5000 BIU/ha over streams, Richardson and

Perrin (1994) noted increased drift only in stonefly larvae. 

U.S. EPA (1998) raises concerns that some batches of B.t. may contain heat labile exotoxins 

that are toxic to Daphnia. The production of these toxins is apparently not well understood and

seems to be an atypical event probably associated with abnormal or poorly controlled

production processes. U.S. EPA (1998) does not require daphnid testing of each commercial 

batch of B.t.; instead, the Agency requires manufacturers to submit a daphnid study on each

new manufacturing process to demonstrate that heat labile exotoxin levels are controlled. 

4.1.3.4. Aquatic Plants – The toxicity of B.t.k. to aquatic plants has not been tested because of

the lack of information suggesting that adverse effects in aquatic plants are plausible (U.S. EPA

1998, p. 30). No relevant data that would call this judgement into question were found in the

available literature. 

4-9
 
4.2. EXPOSURE ASSESSMENT 

4.2.1. Overview. 

The exposure assessment for the ecological risk assessment on B.t.k. are summarized in Table 

4-3. Exposure assessments, based on the hazard identification, are presented for three groups:

small mammals, terrestrial insects, and aquatic species. Although numerous exposure

scenarios could be developed for terrestrial mammals, the only positive hazard identification

for B.t.k. involves inhalation exposures. The ecological risk assessment uses inhalation 3 exposure levels of 100 to 5000 cfu/m , which is the same range used in the human health risk

assessment, to assess potential risks of serious adverse effects in terrestrial vertebrates. These 

concentrations are applied to a 20 g mouse and correspond to inhaled doses of 0.00336 to 0.168

cfu/mouse. While there is no credible basis for asserting that terrestrial invertebrates are likely

to have adverse effects after oral or dermal exposure to B.t.k., an extremely conservative

exposure assessment is developed for combined oral (water and vegetation) and dermal (direct

spray) exposures that yields an estimated maximum dose of approximately 184 mg/kg body

weight. For terrestrial insects, the toxicity values used to assess the consequences of observing

effects is given in units of BIU/ha. Consequently, the exposure assessment for this group is

simply the range of application rates used in USDA programs—i.e., about 49 to 99 BIU/ha. 

For aquatic organisms, toxicity data are expressed in several different units, including mg

formulation/L, IU/L, and cfu/L. Based on application rates used in USDA programs and

conservative assumptions concerning the depth of water over which B.t.k. might be sprayed,

concentrations in water are expected to be less than or equal to 0.24 mg formulation/L. As 

discussed in the hazard identification, there is no basis for concern about adverse effects in 

birds, plants, soil microorganisms or invertebrates ,other than insects, exposed to B.t.k. Hence, 

explicit exposure assessments for these groups are not conducted. 

4.2.2. Terrestrial Animals. 

4.2.2.1. Terrestrial Vertebrates – Terrestrial animals might be exposed to any pesticide from

direct spray, contact with contaminated media (vegetation, water, soil), the ingestion of

contaminated media (vegetation, prey species, or water), or inhalation. Although numerous

exposure scenarios could be developed for each of these types of exposure, the only positive

hazard identification for B.t.k. involves inhalation exposures (see Section 4.1.2.1). As in the 

human health risk assessment (Section 3.4), inhalation exposures of 100 to 5000 cfu/m3 are 

used to assess potential risks of serious adverse effects in terrestrial vertebrates. 

The characterization of the potential risk from inhalation exposure is based on the cumulative

exposure, which is expressed in units of cfu/organism, as in the human health risk assessment. 

The toxicity data are taken from laboratory studies involving B.t.k. exposure to mice

(Hernandez et al. 1999, 2000). In terms of the exposure assessment, the mouse is an

appropriate species on which to base the risk assessment because mice and other small

mammals have a higher breathing rate per unit body weight, compared with larger animals. As 3 noted in Table 3-7, the breathing rate for a 20 g mouse is approximately 0.0000014 m /hour. 3 Taking the concentrations of 100 to 5000 cfu/m and using a 24-hour exposure period (as in the

human health risk assessment), the total cumulative exposure for a 20 g mouse ranges from 3 3 0.00336 to 0.168 cfu/mouse [100 to 5000 cfu/m × 0.0000014 m /hour × 24 hours]. This 

cumulative exposure is used directly in the risk characterization (Section 4.4). 

Although there is no credible evidence that oral or dermal exposure to B.t.k. is likely to cause

adverse effects in terrestrial vertebrates, an extremely conservative exposure assessment for

these routes of exposure can be developed. As noted in Section 4.1.2.1 and discussed further in 

the dose-response assessment (Section 4.3) and risk characterization (Section 4.4), free

standing NOAELs are available for B.t.k. formulations in mammals, which are expressed in 

4-10
 
units of mg formulation/kg body weight/day. The underlying assumption in this exposure

scenario is that a small mammal consumes contaminated vegetation and contaminated water

after having been sprayed directly with B.t.k. over its entire body surface. 

The major routes of oral exposure are the consumption of contaminated vegetation and

contaminated water. Initial residues on vegetation are determined by the type of vegetation and 

application rate. Fletcher et al. (1994) indicate that the highest residues are will be found on

short grass—i.e., 240 mg/kg vegetation at an application rate of 1 lb/acre. As detailed in Table 

2-1, the highest application for any B.t.k. formulation is 2 lbs/acre. Thus, the highest initial

residues on vegetation are expected to be approximately 480 mg/kg on vegetation. General 

allometric relationships dictate that smaller animals, because of their higher metabolic rates,

consume more food than do larger animals. Based on allometric relationships between food

consumption and body weights for rodents (U.S. EPA/ORD 1993, p. 3-6), a small mammal

weighing approximately 20 g will consume about 3.5 g of food per day. Thus, if a small 

mammal were to consume vegetation recently sprayed with a B.t.k. formulation, the dose to the 

animal would be about 84 mg/kg [0.480 mg/g vegetation × 3.5 g ÷ 0.02 kg]. 

An extremely conservative estimate of the dose from contaminated water can be derived in a

similar way. Based on allometric relationships for mammals from U.S. EPA/ORD (1993, Eq.

3-17, p. 3-10), a small mammal will consume about 3 mL of water per day. As noted above, 

the highest application rate for any B.t.k. formulation is 2 lbs/acre, which corresponds to 

224.2 mg/m . 2 Under the assumption that the B.t.k. formulation is sprayed over a shallow (1 cm 2 deep) puddle with a surface are of 1 square meter or 10,000 cm , the volume of water equals

10,000 mL and the initial concentration of the B.t.k. in the water is approximately 0.022 

mg/mL [224.2 mg ÷ 10,000 mL]. Thus, the B.t.k. dose to the 20 g mammal is approximately 

3.3 mg/kg [0.022 mg/mL × 3 mL ÷ 0.02 kg]. 

As a final component of this extreme exposure assessment, assume that the small mammal is

sprayed directly with the B.t.k. formulation. Again using allometric relationships developed by

U.S. EPA (U.S. EPA/ORD 1993, eq. 3-22, p. 3-14), a 20 g mammal has a surface area of about 2 2 0.0086509 m . Thus, at an application rate of 2 lbs/acre or 223.4 mg/m , the maximum dose 2 that could be deposited on a 20 g mammal is about 97 mg/kg body weight [224.2 mg/m × 2 0.0086509 m ÷ 0.02 kg]. It is, of course, somewhat implausible to assume that the complete 

body surface will be covered by a direct spray; however, this calculation is maintained as an

extremely conservative assumption. Furthermore, it is not reasonable to assume that the 

deposited dose will be absorbed. Nonetheless, one of the underlying assumptions for this

conservative exposure assessment is that grooming by the small mammal results in the

ingestion of the entire amount of B.t.k. formulation deposited on the mammal. 

Combining these three routes of exposure, the total dose to the animal is approximately 184

mg/kg body weight [84 mg/kg + 3.3 mg/kg + 97 mg/kg = 184.3 mg/kg bw]. 

4.2.2.2. Terrestrial Invertebrates – As discussed in Section 4.1.2.3 (Hazard Identification for

Terrestrial Invertebrates) and addressed further in Section 4.3 (Dose-Response Assessment),

some terrestrial invertebrates, particularly lepidoptera, appear to be as sensitive to B.t.k. as the 

gypsy moth and other target species. While the dose-response assessment is somewhat

elaborate, it is based on exposure units of BIU/acre or ha; thus, the exposure assessment is

relatively simple—i.e., expressed in units of application rate. As indicated in Section 2.2, the 

application rates considered in this risk assessment are 20 to 40 BIU/acre, which are equivalent

to about 49 to 99 BIU/ha. 

4-11
 
A noteworthy reservation about using an application rate as a measure of exposure is that most

of the toxicity studies do not involve field observations. Instead, different types of vegetation

are treated in a manner equivalent to and expressed as an application rate, most often in units of

BIU/ha. Thus, the effects of drift and canopy interception are not encompassed by the toxicity 

studies. This issue is addressed in the risk characterization (Section 4.4). 

4.2.2.3. Other Terrestrial Species – As discussed in the hazard identification, there is no 

plausible basis for concern regarding adverse effects in birds (see Section 4.1.2.2), plants (see

Section 4.1.2.4), soil microorganisms (see Section 4.1.2.5) or invertebrates other than insects

(see Section 4.1.2.3.3) after exposure to B.t.k.. Thus, as with the previous USDA risk

assessment (USDA 1995), explicit exposure assessments for these species are not conducted. 

The only reservation with this approach involves the used of oil-based formulations. This 

concern is addressed qualitatively in the risk characterization (Section 4.4). 

4.2.3. Aquatic Organisms.

As illustrated in Appendix 5 (Toxicity to Fish) and Appendix 6 (Toxicity to Aquatic

Invertebrates), toxicity data are expressed in several different units. Some field studies (e.g.,

Richardson and Perrin 1994), exposures are expressed application rates. Other studies report

exposures as concentrations in units of mg formulation /L (e.g. Meher et al. 2002; Mayer and

Ellersieck, 1986) and still other studies report exposures in units of cfu/L (e.g., Christensen

1990c,d) or IU/L (Eidt 1985). As noted by Glare and O’Callaghan (2000), this diversity of

units impairs the ability to compare different studies. Nonetheless, as discussed further in the 

dose-response assessment (Section 4.4), the key toxicity values given in IU/L can be converted

to units of mg formulation/L, which are the most useful units of measure for risk

characterization. 

The same approach can be used to derive conservative estimates of B.t.k. concentrations in 

water, expressed in units of mg of formulation/L, as was used to estimate exposure

concentrations for a terrestrial mammal (see Section 4.2.2.1). For the mammal a depth of 1 cm

was used to estimate an extreme worst-case concentration, which is not a reasonable 

assumption for exposure scenarios involving aquatic species. The U.S. EPA typically uses a 

water depth of 6 feet. Because of the apparently low potential for adverse effects, however, the

U.S. EPA (1998) did not conduct an explicit exposure assessment on aquatic species. Most 

Forest Service risk assessments use a somewhat more conservative water depth of 1 m or about

3 feet, and this is the depth used to calculate a plausible concentration of B.t.k. formulation in 

water immediately after a direct spray of B.t.k. at an application rate of 2 lbs/acre or 2 3 224.2 mg/m . At a depth of 1 m, 244.2 mg of formulation would be deposited into 1 m of 

water which is equivalent to 1000 L. Assuming instantaneous mixing, the concentration in

water would be about 0.24 mg formulation/L [244.2 mg ÷ 1000 L]. 

For toxicity studies that are expressed in units of IU/L, the concentration of 0.24 mg

formulation/L can be converted using IU/mg formulation values given in Table 2-1. The 

highest value is 32,000 IU/mg —reported for a number of formulations including Biobit HP,

DiPel DF, and DiPel Pro DF. Thus, the concentration of 0.24 mg formulation/L corresponds to

7680 IU/L or 7.6 IU/mL [0.24 mg formulation/L × 32,000 IU/mg]. 

Some aquatic toxicity data are expressed in units of cfu/L, and these data cannot be converted

readily to other units of exposure. Measurements of B.t.k. formulations are not expressed in 

units of cfu/mg formulation. Consequently, these units of measure are not relevant to those 

involved in the application of B.t.k. formulations. As an alternative, the monitoring study by

Menon and De Mestral (1985) can be used to approximate plausible concentrations of B.t.k. in 

4-12
 
water in terms of cfu/L. In this study, an older formulation of B.t.k., Thuricide 16B, was 

applied at rates of 4.7 to 9.4 L/ha. Concentrations in river water ranged from 22 to 63 cfu/mL 

or 22,000 to 63,000 cfu/L. Menon and De Mestral (1985) do not report the potency of 

Thuricide 16B. Assuming that the nomenclature for Thuricide 16B is the same as that for the

current Thuricide formulations, it is assumed that the Thuricide 16B formulation had a potency

of 16 BIU/gallon. Thus, an application of 4.7 L/ha corresponds to application rate of

approximately 8 BIU/acre [4.7 L/ha × 0.2642 gallon/L × 16 BIU/gallon × 0.4047 acres/ha =

8.0405 BIU/acre], and 9.4 L/ha corresponds to twice that amount or about 16 BIU/acre. It is 

not clear from the publication by Menon and De Mestral (1985) whether the reported cfu/L

concentrations were associated with applications of 4.7 L/ha or 9.4 L/ha. For this component

of the exposure assessment, it is assumed that the reported concentrations were associated with

an application of 4.7 L/ha or 8 BIU/acre. In addition, the upper range of 63,000 cfu/L is used to

calculate a water contamination rate of 7875 cfu/L per BIU/acre [63,000 cfu/L ÷ 8 BIU/acre]. 

As noted in Table 2-1, the maximum application rate of B.t.k. recommended for the control of 

the gypsy moth is 40 BIU/acre. Thus, the expected maximum concentration of B.t.k. in water is 

3.15×105 cfu/L [7875 cfu/L per BIU/acre ×40 BIU/acre = 315,000 cfu/L]. 

Notice that this estimate of B.t.k. in water expressed as cfu/L is based on the most conservative

set of assumptions from the study by Menon and De Mestral (1985) and may grossly

overestimate actual exposure. The magnitude of the potential overestimation can be evaluated

using the more recent monitoring study by Valadares de Amorin et al. (2001), in which B.t.k. 

concentrations in reservoirs were monitored after three applications of B.t.k. (Foray 48B) at a 

rate of 20 BIU/acre. The maximum number of B.t.k. colonies monitored by Valadares de

Amorin et al. (2001) was 200 cfu/L (see Valadares de Amorin et al. 2001, Table 4, p. 1041). 

4-13
 
4.3. DOSE-RESPONSE ASSESSMENT 

4.3.1. Overview. 

The toxicity values used in the ecological risk assessment are summarized in Table 4-4. The 

dose-response assessment parallels the exposure assessment. Specific dose-response

assessments are presented for three groups: small mammals, terrestrial insects, and aquatic

species, both fish and aquatic invertebrates. For small mammals, dose-response assessments 

are given for inhalation and oral exposure. The risk assessment for inhalation exposure is

based a study in which mortality increased in mice exposed to B.t.k. via intranasal instillations 

of the agent. A dose of 107 cfu/mouse is taken as the NOAEL, and 108 cfu/mouse is taken as a 

frank effect level—a dose associated with 80% mortality. The risk assessment for oral 

exposures is based on a free-standing NOAEL, which implies that oral exposure to B.t.k., 

however high the concentration, will not cause adverse effects in mammals or birds. For this 

risk assessment, the dose of 8400 mg/kg/day is used as the NOAEL. For terrestrial 

invertebrates, sufficient data are available to estimate dose-response relationships for sensitive

species and relatively tolerant species. Sensitive species, which consist largely of lepidoptera, 

have an LD50 value of about 21 BIU/ha. Tolerant species, comprised of some lepidoptera and

other kinds of terrestrial insects, have an LD50 value of about 590 BIU/ha, which is 

approximately 28 times greater than the LD50 value for sensitive species, The dose-response

curves developed for sensitive and tolerant species permit mortality estimates for any

application rate. As with terrestrial insects, dose-response assessments are developed for

tolerant and sensitive species of fish and aquatic invertebrates. Fish appear to somewhat less 

sensitive than invertebrates to B.t.k. exposure. For tolerant species of fish, the NOEC of 1000 

mg/L, which corresponds to 2.5×1010 cfu/L, is taken from a study in mosquito fish. For 

sensitive species of fish, the LOEC is based on a trout study in which marginally significant

mortality was observed at 1.4 mg/L or about 2.87×107 cfu/L. The most sensitive invertebrate 

species appears to be Daphnia magna, with a chronic NOEC of 0.45 mg/L or 6.24×108 cfu/L 

for both reproductive effects as well as mortality. The NOEC for tolerant species is taken as 36

mg/L based on bioassays in mayflies and caddisflies. 

4.3.2. Toxicity to Terrestrial Organisms.

4.3.2.1. Terrestrial Vertebrates – As discussed in Section 4.2.2.1, two sets of exposure

assessments are used for terrestrial vertebrates: inhalation exposures expressed in units of 3 cfu/m and oral exposures (including ingestion by grooming of material deposited on body

surface) in units of mg formulation/kg body weight. These two types of exposures represent 

very different potential risks. More precisely, the assessment of the risk from inhalation

exposure is based on the study by Hernandez et al. (2000) in which mortality in mice was

observed after intranasal instillations of B.t.k. The assessment of oral exposures, on the other 

hand, is based on a free-standing NOAEL. 

As discussed in Section 3.3.4, using the study by Hernandez et al. (2000) to assess the potential

risks from inhalation exposures is a tenuous and probably extremely conservative approach—it 

tends to overestimate risk. Notwithstanding this limitation, it is the best available study from

which the potential for serious adverse effects can be assessed. As in the human health risk 

assessment, a dose of 107 cfu/mouse is taken as the NOAEL and 108 cfu/mouse is taken as a 

frank effect level—a dose associated with 80% mortality. 

As discussed in Section 4.1, adverse effects were not observed in mammals or birds after oral 

exposure to B.t.k.. Long-term doses up to 8400 mg/kg/day do not appear to cause adverse

effects in mammals (McClintock et al. 1995b), and multiple oral doses up to 2857 mg

formulation/kg bw are not associated with adverse effects in birds (Lattin et al. 1990a,b,d). For 

4-14
 
this risk assessment, the dose of 8400 mg/kg/day is used as the NOAEL and is compared wth

the exposure assessment developed for the small mammal (see Section 4.2.2.1). 

4.3.2.2. Terrestrial Invertebrates – For terrestrial invertebrates, sufficient data are available to 

estimate dose-response relationships for sensitive species as well as relatively tolerant species. 

The data used in these analyses are summarized in Table 4-5. The sensitive species are all

lepidoptera, and all of the studies used in the analysis involve feeding various lepidopteran

larvae with vegetation treated with various B.t.k. formulations at rates that can be expressed in 

units of BIU/ha. Seven species of lepidoptera are included: two target species (the gypsy moth

and cabage looper) and five non-target species (the Karner blue butterfly, two species of

swallowtail butterfly, the promethea moth, and late instars of the cinnabar moth). The tolerant 

species used in the dose-response assessment involve feeding of early instar cinnabar moth

larvae as well as direct spray of non-lepidopteran insects: green lacewing adults as well as

larvae and direct spray of adult lady beetles. Details of these studies are presented in Section 

4.1.2.3. 

The analysis of these data is somewhat more elaborate than that in other sections of this risk

assessment both because the data are sufficient for a more elaborate analysis and because the

analysis is important. In plain language, the analysis derives dose-response relationships for

both sensitive and insensitive species—i.e., estimates of mortality can be made for any 

application rate. Sensitive species have an LD50 value of about 21 BIU/ha and consist entirely 

of lepidoptera. The tolerant species have an LD50 of about 590 BIU/ha, which is approximately 

28 times greater than the LD50 value for sensitive species. The details of these analyses are 

provided in the remainder of this section. 

In Table 4-5, which summarizes the data used in the dose-response assessment for non-target

insects, the first column specifies the common name of the test organism. This column is 

followed by the application rate in units of BIU/ha, the mortality rate (as a proportion of

organisms) observed in control organisms not exposed to B.t.k., and the mortality rate (again as 

a proportion) in treated organisms. The fifth column gives the mortality rate attributable to 

B.t.k. considering the control response. This rate is calculated using Abbott's formula: 

P = (P* - C) / (1 - C) 

where P is the proportion responding that is attributable to the agent, P* is the observed 

proportion responding in the group exposed to the agent, and C is the proportion responding in 

the control group (Finney 1972, p. 125). This is a common method used to adjust mortality

rates and assumes that the causes of mortality in the control group are independent of mortality

attributable to the agent under study. As noted by Finney (1972), this is the standard approach

for calculating the probability of combinations of independent events. 

For statistical analysis, the probit model was used, which is similar to the approach taken in the

analysis of the mortality data from Hernandez et al. (2000) in Section 3.3.4. Because different 

studies are combined, each with different control response rates, standard probit analysis was

not used. Instead, the responses attributable to B.t.k. based on Abbott’s formula were 

converted to probits using the inverse normal function in EXCEL: 

Probit = 5 + NORMINV(P,0,1) 

where 0 and 1 are the mean and standard deviation of the standard normal curve, and P is as 

defined above. The constant of 5 is the standard constant for converting normal equivalent 

4-15
 
deviates to probits. Thus, a probit of 5 represents a response of 50%, a probit of 6 represents a

response that is one standard deviation above 50% (i.e., a response of about 82%), a probit of 7

represents a response that is two standard deviations above 50% (i.e., a response of about 98%)

and so on. 

While it is beyond the scope of this risk assessment to discuss the probit transformation in

detail, this transformation is simply a method to linearize the proportion responding under the

assumption that the distribution of tolerances in a population (in this case the population of

insects) has a log-normal distribution. Further details regarding the biological and statistical

rationale for the probit transformation are provided in Finney (1972, p. 8 ff). 

Using this transformation, the probit responses (independent variable) and log10 BIU/acre are 

used to estimate the linearized dose-response function: 

Y = a + bx 

using standard linear regression where Y is the probit response, x is the log10 of the BIU/acre 

treatment, b is the slope of the dose-response curve, and a is the intercept. 

The log-dose probit-response model provides a statistically significant fit to data for the 2 2 sensitive (p.0.0004, adjusted r = 0.79) and the tolerant (p.0.00003, adjusted r = 0.95) 

species. In addition, the slopes of the dose-response curves are similar and not significantly

different—i.e., 1.95 with a 95% confidence interval of about 1.2 to 2.7 for sensitive species and

2.6 with a 95% confidence interval of about 2.1 to 3.2 for tolerant species. 

Consequently, the regression analysis was run a second time using a variable, S, assigned a

value of 1 for sensitive species and 0 for tolerant species in order to constrain the slopes of the

two curves to be equal: 

Y = a + bx + cS 

where c is the coefficient for the sensitivity variable, S, and the other terms are as defined 

above. 

The data on both sensitive and tolerant species combined fits the following model: 

Y = -1.48 + 2.34 x + 3.36 S 

-11 2 with a highly significant p-value (8.4×10 ) and an adjusted r of about 0.95—i.e., the model

explains 95% of the variability in the data ,and the probability that the association occurred by

random chance is about 1 in 11 billion. It is worth noting that the p-value for the variable for 

sensitivity is about 2.8×10-11, indicating a highly significant difference between the sensitive

and tolerant species—i.e., the probability that the apparent difference occurred by chance is

about 1 in 36 billion. 

The above equation can be used to calculate the LD50 values for both tolerant and sensitive 

species in order to quantify relative potency, defined as the ratio of equitoxic doses. For 

sensitive species, this is done by setting Y equal to 5 and S equal to 1. With these 

substitutions, the value of x, the log BIU/ha, is about 1.33, corresponding to an LD50 of 21 

BIU/ha [10 1.33]. For tolerant species, the log of the LD50 is calculated by setting Y equal to 5

and S equal to 0 to yield a log BIU/ha of about 2.77, corresponding to an LD50 of about 590 

4-16
 
BIU/ha [10 1.33]. Thus, the relative potency of B.t.k. to sensitive species is about 28, relative to 

tolerant species [590 BIU/ha ÷ 21 BIU/ha]. 

Figure 4-1 also contains data from the study in honey bees by Atkins (1991a) and data from

Peacock et al. (1998) on a number of different non-target lepidoptera exposed to Foray 48B at

89 BIU/ha (Table 4-1 of this risk assessment) and Dipel 8AF at 99 BIU/ha (Table 4-2 of this 

risk assessment). In Peacock et al. (1998) study, several of the bioassays resulted in either 0%

or 100% mortality. Neither of these values can be directly translated to probits. Thus, working

probits of 3 were used for 0% mortality and working probits of 7 were used for 100%

mortality, which reflect the approximate range of probit values from Peacock et al. (1998) in

which partial mortality was observed. These values are used only to illustrate the data and

were not used in any statistical analyses. 

Figure 4-1 illustrates how the models fits the available data on sensitive and tolerant species. 

It is apparent from Figure 4-1 that the variability in sensitivity among the lepidopteran species

reported by Peacock et al. (1998) is encompassed by the dose-response curves for sensitive and

tolerant species derived from the data in Table 4-5, although the use of working probits for 0%

and 100% mortality may obscure some of the more or less sensitive species. Given the 

available data, this apparent confusion cannot be avoided. As illustrated in Figure 4-2, the

number of insensitive species (n=16) is somewhat greater than the number of sensitive species

(n=10). Most species (n=28) appear to have intermediate sensitivity which is nearly uniformly

distributed between that of sensitive and insensitive species. This figure is constructed by

combining the data on both Foray 48B (Table 4-1 of this risk assessment) and Dipel 8AF 

(Table 4-2 of this risk assessment). Although the data on bees by Atkins (1991a) is also

encompassed by the two dose-response curves, the slope of the dose-response relationship for

bees appears to be more shallow than that of either dose-response curve. 

In the context of this analysis, the designations of sensitive and tolerant species are not

intended to imply absolute ranges on tolerance among all possible insects. Instead, the analysis

simply indicates that some non-target species, such as the Karner blue butterfly and cinnabar

moth, appear to be as sensitive to B.t.k. as target species such as the gypsy moth and cabbage 

looper. As illustrated in the data from Peacock et al. (1998), the range of sensitivities among

various insect species appear to follow a continuum and it is possible that some species may be

more or less sensitive to B.t.k. than indicated by the two dose-response curves illustrated in 

Figure 4-1. 

4.3.3. Aquatic Organisms

4.3.3.1. Fish – With the exception of the recent publication by Meher et al. (2002), the

detailed studies regarding the toxicity of B.t.k. and B.t.k. formulations are unpublished. These 

studies are summarized Appendix 5, which also summarizes data from secondary sources

(Abbott Labs 1992; Mayer and Ellersieck 1986) and from the abstract by Martin et al. 1997. 

As discussed in Section 4.1.3.1, the study by Martin et al. (1997) is the only report of adverse

effects on fish at concentrations that might result from the application of B.t.k. As further 

discussed in Section 4.1.3.1, this report is only in abstract form and a full publication of the

study was not found in the literature. The results reported in the abstract are inconsistent with

those reported in several more detailed full studies. Consequently, the information reported by

Martin et al. (1997) is not used in the dose response assessment for fish. Similarly, the

secondary sources (Abbott Labs 1992; Mayer and Ellersieck 1986) do not provide sufficient

detail to evaluate the information reported. Given the availability of detailed primary studies 

on B.t.k. (Meher et al. 2002; Christensen 1990c,d,g,i), information from these secondary

sources are not used in the dose-response assessment. 

4-17
 
 

The study by Meher et al. (2002) involves a standard acute (96-hour) bioassay in mosquito fish

at concentrations ranging from 200 to 1000 mg formulation/L. The study reports that the 7 formulation contained 2.5×10 spores/mg. Assuming that the spores are viable, this range of 9 10 concentrations corresponds to 5×10 to 2.5×10 cfu/L. In this study, none of the fish died and 

there were no signs of sublethal toxicity—i.e., no effects on swimming behavior, reflexes,

general appearance, and gill movement. Since B.t.k. will not persist in water (U.S. EPA 1998;

Glare and O’Callaghan 2000), 1000 mg formulation/L or 2.5×1010 cfu/L is used as an NOEC to 

characterize potential effects in tolerant species of fish. 

The series of studies by Christensen (1990c,d,g,i), however, were conducted over a longer

period of exposure (about 30 days) and marginally significant mortality (p=0.052) was 7 observed in rainbow trout at a nominal concentration of 2.87×10 cfu/L (Christensen 1990d). 

Christensen (1990d) specifies that the B.t.k. powder used in this bioassay contained 2.0×1010 

cfu/g or 2.0×107 cfu/mg. Thus, the nominal concentration of 2.87×107 cfu/L corresponds to 

about 1.4 mg/L. While concentrations of B.t.k. in water will not remain constant for 30-days, 7 the value of 1.4 mg/L or 2.87×10 cfu/L is used to characterize risk to sensitive species of fish.

As discussed further in the risk characterization (Section 4.4), the distinction between sensitive

and tolerant species of fish has no impact on the risk assessment because the concentration of

7 2.87×10 cfu/L is far higher than any plausible concentrations of B.t.k. in water even over very 

brief periods of time. Consequently, there is no need to elaborate on the dose-response 

assessment for fish. 

4.3.3.2. Invertebrates – As with terrestrial invertebrates, the toxicity data on aquatic

invertebrates is much more diverse than the data on fish. As summarized in Appendix 6,

laboratory toxicity bioassays are available in several different groups of aquatic invertebrates,

and several field or field simulation studies are available on mixed populations of invertebrates. 

Comparisons among the different studies are confounded somewhat by the different units in

which the results are reported —i.e., mg formulation, IU, or cfu per volume of water and

application rates in units of BIU per area. Appendix 6 provides some estimated conversions for 

key studies. 

The most sensitive species appears to be Daphnia magna with a 21-day EC50 for 

immobilization of 14 mg/L and a decrease in reproduction rates (number of young per

surviving adult) at 5 mg/L using an unspecified Dipel formulation (Young 1990). Citing this 

study, U.S. EPA (1998) classifies B.t.k. as “moderately toxic” to daphnids. U.S. EPA (1998)

does not cite the chronic study in daphnia by Christensen (1991d). In this study, adverse

effects (mortality and decreased reproduction) were seen at a concentration of 5.9 mg/L or

6.24×108 cfu/L, consistent with the decreased reproduction reported by Young (1990) at 5 

mg/L. The study by Christensen (1991d), however, provides a chronic daphnid NOEC of 0.45

mg/L or 6.24×108 cfu/L for both reproductive effects as well as mortality. This value is used to 

characterize risks in sensitive invertebrates. As noted in Appendix 6, the NOEC of 0.45 mg/L

is somewhat below the estimated NOEC of 0.5 mg/L for effects on larvae of the blackfly

(Prosimulium fascum/mixtum). 

Some invertebrates, including copepods, caddisflies, and glass shrimp appear to be extremely

tolerant to B.t.k. in laboratory bioassays. As noted in the risk characterization (Section 4.4),

selection of a tolerant species has a limited impact on the risk assessment because relatively

sensitive species do not appear to be at substantial risk. For this risk assessment, the NOEC of 

36 mg/L is used to characterize risk for tolerant species of invertebrates. This value is taken 

from a series of 24-hour bioassays conducted by Kreutzweiser et al. (1992) in six species of 

4-18
 
mayflies (Ephemeroptera), three species of stoneflies (Plecoptera), and three species of

caddisflies (Tricoptera). At a concentration of 600 IU/ml, equivalent to a concentration of

about 36 mg Dipel 8AF/L, no mortality was observed in four species of mayflies and three

species of caddisflies. Mortality rates of 4% to 30% were noted in three species of stoneflies,

two species of mayflies, and one species of caddisfly. 

4-19
 
4.4. RISK CHARACTERIZATION 

4.4.1. Overview. 

An overview of the risk characterization for B.t.k. is presented in Table 4-6. The only 

organisms that are likely to be affected by B.t.k. or B.t.k. formulations are terrestrial insects. 

Separate dose-response curves can be generated for both sensitive and tolerant terrestrial

insects. At the application rates used to the control of the gypsy moth, the expected mortality

rates for sensitive terrestrial insects range from about 80% to 94%. All sensitive terrestrial 

insects are comprised of lepidoptera, including some species of butterflies, like the endangered

Karner blue, and some swallowtail butterflies and promethea moths. In some cases, 

lepidopteran sensitivity to B.t.k. is highly dependent on developmental stage. This is 

particularly true for the cinnabar moth, with late instar larvae being as sensitive as target

species to B.t.k. and early instar larvae being among the most tolerant lepidoptera. Given the 

mode of action of B.t.k.—i.e., it must be ingested in order to be highly toxic—effects on even

the most sensitive species are anticipated only when species are in a sensitive larval stage at the

time of or shortly after B.t.k. application. Much lower mortality rates (on the order of less than

1% to about 4%) are anticipated in tolerant species, including non-lepidopteran insects and

certain lepidoptera at a particular stage of development. The risk characterization for terrestrial 

mammals is unambiguous: under foreseeable conditions of exposure, adverse effects are

unlikely. Similarly, based on a very conservative exposure assessment for aquatic species,

effects in fish and aquatic invertebrates appear to be unlikely. As discussed in the hazard 

identification, effects in birds, plants, soil microorganisms or invertebrates other than insects

appear to be of no plausible concern. Thus, quantitative risk characterizations for these groups 

are not conducted. For oil-based formulations of B.t.k. (or any other pesticide), effects are

plausible for some soil invertebrates —i.e., Collembola or earthworms. 

4.4.2. Terrestrial Organisms.

4.4.2.1. Terrestrial Vertebrates – The risk characterization for terrestrial mammals is 

unambiguous: under any foreseeable conditions of exposure, adverse effects are unlikely. The 

potential for serious adverse effects is acknowledged, based on the Hernandez et al. (2000)

study involving the intranasal instillation of B.t.k. to mice. The apparent NOAEL for adverse 

7 effects, however, is 10 cfu/mouse. The maximum concentrations of B.t.k. in ambient air range 3 from 100 to 5000 cfu/m , based on monitoring data and the corresponding maximum dose of 3 0.168 cfu/mouse is based on the upper range of the concentration (5000 cfu/m ) and the 3 -8 breathing rate of the mouse (0.0000336 m /day). The resulting hazard index of 2×10 —0.168 7 cfu/mouse ÷ 10 cfu/mouse rounded to 1 significant digit—is a factor of 50 million below the

level of concern. Therefore, although the risk characterization acknowledges the possibility of

serious adverse effects, the upper range of plausible levels of exposure are far below levels 

associated with serious adverse effects. For oral exposures, the hazard identification is

essentially negative—i.e., there is no indication that oral exposure to B.t.k. at any concentration 

will cause adverse effects. For the purpose of quantitatively expressing risk, the dose of 8400

mg/kg/day is used as a working NOAEL, although it is possible that higher doses might also be

classified as NOAELs. Based on a very conservative exposure assessment involving oral

(vegetation and drinking water) as well as dermal (direct spray) scenarios, the hazard index is

0.02, a factor of 50 below the working NOAEL. 

As noted in the risk characterization for human health effects (see Section 3.4.3), a recent study

by Hernandez et al. (2000) reports a substantial increase in mortality in mice pre-treated with

an influenza virus and then exposed to various doses of B.t.k. In this study, increased mortality

was observed at a very low dose—i.e., 100 cfu/mouse —which is a factor of one-million below 8 the lethal dose in non-viral treated mice of 1×10 cfu/mice. As discussed in Section 3.4.3, the 

significance of the Hernandez et al. (2000) study to potential human health effects is difficult to 

4-20
 
assess. For wildlife, the estimated maximum exposure of 0.186 cfu/mouse is far below the 100

cfu/mouse exposure at which the increased mortality was observed. Nonetheless, the 

Hernandez et al. (2000) study does not identify a NOEC for mice pre-treated with influenza

virus. Thus, as in the human health risk assessment, the potential for interactions between 

B.t.k. and populations infected with influenza virus cannot be well assessed at this time and is

likely to be an area of further study in the coming years. 

4.4.2.2. Terrestrial Invertebrates – Sufficient data are available to estimate dose-response

relationships for both sensitive species as well as relatively tolerant species in units used to

measure application rates—i.e., BIU/ha. As discussed in Section 4.3.2.2, risks for terrestrial 

insects can be expressed using a log-dose probit-response curve: 

Y = -1.48 + 2.34 x + 3.36 S 

where Y is the probit response, x is the common log of the application rate in BIU/ha, and S is 

equal to 1 for sensitive species and 0 for tolerant species. Substituting the application rates of

49 BIU/ha and 99 BIU/ha into the above equation, mortality rates in units of probits can be

explicitly estimated for sensitive and tolerant organisms at both application rates. As 

summarized in Table 4-6, high mortality rates in sensitive species are likely—i.e., rates of

about 80% to 94%. Mortality rates in tolerant organisms are estimated to be much lower, in the

range of 0.6% to 3.6%. Given the experimental scatter (Figure 4-1), these rates should be

regarded as approximate. While confidence intervals could be derived for the dose-response

curves, they would have no impact on the risk characterization. 

The identification of tolerant and sensitive organisms, however, is not always straightforward. 

As summarized in Table 4-5, target species like the gypsy moth and cabbage looper are clearly

sensitive. In addition, some species of butterflies, including the endangered Karner blue and

some swallowtail butterflies and promethea moths appear to be as sensitive as the target

species to B.t.k. exposure. For some lepidoptera, sensitivity to B.t.k. depends primarily on 

developmental stage. This is particularly evident in the case of the cinnabar moth, with late

instar larvae being as sensitive as target species to B.t.k. exposure and early instar larvae being 

among the most tolerant lepidoptera. All of the more sensitive organisms are lepidopteran 

larvae. Given the mode of action of B.t.k.—i.e., it must be ingested in order to be highly

toxic—effects on even the most sensitive species are anticipated only when the species is in a

sensitive larval stage at the time of B.t.k. application or shortly thereafter. 

Tolerant species appear to be comprised of non-lepidopteran insects as well as certain larval

stages of some lepidoptera. As noted above, early instar larvae of the cinnabar moth appear to

among the most tolerant lepidoptera. Based on the study by Peacock et al. (1998), owlet moths

and some looper butterflies also appear to be relatively tolerant to B.t.k. As illustrated in 

Figures 4-1 and 4-2, other lepidopteran species/instars display sensitivities that are intermediate

between those of the most sensitive and most tolerant organisms, and the distribution of

tolerances appears to be nearly uniform. As summarized in Appendix 3, the apparently wide

variability of sensitivity among different lepidopteran species is supported by the recent field

study of Rastall et al. (2003), who noted statistically significant decreases in three nontarget

lepidopteran species but either no change or statistically significant increases in other nontarget

lepidopteran species associated with the application of B.t.k. 

Thus, the risk characterization for terrestrial insects is highly variable. Mortality rates are 

likely to be high among sensitive lepidopteran species after any B.t.k. application that is

effective for controlling the gypsy moth or other target species, whereas mortality rates are not 

4-21
 
likely to be detectable or biologically significant among non-lepidopteran insects or tolerant

lepidoptera at certain stages of development. The response in other lepidopteran species will

be intermediate between sensitive and tolerant species. As discussed in Section 4.1.2.3.2, an 

older oil-based formulation of B.t.k., Dipel 4L, decreased populations of Collembola as well as 

earthworms. Dipel 4L is not used in USDA programs. Nonetheless, any oil-based formulation 

of B.t.k. (or any other pesticide) might be expected to cause adverse effects in some soil

invertebrates. 

As summarized in Table 4-5 and illustrated in Figure 4-1, the toxicity data on honeybees are

encompassed by the dose-response curves for sensitive and tolerant insect species but the

apparent slope of the mortality curve for honeybees is shallower than that for other insect

species. This observation, however, is based on only a single study (Atkins 1991a) and should

not be subject to over interpretation. Nonetheless, the data from Atkins (1991a) suggests that

mortality rates in bees sprayed directly with B.t.k. at application rates used to control the gypsy 

moth could be approximately 20%. In practice, applications of B.t.k. to control the gypsy moth

are not associated with substantial mortality in bees, which may be due to foliar interception of

the applied B.t.k. 

4.4.3. Aquatic Organisms.

The risk characterization for both fish and aquatic invertebrates is based on a maximum

concentration of 0.24 mg formulation/L. As discussed in the exposure assessment (see Section

4.2.4), this concentration is calculated from an application rate of 2 lbs/acre or 224.2 mg/m2 

using a water depth of 1 m. In other words, 0.24 mg formulation/L would be the concentration

in water immediately after direct spray over water. In most applications, actual concentrations

in water would be much less, as suggested by the monitoring data of Valadares de Amorin et al.

(2001). For both fish and invertebrates, this concentration is typically compared to longer-term

toxicity values—i.e., 30 days for fish and 21 days for aquatic daphnids. Thus, the resulting 

hazard quotients are likely to overestimate risk substantially. 

As summarized in Table 4-5, none of the hazard quotients exceed one—i.e., there is no

indication that adverse effects are likely in either tolerant or sensitive species. For tolerant 

species the interpretation is unequivocal: hazard quotients are below a level of concern by

factors of 5000 for fish and more than 140 for invertebrates. For sensitive species of fish, the

hazard quotient of 0.2 is below the level of concern by a factor of 5. Given that the toxicity 

value is based on a 30-day NOEC and given that B.t.k. will not persist in water, there is no 

basis for concern in even sensitive species of fish. The hazard quotient of 0.5 for sensitive

species of invertebrates may be viewed with marginal concern in that it suggests that effects

could be seen in shallower bodies of water. Again, however, the toxicity value is based on a

21-day study and it is not likely that concentrations of B.t.k. would be maintained at levels 

close to 0.24 mg/L for this period of time. 

4-22
 
5. LIST OF STUDIES CONSULTED
 

Addison JA. 1995. Persistence and nontarget effects of Bacillus thuringiensis in soil: a 

review. Can J For Res. 23:2329-2342 

Addison JA; Holmes SB. 1995. Effect of two commercial formulations of Bacillus 

thuringiensis subsp. kurstaki Dipel 8L and Dipel 8AF on the collembolan species Folsomia 

candida in a soil microcosm study. Bull Environ ContamToxicol. 55(5):771-778 

Addison JA; Holmes SB. 1996. Effect of two commercial formulations of Bacillus 

thuringiensis subsp. kurstaki on the forest earthworm Dendrobaena octaedra. Can J For Res. 

26:1594-1601 

Aer'aqua Medicine Ltd. 2001. Health Surveillance following Operation Ever Green: A

Programme to eradicate the white-spotted tussock moth from the eastern suburbs of Auckland. 

Report to the Ministry of Agriculture and Forestry, New Zealand, May 2001 

Aer/aqua Medicine Ltd. 2001. Health Surveillance following Operation Ever Green: A

Programme to eradicate the white-spotted tussock moth from the eastern suburbs of Auckland. 

Report to the Ministry of Agriculture and Forestry, New Zealand, May 2001 

Akiba Y. 1986. Microbial ecology of Bacillus thuringiensis: VI. Germination of Bacillus 

thuringiensis spores in the soil. Appl Entomol Zool. 21:76-80 

Akiba Y. 1991. Assessment of rainwater-mediated dispersion of field-sprayed Bacillus 

thuringiensis in the soil. Appl Entomol Zool. 26:477-483 

Appel HM; McCarthy WJ. 1999. Hostplant incompatibility with biological control: tannins, 

bt, and the gypsy moth. U. S. Department of Agriculture Competitive Research Grant 

Aronson AI; Han ES; Mcgaughey W; Johnson D. 1991. The solubility of inclusion proteins 

from Bacillus thuringiensis is dependent upon protoxin composition and is a factor in toxicity 

to insects. Appl Environ Microbiol. 57:981-986 

Atkins E. 1991a. Bee adult toxicity dusting test evaluating the comparative acute contact and

stomach poison toxicity of BT III dry flowable.(Bacillus thuringiensis var. kurstaki) to honey 

bee worker adults: Lab Project No. 91/838. Unpublished study prepared by Univ. California, 

Riverside. 13 p. MRID 41983301 

Atkins E. 1991b. Bee adult toxicity dusting test evaluating the comparative acute contact and

stomach poison toxicity of BT I dry flowable. (Bacillus thuringiensis var. kurstaki) to honey

worker adults: Lab Project No. 91/836. Unpublished study prepared by Univ. California. 

Riverside. 13 p. MRID 41983501 

Auckland District Health Board. 2002. Health Risk Assessment of the 2002 Aerial Spray

Eradication Programme for the Painted Apple Moth in Some Western Suburbs of Auckland. A 

Report to the Ministry of Agriculture and Forestry. Prepared by: Public Health Service,

Auckland District Health Board, Auckland, New Zealand. Available at: 

http://www.adhb.co.nz/akphp/Services/AppleMoth/PAM%20HRA%20-%20with%20maps.pdf 

5-1
 
 

Barridge B. 1990a. Delta BT-product identity: Lab Project No. DBP 1989-100. Unpublished

study prepared by Delta Biological Products. 16 p. MRID 41751101 

Barridge B. 1990b. Delta BT-formation of unintentional ingredients: Lab Project No. DBP 

1989-102. Unpublished study prepared by Delta Biological Products. 8 p. MRID 41751102 

Barridge B. 1990c. Delta BT-analysis of samples: Lab Project No. DBP 1989-103. 

Unpublished study prepared by Delta Biological Products. 20 p. MRID 41751103 

Barridge B. 1990d. Delta BT-certification of limits: Lab Project No. DBP 1989-104.

Unpublished study prepared by Delta Biological Products. 5 p. MRID 41751104 

Barridge B. 1990e. Delta BT–manufacturing process: Lab Project No. DBP 1989-101.

Unpublished study prepared by Delta Biological Products, Inc. 8 p. MRID 42080101 

Barridge B. 1990f. Delta BT-physical and chemical properties: Lab Project No. DPB 1989­

105. Unpublished study prepared by Delta Biological Products, Inc. 6 p. MRID 42080102 

Barry JW; Skyler PJ; Teske ME; Rafferty JA; Grim BS. 1993. Predicting and measuring drift 

of Bacillus thuringiensis sprays. Env Toxicol Chem. 12(11):1977-1989. 

Bassett J; Watson M. 1999a. Acute ORAL TOXICITY (LD 50) study in rats with DiPel

technical powder: Lab Project No. 7363-98-0123-TX-001: 7636-98-0123-TX-000. 

Unpublished study prepared by Ricerca Inc. 45 p. MRID No. 44791605 

Bassett J; Watson M. 1999b. Primary dermal irritation study in albino rabbits with DiPel

technical powder: Lab Project Nos. 7635-98-0124-TX-001: 7635-98-0124:

7635-98-0124-TX-000. Unpublished study prepared by Ricerca, Inc. 52 p. MRID 44791607 

Beavers J. 1991a. ABG-6305: An avian oral pathogenicity and toxicity in the bobwhite: Lab

Project No. 161-117. Unpublished study prepared by Wildlife International Ltd. 21 p.

MRID 41974804 

Beavers J. 1991b. ABG-6305: An avian oral pathogenicity and toxicity in the bobwhite: Lab

Project No. 161-118. Unpublished study prepared by Wildlife International Ltd. 20 p.

MRID 41974805 

Beavers J; Jaber M. 1987. SAN 418 SC 62 Bacillus thuringiensis Tenebrionis: An avian acute 

oral ld50 pathogenicity study in the mallard : Study No. 131-132. Unpublished study

performed by Wildlife International Ltd. 19 p. MRID 40497409 

Beavers J; Jaber M. 1987. SAN 418 SC 62 Bacillus thuringiensis Tenebrionis: An avian 

intraperitoneal injection pathogenicity study in the mallard: Study No. 131-133. Unpublished

study performed by Wildlife International Ltd. 19 p. MRID 40497410 

Beavers J; Smith G. 1990a. An avian oral pathogenicity and toxicity study in the mallard: Lab 

Project No. 297-106. Unpublished study prepared by Wildlife International Ltd. 19 p.

MRID 41751108 

Beavers J; Smith G. 1990b. An avian oral pathogenicity and toxicity study in the bobwhite: 

Lab Project No. 297-105. Unpublished study prepared by Wildlife International Ltd. 21 p.

MRID 41751109 

5-2
 
Beavers J; Smith G. 1991a. Bacillus thuringiensis var. Israelensis, strain NB31: An avian oral 

pathogenicity and toxicity study in the mallard: Lab Project No. 254-125. Unpublished study

prepared by Wildlife International Ltd. 25 p. MRID 41842702 

Beavers J; Smith G. 1991b. Bacillus thuringiensis var. Israelensis, strain NB31: An avian oral 

pathogenicity and toxicity study in the bobwhite: Lab Project No. 254-124. Unpublished study

prepared by Wildlife International Ltd. 25 p. MRID 41842703 

Beavers J; Clauss B; Jaber M. 1988a. Bacillus thuringiensis Strain EG2348: An Avian Acute 

Oral LD50 Pathogenicity Study in the Bobwhite. Wildlife International Ltd. Project No.: 235­

120. Unpublished study prepared by Wildlife International Ltd. 19 p. MRID 40898807 

Beavers J; Clauss B; Jaber M. 1988b. Bacillus thuringiensis strain EG2424: An avian acute 

oral LD50 pathogenicity study in the mallard. Wildlife International Lt. Project No. 235-122.

Unpublished study prepared by Wildlife International Ltd. 19 p. MRID 40951107 

Beavers J; Clauss B; Jaber M. 1988b. Bacillus thuringiensis strain EG2424: An avian acute 

oral LD50 pathogenicity study in the mallard. Wildlife International Ltd. Project No. 235-122.

Unpublished study prepared by Wildlife International Ltd. 19 p. MRID No. 40951107 

Beevers M. 1990. Effects of Bacillus thuringiensis subp. kurstaki on the insect egg parasitoid, 

Trichogramma pretiosum: Final Report: Lab Project No. CAR/103-90. Unpublished study

prepared by California Agricultural Research, Inc. 41 p. MRID 41443409 

Behle RW; Mcguire MR; Gillespie RL; Shasha BS. 1997. Effects of alkaline gluten on the 

insecticidal activity of Bacillus thuringiensis. J Econ Entomol. 90:354-360 

Behle RW; Mcguire MR; Shasha BS. 1997. Effects of sunlight and simulated rain on residual 

activity of Bacillus thuringiensis formulations. J Econ Entomol. 90:1560-1566 

Bell J. 1994. Personal communication. Agriculture Canada. Telephone conversation with 

Patrick R. Durkin, February 25, 1994 

Bell JL; Whitmore RC. 1997. Bird populations and habitat in Bacillus thuringiensis and 

Dimilin-treated and untreated areas of hardwood forest. Am Midl Nat. 137:239-250 

Bellantoni D; Grimstead S; Roberts C; et al. 1991a. Delta BT: A toxicity and pathogenicity 

test with the rainbow trout (Oncorhynchus mykiss): Final Report: Lab Project No. 297A-101.

Unpublished study prepared by Wildlife International Ltd. 25 p. MRID 41899101 

Bellantoni D; Grimstead S; Holmes C; et al. 1991b. Delta BT: A toxicity and pathogenicity 

test with the cladocern (Daphnia magna): Final Report: Lab Project No. 297A-104. 

Unpublished Study prepared by Wildlife International Ltd. 28 p. MRID 41899102 

Bellantoni D; Grimstead S; Roberts C; et al. 1991c. Delta BT: A toxicity and pathogenicity 

test with the grass shrimp (Palaemonetes pugio): Final Report: Lab Project Number: 297A102. Unpublished study prepared by Wildlife International Ltd. 24 p. MRID 41899103 

Bellantoni D; Grimstead S; Roberts C; et al. 1991d. Delta BT: A toxicity and pathogenicity 

test with the sheepshead minnow (Cyprinodon variegatus): Final Report: Lab Project No.

297A-103. Unpublished study prepared by Wildlife International Ltd. 25 p. MRID 41899104 

5-3
 
Belloq MI; Bendell JF; Cadogan BL. 1992. Effects of the insecticide Bacillus thuringiensis on 

Sorex cinereus (masked shrew) populations, diet, and prey selection in jack pine plantation in

northern Ontario. Can J Zool. 70:505-510 

Ben-Dyke R; Hogan GK; Hoffman CA; et al. 1981. An acute inhalation toxicity and

infectivity study of Thuricide 32-B in therRat: Project No. 80-7472. Unpublished study

received Mar 8, 1982 under 11273-2; prepared by Bio/dyanamics, Inc., submitted by Sandoz,

Inc.--Crop Protection, San Diego, Calif.; CDL:246967-C. MRID 00096529 

Benz G; Altweg A. 1975. Safety of Bacillus thuringiensis for earthworms. J Invert Pathol. 

26:125-126 

Berg N. 1989. Acute dermal toxicity study in rabbits with SP 408, PPQ 2585 in support of

registration of Novodor technical: Lab Project I.D.: 13188. Unpublished study prepared by 

Novo-Nordisk A/S, Enzyme Toxicology Laboratory. 16 p. MRID 41412705 

Berg N. 1990. Bacillus thuringiensis var. kurstaki, batch BBB 0073: Acute intravenous 

toxicity/pathogenicity study in rats. Enzyme Toxicology Laboratory; June 19, 1990; 108 p. 

Sponsored by Novo Nordkisk A/S, Bagsvaerd, Denmark 

Berg N. 1991a. Eye irritation study in rabbits with the end product Foray 48B, batch BBN 

6056. Enzyme Toxicology Laboratory; April 15, 1991; 19 p. Sponsored by Novo Nordisk 

A/S, Bagsvaerd, Denmark 

Berg N. 1991b. Eye irritation study in rabbits with the end product Foray 48B, batch BBN

6057: in support of Formula Amendment of Foray 48B. Enzyme Toxicology Laboratory; April 

18, 1991; 19 p. Sponsored by Novo Nordisk A/S, Bagsvaerd, Denmark 

Berg N; Kiehr B. 1991. Eye irritation study in rabbits with the end product Foray 48B, batch 

BBN 6057. Enzyme Toxicology Laboratory; February 26, 1991; 19 p. Sponsored by Novo 

Nordisk A/S, Bagsvaerd, Denmark 

Berg N; Sorensen E; Overholt J. 1991. Summary of acute toxicology in support of formula 

amendment of Foray 48B:. Lab Project No. NOVO/FFCFA/VOL1. Unpublished study

prepared by Novo Nordisk A/S & Novo Nordisk Bioindustrials, Inc. 21 p. MRID 41884301 

Bernier RL; Gannon DJ; Moser GP; Mazzarello M; Griffiths MM; Guest PJ. 1990. 

Development of a novel B.t. strain for the control of forestry pests. Brighton Crop Protection 

Conference, Pests and Diseases. 1:245-251. (Cited in Glare and O'Callaghan 2000) 

Berstein IL; Bernstein JA; Miller M; Tierzieva S; Sernstein DI; Jummus Z; Selgrade MJK;

Doerfler DL; Seligy VL. 1999. Immune responses in farm workers after exposure to Bacillus 

thuringiensis pesticides. Environ. Hlth Perspect. 107:575-582 

Betz FS; Hammond BG; Fuchs RL. 2000. Safety and advantages of Bacillus thuringiensisprotected plants to control insect pests. Regul Toxicol Pharmacol. 32(2):156-173 

Birkhold D. 1999. Fermentation manufacturing process for Bacillus thuringiensis spp.

kurstaki at Abbott Laboratories. Lab Project No. BTKMAN-28R. Unpublished study prepared 

by Abbott Laboratories. 13 p. MRID 45136503 

5-4
 
Bishop AH; Johnnson C; Perant M. 1999. The safety of Bacillus thuringiensis to mammals 

investigated by oral and subcutaneous dosage. World J. Microbiol Biotechnol. 15:375-380 

Blackmore H. 2003. Painted Apple Moth Eradication Campaign, West Auckland. Interim 

Report of the Community-based health and incident monitoring of the aerial spray programme:

January - December 2002. Report dated February, 2003. Available at: 

http://www.geocities.com/no_spray/index2.html 

Boeri R. 1991. Chronic toxicity of ABG-6305 to daphnid: Daphnia magna: Lab Project No. 

90162-A. Unpublished study prepared by Resource Analysts, Inc. 47 p. MRID 41974802 

Boxenbaum J; D'Souza R. 1990. Interspecies pharmacokinetic scaling, biological design and 

neoteny. Adv Drug Res. 19:139-195 

Broderick NA; Goodman RM; Raffa KF; Handelsman J. 2000. Synergy between zwittermicin 

a and Bacillus thuringiensis subsp. kurstaki against gypsy moth (Lepidoptera: lymantriidae). 

Environ Entomol. 29:101-107 

Buckner CH; Ray DGH; McLeod BB. 1973. The effects of pesticides on small forest

vertebrates of the Spruce Woods Provincial Forest, Manitoba. Manitoba Entomol. 7:37-45 

Buckner CH; Kingsbury PD; McLeod BB; Mortensen KL; Ray DGH. 1974. Impact of aerial

treatment on non-target organisms, Algonquin Park, Ontario, and Spruce Woods, Manitoba,

Section F. In: Evaluation of Commercial Preparations of Bacillus Thuringiensis with and 

Without Chitinase Against Spruce Budworm. .Ottawa, Ontario, Canadian Forestry Service,

Chemical Control Research Institute, pp 1-72. 

Cameron EA; Reeves RM. 1990. Carabidae (Coleoptera) associated with gypsy moth,

(Lymantria dispar (L.) Lepidopter:Lymantriidae), populations subjected to Bacillus 

thuringiensis Berliner treatments in Pennsylvania. Can Entomol. 122:123-129 

Cannon GE; Krize JW. 1975. TH 6040 egg to egg reproduction study in fathead minnows

treated at .1, .05, .025, .0125, to .00675 ppm:. Laboratory No. 5E 6094. Unpublished study

received Feb 10, 1976 under 6G1744; prepared by Cannon Laboratories, Inc., submitted by

Thompson-Hay. MRID 00038616 

Cerf D. 1990a. Susceptibility of four orders of insects (lepidoptera, diptera, coleoptera, and

orthoptera) to technical grade active ingredients (TGAI) and tenebrionis (strain SA10): Final

Report No. 90/03/12. Unpublished study prepared by Sandoz Crop Protection Corp. 40 p.

MRID 41441516 

Cerf D. 1990b. Susceptibility of four orders of insects (lepidoptera, diptera, coleoptera, and

orthoptera) to technical grade active ingredients (TGAI) and tenebrionis (strain SA10): Lab

Project No. 90/03/12. Unpublished study prepared by Sandoz Crop Protection Corp. 40 p.

MRID 41441625 

Cerstiaens A; Verleyen P; Van Rie J; et al. 2001. Effect of Bacillus thuringiensis cry1 toxins 

in insect hemolymph and their neurotoxicity in brain cells of lymantria dispar. Appl Environ

Microbiol. 67(9):3923-3927 

5-5
 
Chandler G. 1990a. Chronic toxicity of Bacillus thuringiensis var. Israelensis technical 

material to the benthic harpacticoid copepod, Amphiascus minutus under static conditions: 

Report No. USC-SPH-2-90: Abbott Lab-VTP-12. Unpublished study prepared by Univ. South 

Carolina. MRID 41439010 

Chandler G. 1990b. Chronic toxicity of Bacillus thuringiensis var. kurstaki technical material 

to the benthic harpacticoid copepod, Amphiascus minutus under static conditions: Toxicity test 

report: Lab Project No. USC-SPH-1-90. Unpublished study prepared by Univ. South Carolina. 

MRID 41443408 

Chen C; Macuga R. 1990a. Plasmid profile of Bacillus thuringiensis subsp. Aizawai, strain 

SA2: Final Report No. 90/02/03E. Unpublished study prepared by Sandoz Crop Protection 

Corp. 40 p. MRID 41441507 

Chen C; Macuga R. 1990b. Plasmid profile of Bacillus thuringiensis subsp. Israelensis, strain 

SA3: Final Report No. 90/02/03C. Unpublished study prepared by Sandoz Crop Protection 

Corp. 40 p. MRID 41441508 

Chen C; Macuga R. 1990c. Plasmid profile of Bacillus thuringiensis subsp. Israelensis, strain

SA3A: Final Report No.90/02/03D. Unpublished study prepared by Sandoz Crop Protection

Corp. 40 p. MRID 41441509 

Chen C; Macuga R. 1990d. Flagella antigen serotyping of Bacillus thuringiensis subsp.

Aizawai, strain SA2: Final Report No. 90/02/12E. Unpublished study prepared by Sandoz 

Crop Protection Corp. 23 p. MRID 41441510 

Chen C; Macuga R. 1990e. Flagella antigen serotyping of Bacillus thuringiensis subsp.

Israelensis, strain SA3: Final Report No. 90/02/12C. Unpublished study prepared by Sandoz 

Crop Protection Corp. 23 p. MRID 41441511 

Chen C; Macuga R. 1990f. Flagella antigen serotyping of Bacillus thuringiensis subsp.

Israelensis, strain SA3A: Final Report No. 90/02/12D. Unpublished study prepared by Sandoz 

Crop Protection Corp. 23 p. MRID 41441512 

Chen C; Macuga R. 1990g. Description of endotoxin proteins produced by Bacillus 

thuringiensis subsp. Aizawai, strain SA2: Final Report No. 90/02/21C. Unpublished study

prepared by Sandoz Crop Protection Corp. 30 p. MRID 41441517 

Chen C; Macuga R. 1990i. Description of endotoxin proteins produced by Bacillus 

thuringiensis subsp. Israelensis, strain SA3A: Final Report No. 90/02/21F. Unpublished study

prepared by Sandoz Crop Protection Corp. 27 p. MRID 41441519 

Chen C; Macuga R. 1990j. Plasmid profile of Bacillus thuringiensis subsp. tenebrionis strain

SA10: Lab Project No: 90/02/03F. Unpublished study prepared by Sandoz Crop Protection

Corp. 40 p. MRID 41441613 

Chen C; Macuga R. 1990k. Plasmid profile of Bacillus thuringiensis subsp. kurstaki strain 

SA11001C98-1-1: Lab Project No. 90/0203A. Unpublished study prepared by Sandoz Crop 

Protection Corp. 40 p. MRID 41441614 

5-6
 
Chen C; Macuga R. 1990l. Plasmid profile of Bacillus thuringiensis subsp. kurstaki strain 

INT-15-313: Lab Project No. 90/02/03. Unpublished study prepared by Sandoz Crop 

Protection Corp. 40 p. MRID 41441615 

Chen C; Macuga R. 1990m. Plasmid profile of Bacillus thuringiensis subsp. kurstaki strain 

SA12: Lab Project No. 90/02/03B. Unpublished study prepared by Sandoz Crop Protection 

Corp. 40 p. MRID 41441616 

Chen C; Macuga R. 1990n. Flagella antigen serotyping of Bacillus thuringiensis subsp.

tenebrionis strain SA10: Lab Project No. 90/02/12F. Unpublished study prepared by Sandoz 

Crop Protection Corp. 23 p. MRID 41441617 

Chen C; Macuga R. 1990o. Flagella antigen serotyping of Bacillus thuringiensis subsp.

kurstaki strain SA11001C98-1-1: Lab Project No. 90/02/12A. Unpublished study prepared by 

Sandoz Crop Protection Corp. 23 p. MRID 41441618 

Chen C; Macuga R. 1990p. Flagella antigen serotyping of Bacillus thuringiensis subsp.

kurstaki strain SA11001C98-1-1: Lab Project No. 90/02/12A. Unpublished study prepared by 

Sandoz Crop Protection Corp. 23 p. MRID 41441619 

Chen C; Macuga R. 1990q. Flagella antigen serotyping of Bacillus thuringiensis subsp.

kurstaki strain SA12: Lab Project No. 90/02/12B. Unpublished study prepared by Sandoz Crop 

Protection Corp. 23 p. MRID 41441620 

Chen C; Macuga R. 1990r. Description of endotoxin proteins produced by Bacillus 

thuringiensis subsp. kurstaki strain INT-15-313: Lab Project No. 90/02/21A. Unpublished

study prepared by Sandoz Crop Protection Corp. 30 p. MRID 41441626 

Chen C; Macuga R. 1990s. Description of endotoxin proteins produced by Bacillus 

thuringiensis subsp. kurstaki strain -1-1: Lab Project No. 90/02/21. Unpublished study

prepared by Sandoz Crop Protection Corp. 30 p. MRID 41441627 

Chen C; Macuga R. 1990t. Description of endotoxin proteins produced by Bacillus 

thuringiensis subsp. kurstaki strain SA12: Project No. 900221B. Unpublished study prepared 

by Sandoz Crop Protection Corp. 30 p. MRID 41441628 

Chen C; Macuga R. 1990u. Description of endotoxin proteins produced by Bacillus 

thuringiensis subsp. tenebrionis strain SA10: Lab Project No. 90/02/21D. Unpublished study

prepared by Sandoz Crop Protection Corp. 25 p. MRID 41441629 

Chen C; Macuga R. 1990v. Description of Endotoxin Proteins Produced by Bacillus 

thuringiensis subsp. Israelensis, Strain SA3: Final Report: Final Report No. 90/02/21E.

Unpublished study prepared by Sandoz Crop Protection Corp. 27 p. MRID 41441518 

Chen C; Macuga R; Cerf D. 1990a. Insecticidal toxins produced by Bacillus thuringiensis

subsp. Aizawai, strain SA2. I. Effect of autoclaving: Final Report No. 90/01/31E. Unpublished

study prepared by Sandoz Crop Protection Corp. 19 p. MRID 41441520 

Chen C; Macuga R; Cerf D. 1990b. Insecticidal toxins produced by Bacillus thuringiensis

subsp. Israelensis, strain SA3. I. Effect of autoclaving: Final Report No. 90/01/31. 

Unpublished study prepared by Sandoz Crop Protection Corp. 19 p. MRID 41441521 

5-7
 
Chen C; Macuga R; Cerf D. 1990c. Insecticidal toxins produced by Bacillus thuringiensis

subsp. Israelensis, strain SA3A. I. Effect of autoclaving: Final Report No. 90/01/31D.

Unpublished study prepared by Sandoz Crop Protection Corp. 19 p. MRID 41441522 

Chen C; Cerf D; Sjolander A; et al. 1990d. Insecticidal toxins produced by Bacillus 

thuringiensis subsp. Aizawai, strain. II. Concentration of â-exotoxin: Final Report No. 

90/02/07E. Unpublished study prepared by Crop Protection Corp. 36 p. MRID 41441523 

Chen C; Cerf D; Sjolander A; et al. 1990e. Insecticidal Toxins produced by Bacillus 

thuringiensis subsp. Israelensis, strain SA3. II. Concentration of â-exotoxin: Final Report No. 

90/02/07C. Unpublished study prepared by Sandoz Crop Protection Corp. 36 p.

MRID 41441524 

Chen C; Cerf D; Sjolander A; et al. 1990f. Insecticidal toxins produced by Bacillus 

thuringiensis subsp. Israelensis, strain SA3A. II. Concentration of â-exotoxin: Final Report No. 

90/02/07D. Unpublished study prepared by Sandoz Crop Protection Corp. 36 p.

MRID 41441525 

Chen C; Macuga R; Cerf D. 1990g. Insecticidal toxins produced by Bacillus thuringiensis

Subsp. Tenebrionis strain SA10. I. Effect of autoclaving: Final Report: Lab Project No.

90/01/31F. Unpublished study prepared by Sandoz Crop Protection Corp. 19 p.

MRID 41441605 

Chen C; Macuga R; Cerf D. 1990h. Insecticidal toxins produced by Bacillus thuringiensis 

subsp. kurstaki strain SA11001C98-1-1. I. Effect of sutoclaving: Final Report: Lab Project No. 

90/01/31A. Unpublished study prepared by Sandoz Crop Protection Corp. 19 p.

MRID 41441606 

Chen C; Macuga R; Cerf D. 1990i. Insecticidal toxins produced by Bacillus thuringiensis 

subsp. kurstaki strain INT-15-313. I. Effect of autoclaving: Final Report: Lab Project No.

90/01/31. Unpublished study prepared by Sandoz Crop Protection Corp. 19 p. MRID 

41441607 

Chen C; Macuga R; Cerf D. 1990j. Insecticidal toxins produced by Bacillus thuringiensis 

subsp. kurstaki strain SA12. I. Effect of autoclaving: Lab Project No. 90/01/31B. Unpublished

study prepared by Sandoz Crop Protection Crop. 19 p. MRID 41441608 

Chen C; Cerf D; Sjolander A; et al. 1990k. Insecticidal toxins produced by Bacillus 

thuringiensis subsp. kurstaki strain SA11001C98-1-1. II. Concentratration of â-exotoxin: Lab 

Project No. 90/02/07A. Unpublished study prepared by Sandoz Crop Protection Corp. 36 p.

MRID 41441630 

Chen C; Cerf D; Sjolander A; et al. 1990l. Insecticidal toxins produced by Bacillus 

thuringiensis subsp. kurstaki strain INT-15-313. II. Concentration of â-exotoxin: Lab Project

No. 90/02/07. Unpublished study prepared by Sandoz Crop Protection Corp. 36 p.

MRID 41441631 

Chen C; Cerf D; Sjolander A; et al. 1990m. Insecticidal toxins produced by Bacillus 

thuringiensis subsp. tenebrionis strain SA10. II. Concentration of â-exotoxin: Lab Project No. 

90/02/07F. Unpublished study prepared by Sandoz Crop Protection. Corp. 36 p.

MRID 41441632 

5-8
 
Chen C; Cerf D; Sjolander A; et al. 1990n. Insecticidal toxins produced by Bacillus 

thuringiensis subsp. kurstaki strain SA12. II. Concentration of â-exotoxin: Lab Project No. 

90/02/07B. Unpublished study prepared by Sandoz Crop Protection Corp. 36 p.

MRID 41441633 

Chen CH; Ding HC; Chang TC. 2001. Rapid identification of bacillus cereus based on the

detection of a 28.5-kilodalton cell surface antigen. J Food Prot. 64(3):348-354 

Cheon HM; Kim HJ; Kang SK; Seo SJ. 1997. Effects of Bacillus thuringiensis deltaendotoxin on insect fat body structure. Korean J Bio Sci. 1(3):507-513 

Christensen K. 1990a. Vectobac technical material (Bacillus thuringiensis var. Israelensis): 

Infectivity and pathogenicity to bluegill sunfish (Lepomis macrochirus) during a 30-day static

renewal test: Final Report: SLI Report 90-2-3228; SLI Study 2439.0889.6104.158. 

Unpublished study prepared by Springborn Laboratories, Inc. 55 p. MRID 41439007 

Christensen K. 1990b. Vectobac technical material (Bacillus thuringiensis var. Israelensis): 

Infectivity and pathogenicity to rainbow trout (Oncorhynchus mykiss) during a 32-day static

renewal test: Final Report: SLI Report 90-2-3242; SLI Study 2439.0889.103.157. Unpublished

study prepared by Springborn Laboratories, Inc. 55 p. MRID 41439008 

Christensen K. 1990c. Dipel Technical material (Bacillus thuringiensis var. kurstaki): 

Infectivity and pathogenicity to bluegill sunfish (Lepomis macrochirus) during a 32-day static 

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Yang CY; Pang JC; Kao SS; Tsen HY. 2003. Enterotoxigenicity and cytotoxicity of Bacillus 

thuringiensis strains and development of a process for cry1ac production. J Agric Food Chem. 

51(1):100-105 

Young B. 1990. 21-Day Prolonged static renewal toxicity of dipel technical to Daphnia 

magna: Lab Project No. 38417. Unpublished study prepared by Analytical Bio-chemistry 

Laboratories, Inc. 123 p. MRID 41443407 

Youston AA. 1973. Effects of Bacillus thuringiensis delta-endotoxin on an insect predator

which has consumed intoxicated cabbage looper larvae. J Invert Pathol. 21:312-314 

5-36
 
Yu L; Berry RE; Croft BA. 1997. Effects of Bacillus thuringiensis toxins in transgenic cotton 

and potato on Folsomia candida (Collembola: Isotomidae) and Oppia nitens (Acari: 

oribatidae). J Econ Entomol. 90(1):113-118 

Zhioua E; Heyer K; Browning M; Ginsberg HS; Lebrun RA. 1999. Pathogenicity of Bacillus 

thuringiensis variety kurstaki to ixodes scapularis (Acari: ixodidae). J Med Entomol. 

36:900-902 

5-37
 
Figure 3-1: Number of symptoms per worker based on total exposure to B.t.k. (millions of cfu

hours) and the use of protective masks (data from Cook 1994 as summarized in Table 3-6 of this

risk assessment) 

Figures - 1 
Figure 3-2: Dose-response relationships in mice after intranasal administration of B.t.k. with or 

without previous challenge with influenza virus at 4% of the LD50 (data from Hernandez et al. 

1999 and 2000). 

Figures - 2
 
Figure 4-1: Dose-Response Assessment for non-target terrestrial invertebrates. 

Figures - 3
 
Figure 4-2: Distribution of sensitivity in various non-target lepidoptera (data from Peacock et 

al. 1998) 

Figures - 4
 
Table 2-1: Commercial formulations of B.t.k. that may be used in Forest Service Programs 1 

Formulation/ 

Producer 

Type of 

formulation 

% a.i. 2 Potency Application Rates 3 Type 

application 

Biobit HP/ 

Valent USA Corp 

Wettable 

power 

6.4 32,000 IU/mg 

14.52 BIU/lb 

0.5-2 lb/acre Ground or 

aerial 

DiPel DF/ 

Valent USA Corp 

Dry flowable 10.3 32,000 IU/mg 

14.5 BIU/lb 

0.5-2 lb/acre Ground only 

DiPel ES/ 

Valent USA Corp 

Emulsified 

suspension 6 

3.5 17,600 IU/mg 

64 BIU/gallon 

1-4 pints/acre Ground only 

DiPel Pro DF/ 

Valent USA Corp 

Dry flowable 10.3 32,000 IU/mg 

14.5 BIU/lb 

1-4 lb/100 gallons Ground only 

DiPel 2X/ 

Valent USA Corp 

Wettable 

powder 

6.4 32,000 IU/mg 

14.52 BIU/lb 

0.5-2 lb/acre Ground or 

aerial 

Foray 48B/ 

Valent 

BioSciences 

Flowable 

concentrate 

2.1 10,600 UI/mg 

48 BIU/gallon 

1.3-6.7 pts/acre 

8-40 BIU/acre 

Ground or 

aerial 

Foray 48F/ 

Valent 

BioSciences 

Flowable 

concentrate 

5.7 11,800 FTU/mg 

48 BFTU/gallon 

21-128 oz/acre 

8-48 BFTU/acre 

Ground or 

aerial 

Foray 76B/ 

Valent 

BioSciences 

Flowable 

concentrate 

3.3 16,700 IU/mg 

76 BIU/gallon 

13.5-67.5 oz/acre 

8-40 BIU/acre 

Ground or 

aerial 

5

 Thuricide 48LV/ 

Valent 

BioSciences 

Aqueous 

concentrate 

2.4 48 BIU/gallon 14-87 oz/acre 

8-40 BIU/acre 

Ground or 

aerial 

5

 Thuricide 76LV/ 

Valent 

BioSciences 

Aqueous 

concentrate 

14.4 76 BIU/gallon 14-67 oz/acre 

8-40 BIU/acre 

Ground or 

aerial 

1

 Source: Specimen labels from C&P Press, 2001.


2 Includes B.t.k. solids, spores, and toxins. The remainder of the product formulation is classified as inerts. See
 

text for discussion.
 

3

 All application rates expressed in amount (lb or oz) of formulation not amounts of active ingredient.


4 Potency expressed as Forestry Toxic Units (FTU). Application rate corresponds to approximately 0.16 to 1
 

gallons/acre.


5

 Information based on Certis (2002) labels.
 

6

 Oil based formulation
 

Tables - 1 
TABLE 2-2: Use of B.t.k. from 1995 to 2001 for Suppression, Eradication, 

and Slow the Spread 1 

 Source: GMDigest, Morgantown, WV 

Year 

1995 

Suppression 

271,961 

Eradication 

332,276 

Slow the Spread 

32,528 

1996 201,540 154,572 18,949 

1997 46,703 200,720 18,744 

1998 91,672 174,840 34,534 

1999 153,198 164,856 7,252 

2000 227,688 1,996 84,127 

2001 273,384 1,440 62,398 

2002 149,772 9,961 28,705 

1 

Total 1,415,918 1,040,661 287,237 

(http://fhpr8.srs.fs.fed.us/wv/gmdigest/gmdigest.html) 

Total 

636,765 

375,061 

266,167 

301,046 

325,306 

313,811 

337,222 

188,438 

2,743,816

Tables - 2 
Table 3-1: Epidemiology Studies on B.t.k. Formulations 

Formulation, Location, Observations, Response Reference(s) 

Population, Exposure 

Dipel, Oregon, USA, about 

80,000 residents in spray area, 

3 applications at 16 BIU/acre. 

About 180,000 residents in 

unsprayed area. 

Foray 48B, British Columbia, 

Canada, residents in sprayed 

and unsprayed areas and 

workers, 20.2 BIU/acre. 

Javelin (B.t.k. 17 BIU per lb), 

application rate not specified 

but probably in range of 2 

BIU/acre to 25.5 BIU/acre, 

workers harvesting treated 

crops (groups of 20 to 48) 

Foray 48B, Auckland, New 

Zealand, 88,000 residents in 

sprayed area, 4.3 pints per acre 

(about 0.5375 gal./acre or 25.8 

BIU/acre). Multiple 

applications in different areas. 

Foray 48B, British Columbia, 

Canada, 29 children in spray 

area and 29 children in 

unsprayed area, 3.4 pints/acre 

(about 0.425 gal./acre or 20.4 

BIU/acre), 3 applications over 

10 days. 

Foray 48B, Auckland, New 

Zealand, 292 individuals 

surveyed before and after 

spray, 4.3 pints per acre (about 

0.5375 gal./acre or 25.8 

BIU/acre). Three applications. 

Surveillance program in four clinical laboratories for B.t.k. 

in clinical samples. Seven B.t.k. in clinical samples (other 

than incidental contamination) in sprayed area. None in 

unsprayed area. No significant adverse effects. 

Survey of 1,140 visits to family practice physicians and 

3,500 hospital admissions. Analysis of Bacillus isolates. 

B.t.k. not implicated as disease agent. Cellular fatty acid 

profiles of B.t.k. cultures from humans as well as plants 

differed from B.t.k. in formulation. Some workers 

involved in ground applications evidenced nasal swabs 

positive for B.t.k. for up to 120 days after application. 

Respiratory and dermal irritation in workers. 

No signs of respiratory impairment or other adverse 

effects associated with exposure. A significant increase in 

skin-prick test responses to B.t.k. 1-4 months after 

exposure. Increase in IgE antibodies in highest exposure 

groups consistent with a potential for allergic sensitization. 

Surveillance program of sentinel physicians. Selfreporting survey of adverse effects after exposure. 

Surveillance of births and incidence of meningococcal 

disease and reported infections. Self-reports of headache 

and respiratory irritation (sore throat). No effects 

demonstrated in review of sentinel physicians. 

No differences between the children (all with a history of 

asthma) in treated and untreated areas in terms of asthma 

symptoms or peak respiratory flow rates. No increase in 

symptoms of asthma in either group after spray. 

Increase in incidence of B.t.k. HD-1 from nasal swabs 

after B.t.k. spray. Relatively few B.t.k. HD-1 identified in 

water (2.9%). 

Self-reports before spray (n=292) and after spray (181 of 

292 respondents). Increase in symptoms grouped as 

irritant, gastrointestinal, and neuropsychiatric effects that 

were significant at p<0.05 based on pair-wise 

comparisons. 

Elliott et al. 

1988; Elliott 

1986; Green et 

al. 1990 

Cook (1994); 

Noble et al. 

(1992) 

Bernstein et al. 

1999 

Aer’aqua 

Medicine Ltd. 

2001 

Pearce et al. 

2002 

Valadares de 

Amorim et al. 

2001 

Petrie et al. 

2003 

Tables - 3 
 

 

 

Table 3-2: Publically available information on inerts used in B.t.k. formulations. 

Ingredient	 Description 

Benzoic acid/sodium CAS No. 65-85-0. GRAS compound and approved food additive. Functions in pH 

benzoate 1 control and as an antimicrobial (Clydesdale 1997). 

Hydrochloric acid 1	 CAS No. 7647-01-0. GRAS compound and approved food additive. Functions in pH 

control (Clydesdale 1997). 

Methyl paraben 1,2 CAS No. 7775-19-1. U.S. EPA List 3 Inert 3. Uses: Pharmaceutical aid (antimicrobial 

(methyl preservative). Used in some suntan lotions, hand lotions, and bubble bath 

hydroxybenzoate) formulations. Occurs naturally in some berries and fruits (Burdock et al. 2002). There 

appears to be adequate data on this compound to remove it from List 3. 

Phosphoric acid	 CAS No.7664-38-2. GRAS compound and approved food additive. Functions in pH 

control, fermentation aid, fumigant, antimicrobial, and sweetener (Clydesdale 1997). 

Polyacrylic acid CAS No.25987-55-7 (calcium polyacrylate). U.S. EPA List 3 Inert 3

. Toxicity data on 

(carbopol) 1 this compound appears to be incomplete. 

Potassium phosphate 2	 CAS No.7778-77-0. GRAS compound and approved food additive. Functions in pH 

control agent, nutrient supplement, stabilizer or thickener, malting or fermenting aid 

(Clydesdale 1997). 

Potassium sorbate 1	 CAS No. 24634-61-5. GRAS compound and flavoring agent. Functions as 

antimicrobial agent, pH control agent, antioxidant, flavor Flavoring agent or adjuvant, 

nutrient supplement, or coloring adjunct (Clydesdale 1997). 

Propylene glycol 1	 CAS No. 57-55-6. GRAS compound and food additive. Functions as solvent 

antimicrobial agent, anti-caking agent or free-flow agent, drying agent, flavoring agent 

or adjuvant, antioxidant, emulsifier, or formulation aid (NOS) (Clydesdale 1997). 

Sodium hydroxide 2	 CAS No. 1310-73-2. GRAS compound and food additive. Functions as pH control 

agent, processing aid, fumigant, washing or surface removal agent, dough 

strengthener, flour treating agent, oxidizing or reducing agent, flavoring agent, 

coloring adjunct (Clydesdale 1997). 

Sodium sulfite 2	 CAS No.7757-83-7. GRAS compound and food additive. Functions as dough 

strengthener, flour treating agent, oxidizing or reducing agent, color or coloring 

adjunct, ph control agent, antioxidant, formulation aid (NOS) (Clydesdale 1997). 

Sorbitol 1	 CAS No.50-70-4. GRAS compound and food additive. Functions as stabilizer or 

thickener, nutritive sweetener, flavoring agent, drying agent, pH control agent, 

solvent, coloring adjunct, texturizer, nutrient supplement (Clydesdale 1997). 

Sulfuric acid 2	 CAS No.7664-93-9. GRAS compound and food additive. Functions as pH control 

agent, formulation aid, flavoring agent, flavor enhancer, processing aid (Clydesdale 

1997). 

1 Painted Apple Moth Community Coalition (CC-PAM), http://www.moth.co.nz/homepage.htm 2

 Swadener 1994 

3

 The U.S. EPA inerts list is available at http://www.epa.gov/opprd001/inerts/ 

Tables - 4 
 

Table 3-3: Overview of exposure data for workers and members of the general public. 1 

Concentrations of	 Description Reference 

B.t.k. in air 2 

WORKERS 

0.2 to 15.8 × 106	 Highest exposures in ground spray workers. Lower range 

cfu/m3 associated with support personnel – i.e., auditors, public 

relations personnel, and card handlers. 

400 to 11,000 cfu/m 3	 No clear association between applicators (pilots) in aerial 

application and support personnel. Five of 15 workers, 

including one pilot, had no detected exposure. 

GENERAL PUBLIC 

1000 and 1600 cfu/m 3	 Personal air samples of four individuals. Exposure noted in 

two – a grocery store clerk and a service station attendant. 

Two individuals had no detectable exposures (a church 

custodian and a mail carrier). 

200 to 4,200 cfu/m 3	 Twelve general air samples at various locations. No 

colonies in seven samples, some of which were in work area 

– i.e., helicopter loading area. 

739 cfu/m 3	 The average in the spray zone during spraying. 

77 and 244 cfu/m 3	 Average outdoor and indoor concentrations at 5 to 6 hour 

after spraying. Note: Indoor concentrations were higher. 

739-770 cfu/m 3	 96% of samples positive for B.t.k. inside spray area during 

spray. 

484-551 cfu/m 3	 95% of samples positive for B.t.k. outside spray area during 

spray. 

1

See Table 3-1 for a description of the epidemiology studies. 

2

Excluding non-detects which are discussed in the description column. 

Cook 1994 

Elliott et al. 1988, 

Elliott 1986 

Elliott et al. 1988, 

Elliott 1986 

Elliott et al. 1988, 

Elliott 1986 

Teschke et al. 2001 

Teschke et al. 2001 

Valadares de Amorim 

et al. 2001 

Valadares de Amorim 

et al. 2001 

Tables - 5 
Table 3-4: Post-spray symptoms reported by ground-spray workers and controls 1 

Number (%) 

Symptom 

Controls (n=29) Workers 

(n=120) 

p-value 2 

Dermal (dry or itchy skin, chapped lips) 3 (10%) 41 (34%) 0.007630 

Eyes (redness, itch, burning, puffiness) 4 (13%) 24 (20%) 0.317398 

Headache 3 (10%) 8 (7%) 0.858536 

Throat (dry, sore) 2 (7%) 35 (29%) 0.007868 

Runny nose or stuffiness 4 (13%) 32 (27%) 0.109883 

Respiratory (cough, tightness) 1 (3%) 24 (20%) 0.021899 

Digestive (nausea, diarrhea) 3 (10%) 8 (7%) 0.858536 

Total (all symptoms combined) 11 (38%) 76(63%) 0.011638 

1

 Data from Cook (1994), Table 3, p. 22.
 

2 p-value calculated using Fischer-Exact Test [p-value = 0.05 ÷ 7 = 0.0071].
 

Tables - 6 
Table 3-5: Summary of the number of symptoms per worker in 120 ground-spray workers

segregated by exposure groups and use of protective masks 1 

Mask Use 3 

2 3 6 

Exposure Group 2 

Regular Occasional None 

<1 to 100 1.7 [3] 3.7 [7] 1.5 [33] 

101 to 300 2.0 [3] 3.3 [3] 1.4 [43] 

> 300 2.0 [1] 4.0 [3] 2.8 [24] 

1

 Data from Cook (1994), Table 3, p. 23. 

B.t.k. exposure in cfu/m × 10 × hours
 3

 Number of symptoms per worker [number of workers per group]
 

Tables - 7 
Table 3-6: Self-reported symptoms in individuals before and after the aerial application 

of B.t.k. 1 

Baseline After Spray Reported p- Fisher Exact Health Problem (n of 292) (n of 181) value Test 

Headache 133 93 0.06 0.127201 

Back pain 

Coughing 

Cold, flu 

Sleep problems 

Neck pain 

Leg pain during physical activity 

Shoulder pain 

Arm pain 

Stomach discomfort 

105 

85 

84 

78 

70 

69 

59 

50 

48 

57 

60 

54 

66 

45 

35 

43 

34 

46 

0.06 

0.1 

0.6 

0.03 

0.89 

0.37 

0.26 

0.48 

0.03 

0.863310 

0.204836 

0.441418 

0.016637 

0.454930 

0.887366 

0.211994 

0.366523 

0.012472 

Irritated throat 47 58 0.0001 0.000048 

Itchy nose 

Migraine 

Dizziness 

47 

37 

32 

42 

27 

31 

0.04 

0.18 

0.01 

0.036631 

0.287439 

0.038634 

Wheezing 

Diarrhoea 

29 

27 

24 

30 

0.11 

0.03 

0.167014 

0.013527 

Gas discomfort 25 30 0.02 0.006847 

Chronic eye irritation 

Eczema 

24 

23 

25 

13 

0.07 

0.99 

0.038379 

0.671774 

Pain in ears 23 19 0.49 0.208708 

Chest pain 

Extra heartbeats 

21 

20 

16 

19 

0.49 

0.05 

0.315260 

0.110163 

Constipation 

Difficulty concentration 

Blurred or double vision 

18 

15 

15 

12 

23 

18 

0.32 

0.001 

0.2 

0.491525 

0.003170 

0.036674 1

 The number of responders per effect is based on the percent responses and numbers of individuals 

reported in Petrie et al. 2003. The p-values in column 3 are those reported by Petrie et al. (2003). Fisher 

exact tests calculated on-line at http://www.matforsk.no/ola/fisher.htm. [p-value 0.05 ÷ 25 = 0.002] 

Tables - 8 
Table 3-7: Exposure conversions for mice and humans with effects noted in mice after

intranasal instillations. 

cfu/mouse Mouse 

cfu/m × hour 3 (1) 

Equivalent 

cfu/person (2) 

Equivalent human

cfu/m × hour 3 (3) 

Effect in Mice (4) 

1e+02 

1e+04 

1e+07 

7.14e+07 

7.14e+09 

7.14e+12 

3.5e+05 

3.5e+07 

3.5e+10 

1.4e+05 

1.4e+07 

1.4e+10 

inflamation, no 

mortality 

1e+08 7.14e+13 3.5e+11 1.4e+11 80% mortality 

(1) Based on a breathing rate of 0.0014 L/hour for a 0.020 g mouse, derived from U.S. EPA (1988a), 

Recommendations for and Documentation of Values for Use in Risk Assessment, Table 1-3, p. 1-11: L/day = 1.0496 3 3 1.99 Bwkg . Note that 0.0014 L/hour is equivalent to 0.0000014 m /hour [1 m = 1000 L ] or 

3 0.0000336 m /day.

(2) cfu/mouse × 70 kg/0.02 kg. (3) 3 Based on a human breathing rate for moderate activity of 2.5 m /hour from U.S. EPA (1989d), Exposure

Factors Handbook, Table 3-1, p. 3-4.

(4) From Hernandez et al. (1999, 2000), intranasal instillations in mice without exposure to influenza virus. 

Tables - 9 
Table 3-8: Risk characterization for serious health effects from exposure to B.t.k. 

Exposure cfu/m3 Duration 

(hours) 

Cumulative 

Exposure 

(hours × cfu/m )3 

Hazard Index 

General public, 

lower range 100 24 2,400 0.00000024 

upper range 5,000 24 360,000 0.000036 

Aerial Workers, 

lower range 400 8 3,200 0.00000032 

higher range 11,000 8 88,000 0.000009 

Ground Workers, 

lower range 200,000 8 1,600,000 0.00016 

higher range 15,800,000 8 126,400,000 0.01264 

extreme range 400,000,000 0.04 

Human NOAEL 1.00e+10 hours × cfu/m3 

Tables - 10 
Table 4-1: Mortality in species subject to foliage treated with Foray 48B at 89 BIU/ha

(Peacock et al. 1998). 

Family Species Instar 1 

Control 

No. 

Alive 

No. 

Dead 

Foray 48B at 

89 BIU/ha 

No. 

Alive 

No. 

Dead 

p-value 3 

Papilionidae, 

Swallowtail 

Butterflies 

Papilio glaucus 1-3 10 0 0 20 <0.00001 

Nymphalidae, 

Danaid and 

Brown 

Butterflies 

Speyeria diana 

Limenitis arthemis astyanax 

Astercampa clyton 

2-3 

n/n-1 

4-5 

10 

10 

21 

0 

0 

1 

1 

0 

1 

15 

20 

40 

<0.00001 

<0.00001 

<0.00001 

Geometridae, 

Looper 

Butterflies 

Alsophila pometaria 

Phiglia titea 

Euchlaena obtusaria 

n 

n/n-1 

n-1 

19 

20 

12 

1 

0 

0 

11 

43 

18 

7 

7 

0 

0.0164 

0.1801 

1 

Ennomos magnaria 

Ennomos magnaria 

Lambdina fervidaria 

1 

1 

1 

22 

17 

17 

1 

14 

1 

0 

0 

10 

66 

27 

26 

<0.00001 

<0.00001 

<0.00001 

Eutrapela clemataria 

Prochoerodes transversata 

H 2 

2 

20 

19 

0 

1 

4 

28 

31 

13 

<0.00001 

0.0237 

Lasiocampidae, 

Lappet Moths 

Malacosoma disstria 

Malacosoma disstria 

2 

n 

23 

20 

4 

0 

4 

1 

26 

44 

<0.00001 

<0.00001 

Saturniidae, Silk 

Moths 

Hemileuca maia H 47 0 5 53 <0.00001 

Hemileuca maia 1 70 1 48 312 <0.001 

Hemileuca maia 1 20 0 0 51 <0.00001 

Hemileuca maia 2 109 1 0 111 <0.00001 

Antheraea polyphemus 

Actias luna 

1 

1 

16 

26 

4 

14 

3 

0 

57 

96 

<0.00001 

<0.00001 

Lymantriidae, 

Tussuck Moths 

Dasychira obliquata 4 20 0 27 1 0.9999 

Noctuidae, 

Owlet moths 

Noctuidae, 

Owlet moths 

(continued) 

Amphipyra pyramidoides 

Amphipyra pyramidoides 

Xystopeplus rufago 

Psaphida rolandi 

Psaphida resumens 

Egira alternans 

Egira alternans 

Zale aeruginosa 

Eupsilia vinulenta 

Eupsilia vinulenta 

Sericaglaea signata 

Metaxaglaea semitaria 

Chaetaglaea sericea 

Chaetaglaea sericea 

Sunira biclorago 

Sunira biclorago 

n-1 

n-1 

1,2 

n-1 

1,2 

1 

2-3 

H 

n-1/n-2 

n-1/n-2 

4 

n 

n-1 

n-1 

n/n-1 

n 

19 

20 

28 

19 

20 

20 

18 

12 

20 

20 

18 

20 

20 

19 

20 

20 

2 

0 

0 

1 

0 

5 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0 

6 

11 

12 

18 

9 

22 

35 

19 

19 

43 

48 

51 

20 

48 

45 

29 

24 

37 

21 

22 

41 

27 

2 

11 

1 

1 

0 

1 

0 

1 

3 

0 

<0.00001 

0.0001 

<0.00001 

0.0001 

<0.00001 

0.0059 

1 

0.0173 

0.9999 

0.9999 

1 

0.9999 

1 

0.9999 

0.5498 

1 

Tables - 11 
Table 4-1: Mortality in species subject to foliage treated with Foray 48B at 89 BIU/ha

(Peacock et al. 1998). 

Family Species Instar 1 

Control 

No. No. 

Foray 48B at 

89 BIU/ha 

No. No. 

p-value 3 

Alive Dead Alive Dead 

Xylotype capax n-1 19 1 48 0 0.2941 

Orthosia alurina n-2 19 1 29 0 0.9999 

Orthosia alurina n-1 18 0 30 7 0.0823 

Orthosia hibisci n-1 20 0 39 0 1 

Abagrotis alternata n/n-1 29 0 50 0 1 

Abagrotis alternata n/–1 18 0 13 0 1 1

 n designates last instar 

2

 H designate hatchling 

3

 Fischer Exact test 

Tables - 12 
Table 4-2: Mortality in species subject to foliage treated with Dipel 8AF at 99 BIU/ha 

(Peacock et al. 1998). 

Family Species Instar 1 

Control 

No. No. 

Dipel 8AF at 

99 BIU/ha 

No. No. 

p-value 3 Compariso 

n to Foray 

Alive Dead Alive Dead 

Geometridae, 

Looper 

Butterflies 

Asterocampa 

clyton 

Alsophila 

pometaria 

4,5 

n 

21 

19 

1 

1 

2 

11 

20 

21 

<0.00001 

<0.00001 Match 

Ennomos 1 17 14 0 47 <0.00001 Match 

magnaria 

Lasiocampid 

ae, Lappet 

Moths 

Malacosoma 

disstria 

2 23 4 0 28 <0.00001 Match 

Lymantriidae 

, Tussuck 

Moths 

Dasychira 

obliquata 

4 20 0 26 0 1 Match 

Noctuidae, 

Owlet moths 

Catocala vidua 1 17 2 0 31 <0.00001 

Amphipyra 

pyramidoides 

n-1 19 2 3 35 <0.00001 Match 

Lithophane grotei n-1/n-2 20 0 22 28 <0.00001 

Lithophane 

unimoda 

n-1 19 1 38 9 0.1423 

Eupsilia vinulenta n-2 20 0 19 9 0.0063 No match, 

different 

instars 

Chaetaglaea 

sericea 

n-1 20 0 30 0 1 Match 

Sunira biclorago n/n–1 20 0 41 0 1 Match 

Orthosia alurina n–2 19 1 14 4 0.1698 Match 

Abagrotis 

alternata 

n/–1 18 0 31 1 0.9999 Match 

1

 n designates last instar 

2

 H designate hatchling 

3

 Fischer Exact test 

Tables - 13 
Table 4-3: Summary of exposures used in ecological risk assessment. 

Organism 

Small mammal 

Terrestrial 

Invertebrates 

Exposure(s) 

Inhalation: 100 to 5000 cfu/m3

 or 

0.00336 to 0.168 cfu/mouse

Food/Water/Dermal: 184 mg/kg bw 

20 to 40 BIU/acre [49 to 99 BIU/ha] 

Section 

4.2.2.1. 

4.2.2.2. 

Aquatic Species 0.24 mg formulation/L

7680 IU/L 

4.2.4. 

Tables - 14 
Table 4-4: Summary of toxicity values used in ecological risk assessment. 

Organism Toxicity Value(s) Section 

Small mammal Inhalation 4.3.2.1. 

7 10 cfu/mouse – NOAEL and 3.3.4 

8 10 cfu/mouse – Frank Effect Level

Oral 

8400 mg/kg/day – NOAEL 

Terrestrial Sensitive Species: 21 BIU/ha [. 8.4 BIU/acre] LD 50 4.3.2.2. 

Insects Tolerant Species: 590 BIU/ha [.240 BIU/acre] LD50

(see text for discussion dose-response curves) 

Fish Sensitive Species: 1.4 mg formulation/L or 4.3.3.1. 7 1.51×10 cfu/L – LOEC

Tolerant Species: 1000 mg formulation/L or

2.5×1010 cfu/L – NOEC 

Aquatic Sensitive Species: 0.45 mg/L or 4.3.3.2. 

8 Invertebrates 6.24×10 cfu/L – NOEC

Tolerant Species: 36 mg/L – NOEC 

Tables - 15 
Table 4-5: Data used in dose-response assessment for non-target insects. 

Mortality 

Exposure Control Exposed Attributable to 

Common Name (BIU/ha) Response Response B.t.k. Reference 

Sensitive Insects 

Gypsy moth 1st instar 33.5 0.2 0.67 0.5875 Herms et al. 1997 

Gypsy moth 1st instar 90 0.2 0.95 0.9375 

Karner blue butterfly larvae 33.5 0 0.72 0.72 

Karner blue butterfly larvae 90 0 0.86 0.86 

Swallowtail butterfly larvae 40 0.67 0.94 0.8182 Johnson et al. 1995 

Swallowtail butterfly larvae 40 0.58 0.93 0.8333 

Promethea moth larvae 40 0.66 0.89 0.6765 

Cabbage looper larvae 16 0 0.5 0.5 James et al. 1993 

Cinnabar moth, 4th instar 26 0 0.5 0.5 

Cinnabar moth, 5tht instar 19 0 0.5 0.5 

Tolerant Insects 

Cinnabar moth, 1st instar 427 0 0.5 0.5 James et al. 1993 

Cinnabar moth, 2nd instar 437 0 0.5 0.5 

Cinnabar moth, 3rd instar 575 0 0.5 0.5 

Green lacewing, larvae 79 0.116 0.135 0.0215 Haverty 1982 a 

Green lacewing, adult 79 0.037 0.056 0.0197 

Green lacewing, larvae 158 0.116 0.175 0.0667 

Green lacewing, adult 158 0.037 0.088 0.0530 

Lady beetle, adult 158 0.335 0.424 0.1338 

Other Insects b 

Honey bee, adult worker 25 0 0.127 0.127 Atkins 1991a a 

50 0 0.192 0.192 

75 0 0.191 0.191 a These studies involved direct spray of adults or larvae as specified in column 1. All other studies involved
 

consumption of contaminated vegetation by larvae.


b Not used quantitatively in dose-response assessment. See text for discussion.
 

Tables - 16 
Table 4-6: Risk characterization for ecological risk assessment of B.t.k. 

Species Scenario or 

Group 

Exposure Toxicity

Value 

Risk 

Characterization 1 

Small Mammal Inhalation 0.168 cfu 10 cfu 7 HQ = 2×10 -8 

Oral/Dermal 184 mg/kg 8400 mg/kg HQ = 0.02 

Terrestrial Insects Sensitive Species 49 to 99 BIU/ha Dose-response 

curve 2 

80% to 94% [Probit 

5.84 to 6.55] 

Tolerant Species 0.6% to 3.6% 

[Probit 2.47 to 3.19] 

Other terrestrial 

invertebrates 

All No effects anticipated from B.t.k. Oil based formulations may 

cause adverse effects in some soil invertebrates. 

Fish Sensitive Species 0.24 mg/L 1.4 mg/L HQ = 0.2 

Tolerant Species 1000 mg/L HQ = 0.0002 

Aquatic 

Invertebrates 

Sensitive Species 

Tolerant Species 

0.24 mg/L 0.45 mg/L 

36 mg/L 

HQ = 0.5 

HQ = 0.007 

1

 For all groups except terrestrial invertebrates, the risk characterization is given as the hazard quotient (HQ),
 

the exposure divided by the toxicity value.


2 Estimated mortality based on dose response equation: Y = -1.48 + 2.34 x + 3.36 S. In this equation, Y is the
 

probit response, x is the common log of the application rate in BIU/ha, and S is equal to 1 for sensitive species
 

and 0 for tolerant species. See text for discussion.
 

Tables - 17 
APPENDICES 

Appendix 1: Toxicity in Mammals

Appendix 2: Toxicity in Birds

Appendix 3: Toxicity in Non-target Lepidoptera

Appendix 4: Toxicity in Non-Target Terrestrial Insects Other Than Lepidoptera

Appendix 5: Toxicity of B.t.k. and B.t.k. Formulations to Fish 

Appendix 6: Toxicity of B.t.k. and B.t.k. Formulations to Aquatic Invertebrates 
Appendix 1: Toxicity in Mammals 

Product Species/Exposure Observations Reference 

ORAL 

DiPel 

“technical 

material” 

Rat/Sprague-Dawley, 

21/male 

21/female, 10 cfu, 8 

gavage 

No mortality and no signs of toxicity. Total 

clearance estimated at 47 days based on fecal 

excretion. Some samples from tissues (kidney and 

spleen) contained B.t.k. but this was seldom 

demonstrated on duplicate plates. This was also seen 

in some control animals and attributed to 

contamination of plates. 

David 

1990b 

DiPel 

Technical 

Powder 

Rat/Sprague-Dawley, 

4/male 

5/female, 5050 mg/kg 

gavage 

Mortality in one male rat on Day 1, probably due to 

aspiration of material during dosing. No treatment 

related signs of toxicity. 

Bassett and 

Watson 

1999a 

Dipel ES Rat/Sprague-Dawley, 

5/male 

5/female, 5050 mg/kg 

gavage 

No mortality, no gross pathology, and no clinical 

signs of toxicity. 

Kuhn 1998b 

Foray 48B Rat/Sprague-Dawley, 

5/male 

5/female, 5000 mg/kg 

gavage 

No mortality; no clinical signs; no abnormalities at 

necropsy. 

[Identical data cited in summary by Berg et al. 

1991.] 

Cuthbert 

and Jackson 

1991 

Foray 76B Rat/HSD, 5/male 

5/female, 5050 mg/kg 

gavage 

No mortality; all rats appeared normal for the 

duration of the study; gross necropsy revealed no 

abnormalities in any of the rats 

Kuhn 1991 

Foray 48B Rat/Wistar 14/male 

14/female, 1 mL/rat 

No mortality; there was no treatment related 

pathology; after 4 days, B.t.k. was isolated from the 

lungs and spleen in one rat, which indicates a 

technical error at dosing; two other rats also showed 

the microorganism in the lungs after 15 and 22 days, 

respectively; the microbial count in feces decreased 

rapidly during the first 3 days after exposure. 

Harde 

1990a 

B.t.k. 

(NOS) from 

Novo 

Nordisk 

Rats, SPF Wistar, 

4M/4/F, 1 mL dose (cfu 

counts in dose illegible 

on fiche). Gavage 

No mortality or signs of toxicity. No B.t.k. found in 

blood. B.t.k. in feces and organs dropped by a factor 

of 100 in 24 hours. 

Harde 

1990a 

sB.t.k. 

powder 

Rats, Wistar, 10 cfu per 8 

rat, gavage. Groups of 

3-4 rats per sex 

No effect on mortality, organ weights, gross 

pathology, and clinical signs. B.t.k. not found in 

blood of any animal. B.t.k. decreased by factor of 

about 100 per day. No indication of infectivity 

based on microbial counts in kidney, liver, spleen, 

lymph nodes, lungs, brain, blood and feces. 

Harde 

1990b 

B.t.k. Rats, HA albino. 

20M/20F, 7.5×10 ,7 

1×10 , 1.25×10 6 6 

spores/rat, single oral 

dose (presumably 

gavage) 

No signs of toxicity over 21-day observation period 

based on mortality, body and organ weights, clinical 

biochemistry and hematology, and reflexes. 

Meher et al. 

2002 

Note on Meher et al. 2002: B.t.k. characterized as a wettable powder formulation produced in India. 

Appendix 1-1 
 

Appendix 1: Toxicity in Mammals 

Product Species/Exposure Observations Reference 

DERMAL 

Dipel ES Rabbits, 5/male 

5/female, 5050 mg/kg, 

intact skin 

No mortality. Decreased body weight in 6 animals. 

Signs of dermal irritation included erythema, edema, 

and desquamation. 

Kuhn 1998b 

Dipel ES Rabbits, 3/male 

3/female, 0.5 mL, intact 

skin, covered with 

patch. Removed after 6 

hours. 

Very slight erythema a 1 and 24 hours. Kuhn 1999a 

NOTE on Kuhn 1998b and Kuhn 1999a: Study titles on title page indicate that the studies were done 

on rats. This is clearly an error. The studies were conducted on New Zealand White rabbits. 

B.t.k. 

formulation 

Rabbits, albino. 6M/6F, 

2.5×107

 spores in 1 mL 

on shaved and abraded 

skin 

“Low-grade” reddening of skin which reversed after 

72 hours. No signs of toxicity over 21-day 

observation period. 

Meher et al. 

2002 

B.t.k. 

formulation 

Rabbits, albino. 6M/6F, 

5×107

 spores in 0.5 mL 

on shaved and abraded 

skin. Treated area 

covered. 

“Low-grade” reddening of skin which reversed after 

72 hours. 

Meher et al. 

2002 

DiPel 

Technical 

Powder 

Rabbits, 

6/female, 0.5 g on 

abraded skin 

Well-defined erythema at 30 minutes to 24 hours in 

3 rabbits, which reduced during the 14-day period. 

On rabbit with initial slight erythema from 30 

minutes had well-defined erythema by Day 14. 

Bassett and 

Watson 

1999b 

Foray 48B Rabbit/Mol: Russian, 

6/female, 0.5 mL, 4 

hours 

Very slight erythema in one rabbit Jacobsen 

1993 

Foray 48B Rabbit, 10 cfu/rabbit 10 Mild irritation which cleared after 4 days. Berg et al. 

1991 

Foray 76B Rabbit/New Zealand 

White, 5/male 

5/female, 2.0 g 

(1x1010 units/rabbit), 24 

hours 

No systemic effects; only mild skin reactions that 

cleared within 2 days after exposure. Behavior and 

appearance of all rabbits were normal throughout the 

study; agent was classified as "mild irritant" 

Kiehr 1991a 

OCULAR 

Dipel ES Rabbits, 3M/3F, 0.1 mL 

formulation in right eye 

for 1 minute and then 

washed. 

At 1 hour post-exposure, redness in conjunctiva of 2 

rabbits. Normal after 24 hours. No other effects on 

conjunctiva, iris, or cornea. 

Kuhn 1999b 

Foray 48B 

(Batch 

BBN 6056) 

Rabbit/New Zealand 

White, 6/male, 0.1 mL 

Conjunctival reactions in the form of redness and 

discharge that cleared within 7 days after application 

Berg 1991a 

Appendix 1-2 
Appendix 1: Toxicity in Mammals 

Product Species/Exposure Observations Reference 

Foray 48B 

(Batch 

BBN 6057) 

Rabbit/New Zealand 

White, 6/male, 0.1 mL 

At day 7 mild redness was seen in 3/6 rabbits 

accompanied by small amounts of discharge in one 

of them; at day 8 mild redness was still seen in 1 

rabbit and small of amounts of discharge were seen 

in another; lesions were temporary and cleared 

within 9 days after application. 

Berg 1991b 

Foray 48B 

(Batch 

BBN 6057) 

Rabbit/New Zealand 

White, 6/male, 0.1 mL 

Substantial conjunctival reactions; lesions were of 

temporary nature and cleared within 10 days after 

application 

Berg and 

Kiehr 1991 

B.t.k. 

formulation 

Rabbits, albino. 3M/3F, 

2.5×10 spores in 0.1 6 

mL into one eye. 

No signs of irritation or other effects over 14-day 

observation period. At 14 days but not 20 day, B.t.k. 

could be detected in cultures from the treated eye. 

Meher et al. 

2002 

INHALATION 

B.t.k. 

(Biobit 

concentrate) 

Rats, Sprague-Dawley: 

14M/14F per dose. 

0.47 and 2.17 mg/L, 4 

hours, nose only. 

No mortality. Respiratory depression during 

exposure. Transient body weight loss. Dose related 

increase in mottled lungs. Poorly eliminated from 

lungs over 28 days – i.e., very little change at low 

dose and decrease by a factor of about 10 at high 

dose (Appendix 3 of study). 

Oshodi and 

Macnaughtan 

1990a 

Note: Oshodi and Macnaughtan 1990c has different MRID number but appears to be identical to 

Oshodi and Macnaughtan 1990a. Probably two different submissions. 

Dipel ES Rats, Sprague-Dawley: 

5M/5F. 2.95 mg/L for 4 

hours. 

No mortality or clinical signs of toxicity. Gross 

necropsy noted discolored lungs in one male and two 

females. 

Leeper 

1999a 

Dipel 

Technical 

Powder 

Rat/Sprague-Dawley, 

5/male 

5/female, 5.95 mg/L for 

4 hours. Whole body. 

No mortality. Decrease in activity and piloerection 

on Day 1 only. No signs of toxicity over 14-day 

observation period. 

Leeper 

1999b 

Foray 76B Mice (M/F): aerosol 

whole body exposure, 4 

hours, 

3.22 mg/L. (3.13x109 

cfu/L) 

Decreased activity, alopecia, piloerection, polyuria. 

Alopecia at necropsy was considered unusual and 

possibly related to exposure; no rats died during the 

study; during exposure period the rats were heavily 

coated with the thick test material. 

Holbert 

1991 

Foray 48B Rat/Sprague-Dawley, 

14/male 

14/female, 0.47 mg/L 

for 4 hours 

Respiratory depression during exposure; wet and 

unkempt appearance after exposure; gross pathology 

included mottled lungs (sometimes dark) in a 

majority of rats; histopathology revealed alveolitis, 

interstitial pneumonitis, perivascular eosinophils and 

focal intra-alveolar hemorrhage; minimal 

bronchiolitis was observed in a few animals. 

Oshodi and 

Macnaughtan 

1990b 

Foray 48B Rat/Sprague-Dawley, 

5/male 

5/female, 6.81 mg/L for 

4 hours, nose only 

There was no mortality; necropsy revealed no 

observable abnormalities; all values for lung:body 

weight ratio were within normal limits 

McDon-ald 

and Scott 

1991 

Appendix 1-3 
 

Appendix 1: Toxicity in Mammals 

Product Species/Exposure Observations Reference 

INTRATRACHEAL 

Dipel 

technical 

powder, 

2.01×1010 

spores/g 

Rat/Sprague-Dawley, 

0.06 mL of 9×10 or 9 

1.55×1010 cfu/mL to 

groups of 9M/9F and 

24M/24F, respectively. 

Respiratory distress, lethargy, hunched body 

position, and ruffled coat on Day 1. 10/33 males and 

15/33 females died on Day 2. Sporadic deaths 

thereafter. B.t.k. found in spleen, liver, lymph nodes 

and kidney. On necropsy, severe pulmonary 

hemorrhaging and edema. 

Clearance time in surviving animals estimated at 235 

days. 

David 

1990c 

PARENTERAL 

Foray 48B Rat/Wistar, 5/Male, i.v., 

1 mL 

(3x10 cfu/g) 9 

[vehicle=0.9% sterile 

NaCl] 

Four of five rats died within 23 hours. Edema and 

hemorrhages were seen in the pyloric part of the 

stomach in all rats; two rats had enlarged spleens; the 

rat that was killed had a necrotic tail and extensive 

oedema and hemorrhages on the hindquarters 

stretching down on the hind legs. 

Berg 1990 

Foray 48B Rat/Wistar, 16/Male, 

16/Female, iv, 1 mL 

(4x10 cfu/g) 8 

[vehicle=0.9% sterile 

NaCl] 

No mortality; transient decreased motor activity and 

cyanotic appearance 30 minutes after exposure; 

enlarged spleen in 2 rats; treatment-related 

unspecific reactive hepatitis; A higher incidence of 

histopathological findings in the liver and the 

reticuloendothelial system was found in the treated 

group compared to the controls. These were 

attributed to a background viral infection suggesting 

that the treatment with high levels of B.t.k. 

aggravated a preexisting disease. Over 167 days, a 

complete elimination of the test organism from all 

tissues except the spleen, which on average 

contained 3x10 B.t.k./g at the end of the study. 2 

Berg 1990 

B.t. strain 

SA-3 

Mice, 3M/3F per dose, 

i.p. injections of 10 ,6 

10 , and 10 cfu/mouse. 7 8 

No mortality or clinical signs of toxicity. Schindler 

1990a 

B.t. strain 

SA-3 

Mice, 5M/5F per dose, 

i.p. injections of 10 ,6 

10 , and 10 cfu/mouse. 7 8 

No mortality or clinical signs of toxicity. Enlarged 

spleen and kidney in one female at low dose not 

attributed to treatment. 

Schindler 

1990b 

B.t. strain 

SA-10 

Mice, 5M/5F per dose, 

i.p. injections of 10 ,6 

10 , and 10 cfu/mouse. 7 8 

No mortality or clinical signs of toxicity. Enlarged 

spleen in 1/5, 1/5, and 3/5 animals in the low, mid, 

and high dose groups. Variable changes in kidney 

weight. These effects were not attributed to 

treatment. 

Schindler 

1990c 

B.t. strain 

SA-12 

Mice, 5M/5F per dose, 

i.p. injections of 10 ,6 

10 , and 10 cfu/mouse. 7 8 

4/5 males and 3/5 females died 1 to 3 days after 

injections at the highest dose. Signs of toxicity 

observed in surviving animals – including 

hypoactivity, enlarged spleens, and effects on the 

kidneys. 

Schindler 

1990d 

NOTE: SA-12 is 3a3b, B.t.k. (Chen and Macuga 1990o,p,q) 

Appendix 1-4 
Appendix 1: Toxicity in Mammals 

Product Species/Exposure Observations Reference 

B.t.k. CGA237218 Mice (5M/5F): 10 , 10 , 6 7 

10 cfu/mouse. Five 8 

different production 

batches. 

No mortality in any batch at lowest dose. At middose, no mortality in 3 batches and 10% and 40% 

mortality in two batches. At highest dose, 50% to 

100% mortality. 

Vlachos 

1991 

NOTE: CGA-237218 is not identified in Vlachos (1991) but is clearly identified as B.t.k. in 

Christensen (1991c). 

FIELD STUDIES 

B.t.k. (Dipel 

8L and red 

dye) 

Masked shrew (Sorex 

cinereus) exposed to 

aerial application of 1.8 

L/ha (30 BIU/ha or ca. 

12 BIU/acre) Dipel 8L 

on a 22-year-old jack 

pine plantation in 

northern Ontario 

between May and July 

1989. 

Treatment had no effect on the total abundance of S. 

cinereus; however, the investigators observed 

treatment-related effects on the abundance and diet 

of certain sex and age groups: there were fewer adult 

males and more juveniles in the treated areas, 

compared with the control areas. In addition, adult 

males in the treated area at the same proportion of 

lepidopteran larvae as in the control area, while 

females and juveniles shifted their diet form 

lepidopteran larvae to alternate prey, which may 

have been due to the significant reduction in 

lepidopteran larvae as a result of treatment. 

Belloq et al. 

1992 

B.t.k. 

(Thuricide 

48 LV) 

Populations of small 

rodents and shrews. 20 

BIU/ha (ca. 8 BIU/acre) 

No detectable impact on populations. Innes and 

Bendell 

1989 

Omitted some studies in which the B.t. strain was not identified (Robbins 1991a,b). Omitted studies of Abbott 

ABT-6305 in this and other tables. Abbott ABT-6305 is B.t. aizawai 

(www.epa.gov/pesticides/foia/reviews/006403.htm). 

Appendix 2: Toxicity in Birds 

Product Species/Exposure Observations Reference 

ORAL 

B.t. 

EG2348 

Bobwhite Quail, 

3333mg/kg gavage 

No mortality or signs of toxicity/pathogenicity. Beavers et al. 

1988a 

B.t. 

EG2348 

Mallard Duck, 

3333mg/kg gavage 

No mortality or signs of toxicity/pathogenicity. Beavers et al. 

1988a 

Biobit WP Mallard Duck, 2500 

mg/kg or about 5.7×1011 

cfu/kg for 5 days by 

gavage. 

No signs of toxicity or pathogenicity. Lattin et al. 

1990c 

Biobit WP Mallard duck, 2500 

mg/kg or about 2×1011 

spores/kg by gavage for 

5-days 

No signs of toxicity or pathogenicity. Lattin et al. 

1990g 

Dipel B.t.k. Bobwhite quail, 2857 

mg/kg or about 5.7×1010 

spores/kg for 5 days by 

gavage. 

No signs of toxicity or pathogenicity. Lattin et al. 

1990a 

Dipel B.t.k. Mallard Duck, 2857 

mg/kg or about 5.7×1010 

spores/kg for 5 days by 

gavage. 

No signs of toxicity or pathogenicity. Lattin et al. 

1990b 

Appendix 2-1 
Appendix 2: Toxicity in Birds 

Product Species/Exposure Observations Reference 

Dipel Bobwhite quail, 2857 No signs of toxicity or pathogenicity. Lattin et al. 

Technical mg/kg or about 5.7×1010 1990d 

Material spores/kg for 5 days by 

gavage. 

Biobit B.t.k. Bobwhite quail, 2500 

mg/kg or about 2×1011 

spores/kg for 5 days by 

gavage. 

No signs of toxicity or pathogenicity. Lattin et al. 

1990e 

Biobit B.t.k. Mallard duck, 2500 

mg/kg or about 2×1011 

spores/kg for 5 days by 

gavage. 

No signs of toxicity or pathogenicity. Lattin et al. 1990f 

B.t. Abbott Bobwhite quail, 1714 No signs of toxicity or pathogenicity. Lattin et al. 1990f 

ABG-6305 mg/kg or about 3.4×1011 

cfu/kg for 5 days by 

gavage. 

B.t. Abbott Mallard duck, 1714 No signs of toxicity or pathogenicity. Beavers 1991b 

ABG-6305 mg/kg or about 3.4×1011 

cfu/kg for 5 days by 

gavage. 

Omitted studies by Beavers and Smith 1990a,b on Delta BT. Cannot identify as B.t.k. Omitted Beavers 

1991a,b on B.t. Abbott ABG-6305. This is B.t.a. 

FIELD STUDIES 

B.t.k. Black-throated blue In 1983, caterpillar biomass was significantly Rodenhouse and 

Thuricide warblers (Dendroica different throughout the breeding season in one Holmes 1992 

23LV with caerulesceus), aerial sprayed plot, compared with two unsprayed 

Rhoplex application of 3.5 L/ha plots. Other adverse effects on the reduced 

sticker to four 30-hectare 

forested plots of White 

Mtn. National Forest, 

NH consisting of 

second-growth northern 

hardwoods 

(predominantly sugar 

maple, american beech, 

and yellow birch). The 

study was conducted 

between 1982 and 1985. 

caterpillar plot included significantly fewer 

nesting attempts and significantly fewer 

caterpillars in the diets of nestlings. No 

adverse effects were observed on clutch size, 

hatching success, or the number of fledglings 

per nest in the reduced food site, compared 

with controls. Spraying had no detectable 

effects on caterpillar biomass in 1984 or 1985 

because the natural abundance of caterpillars 

was already low. 

Investigators conclude that neotropical migrant 

bird species are probably limited periodically 

by food when breeding in north-temperate 

habitats. 

B.t.k. Hooded warbler B.t.k. application appeared to have only Nagy and Smith 

(NOS) (Wilsonia citrina) on 

two treatment plots in 

the Arkansas Ozards 

following two 

applications of B.t. in 

1994 

minimal adverse effects on reproduction, in as 

much as the decreased numbers of lepidotperan 

larvae appeared to have a negative effect on 

nestling masses early in the season and 

appeared to alter feeding rates only in small 

clutches. 

1997 

Appendix 2-2 
Appendix 2: Toxicity in Birds 

Product Species/Exposure Observations Reference 

B.t.k. Chestnut-backed and No effects on growth rate of fledgling success Gaddis 1987; 

(NOS) black-capped 

chickadees (Parus 

rufescens, and P. 

atricapillus), 

application of 

unspecified product at 

60 BIU/ha in Portland, 

OR area and 

surrounding counties. 

in 1 year. Reduced fledgling success 2 year st nd 

due to unexplained nest abandonment on 3 

treatment plots (also 1 nest on control plot). 

Significantly smaller proportion of caterpillars 

brought as food on treatment sites both years, 

but provisioning rate no different. 

Gaddis and 

Corkran 1986 

as cited in 

USDA/FS 1995 

B.t.k. , 20 BIU/ha for control of Assay of secondary effects on chicks of spruce Norton et al. 

Thuricide jack pine budworm. grouse (Dendragapus canadensis). Chicks 2001 

48 LV Aerial and hand spray. (dependent on larvae for first two weeks) were 

allowed to graze freely on either treated or 

untreated plots. About a 50% decrease in 

lepidopteran larvae on treated plots. Slower 

growth rate for chicks on treated plots. Based 

on linear slopes (Figure 2), growth rate was 

decrease by about 33%. Attributed to change 

in larvae availability on treated plots. 

B.t.k. , Foray 48B applied at 50 Assayed song bird populations on treated and Sopuck et al. 

Foray 48B BIU/ha. Three 

applications. 

untreated plots before and after applications in 

the same year as well as assay approximately 

one year after applications. In general, no 

adverse effects on songbird populations in 

terms of species richness and relative 

abundance of song birds despite a decrease in 

caterpillar populations. In one species of 42 

species surveyed, the spotted towhee (Pipilo 

maculatus), a statistically significant decrease 

in abundance was noted in the spray year but 

not one year following the spray. 

2002 

Appendix 2-3 
Appendix 3: Toxicity in Non-target Lepidoptera 

Product Species/Exposure Observations Reference 

B.t.k. Spruce budworm No differences in treated or control Buckner et al. 

(Thuricide (Choristoneura fumiferana) plots regarding the number of hand­ 1974 

16B; Dipel exposed to applications of 2 picked larvae from aspen, alder, and 

WP, with or or 4 lbs/acre in Algonquin maple. 

without Park, Ontario and Spruce 

chitinase) Woods Manitoba (Spruce-Fir 

forests). 

B.t.k. (NOS) 32 Species of Lepidoptera on 

tobacco brush (Ceanothus 

velutinusI) treated with 20 

BIU/ha (product not 

specified) in program to 

control spruce budworm 

(Choristoneura occidentalis) 

in Estacada, Clackamas 

County, OR 

Number of larvae on shrubs in treated 

site decreased 80% between pre- and 

post-treatment surveys, compared 

with controls site where the number 

of larvae increased 6% in the same 

time period, 2 weeks after treatment; 

there were no differences between 

spray and control sites 2 months after 

treatment. 

Miller 1990a 

B.t.k. (NOS) 35 Species belonging to 10 

families in the guild of 

nontarget leaf-feeding 

Lepidoptera (caterpillars) on 

Garry oak (Quercus 

garryana) monitored in the 

field from 1986 to 1988 in 

Elmira, Lane County, OR 

after three aerial (via 

helicopter) applications of 16 

BIU/2.8 L water/0.4 ha B.t.k. 

Target species was the gypsy 

moth. 

Target species was significantly 

reduced in treated plots during all 3 

years of the study; species richness 

was reduced in the treated plots 

during all 3 years of the study; and 

the total number of individual nontarget Lepidoptera was significantly 

reduced in treated plots in years 1 

and 2 but not in year 3. 

Miller 1990b 

B.t.k. Thuricide Forest Lepidoptera, aerial Significant decrease in caterpillar Rodenhouse and 

23LV with application of 3.5 L/ha to four biomass in treated plots, compared Holmes 1992 

Rhoplex sticker 30-hectare forested plots of 

White Mtn. National Forest, 

NH consisting of secondgrowth northern hardwoods 

(predominantly sugar maple, 

American beech, and yellow 

birch). The study was 

conducted between 1982 and 

1985. 

with untreated plots, in 1983; no 

significant decreases in caterpillar 

biomass between treated and 

untreated plots in 1984 or 1985 

because natural abundance was 

already low. 

B.t.k. (NOS) Non-target moths in Asian 

gypsy moth eradication 

program in Pierce and King 

Counties, WA exposed to 60 

BIU/ha (24 BIU/acre). 

Full spectrum lights; 49-97% lower 

catches at treated sites in 1993 versus 

same sites in 1992; statistically 

significant decrease; three sites 

(Orthosia hibisci, Protorthodes 

rufula, Perizoma curvilinea) 

eliminated from site? Overall, moth 

diversity unaffected. 

Crawford et al. 

1993 

Appendix 3-1 
Appendix 3: Toxicity in Non-target Lepidoptera 

Product Species/Exposure Observations Reference 

B.t.k. (NOS) Micro-and MacroLepidoptera exposed to 89 

BIU/ha (36 BIU/acre) in 50 

acre plots of oak woodland in 

Rockbridge County, VA 

Sampled in 1992 and 1993. Pre- and 

post (day 6 and 12) foliage samples 

from canopy, subcanopy and shrublayer show reductions in the relative 

abundance of 12/19 most common 

taxa. 12/16 were micro-Lepidoptera. 

In 1992, larval abundance reduced on 

3/5 B.t.k. sites in canopy and 

subcanopy. Reduction in microLepidoptera in 4/5 sites in canopy 

and 3/5 sites in subcanopy. Uneven 

application accounted for variable 

effects. Two plots consistently 

showed the greatest effects. No 

differences observed in total numbers 

of Lepidoptera on foliage in treated 

sites, compared with control sites in 

1993. Micro-Lepidoptera accounted 

for 95% of the individuals collected 

from foliage in 1992 and about 85% 

in 1993. 

Peacock et al. 

1994 

6/8 most common macro-Lepidotpera 

species trapped under burlap bands 

were reduced by treatment. Three of 

these species were nearly absent in 

treated plots (Satyrium calanus, 

Malacosoma disstria, Orthosia 

rubescens). Other less common 

species appeared to be significantly 

less on treated plots. Dasychira 

obliquatc was not affected 

apparently. Noctuidae also lower in 

1993. 

B.t.k. (Foray 

48B) 

Gypsy Moth and non-targets 

lepidoptera (sampled in 1991­

1992) exposed to 14.4 

BIU/ha (36 BIU/acre) 

(sprayed in May 1991) on 24 

50 acre plots in oak, hickory 

with pine, and blueberry 

shrub layer in and Grant and 

Pendleton Counties, WV 

Four treatments: control; B.t. sprayed 

without gypsy moth; B.t. with gypsy 

moth; gypsy moth alone (defoliated). 

Total larval abundance reduced 

following B.t.k. application in 1991. 

No effects of B.t.k. and gypsy moth 

on several Lepidoptera. 

Short-term effects of B.t.k. on nontarget lepidoptera are detrimental but 

longer term effects are beneficial. 

Minor effect on some species of 

lepidoptera consumed by bats 

(Noctuidae and Notodontidae). 

Sample et al. 1996 

Appendix 3-2 
Appendix 3: Toxicity in Non-target Lepidoptera 

Product Species/Exposure Observations Reference 

B.t.k. (Foray 

48B) 

Karner blue butterlfy 

(Lycaeides melissa samuelis) 

larvae (early and late instars) 

reared on wild lupine foliage 

treated in laboratory bioassay 

with B.t.k. at rate of 30-37 or 

90 BIU/ha for 7 days. 

A concurrent laboratory 

bioassay involving gypsy 

moth 2nd instars on similarly 

treated white oak for 7 days. 

Survival rates for Karner blue larvae 

were: 100% for controls, 27% at 30­

37 BIU/ha treatment rate, and 14% at 

90 BIU treatment rate. 

Survival rates for gypsy moth larvae 

were: 80% for controls; 33% for lowdose treatment, and 5% for high-dose 

treatment. 

Investigators conclude that the 

Karner blue is both phenologically 

and physiologically susceptible to 

B.t. used for gypsy moth suppression, 

although the larval generation at risk 

and extent of phenological overlap 

may vary from year to year. 

Herms et al. 1997 

B.t.k. (Dipel: Mulberry silkworm (Bombyx LC50 = 1.40x10 spores/L (larval Jayanthi and 

wettable mori) larvae exposed to instar I) Padmavathamma 

powder) laboratory concentrations of 

1x10, 1x10 , 1x10 , 1x10 , 2 3 4 

1x10 , 1x10 , 1x10 , 1x10 , 5 6 7 8 

or 1x10 spore/mL applied to 9 

mulberry leaves 

LC50 = 4.20x102

 spores/L(larval 

instar II) 

LC50 = 1.0x103

 spores/L(larval instar 

III) 

LC50 = 2.0x105

 spores/L(larval instar 

IV) 

LC50 = 6.3x106

 spores/L(larval instar 

II) 

1997 

Larval mortality was dose-dependent 

with highest % mortality observed at 

highest concentrations of B.t. The 

highest % of mortality was observed 

in the early instars, compared with 

the later instars, and a longer 

incubation period was observed at the 

lower concentrations. The higher 

concentrations of B.t. were associated 

with decreased pupation, greater 

pupal mortality, increased incidences 

of malformed adult emergence and 

lower emergence of normal adults in 

all instars. 

Appendix 3-3 
Appendix 3: Toxicity in Non-target Lepidoptera 

Product Species/Exposure Observations Reference 

B.t.k. (Foray Swallowtail butterflies Significant differences in larval Johnson et al. 

48B) (Papilio glaucus and Papilio 

canadensis) and promethea 

moth (Callosamia 

promethea) (1 and 2 instars st nd 

of the three nontarget species) 

exposed to Foray 48B applied 

at a rate of 40 BIU/ha to 

individual trees using a B.t.­

dedicated backpack sprayer 

to eliminate possibility of 

contamination from other 

insecticides. Larvae were 

placed on the tree at 0 or 1 

day after spray and monitored 

for 7-8 days. 

survival by day 5 between sprayed 

and control trees; nearly all larvae 

died or disappeared by day 8 from 

sprayed foliage. See text for 

additional details. 

1995 

B.t.k. (Foray Long-term persistence field Tree survival was lower in the below- Johnson et al. 

48B) studies in which Foray 48B 

was applied at a rate of 40 

BIU/ha to 5-year-old, 1-2 m 

high potted tulip trees which 

were randomly assigned to 

full sun or below-canopy 

locations in the field sites. 

canopy locations, but the differences 

were not always significant. Toxicity 

toward early instar P. glaucus 

persisted for up to 30 days. 

1995 

Dipel 8AF Laboratory bioassays 

equivalent to application rate 

of 89 BIU/ha. 

18 species of lepidoptera native to 

U.S. 

8 species of larvae (44%) evidenced 

significant mortality. 

Peacock et al. 

1998 

See text and 

Tables 4-1 and 4-2 

to additional 

details. 

Foray 48B Laboratory bioassays 

equivalent to application rate 

of 99 BIU/ha. 

42 species of lepidoptera native to 

U.S. 

27 species of larvae (61%) evidenced 

significant mortality. 

Foray 48F Field study in which Foray 

48F was applied at a rate of 

40 BIU/acre in May of 1997 

and 1998 to two forests 

susceptible to gypsy moth. 

Nontarget lepidoptera 

monitored in two pretreatment year as well as in 

treatment years. 

Larvae of three lepidopteran species 

were significantly decreased in 

treatment years: Lambdina fervidaria 

[geometrid], Heterocampa 

guttivitta [notodontid], and Achatia 

distincta [noctuid]. For 19 other 

species, larval counts were 

significantly higher in treatment years 

as were the total number of noctuids 

combined and the total number of all 

nontarget lepidopteran species 

combined. 

Rastall et al. 2003 

Dipel 6AF Applied aerially at 59 BIU/ha Two non-target lepidoptera: Incisalia Whaley et al. 

(12,000 IU/mg) (ca. 24 BIU/acre). fotis (Desert Elfin butterfly) and 

Callophrys sheridanii (Sheridan's 

Hairstreak butterfly). Significant 

mortality in larvae that was doserelated. 3,473 cfu/mm2

 lead to nearly 

80% mortality in 7 days. 

1998 

Appendix 3-4 
Appendix 3: Toxicity in Non-target Lepidoptera 

Product Species/Exposure Observations Reference 

B.t.k. (Dipel- Cinnabar moth (Tyria LC50 = 26 BIU/ha (4th instar) James et al. 1993 

HG) potency of jacobaeae) larvae (1st - 5th (95% CI = 9.6-62 BIU/ha) 

4320 IU/mg instar) allowed to feed on 

tansy ragwort leaf pieces 

dipped in concentrations of 0, 

0.24, 0.094, 0.295, 0.943, or 

2.95 mg formulation/mL 

water (corresponding to field 

rates of 0, 2, 8, 25, or 250 

BIU/ha); Cabbage looper 

(Trichoplusia ni) used as 

postive control. 

LC50 = 19 BIU/ha (5th instar) 

(95% CI = 5.9-44 BIU/ha) 

LC50 = 16 BIU/ha (Trichoplusia ni) 

(95% CI = 5.6-30 BIU/ha) 

Treatment had little effect on 1st 

50 through 3rd instar survival) – LC 

values of 427 to 575 BIU/ha. 

See text for discussion. 

B.t.k. (Dipel Diamondback moth exposed Direct dip LC50 >100 mg/mL Idris and Grafius 

2X) to topical application Leaf dip LC50 = 0.014 mg/mL 1993 

Summarized in 

USDA 1995 

B.t.k. HD-1 White-marked tussock moth LC50 = 12 IU/mL diet Frankenhuyszen et 

strain (Orgyia leucostigma) larvae (95% CI = 9-13 IU/mL) al. 1992 

(Thuricide 32 (early 3rd instar) via dietary 

LV) exposure 

B.t.k. HD-1 Eastern hemlock looper LC50 = 162 IU/mL diet Frankenhuyszen et 

strain (Lambdina fiscellaria (95% CI = 129-343 IU/mL) al. 1992 

(Thuricide 32 fiscellaria) larvae (early 3rd 

LV) instar) via dietary exposure 

B.t.k. HD-1 Jack pine budworm LC50 = 145 IU/mL diet Frankenhuyszen et 

strain (Choristoneura pinus) larvae (95% CI = 121-169 IU/mL) al. 1992 

(Thuricide 32 via dietary exposure 

LV) 

B.t.k. HD-1 Western spruce budworm LC50 = 11 IU/mL diet Frankenhuyszen et 

strain (Choristoneura (95% CI = 9-13 IU/mL) al. 1992 

(Thuricide 32 occidentalis)larvae via 

LV) dietary exposure 

B.t.k. HD-1 Spruce budworm LC50 = 63 IU/mL diet Frankenhuyszen et 

strain (Choristoneura fumiferana) (95% CI = 46-82 IU/mL) al. 1992 

(Thuricide 32 larvae (early 4th instar) via 

LV) dietary exposure 

B.t.k. 

(Thuricide 32 

LV) (84 BIU/L 

Spruce budworm 

(Choristoneura fumiferana) 

exposed via diet for 14 days 

LC50 = 160 IU/mL diet 

(95% CI = 139-183 IU/mL) 

Frankenhuyszen 

and Fast 1989 

B.t.k. 

(Thuricide 32 

LV) (84 BIU/L 

Western spruce budworm 

(Choristoneura occidentalis) 

exposed via diet for 14 days 

LC50 = 26 IU/mL diet 

(95% CI = 20-33 IU/mL) 

Frankenhuyszen 

and Fast 1989 

Appendix 3-5 
Appendix 4: Toxicity in Non-Target Terrestrial Insects Other Than Lepidoptera (sorted by insect order). 

Product Species/Exposure Observations Reference 

Coleoptera (Beatles) 

B.t.k. (Dipel 

4L) [] 

Convergent lady beetle 

(Hippodamia convergens 

Guerin) adults only 

exposed to 9.4 or 18.7 

L/ha Dipel 4L and water 

(1:3) 

No significant mortality at 9.4 L/ha [79 

BIU/ha] for up to 7 days. 

At 18.7 L/ha [158 BIU/ha], 13.4% 

mortality attributable to B.t.k. at 7-days 

post-exposure. 

Haverty 1982 

Note on Haverty (1982): Dipel 4L is not used in USDA programs. This is an oil based formulation 

with 32 BIU/gallon (http://www.greenbook.net/docs/LABEL/L16533.PDF) or 8.45 BIU/L. 

The only oil based formulation used in USDA programs is Dipel ES (64 BIU/gallon). 

B.t.k. CGA237218 Ladybird beetles 

(Coccinella 

septempunctata), 5-days, 

dietary, 10 , 10 , 10 5 7 9 

cfu/g food. 

Concentrations characterized as 80 to 

1400X ECC. No observation period 

beyond dosing period. No increase in 

mortality. Mortality in exposed beetles 

consistently less than controls. This is not 

discussed in study. 

Winter et al. 1990 

Thompson 1991a 

NOTE: Winter et al. 1990 and Thompson 1991a have identical data. Appears to be the same study. 

Collembola (snow-fleas, springtails) 

Dipel 8L (oil 

based) as well 

as formulation 

(oil) blank 

Microcosm study using 

Collembola: 1000X EEC 

– i.e., 20,289 I.U./cc OM 

in soil. Observations at 

weeks 2,3,4, and 6 after 

treatment. 

Collembolan populations significantly 

decreased with both B.t.k. formulation 

and oil blank. 

Addison and 

Holmes 1995 

Dipel 8AF 

(aqueous) as 

well as 

unformulated 

B.t.k. 

No effects on Collembolan populations. 

Dermaptera (earwigs) 

B.t.k. (Dipel 

WP) 

Striped earwig (Labidura 

riparia) exposed to 10x 

label application rate 

No mortality observed Workman 1977 as 

summarized in 

USDA 1995 

Appendix 4-1 
Appendix 4: Toxicity in Non-Target Terrestrial Insects Other Than Lepidoptera (sorted by insect order). 

Product Species/Exposure Observations Reference 

Diptera (flies) 

B.t.k. HD-1 

(serovar 3a3b) 

Laboratory bioassay in 

Mexican fruit fly 

(Anastrepha ludens ). 

Significant mortality from both pellet and 

supernatant preparations of B.t.k. in agar. 

Screening study using a variety of 

different B.t. strains to test for efficacy. 

Not directly useful for dose-response 

comparisons. 

Robacker et al. 

1996 

Hemiptera (Bedbugs, aphids, cicadas) 

B.t.k. 

(Bactospeine 

WP) produced 

in the 

Netherlands 

Spined soldier bug 

(Podisus maculiventris) 

(4th instars and 7-day-old 

female adults) exposed to 

B.t.k. formulation 

(16,000 IU mg ) via -1 

ingestion for 48 hours 

No adverse effects and no mortality 

observed at the highest dose tested 

(10,000 mg formulated material/L). 

Mohaghegh et al. 

2000 

Hymenoptera (ants, bees, wasps, sawflies, chalcids, and ichneumons) 

Bees 

B.t.k. , 

Bactec Corp. 

14.5 BIU per 

lb 

Honey bees (Apis 

mellifera): Contact 

toxicity. 0, 7.7 , 15.4, and 

23.2 μg/bee 

corresponding to 0.7, 1.4, 

and 2.1 lb/acre. 

Application rates 

correspond 1.73, 3.45, or 

5.19 lb/ha which also 

corresponds to 25, 50, 

and 75 BIU/ha. 

Mortality at 48 hours: 

BIU/ha Mortality Corrected 

0: 7.17% 

25 19% 12.7% 

50 25% 19.2% 

75 24.9% 19.1% 

See text for additional discussion. W1 

Atkins 1991a 

[Atkins 1991b 

appears to be the 

same study but 

with a different 

MRID number.] 

B.t.k. NOS Honey bees 10-day LC 118 ug/bee (consumed) MRID 435681-01 

summarized but 

not referenced in 

U.S. EPA 1998 

B.t.k. NOS Honey bees No significant effects at 10X field rate 

(NOS). 

MRID 434917-02 

summarized but 

not referenced in 

U.S. EPA 1998 

Ants 

Foray 48F Ants, various species. 

Field study involving 18 

plots in Augusta County, 

VA. 16 BIU/ha (ca. 6.5 

BIU/acre) in May 1997. 

No substantial effects on ant populations: 

abundance, species richness, composition 

and diversity over a 3 year sampling 

period. A decrease of abundance was 

noted in the third year but was attributed 

to over-trapping. 

Wang et al. 2000 

Mantodea (mantids sometimes included with Dictyoptera/roaches) 

Appendix 4-2 
 

Appendix 4: Toxicity in Non-Target Terrestrial Insects Other Than Lepidoptera (sorted by insect order). 

Product Species/Exposure Observations Reference 

B.t.k. Chinese praying mantis No effect on mortality or survival Yousten 1973 

(Commercial (Tenodera aridifolia 

formulation sinensis) exposed via 

containing consumption of cabbage 

18,000 IU/mg) looper larvae that had 

consumed B.t.k. for 15 

hours in 150 μg/mL diet 

Neuroptera (antlions, lacewings, and Dobsonflies) 

Dipel, Common green lacewing Increased mortality in high dose group O'Leary 1990 

specified only (Chrsoperla carnea) but not significantly different from 

as “technical 0.1X, 1X, and 10X field controls. Higher than expected mortality 

powder”. No application rate. Direct in control groups and high variability 

BIU spray and residue among replicates. 

equivalents exposure. 

given. 

B.t.k. (Dipel Common green lacewing Low mortality in larvae (2.1%) and adults Haverty 1982 

4L) (Chrysopa carnea 

Stephens) adults and 

larvae exposed to 9.4 or 

18.7 L/ha Dipel 4L and 

water (1:3) 

(2.0%) at 9.4 L/ha [79 BIU/ha] for up to 7 

days. 

At 18.7 L/ha [158 BIU/ha], mortality 

increased for both adults (5.3%) and 

larvae (6.7). 

B.t.k. Common green lacewing No mortality in control group (0/30). Hoxter et al. 

Biobit (Chrsoperla carnea), 9­ Mortality in dosed groups of 3/30, 4/30, 1990a 

days dietary, 4×10 ,4 and 4/30. [Note: P-value of 0/30 vs 4/30 

2×10 , and 10 cfu/g 6 8 

feed. 

is 0.0562 using Fisher Exact test.] 

B.t.k. CGA- Green lacewing No dose-related increase in mortality. Thompson 1991b 

237218 (Chrsoperla carnea), 5­

days dietary, 10 , 10 , 6 7 

and 10 cfu/g feed. 9-day 8 

post observation period 

Mortality rates in dosed groups ranged 

from 3% (mid-dose) to 33% (low-dose). 

Mortality rates in control groups ranged 

from 23% to 37%. 

Omitted studies by Winter et al. 1991a, Hoxter and Smith 1991 on Delta BT. Cannot identify as B.t.k. 

Omitted Kirkland 1991, Nelson 1991b, and Palmer and Beavers 1993 studies on B.t. Abbott ABG-6305. This 

is B.t.a. 

Appendix 4-3 
Appendix 5: Toxicity of B.t.k. and B.t.k. Formulations to Fish. 

Product Species/Exposure Observations Reference 

Dipel 

Technical 

Material 

Bluegill sunfish (n=30), 32 days, 

static renewal, at 2.87×107

 cfu/L 

nominal (1.45×107

 cfu/L measured) 

No mortality, abnormal gross 

pathology, and no effects on body 

weight or length. 

Christensen 

1990c 

Dipel Rainbow trout (n=30), 32 days, 6/30 treated fish and 1/30 control Christensen 

Technical static renewal, at 2.87×107

 cfu/L fish died, most during the last 14 1990d 

Material, nominal (1.51×107

 cfu/L days of the study [p-value of 0.052 

2.0×1010 measured). using Fisher Exact test]. Mortality 

cfu/g and attributed to 

88,200 The nominal concentration of aggression/competition for food in 

IU/mg. 2.87×107

 cfu/L corresponds to 1.4 

mg/L or 123,480 IU/L. 

cloudy test solution. No abnormal 

gross pathology and no effects on 

body weight or length. [Water pH 

and dissolved oxygen were within 

normal limts.] 

Dipel Sheepshead minnow (n=52), 30 Concentrations characterized as Christensen 

Technical days, static renewal, at aqueous 100X and 1000x expected 1990g 

Material concentration of 2.87×1010 cfu/L 

and dietary concentration of 

 2.87×107

 cfu/L. 

environmental concentrations 

(EEC). 

Four fish died. In one fish, body 

burden of B.t.k. was higher than 

anticipated based on aqueous and 

dietary concentrations – it is 

unclear how this determination was 

made. No inflamation or necrosis. 

B.t.k. Rainbow trout (n=30), 31 days, at Aqueous and dietary Christensen 

Biobit aqueous concentration of 3.67×1010 

cfu/L and dietary concentration of 

1.41×1010 cfu/g. 

concentrations characterized as 

1000x and 40,0000x expected 

environmental concentrations 

(EEC). 

Decreased mean body length and 

weight in exposed fish. No other 

signs of toxicity. 

1990i 

B.t.k. CGA- Rainbow trout (n=30), 32 days, at a Concentrations in water and diet Christensen 

237218 nominal aqueous concentration of 

3.9×1010 cfu/L and dietary 

concentration of 1.52×1010 cfu/g 

characterized as 500X and 

200,000x EEC. 1/30 fish died 

during exposure. No B.t.k. found 

in dead fish. Two fish has gill 

lesions from which B.t.k. could be 

cultured. The concentration in gills 

was less than the concentration in 

water. 

1991c 

B.t.k. CGA- Sheepshead minnow (n=30), 30 Concentrations in water and diet Christensen 

237218 days, at a nominal aqueous 

concentration of 7.8×10 cfu/L and 7 

dietary concentration of 1.56×1010 

cfu/g 

characterized as 50X and 200,000x 

EEC. No mortality. No signs of 

toxicity or infectivity. 

1991e 

Appendix 5-1 
Appendix 5: Toxicity of B.t.k. and B.t.k. Formulations to Fish. 

Product Species/Exposure Observations Reference 

B.t.k. 

(wettable 

powder 

formulation 

manufacture 

d in India) 

Mosquito fish (Gambussia affinis) 

10 fish/group exposed to 0, 200, 

400, 600, 800, or 1000 mg/L for 96 

hours. The formulation contained 

2.5×10 spores/mg. Thus, these 7 

doses correspond to 0, 5×10 , 9 

1×10 , 1.5×10 , 2×10 , and 10 10 10 

2.5×1010 spores/L. 

No mortality observed. No signs 

of sublethal toxicity – i.e., no 

effects on swimming behavior, 

reflexes, general appearance, and 

gill movement. 

Meher et al. 2002 

B.t.k. Rainbow trout, 96 hour exposure LC > 10 mg/L 50 Mayer and 

Ellersieck, 1986 

B.t.k. Bluegill sunfish, 96 hour exposure LC = 95 mg/L 50 Mayer and 

Ellersieck, 1986 

B.t.k. as Koi carp (Cyprinus carpio) Small quantities of bacteria Martin et al. 1997 

unformulated exposed to 1x or 10x ECC via food unrelated to B.t. were recovered 

product in and water in experimental tanks for from various fish organs; bacteria NOTE: This is 

Foray 48B 32 days occurred predominantly in the 

intestine; B.t. found intermittently; 

some of the B.t. strains isolated 

were not the strain applied to the 

tank; sublethal effects observed in 

the treated fish were independent 

of B.t. recovery; sublethal adverse 

effects included significant 

decreases in plasma protein values 

and body weight. 

an abstract and 

the reported 

finding cannot be 

well evaluated. 

A full publication 

has not been 

encountered in 

the literature. 

See Section 

4.1.3.1 for 

discussion. 

B.t.k. Bluegill sunfish, 100x MEEC no evidence of pathogenicity Abbott Labs 

technical (maximum expected environmental 1992 

material concentration) in water and diet for 

30 days Note: This is a 

non-detailed 

summary and 

cannot be well 

evaluted. 

Omitted Bellantoni et al. 1991a,d on Delta BT. Cannot identify strain. 

Appendix 5-2 
Appendix 6: Toxicity of B.t.k. and B.t.k. Formulations to Aquatic Invertebrates (sorted by specified group – 

phylum, order, or subclass – followed by studies on mixed populations). 

Cladocera 

Dipel, NOS Daphnia magna, 21-day static 

renewal, 0, 5, 50, and 100 mg/L. 

Constant aeration. 

Increased BOD of test chambers at 

50 and 100 mg/L. 

21 Day EC50 of 14 mg/L based on 

immobilization. 

Delayed in time to first brood and 

number of young per adult at 5 

mg/L. 

Young 1990 

B.t.k. CGA- Daphnia magna, 21-day static No daphnids survived at two Christensen 

237218 renewal. Measured concentrations highest concentrations. Decreased 1991d 

[Specified as of 0, 4.85×10 , 1.57×10 , 7 8 survival at three lower 

containing 6.24×10 , 1.77×10 , 5.71×10 8 9 9 concentrations: 85% (low), 10% 

1.06×1011 cfu/L. Aeration not specified. (mid), and 30% (high). Decrease 

cfu/g These concentrations are significant only at midequivalent to equivalent to about 0, 0.45, 1.4, concentration group. No 

1.06×108 5.9, 17, and 54 mg/L. difference in reproduction at the 

cfu/mg]. two lower concentrations. 

Substantial decreases in dissolved 

oxygen at two highest 

concentrations [Table 1, p. 28/90]. 

Copepoda 

B.t.k. 

technical 

material 

Amphiascus minutus (copepod). 5, 

50, and 500 mg/kg sediment for 10 

days. (1×10 , 1×10 , and 1×10 5 6 7 

cfu/g sediment) 

No adverse effects at any 

concentration on survival or 

reproduction. Number of offspring 

at 500 mg/kg was significantly 

greater than controls, probably due 

to the utilization of B.t.k. as a food 

source. 

Chandler 1990b; 

Abbott Labs 

1992 

Glass Shrimp (Palaemonetes) 

Dipel 

technical 

material 

Grass shrimp (n=60), 30-day static 

renewal, 100X EEC in water and 

food: 2.87×109

 cfu/L and 2.87×109 

cfu/g food. 

One shrimp died in both exposed 

and control groups. No significant 

differences in body weight or 

length. No apparent adverse 

effects. 

Christensen 

1990h 

B.t.k. CGA237218 Grass shrimp (n=60), 30-day static 

renewal, dietary: 1.58×1010 cfu/g 

food. Concentration characterized 

as 200,000 EEC. 

Mortality of 12/60 in treatment 

groups and 14/60 in control group. 

No effect on survival or growth. 

No signs of infectivity or 

pathogenicity. 

Christensen 

1991f 

Appendix 6-1 
Appendix 6: Toxicity of B.t.k. and B.t.k. Formulations to Aquatic Invertebrates (sorted by specified group – 

phylum, order, or subclass – followed by studies on mixed populations). 

Glass Shrimp (Palaemonetes) (continued) 

B.t. technical Grass shrimp, 100x MEEC no adverse effects Abbott Labs 

material (maximum expected environmental 

concentration) in diet for 30 days 

1992 

[appears to refer 

to Christensen 

1990h] 

Trichoptera 

B.t.k. (Dipel Caddisfly (Hydatophylas argus) Treatment had no apparent effect Kreutzweiser and 

64 AF) larvae exposed to aqueous flowable 

formulation applied to leaf disks 

treated with 20 IU/mL (maximum 

expected environmental 

concentration) or 20,000 IU/mL 

(1000x expected environmental 

concentration) for 2 days under 

flow-through conditions. 

on the palatability of the leaf disks; 

no significant differences among 

treatment levels with regard to leaf 

consumption; no mortality 

observed 

Capell 1996 

Mixed Populations 

B.t.k. Larvae of Simulidae, Clear signs of toxicity observed Eidt 1985 

(Thuricide Chironomidae, Trichoptera, only in Simulium vittatum (black 

32 LV Megaloptera, and nymphs of fly) in which only 6 adults emerged 

containing Ephemeroptera and Plecoptera at at 430 IU/mL; possible signs of 

8.45 BIU/L) continuous exposure to 4.3, 43, or 

430 IU/mL. These concentrations 

correspond to 4300, 43,000, and 

430,000 IU/L. Assuming a density 

of 1 for the formulation, 8.45 

BIU/kg corresponds to 0.00012 

mg/IU. Thus, the concentrations 

correspond to about 0.5 mg/L, 5 

mg/L, and 50 mg/L. 

toxicity were observed in 

Prosimulium fascum/mixtum (black 

fly) in which survival was 

decreased at 43 and 430 IU/mL, 

compared with 4.3 IU/mL 

concentration and with the 

controls. 

Appendix 6-2 
Appendix 6: Toxicity of B.t.k. and B.t.k. Formulations to Aquatic Invertebrates (sorted by specified group – 

phylum, order, or subclass – followed by studies on mixed populations). 

Mixed Populations (continued) 

B.t.k. (Dipel Ephemeroptera (mayflies) (6 taxa); No significant mortality in 11 Kreutzweiser et 

8AF with Plecoptera (stoneflies) (3 taxa); species after 9 days; average al. 1992 

potency of Trichoptera (caddisflies) (4 taxa) mortality of 30% in stoneflies 

16.9 BIU/L) exposed to maximum concentration 

of 600 IU/mL (considered to be 

100x the expected environmental 

concentration in 50 cm of water 

resulting from direct over spray) 

for 24 hours in continuous flowthrough bioassay 

(Taeniopteryx nivalis) after 9 days. 

B.t.k. (Dipel Ephemeroptera (mayflies) (6 taxa); No effect on invertebrate drift; by 1 Kreutzweiser et 

8AF with Plecoptera (stoneflies) (3 taxa); hour after exposure, the % drift al. 1992 

potency of Trichoptera (caddisflies) (4 taxa) was slightly but not significantly 

16.9 BIU/L) exposed to maximum concentration 

of 600 IU/mL for 2.5 hours in 

higher (p>0.05), compared with 

controls, in 5 of 10 species; no 

About outdoor stream channels to effect on survival of drifted insects 

0.00006 measure lethal and drift response. 1 hour after applications. 

mg/BIU. Exposure considered to be 100x 

the expected environmental 

concentration in 50 cm of water 

resulting from direct over spray. 

24-hour LC50 values >600 IU/mL 

(600,000/L or 36 mg/L). No 

mortality in four species of 

Ephemeroptera and three species of 

Trichoptera. 4-30% mortality in 3 

species of Plecoptera, 2 species of 

Ephemeroptera, and one species of 

Trichoptera. 

B.t.k. (Dipel caddisflies, mayflies, stoneflies (12 Only the stonefly (Leuctra tenuis) Kreutzweiser et 

64AF) taxa) exposed to 10x label 

application 

was reduced at 4 days after 

treatment 

al. 1993. 

Summarized in 

USDA 1995 

Appendix 6-3 
Appendix 6: Toxicity of B.t.k. and B.t.k. Formulations to Aquatic Invertebrates (sorted by specified group – 

phylum, order, or subclass – followed by studies on mixed populations). 

Mixed Populations (continued) 

B.t.k. (Dipel Macro invertebrate community in a No significant effects on Kreutzweiser et 

64 AF) section of forest stream (Icewater 

Creek, Ontario) exposed to direct 

application of nominal 

concentration of 200 IU/mL (10x 

expected environmental 

concentration) 

abundance of most benthic 

invertebrates; limited impact of 

B.t.k. application on the stream 

invertebrate community includes a 

slight increase in invertebrate drift 

density at 0.5 hour application and 

only at the site 10 m below the 

application point and the 

significant reduction of the stonefly 

(L. tenuis) (-70%) 4 days after 

application. Although the 

abundance of the stonefly remained 

considerably lower at the treated 

site, compared with the reference 

site, for at least 18 days, the 

difference was not significant. 

al. 1994 

B.t.k. 50-5000 BIU/ha over streams. No effect on benthic stream 

communities or insect emergence. 

Increased drift rates in mayfly 

(Baetis sp) 

Richardson and 

Perrin 1994 

B.t.k. Field trial for control of the spruce 

budworm 

No effects 28 days after treatment 

relative to 14 days prior to 

treatment in populations of a 

number of aquatic invertebrates: 

Amphipoda, Decapoda, 

Hydracarina, Hirudinea, Hydrozoa, 

Nematoda, Oligochaeta, Porifera, 

Pulmonata and Turbellaria. 

Buckner et al. 

1974 

Omitted Bellantoni et al. 1991b,c on Delta BT. Cannot identify strain. Omitted Boeri 1991, B.t.a. 

Appendix 6-4 
Appendix G 

Gypchek 

(Nucleopolyhedrosis Virus) 

Risk Assessment 

Figure G-1. Creosote was used in 1895 to treat gypsy moth egg masses. 

SERA TR 04-43-05-03b 

Control/Eradication Agents for the


Gypsy Moth ­

Human Health and Ecological Risk Assessment
 

for Gypchek – a Nuclear Polyhedrosis Virus (NPV)
 

FINAL REPORT
 

Prepared for: 

USDA, Forest Service
 

Forest Health Protection 

GSA Contract No. GS-10F-0082F
 

USDA Forest Service BPA: WO-01-3187-0150
 

Requisition No.: 43-3187-1-0269
 

Task No. 5
 

Submitted to:
 

Dave Thomas, COTR
 

Forest Health Protection Staff
 

USDA Forest Service
 

Rosslyn Plaza Building C, Room 7129C
 

1601 North Kent Street
 

Arlington, VA 22209
 

Submitted by:
 

Patrick R. Durkin
 

Syracuse Environmental Research Associates, Inc.
 

5100 Highbridge St., 42C
 

Fayetteville, New York 13066-0950
 

Telephone: (315) 637-9560
 

Fax: (315) 637-0445
 

E-Mail: SERA_INC@msn.com
 

Home Page: www.sera-inc.com
 

June 16, 2004 
TABLE OF CONTENTS
 

TABLE OF CONTENTS ...................................................... ii
 

LIST OF TABLES .......................................................... iii
 

LIST OF APPENDICES ..................................................... iii
 

GENERAL ACRONYMS, ABBREVIATIONS, AND SYMBOLS . . . . . . . . . . . . . . . . . . . iv
 

COMMON UNIT CONVERSIONS AND ABBREVIATIONS . . . . . . . . . . . . . . . . . . . . . . . . v
 

CONVERSION OF SCIENTIFIC NOTATION ................................... vi
 

EXECUTIVE SUMMARY ................................................... vii
 

1. INTRODUCTION ....................................................... 1-1
 

2. PROGRAM DESCRIPTION .............................................. 2-1
 

2.1. Overview ...................................................... 2-1
 

2.2. Description and Commercial Formulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-1
 

2.3. Application Methods, Rates, and Mixing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-1
 

2.4. Use Statistics ................................................... 2-2
 

3. HUMAN HEALTH RISK ASSESSMENT ................................... 3-1
 

3.1. HAZARD IDENTIFICATION ........................................... 3-1
 

3.1.1. Overview. .................................................... 3-1
 

3.1.2. Epidemiology Studies a nd other Human Data ......... . ..... 3-1
 

3.1.3. Mechanism of Action (Persistence and Pathogenicity). . . . . . . . . . . . . . . . . . 3-2
 

3.1.4. Acute Oral Toxicity. ............................................ 3-2
 

3.1.5. Subchronic or Chronic Systemic Toxic Effects. . . . . . . . . . . . . . . . . . . . . . . . 3-3
 

3.1.6. Effects on Nervous System. ...................................... 3-3
 

3.1.7. Effects on Immune System. ...................................... 3-4
 

3.1.8. Effects on Endocrine System. ..................................... 3-4
 

3.1.9. Reproductive and Teratogenic Effects. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-5
 

3.1.10. Carcinogenicity and Mutagenicity. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-5
 

3.1.11. Irritation (Effects on the Skin and Eyes) . . . . . . . . . . . . . .. . . . . . . . . . . . . . 3-5
 

3.1.12. Systemic Toxic Effects from Parenteral Exposure . . . . . . . . . . . . . . . . . . . . 3-6
 

3.1.13. Respiratory Effects and Inhalation Exposure. . . . . . . . . . . . . . . . . . . . . . . . . 3-6
 

3.1.14. Impurities and Contaminants. .................................... 3-7
 

3.1.15. Inerts and Adjuvants. .......................................... 3-6
 

3.2. EXPOSURE ASSESSMENT ............................................ 3-8
 

3.2.1. Overview. .................................................... 3-8
 

3.2.2. LdNPV and Gypsy Moth Parts in Gypchek . . . . . . . . . . . . . . . . . . . . . . . . . . 3-8


3.2.3. Supplemental Extreme Exposures. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-9
 

3.3. DOSE-RESPONSE ASSESSMENT ...................................... 3-10
 

3.3.1. Overview. ................................................... 3-10
 

3.3.2. Surrogate RfD for Acute Exposure. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-10
 

3.3.3. Eye Irritation. ................................................ 3-10
 

3.4. RISK CHARACTERIZATION .......................................... 3-12
 

3.4.1. Overview. ................................................... 3-12
 

3.4.2. Pathogenicity and Systemic Toxicity. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-12
 

3.4.3. Extreme Exposure Scenarios. .................................... 3-12
 

ii 
 

 

4. ECOLOGICAL RISK ASSESSMENT ....................................... 4-1
 

4.1. HAZARD IDENTIFICATION ..................................... 4-1
 

4.1.1. Overview. .................................................... 4-1
 

4.1.2. Toxicity to Terrestrial Organisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-1
 

4.1.2.1. Mammals ............................................. 4-1
 

4.1.2.2. Birds ................................................. 4-2
 

4.1.2.3. Terrestrial Invertebrates .................................. 4-2
 

4.1.2.4. Terrestrial Plants (Macrophytes) . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-3
 

4.1.2.5. Terrestrial Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-3
 

4.1.3. Aquatic Organisms.............................................. 4-3
 

4.1.3.1. Fish .................................................. 4-3
 

4.1.3.2. Amphibians ........................................... 4-3
 

4.1.3.3. Aquatic Invertebrates .................................... 4-3
 

4.1.3.4. Aquatic Plants ......................................... 4-4
 

4.2. EXPOSURE ASSESSMENT ............................................ 4-5
 

4.2.1. Overview. .................................................... 4-5
 

4.2.2. LdNPV and Gypsy Moth Parts in Gypchek. . . . . . . . . . . . . . . . . . . . . . . . . . . 4-5


4.2.3. Supplemental Extreme Exposures. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-5
 

4.2.3.1. Contaminated Vegetation ................................ 4-5
 

4.2.3.2. Small Pond ........................................... 4-6
 

4.3. DOSE-RESPONSE ASSESSMENT ....................................... 4-7
 

4.3.1. Overview. .................................................... 4-7
 

4.3.2. Qualitative Assessment .......................................... 4-7
 

4.3.3. Quantitative Assessments. ....................................... 4-7
 

4.4. RISK CHARACTERIZATION ........................................... 4-9
 

4.4.1. Overview. .................................................... 4-9
 

4.4.2. Qualitative Assessment.............. .................. 4-9
 

4.4.3. Quan t itative Assessment............................... 4-9
 

5. REFERENCES ......................................................... 5-1
 

LIST OF TABLES 

Table 2-1: Use of Gypsy Moth NPV from 1995 to 2001 for Suppression, Eradication, and Slow 


the Spread .......................................................... 2-2
 

LIST OF APPENDICES
 

Appendix 1: Gypsy Moth NPV Toxicity in Mammals........................................................Appendix 1-1
 

Appendix 2: Toxicity of Gypsy Moth LdNPV to Birds........................................................Appendix 2-1
 

Appendix 3: Gypsy Moth NPV Toxicity in Nontarget Terrestrial Insects..........................Appendix 3-1
 Appendix 4: Toxicity of NPV to Aquatic Invertebrates........................................................Appendix 4-1
 

iii 
GENERAL ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

a.i. active ingredient 

A.U. activity units 

AEL adverse-effect level 

ACGIH American Conference of Governmental Industrial Hygienists 

bw body weight 

CBI confidential business information 

cm centimeter 

F female 

FS Forest Service 

g gram 

HQ hazard quotient 

kg kilogram 

L liter 

lb pound 

LC50 lethal concentration, 50% kill 

LD50 lethal dose, 50% kill 

LdNPV Lymantria dispar (gypsy moth) nuclear polyhedrosis virus 

LOAEL lowest-observed-adverse-effect level 

m meter 

M male 

mg milligram 

mg/kg/day milligrams of agent per kilogram of body weight per day 

mL milliliter 

mM millimole 

MNPV multinucleocapsid nuclear polyhedrosis virus 

MW molecular weight 

MOS margin of safety 

MSDS material safety data sheet 

NCI National Cancer Institute 

NOAEL no-observed-adverse-effect level 

NOEL no-observed-effect level 

NPV nuclear polyhedrosis virus 

NRC National Research Council 

OB occlusion body 

OpNPV Orgyia pseudotsugata (Douglas-fir tussock moth) nuclear polyhedrosis virus 

OPPTS Office of Pesticide Planning and Toxic Substances 

PIBs polyhedral inclusion bodies 

ppm parts per million 

RED reregistration eligibility decision 

RfD reference dose 

TGAI technical grade active ingredient 

UF uncertainty factor 

U.S. United States 

U.S. EPA U.S. Environmental Protection Agency 

USDA U.S. Department of Agriculture 

> greater than 

$ greater than or equal to 

< less than 

# less than or equal to 

= equal to 

� approximately equal to 

- approximately 

iv 
COMMON UNIT CONVERSIONS AND ABBREVIATIONS
 

To convert ... Into ... Multiply by ... acres 

acres 

hectares (ha) 

square meters (m ) 2 

0.4047 

4,047 

atmospheres millimeters of mercury 760 

centigrade Fahrenheit 1.8C°+32 

centimeters inches 0.3937 

cubic meters (m ) 3 liters (L) 1,000 

Fahrenheit centigrade 0.556F°-17.8 

feet per second (ft/sec) miles/hour (mi/hr) 0.6818 

gallons (gal) liters (L) 3.785 

gallons per acre (gal/acre) liters per hectare (L/ha) 9.34 

grams (g) ounces, (oz) 0.03527 

grams (g) pounds, (oz) 0.002205 

hectares (ha) acres 2.471 

hectares (ha) square meters 10,000 

kilograms (kg) ounces, (oz) 35.274 

kilograms (kg) pounds, (lb) 2.2046 

kilograms per hectare (hg/ha) pounds per acre (lb/acre) 0.892 

kilometers (km) 

liters (L) 

miles (mi) 

cubic centimeters (cm ) 3 

0.6214 

1,000 

liters (L) gallons (gal) 0.2642 

liters (L) ounces, fluid (oz) 33.814 

miles (mi) kilometers (km) 1.609 

miles per hour (mi/hr) cm/sec 44.70 

milligrams (mg) ounces (oz) 0.000035 

meters (m) feet 3.281 

ounces (oz) grams (g) 28.3495 

ounces per acre (oz/acre) grams per hectare (g/ha) 70.1 

ounces per acre (oz/acre) 

ounces fluid 

kilograms per hectare (kg/ha) 

cubic centimeters (cm ) 3 

0.0701 

29.5735 

pounds (lb) grams (g) 453.6 

pounds (lb) kilograms (kg) 0.4536 

pounds per acre (lb/acre) 

pounds per acre (lb/acre) 

pounds per acre (lb/acre) 

kilograms per hectare (kg/ha) 

mg/square meter (mg/m ) 2 

μg/square centimeter (μg/cm ) 2 

1.121 

112.1 

11.21 

pounds per gallon (lb/gal) 

square centimeters (cm ) 2 

square centimeters (cm ) 2 

square meters (m ) 2 

grams per liter (g/L) 

square inches (in ) 2 

square meters (m ) 2 

square centimeters (cm ) 2 

119.8 

0.155 

0.0001 

10,000 

yards meters 0.9144 

Note: All references to pounds and ounces refer to avoirdupois weights unless otherwise specified. 

v 
CONVERSION OF SCIENTIFIC NOTATION
 

Scientific Decimal Verbal 

Notation Equivalent Expression 

1 @ 10-10 0.0000000001 One in ten billion 

1 @ 10-9 0.000000001 One in one billion 

1 @ 10-8 0.00000001 One in one hundred million 

1 @ 10-7 0.0000001 One in ten million 

1 @ 10-6 0.000001 One in one million 

1 @ 10-5 0.00001 One in one hundred thousand 

1 @ 10-4 0.0001 One in ten thousand 

1 @ 10-3 0.001 One in one thousand 

1 @ 10-2 0.01 One in one hundred 

1 @ 10-1 0.1 One in ten 

1 @ 100 1 One 

1 @ 101 10 Ten 

1 @ 102 100 One hundred 

1 @ 103 1,000 One thousand 

1 @ 104 10,000 Ten thousand 

1 @ 105 100,000 One hundred thousand 

1 @ 106 1,000,000 One million 

1 @ 107 10,000,000 Ten million 

1 @ 108 100,000,000 One hundred million 

1 @ 109 1,000,000,000 One billion 

1 @ 1010 10,000,000,000 Ten billion 

vi 
 

EXECUTIVE SUMMARY
 

OVERVIEW 

Gypchek is a preparation of polyhedral inclusion bodies (PIBs) of the Gypsy moth nuclear

polyhedrosis virus (LdNPV). Gypchek is a control agent for the gypsy moth developed and

registered by the USDA Forest Service. This risk assessment is an evaluation of the potential

consequences of using Gypchek and is an update to a previous risk assessment conducted for the

Forest Service as part of the 1995 Final Environmental Impact Statement (FEIS) for the

Cooperative Gypsy Moth Management Program. LdNPV is a naturally occurring baculovirus 

that is clearly pathogenic to gypsy moth larvae. There is no indication, however, that LdNPV is 

pathogenic or otherwise toxic to other species including other Lepidoptera humans. While the 

lack of toxicity displayed by Gypchek somewhat limits the quantitative expression of risk, very

conservative estimates of exposure are below a plausible level of concern by factors of about 750

for humans, 1000 for terrestrial wildlife species, and 30,000 for aquatic species. 

PROGRAM DESCRIPTION 

The active ingredient in Gypchek is the gypsy moth nucleopolyhedrosis virus (NPV), commonly

abbreviated as LdNPV. LdNPV is a naturally occurring baculovirus that is pathogenic to gypsy 

moth (Lymantria dispar) larvae causing a dissolution of tissues and the accumulation of

polyhedral granules in the resultant fluid. The recommended application rate is 0.43 oz

Gychek/acre for suppression and 1.08 oz Gypchek/acre for eradication. The application rate of 

0.43 oz/acre corresponds to about 4×1011 PIB/acre and the application rate of 1.08 oz/acre 

corresponds to about 1×1012 PIB/acre. The production of Gypchek is very expensive and the 

application of this agent is currently limited to areas that are considered environmentally

sensitive. 

HUMAN HEALTH RISK ASSESSMENT 

Hazard Identification – Gypchek does contain substantial amounts ($80% by weight) of gypsy

moth larvae parts, including hairs which are known to cause skin and respiratory irritation in

humans. Based on the available animal data, there is clear evidence that Gypchek can cause eye 

irritation. There is little indication that Gypchek is likely to cause dermal or respiratory irritation. 

The toxicity data on LdNPV are reasonably complete and cover standard acute and chronic

studies for systemic toxicity, standard assays for irritation of the skin and eyes, and basic

pathogenicity studies required of most biological pesticides. While some new studies on eye

irritation have been completed on Gypchek and LdNPV, most of the available studies are

relatively old; they were conducted in the 1970's for the initial registration of Gypchek and most 

of the studies are unpublished. Nonetheless, these unpublished studies have been reviewed and

accepted by U.S. EPA and have been re-reviewed in the preparation of this risk assessment. 

Also as with most pesticides, the toxicity data base on Gypchek is extremely limited for certain

types of biological effects for which the U.S. EPA does not routinely require testing – i.e.,

immunotoxicity, endocrine effects, and neurotoxicity. 

In terms of systemic toxicity or pathogenicity, there is not basis for asserting that Gypchek has

the potential cause adverse effects at any exposure level. There is no indication that LdNPV is 

pathogenic in any mammalian species, even when the animal’s immune function is

compromised. Very high concentrations of Gypchek in the diet of rats – i.e., 500 mg/kg – have

been associated with decreased food consumption and consequent loss of body weight but it is

not clear that the effect was attributable to a toxic response to LdNPV since adverse effects,

including mortality, were noted in the control group. Standard longer term toxicity studies in 

both rodents and dogs revealed no signs of toxicity. 

vii 
Gypchek is typically applied with a carrier, either Carrier 038A or a lignosulfonate-molasses

carrier and another product, Blankophor, may also be included in Gypchek applications. 

Toxicity data on these adjuvants are extremely limited. Carrier 038A is a proprietary surfactant 

formulation. Surfactants are soap-like materials that can have a spectrum of toxic effects, most

of which involve irritation to biological membranes. This appears to be the case for Carrier 

038A. Toxicity data on this material is scant. One available bioassay indicates that Carrier 038A 

is practically nontoxic to rainbow trout. Blankophor serves primarily to protect the LdNPV virus

from sunlight but may also enhance the toxicity of the LdNPV to the gypsy moth. There is 

limited toxicity data on this compound that indicates a very low toxicity. 

Exposure Assessment – Given the failure to identify any hazard associated with Gypchek and

LdNPV, there is little basis for conducting a detailed exposure assessment for Gypchek. 

Gypchek does contain gypsy moth parts and these constituents, as with gypsy moth larvae

themselves, have irritant effects in humans. The use of Gypchek, however, will not add

substantially to exposures to gypsy moth parts in infested areas and will serve to reduce exposure

to gypsy moth larvae by reducing larval populations. 

Based on simply physical processes associated with the application of any pesticide, it is possible

to construct any number of exposure scenarios for Gypchek. The current risk assessment focuses 

on one extreme exposure scenario involving the accidental spray of a home garden. While 

Gypchek is not intentionally applied to such vegetation, the inadvertent spray scenario is

plausible. Based on this accidental exposure scenario, the estimated dose to an individual is

0.034 mg Gypchek/kg bw, with an upper range of 0.66 mg Gypchek/kg bw. 

Dose-Response Assessment – Because no systemic toxic effects can be qualitatively identified

for any plausible routes of exposure (i.e., oral, dermal, or inhalation), the U.S. EPA has not

derived either an acute or chronic RfD for Gypchek. While this is a reasonable approach, the

current risk assessment derives a surrogate acute RfD of 26 mg/kg bw based on an experimental

acute NOAEL of 2,600 mg/kg bw in rats and the application of an uncertainty factor of 100. 

This approach is taken simply to provide a more quantitative basis for comparing the extremely

low risks associated with the application of Gypchek to the risks posed by other agents that may

be used to control the gypsy moth. 

Technical grade Gypchek is an eye irritant. While not quantitatively considered in this risk

assessment, the distinction between the irritant properties of technical grade Gypchek and the

lack of eye irritation with Gypchek formulations as applied in the field is emphasized in order to

highlight areas in which prudent handling practices are likely to be most important. 

Risk Characterization – There is no basis for asserting that any risk is plausible to either

workers or members of the general public in the use of Gypchek to control the gypsy moth. This 

statement follows from the failure to identify any hazard associated with exposures to Gypchek

or LdNPV and is essentially identical to the risk characterization given by the U.S. EPA. 

As discussed in both the exposure and dose-response assessments, the current risk assessment

extends the U.S. EPA risk assessment by proposing a surrogate acute RfD and presenting a very

conservative exposure assessment based on the accidental spray of a home garden. This 

approach is taken simply to facilitate the comparison of risks (or lack of risk) associated with

Gypchek to the risks associated with other agents used to control the gypsy moth. Based on a 

relatively standard dose-response assessment and very conservative exposure assumptions,

plausible exposures to Gypchek are below a level of concern by factors of about 50 to over 750. 

While more typical exposures – i.e., incidental exposure to Gypchek in water or air – are not 

viii 
provided, they will be substantially less than the range of accidental exposure scenarios used to

quantify risk. 

ECOLOGICAL RISK ASSESSMENT 

Hazard Identification – Similar to the hazard identification for the human health risk 

assessment, the hazard identification for nontarget wildlife species fails to identify any adverse

effects of concern – i.e., there is no indication that LdNPV or the Gypchek formulation of

LdNPV has the potential to cause any adverse effects in any nontarget species. The mammalian 

toxicity data base for LdNPV is reasonably complete and indicates that LdNPV is not pathogenic

or otherwise toxic to mammals. One specific study conducted on wildlife mammals that may

consume contaminated gypsy moth larvae indicates no adverse effects in mice, shrews, and 

opossums. Relative to the large number available studies in mammals, few studies are available

in birds but the results of these studies are essentially identical to those in mammals indicating

that exposures to LdNPV at levels that are substantially higher than those likely to occur in the

environment will not be associated with any adverse effects. Based bioassays of LdNPV on the

large number of nontarget insect species and supported by the generally high species specificity

of related baculoviruses, the hazard identification for LdNPV in nontarget insects is essentially

identical to that in birds and mammals. There is no indication that adverse effects will be caused 

in nontarget insects at any level of exposure. Relatively few studies have been conducted in fish

and aquatic invertebrates but these studies are consistent with studies in terrestrial species and

indicate that effects on fish or aquatic invertebrates are unlikely. No data are available on the 

effects of LdNPV on amphibians, aquatic or terrestrial plants or other microorganisms. While 

this lack of information does, by definition, add uncertainty to this risk assessment, there is no

basis for asserting that effects on these or other organisms are plausible. 

Exposure Assessment – In ground or aerial applications, it is likely that a large number of

species could be exposed to Gypchek/LdNPV. The need for any formal risk assessment is

questionable, however, because neither Gypchek nor LdNPV appear to cause systemic adverse

effects. Nonetheless, in an attempt to provide some bases for comparing the potential risks of

Gypchek to other agents used to control the gypsy moth, two extreme exposure assessments are

developed: one for a terrestrial herbivore consuming contaminated vegetation and the other for

aquatic organisms in a small pond directly sprayed with Gypchek at the highest application rate. 

For the terrestrial herbivore, the dose estimates range from 1.1 mg Gypchek /kg bw to 3.2 mg

Gypchek /kg bw. For aquatic organisms, concentrations are expressed in units of PIB/liter

because this unit is used in the corresponding toxicity studies. For a small pond directly sprayed 5 with Gypchek at the highest application rate, the estimated initial concentration is 2.5×10 PIB/L. 

A large number of other less extreme exposure assessments could be developed but these would

not alter the assessment of risk since these extreme exposure assessments are substantially below

any level of concern. 

Dose-Response Assessment – Because no hazards can be identified for any species, a

quantitative dose-response assessment is not required and no such assessments have been

proposed by U.S. EPA and no quantitative dose-response assessments were used in the previous

gypsy moth risk assessment for Gypchek. In order to provide a clear comparison of the risks of

using Gypchek relative to other agents, dose-response assessments are proposed in the current

risk assessment for both terrestrial mammals and aquatic species. For terrestrial mammals, the 

NOAEL of 2,600 mg/kg bw is used. This is the same NOAEL that served as the basis for the 

surrogate acute RfD in the human health risk assessment. For aquatic species, only NOEC 9 values are available and the highest NOEC of 8×10 PIB/L is used to characterize risk.

ix 
Risk Characterization – There is no basis for asserting that the use of Gypchek to control or

eradicate gypsy moth populations is likely to cause any adverse effects in any species other than

the gypsy moth. While no pesticide is tested in all species under all exposure conditions, the data

base on LdNPV and related viruses is reasonably complete and LdNPV has been tested

adequately for pathogenicity in a relatively large number of species, particularly terrestrial

invertebrates. LdNPV appears to be pathogenic and toxic to the gypsy moth and only to the 

gypsy moth. 

For Gypchek, quantitative expressions of risk are in some respects more difficult because clear

NOEC and LOEC values cannot be defined – i.e., if an agent is not shown to cause an effect, the

threshold exposure level is not a meaningful concept. Nonetheless, general but very conservative

exposure assessments demonstrate that plausible upper ranges of exposures are clearly below any

level of concern by a factor of 1000 for terrestrial species and 30,000 for aquatic species. 

x 
1. INTRODUCTION
 

This risk assessment is an evaluation of the potential consequences of using Gypchek and is an

update to a previous risk assessment conducted for the Forest Service as part of the 1995 Final

Environmental Impact Statement (FEIS) for the Cooperative Gypsy Moth Management Program

(Durkin et al. 1994; USDA 1995). The USDA Forest Service uses Gypchek in the control of the 

Gypsy moth (Lymantria dispar). Gypchek is a preparation of polyhedral inclusion bodies (PIBs)

of the Gypsy moth nuclear polyhedrosis virus (NPV). Based on the recent re-registration

eligibility decision (RED, U.S. EPA 1996) and a few more recent studies not cited in the RED,

the present document provides risk assessments for human health effects and ecological effects

of LdNPV to support an assessment of the environmental consequences of using Gypchek in

Forest Service programs. In the re-registration process, the U.S. EPA (1996) combined data from

the Gypsy Moth NPV (LdNPV) and a related virus, Tussock Moth NPV (OpNPV). 

In addition to this introduction, this document includes a program description, a risk assessment

for human health effects, and a risk assessment for ecological effects or effects on non-target

wildlife species. Each of the two risk assessment chapters has four major sections, including an

identification of the hazards associated with LdNPV, an assessment of potential exposure to the

virus, an assessment of the dose-response relationships, and a characterization of the risks

associated with plausible levels of exposure. These are the basic steps recommended by the

National Research Council of the National Academy of Sciences (NRC 1983) for conducting and

organizing risk assessments. 

Nonetheless, this risk assessment of LdNPV is qualitatively different in some ways from risk

assessments of chemical agents. Because NPVs are biological organisms rather than chemicals,

many standard physical and chemical properties used to characterize chemical compounds and

estimate certain exposure parameters (e.g., SERA 2001) simply do not apply to LdNPV or other

NPVs. More significant is the fact that most NPVs including LdNPV are highly host specific. 

LdNPV is pathogenic to the gypsy moth. In this species, LdNPV produces a well-characterized

effect for which the most meaningful exposure metameter is clearly the number of active

polyhedral inclusion bodies (PIBs). For other species, including humans, PIBs are a less

meaningful measure of exposure because LdNPV does not appear to affect non-target species. 

Instead, the available information suggests that most adverse effects in non-target species

associated with exposure to Gypchek are likely to be associated with insect parts in the

commercial formulation. 

The human health and ecological risk assessments presented in this document are not, and are not

intended to be, comprehensive summaries of all of the available information (e.g., efficacy

studies) but are focused on the information that most clearly impacts an assessment of risk. Most 

of the mammalian toxicology studies and some ecotoxicology and environmental fate studies are

unpublished reports submitted to the U.S. EPA as part of the registration or re-registration of

LpNPV. Full text copies of studies submitted to the U.S. EPA were kindly provided by U.S.

EPA/OPP (n=81). These studies were reviewed and are discussed in this document. 

This is a technical support document and it addresses some specialized technical areas. 

Nevertheless, an effort has been made to ensure that the document can be understood by

individuals who do not have specialized training in the chemical and biological sciences. Certain 

technical concepts, methods, and terms common to most risk assessments are described in a

separate document (SERA 2001). In addition, technical terms commonly used in this document

and other risk assessments are defined in a glossary (SERA 2003) and more specialized terms are

defined in the text as necessary. 

1-1
 
2. PROGRAM DESCRIPTION
 

2.1. Overview 

The active ingredient in Gypchek is the gypsy moth nucleopolyhedrosis virus (NPV), commonly

abbreviated as LdNPV. LdNPV is a naturally occurring baculovirus that is pathogenic to gypsy

moth larvae causing a dissolution of tissues and the accumulation of polyhedral granules in the

resultant fluid. The recommended application rate is 0.43 oz Gychek/acre for suppression and

1.08 oz Gypchek/acre for eradication. The application rate of 0.43 oz/acre corresponds to about 

11 12 4×10 PIB/acre and the application rate of 1.08 oz/acre corresponds to about 1×10 PIB/acre. 

The production of Gypchek is very expensive and the application of this agent is currently 

limited to areas that are considered environmentally sensitive. 

2.2. Description and Commercial Formulation

Gypsy moth nucleopolyhedrosis virus (LdNPV) is a naturally occurring baculovirus that is

usually important in bringing about the collapse of gypsy moth populations (Cook et al. 1997;

Podgwaite 1979;Webb et al. 1999a,b). Gypchek is a powdered formulation of LdNPV developed

and registered by USDA for control of the gypsy moth (Podgwaite 1999). 

The active ingredient in Gypchek is about 12% (by weight) polyhedral inclusion bodies (PIB’s) 

of LdNPV (USDA/FS 2003a). Some earlier preparations of Gypchek were about 20% LdNPV

by weight (USDA/FS 19??c, MRID 00066097). [Note: Designations such as 19??c are used by

U.S. EPA to identify submissions whose date is unclear. This designation is also used in this risk

assessment for consistency with U.S. EPA.] The powder is produced by culturing and processing

gypsy moth larvae infected with LdNPV (Lewis 1971; USDA/FS 1975). The average yield of 9 PIB’s in mass production is about 2×10 PIB/larva (Lewis 1971) and the average weight of each

PIB is about 3.66×10-12 grams (Adamson 1991). The active material is sometimes referred to as 

occulsion bodies (OBs) because the virus particles occluded, containing variable numbers of

nucleocapsids (genetic material) within one protein envelope. The rest of the Gypchek

formulation consists of gypsy moth parts (USDA/FS 19??a,b,c; USDA/FS 2003a). A similar 

product, Disparvirus, was developed in Canada (Nealis and Erb 1993). Gypchek causes

polyhedrosis, a viral disease of insect larva, which is characterized by dissolution of tissues and

the accumulation of polyhedral granules in the resultant fluid. 

2.3. Application Methods, Rates, and Mixing

Gypchek is usually applied against first or second instars of the gypsy moth. Application rates or

other measures of exposure to Gypchek can be expressed in various units, the most common of

which are weight of formulation, weight of the virus PIBs, or counts of the polyhedral inclusion

bodies. Based on the most recent product label (USDA/FS 2003a), the recommended application

rate for aerial spray is 0.43 oz/acre for suppression and 1.08 oz/acre for eradication. For ground 

applications, a rate of 0.54 oz/acre is recommended. The current product label does not specify

an application rate in PIBs per acre but does provide a reference value of 929.3 billion 11 11 [9.293×10 ] PIB per ounce. The application rate of 0.43 oz/acre corresponds to about 4×10 

PIB/acre and the application rate of 1.08 oz/acre corresponds to about 1×1012 PIB/acre. This is 

very similar to the application rates considered in the 1995 risk assessment. In all applications, 

the Gypchek formulation is applied at particles sizes of 100–150 : (Podgwaite 1994). 

Gypchek is applied in a carrier. A number of different carriers and adjuvants have be evaluated 

for Gypchek including Carrier 244 from Novo Nordisk (Cunningham et al. 1996) and

Blankophor BBH, supplied by Burlington Chemical Company (Thorpe et a. 1999; Webb et al.

1998, 1999a). Carrier 038 or a lignosulfonate-molasses formulation has been used with Gypchek

(Podgwaite 1999). Both Carrier 038 and a lignosulfonate-molasses formulation are listed as

agents that can be used with Gypchek on the current product label (USDA/FS 2003a). Carrier 

2-1
 
 

038 is produced by Novo Nordisk (Webb et al. 1999b). A presumably related carrier, Carrier

038-A, is currently listed at the USDA Forest Service web site

(http://www.dnr.state.wi.us/org/land/forestry/ fh/GM/). This carrier is produced by OMNOVA

Solutions (1999) and is identified only as a proprietary mixture. No additional information on 

the constituents of Carrier 038 or Carrier 038-A have been located in the open literature or the

U.S. EPA/OPP FIFRA files. 

Applications of Gypchek vary depending on the carrier used. For Carrier 038, 0.95 gallons of the

carrier are mixed with a small amount of water (0.05 gal.) and 6.4 grams of Gypchek. For the 

lignosulfonate-molasses carrier, 1.7 gallons of water are mixed with 1 lb of Lignosite AN, 0.26 lb

of feed-grade molasses,0.04 gallons of Bond, and 15.9 grams of Gypchek (USDA/FS 2003a). 

2.4. Use Statistics 

Gypchek was applied to only 53,034acres – about 6600 acres per year between 1995 and 2003

(Table 2-1). As indicated in Table 2-1, this figure does not include the number of acres that were

treated twice. Including these repeated applications, a total of 54,034 acres were treated between

1995 and 2003 (Onken 2004). 

As noted by Podgwaite (1999), the application of Gypchek is very expensive and is limited to

areas that are considered environmentally sensitive. Gypchek is highly specific to the gypsy

moth and there is no indication that LdNPV will effect any nontarget species (Sections 3.1 and

4.1). 

TABLE 2-1: Use of Gypchek from 1995 to 2001 for Suppression, Eradication, and Slow the Spread* 

1995 1996 1997 1998 1999 2000 2001 2002 2003 Total 

(acres) 

Suppression 2,127 791 4,367 3,956 2,306 5,882 2,280 4,794 10,015 36,518 

Eradication 0 0 0 2,122 5,254 0 0 0 0 7,376 

Slow the 262 0 374 0 500 0 0 0 8,004 9,140 

Spread 

Total 2,389 791 4,741 6,078 8,060 5,882 2,280 4,794 18,019 53,034 

*Source: GMDigest, Morgantown, WV (http://fhpr8.srs.fs.fed.us/wv/gmdigest/gmdigest.html). Does not include 

areas that were treated twice. 

2-2
 
3. Human Health Risk Assessment 

3.1. HAZARD IDENTIFICATION 

3.1.1. Overview 

LdNPV is a naturally occurring baculovirus that is clearly pathogenic to gypsy moth larvae. 

There is no indication, however, that LdNPV is pathogenic to other species, including humans or

other mammals. Gypchek, the commercial formulation of LdNPV, is produced by culturing

infected gypsy moth larvae and Gypchek does contain substantial amounts (>80% by weight) of

gypsy moth larvae parts, including hairs which are known to cause skin and respiratory irritation

in humans. Based on the available animal data, there is clear evidence that Gypchek can cause

eye irritation. There is little indication that Gypchek is likely to cause dermal or respiratory 

irritation. 

Information on the toxicity data of LdNPV is reasonably complete and covers standard acute and

chronic studies for systemic toxicity, standard assays for irritation of the skin and eyes, basic

pathogenicity studies required of most biological pesticides. While some new studies on eye

irritation have been completed on Gypchek and LdNPV, most of these studies are relatively old,

being conducted in the 1970's for the initial registration of Gypchek and most of the studies are

unpublished. Nonetheless, these unpublished studies have been reviewed and accepted by U.S.

EPA and have been re-reviewed in the preparation of this risk assessment. Also as with most 

pesticides, the toxicity data base on Gypchek is extremely limited for certain types of biological

effects for which the U.S. EPA does not routinely require testing – i.e., immunotoxicity,

endocrine effects, and neurotoxicity. 

There is no indication that LdNPV is pathogenic in any mammalian species, even when the

animal’s immune function is compromised. Very high concentrations of Gypchek in the diet of

rats – i.e., 500 mg/kg – have been associated with decreased food consumption and consequent

loss of body weight but it is not clear that the effect was attributable to a toxic response to

LdNPV since adverse effects, including mortality, were noted in the control group. Standard 

longer term toxicity studies in both rodents and dogs have not identified adverse effects at any

dose level tested. 

Gypchek is typically applied with a carrier (Section 2). Toxicity data on the adjuvants are 

extremely limited. Carrier 038A is a proprietary surfactant formulation. Surfactants are soaplike materials that can have a spectrum of toxic effects, most of which involve irritation to

biological membranes. This appears to be the case for Carrier 038A as well as many household 

soaps. Toxicity data on Carrier 038A is scant. One available bioassay indicates that the material 

is practically nontoxic to rainbow trout. Blankophor serves primarily to protect the LdNPV virus

from sunlight but may also enhance the toxicity of the LdNPV to the gypsy moth. There is some 

limited toxicity data on this compound that indicates a very low toxicity. 

3.1.2. Epidemiology Studies and Other Human Data

Epidemiology studies regarding health effects in humans after exposure to LdNPV were not

located in the available literature. Gypchek contains substantial amounts of gypsy moth larvae

parts and exposure to gypsy moth larvae has been associated with dermal and respiratory effects

in humans (Durkin et al. 1995). Based on the available animal data, it is plausible that exposure

to Gypchek could be associated with ocular irritation in humans (Section 3.1.11). The 

plausibility of respiratory irritation (Section 3.1.13) or dermal irritation (Section 3.1.11) is less

clear. 

3-1
 
 

3.1.3. Mechanism of Action (Persistence and Pathogenicity)

As discussed in the following subsections, LdNPV has been subject to a large number of

relatively standard toxicity studies and there is no indication that LdNPV exposures are

pathogenic in mammals. In addition, as detailed further in Section 4.1, LdNPV appears to be

highly specific to the gypsy moth and does not appear to be pathogenic to other species. In

addition, a series of experiments were conducted to determine if NPV could infect or otherwise

affect mice immunosuppressed with cyclophosphamide, thymectomy, or anti-lymphocyte serum

and guinea pigs immunosuppressed with cortisone or cobra venom factor. No lesions, 

histopathological changes, or signs of infection associated with treatment were noted (Shope

1976; Shope and others 1977). Circulating antibodies to the insect viral subfractions have not

been observed in laboratory workers (Mazzone et al. 1976; Tignor et al. 1976). Thus, there is no 

basis for asserting that LdNPV poses a risk of pathogenicity in humans. 

Persistence in lung tissue has been examined in a study submitted to the U.S. EPA by the U.S.

Forest Service. Several summaries of this study are available but are poorly documented

(USDA/FS 19??d, MRID 00066105; USDA/FS 19??g, MRID 00060701; USDA/FS 1975?,

MRID 00090598). Only one of these studies, MRID 00066105, is explicitly cited in the U.S.

EPA (1996) although a later submission, MRID 00090598, gives a somewhat fuller description

of the study. As indicated in Appendix 1, rats were exposed to LdNPV via inhalation for 1 hour 8 8 at a concentration of 6.12 ± 2.087 mg/L (= 4.04x10 ± 1.38x10 PIBs/L) and sacrificed 1, 7, or 14

days after exposure. Recovery of LdNPV from the lung, relative to amounts recovered

immediately after exposure, were about 96% at day1, 68% at day 7, and 18% at day 14. 

Assuming first-order clearance, this corresponds to a clearance rate of 0.13 days-1 or a halftime of 

about 5 days. 

3.1.4. Acute Oral Toxicity

The U.S. EPA requires standard acute oral toxicity studies for the registration of most pesticides,

including Gypchek. For microbial pesticides, additional requirements include assays for 

pathogenicity. The standard assays involving LdNPV or Gypchek are summarized in 

Appendix 1. A large number of studies have been submitted to U.S. EPA. As detailed in 

Appendix 1, many of these are duplicate submissions or submissions of preliminary results. 

Some of these refer to the test agent as P. dispar NPV, referring to Porthetria dispar, a former 

designation for the gypsy moth. Thus, P. dispar NPV is identical to LdNPV. 

A single dose of LdNPV at 400 mg was not associated with any adverse effects in male or female

rats over a 30-day observation period (Terrell and Parke 1976a,b). At a somewhat higher dose,

500 mg per rat, a transient (2 week) but statistically significant decrease was noted in body

weights over a 35-day observation period (Terrell et al. 1976c). This effect was associated with 

decreased food consumption. As noted in Appendix 1, mortality was noted in both control

(8/20) and treated (3/20) animals. Thus, it appears that the health of the animals may have been

compromised by factors other than treatment with LdNPV. As noted above, no effects were seen 

in immunosuppressed mice at a dose of 0.02 g/mouse over a 21-day observation period (Shope et

al. 1975, 1977). Hart and coworkers (Hart 1976; Hart and Thornett 1975a,c) also observed no

signs of toxicity or pathogenicity in groups of 20 to 30 rats after single gavage doses of up to 1

mL of a 4×1010 solution of LdNVP per rat. The U.S. EPA (1986) indicates an additional acute 

oral/pathogenicity study (MRID 41738701) is available for LdNPV. This study, however, 

involved exposures to OpNPV and not LdNPV.] 

3-2
 
3.1.5. Subchronic or Chronic Systemic Toxic Effects

No recent studies have been conducted on the subchronic or chronic toxicity of Gypchek. As 

detailed in Appendix 1, two standard longer term toxicity studies are available on Gypchek: a 90­

day subchronic feeding study in dogs (Hart 1975a) and a two-year chronic feeding study in rats

(Hart 1975b). Both of these studies were submitted for the initial registration of Gypchek and

have been reviewed by U.S. EPA (1996) and accepted as supplemental in the reregistration of

both Gypchek and TM-Biocontrol. 

In the subchronic study, purebred beagles were given LdNPV in the diet at concentrations that 7 8 9 resulted in average daily doses of 0, 10 , 10 , or 10 OB of LdNPV/dog for 90 days. These doses 

correspond to Gypchek doses of 0, 1.8, 18, or 180 mg formulation/dog. The terminal body

weights reported in the study were 9.5 kg for the low dose group, 11.1 kg for the middle dose

group, and 10.3 kg for the high dose group. These doses expressed in mg Gypchek/kg bw equal

0.2 mg/kg for the low dose group, 1.6 mg/kg for the middle dose group, and 17 mg/kg for the

high dose group. Each dog was observed at least once daily for gross effects. Standard 

hematology, clinical biochemistry, and urinalysis were conducted on each animal at or before the

start of exposure and at 2, 4, and 6 months after the start of exposure. After sacrifice, standard 

examinations were conducted for signs of gross pathology or histopathology. No treatment 

related effects were observed (Hart 1975a). 

In the chronic study, Dublin (Sprague-Dawley derived) rats were given LdNPV in chow at levels 7 8 that resulted in daily doses of 10 or 10 OB/rat for 2 years. This exposure corresponded to 

Gypchek daily doses of 1.8 or 18 mg/rat. The average terminal body weights (both sexes 

combined) was approximately 400 g. Thus, the dose rate was 4.5 or 45 mg Gypchek/kg bw. 

Each of the treated and control groups consisted of 50 males and 50 females. Observations 

included body weight, food consumption, gross signs of toxicity, and pathology. No increased 

mortality was observed and no pathological changes were attributed to treatment (Hart 1975b). 

As discussed in Section 4.1.2.1 and also summarized in Appendix 1, mammalian feeding studies

have been conducted on various mammalian predators of the gypsy moth (Lautenschlager et al.

1977) but the exposure data from this study is not sufficiently detailed to permit a clear

assessment of the actual doses that were used. Nonetheless, this study is consistent with the

above standard studies in that no signs of toxicity were observed in any species. 

3.1.6. Effects on Nervous System 

A neurotoxicant is chemical that disrupts the function of nerves, either by interacting with nerves

directly or by interacting with supporting cells in the nervous system (Durkin and Diamond

2002). This definition of neurotoxicant is critical because it distinguishes agents that act directly 

on the nervous system (direct neurotoxicants) from those agents that might produce neurologic

effects that are secondary to other forms of toxicity (indirect neurotoxicants). Virtually any agent

(microbial or chemical) will cause signs of neurotoxicity in severely poisoned animals and thus

can be classified as an indirect neurotoxicant. 

Studies designed specifically to detect impairments in motor, sensory, or cognitive functions in

mammals exposed to Gypchek or purified preparations of LdNPV have not been encountered in

the open literature or in submissions to U.S. EPA. The U.S. EPA/OPTS (2003) has standard 

protocols for a number of types of neurotoxicity studies including a neurotoxicity screening

battery (Guideline 870.6200), acute and 28-day delayed neurotoxicity of organophosphorus

substances (Guideline 870.6100). Neither of these types of studies have been conducted on 

Gypchek. Further, the RED for LdNPV (U.S. EPA 1996) does not specifically discuss the 

potential for neurologic effects. 

3-3
 
As summarized in Appendix 1, one early study on Gypchek, Terrell et al. (1976c), reports

symptoms that are consistent either with either direct or indirect neurotoxicity – i.e., piloerection

and decreased locomotor activity. These effects, however, occurred in both exposed and control 

animals. Based on both the acute and longer-term studies on Gypchek, there is no indication that

exposure to LdNPV will be associated with either direct or indirect signs of neurotoxicity. 

3.1.7. Effects on Immune System

With LdNPV or any other biological agent that may be pathogenic, the response of or

pathological activity in immunocompromised animals – i.e., animals with impaired immune

function – is a concern. In addition, some chemical or biological agents may act as

immunotoxicants – i.e., chemical agents that disrupt the function of the immune system. Two 

general types of immunotoxic effects, suppression and enhancement, may be seen and both of

these are generally regarded as adverse. Agents that impair immune responses (immune 

suppression) enhance susceptibility to infectious diseases or cancer. Enhancement or 

hyperreactivity can give rise to allergy or hypersensitivity, in which the immune system of

genetically predisposed individuals inappropriately responds to chemical or biological agents

(e.g., plant pollen, cat dander, flour gluten) that pose no threat to other individuals or

autoimmunity, in which the immune system produces antibodies to self components leading to

destruction of the organ or tissue involved (Durkin and Diamond 2002). 

As summarized in Appendix 1, Shope et al. (1975) assayed the effects of LdNPV on normal and

immunosuppressed animals by several routes of exposure: oral intubation, dermal application,

ocular or intranasal installation, and footpad inoculation. The dermal studies were conducted on 

guinea pigs and other studies were conducted in mice. Differences in responses were observed

between immunocompetent animals and immunosuppressed animals but these differences are

attributable to the immunosuppressive agents rather than to any increased toxicity of LdNPV. 

Specifically, immunocompetent guinea pigs exhibited a greater skin irritant response to LdNPV

than did immunosuppressed guinea pigs, indicating a general allergic reaction to the LdNPV in

which a greater response in immunocompetent individuals would be expected. In mice, 

immunocompetent individuals evidenced a greater antibody titre than did immunosuppressed

individuals after both oral exposure and intranasal installation (Shope et al. 1975). Again, this

difference in response between immunocompetent and immunosuppressed mice would be

expected after exposure to any antigenic material. In mice treated by footpad inoculation, 

secondary bacterial infections were noted. The study does not specify whether or not there were

any differences in the incidence of bacterial infections between immunocompetent and

immunosuppressed mice. Based on this study, the lack of marked dermal irritation (Section

3.1.11) and the low acute and chronic systemic toxicity of LdNPV (Sections 3.1.4 and 3.1.5), the

U.S. EPA (1996) elected not to require additional testing on the immunologic effects of LdNPV. 

3.1.8. Effects on Endocrine System

In terms of functional effects that have important public health implications, effects on endocrine

function would be expressed as diminished or abnormal reproductive performance. As discussed 

in the following section (Section 3.1.9), however, very limited data are available on the

reproductive effects of LdNPV. The potential for direct endocrine effects are typically assessed

by various mechanistic assays (Durkin and Diamond 2002). LdNPV or other related NPV have 

not been tested for activity as an agonists or antagonists of the major hormone systems (e.g.,

estrogen, androgen, thyroid hormone). In the re-registration review for LdNPV, the U.S. EPA

(1996) does not discuss the potential for effects on endocrine function. Thus, in the absence of 

direct experimental data on endocrine function or related toxicity studies that might be useful for

assessing effects on endocrine function, no definitive hazard identification is possible. This does 

not imply that a risk is plausible. To the contrary, most endocrine active agents are synthetic 

3-4
 
organic chemicals that mimic or otherwise interfere with the function of naturally occurring

hormones. There is no basis for asserting that LdNPV is likely to have such an effect. 

3.1.9. Reproductive and Teratogenic Effects

A number of standard tests for reproductive effects – i.e., effects on fertility – as well as tests for

the potential to cause birth defects – i.e., teratogenicity – are available and are often required for

pesticides. Examples of protocols for such tests are available from the U.S. EPA’s web site:

http://www.epa.gov/OPPTS_Harmonized/. These tests have not been required for LdNPV or 

OpNPV by the U.S. EPA (1996). 

The only available information on the reproductive effects of LdNPV is the early study by

Lautenschlager et al. (1977). This study reports no effects on reproduction in mice after they

were fed diets containing LdNPV over a 20 day period. In the treated group, consisting of 8

males and 9 females, 5 litters with a total of 20 young were produced. In the control group,

consisting of 10 males and 10 females, only 1 litter with 4 young was produced. While all 

exposures were dietary, the exposure regime was complex consisting of gypsy moth larvae

infected with LdNPV, followed by a purified formulation of LdNPV, that was in turn followed

by a diet containing a spray preparation of LdNPV. In any event, this study does provide a basis

for asserting that relatively prolonged exposures to LdNPV did not cause adverse reproductive

effects in mice. 

3.1.10. Carcinogenicity and Mutagenicity

The two-year chronic feeding study in rats (Hart 1975b), which is discussed in Section 3.1.5 and

summarized further in Appendix 1, is a standard in vivo assay for both chronic toxicity and 

carcinogenicity. As noted in Appendix 1, no increase in the incidence of tumors was noted in 

this study. This is the only long term study that is appropriate for assessing the potential

carcinogenic effects of LdNPV. 

3.1.11. Irritation (Effects on the Skin and Eyes)

LdNPV does not appear to be a marked skin irritant. As summarized in Appendix 1, relatively

standard assays for dermal irritation noted no dermal irritation (Hart and Thornett 1975b,d,e;

Becker and Parke 1976d) and, based on these studies, the U.S. EPA (1996) has classified LdNPV 

as not a dermal irritant (Category IV) (U.S. EPA 1996, p. 13). 

The U.S. EPA (1996) has classified LdNPV as a Category I Eye Irritant – i.e., irritation with

corneal involvement not cleared by day 14 after treatment. While the U.S. EPA (1996) cites

many of the studies included in Appendix 1 in support of this determination, some studies (e.g.,

Hart and Thronett 1975f; Becker and Parke 1976c) noted little or only slight irritation. The most 

severe irritation and the only study consistent with the Category I designation is the study by

Imlay and Terrell (1978) in which rabbits did evidence irritation with corneal opacity and

conjunctival irritation that persisted through day 14 after treatment. This effect was seen, 

however, only in animals whose eyes were not washed at all after the instillation of a LdNPV

formulation – i.e., Group 4 from the Imlay and Terrell 1978 study as summarized in Appendix 1. 

In other groups of rabbits whose eyes were flushed after treatment, signs of eye irritation were

evident but much less severe. 

Subsequent to the RED (U.S. EPA 1996), the Forest Service funded two studies on the ocular

irritation of Gypchek, the commercial formulation of LdNPV. One study used the commercial

formulation (Kuhn 1997a) and the other study used an aqueous solution at twice the anticipated

field concentration (Kuhn 1997b). Both studies identify the test material as a 3.65×1010 PIBs/g 

LdNPV preparation [Lot GR-14A], a wettable powder. The study by Kuhn (1997a) characterizes 

the applied material as a “Gypchek TGAI”, presumably referring to technical grade active 

3-5
 
ingredient (i.e., the mixture of virus, insect parts and other ingredients). The study by Kuhn 

(1997b) characterizes the applied material as a “Gypchek Solution 2X”, presumably indicating

that the test solution was diluted to a concentration that is twice that used in field applications. 

Kuhn (1997b) does not specify the actual concentration of the test solution. In a letter of 

clarification to the U.S. EPA, Kuhn (1997c) indicates that the 2X solution was a concentration of 

2.92 mg technical product/mL. This dose is characterized as twice the field concentration based 

on a letter from Podgwaite (1996) indicating that the batch of Gypchek tested by Kuhn (1997a,b)

would be diluted to 2×1011 PIBs/gallon and that this would correspond to 1.45 mg/mL. 

In both studies, New Zealand White rabbits were dosed with 0.1 mL by volume of the test

substance which was placed into the right eye of each of six males and six females. In the TGAI 

study (Kuhn 1997a), the eyes were washed for 1 minute beginning 30 seconds after treatment in

three each of the males and females. None of the eyes were washed in the 2X study (Kuhn 

1997b). The rabbits were examined at 1, 24, 48, and 72 hours as well as 4, 7, 10, 14, and 17 days

after treatment. 

In the TGAI study (Kuhn 1997a), the maximum average irritation score was 5.3 after 1 hour

(minimally irritating) in the washed eyes and the maximum irritation score was 37.3 (moderately

irritating) in the unwashed eyes. All effects cleared by day 17 after exposure. Based on U.S. 

EPA’s classification scheme for ocular irritation, Kuhn (1997a) characterized the LdNPV

preparation as Category II for non-washed eyes and Category IV for washed eyes. In the 2X 

study, no indication of eye irritation was noted and the test substance was assigned to Category

IV, no or minimal effects. 

Thus, while it is clear that LdNPV does have the potential to cause severe eye irritation, as

demonstrated in the study by Imlay and Terrell (1978), it is less clear that such effects will be

evident in the normal use of Gypchek with prudent use of protective measures to limit exposure

to the eyes and to clean contaminated eyes in the event of unintended ocular exposure. This is 

discussed further in the risk characterization (Section 3.4). 

3.1.12. Systemic Toxic Effects from Parenteral Exposure

Parenteral exposures involving injecting a substance into animal, typically into a vein (i.v.) or

into the abdominal cavity (intraperitoneal or i.p. administration). These studies are used 

primarily as qualitative screening tools to assess general toxicity for both biological and chemical

agents as well as pathogenicity and infectivity for biological agents. Two studies are listed in 

the U.S. EPA (1996) RED: Terrell and Parke 1976c and Terrell and Parke 1976d. Both of these 

studies appear to be identical, indicating no mortality or signs of toxicity in mice after a single

intraperitoneal dose of about 125 mg/kg bw (Appendix 1). 

3.1.13. Respiratory Effects and Inhalation Exposures

Two standard acute inhalation studies have been conducted on Gypchek and are summarized in

Appendix 1. Neither of these studies gives a direct indication of toxicity. In one study, no overt

signs of toxicity were observed in a group of 10 male rats exposed to 6.12 mg/L Gypchek for 1

hour. During exposure, the rats were inactive and had closed eyes and labored respiration. 

Examinations for lung and trachea pathology 1, 7, and 14 days after recovery revealed no effects

attributable to exposure (Brown 1976). In the other inhalation study, rats were subjected to heads

only exposure to avoid ingestion during grooming (Thornett 1975). The test material was a 

white dust with 1.76 @ 1011 OB/g. The exposure concentrations ranged from 0.028 to 0.81 mg/L. 

No signs of toxicity were observed in any of the rats during exposure or upon necropsy. 

As noted in Section 3.1.7, Shope et al. (1975) used intranasal instillations to assess differences in

response between immunosuppressed and immunocompetent mice. Intranasal instillations are 

3-6
 
 

sometimes used as surrogates for inhalation exposures, particularly for biological agents that

have a low order of toxicity and pathogenicity. Other than expected changes in

immunocompetent mice associated with exposure to a foreign protein, no signs of pathogenicity 

were apparent. 

3.1.14. Impurities and Contaminants

As indicated in Section 2.2, Gypchek is produced by culturing and processing gypsy moth larvae

infected with LdNPV (Lewis 1971; USDA/FS 1975). The main contaminant in Gypchek is

gypsy moth parts, which account for a substantial proportion (80-88%) by weight of the

formulation (USDA/FS 19??a,b,c; USDA/FS 2003). In response to the potential for Gypchek to

become contaminated with bacteria, a quality control program has been developed to ensure that

batch preparations of NPV do not contain harmful bacteria (Podgwaite and Bruen 1978). The 

program consists of tests to determine bacterial counts of total aerobes, anaerobes, and bacterial

spores; an enumeration of total and fecal coliform bacteria, assays for primary pathogens (that is,

Salmonella, Shigella, Vibrio, Streptococcus, Staphylococcus, and Clostridium) and an in vivo 

pathogenicity test in mice. These tests are performed on each batch of Gypchek before it is used. 

3.1.15. Inerts and Adjuvants

As indicated in Section 2.3, Gypchek is typically applied with a carrier, either Carrier 038A or a

lignosulfonate-molasses carrier (Web et al. 1999c). Another product, Blankophor, may also be

included in Gypchek applications to enhance the persistence and activity of LdNPV (Thorpe et

al. 1999; Webb et al. 1999a,b). 

Carrier 038A is an aqueous surfactant mixture consisting of 58.5% water and 41.5% proprietary

surfactant mixture (Omnova Solutions 1999). Further details on the nature of the surfactant 

mixture are not available. The MSDS for Carrier 038A indicates that the surfactant mixture may

cause mild to moderate eye, skin, and respiratory tract irritation. This is true for most surfactants, 

including household soaps, which may disrupt the lipid structure in biological membranes

including those of the skin, eyes, and respiratory tract. The only specific information of the

toxicity of Carrier 38A is a standard acute toxicity study in rainbow trout (Drottar and Krueger

2001) in which the 96-hour LC50 value was 914 mg/L with a corresponding NOEC of 600 mg/L. 

Based on the categorization system currently used by U.S. EPA/EFED (2001), Carrier 038A

would be classified as practically nontoxic to rainbow trout. 

Blankophor is the common or trade name for the disodium salt of 2,2'-stilbendisulfonic acid,

4,4'-bis( (4-anilino- 6-morpholino-s-triazin-2-yl)amino) (NIOSH 2003). The toxicity data

available on this compound indicates that the compound has a very low acute oral toxicity with

reported LD50 values in excess of 80,000 mg/kg. In repeated dose skin exposures in rats at a dose

of 21,000 mg/kg bw, changes were seen in kidney and serum. This study is summarized by

NIOSH (2003) and is a 1966 study from the Bulgarian literature. Blankophor serves primarily to

protect the LdNPV virus from sunlight but may also enhance the toxicity of the LdNPV to the

gypsy moth (Thorpe et al. 1999). The U.S. EPA is in the process of registering Blankophor as a 

new pesticide inert (www.bnckay.com/inerts.htm). 

3-7
 
3.2. EXPOSURE ASSESSMENT 

3.2.1. Overview 

Because adverse effects associated with Gypchek or LdNPV, there is little basis for conducting a

detailed exposure assessment for Gypchek. Gypchek does contain gypsy moth parts and these

constituents, as with gypsy moth larvae themselves, have irritant effects in humans. The use of 

Gypchek, however, will not add substantially to exposures to gypsy moth parts in infested areas

and will serve to reduce exposure to gypsy moth larvae by reducing larval populations. 

Based on simple physical processes associated with the application of any pesticide, it is possible

to construct any number of exposure scenarios for Gypchek. The current risk assessment focuses 

on one extreme exposure scenario involving the accidental spray of a home garden. While 

Gypchek is not intentionally applied to such vegetation, the inadvertent spray scenario is

plausible. Based on this accidental exposure scenario, the estimated dose to an individual is

0.034 mg Gypchek/kg bw, with an upper range of 0.66 mg Gypchek/kg bw. 

3.2.2. LdNPV and Gypsy Moth Parts in Gypchek

In the re-registration of both LdNPV and OpNPV, the related virus used to control the Douglasfir Tussock moth, the U.S. EPA (1996) determinated that formal exposure assessments for the

general public and workers were not required. Two reasons for this decision are given. First, 

there is essentially no reason to assert that any adverse effects are plausible, and, as subsequently

detailed in section 3.3, there is no standard dose-response assessment. In other words, there is no 

indication that LdNPV will cause systemic adverse effects; therefore, a formal exposure

assessment would serve little purpose. 

Secondly, the use of LdNPV to control gypsy moth populations is likely to reduce rather than

increase exposure to the insect parts that are in Gypchek preparations: 

Spraying of the PIBs of OpNPV and LdNPV will not significantly

increase exposure to larval hairs, microbes, or other by-products

that occur in the preparation of the ai’s [active ingredients]. Pest 

densities that necessitate spraying have a natural high background

of these factors; moreover, dilution of the ai’s in the spraying

preparation and its sticking to the forest foliage reduce the

likelihood of exposure to a negligible level. (U.S. EPA 1996, p.

17) 

In other words, the use of either LdNPV will not increase exposure to both the viruses in these

products and the insects that they control. 

The potential for Gypchek to reduce exposure to both the LpNPV and the moth larvae can be

discussed in some detail. As summarized in Section 2.2, the application rates for Gypchek range 11 12 from 4 @ 10 PIB/acre per application to 1 @ 10 PIB/acre per application. As noted in Section 9 2.2, the average yield in the production of Gypchek is about 2×10 PIBs per larva (Lewis 1971). 

Thus, at the lower application rate of 4 @ 1011 PIB/acre, the number of larval equivalents applied 11 9 at the nominal application rate is about 200 larvae/acre [4 @ 10 PIB/acre ÷ 2×10 PIBs/larva]. At 

the higher application rate, the corresponding value is 500 larvae/acre [1 @ 1012 PIB/acre ÷ 2×109 

PIBs/larva]. This is actually a substantial overestimate because it does not consider the partial

removal of insect parts during the production of Gypchek. By comparison, the density of gypsy

moth larvae can be on the order of 10,000–100,000 larvae/acre. Thus, treatment during a severe

infestation would increase exposure to the larvae by only about 0.2% [200 larvae/acre ÷ 100,000

larvae/acre = 0.002] to 2%[200 larvae/acre ÷ 10,000 larvae/acre = 0.02]. Treatment of areas 

3-8
 
with a lower infestation rates would reduce exposure by inhibiting the increase in the larval

population by a substantial amount with a subsequent reduction in LdNPV exposure. 

3.2.3. Supplemental Extreme Exposures

While the approach taken by U.S. EPA (1996) is reasonable – i.e., provide no formal exposure

assessment because no hazard is apparent – this risk assessment of LdNPV is part of a series of

risk assessments involving several different control agents and at least a partial exposure

assessment is developed in order to facilitate a comparison of risk among the different control

agents that may be used by the Forest Service. For this risk assessment on Gypchek, the most

plausible route of exposure for humans will involve the consumption of contaminated vegetation. 

While Gypchek is not used directly on food crops, it is plausible that home-grown vegetation

could be incidentally contaminated in the aerial application of Gypchek. 

As indicated in Section 2.3, Gypchek is applied at a rate of up to about 0.03 kg/acre – i.e., 30.6

g/acre for eradication – or about 0.066 lb/acre. The concentration of any material deposited on 

vegetation will depend on the characteristics of the vegetation (i.e., effective surface area to

weight ratio) and application rate. In most Forest Service risk assessments (SERA 2001) as well

as risk assessments conducted by U.S. EPA, empirical relationships proposed by Fletcher et al.

(1994) are used to estimate initial concentrations on vegetation. For broadleaf forage plants,

similar to those that might be grown in a domestic garden, Fletcher et al. (1994) estimate residue

rates of 45 to 135 mg pesticide/kg vegetation per pound active ingredient applied. The 

consumption of homegrown vegetation is relatively well documented (U.S. EPA/ORD 1996). 

Individuals between the ages of 20 and 39 will typically consume about 0.000761 kg of

homegrown vegetation per kg of body weight with 95% confidence intervals on consumption

ranging from 0.0000777 to 0.00492 kg veg/kg bw (U.S. EPA/ORD 1996, Table 12-15, p. 9-14). 

Thus, taking the typical residue rate of 45 mg/kg vegetation and the typical consumption rate of

0.000761 kg veg/kg bw, the typical dose for an individual would be 0.034 mg Gypchek/kg bw. 

As an upper range on exposure, the 135 mg/kg residue rate may be used with the upper range on

consumption, 0.00492 kg veg/kg bw, to calculate a dose of 0.66 mg Gypchek/kg bw. 

A large number of other less extreme exposure scenarios could be developed for Gypchek but

would serve little purpose in terms of assessing potential risk. As noted in Section 3.4, the upper

range dose of 0.66 mg/kg bw is far below the no observed effect levels for Gypchek. 

3-9
 
3.3. DOSE-RESPONSE ASSESSMENT 

3.3.1. Overview 

Because no systemic toxic effects can be qualitatively identified for any plausible routes of

exposure (i.e., oral, dermal, or inhalation), the U.S. EPA has not derived either an acute or

chronic RfD for Gypchek. While this is a reasonable approach, the current risk assessment

derives a surrogate acute RfD of 26 mg/kg bw based on an experimental acute NOAEL of 2,600

mg/kg bw in rats and the application of an uncertainty factor of 100. This approach is taken

simply to provide a more quantitative basis for comparing the extremely low risks associated

with the application of Gypchek to the risks posed by other agents that may be used to control the

gypsy moth. 

Technical grade Gypchek is an eye irritant. While this is not quantitatively considered in this risk

assessment, the distinction between the irritant properties of technical grade Gypchek and the

lack of eye irritation with Gypchek formulations as applied in the field is emphasized in order to

highlight areas in which prudent handling practices are likely to be most important. 

3.3.2. Surrogate RfD for Acute Exposures

The U.S. EPA (1996) did not propose a dose-response assessment for Gypchek or LdNPV. This 

approach is reasonable because no systemic toxic effects can be qualitatively identified for any

plausible routes of exposure (i.e., oral, dermal, or inhalation). As noted in the exposure

assessment, however, the current risk assessment on Gypchek is part of a series of risk

assessments on several different agents. In order to facilitate an at least crude risk comparison

among the different agents, a dose-response assessment for oral exposures will be developed. 

As noted in Section 3.1.4, a single dose of LdNPV at 400 mg per rat was not associated with any

adverse effects in male or female rats over a 30-day observation period (Terrell and Parke

1976a,b). At a somewhat higher dose, 500 mg per rat, a transient (2 week) but statistically

significant decrease was noted in body weights over a 35-day observation period (Terrell et al.

1976c). For the purposes of this risk assessment, 400 mg will be taken as an acute NOAEL. 

Taking the upper range of the reported body weights of the rats – i.e., 150 grams or 0.15 kg – the

400 mg dose corresponds to a NOAEL of about 2,600 mg/kg bw. Following the general

approach of a 10 fold-safety factor for sensitive subgroups and a 10 fold safety factor of for

animal to human extrapolation, the 2,600 mg/kg bw dose will be divided by an uncertainty factor

of 100 and a dose of 26 mg/kg bw will be adopted as a surrogate acute RfD for the risk

characterization (Section 3.4). 

3.3.3. Eye Irritation

Although Gypchek has a very low order of systemic toxicity, Gypchek may cause eye irritation

and this endpoint is a concern at least for occupational exposures. This judgment is consistent

with the assessment made by U.S. EPA (1996) in the re-registration of Gypchek. As discussed in 

Section 3.1.11, Gypchek is moderately irritating to the eyes when assayed at full strength (TGAI)

in the rabbit eye (see discussion of Kuhn 1997a in Section 3.1.11). In the RED, the U.S. EPA 

(1996) noted the requirement for the following label warning concerning eye irritation for

Gypchek: 

a label statement is required indicating that these products are

severe eye irritants and specifying appropriate eye protection. 

Toxicity Category I for primary eye irritation requires products

containing the ais [active ingredients] to be labeled with the signal

word "Danger" and the appropriate Statements of Precaution and

Personal Protective Equipment, Practical Treatment, and Note to

Physician. 

3-10
 
On review of the study using 2X Gypchek (Kuhn 1997b) in which no eye irritation was noted

(Section 3.1.11), the U.S. EPA (Williams 1998) revised this assessment and concluded that: 

The study [2X] demonstrated that the products, Gypchek and TMBiocontrol, at concentrations twice standard dilution rate are 

“non-irritating”. 

Thus, eye irritation may remain a concern in the manufacture or mixing of Gypchek and prudent

industrial hygiene practices should be used to limit the possibility of contamination of the eyes. 

3-11
 
3.4. RISK CHARACTERIZATION 

3.4.1. Overview 

There is no basis for asserting that any risk is plausible to either workers or members of the

general public in the use of Gypchek to control the gypsy moth. As discussed in both the 

exposure and dose-response assessments, the current risk assessment extends the U.S. EPA risk

assessment by proposing a surrogate acute RfD and presenting a very conservative exposure

assessment based on the accidental spray of a home garden. This approach is taken simply to

facilitate the comparison of risks (or lack of risk) associated with Gypchek to the risks associated

with other agents used to control the gypsy moth. Based on a relatively standard dose-response

assessment and very conservative exposure assumptions, plausible exposures to Gypchek are

below a level of concern by factors of about 50 to over 750. While more typical exposures – i.e.,

incidental exposure to Gypchek in water or air – are not provided, they will be substantially less

than the range of doses in the accidental exposure scenarios used to quantify risk. 

3.4.2. Pathogenicity and Systemic Toxicity

Because Gypchek and LdNPV do not appear to cause adverse effects (Section 3.1), there is no

basis for asserting that any risk is plausible to either workers or members of the general public in

the use of Gypchek to control the gypsy moth. This conclusion is concurrent with the 

conclusions reach by U.S. EPA (1996) concerning the use of Gypchek as well as a related

product, TM-Biocontrol: 

The Agency does not expect any risk to humans or the environment

from use of these biopesticides; therefore, all uses are eligible for

reregistration. The bases of this decision are: 

evaluation of the submitted data and published scientific

literature for the RED indicate the data base is complete

and acceptable for all data requirements; 

the fact that PIBs of OpNPV and LdNPV are

naturally-occurring pathogens of gypsy moth and Douglas

fir tussock moth and are selective for Lymantriids with no

known adverse effects to any species other than the hosts,

gypsy moth and Douglas fir tussock moth; and 

the fact that in approximately 20 years of use, there have

been no reports of adverse human health and ecological

effects, with the exception of possible dermal sensitivity and

eye irritation in exposed humans during manufacture.

–U.S. EPA, 1996, pp. 24-25 

In other words, there is no basis for asserting that any exposures to Gypchek are likely to harm

either workers or members of the general public. 

3.4.3. Extreme Exposure Scenarios

Notwithstanding the above assertions, this risk assessment does attempt to quantify risk from one

extreme exposure scenario – the inadvertent spray of a home garden. This is an extreme scenario 

because Gypchek should not be applied to any vegetation other than tree species that contain

gypsy moth larvae (U.S. EPA 1996). Nonetheless, in aerial applications, an accidental spray of a 

home garden could occur. Based on the upper range of the application rate, the upper range of

contamination rates, and the upper range of the consumption of homegrown vegetation, the

highest estimated dose is 0.66 mg/kg bw (Section 3.2.3). Based on the surrogate acute RfD of 26 

3-12
 
mg/kg bw (Section 3.3.2), this results in a hazard quotient of 0.02, below the level of concern

(i.e., a hazard quotient of one) by a factor of 50. Other more plausible exposure scenarios would 

lead to much smaller hazard quotients. For example, based on the upper range of the application

rate but using the typical residue rate typical consumption rate, the typical dose for an individual

would be 0.034 mg Gypchek/kg bw, with a corresponding hazard quotient of 0.0013, which is

below the level of concern by a factor of over 750. 

3-13
 
4. ECOLOGICAL RISK ASSESSMENT 

4.1. HAZARD IDENTIFICATION 

4.1.1. Overview. 

Similar to the hazard identification for the human health risk assessment, there is no indication 

that LdNPV or the Gypchek formulation of LdNPV has the potential to cause any adverse effects

in any nontarget species. The mammalian toxicity data base for LdNPV is reasonably complete

and indicates that LdNPV is not pathogenic or otherwise toxic to mammals. One specific study

conducted on wildlife mammals that may consume contaminated gypsy moth larvae indicates no

adverse effects in mice, shrews, and opossums. Relative to the large number of available studies

in mammals, few studies are available in birds but the results of these studies are essentially

identical to those in mammals indicating that exposures to LdNPV at levels that are substantially

higher than those likely to occur in the environment will not be associated with any adverse

effects. Based on bioassays of LdNPV on the large number of nontarget insect species and

supported by the general high species specificity of related baculoviruses, the hazard

identification for LdNPV in nontarget insects is essentially identical to that in birds and

mammals. There is no indication that adverse effects will be caused in nontarget insects at any

level of exposure. Relatively few studies have been conducted in fish and aquatic invertebrates

but these studies are consistent with studies in terrestrial species and indicate that effects on fish

or aquatic invertebrates are unlikely. No data are available on the effects of LdNPV on 

amphibians, aquatic or terrestrial plants or other microorganisms. While this lack of information 

does, by definition, add uncertainty to this risk assessment, there is no basis for asserting that

effects on these or other organisms are plausible. 

4.1.2. Toxicity to Terrestrial Organisms. 

4.1.2.1. Mammals – The hazard identification for mammals is closely related to the hazard

identification for the human health risk assessment (Section 3.1) in that both may be based, at

least partially, on a number of standard toxicity studies in experimental mammals (Appendix 1). 

As summarized in Appendix 1 and discussed in Section 3.1, adverse systemic effects caused by

Gypchek or LdNPV have not been observed in mammals. Except for eye irritation, there is little 

indication that LdNPV or the Gypchek formulation of LdNPV will have any effect in mammals 

even at extremely high levels of the exposure. The relationship of plausible exposures to any

potential effect is discussed further in Section 4.4 (Risk Characterization). 

One study has been specifically conducted on wildlife mammals – i.e., mammals other than the

common test species used in the human health risk assessment. As summarized in Appendix 1,

Lautenschlager et al. (1977) exposed mice, short-tailed shrews, and opossums to various forms of

LdNPV: gypsy moth larvae infected with LdNPV, a purified formulation of LdNPV, and a spray

preparation of LdNPV. Based on both gross observations as well as necropsy and microscopic

examination of several different tissues, no effects were seen in any species. Again, this is

consistent with the relatively complete set of standard toxicity studies available on commonly

used laboratory mammals (Section 3.1). In addition, as discussed in Section 3.1.9, reproduction

in paired mice was higher in the LdNPV treated mice than the control group. While this study

was not a formal or standard assay for reproductive performance, it is the only reproduction study

available. Consistent with the other toxicity studies on LdNPV, the results provide no basis for

asserting any plausible hazard in mammals exposed to LdNPV or the Gypchek formulation. 

4-1
 
4.1.2.2. Birds – The available studies in birds are detailed in Appendix 2. Relative to the large

number available studies in mammals, few studies are available in birds but the results of these 

studies are essentially identical to those in mammals indicating that exposures to LdNPV at

levels that are substantially higher than those likely to occur in the environment will not be

associated with any adverse effects. 

One relatively standard dietary exposure study has been conducted in mallard ducks, a common

test species for assessing the effects of pesticides on birds (Roberts and Wineholt 1976). At 9 exposure levels of up to 1.04x10 PIBs/g of feed (estimated by the authors to represent exposures

equivalent to 100 times the normal application rate), no adverse effects associated with treatment

were observed. As with most toxicity studies in birds, clinical biochemistry and histopathology

were not conducted. 

In a field simulation study (Podgwaite and Galipeau 1978), black-capped chickadees and house

sparrows were fed LdNPV infected gypsy moth larvae every other day for 3 weeks. This study

included histopathology and, as with the comparable studies in mammals, no adverse effects

were noted based on histopathology, changes in body weight or gross signs of toxicity. 

Lautenschlager et al. (1976b) conducted a field study on resident songbirds and caged quail in

areas treated with two different formulations of LdNPV (see Appendix 2 for details). Consistent 

with the standard toxicity studies, no evidence of direct adverse effects from exposure to LdNPV

were noted. In addition, the study noted no secondary adverse effects on birds that use gypsy

moth larvae as a food source. Compared to untreated plots that were infested with gypsy moth

larvae, the secondary effect of LdNPV treatments appeared to be an enhancement songbird

habitat secondary to a reduction in defoliation from gypsy moth larvae. 

4.1.2.3. Terrestrial Invertebrates – The primary characteristic of LdNPV as well as many

related viruses involves a very high degree of host specificity – i.e., the virus is pathogenic to one

or only a very small number of species. LdNPV specifically is a member of the Baculoviridae

that includes both nucleopolyhedroviruses, such as LdNPV and OpNPV, as well as

granuloviruses (Döller 1985). Both budded viruses and occluded viruses are produced by 

baculoviruses. The budded viruses participate in cell to cell spreading of the infection, and the

occluded viruses participate in the spread of the infection among individual insects in a

population (Russell and Rohrmann 1997, Theilmann et al. 1996). Baculoviruses have been 

isolated only from arthropods and are characterized by a very limited host range (Chou et al.

1996). 

This general tendency for host specificity in baculoviruses has been demonstrated for LdNPV. 

As summarized in Appendix 3, LdNPV has been assayed in 46 species of nontarget Lepidoptera

(Barber et al. 1993), 17 genera and 31 species of ants (Wang et al. 2000), as well as a species of

fly (Barber et al. 1993), the common honey bee (Cantwell et al. 1972; Knoz 1970), and the

leafcutting bee (Barber et al. 1993). The studies by Barber et al. (1993) specifically assayed for

infectivity and found no indication that LdNPV is pathogenic to any insect species except the

gypsy moth. No adverse effects were observed in any species tested in any of these studies. In 

addition, the recent field study by Rastall et al. (2003) noted no effects in nontarget insects after

the application of Gypchek. In this study, Gypchek was applied at a rate of 2×1011 OB/acre in 

May of 1997 and 1998 to two forests susceptible to gypsy moth. Nontarget lepidoptera were 

monitored in two pre-treatment year as well as in treatment years. No statistically significant 

effects were associated with the Gypchek applications. 

4-2
 
Thus, based on the large number of species assays with LdNPV, a recent field study, and

supported by the general high species specificity of related baculoviruses, the hazard

identification for nontarget insects is essentially identical to that in birds and mammals. There is 

no indication that adverse effects will be caused in nontarget insects at any level of exposure. 

4.1.2.4. Terrestrial Plants (Macrophytes) – No phytotoxicity studies on LdNPV were

encountered and the U.S. EPA waived the requirement for such tests (U.S. EPA 1996). This 

appears to be a reasonable approach in that there is no basis for supposing that LdNPV is likely

to be toxic to any form of vegetation. The only effect that is plausible is the protective effect that

LdNPV will have in terms of preventing damage to vegetation from gypsy moth larvae. 

4.1.2.5. Terrestrial Microorganisms – No studies have been encountered on the effects of 

LdNPV on terrestrial microorganisms. There is no apparent basis for asserting that direct effects 

– i.e., microbial toxicity – are plausible. The protective effect of LdNPV on vegetation is likely

to affect soil microorganisms in that the microbial soil community is likely to change secondary

to changes in terrestrial vegetation. 

4.1.3. Aquatic Organisms.

4.1.3.1. Fish – Two studies are available on the toxicity of LdNPV to fish (Moore 1977;

Kreutzweiser et al. 1997) and the results of both studies are consistent with the data on terrestrial

species: there is no indication of toxicity or pathogenicity. 

In the study by Moore (1977), a “crude nuclear-polyhedrosis virus preparation” was tested in

both bluegill sunfish and brown trout. Fish were exposed to LdNPV for 96 hours and observed 8 for 30 days after exposure. The test concentrations are given in the study as 7.5×10 PIB/gram 9 of fish or 1.5×10 PIB/gram of fish (Moore 1977, Table 2, p. 10). Details on how these 

exposures are calculated are not given. In addition to standard observations for mortality,

appearance and general behavior, histopathology was conducted on gill arches, stomach, liver,

and intestines. Fish were equally divided among control groups, low concentration and high

concentration groups. A total of 240 fish of each species were used and no treatment related

effects were noted in either species. 

Kreutzweiser et al. (1997) assayed LdNPV in rainbow trout after the viruses were fed to the trout

in standard feed pellets at a dose of 1.6×106 occlusion bodies (OBs)/fish. Since each fish 

8 weighed approximately 6 g, this corresponds to a dose of about 2.7×10 OBs/kg bw. The study

covered a 21-day treatment period in which the fish were fed on days 1, 3, 5, 8, 10, 12, 15, 17,

and 19. No effects were noted on mortality, behavior, growth rate, or gross pathological

examination of the internal organs. In addition, no viable NPV was detected in the stomach or 

intestinal tract. As reviewed by Kreutzweiser et al. (1997), these results are consistent with the

general observation that “NPVs cannot induce protein production nor reproduce in vertebrate 

cells in general”. (Kreutzweiser et al. 1997, p. 68, column 1). 

4.1.3.2. Amphibians – No data have been encountered on the effects of NPV exposures to 

amphibians. 

4.1.3.3. Aquatic Invertebrates – Only one study (Streams 1976) has been encountered on the

toxicity of LdNPV to aquatic invertebrates. This study, however, involved five species: Daphnia 

magna (a commonly used test species in aquatic toxicology), backswimmers (Notonecta 

undulata), midge larvae (Chironomus thummi), and two species of water boatmen (adult 

Hesperocorixa interrupta and Sigara gordita). As detailed in Appendix 4, no effects were

observed on mortality or reproduction in any species over exposure periods of up to four weeks. 

4-3
 
While this study is not a standard bioassay typically conducted on pesticides, it provides much

more detailed information than standard bioassays and has been accepted by U.S. EPA (1996) as

indicating no apparent toxicity to aquatic invertebrates. 

4.1.3.4. Aquatic Plants – As with terrestrial plants, no studies have been conducted on the

toxicity of LdNPV to aquatic plants. Given the lack of any biological basis for asserting that

direct effects on aquatic plants are plausible, this does not add substantial uncertainty to the risk 

assessment. The U.S. EPA (1996) has explicitly waived the requirements for toxicity testing in

nontarget plant species. 

4-4
 
4.2. EXPOSURE ASSESSMENT 

4.2.1. Overview 

In ground or aerial applications, it is likely that a large number of species could be exposed to

Gychek/LdNPV. Because of the apparently very low toxicity of Gypchek and LdNPV, the need

for any formal exposure assessment is questionable. Nonetheless, in an attempt to provide some

bases for comparing the potential risks of Gypchek to other agents used to control the gypsy

moth, two extreme exposure assessments are developed: one for a terrestrial herbivore

consuming contaminated vegetation and the other for aquatic organisms in a small pond directly

sprayed with Gypchek at the highest application rate. For the terrestrial herbivore, the dose 

estimates range from 1.1 mg Gypchek /kg bw to 3.2 mg Gypchek /kg bw. For aquatic organisms,

concentrations are expressed in units of PIB/liter because this unit is used in the corresponding

toxicity studies. For a small pond directly sprayed with Gypchek at the highest application rate, 5 the estimated initial concentration is 2.5×10 PIB/L. A large number of other less extreme 

exposure assessments could be developed but these would not alter the assessment of risk since

these extreme exposure assessments are substantially below any level of concern. 

4.2.2. LdNPV and Gypsy Moth Parts in Gypchek

As with the human health risk assessment, a formal exposure assessment for Gypchek is not

necessary because of the failure to identify any adverse effects. As discussed in section 3.2, the 

application of Gypchek in areas infested by the gypsy moth will not substantially increase

exposure to either LdNPV or the larval parts (e.g., hairs) that contaminate Gypchek. To the 

contrary, treatment of gypsy moth infestations with Gypchek is likely to reduce longer term

exposures to both the larval parts and the virus by reducing the population of gypsy moth and

lessening the chance of a substantial increase in the gypsy moth population (U.S. EPA 1996). 

4.2.3. Supplemental Extreme Exposures

As with the human health risk assessment (Section 3.2), some extreme exposure scenarios will

be developed for Gypchek and used in the risk characterization (Section 4.4). Again, this

approach is taken to facilitate comparisons of risk among the various agents that may be used to

control or eradicate gypsy moth infestations. Two specific exposure scenarios are developed:

one for a large vertebrate consuming vegetation directly sprayed with Gypchek and the other for

aquatic species in a small pond directly sprayed with Gypchek. Both of these scenarios should be 

regarded as extreme, since efforts are made in the application of Gypchek to avoid contamination

of vegetation that will not be habitat for the gypsy moth (e.g., grasses) as well as incidental

contamination of open water. 

4.2.3.1. Contaminated Vegetation – For terrestrial species, an exposure assessment is developed

for a large herbivore, such as a deer, consuming contaminated vegetation. The general approach

is similar to that used in the human health risk assessment except that the deer is assumed to

consume contaminated grass rather than broadleaf vegetables. This approach is taken because

contaminated grass is estimated to have higher residue rates – i.e., 85 and 240 mg pesticide/kg

vegetation per pound active ingredient applied per acre – than the corresponding values for

broadleaf vegetation – i.e., 45 mg pesticide/kg vegetation to 135 mg pesticide/kg vegetation per

pound active ingredient applied per acre (Fletcher et al. 1994). Thus, at an application rate of

0.066 lb Gypchek/acre (Section 2.3), the estimated initial residues on vegetation would be in the

range of about 5.6 mg Gypchek/kg vegetation [85 mg pesticide/kg vegetation per lb/acre × 0.066

lb/acre = 5.61 mg/kg] to 16 mg Gypchek/kg vegetation [240 mg pesticide/kg vegetation per

lb/acre × 0.066 lb/acre = 15.84 mg/kg]. 

In order to estimate the dose to the deer, the amount of vegetation consumed must be estimated. 

This will be highly variable, depending on the amount of grass consumed relative to other types 

4-5
 
of vegetation and the amount of time spent grazing at the treated site. As a very conservative

upper limit, it will be assumed that the deer consumes its caloric requirement for food totally as

contaminated grass. Caloric requirements for mammals are well-characterized. The U.S. 

EPA/ORD (1993, p. 3-6), recommends the following relationship based on body weight (BW):

kcal/day = 1.518 × W(g) 0.73. Based on this relationship, a 70 kg deer would require 

approximately 5226 kcal/day [1.518 × 70,000 g0.73 = 5226.288]. The caloric content of 

vegetation is given by U.S. EPA/ORD (1993, p. 3-5) as 2.46 kcal/gram vegetation dry weight

with a corresponding water content of 85% (U.S. EPA/ORD 1993, p. 4-14). Correcting the dry

weight caloric content to wet weight, the caloric content of the grass will be taken as 0.369 kcal/g

[2.46 kcal/gram vegetation dry weight × (1-0.85) = 0.369 kcal/g]. Thus, the 70 kg deer would

consume about 14.2 kg of grass per day [5226 kcal/day ÷ 0.369 kcal/g = 14,162.6 g, which is

equal to about 14.2 kg]. 

At the lower range of the estimated residue rate of 5.6 mg Gypchek/kg vegetation, the estimated

dose to the deer would be 1.1 mg Gypchek /kg bw [5.6 mg Gypchek/kg vegetation × 14.2 kg

vegetation ÷ 70 kg bw = 1.136 mg Gypchek /kg bw]. At the upper range of the estimated residue

rate of 16 mg Gypchek/kg vegetation, the estimated dose to the deer would be about 3.2 mg

Gypchek /kg bw [16 mg Gypchek/kg vegetation × 14.2 kg vegetation ÷ 70 kg bw = 3.2457 mg/kg

bw]. 

4.2.3.2. Small Pond – For the risk characterization of aquatic species, one extreme exposure

scenario is developed in which a small pond is directly sprayed with Gypchek at the highest

application rate. As discussed in Section 4.3.3, the toxicity data for aquatic species is expressed

in units of PIB/L. The highest application rate for Gypchek is 1×1012 PIB/acre (Section 2.3). 

2 For this exposure scenario, the small pond will be characterized as 1000 m in surface area with

an average depth of 1 meter. An application rate of 1×1012 PIB/acre corresponds to about 8 2 12 2 8 2 2.5×10 PIB/m [1×10 PIB/acre ÷ 4047 m /1 acre = 2.471×10 PIB/m ]. At a depth of 1 meter, 

each square meter of pond surface would correspond to 1 cubic meter of water or 1,000 liters. 5 Thus, assuming instantaneous mixing, the concentration in the water would be 2.5×10 PIB/L 8 [2.5×10 PIB ÷ 1000 L]. This concentration will be used directly to characterize risks to aquatic 

species. 

4-6
 
4.3. DOSE-RESPONSE ASSESSMENT 

4.3.1. Overview 

Because no hazards can be identified for any species, a quantitative dose-response assessment is

not required and no such assessments have been proposed by U.S. EPA and no quantitative doseresponse assessments were used in the previous USDA risk assessment for Gypchek. In order to 

provide a clear comparison of the risks of using Gypchek relative to other agents, dose-response

assessments are proposed in the current risk assessment for both terrestrial mammals and aquatic

species. For terrestrial mammals, the NOAEL of 2,600 mg/kg bw is used. This is the same 

NOAEL that served as the basis for the surrogate acute RfD in the human health risk assessment. 9 For aquatic species, only NOEC values are available and the highest NOEC of 8×10 PIB/L is

used to characterize risk. 

4.3.2. Qualitative Assessment 

There is no basis for asserting that Gypchek poses any risk to nontarget species. Consequently, a

standard dose-response assessment is not required for any species or groups of species and the

previous USDA (1995) risk assessment does not propose a quantitative dose-response

assessment for any wildlife species. This is essentially identical to the approach and conclusions

reached by U.S. EPA (1996) in the re-registration eligibility decision for both Gypchek and TMBiocontrol: 

The available avian and aquatic data and other relevant literature and

information show that PIBs of OpNPV and LdNPV do not cause adverse effects

on avian, mammalian and aquatic wildlife. No mortalities were seen when these

viruses were fed to mallard ducks, house sparrows, bobwhite quail and

black-capped chickadees. No mortalities or other adverse effects were seen in 

brown trout, bluegill sunfish, and a variety of aquatic invertebrates. Similarly,

tests with mule deer, Virginia opossums, short-tailed shrews and white-footed

mice, resulted in no evidence of pathogenicity or toxicity. Known insect host 

range and scientific literature on honey bee mortality demonstrate that these

baculoviruses do not have adverse effects on honeybees and should not pose a

significant risk to nontarget insects (Cantwell et al. 1972; Knox 1970). NPV 

effects on endangered species are considered a low risk based on the absence of

threat to nontarget organisms. (U.S. EPA 1996, pp. 23-24) 

4.3.3. Quantitative Assessments 

While the qualitative approach to assessing the potential effects in nontarget species is clearly

justified, the current risk assessment quantifies extreme exposures to Gypchek for both a

terrestrial herbivore and aquatic species (Section 4.2.3). As in the human health risk assessment, 

this approach is taken to permit a clearer comparison of risks among the different agent that may

be used in response to gypsy moth infestations. 

For a large herbivore consuming vegetation, exposures are expressed in units of mg Gypchek/kg

vegetation and the NOAEL of 2,600 mg Gypchek/kg bw used as the basis for the surrogate acute

RfD (Section 3.3.2) can used to characterize risks for the large herbivore. As discussed in 

Section 3.3.2, this NOAEL of 2,600 mg Gypchek/kg bw is based on the study by (Terrell and

Parke 1976a,b) in which rats weighing 100 to 150 grams were dosed with 400 mg Gypchek and

no adverse effects were noted over a 30-day observation period. At a somewhat higher dose, 500

mg Gypchek/rat, decreased food consumption with a corresponding decrease in body weight was

observed in a study by the same investigators (Terrell et al. 1976c). These studies are detailed 

further in Appendix 1. 

4-7
 
As discussed in Section 4.1.3, there are no studies indicating that Gypchek will be toxic or

pathogenic to any aquatic organisms under any exposure conditions. The most recent study,

Kreutzweiser et al. (1997), involved feeding trout with contaminated food pellets. While this 

study is useful for the qualitative assessment of pathogenicity and toxicity, the route of exposure

is not suitable for use in a quantitative risk assessment. 

The other two studies that could be used both involved exposures to Gypchek in water. The 

study in invertebrates (Streams 1976) used concentrations of 250 polyhedra/mL or 2.5×105 

PIB/L. The study in fish (Moore 1977) expresses exposures in units of PIB/gram of fish (Section

4.1.3.1). Moore (1977) does not specifically convert the exposure units in PIB/g fish to more

typical concentrations (e.g., PIB/liter of water) but does indicate loadings in units of grams of

fish per liter of water. For bluegills, the loading factor was 0.23 grams of fish per liter of water. 8 8 Thus, the concentrations would correspond to approximately 1.7×10 PIB/liter [7.5×10 8 8 9 PIB/gram of fish × 0.23 grams fish/L = 1.725×10 PIB/liter] and 3.45×10 PIB/liter [1.5×10 9 PIB/gram of fish × 0.23 grams fish/L = 0.345×10 PIB/liter]. For trout, the loading factors were 

5.31 grams of fish per liter of water and the corresponding concentrations were about 4×109 

8 8 9 PIB/liter [7.5×10 PIB/gram of fish × 5.31 grams fish/L = 39.825×10 PIB/liter] and 8×10 9 9 PIB/liter [1.5×10 PIB/gram of fish × 5.31 grams fish/L = 7.965×10 PIB/liter]. 

All of these exposures are essentially NOEC’s values – i.e., no effects were observed at any

concentrations. In the absence of an LOEC, the most appropriate value to use in risk 9 characterization is the highest NOEC, in this case 8×10 PIB/liter from trout in the study by

Moore (1977). In other words, if a large number of NOEC values are available with no

indication that any concentration will cause an adverse effect, it is appropriate and conservative

to use the highest NOEC because this NOEC is still below any concentration that would be

anticipated to cause an adverse effect. While the use of the lowest NOEC would be “more 

conservative”, it would tend to distort rather than clarify risk. 

4-8
 
4.4. RISK CHARACTERIZATION 

4.4.1. Overview 

There is no basis for asserting that the use of Gypchek to control or eradicate gypsy moth

populations is likely to cause any adverse effects in any species other than the gypsy moth. 

While no pesticide is tested in all species under all exposure conditions, the data base on LdNPV

and related viruses is reasonably complete and LdNPV has been tested adequately for

pathogenicity in a relatively large number of species, particularly terrestrial invertebrates. 

LdNPV appear to be pathogenic and toxic to the gypsy moth and only to the gypsy moth. 

Because Gypchek does not appear to cause adverse effects, quantitative expressions of risk are in

some respects more difficult because clear NOEC and LOEC values cannot be defined – i.e., if

an agent is not shown to cause an effect, the threshold exposure level is not a meaningful 

concept. Nonetheless, general but very conservative exposure assessments demonstrate that

plausible upper ranges of exposures are clearly below any level of concern by a factor of 1000 for

terrestrial species and 30,000 for aquatic species. 

4.4.2. Qualitative Assessment 

Gypchek does not appear to be capable of causing adverse effects in any species other than the

gypsy moth. Thus, the use of Gypchek to control or eradicate gypsy moth infestations appears to

carry no identifiable risk. This is essentially identical to the conclusions reached by U.S. EPA

(1996) in the re-registration of LdNPV and OpNPV: 

Due to the lack of adverse effects on avian, mammalian and 

aquatic wildlife, plants and nontarget insects documented in the

submitted studies and scientific literature after 20 years of use, the

Agency finds that the PIBs of L. dispar and O. pseudotsugata

NPVs pose minimal or no risk to nontarget wildlife, including

endangered species. 

The current re-evaluation of the available information supports this basic conclusion with no

reservations. 

As in the human health risk assessment, there are basically two agents that could be of concern in

the use of Gypchek: the virus and the insect parts. As discussed in Section 3.1 and 4.1, there is 

no indication that LdNPV is pathogenic or otherwise toxic to any species other than the gypsy

moth. To the contrary, experience with this as well as other related NPVs indicate that these

viruses have a very narrow host range. As is also true for the human health risk assessment, the 

overriding consideration in the risk characterization for nontarget species is that the use of

Gypchek will decrease rather than increase exposure to the gypsy moth and LdNPV (Section

3.2.2). 

4.4.3. Quantitative Assessments 

The above qualitative assessment is adequate for assessing the plausibility of intended harm from

the use of Gypchek to control or eradicate gypsy moth populations. This risk assessment, 

however, is part of a larger effort to review the risks associated with the use of several different

and diverse agents and some quantitative expression of risk for Gypchek is useful both in further

demonstrating the apparent safety of this agent and in comparing potential risks among the

different agents that may be used. 

Based on the exposure assessment (Section 4.2) and dose-response assessment (Section 4.3), two

such expressions of risk may be made: one for a large mammal consuming contaminated

vegetation and the other for aquatic species in a small pond directly sprayed with Gypchek. As 

4-9
 
detailed in Section 4.2.3.1, a large mammal grazing exclusively on grass directly sprayed with

Gypchek at the highest application rate might consume as much as 3.2 mg Gycheck/kg body

weight. Using the acute NOAEL of 2,600 mg Gypchek/kg bw (Section 4.3.3), this exposure

would correspond to a hazard quotient of 0.001 [3.2 mg Gycheck/kg body weight ÷2,600 mg

Gypchek/kg bw = 0.00123]. In other words, the maximum level of exposure is below the

NOAEL by a factor of about 1000. This numeric expression of risk is thus consistent with the

qualitative risk characterization offered by U.S. EPA (1996) and the previous risk assessment on

Gypchek (USDA 1995). 

For aquatic species, the direct spray of a small pond is estimated to result in initial concentrations 5 of about 2.5×10 PIB/L. This is a reasonable worst case scenario in that direct spray of the pond 

at the highest application rate is assumed. Because there is no indication that any concentration

of Gypchek will cause any effect in any aquatic species, the highest available NOEC is used to 9 characterize risk – i.e., 8×10 PIB/liter from the trout study by Moore (1977), as discussed in 5 9 Section 4.3.3. Thus, the hazard quotient is 0.00003 [2.5×10 PIB/L ÷ 8×10 PIB/liter =

0.00003125], as factor of over 30,000 below the NOEC. Again, this numeric expression of risk

is in agreement with the qualitative conclusions reached by U.S. EPA (1996) and USDA (1995). 

4-10
 
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lymantriidae). J Entomol Sci. 25(3):383-393. 

Webb RE; Shapiro M; Podgwaite JD; Lynn DE; Dougherty EM; Ridgway RL; Venables L;

Cohen DL. 1993. Field comparison of different strains of gypsy moth nuclear polyhedrosis virus

against gypsy moth (Lepidoptera: lymantriidae) in western Maryland in 1990. J Econ Entomol. 

86: 1185-1190. 

Webb RE; Shapiro M; Podgwaite JD; Ridgway RL; Venables L; White GB; Argauer RJ; Cohen

DL; Witcosky J; Kester KM. 1994a. Effect of optical brighteners on the efficacy of gypsy moth

(Lepidoptera: lymantriidae) nuclear polyhedrosis virus in forest plots with high or low levels of

natural virus. J Econ Entomol. 87: 134-143. 

Webb RE; Dill NH; Podgwaite JD; Shapiro M; Ridgway RL; Vaughn JL; Venables L; Argauer

RJ. 1994b. Control of third and fourth instar gypsy moth (Lepidoptera: lymantriidae) with

Gypchek combined with a stilbene disulfonic acid additive on individual shade trees. J Entomol 

Sci. 29: 82-91. 

Webb RE; Peiffer RA; Fuester RW; Valenti MA; Thorpe KW; White GB; Shapiro M. 1999a. 

Effects of Blankophor BBH, a virus-enhancing adjuvant, on mortality of gypsy moth

(Lepidoptera: lymantriidae). J Entomol Sci. 34: 391-403. 

Webb RE; Thorpe KW; Podgwaite JD; Reardon RC; White GB; Talley SE. 1999b. Field 

evaluation of an improved formulation of Gypchek (a nuclear polyhedrosis virus product) against

the gypsy moth (Lepidoptera: lymantriidae). J Entomol Sci. 34: 72-83. 

Webb RE; Thorpe KW; Podgwaite JD; Reardon RC; White GB; Talley SE. 1999c. Efficacy of

Gypchek against the gypsy moth (Lepidoptera: lymantriidae) and residual effects in the year

following treatment. J Entomol Sci. 34: 404-414. 

Woods SA; Elkinton JS. 1987. Bimodal patterns of mortality from nuclear polyhedrosis virus in 

gypsy moth (Lymantria dispar) populations. J Invert Pathol. 50:151-157. 

5-13
 
Woods SA; Elkinton JS; Podgwaite JD. 1989. Acquisition of nuclear polyhedrosis virus from

tree stems by newly emerged gypsy moth larvae. Environ Entomol. 18(2):298-301. 

Woods SA; Elkinton JS; Shapiro M. 1990. Factors affecting the distribution of a nuclear

polyhedrosis virus among gypsy moth (Lepidoptera: Lymantriidae) egg masses and larvae. 

Environ Entomol. 19(5):1330-1337. 

5-14
 
APPENDICES
 

Appendix 1: Toxicity of LdNPV in Mammals 

Appendix 2: Toxicity of LdNPV in Birds 

Appendix 3: Toxicity of Gypsy Moth LdNPV in Nontarget Insects 

Appendix 4: Toxicity of Gypsy Moth NPV in Aquatic Invertebrates 

NOTE: Several of the studies summarized in these appendices appear to have been submitted to

U.S. EPA on more than one occasion and some with an inconsistent list of authors. This is 

indicated in the appendices by multiple references given for the same data summary. Unless 

otherwise specified, the multiple cited references for the same data are identical study

submissions. The multiple references are maintained in the appendices simply to avoid

confusion that might be associated with “missing” MRID numbers. 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

ACUTE ORAL
 

Gypsy Moth Single oral dose of 400 mg test No mortality and no adverse effects on Terrell and 

NPV prepared as material to 20 male and 20 behavior throughout the 30-day Parke 1976b 

20% suspension female Sprague Dawley rats. observation period. No treatment- MRID 

in distilled water Negative control group 

consisted of 20 males and 20 

females. All rats were observed 

for 30 days. Animals weighted 

between 100 and 150 grams. 

related gross pathological findings. 

NOTE: Although this is called a 

“feeding study” the precise route of 

exposure is not specified. 

00048862 

Terrell and 

Parke 1976a 

MRID 

00055915 

Gypsy Moth Single oral dose of 500 mg test Mortality in 8 control animals and 3 Terrell et al. 

NPV prepared as material to 20 male and 20 treated animals, all of which exhibited 1976c 

20% suspension female Sprague Dawley rats. overt physical and or behavioral MRID 

in distilled water Negative control group 

consisted of 20 males and 20 

females. All rats were observed 

for 35 days. Animals weighted 

between 100 and 150 grams. 

changes including piloerection, 

decreased locomotor activity, increased 

respiratory rate, and decreased body 

weight. 

Adverse treatment-related effects 

included statistically significant 

decreases in body weights of males for 

the first 2 weeks and statistically 

significant decreases in food 

consumption for males and females 

during the first week. 

No treatment-related adverse effects 

were noted regarding body 

temperature, hematological and clinical 

chemistry results, urinalysis parameters 

or necropsy examinations. 

00048863 

L. dispar NPV Single oral gavage dose of No signs of toxicity observed; no Hart 1976 

(Lot 33) NPV suspended in 0.9% saline 

at a concentration of 0.2 g/mL 

(equivalent to 1.32 PIB/mL) 

administered to fasted young 

adult rats (30 males and 30 

females, weighing 

approximately 125 g). Rats 

were observed daily for 30 

days. 

mortality. MRID 

00068401 

Appendix 1 - 1 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

ACUTE ORAL (continued) 

P. dispar NPV 1 Single oral gavage dose of test 

compound in 0.8% saline at a 

concentration of 40x109 

polyhedra/mL (dosage was 1 

mL of the stated suspension per 

rat) to 20 male and 20 female 

Sprague Dawley weanling 

albino rats. Negative controls 

(20 males and 20 females) 

received saline 

No mortality and no overt signs of 

toxicity during the 35-day observation 

period. 

Hart and 

Thornett 1975c 

MRID 

00049263 

Hart et al.1975a 

MRID 

00060702 

[Final Report] 

P. dispar1

 NPV Single virus exposure (gastric No treatment related adverse effects Shope et al. 

intact polyhedra intubation) to 0.02 g/animal observed; no mortality among 1975 

(suspensions polyhedra to adult mice [10 immunosuppressed mice; no lesions MRID 

contained males (5 untreated and 5 noted grossly post-mortem. 000606700 

1.8x1011 immunosuppressed) and 10 

polyhedra/g) females (5 untreated and 5 

immunosuppressed)]. 

Immunosuppressed mice were 

selectively depleted of cellmediated immune function by 

thymectomy and treatment with 

anit-lymphoctye serum 

(cytoxan administered ip at 300 

mg/kg/mouse). Positive 

controls treated with 

autoclaved polyhedra; negative 

controls treated with saline. 

All animals observed for 21 

days. 

Serological results indicated that the 

animals with intact immune systems 

were exposed to NPV antigen, since 

positive reactions were apparent with 

autoclaved and non-autoclaved NPV 

preparations. Control (saline) exposure 

did not produce antibody responses. 

Appendix 1 - 2 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

LONGER TERM ORAL
 

NPV of the Mammalian predators of the No adverse effects were observed Lautenschlager 

gypsy moth gypsy moth (40 white-footed 

mice caged in pairs; 6 shorttailed shrews caged 

individually; and 2 Virginia 

opossums caged individually) 

were collected in the field and 

exposed orally to NPV in the 

form of NPV-infected 5 gypsy th 

moth larvae, PIBs mixed in dog 

food, and PIBs mixed in a 

standard spray formulation for 

20 days. All animals were 

sacrificed on day 32. 

related to general body condition, 

weight, or reproductive efficiency 

(mice only species tested). In addition, 

necropsy and microscopic examination 

revealed no abnormalities resulting 

from exposure to NPV. 

et al. 1977 

MRID 

00134314 

The total amount of NPV 

consumed by each test mouse 

and shrew was equivalent to 

more than a 40-ha exposure for 

a 70 kg person assuming that 

NPV was applied at the rate of 

5.0x10 PIB/ha. No further 11 

details regarding these 

estimates are provided. 

NPV of the 

Gypsy Moth in 

distilled water 

Administration of daily doses 

of 0, 10 , 10 , or 10 7 8 9 

PIBs/animal to young adult, 

purebred beagles (13 males and 

14 females) over a period of 90 

days. These doses correspond 

Gypchek doses of 0, 1.8, 18, 

and 180 mg/dog or 

approximately 0.2, 1.6, and 17 

mg/kg/day based on terminal 

body weights in each dose 

group. The doses were 

delivered directly into the 

mouth of each dog and small 

amounts of sugar were added 

just before dosing to increase 

palatability. 

No evidence of toxicity. All treated 

and control animals were in good 

health throughout the study. 

Standard hematology, clinical 

biochemistry, and urinalysis were 

conducted on each animal at or before 

the start of exposure and at 2, 4, and 6 

months after the start of exposure. 

After sacrifice, standard examinations 

were conducted for signs of gross 

pathology or histopathology. No 

treatment related effects were observed. 

Hart and Wosu 

1975 

MRID 

00060698 

Hart 1975a 

MRID 

00067103 

[Final Report] 

Appendix 1 - 3 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

LONGER TERM ORAL (continued) 

P. dispar1 NPV Sprague Dawley rats (50 males Observations included body weight, Hart 1975b 

and 50 females/dose group) food consumption, gross signs of MRID 

exposed to dietary 

concentrations of 0, 107

 or 108 

toxicity, and pathology. No treatmentrelated effects on survival and no 00049267 

PIB/rat/day for 2 years. These significant differences in tumor Hart and 

doses correspond to Gypchek incidence or other lesions in treated Cockrell 1975 

daily doses of 1.8 or 18 mg/rat. rats, compared with controls. MRID 

The average terminal body 00060699 

weights (both sexes combined) Authors indicate overall survival to 

was approximately 400 g. termination at 104 weeks was 137/299 

Thus, the approximate average or 46%. Individual groups ranged 

dose rate was 4.5 or 45 mg from 32 to 60% with both extremes 

Gypchek/kg body weight. falling in the high dosage group. It 

seems clear that treatment did not 

influence survival. 

DERMAL
 

P. dispar NPV 1 Dermal application of 1/10 of 1 No mortality and no evidence of Hart and 

mL of test compound in 0.8% irritation (either erythema or edema) Thornett 1975d 

saline at a concentration of 

40x10 polyhedra/mL or freed 9 

resulting from exposure to NPV of the 

Gypsy Moth either as the polyhedra 

MRID 

00049263 

virus rods prepared from dry themselves or as virus rods freed from 

polyhedra to shaved and the polyhedra throughout observation Hart et al. 

abraded or shaved and intact period. No evidence of systemic 1975b 

skin of albino guinea pigs (5 toxicity. MRID 

males and 5 females/dose 00060703 

group). Treated sites were [Final Report] 

covered by 1"x1" gauze pads 

held in place by tape and 

covered by impermeable 

binding (rubber dam) for 24 

hours. Animals were observed 

for 21 days after treatment. 

Appendix 1 - 4 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

DERMAL (continued) 

P. dispar NPV 1 Dermal application of 0.5 mL 

test material (P. dispar1

 NPV 

suspended in 0.8% saline at the 

rate of 40x109 

polyhedra/animal) to shaved 

and abraded skin (3 rabbits) or 

shaved and intact skin (3 

rabbits). Treated sites were 

covered with 1" sq gauze patch 

and held in place with adhesive 

tape. Entire trunks were 

wrapped with nonabsorbent 

binder for 24 hours. After 24­

hour exposure, the skin was 

cleaned and the reactions were 

scored immediately and again 

at 72 hours after exposure. 

Primary irritation score = 0; there was 

no evidence of irritation in either intact 

or abraded skin and no edema was 

observed. Body temperatures were 

within normal temperature range except 

in one rabbit whose temperature was 

slightly depressed at 24, 48, and 72 

hours. This finding is judged to be 

idiosyncratic and not significant. 

Hart and 

Thornett 1975b 

MRID 

00066104 

P. dispar1

 NPV Dermal application of 0.04 g NPV treatment to ears caused positive Shope et al. 

intact polyhedra saline (negative controls), 

autoclaved polyhedra (positive 

controls) or polyhedra to 

shaved backs of 5 male and 5 

female albino guinea pigs with 

depressed cell-mediated 

immune functions after 

cortisone treatment (300 mg/kg 

ip)on two areas of intact skin 

and one ear. Exposed ears were 

measured for 7-10 days; areas 

larger than 16mm were 

considered positive. 

responses in 3/5 males and 5/5 females 

without immunosuppressive treatment. 

In animals with depressed cellmediated immune functions due to 

cortisone treatment, NPV caused 

positive responses in 3/5 males and 2/5 

females. 

None of the immunosuppressed 

animals died during the observation 

period. 

1975 

MRID 

000606700 

Shope et al. 

1977 

P. dispar NPV 1 Dermal application of 40x10 9 

polyhedra suspended in 0.8% 

saline (dose = 0.5 mL) to 

shaved abraded or intact skin of 

New Zealand white rabbits 

(3/dose group) occluded for 24 

hours. Skin cleaned after 24­

hour exposure and observed at 

24 and 72 hours. 

No irritation or edema at 24 or 72 

hours after exposure on abraded or 

intact skin. Primary skin irritation score 

is zero. 

Hart and 

Thornett 1975e 

MRID 

00049265 

L. dispar NPV Dermal application of 1 No mortality. Test compound did not Becker and 

(Bioserv Lot 33) g/animal to abraded and intact 

skin on approximately 10% of 

the body surface of New 

Zealand white rabbits (2 males 

and 2 females/dose group). 

Daily observations were made 

for 21 days after treatment. 

cause dermal toxicity or abnormal 

behavior in any of the animals 

throughout the 21-day observation 

period. No treatment-related gross 

pathological or histopathological 

effects were observed. 

Parke 1976b 

MRID 

00060694 

Becker et al. 

1976 

MRID 

00066101 

Appendix 1 - 5 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

OCULAR
 

P. dispar1 NPV Single virus exposure (eye Immunosuppressed mice were Shope et al. 

intact polyhedra irritation study, NOS) to 0.01 

g/animal polyhedra to adult 

mice [10 males (5 untreated 

and 5 immunosuppressed) and 

10 females (5 untreated and 5 

immunosuppressed)]. Positive 

controls treated with 

autoclaved polyhedra; negative 

controls treated with saline. 

All animals observed for 21 

days. 

selectively depleted of cell-mediated 

immune function by thymectomy and 

treatment with anti-lymphocyte serum 

(cytoxan administered i.p. at 300 

mg/kg/mouse). No eye irritation noted. 

1975 

MRID 

000606700 

P. dispar NPV 1 Administration of test 

compound in 0.8% saline at a 

rate of 40x10 polyhedra per 9 

animal to the left eye 

(conjunctival sac) (dose = 0.1 

mL per animal) of 5 male and 5 

female New Zealand white 

rabbits. Right eye served as 

control and received 0.1 mL of 

0.8% saline. Animals 

examined for injury at 24, 48, 

and 72 hours. 

No significant signs of irritation. Hart and 

Thornett 1975a 

MRID 

00049264 

Hart and 

Thronett 1975f 

MRID 

00060704 

[Final Report] 

Appendix 1 - 6 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

OCULAR (continued) 

P. dispar NPV 1 Administration of freed virus 

rods at a concentration 

corresponding to 40x109 

polyhedra/mL of 0.8% saline to 

the left eye (conjunctival sac) 

(dose = 0.1 mL per animal) of 

5 male and 5 female New 

Zealand white rabbits. Right 

eye served as control and 

received 0.1 mL of 0.8% saline. 

Animals examined for injury at 

24, 48, and 72 hours. 

No significant signs of irritation. Hart and 

Thornett 1975a 

MRID 

00049264 

Hart and 

Thronett 1975f 

MRID 

00060704 

[Final Report] 

“Gypsy Moth 

Virus” 

(6.48x1010/g) 

(Lot 35) 

described as 

light grey 

powder 

Administration of 50 mg of test 

compound in to one eye of each 

of 9 male New Zealand white 

(albino) rabbits, other eye of 

each rabbit served as control. 

After administration, treated 

eyes of 3 rabbits were washed 

with 20 mL of lukewarm 

dionized water 1 minute after 

treatment. The eyes of 3 other 

rabbits were washed 5 minutes 

after treatment and the eyes of 

the remaining 3 rabbits were 

not washed after treatment. 

One rabbit from the 1-minute wash 

died after 1 day, but the death was not 

considered to be treatment related. 

Clinical and necropsy findings showed 

the presence of diarrhea. 

Although early washing significantly 

lessened the discharge noted after 24 

hours in two rabbits, the investigators 

indicate that 20 mL of water was not 

sufficient to ensure that all the powdery 

test material as completely washed out 

of the treated eye. 

In short, the most significant finding 

was that of corneal opacity which did 

not always clear by day 14. 

In this study, “Gypsy Moth Virus” was 

judged to be a moderate eye irritant, 

and the test material was judged not to 

be corrosive. 

Gordon and 

Kinsel 1977 

MRID 

00068404 

Litton Bionetics 

1977 

Appendix 1 - 7 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

OCULAR (continued) 

“Insect Virus L. 

dispar NPV 

Bioserv Lot 

#33" 

Administration of 3 mg of test 

material in left eye of each of 

six New Zealand albino rabbits 

(weighing 2.0-2.5 kg). Right 

eyes served as controls. 

Rabbits were separated into 3 

groups with 2 animals/group: 1 

minute wash; 5 minute wash; 

and no wash. Treated eyes 

were scored at 24, 48, and 72 

hours and at 4 and 7 days after 

treatment. 

Slight conjunctival irritation was 

observed at 24 hours in the two rabbits 

in the “no wash” group, but the 

irritation cleared at 48 hours. No 

irritation was observed when the test 

material was washed out of the eyes at 

1 minute and 5 minutes. 

The irritation observed in the “no 

wash” group was not considered to be 

significant by the investigators. 

Becker and 

Parke 1976c 

MRID 

00068403 

Cannon Labs 

1976e 

L. dispar NPV 

(Bioserv Lot 

#33) 

Administration of 20 mL test 

compound to left eye of each of 

six New Zealand white rabbits 

(weight range of 2.0-2.5 kg). 

Right eyes served as controls. 

Treated eyes were observed 

and scored at 24, 48, and 72 

hours and 4 and 7 days after 

exposure. 

Positive reaction in all six rabbits at 24, 

48, and 72 hours and 4 and 7 days. 4/6 

animals had corneal involvement at 24, 

48, and 72 hours and 4 and 7 days. 

Conjunctival involvement was present 

at 24, 48, and 72 hours and 4 and 7 

days. 

Becker and 

Parke 1976a 

MRID 

00060696 

Gypchek TGAI New Zealand white rabbits, 6 In the unwashed eyes, the maximum Kuhn 1997a 

(Gypchek males and 6 females received average irritation score was 37.3 and MRID 

Lymantria undiluted test substance (0.1 was reached at 24 hours after exposure. 44354301 

dispar NPV) mL by volume) in the Gypchek TGAI in unwashed eyes was 

(Lot GR-14A) conjunctival sac of the right rated moderately irritating. Fluorescein 

wettable powder eye. Three treated eyes were 

each washed with deionized 

water for 1 minute, beginning 

30 seconds after treatment. 

Three treated eyes were left 

unwashed for 24 hours. 

staining, which was observed in all six 

treated unwashed eyes at 24 hours, was 

not observed in any eyes on day 17. 

In washed eyes, the maximum average 

irritation score was 5.3 and was 

reached at 1 hour after treatment. 

Gypchek TGAI in washed eyes was 

rated minimally irritating. Fluorescein 

staining was not observed in any of the 

treated washed eyes. 

Appendix 1 - 8 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

OCULAR (continued) 

Gypchek New Zealand white rabbits, 3 No positive effects were observed in Kuhn 1997b 

Solution 2X males and 3 females received a any of the treated eyes at any time MRID 

(Gypchek dose of 0.1 mL of the test during the study. 44354302 

Lymantria substance mixed with sterile 

dispar NPV) water in the conjunctival sac of Gypchek Solution 2X was rated non­

(Lot GR-14A) the right eye. All treated eyes irritating with a maximum irritation 

wettable powder were washed with deionized 

water for 1 minute immediately 

after recording the 24-hour 

observation. 

score of 0.0. 

See Section 3.1.11 for additional 

discussion. 

LDP 53 air dried 

sample 

(3.73x1010 

PIBs/g) 

Adult New Zealand albino 

rabbits (weighing between 2.0 

and 2.5 kg) 3 rabbits/test group, 

received 50 mg of “LDP 53" in 

the right eye with the untreated 

eye serving as a control. The 

test groups were treated as 

follows: Group I: 10 second 

wash; Group II: 1 minute wash; 

Group III: 5 minute wash; and 

Group IV: no wash. The 

treated eyes were observed and 

scored at 24, 48, and 72 hours 

as well as 4, 7, and 14 days 

after exposure. In addition, the 

treated and control eyes were 

swabbed before exposure and 

again at 4, 7, and 14 days after 

exposure for microbiological 

evaluation after a 48-hour 

incubation period. 

In Group I (10 second wash), one 

rabbit had eye irritation limited to 

conjunctival redness that lasted through 

day 4. 

In Group II (1 minute wash), all three 

rabbits exhibited conjunctival redness 

of grade 2 at 24 hours and grade 1 at 48 

hours. All irritation in this group 

cleared after 4 days. 

In Group III (5 minute wash) all three 

rabbits had corneal opacity of grade 1 

throughout the test. Iritis was present 

in two rabbits throughout the test and in 

one rabbit for 4 days. Conjunctival 

irritation was present in all rabbits 

throughout the test. 

In Group IV (no wash), all three rabbits 

had corneal opacity, but one of the 

cases cleared after 48 hours while the 

remaining two exhibited corneal 

opacity throughout the study. 

Iritis cleared after 72 hours in one 

rabbit, after 7 days in another rabbit, 

and continued in the third rabbit for the 

duration of the test. Conjunctival 

irritation persisted in all three rabbits 

through day 14. 

Imlay and 

Terrell 1978 

MRID 

00091124 

Cannon Labs 

1978 

Microbial evaluation revealed Staph 

epidermidis, Corynebacteria xerosis, 

Bacillus cereus, and Bacillius subtillis, 

but the findings were not considered to 

be significant. 

Appendix 1 - 9 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

INHALATION
 

P. dispar1 Sprague Dawley rats (9 males No mortality and no evidence of Thronett 1975 

nuclear PIB’s, and 9 females) exposed for 60 toxicity resulting from exposure. MRID 

Hamden minutes (heads only) to 0.028 00049266 

Standard to 0.81 mg LdNPV/L. 

Litton Bionetics 

1975d 

L. dispar NPV Rats (10 males, weighing 125­ No mortality and no treatment-related Brown 1976 

(Lot #33) 146 g) exposed to average effects on lung or trachea tissue. MRID 

analytical concentration of 6.12 00060695 

± 2.087 mg/L for 1 hour. Appendix to the study in the open 

Recovery period of 14 days. literature (Cannon Labs 1976c) 

indicates that alveolar thickening and a 

single finding of low grade pneumonitis 

were considered coincidental and not 

statistically significant by a pathologist 

at Cannon Labs who reviewed lung and 

trachea sections from the exposed rats. 

Cannon Labs 

1976c 

P. dispar1

 NPV Single virus dose exposure to Negative results. Shope et al. 

intact polyhedra (intranasal instillation) 0.02 

g/animal polyhedra to adult 

mice [10 males (5 untreated 

and 5 immunosuppressed) and 

10 females (5 untreated and 5 

immunosuppressed)]. 

Immunosuppressed mice were 

selectively depleted of cellmediated immune function by 

thymectomy and treatment with 

anit-lymphoctye serum 

(Cytoxan administered ip at 

300 mg/kg/mouse). Positive 

controls treated with 

autoclaved polyhedra; negative 

controls treated with saline. 

All animals observed for 21 

days. 

Serological results indicated that the 

animals with intact immune systems 

were exposed to NPV antigen, since 

positive reactions were apparent with 

autoclaved and non-autoclaved NPV 

preparations. Control (saline) exposure 

did not produce antibody responses. 

Investigators indicated that serology 

(characterization of P. dispar1

 NPV) 

and histopathology are incomplete. 

1975 

MRID 

000606700 

Appendix 1 - 10 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

INHALATION (continued) 

L. dispar NPV Rats, 10 males (initial weights Average persistence in lung tissue of USDA/FS 

(BioServ of 125-146 g) exposed to L. sacrificed animals: 19??g 

Lot#33; 6.6x1010 dispar NPV via inhalation for 1 MRID 

PIBs/g as dust) hour at a concentration of 6.12 

± 2.087 mg/L (= 4.04x10 ±8 

1.38x10 PIBs/L) for 1 hour 8 

and sacrificed 1, 7, or 14 days 

after exposure 

day 1 sacrifice: 95.96% (190/198) 

day 7 sacrifice: 68.0% (68/100) 

day 14 sacrifice: 18.09 % (36/199) 

00060701 

USDA/FS 

19??d 

MRID 

00066105 

USDA/FS 

1975? 

MRID 

00090598 

[most complete 

discussion of 

protocol and 

results] 

INTRAPERITONEAL
 

L-Dispar. Lot 33 10 Male ICR mice weighing 

18-25 g given single i.p. 

injection of 0.5 mL/mouse. To 

achieve dose, 50 mg of test 

material was suspended in 10 

mL of saline or 5 mg/mL. 

Thus, the dose was about 2.5 

mg LdNPV per mouse or about 

125 mg/kg bw using an average 

bw of 0.02 kg. 

No mortality and no adverse effects 

observed at 1,3, or 6 hours after 

treatment or at daily observations 

thereafter for 7 days. 

Terrell and 

Parke 1976c 

MRID 

00066103 

Terrell and 

Parke 1976d 

MRID 

00066109 

Appendix 1 - 11 
Appendix 1: Gypsy Moth NPV Toxicity in Mammals 

Product Species/Exposure Observations Reference 

OTHER
 

P. dispar1 NPV Single virus dose exposure Mice developed bacterial abscess Shope et al. 

intact polyhedra (footpad inoculation, not 

otherwise specified) to 0.02 

g/animal polyhedra to adult 

mice [10 males (5 untreated 

and 5 immunosuppressed) and 

10 females (5 untreated and 5 

immunosuppressed)]. Immunosuppressed mice were 

selectively depleted of cellmediated immune function by 

thymectomy and treatment with 

anit-lymphoctye serum 

(Cytoxan administered ip at 

300 mg/kg/mouse). Positive 

controls treated with 

autoclaved polyhedra; negative 

controls treated with saline. 

All animals observed for 21 

days. 

localized at the site of inoculation, but 

showed no other signs of toxicity. The 

study does not specify whether the 

incidence of bacterial infection was 

different between immunosuppressed 

and immunocompetent mice. 

1975 

MRID 

000606700 

1 P. dispar refers to Porthetria dispar, a former designation for the gypsy moth. 

Appendix 1 - 12 
Appendix 2: Toxicity of Gypsy Moth LdNPV to Birds 

Product Species/Exposure Observations Reference 

ORAL
 

Gypsy Moth Virus 

(Lot #33) (NOS) 

Mallard ducks (between 10 

and 15 days old) 10/dose 

group exposed to dietary 

concentrations of LdNPV 

ranging from 0.1x to 100x 

field usage (i.e., 1.04x10 , 6 

5.2x10 , 1.04x10 , 6 7 

1.04x10 , 1.04x10 PIBs/g 8 9 

of feed). Controls were not 

exposed to virus in the diet. 

No signs of abnormal behavior such 

as decreased locomotor activity, 

feather erection, or loss of righting 

reflex. No mortality except for one 

death at the 1x level that was not 

considered to be treatment related. 

Roberts and 

Wineholt 1976 

MRID 00068410 

NPV of the gypsy 

moth 

Gypsy moth avian predators 

(6 black-capped chickadees, 

Parus atricapillus, and 9 

house sparrows, Passer 

domesticus) fed LdNPVinfected 4th instar gypsy 

moth larvae on day 1 and on 

alternate days for 3 weeks. 

Each infected larva 

contained from 3.3x107

 to 

2.1x10 PIB. During the test 8 

period, each chickadee ate 

70-80 infected larvae (from 

2.3x10 to 1.7x10 PIB) 9 10 

and each treated sparrow ate 

90-100 infected larvae 

(from 3.0x109

 to 2.1x1010 

PIB). 

No signs of disease were observed 

in the birds during the test period; 

body weight and results of 

histological examination of organs 

of treated birds indicated that 

LdNPV exposure caused no 

apparent short-term adverse effects. 

Podgwaite and 

Galipeau 1978 

MRID 00134318 

Appendix 2-1 
Appendix 2: Toxicity of Gypsy Moth LdNPV to Birds 

Product Species/Exposure Observations Reference 

FIELD STUDIES
 

NPV molasses-based 

formulation 

containing “k” rotor 

purified polyhedral 

inclusion bodies 

(PIBs) (0.25 gal 

Cargill insecticide 

base; 6.0 oz Chevron 

spray sticker; 1.0 lb 

IMC 900001; 1.75 

gal water) 

Resident songbird 

populations, caged quail 

(Colinus virginianus) in 

woodland plots in central 

mountain region of 

Pennsylvania treated with 

two aerial applications 

(May 28 and June 2, 1975) 

of LdNPV at the rate of 

2.5x1012 PIBs/ha (18.7 

L/ha). Applications were 

made with 450 hp Grumman 

AgCat aircraft equipped 

with 6 Beecomist nozzles. 

Elevations of treated plots 

ranged from 1500 to 1800 ft 

(550-650 m) above sea level 

and supported 300-2000 

egg masses/acre (750­

5000/ha). Untreated plots 

were used as a negative 

control. 

No significant differences in 

population trends between treated 

and control plots at either 1 or 2 

months after LdNPV applications. 

LdNPV treatment had no adverse 

effects on the resident song birds or 

caged quail. In fact, it appeared 

that the LdNPV application, by 

reducing defoliation, helped to 

maintain significantly higher 

densities of the yellow throat 

warblers; once bird species which 

utilizes a niche close to the ground. 

Investigators concluded that aerial 

application of LdNPV at the rates 

used in this study had no adverse 

effects on birds that use gypsy 

moths as a food source or birds that 

contact the virus from the LdNPV 

spray, spray residue, or the dying 

larvae. 

Lautenschlager et 

al. 1976b 

MRID 00066108 

Lautenschlager et 

al. 1978b 

MRID 00134316 

[This is an 

abstract of the 

Lautenschlager et 

al. 1976b study 

that was 

submitted 

separately to 

EPA] 

Lautenschlager 

and Podgwaite 

1979b 

NPV formulation 

containing a 

commercial adjuvant 

and “k” rotor purified 

PIBs (1.0 gal Sandoz 

Virus Adjuvant; 1.0 

gal water). 

Resident songbird 

populations caged quail 

(Colinus virginianus) in 

woodland plots in central 

mountain region of 

Pennsylvania treated with 

two aerial applications 

(May 28 and June 2, 1975) 

of LdNPV at the rate of 

2.5x1012 PIBs/ha (18.7 

L/ha). Applications were 

made with 450 hp Grumman 

AgCat aircraft equipped 

with 6 Beecomist nozzles. 

Elevations of treated plots 

ranged from 1500 to 1800 ft 

(550-650 m) above sea level 

and supported 300-2000 

egg masses/acre (750­

5000/ha Untreated plots 

were used as a negative 

control. 

No significant differences in 

population trends between treated 

and control plots at either 1 or 2 

months after LdNPV applications. 

LdNPV treatment had no adverse 

effects on the resident song birds or 

caged quail. In fact, it appeared 

that the NPV application, by 

reducing defoliation, helped to 

maintain significantly higher 

densities of the yellow throat 

warblers; once bird species which 

utilizes a niche close to the ground. 

Investigators conclude that aerial 

application of LdNPV at the rates 

used in this study had no adverse 

effects on birds that use gypsy 

moths as a food source or birds that 

contact the virus from the LdNPV 

spray, spray residue, or the dying 

larvae. 

Lautenschlager et 

al. 1976b 

MRID 00066108 

[This is the same 

study as above 

but using a 

different 

formulation of 

LdNPV] 

Lautenschlager et 

al. 1978b 

MRID 00134316 

Lautenschlager 

and Podgwaite 

1979b 

Appendix 2-2 
Appendix 3: Gypsy Moth NPV Toxicity in Nontarget Terrestrial Insects 

Product Species/Exposure Observations Reference 

LdNPV 46 species of nontarget No statistically significant mortality, Barber et al. 1993 

(aqueous Lepidoptera exposed to compared with controls; 0.0% infection in 

suspension) four successive 24- to 48­

hour doses of 3x10 PIBs 4 

in 2μL applied to pellets 

of artificial diet or 

isolated surfaces of 

foliage 

all treated species. 

LdNPV Adult fly, Cyrtophleba No statistically significant motality, Barber et al. 1993 

(aqueous coquilletti Aldr. exposed compared with controls; 0.0% infection. 

suspension) to single dose of 12x105 

PIBs in 2μL of 30% 

sucrose solution. Those 

that completely consumed 

the dose were transferred 

to appropriate 

maintenance conditions 

for 7-10 days and then 

frozen. 

LdNPV 

(aqueous 

suspension) 

Adult male bees, 

Megachile rotundata 

(Fabr). exposed to single 

dose of 12x10 PIBs in 5 

2μL of 30% sucrose 

solution. Those that 

completely consumed the 

dose were transferred to 

appropriate maintenance 

conditions for 7-10 days 

and then frozen. 

No statistically significant motality, 

compared with controls; 0.0% infection. 

Barber et al. 1993 

Gypsy Moth 

NPV Porthetria 

dispar (L). 

Adult honey bees exposed 

to estimated dose of 

1x10 polyhedra in 6 

sucrose solution 

No indication of detrimental effects 

resulting from exposure to test substance. 

Cantwell et al. 

1972 

Gypsy Moth Honeybee (Apis No differences were observed between Knox 1970 

NPV melliferai) in observation treated and untreated bee colonies 

(Porthetria hives fed 10x109 

dispar) polyhedra mixed with 200 

mL sucrose solution 

(sugar-water 1:1) (total 

dose/hive) over 4-month 

period. 

Gypchek Application at a rate of 

8x1010 PIB/ha on ant 

communities. Pitfall traps 

operated for 45 weeks 

during summers of 1995­

1997 in George 

Washington national 

Forest, Augusta County, 

VA and Monongahela 

National Forest in 

Pocahontas County, WV. 

Ants representing 17 genera and 31 

species were collected, indicating that 

species richness, diversity, abundance, and 

species composition were not adversely 

affected by treatment. 

Wang et al. 2000 

Appendix 3-1 
Appendix 4: Toxicity of NPV to Aquatic Invertebrates 

Product Species/Exposure Observations Reference 

NPV containing 1.7x1011 Daphnia (D. magna), 15, #24 Treatment had no significant Streams 1976 

polyhedra/g and some hours old exposed to test effect on either survival MRID 

bacterial impurities. concentration of 250 

polyhedra/g. Virus was added 

initially and anew every 2 days. 

Complete experiment was 

replicated 3x (conducted several 

weeks apart in time). Surviving, 

mature Daphnia produced 

young, which were counted. 

(p>0.05) or reproduction 

(p>0.05). 

00068408 

NPV containing 1.7x1011 Daphnia (D. magna) surviving The average mortality rate Streams 1976 

polyhedra/g and some the acute toxicity study were for gypsy moth larvae fed MRID 

bacterial impurities. randomly frozen for bioassay or 

transferred to a virus-free 

medium with samples taken at 6­

to 12-hour intervals. The 

purpose of the bioassays was to 

determine whether NPV could 

be detected in a apparently 

healthy Daphnia reared in water 

with a high concentration of 

polyhedra and , if so, how soon 

the NPV disappeared from 

Daphnia when placed in a virus 

free medium. 

Daphnia reared in virustreated water was similar to 

that of larvae fed Daphnia 

reared in virus free water 

(2.2% vs.3.1%); the average 

percent mortality rate for 

gypsy moth larvae fed a 

sterile diet was 0.5%. 

Mortality rate was not 

affected when gypsy moth 

larvae were fed Daphnia 

removed from virus-treated 

medium and reared in virus 

free medium for up to 48 

hours. 

Daphnia did not accumulate 

gypsy moth NPV under the 

test conditions. 

00068408 

NPV containing 1.7x1011 Backswimmers (Notonecta No significant effects of Streams 1976 

polyhedra/g and some undulata), newly hatched NPV on N. undulata were MRID 

bacterial impurities. nymphs reared for the first 2 

instars in virus-free water after 

which time NPV at a 

concentration of 250 

polyhedra/mL was added to the 

containers. The treated 

backswimmers were fed live, 

virus-treated Daphnia. The 

Daphnia fed to the treated 

backswimmers were reared in 

water containing virus at a 

concentration of 250 

polyhedra/mL and the treated 

water was renewed about 

3x/week. 

observed with regard to 

survival or reproduction. 

Data are presented in Table 3 

of the study. 

Bioassay results are recorded 

in Table 7 of the study and 

indicate that N. undulata 

reared in water with 250 

polyhedra/mL of gypsy moth 

NPV or fed Daphnia reared 

in similar concentrations do 

not accumulate the NPV 

virus. 

00068408 

Appendix 4-1 
Appendix 4: Toxicity of NPV to Aquatic Invertebrates 

Product Species/Exposure Observations Reference 

NPV containing 1.7x1011 Midge (Chironomus thummi), No significant difference Streams 1976 

polyhedra/g and some newly hatched larvae reared to (p>0.05) in survival of MRID 

bacterial impurities. pupation in containers in which 

NPV was mixed with the water 

and the food at a concentration 

of 250 polyhedra/mL. Emerging 

adults were set up in screened 

breeding cages for 1 week to 

obtain reproduction and to check 

on the viability of any eggs 

produced. 

treated midge, compared 

with controls; developmental 

time was identical in treated 

and in untreated replicates; 

and reproduction by adults 

reared from treated replicates 

was similar to that observed 

in controls (all egg masses 

were fertile). 

00068408 

NPV containing 1.7x1011 Water boatmen (adult No significant difference in Streams 1976 

polyhedra/g and some Hesperocorixa interrupta survival of either species in MRID 

bacterial impurities. [n=10/replicate] and Sigara 

gordita n=20/replicate]) exposed 

to NPV at a concentration in 

water of 250 polyhedra/mL for 4 

weeks. 

among treated and control 

adults and no apparent 

adverse effects on 

reproduction were observed 

in Sigara, which produced 

eggs, many of which hatched 

before the end of the study. 

Results of the bioassay 

indicate that the water 

boatmen did not accumulate 

NPV under the conditions of 

the study. 

00068408 

Appendix 4-2 
Appendix H 

Disparlure 

Risk Assessment 

Figure H-1. Female gypsy moth pupae were gathered in Massachusetts in 1948 in 

order to obtain sex attractant for trapping programs. 

SERA TR 06-52-07-02d 

Control/Eradication Agents for the Gypsy Moth -

Human Health and Ecological Risk Assessment for 

Disparlure (a.i.) and Disrupt II formulation 

– FINAL REPORT

 Prepared for: 

USDA, Forest Service 

Forest Health Protection 

USDA Forest Service Contract No: AG-3187-C-06-0010
 

USDA Order No. AG-43ZP-D-06-0021
 

SERA Task No. 52-07

 Submitted to: 

Paul Mistretta, COR
 

Kay A. Matthews, Contracting Officer
 

USDA/Forest Service, Southern Region 

1720 Peachtree RD, NW 

Atlanta, Georgia 30309 

Prepared by Patrick Durkin 

 Submitted by: 

Syracuse Environmental Research Associates, Inc. 

5100 Highbridge St., 42C 

Fayetteville, New York 13066-0950 

 Telephone: (315) 637-9560 

Fax: (315) 637-0445 

E-Mail: SERA_INC@msn.com

 Home Page: www.sera-inc.com
 

August 28, 2006 (Risk Assessment) 

July 20, September 10, and September 19, 2007 (Editorial Corrections) 
 

 

 

 

 

PREFACE
 

This document is a revision to a risk assessment that was originally prepared by Syracuse 

Environmental Research Associates, Inc. (SERA Inc.) under GSA Contract No. GS-10F-0082F, 

USDA Forest Service BPA: WO-01-3187-0150, USDA Purchase Order No.: 43-3187-1-0269. 

The SERA documented was prepared by Drs. Patrick R. Durkin (SERA Inc.) and Julie 

Klotzbach (currently with Syracuse Research Corporation). The SERA document was submitted 

to the USDA Forest Service as Control/Eradication Agents for the Gypsy Moth - Human Health 

and Ecological Risk Assessment for Disparlure (a.i.) - FINAL REPORT, SERA TR 04-43-05­

04b, reported dated August 27, 2004. As indicated in the title, SERA TR 04-43-05-04b covered 

only the active ingredient – i.e., disparlure – and did not address the formulation of disparlure in 

Disrupt II flakes. The original SERA document was reviewed by Dr. Rolf Hartung (Univ. 

Michigan, retired) and by USDA/Forest Service personnel: Dr. Paul Mistretta, Mr. Joseph Cook, 

and Ms. Donna Leonard. 

Under USDA Order No. AG-43ZP-D-06-0015, USDA Forest Service Contract No: AG-3187-C06-0010, SERA revised the above report to include Disrupt II flakes. The subsequent revision 

(SERA TR 06-52-02-01a) was submitted to the USDA on June 30, 2006). This revision was 

based on new information provided by the USDA/Forest Service. The listing below indicates the 

specific references that were added to the June 30, 2006 revised risk assessment concerning 

Disrupt II: 

Hercon Environmental. 2006a. Hercon Disrupt II Product Label. Copy courtesy of Donna Leonard, 

USDA Forest Service, Forest Health Protection, PO Box 2680, Asheville, NC 28802. e-mail: 

dleonard@fs.fed.us. Received June 27, 2006. 

Hercon Environmental. 2006b. Hercon Disrupt II Material Safety Data Sheet. Copy courtesy of Priscilla 

MacLean, Product Development Manager, Hercon Environmental, P.O. Box 435, Emigsville PA, 17318. 

e-mail: pmaclean@herconenviron.com. Received June 27, 2006. 

Leonard D. 2006a. Comments on Application Rates for Disparlure in STS (Slow-The-Spread) Programs. 

Comments by Donna Leonard, USDA Forest Service, Forest Health Protection, Asheville, NC. Comments 

received via email from dleonard@fs.fed.us on June 27, 2006. 

Leonard D. 2006b. Comments on The Use of Disparlure in STS (Slow-The-Spread) Programs. Comments 

by Donna Leonard, USDA Forest Service, Forest Health Protection, Asheville, NC. Comments received 

via email from dleonard@fs.fed.us on June 27, 2006. 

MacLean P. 2006. Comments on Inerts in Disrupt II, Product Development Manager, Hercon 

Environmental, P.O. Box 435, Emigsville PA, 17318. e-mail: pmaclean@herconenviron.com. Received 

June 27, 2006. 

Palmer SJ; Krueger HO. 2006a. SF 2003 and SF 2005: A 48-Hour Static-Renewal Acute Toxicity Test 

with the Cladoceran (Daphnia magna). Wildlife International, Ltd. Project Number: 6 L4a- 102. Study 

completion date: Jan. 12, 2006. Copy courtesy of Paul Mistretta, USDA/FS. 

Palmer SJ; Krueger HO. 2006b. MF 2003 and MF 2005: A 48-Hour Static-Renewal Acute Toxicity Test 

with the Cladoceran (Daphnia magna). Wildlife International, Ltd. Project Number: 6 L4a- 101. Study 

completion date: Jan. 12, 2006. Copy courtesy of Paul Mistretta, USDA/FS. 

ii 
 

 

 

Because of limitations in the available toxicity data on disparlure and Disrupt II, more extensive 

use has been made of quantitative structure activity relationships (QSAR) and the following 

additional references (not specific to disparlure) have been added: 

Bintein S, Devillers J, and Karcher W. 1993. Nonlinear dependence of fish bioconcentration on noctanol/water partition coefficient. SAR QSAR Environ Res. 1(1):29-39. 

Clements RG, Nabholz JV, and Zeeman M. 1996. Estimating Toxicity of Industrial Chemicals to Aquatic 

Organisms Using Structure-activity Relationships. Environmental Effects Branch, Health and 

Environmental Review Division, Office of Pollution Prevention and Toxics, U.S. Environmental Protection 

Agency. Report dated August 30, 1996. 

Jeppsson R. 1975. Parabolic Relationship between Lipophilicity and Biological Activity of Aliphatic 

Hydrocarbons, Ethers and Ketones after Intravenous Injections of Emulsion Formulations into Mice. Acta 

Pharmacol. Et Toxicol. 37: 56-64. 

U.S. EPA/OPPT (U.S. Environmental Protection Agency/Office of Pollution Prevention and Toxics). 

2000. On-Line EPI Suite User's Guide, Version 3.12. Developed by the EPA's Office of Pollution 

Prevention Toxics and Syracuse Research Corporation (SRC). Available at: 

http://www.epa.gov/opptintr/exposure/docs/episuite.htm 

SERA TR 06-52-07-01a was then submitted based on comments from Forest Service and APHIS 

personnel. A consolidation of comments was prepared by Joe Cook (USDA/FS). This was the 

primary source for the current revisions. Comments from various Forest Service personnel were 

provided and consulted as needed, including comments from Hank Appleton, Jesus Cota, John 

Kyhl, and Donna Leonard. A PDF copy of the risk assessment with annotations from APHIS 

personnel was also consulted. Lastly, an unpublished synopsis of the following study was 

provided by Donna Leonard, reviewed and incorporated into this risk assessment as appropriate: 

Thwaits BF; Sorensen PW. 2005. Olfactory sensitivity of rainbow trout to 

racemic disparlure. Unpublished synopsis dated April 1, 2005. Copy courtesy of 

Donna Leonard, USDA/Forest Service. 2 pp. 

The current report, SERA TR 06-52-07-02a, is based on editorial comments from Joe Cook, 

some additional comments on formulations from Donna Leonard (cited as Leonard 2006e), and 

internal review. There are no substantial technical changes from SERA TR 06-52-07-01a. 

iii 
 

TABLE OF CONTENTS
 

PREFACE...................................................................................................................................... iii 

TABLE OF CONTENTS............................................................................................................... iv 
 

LIST OF APPENDICES................................................................................................................ vi 

LIST OF TABLES......................................................................................................................... vi 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS .............................................................. viii 

COMMON UNIT CONVERSIONS AND ABBREVIATIONS................................................... ix 

CONVERSION OF SCIENTIFIC NOTATION .............................................................................x 

EXECUTIVE SUMMARY ........................................................................................................... xi 

1. INTRODUCTION .................................................................................................................. 1-1 

2. PROGRAM DESCRIPTION.................................................................................................. 2-1 

2.1. OVERVIEW ............................................................................................................ 2-1 

2.2. CHEMICAL DESCRIPTION.................................................................................. 2-2 

2.3. APPLICATION METHODS AND RATES............................................................ 2-3 

2.4. USE STATISTICS................................................................................................... 2-3 

3. HUMAN HEALTH RISK ASSESSMENT............................................................................ 3-1 

3.1. HAZARD IDENTIFICATION................................................................................ 3-1 

3.1.1. Overview................................................................................................... 3-1 

3.1.2. Mechanism of Action................................................................................ 3-1 

3.1.3. Kinetics and Metabolism. ......................................................................... 3-1
 

3.1.4. Acute Oral Toxicity. ................................................................................. 3-2 

3.1.5. Subchronic or Chronic Systemic Toxic Effects........................................ 3-2 

3.1.6. Effects on Nervous System....................................................................... 3-3
 

3.1.7. Effects on Immune System. ...................................................................... 3-3 

3.1.8. Effects on Endocrine System.................................................................... 3-3 

3.1.9. Reproductive and Teratogenic Effects...................................................... 3-3 

3.1.10. Carcinogenicity and Mutagenicity.......................................................... 3-3 

3.1.11. Irritation and Sensitization (Effects on the Skin and Eyes). ................... 3-3 

3.1.12. Systemic Toxic Effects from Dermal Exposure. .................................... 3-4 

3.1.13. Inhalation Exposure. ............................................................................... 3-4 

3.1.14. Inerts and Adjuvants. .............................................................................. 3-4 

3.1.15. Impurities and Metabolites. .................................................................... 3-5
 

3.1.16. Toxicological Interactions ...................................................................... 3-5 

iv 
TABLE OF CONTENTS (continued) 

3.2. EXPOSURE ASSESSMENT .................................................................................. 3-6 

3.2.1. Overview.................................................................................................... 3-6 

3.2.2. Dermal Exposure ....................................................................................... 3-6 

3.2.3. Inhalation Exposure . ................................................................................. 3-6 

3.2.4. Oral Exposure . .......................................................................................... 3-7
 

3.3. DOSE-RESPONSE ASSESSMENT ....................................................................... 3-8
 

3.4. RISK CHARACTERIZATION............................................................................. ..3-9
 

3.4.1. Overview................................................................................................... 3-9
 

3.4.2. Workers and the General Public ............................................................... 3-9
 

3.4.3. Sensitive Subgroups.................................................................................. 3-9
 

3.4.4. Cumulative Effects.................................................................................. 3-10
 

3.4.5. Connected Actions .................................................................................. 3-10
 

4. ECOLOGICAL RISK ASSESSMENT .................................................................................. 4-1 

4.1. HAZARD IDENTIFICATION................................................................................ 4-1 

4.1.1. Overview................................................................................................... 4-1 

4.1.2. Toxicity to Terrestrial Organisms............................................................. 4-1 

4.1.2.1. Mammals.................................................................................... 4-1 

4.1.2.2. Birds........................................................................................... 4-2 

4.1.2.3. Terrestrial Invertebrates............................................................. 4-2 

4.1.2.4. Terrestrial Plants (Macrophytes)................................................ 4-2
 

4.1.2.5. Terrestrial Microorganisms........................................................ 4-3 

4.1.3. Aquatic Organisms.................................................................................... 4-3 

4.1.3.1. Fish............................................................................................. 4-3 

4.1.3.2. Amphibians................................................................................ 4-4 

4.1.3.3. Aquatic Invertebrates................................................................. 4-4 

4.1.3.4. Aquatic Plants ............................................................................ 4-8 

4.1.3.5. Other Aquatic Microorganisms.................................................. 4-8 

4.2. EXPOSURE ASSESSMENT .................................................................................. 4-9 

4.2.1. Overview................................................................................................... 4-9 

4.2.2. Exposure of Aquatic Animals................................................................... 4-9 

4.3. DOSE-RESPONSE ASSESSMENT ..................................................................... 4-11 

4.3.1. Overview................................................................................................. 4-11 

4.3.2. Fish.......................................................................................................... 4-11 

4.3.3. Aquatic Invertebrates.............................................................................. 4-11 

v 
 

TABLE OF CONTENTS (continued) 

4.4. RISK CHARACTERIZATION............................................................................. 4-14
 

4.4.1. Overview................................................................................................. 4-14
 

4.4.2. Terrestrial Species................................................................................... 4-14
 

4.4.3. Fish........................................................................................................... 4-15
 

4.4.4. Aquatic Invertebrates............................................................................... 4-15
 

5. REFERENCES ....................................................................................................................... 5-1 

LIST OF APPENDICES 

Appendix 1: Acute toxicity of disparlure to experimental mammals 

Appendix 2: Toxicity of disparlure to birds 

Appendix 3: Toxicity of disparlure aquatic species 

Appendix 4: EPI Suite Output for Disparlure 

LIST OF TABLES 

Table 2-1: Identification and physical/chemical properties of Disparlure................................ Tables-1 

Table 2-2: Use of disparlure by USDA to control the North American Gypsy Moth

 from 1995 to 2003 .................................................................................................. Tables-2 

Table 3-1: Summary of acute toxicity data of Disparlure in mammals ................................... Tables-3 

Table 4-1: Summary of acute toxicity data of Disparlure in avian 

 and aquatic species ................................................................................................. Tables-4 

Table 4-2: Summary of QSAR Toxicity Estimates for Disparlure to Aquatic Species 

and Algorithms for Estimating the Toxicity of Mono-Epoxide 

Compounds to Aquatic Species Developed by Clements et Al. (1996) ................ Tables-5 

NOTE: Tables are placed after Section 5, References. 

Workbook 

Disparlure: Simplified EXCEL Worksheets for Calculating Risks to Small Aquatic Invertebrates 

SERA EXWS 06-52-07-01a. Worksheet dated August 25, 2006. 

Located at: http://www.fs.fed.us/foresthealth/pesticide/risk.shtml 

vi 
 

 

 

 

 

 

 

 

 

 

 

 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

AEL adverse-effect level 

AGM Asian Gypsy Moth 

a.i. active ingredient 

BCF bioconcentration factor 

bw body weight 

CBI confidential business information 

cm centimeter 

CNS central nervous system 

ECx concentration causing X% inhibition of a process 

EC25 concentration causing 25% inhibition of a process 

EC50 concentration causing 50% inhibition of a process 

F female 

FH Forest Health 

FIFRA Federal Insecticide, Fungicide and Rodenticide Act 

FQPA Food Quality Protection Act 

g gram 

ha hectare 

HQ hazard quotient 

IARC International Agency for Research on Cancer 

IRIS Integrated Risk Information System 

ka absorption coefficient 

ke elimination coefficient 

kg kilogram 

Ko/c organic carbon partition coefficient 

Ko/w octanol-water partition coefficient 

Kp skin permeability coefficient 

L liter 

lb pound 

LC50 lethal concentration, 50% kill 

LD50 lethal dose, 50% kill 

LOAEL lowest-observed-adverse-effect level 

m meter 

M male 

vii 
 

 

 

 

 

 

 

 

 

 

 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS (continued) 

mg milligram 

mg/kg/day milligrams of agent per kilogram of body weight per day 

mL milliliter 

mM millimole 

MRID Master Record Identification Number 

MSDS material safety data sheet 

MW molecular weight 

NAGM North American Gypsy Moth 

NOAEL no-observed-adverse-effect level 

NOEC no-observed-effect concentration 

NOEL no-observed-effect level 

NOS not otherwise specified 

NRC National Research Council 

NTP National Toxicology Program 

OPP Office of Pesticide Programs 

OPPTS Office of Pesticide Planning and Toxic Substances 

ppm parts per million (used in expressing dietary concentrations only) 

QSAR quantitative structure activity relationship 

RfD reference dose 

SERA Syracuse Environmental Research Associates 

SRC Syracuse Research Corporation 

UF uncertainty factor 

U.S. United States 

USDA U.S. Department of Agriculture 

U.S. EPA U.S. Environmental Protection Agency 

WHO World Health Organization 

μ micron 

< greater than 

$ greater than or equal to 

< less than 

# less than or equal to 

= equal to 

• approximately equal to 

- approximately 

viii 
COMMON UNIT CONVERSIONS AND ABBREVIATIONS
 

To convert ... Into ... Multiply by ...
 

centigrade Fahrenheit 1.8 �C + 32 

atmospheres millimeters of mercury 760 

centimeters inches 0.3937 

acres hectares (ha) 0.4047 

square meters (m2 acres ) 4,047 

cubic meters (m3) liters (L) 1,000 

feet per second (ft/sec) miles/hour (mi/hr) 0.6818 

Fahrenheit centigrade 5/9 (�F-32)
 

gallons per acre (gal/acre) liters per hectare (L/ha) 9.34

gallons (gal) liters (L) 3.785 

grams (g) pounds, (oz) 0.002205 

grams (g) ounces, (oz) 0.03527 

inches (in) centimeters (cm) 2.540 

hectares (ha) acres 2.471 

kilograms (kg) pounds, (lb) 2.2046 

kilograms (kg) ounces, (oz) 35.274 

kilometers (km) miles (mi) 0.6214 

kilograms per hectare (hg/ha) pounds per acre (lb/acre) 0.892 

liters (L) gallons (gal) 0.2642 

liters (L) cubic centimeters (cm3) 1,000 

miles (mi) kilometers (km) 1.609 

liters (L) ounces, fluid (oz) 33.814 

milligrams (mg) ounces (oz) 0.000035 

miles per hour (mi/hr) cm/sec 44.70
 

ounces (oz) grams (g) 28.3495 

meters (m) feet 3.281 

ounces per acre (oz/acre) kilograms per hectare (kg/ha) 0.0701 

ounces per acre (oz/acre) grams per hectare (g/ha) 70.1 

pounds (lb) grams (g) 453.6 

ounces fluid cubic centimeters (cm3) 29.5735 

pounds per acre (lb/acre) kilograms per hectare (kg/ha) 1.121 

pounds (lb) kilograms (kg) 0.4536 

μg/square centimeter (μg/cm2 pounds per acre (lb/acre) ) 11.21 

pounds per acre (lb/acre) mg/square meter (mg/m2) 112.1 

square centimeters (cm2) square inches (in2) 0.155 

pounds per gallon (lb/gal) grams per liter (g/L) 119.8 

square meters (m2) square centimeters (cm2) 10,000 

square centimeters (cm2) square meters (m2) 0.0001 

yards meters 0.9144 

Note: All references to pounds and ounces refer to avoirdupois weights unless otherwise specified. 

ix 
CONVERSION OF SCIENTIFIC NOTATION
 

Scientific Decimal Verbal 

Notation Equivalent Expression 

1 x 10-10 0.0000000001 One in ten billion 

1 x 10-8 0.00000001 One in one hundred million 

1 x 10-9 0.000000001 One in one billion 

1 x 10-7 0.0000001 One in ten million 

1 x 10-6 0.000001 One in one million 

1 x 10-5 0.00001 One in one hundred thousand 

1 x 10-4 0.0001 One in ten thousand 

1 x 10-3 0.001 One in one thousand 

1 x 10-2 0.01 One in one hundred 

1 x 10-1 0.1 One in ten 

1 x 100 1 One 

1 x 101 10 Ten 

1 x 102 100 One hundred 

1 x 103 1,000 One thousand 

1 x 104 10,000 Ten thousand 

1 x 105 100,000 One hundred thousand 

1 x 106 1,000,000 One million
 

1 x 107 10,000,000 Ten million 

1 x 108 100,000,000 One hundred million 

1 x 109 1,000,000,000 One billion 

1 x 1010 10,000,000,000 Ten billion 

x 
EXECUTIVE SUMMARY 

OVERVIEW 

Disparlure is a naturally occurring insect pheromone used to disrupt mating of gypsy moths by 

confusing male moths. Disparlure is also used as an attractant in traps. There are limited data 

available on the toxicity of disparlure. Only a small number of acute exposure studies have been 

conducted; no chronic toxicity studies in any species were identified in the available literature. 

Based on the results of the available data, the toxicity profile of disparlure in terrestrial animals 

does not suggest that disparlure is likely to cause adverse effects at plausible levels of exposure. 

Similarly, disparlure is not likely to cause any toxic effects in aquatic species at the limit of 

solubility of disparlure in water. Thus, under normal conditions of exposure, no hazard to 

aquatic species can be identified. In cases of an accidental application of disparlure to a small 

body of standing water, such as a pond, no effects are likely in fish. An accidental application or 

some other similar event such as an accidental spill could lead to an insoluble film of disparlure 

at the air-water interface of a standing body of water. This could result in some small 

invertebrates becoming trapped in the film of disparlure. While the entrapment of daphnids has 

been observed in laboratory studies of both disparlure and Disrupt II formulations, the likelihood 

of this occurring in the field to an extent that detectable effects would be observed is difficult to 

determine. The formation of a film that could trap small invertebrates in rapidly moving bodies 

of water does not seem plausible. 

PROGRAM DESCRIPTION 

Disparlure is a naturally occurring insect pheromone (attractant) synthesized by the female gypsy 

moth to attract the male gypsy moth. Disparlure can take two enantiomer forms, referred to as 

(+)disparlure and (-)disparlure. Enantiomers are mirror-image molecules with identical gross 

structures. The (+)enantiomer is the form produced by the female gypsy moth and is the only 

form that is biologically active as an attractant. In gypsy moth programs, two forms of disparlure 

are used: the (+)enantiomer and the racemic mixture, a 50:50 blend of the (+)enantiomer and (­

)enantiomer. Racemic disparlure is used as a control agent. It is broadcast over relatively large 

areas and disrupts mating by confusing male moths – i.e., the male moth has difficulty in 

locating the female moth. 

Disparlure is always formulated in a slow release matrix and several different formulations have 

been tested including polyvinyl chloride flakes, microcapsules, and polyvinyl chloride twine. 

Disrupt II, a formulation of disparlure in polyvinylchloride flakes, has been used by the USDA 

Forest Service for many years. The specific formulation has evolved over time. This risk 

assessment considers the available information both on the current and some previous Disrupt II 

formulations. As noted by Leonard (2006e), it is possible that the U.S. EPA will require 

different labels for the two different Disrupt formulations, with the previous formulation 

designated as Disrupt II and the newer formulation designated as Disrupt III. Because this 

decision has not yet been made, this risk assessment will refer to the older Disrupt formulation as 

standard flakes and the newer Disrupt formulation as modified flakes. These designations are 

discussed further in Section 4.1.3.3 in terms of differences in toxicity to Daphnia. 

xi 
Since 1995, the use of disparlure in programs intended to slow the spread of gypsy moths has 

increased over 250-fold, from 2,448 acres treated in 1995 to a maximum of 647,394 acres treated 

in 2003. The (+)enantiomer of disparlure is used as an attractant or bait in two types of traps: 

milk carton traps that also contain DDVP and delta traps that do not contain an insecticide. 

These traps are used to monitor existing (endemic) populations and detect new infestations. 

HUMAN HEALTH RISK ASSESSMENT 

Hazard Identification – Insect pheromones are generally regarded as nontoxic to mammals and 

these pheromones are commonly employed in very low environmental concentrations. 

Consequently, U.S. EPA requires less rigorous testing of these products than is required of 

insecticides. Except for some standard acute toxicity studies in laboratory mammals, few data 

are available regarding the toxicity of disparlure to terrestrial species. Results of acute exposure 

studies for oral, dermal, ocular and inhalation exposure to disparlure show no indication of 

adverse effects. The LD50 of a single dose administered to rats by gavage exceeds 34,600 mg/kg. 

With the exception of one acute gavage study in rats using the 50:50 racemic mix, none of the 

toxicity studies specified whether the 50:50 racemic mix or the (+)enantiomer was tested. Based 

on the results of studies on disparlure itself (i.e., the active ingredient), acute exposure to 

disparlure has very low toxicity in mammals. No studies investigating the effects of chronic 

exposure of mammals to disparlure or studies investigating the effects of disparlure on the 

nervous system, immune system, reproductive system or endocrine system were identified. The 

carcinogenic potential of disparlure has not been assessed. In a single study on mutagenicity, 

there was no indication that disparlure is mutagenic. There is no information available regarding 

the kinetics and metabolism of disparlure in mammals. The kinetics of absorption of disparlure 

following dermal, oral or inhalation exposure are not documented in the available literature. A 

case report of an accidental exposure indicates that disparlure may persist in humans for years. 

Exposure Assessment – For both occupational exposure of workers and accidental exposure of 

the general public, exposure to disparlure may involve multiple routes of exposure (i.e., oral, 

dermal, and inhalation). Nonetheless, dermal exposure is generally most likely to be the 

predominant route. While exposure scenarios can be developed and exposures quantified for 

each potential exposure route based on application rates of disparlure and limited monitoring 

data, given the low toxicity of disparlure to laboratory mammals and the lack of chronic toxicity 

studies, detailed quantitative estimates of exposure will not significantly add to the assessment of 

risk associated with disparlure. 

Dose-Response Assessment – The toxicity data on disparlure are not adequate for making a 

standard dose-response assessment. The limited available data indicate that disparlure has a low 

order of acute toxicity based on mortality as follows: oral LD50 >34,600 mg/kg, dermal LD50 

>2,025 mg/kg, and inhalation LC50 >5 mg/L x 1 hour. Data regarding the toxicity of disparlure 

to animals or humans after subchronic or chronic exposures were not located. Moreover, the 

acute toxicity of this compound for endpoints other than mortality is poorly characterized. Thus, 

due to insufficient data, the U.S. EPA has not derived either an RfD for acute or chronic 

exposure. 

xii 
Risk Characterization – Although studies on the acute toxicity of disparlure have been 

conducted in laboratory animals, the lack of subchronic or chronic toxicity data precludes a 

quantitative characterization of risk. The available data regarding the acute toxicity of disparlure 

indicate that the potential hazard from exposure to the compound is low. 

The reliance on acute toxicity data introduces uncertainties into the risk assessment that cannot 

be quantified. Other uncertainties in this analysis are associated with the exposure assessment 

and involve environmental transport and dermal absorption. These uncertainties are relatively 

minor compared to the lack of subchronic or chronic toxicity data. Thus, while there is no 

reason to believe that longer-term exposure to disparlure will produce adverse effects, this 

assumption can not be substantiated due to the lack of chronic toxicity data. The significance of 

this uncertainty is at least partially offset by the very low exposures that are plausible given the 

low application rates and the nature of plausible exposures of humans to disparlure. 

ECOLOGICAL RISK ASSESSMENT 

Hazard Identification – There is very little information regarding the toxicity of disparlure to 

nontarget wildlife species. As discussed above, rigorous toxicity testing of disparlure has not 

been required by the U.S. EPA. Thus, the only studies available are acute toxicity studies in 

bobwhite quail, mallard ducks, rainbow trout, bluegill sunfish, Daphnia magna and Eastern 

oysters. No chronic toxicity studies were identified in the literature or in the studies submitted to 

the U.S. EPA. 

Results of acute gavage and dietary toxicity studies in mallard ducks and bobwhite quail show 

that disparlure has very low toxicity in these species, with no mortalities observed following 

exposure to up to 2510 mg/kg bw in bobwhite quail. 

Limited data are available regarding the toxicity of disparlure to aquatic animals. A major issue 

in the interpretation of the aquatic toxicity data on disparlure involves the solubility of disparlure 

in water. While no measured values for the solubility of disparlure in water are available, 

estimates based on quantitative structure-activity relationships developed by the U.S. EPA 

suggest that the solubility of disparlure in water is in the range of 0.0019 to 0.0028 mg/L. The 

bioassays that have been conducted on disparlure and Disrupt II formulations of disparlure have 

not measured concentrations of disparlure in the test water but report nominal concentrations of 

disparlure that exceed the water solubility of disparlure by factors of about 10 [0.028 mg/L] to 

over 150,000 [300 mg/L]. Based on the results of the available bioassays and considerations of 

water solubility, disparlure does not appear to present any toxic hazards to aquatic species. In 

toxicity tests of small aquatic invertebrates (i.e., daphnids), trapping of the organism at the 

surface of the water has been noted in bioassays of both technical grade disparlure and Disrupt II 

formulations. The trapping of small invertebrates at surface of the water can present a physical 

hazard to the organism. The significance of this physical hazard observed in bioassays to 

potential hazards in field applications is unclear. 

Exposure Assessment – Disparlure appears to be essentially nontoxic to mammals and birds. 

While this assessment is limited by the lack of chronic toxicity data in terrestrial species, it is not 

expected that acute or chronic exposure of terrestrial mammals or birds to disparlure would result 

xiii 
in the development of significant adverse effects. Given the low toxicity of disparlure and 

limited available data, an exposure assessment for terrestrial species would not add to the 

assessment of risk for terrestrial species. Thus, an exposure assessment for terrestrial species is 

not included in this risk assessment. For aquatic species, the range of plausible nominal 

concentrations of disparlure in water are calculated at 0.0015 mg/L to 0.0037 mg/L over the 

range of application rates considered in this risk assessment. These concentrations apply to a 1 

meter deep body of water. The lower end of this range is within the estimated solubility of 

disparlure in water – i.e., 0.0019 to 0.0028 mg/L. 

Dose-Response Assessment – Given the low toxicity of disparlure to terrestrial animals coupled 

with the limitations imposed due to lack of chronic toxicity data, no standard dose-response can 

be made for disparlure for terrestrial species. Disparlure is produced by other species in the 

genus Lymantria that are closely related to the gypsy moth (http://www.pherobase.com) such as 

the nun moth (Lymantria monacha), a Eurasian pest of conifers that is considered a serious risk 

for introduction into North America (http://www.na.fs.fed.us/spfo/pubs/pest_al/nunmoth/ 

nun_moth.shtm). However, since there are no quantitative data available regarding the efficacy 

of disparlure in nontarget moths, a dose-response assessment for this effect in a nontarget species 

cannot be made. Similarly, no explicit dose-response relationship is proposed for fish. There is 

no basis for asserting that adverse effects in fish are plausible under any foreseeable conditions. 

For aquatic invertebrates, there is no basis for asserting that toxic effects are likely at the limit of 

the solubility of disparlure in water. At nominal concentrations that exceed the solubility of 

disparlure in water (e.g., as the result of an accidental spill or application to water), small 

invertebrates that may interact with the water-surface interface could become trapped in this 

interface due to a layer of undissolved disparlure at the air-water interface. 

Risk Characterization – There is little data available on terrestrial and aquatic animals to allow 

for a quantitative characterization of risk. Furthermore, the lack of chronic toxicity data in any 

species adds significant uncertainty to any risk characterization. Thus, for both terrestrial and 

aquatic species, the potential for the development of toxicity from long-term exposure to 

disparlure cannot be assessed. Nonetheless, given the low toxicity of disparlure based on acute 

toxicity studies, it is unlikely that exposure to disparlure will result in the development of serious 

adverse effects in terrestrial and aquatic species. Regarding potential effects on terrestrial 

invertebrates, disparlure is able to disrupt mating of some other closely related species of moths 

other than the gypsy moth. These other closely related species, however, are all Asian or 

Eurasian species and are not known to exist in North America. Thus, there is no basis for 

asserting that mating disruption is plausible in nontarget species in North America. 

Under normal conditions, aquatic species will not be exposed to substantial levels of disparlure. 

At the limit of the solubility of disparlure in water, there is no indication that toxic effects are 

likely in any aquatic species. If Disrupt II flakes are accidently applied to water, the amount of 

disparlure in the water could result in the formation of an insoluble layer of disparlure at the airwater interface. There is no indication that this would impact fish. Based on toxicity studies 

conducted in the laboratory, small invertebrates that come into contact with the air-water 

interface might become trapped in an insoluble film of disparlure. The likelihood of this 

occurring and the likelihood of this causing any detectable impact in a body of water is difficult 

xiv 
to determine and would vary with the quantity of flakes applied to the body of water and the 

depth of the body of water. Based on variability in the experimental data as well as the range of 

application rates used in the USDA programs, hazard quotients would vary from about 0.15 to 

about 0.37 below the level of concern by factors of about 3 to 10. This risk characterization 

applies to accidental application of disparlure to a 1 meter deep body of water. 

xv 
1. INTRODUCTION 

The USDA Forest Service uses disparlure and the formulation of disparlure as Disrupt II in 

programs to control or eradicate gypsy moth populations. This document is an update to a risk 

assessment prepared in 1995 (USDA 1995) and provides risk assessments for human-health 

effects and ecological effects to support an assessment of the environmental consequences of 

these uses. 

This document has four chapters, including the introduction, program description, risk 

assessment for human health effects, and risk assessment for ecological effects or effects on 

wildlife species. Each of the two risk assessment chapters has four major sections, including an 

identification of the hazards associated with disparlure, an assessment of potential exposure to 

the product, an assessment of the dose-response relationships, and a characterization of the risks 

associated with plausible levels of exposure. These are the basic steps recommended by the 

National Research Council of the National Academy of Sciences (NRC 1983) for conducting and 

organizing risk assessments. 

Although this is a technical support document and addresses some specialized technical areas, an 

effort was made to ensure that the document can be understood by individuals who do not have 

specialized training in the chemical and biological sciences. Certain technical concepts, 

methods, and terms common to all parts of the risk assessment are described in plain language in 

a separate document (SERA 2006). 

The human health and ecological risk assessments presented in this document are not, and are 

not intended to be, comprehensive summaries of all of the available information. No published 

reviews regarding human health or ecological effects of disparlure have been encountered. 

Moreover, almost all of the mammalian toxicology studies and most of the ecotoxicology studies 

are unpublished reports submitted to the U.S. EPA as part of the registration process for 

disparlure. 

Because of the lack of a detailed, recent review concerning disparlure and the preponderance of 

unpublished relevant data in U.S. EPA files, a complete search of the U.S. EPA FIFRA/CBI files 

was conducted. Full text copies of relevant studies were kindly provided by the U.S. EPA Office 

of Pesticide Programs. These studies were reviewed, discussed in Sections 3 and 4 as necessary, 

and synopses of the most relevant studies are provided in the appendices to this document. 

The Forest Service will update this and other similar risk assessments on a periodic basis and 

welcomes input from the general public on the selection of studies included in the risk 

assessment. This input is helpful, however, only if recommendations for including additional 

studies specify why and/or how the new or not previously included information would be likely 

to alter the conclusions reached in the risk assessments. 

1-1 
2. PROGRAM DESCRIPTION 

2.1. OVERVIEW 

Disparlure is a naturally occurring insect pheromone (attractant) synthesized by the female gypsy 

moth to attract the male gypsy moth. Disparlure can take two enantiomer forms, referred to as 

(+)disparlure and (-)disparlure. Enantiomers are mirror-image molecules with identical gross 

structures. The (+)enantiomer is the form produced by the female gypsy moth and is the only 

form that is biologically active as an attractant. In gypsy moth programs two forms of disparlure 

are used: the (+) enantiomer that is used as an attractant or bait in traps and the racemic mixture, 

a 50:50 blend of the (+) and (-) enantiomers that is used as a control agent. When it is used as a 

control agent, racemic disparlure is broadcast over relatively large areas to disrupt mating by 

confusing the male moths. 

Disparlure is always formulated in a slow release matrix and several different formulations have 

been tested including polyvinyl chloride flakes, microcapsules, and polyvinyl chloride twine. 

Disrupt II, a formulation of disparlure in polyvinylchloride flakes, has been used by the USDA 

Forest Service for many years. The specific formulation has evolved over time. This risk 

assessment considers the available information both on the current and some previous Disrupt II 

formulations. 

Since 1995, the use of disparlure in programs intended to slow the spread of gypsy moths has 

increased over 250-fold, from 2,448 acres treated in 1995 to 647,394 acres treated in 2003. 

(+)disparlure is used as an attractant or bait in two types of traps: milk carton traps that also 

contain DDVP and delta traps that do not contain an insecticide. These traps are used to monitor 

existing (endemic) populations and detect new infestations. 

2.2. CHEMICAL DESCRIPTION 

Disparlure is the common name for cis-7,8-epoxy-2-methyloctadecane: 

H C CH 

CH3 

3 3

O 

H H 

Disparlure can take two enantiomer forms, referred to as (+)disparlure and (-)disparlure. The 

term enantiomer refers to molecules that are structurally identical except for differences in the 3­

dimensional configuration such that one form is the mirror image of the other. 

(+)Disparlure is a naturally occurring insect pheromone (attractant) synthesized by the female 

gypsy moth to attract the male gypsy moth. (+)Disparlure is also a natural constituent of and is a 

pheromone for other species including the nun moth (Lymantria monacha, Morewood et al. 

1999, 2000) and Lymantria fumida [the pink gypsy moth which is a species native to Japan] 

2-1 
(Schaefer et al. 1999). As with the gypsy moth, both of these Lymantria species are forest pests 

and adverse effects on these species are not a substantial concern for this risk assessment. 

Selected chemical and physical properties of disparlure are summarized in Table 2-1. Due to the 

lack of experimental data, most of the values given in Table 2-1 are estimated from EPI Suite, an 

estimation program developed by Meylan and Howard (2000) in conjunction with the U.S. EPA 

(U.S. EPA/OPPT 2000). For convenience, the specific estimates for disparlure that were 

obtained from EPI Suite are referenced in this document as EPI Suite (2006) and a full copy of 

this run is included as Appendix 4. 

In gypsy moth programs, two forms of disparlure are used: the (+)enantiomer and the racemic 

mixture, a 50:50 blend of the (+)enantiomer and (-)enantiomer. For disparlure, the 

(+)enantiomer is the biologically active form (that is, the form that attracts the male gypsy moth). 

Racemic disparlure is used as a control agent. It is broadcast over relatively large areas and 

disrupts mating by confusing male moths. This product is typically aerially applied in a single 

application just before the emergence of adult gypsy moths. Although the label for Disrupt II 

allows a second application later in the season, operational programs never use a second 

application. 

As discussed in Section 3 and Section 4, most toxicity studies conducted on disparlure do not 

specify whether the racemic mix or the (+)enantiomer of disparlure was tested. Except for the 

attractant effects of (+)disparlure, there is no clear indication that toxicity profiles differ between 

the (+)enantiomer of disparlure and the 50:50 racemic mix. For the purposes of this risk 

assessment, no distinction is made between (+)disparlure and the racemic mix. All references to 

the active ingredient (a.i.) refer to disparlure and do not distinguish between (+)disparlure and 

the 50:50 racemic mix. 

When used as a control agent, disparlure is formulated in a slow release matrix and several 

different formulations have been tested including polyvinyl chloride flakes, microcapsules, and 

twine (Caro et al. 1977, 1981; Taylor 1982). In recent programs, the USDA used Disrupt II 

(Leonhardt et al. 1996) and this formulation is currently registered by U.S. EPA (Hercon 

Environmental 1993). This formulation contains 17.9% disparlure and 82.1% carrier flakes. 

Disrupt II flakes are about 1/32 inch by 3/32 inch and consist of polyvinyl chloride films, 

polyvinyl chloride resin and a plasticizer (Hercon Environmental 2004). The USDA has 

participated in the development of new formulations of disparlure in either new flake 

formulations developed by Hercon or new microcapsule formulation being developed by 3M 

(Leonard 2004). 

Currently, the USDA has elected to use a new Disrupt II flake formulation (Leonard 2006a,b). 

As with past formulations of Disrupt II, this flake formulation contains 17.9% disparlure and 

82.1% polyvinylchloride carrier flakes and other inerts (Hercon 2006a,b). As detailed further in 

Section 4.1.3.3, toxicity data are available on the current formulation of Disrupt II as well as a 

previous formulation. Available information on the inerts in Disrupt II is discussed in Section 

3.1.14. 

2-2 
 

2.3. APPLICATION METHODS AND RATES 

The application rates recommended on the label of Disrupt II (Hercon 2006a), range from 6 

grams a.i./acre to 30 grams a.i./acre, corresponding to about 0.0132 lb a.i./acre to 0.066 lb 

a.i./acre[1 gram = 0.0022 lb (avdp)]. 

The USDA uses disparlure in two different types of programs: slow the spread and eradication. 

Slow the spread programs involve the control of the North American Gypsy Moth (NAGM), a 

species that is already established in the US. Slow the spread programs are typically 

administered by the USDA/Forest Service using application rates of 6 grams a.i./acre and 

occasionally using an application rate of 15 g a.i./acre. Tobin and Leonard (2006) have 

estimated that this range of application rates will result in the release of disparlure that is 

substantially greater than the amounts released by female gypsy moths during a major outbreak. 

Eradication efforts are administered by USDA/APHIS (Animal and Plant Health Inspection 

Service). Eradication efforts are focused on the Asian strain of the gypsy moth (AGM) that is 

not known to be established in the United States as well as small and isolated infestations of the 

NAGM that could be eradicated. For purposes of exclusion and eradication, APHIS considers 

AGM to be a separate species from NAGM. With NAGM, eradication uses applications of up to 

15 g a.i./acre. The maximum labeled application rate of 30 g a.i./acre has only been used once 

for AGM eradication. This application involved only 600 acres out of a total of approximately 

2.5 million acres treated between 1995 and 2005 – i.e., less than 0.03% of the total acres treated. 

Because the application rate of 30 g a.i./acre is used only rarely, the current risk assessment will 

explicitly consider application rates in the range of 6 grams a.i./acre and 15 g a.i./acre. If other 

application rates need to be considered in certain applications, the Worksheet A02 of the EXCEL 

workbook that accompany this risk assessment may be modified. This workbook is described in 

Section 4.4.2 of this risk assessment. 

(+)Disparlure is used as an attractant or bait in two types of traps: milk carton traps that also 

contain DDVP and delta traps that do not contain an insecticide. These traps are used to monitor 

existing (endemic) populations and detect new infestations. Since the early 1980s, (+)disparlure 

has been formulated as 3 x 25 mm plastic laminates (two outer layers of 50 μm PVC with an 

inner polymeric layer containing 500 μg (+)disparlure). 

2.4. USE STATISTICS 

Use statistics for the number of acres treated with disparlure according to type of use are 

summarized in Table 2-2 (USDA/FS 2005). From 1995 to 2003, the use of disparlure to slow 

the spread of gypsy moths increased substantially. In 1995, 2,448 acres were treated with 

disparlure flakes and in 2003, 647,394 acres were treated; this is an increase in acres treated of 

over 250-fold. It is anticipated that slow the spread applications will typically entail about 

500,000 acres per year and that these applications will account for 99.9% of all mating disruption 

applications (Leonard 2005a). 

2-3 
 

3. HUMAN HEALTH RISK ASSESSMENT 

3.1 HAZARD IDENTIFICATION 

3.1.1 Overview. 

Insect pheromones are generally regarded as nontoxic to mammals (Jacobson 1976) and, as with 

disparlure, application rates of insect pheromone are generally very low – i.e., pheromones are 

active at very low concentrations. Consequently, U.S. EPA requires less rigorous testing of these 

products than is required of insecticides (U.S. EPA 1994). Except for some standard acute 

toxicity studies in laboratory mammals, little information is available regarding the biological 

activity of disparlure. The USDA has funded acute toxicity studies on disparlure during its 

development for use in the gypsy moth control program. The studies were conducted by 

Industrial Bio-test and were submitted to the U.S. EPA by Hercon Environmental Company as 

part of the registration package (Kretchmar 1972). Summaries of these studies are published in 

the open literature (Beroza et al. 1975). 

Results of acute toxicity studies for oral, dermal, ocular and inhalation exposure to disparlure are 

summarized in Table 3-1. With the exception of one acute gavage study in rats using the 50:50 

racemic mix (Coleman 2000), none of the toxicity studies specified whether the 50:50 racemic 

mix or the (+)enantiomer was tested. Based on the results of studies on disparlure, acute 

exposure to disparlure appears to pose a very low risk to mammals. No studies investigating the 

effects of chronic exposure of mammals to disparlure or studies investigating the effects of 

disparlure on the nervous system, immune system, reproductive system or endocrine system 

were identified. The carcinogenic potential of disparlure has not been assessed. The results of a 

single study show that disparlure is not mutagenic. 

3.1.2 Mechanism of Action 

As discussed in Section 4.1.2.3, the mechanism of action for the efficacy of disparlure as an 

attractant for male gypsy moths has been well characterized. However, since disparlure has very 

low toxicity to mammals, studies on the mechanism of action for toxicity of disparlure in 

mammals have not been conducted. Thus, there is no information available in the FIFRA files or 

in the open literature regarding the mechanism of toxicity (if any) of disparlure in mammals. 

3.1.3 Kinetics and Metabolism 

No studies designed specifically to obtain information on the kinetics or metabolism of 

disparlure were identified. The kinetics of absorption of disparlure following dermal, oral or 

inhalation exposure are not documented in the available literature. Disparlure appears to persist 

in humans for long periods of time. This supposition is based on a case report of an individual 

who had direct dermal contact with disparlure in 1977 (Cameron 1981, 1983, 1995). This 

individual appears to have attracted male gypsy moths for a period of over 15 years. It is 

estimated that the exposure level of this individual to disparlure was very low, although no 

quantitative estimates of exposure were reported. 

3-1 
Assays have been conducted using disparlure and several natural and xenobiotic epoxides to 

determine the ability of each to induce epoxide metabolizing enzymes (Moody et al. 1991). 

Male mice were given 500 mg a.i./kg/day disparlure by intraperitoneal injection for 3 days. This 

was the maximum dose tested in preliminary range finding studies. Exposure to the compound 

had no effect on relative liver weight, using matched controls, or microsomal protein. Relative 

cytosolic protein was significantly (p<0.05) increased by 18% over control values. Disparlure 

also caused a moderate but statistically significant (p<0.05) increase in microsomal cholesterol 

epoxide hydrolase activity. This study suggests that very high doses of disparlure may induce 

enzymes involved in the metabolism of disparlure. Given the very low levels of exposure to 

disparlure that are likely in the use of this agent in gypsy moth control programs, this study has 

no direct relevance to this risk assessment. 

3.1.4 Acute Oral Toxicity 

Other than standard bioassays for acute toxicity that were conducted as part of the registration 

process, no information regarding the acute toxicity of disparlure was identified. The most 

common measure of acute oral toxicity is the LD50, the estimate of a dose that causes 50% 

mortality in the test species. As summarized in Appendix 1, there are two studies investigating 

the acute oral toxicity of high doses of disparlure in rats (Coleman 2000; Kretchmar 1972). 

Acute oral exposure to 10,250–34,600 mg a.i./kg body weight was not lethal to rats (LD50 greater 

than 34,600 mg a.i./kg) (Kretchmar 1972). Disparlure was administered, undiluted, by gavage, 

and the rats were observed for 14 days following exposure. This report does not specify whether 

the test material used was the 50:50 racemic mix or the (+)enantiomer. Necropsy revealed no 

pathological alterations in any of the treated rats. At all dose levels, however, the animals 

exhibited hypoactivity, ruffed fur, and diuresis. The significance of these observations cannot be 

assessed because no control group was used. The apparent NOAEL for mortality and serious 

clinical toxicity is 34,600 mg a.i./kg, the highest dose tested. 

In a more recent study in which rats were administered 5000 mg a.i./kg of a racemic preparation 

of disparlure, no deaths or pathological abnormalities were observed (Coleman 2000). Clinical 

signs of toxicity, including piloerection, hunched posture and ungroomed appearance were 

observed during the first three days following exposure; however, no clinical signs of toxicity 

were noted during the remaining 11 days of the observation period. As in the study by 

Kretchmar (1972), no control group was used in the Coleman (2000) study. In this study the 

LC50 is > 5000 mg a.i./kg and the NOAEL is 5000 mg a.i./kg. Thus, with the acute oral LD50 

exceeding 5,000mg a.i./kg, disparlure would be classified as practically non-toxic using the 

scheme adopted by U.S. EPA (2003). 

3.1.5 Subchronic and Chronic Systemic Toxic Effects 

No studies investigating the subchronic or chronic effects of disparlure in mammals were 

identified. As discussed in Section 8.1.1, studies investigating subchronic and chronic exposures 

were not required for registration of disparlure (Jacobson 1976; U.S. EPA 1994). 

3-2 
3.1.6 Effects on Nervous System 

As discussed in Durkin and Diamond (2002), a neurotoxicant is a chemical that disrupts the 

function of nerves, either by interacting with nerves directly or by interacting with supporting 

cells in the nervous system. This definition of neurotoxicant is critical because it distinguishes 

agents that act directly on the nervous system (direct neurotoxicants) from those agents that 

might produce neurologic effects that are secondary to other forms of toxicity (indirect 

neurotoxicants). Virtually any chemical will cause signs of neurotoxicity in severely poisoned 

animals and thus can be classified as an indirect neurotoxicant. 

By this definition, disparlure may be classified as an indirect neurotoxicant. As noted in Section 

3.1.4, hypoactivity and piloerection were observed following acute oral exposure to very high 

doses of disparlure (Coleman 2000; Kretchmar 1972). These observations, however, do not 

implicate disparlure as a direct neurotoxicant. No studies designed specifically to detect 

impairments in motor, sensory, or cognitive functions in animals or humans exposed to 

disparlure were identified. No evidence for disparlure producing direct effects on the nervous 

system was found. 

3.1.7 Effects on Immune System 

No studies investigating the effects of disparlure on immune system function in mammals were 

identified. 

3.1.8 Effects on Endocrine System 

No studies investigating the effects of disparlure on endocrine system function in mammals were 

identified. 

3.1.9. Reproductive and Teratogenic Effects 

No studies investigating the reproductive or teratogenic effects of disparlure in mammals were 

identified. 

3.1.10. Carcinogenicity and Mutagenicity 

No studies investigating the carcinogenic activity of disparlure in mammals were identified. A 

single study investigated the mutagenicity of disparlure with and without metabolic activation in 

Salmonella typhimurium and Esherichia coli (Oguma 1998). There was no evidence of 

mutagenic activity under any of the experimental conditions of this study. This report does not 

specify whether the test material used was the 50:50 racemic mix or the (+)enantiomer. 

3.1.11. Irritation and Sensitization (Effects on Skin and Eyes) 

The primary skin irritation of disparlure was evaluated in a single study using young albino New 

Zealand rabbits (Kretchmar 1972). Details are provided in Appendix 1. The test sites, located 

lateral to the midline of the shaved back, were approximately 10 cm apart from one another, and 

one site was abraded while the other remained intact. The sites were occluded with gauze 

patches for the duration of the 24-hour exposure period, after which the intact and abraded test 

sites were examined. The sites were examined and scored again after 72 hours. Signs of mild 

skin irritation, including erythema and edema, were noted at 24 and 72 hours after application of 

disparlure. Based on the results of this single study, dermal exposure to a high dose of disparlure 

appears only mildly irritating to skin and is not a primary skin irritant. 

3-3 
Eye irritation was assayed in a single study in six young New Zealand rabbits exposed to 0.1 mL 

disparlure (Kretchmar 1972). Details of this study are provided in Appendix 1. Disparlure was 

instilled into the right eye of each rabbit (the left eye served as a control) to determine the extent 

of irritation or damage to cornea, iris, and conjunctiva. The severity of ocular lesions was 

monitored at intervals of 24, 48, and 72 hours. Three of the six rabbits had redness of the 

conjunctiva at 24 hours, but no effects were observed in any of the rabbits at the later observation 

periods. No effects were observed 7 days after exposure. Based on the results of this study, 

disparlure would be classified as a non-irritant for eyes using the scheme proposed by U.S. EPA 

(2003). 

3.1.12. Systemic Toxic Effects from Dermal Exposure 

The acute dermal toxicity of disparlure was tested using four young adult New Zealand rabbits 

(Kretchmar 1972). Study details are provided in Appendix 1. When applied, undiluted, to the 

shaved backs of the rabbits, 2,025 mg a.i./kg caused local skin reactions after 24 hours of contact 

with the epidermis. No other dose levels were tested. The rabbits were observed for 14 days 

after exposure, and the effects observed during this period included dryness (escharosis), skin 

flaking (desquamation), hemorrhaging, and fissures after 7 days and desquamation, fissures, and 

pustules after 14 days. Necropsy revealed no pathological alterations other than the effects on 

the skin. None of the rabbits died as a result of treatment (dermal LD50 greater than 2,025mg 

a.i./kg). 

3.1.13. Inhalation Exposure 

The acute toxicity of inhalation exposure to disparlure was assessed in rats (Grapenthien 1972). 

Study details are provided in Appendix 1. Rats were exposed to an aerosol of disparlure for 1 

hour, with a calculated average concentration of the aerosol was 5.0 mg a.i./L air. The rats were 

observed for 14 days after exposure. None of the rats died as a result of exposure. No clinical 

signs of toxicity were reported. The LC50 for inhalation exposure is > 5.0 mg a.i./L air. 

3.1.14. Inerts and Adjuvants 

As discussed in Section 2, disparlure is typically applied in a slow release polyvinyl chloride 

formulation and various formulations have been tested and used in USDA programs. As also 

discussed in Section 2, the USDA uses Disrupt II, a formulation of polyvinyl chloride flakes. 

The precise composition of the flake formulation is considered proprietary by Hercon. In the 

preparation of the current risk assessment, the product manager at Hercon for Disrupt II was 

contacted and some information on the inerts has been disclosed. The new formulation of 

Disrupt II contains 5 inert ingredients. Two of the inerts, one of which is identified as 

diatomaceous earth, are on the U.S. EPA List 4A list and another is on List 4B. A new inert is 

listed on the exemptions from requiring tolerances 40 CFR 180.910 and 180.930. 

Polyvinylchloride itself is exempt from tolerance under 40 CFR 180.960 (MacLean 2006). 

The reference to the U.S. EPA List 4 refers to the U.S. EPA method for classifying inert 

ingredients that are used in pesticide formulations. U.S. EPA classifies inerts into four lists 

based on the available toxicity information: toxic (List 1), potentially toxic (List 2), 

3-4 
unclassifiable (List 3), and non-toxic (List 4). These lists as well as other updated information 

on pesticide inerts are maintained by the U.S. EPA at the following web site: 

http://www.epa.gov/opprd001/inerts/. Any compound classified by U.S. EPA as toxic or 

potentially toxic must be identified on the product label if the compound is present at a level of 

1% or greater in the formulation. If the compounds are not classified toxic, all information on 

the inert ingredients in pesticide formulations is considered proprietary under Section 10(a) of 

the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA). In that case, the formulators 

of the pesticide need not and typically do not disclose the identity of the inert or adjuvant. List 

4A is classified as minimal risk inert ingredients. List 4B is defined by the U.S. EPA as follows: 

Other ingredients for which EPA has sufficient information to 

reasonably conclude that the current use pattern in pesticide 

products will not adversely affect public health or the environment 

(http://www.epa.gov/opprd001/inerts/lists.html) 

As discussed further in Section 4.1.3.3, some information is available on the toxicity of 

disparlure, the Disrupt II formulation of disparlure, and Disrupt II flakes that contain only the 

PVC flakes and other inerts (i.e., no disparlure). While limited, this information suggests that the 

PVC flakes and other inerts do not contribute to the toxicity of Disrupt II. 

3.1.15. Impurities and Metabolites 

3.1.15.1. Impurities –Virtually no chemical synthesis yields a totally pure product. Technical 

grade disparlure does contain low concentrations of four compounds that are structurally related 

to disparlure – i.e., three octadecenes (all at less than 1%) and one octadecyne (at less than 0.5%) 

(MTM Chemicals 1991). Additional data regarding impurities in disparlure have been identified 

in the FIFRA/CBI files (Shin-Etsu Chemical Company 2002; Oguma 2000). The specific 

information contained in these files is protected under FIFRA Section 12(a)(2)(D) and this 

information cannot be disclosed in this risk assessment. Nonetheless, concern for impurities is 

reduced by the fact that the toxicity of impurities should be encompassed in the acute toxicity 

studies conducted on technical grade disparlure – i.e., disparlure that contains these impurities. 

3.1.15.2. Metabolites – No studies on the metabolism of disparlure in mammals were identified 

in the open literature or the FIFRA/CBI files. Acute toxicity studies, however, typically involve 

a single exposure followed by a period of observation, most often a 14-day post-dosing period 

(e.g., U.S. EPA/OPPTS 2003). Because of this, the effects of metabolites formed during the 

observation period should be encompassed in the acute toxicity studies conducted on disparlure. 

3.1.16. Toxicological Interactions. 

DDVP pest strips (Vaportape II strip) are contained in the milk carton trap together with a carrier 

containing disparlure. These milk carton traps are placed in selected areas to monitor gypsy 

moth infestations. No published literature or information in the FIFRA files permit an 

assessment of potential toxicological interactions between disparlure and DDVP or any other 

compounds. A separate risk assessment on DDVP has been prepared as part of the series of risk 

assessments on the control/eradication agents used for the gypsy moth. 

3-5 
3.2. EXPOSURE ASSESSMENT 

3.2.1. Overview 

For both workers and the general public, exposures to disparlure may involve multiple routes of 

exposure (i.e., oral, dermal, and inhalation). Because of the limited toxicity data on disparlure – 

i.e., no chronic toxicity data are available – no chronic exposure scenarios are developed. 

3.2.2. Dermal Exposure 

Dermal exposure is most likely to be the predominant route for occupational exposure to 

disparlure and is also a possible route of exposure for the general public. As discussed in Section 

3.1.3, a case report of an accidental exposure of a worker to disparlure shows that no signs of 

toxicity developed; the only notable effect of disparlure exposure in this worker was the 

persistent attraction of gypsy moths (Cameron 1981, 1983, 1995). Exposure of this worker was 

most likely by the dermal route, although the possibility of inhalation exposure cannot be ruled 

out (Cameron 1995). Since the systemic toxicity of disparlure in mammals is very low, the 

absence of dermal absorption data does not add significant uncertainty to this risk assessment 

since no systemic toxicity would be expected to occur, even at very high exposure levels of 

disparlure. While dermal exposure of workers is expected to be non-toxic, dermal exposure is 

likely to cause the persistent attraction of gypsy moths. 

3.2.3. Inhalation Exposure 

Both workers and the public may be exposed to disparlure by inhalation and the magnitude of the 

exposure can be estimated from available monitoring studies. In these studies, high application 

rates, relative to the projected rates used in program activities (29.1 g/acre, Section 2.3), were 

used in order to be able to detect disparlure in air. 

Caro et al. (1981) investigated the distribution and persistence of three disparlure formulations 

including gelatin microcapsules, laminated plastic flakes, and hollow fibers. Each formulation 

was applied at a rate of 500 g a.i./hectare (approximately 0.45 lb a.i./acre). Release of disparlure 

from these formulations was most rapid during the first 2 days after application. Initially, air 

concentrations ranged from approximately 22 to 30 ng/m3

 (nanograms per meter cubed) for 

microcapsules and fibers and from 7.3 to 8.2 ng/m3

 for flakes. Other investigators using the 

same application rate reported similar initial concentrations of disparlure in air, approximately 

28-30 ng/m3

 for gelatin microcapsules and laminated plastic flakes (Taylor 1982). At a lower 

application rate (250 g/hectare), there were somewhat higher levels, 44.5-99.3 ng/m3

, using 

gelatin microcapsules (Plimmer et al. 1977). 

Over time, the concentrations of disparlure in air will decrease as the disparlure dissipates. After 

30 days, air concentrations ranged from approximately 0.4 to 2.5 ng/m3

 for all formulations 

(Caro et al. 1981). Flakes that originally contained 7.1% disparlure (w/w) contained 6.0% (w/w) 

disparlure (85% of the original level) by 30 days after treatment. Results of a study using a 

disparlure gelatin microcapsule formulation show that release rates increase with increasing 

temperature (Caro et al. 1977). 

3-6 
The highest reported air concentration after aerial application of 250 g/hectare racemic disparlure 

on flakes is slightly less than 100 ng/m3

 (Taylor 1982). At an application rate of nearly 30 

g/acre, concentrations of approximately 30 ng/m3 can be expected. Since this estimate is based 

on the highest levels of disparlure in air, which occur within the first 5 days after application 

(Caro et al. 1981, Taylor 1982), actual levels of exposure could be lower. 

Air concentrations resulting from the release of disparlure from traps are expected to be low 

relative to air concentrations resulting from aerial application of disparlure. Traps contain only 

0.5 mg disparlure/trap. The rate of dissipation of disparlure from traps is dependent upon many 

factors, including dispenser design, lure type, and air temperature and flow (Bierl 1977, BierlLeonbardt 1979, Leonhardt et al. 1990). Thus, air concentrations resulting from volatilization of 

disparlure from traps are expected to be very low and highly variable. 

Over a 120-day period, 38 to 68% of disparlure was lost from lures in laminated plastic 

dispensers, with loss varying over a variety of experimental conditions (Bierl-Leonbardt 1979). 

Loss of (+)disparlure was reduced with the use of thicker plastic dispensers and increased with 

increasing air flow rate and increasing temperature. Greenhouse studies have shown that 

approximately 50%–80% of (+)disparlure is released from PVC twine or laminates during a 16­

week aging process (Kolodny-Hirsch and Webb 1993). Release rates 30 to 40 ng/hr were noted 

from cotton wicks containing 100 μg (+)disparlure, with increased rates observed at higher 

temperatures. 

3.2.4. Oral Exposure 

Although the efficacy of disparlure depends on its volatility, the studies summarized above 

demonstrate that 70%–85% of disparlure may remain in the carrier matrix after prolonged 

periods of time. Consequently, oral exposure may occur from consumption of disparlure flakes 

or tape. At an application rate of approximately 30 g/acre, an individual would have to consume 

all of the flakes in a 1 m2

 area to receive a dose of 7.4 mg. If this were done by a 10 kg child, 

the dose would be 0.74 mg/kg. 

3-7 
3.3. DOSE-RESPONSE ASSESSMENT 

The toxicity data on disparlure are not adequate for making a standard dose-response assessment. 

As detailed in Section 3.1, the limited available data indicate that disparlure has a low order of 

acute toxicity, based on mortality as the endpoint: 

 Oral LD50 >34,600 mg/kg 

 Dermal LD50 >2,025 mg/kg 

 Inhalation LC50 >5 mg/L x 1 hour 

Data regarding the toxicity of disparlure to animals or humans after subchronic or chronic 

exposures were not located in the available literature. Moreover, the acute toxicity of this 

compound for endpoints other than mortality is poorly characterized. 

3-8 
3.4. RISK CHARACTERIZATION 

3.4.1 Overview 

Although studies on the acute toxicity of disparlure have been conducted in laboratory animals, 

the lack of subchronic or chronic toxicity data precludes a quantitative assessment of risk for 

longer-term exposures. The available data regarding the acute toxicity of disparlure indicate that 

the potential hazard from exposure to the compound is low. 

The reliance on acute toxicity data introduces uncertainties into the risk assessment that cannot 

be quantified. Other uncertainties in this analysis are associated with the exposure assessment 

and involve environmental transport and dermal absorption. Thus, while there is no reason to 

believe that longer-term exposure to disparlure will produce adverse effects, this assumption can 

not be substantiated due to the lack of chronic toxicity data. The significance of this uncertainty 

is at least partially offset by the very low exposures that are plausible given the low doses of 

disparlure used in programs to control the gypsy moth. 

3.4.2. Workers and the General Public 

It is not possible to develop a reference dose (RfD); therefore, the calculation of a hazard 

quotient (level of exposure divided by the RfD) and a standard risk characterization cannot be 

developed. Nonetheless, the limited information that is available regarding the use and toxicity 

of disparlure gives no clear indication of hazard. For example, the plausible level of oral 

exposure to a small child is less than 1 mg/kg (Section 3.1.4). This is a factor of 10,000–35,000 

less than the exposure levels that were not lethal to rats (Kretchmar 1972, Section 3.1.4). 

Empirical relationships between acute exposure levels that are lethal to experimental mammals 

and subchronic or chronic NOAELs in experimental mammals (for example, Dourson and Stara, 

1983) do not suggest that the use of disparlure to control of the gypsy moth is likely to pose a 

substantial hazard to humans. 

The only clear and unequivocal biological activity of disparlure is its ability to attract the male 

gypsy moth. Because disparlure appears to be highly persistent in humans, dermal contact with 

the compound might make an individual an attractant to male moths over a period of many years. 

Although this is not likely to cause adverse health effects, it is likely to be a nuisance. 

3.4.3. Sensitive Subgroups 

The toxic effects of disparlure, if any, have not been identified. Consequently, groups at special 

risk, if any, cannot be characterized. Because disparlure attracts the male gypsy moth, 

individuals who have an aversion to insects might be considered to be a sensitive subgroup. 

Nonetheless, this aversion and sensitivity would not be related to any frank health effect. 

3-9 
3.4.4. Cumulative Effects 

Very little information is available on the toxicity of disparlure. As noted above, the ability to 

attract the male gypsy moth is the only clear biological activity of this compound. Since this 

compound seems to persist in humans for prolonged periods, repeated exposures are more likely 

than single exposures to transfer sufficient quantities of disparlure to the individual to attract the 

moth. 

3.4.5. Connected Actions 

No information is available on the interaction of disparlure with other control agents or other 

chemicals usually found in the environment. There is an obvious and substantial interaction of 

disparlure with the adult male gypsy moth. Individuals who are exposed to sufficient quantities 

of disparlure and who live in an area in which male gypsy moths reside will attract the moth. 

The definition of a sufficient quantity of disparlure, however, cannot be characterized from the 

available data. 

3-10 
4. ECOLOGICAL RISK ASSESSMENT 

4.1. HAZARD IDENTIFICATION 

4.1.1. Overview 

There is very little information regarding the toxicity of disparlure to nontarget wildlife species. 

As discussed in Section 3.1, rigorous toxicity testing of disparlure was not required by the U.S. 

EPA (U.S. EPA 1994). Thus, the only studies identified in the available literature are acute 

toxicity studies in bobwhite quail, mallard ducks, rainbow trout, bluegill sunfish, Daphnia 

magna and Eastern oysters. No chronic toxicity studies were identified in the available 

literature. 

Results of acute gavage and dietary toxicity studies in mallard ducks and bobwhite quail show 

that disparlure has very low toxicity in these species, with no mortalities observed following 

exposure to up to 2510 mg/kg bw in bobwhite quail. 

Limited data are available regarding the toxicity of disparlure to aquatic animals. A major issue 

in the interpretation of the aquatic toxicity data on disparlure involves the solubility of disparlure 

in water. While no measured values for the solubility of disparlure in water are available, 

estimates based on quantitative structure-activity relationships developed by the U.S. EPA 

suggest that the solubility of disparlure in water is in the range of 0.0019 to 0.0028 mg/L. The 

bioassays that have been conducted on disparlure and Disrupt II formulations of disparlure have 

not measured concentrations of disparlure in the test water but report nominal concentrations of 

disparlure that exceed the water solubility of disparlure by factors of about 10 [0.028 mg/L] to 

over 150,000 [300 mg/L]. Based on the results of the available bioassays and considerations of 

water solubility, disparlure does not appear to present any toxic hazards to aquatic species. In 

toxicity tests of small aquatic invertebrates (i.e., daphnids), trapping of the organism at the 

surface of the water has been noted in bioassays and this can present a physical hazard to the 

organism. The significance of this physical hazard observed in bioassays to potential hazards in 

field applications is unclear. 

4.1.2. Toxicity to Terrestrial Organisms 

4.1.2.1. Mammals– As discussed in Section 3.1, there is very little information on the toxicity of 

disparlure in mammalian species. Results of acute toxicity studies for oral, dermal, ocular and 

inhalation exposure to disparlure show that disparlure has very low toxicity to mammals. Other 

than some minor clinical signs of toxicity (i.e., piloerection, hunched posture and ungroomed 

appearance in rats), acute oral exposure of rats to very high doses of disparlure (up to 34,600 mg 

a.i./kg bw) did not result in death or signs of systemic toxicity in rats (Kretchmar 1972). Thus, 

acute exposure to disparlure does not appear to exhibit any organ-specific toxicity. There is no 

information available regarding the effects of chronic exposure of mammals to disparlure. No 

field studies are available in which the impact of disparlure were assessed on mammalian 

wildlife communities. 

4-1 
4.1.2.2. Birds– As summarized in Appendix 2, the acute toxicity of disparlure administered by 

gavage has been studied in bobwhite quail (Fink et al. 1980) and acute exposure to dietary 

disparlure has been studied in bobwhite quail chicks and mallard ducklings (Hudson 1975). In 

adult bobwhite quail administered single doses of disparlure ranging from 398 to 2510 mg a.i./kg 

by gavage, no mortalities were observed at any dose level (Fink et al. 1980). In the highest dose 

group, lethargy was observed in 3 of 10 birds; it is unclear if this observation was treatment 

related. In quail chick and mallard ducklings exposed to 313 to 5000 ppm disparlure in the diet 

for 5 days, no mortalities were observed and no clinical signs of toxicity were reported during 

the 14-day observation period. Based on the results of these studies, the LD50 for a single dose 

of disparlure administered by gavage to bobwhite quail is > 2510 mg a.i./kg bw and the 

corresponding value for 5-day dietary exposure to quail chicks and mallard ducklings is > 5000 

ppm. 

4.1.2.3. Terrestrial Invertebrates– As discussed in Section 2, disparlure is a naturally occurring 

insect pheromone. The mechanism of action of disparlure in disrupting gypsy moth mating is 

well characterized. The (+)disparlure enantiomer, which is produced and released by female 

gypsy moths, is a powerful attractant to male gypsy moths. Male gypsy moths detect disparlure 

through highly specific detectors located on antennae (Murlis et al. 2000, Plettner et al. 2000). 

The (–)disparlure enantiomer is a receptor antagonist to (+)disparlure and has slight repellent 

activity (Plettner et al. 2000). When sprayed over a large area, disparlure disrupts mating by 

confusing male moths. There are a large number of greenhouse and field studies showing that 

disparlure is an effective agent in decreasing gypsy moth populations (Beroza et al, 1975, 

Campbell 1983, Herculite Products Inc., 1978, Kolodny-Hirsch and Webb 1993, Leonhardt et 

al. 1990, Leonhardt et al. 1993, Leonhardt et al. 1996, Plimmer et al. 1977, Schwalbe et al. 

1978, Schwalbe et al. 1979, Sharov et al. 2002, Thorpe et al. 1993, US Department of 

Agriculture 1973). 

Although disparlure is considered highly selective for gypsy moths, there is some evidence 

showing that disparlure may have effects on the mating of other species of moths. As part of the 

reproductive communication between male and female nun moths, female nun moths produce a 

blend of pheromones that contains disparlure (Gries et al. 2001). Studies show that lures 

containing disparlure are effective in attracting male nun moths (Gries et al. 2001, Morewood et 

al. 1999, Morewood et al. 1999). The potency of disparlure in attracting male gypsy moths 

relative to nun moths has not been assessed. Disparlure is also produced by L. fumida [a species 

native to Japan] (Schaefer et al. 1999). Thus, based on the results of these studies, it appears that 

disparlure is not completely selective for the gypsy moth. Although studies have not been 

conducted, it is possible that other closely related species of moths could also respond to 

disparlure. 

No laboratory or field studies on the effects of acute or chronic exposure of disparlure to other 

terrestrial invertebrates were identified in the available literature. 

4.1.2.4. Terrestrial Plants (Macrophytes)–Neither the published literature nor the U.S. EPA files 

include data regarding the toxicity of disparlure to terrestrial plants. 

4-2 
 

4.1.2.5. Terrestrial Microorganisms– Neither the published literature nor the U.S. EPA files 

include data regarding the toxicity of disparlure to terrestrial microorganisms. 

4.1.3. Aquatic Organisms 

4.1.3.1. Fish – As summarized in Appendix 3, acute toxicity studies of disparlure were 

conducted in rainbow trout and bluegill sunfish (Knapp and Terrell 1980, Rausina no date). No 

effect on survival was observed in bluegill sunfish exposed to disparlure at a nominal 

concentration of 100 mg/L (Rausina no date) or 300 mg/L (Knapp and Terrell 1980) for up to 96 

hours. The 96-hour LC50 for bluegill sunfish is >300 mg/L. In rainbow trout, no effect on 

survival was observed following exposure to 100 mg/L disparlure for 48 hours (Rausina no 

date). However, after 72 hours of exposure to 100 mg/L disparlure, only 8 of 10 trout survived. 

Survival of trout was not affected at disparlure concentrations of 0.1 to 10 mg/L. Under these 

experimental conditions, the NOEC for mortality in rainbow trout is 10 mg/L. 

Neither of these studies would be considered acceptable by current standards for toxicity studies 

in fish (e.g., U.S. EPA/OPPTS 2006). For example, the U.S. EPA guidelines for acute toxicity 

studies in fish require information on the solubility of test compound in water and require that 

the test substance not be tested as concentrations in excess of the solubility of the compound in 

water. 

As noted above and detailed further in Appendix 3, neither Rausina (no date) nor Knapp and 

Terrell (1980) measured the concentration of disparlure in the test water. As noted in Section 2, 

no measured values are available for the solubility of the disparlure in water. Based on 

quantitative structure activity relationships (QSAR), however, it is likely that the solubility of 

disparlure in water is very low. As indicated in Table 2-1, the QSAR package developed by the 

U.S. EPA estimates a water solubility for disparlure of 0.0019 to 0.0028 mg/L (EPI Suite 2006). 

In the preparation of this risk assessment, Hercon (the company that manufactures the Disrupt II 

flakes) was contacted and the chemists at Hercon indicated that they were not aware of any 

measured water solubility values for disparlure but, consistent with the estimates from EPI Suite 

(2006), the chemists at Hercon indicated that the water solubility is likely to be very low. 

The importance of considering water solubility in the assessment of a chemicals toxicity to 

aquatic species is discussed by Clements et al. (1996), the individuals who developed the toxic 

estimation algorithms used in EPI Suite. Essentially, if a compound is non-toxic at the limit of 

water solubility, then the compound can be classified as presenting no plausible toxic risk to the 

organism. Physical hazards may still be plausible. This is discussed further in Section 4.1.3.3 

(Aquatic Invertebrates). 

The toxicity values estimated by EPI Suite (2006) using algorithms of Clements et al. (1996) are 

summarized in Table 4-2. The algorithms used to estimate the toxicity values were developed by 

Clements et al. (1996) and are based on regression equations which take the general form of: 

Log10(TV) = m Log10(Kow) + b 

4-3 
where TV is the toxicity value in units of millimoles/liter (mM/L), Kow is the octanol/water 

partition coefficient, and m and b are model parameters (slope and intercept, respectively). 

While the algorithms are based on molar concentrations, EPI Suite converts these concentrations 

to units of mg/L for the output files. The specific model parameters are summarized in Table 4-2 

and are based on QSAR estimates for mono-epoxides – i.e., compounds structurally similar to 

disparlure. 

A very important feature of these estimates concerns the limiting values for the Kow of the 

compound. As discussed by Clements et al. (1996), this recommended limiting value is based on 

the range of Kow values on which the QSAR estimates are based. For mono-epoxides, the limit 

recommended by Clements et al. (1996) is 5. As noted in Table 2-1, the estimated log Kow 

value for disparlure is 8.08 – i.e., higher than the recommended cut off value by a factor of about 

1000. 

This cutoff value is very important in the interpretation of the estimated toxicity values. As 

indicated in Table 4-2, the estimated toxicity values for fish range from about 0.12 to 0.14 mg/L 

based on the Kow. Although the studies by Knapp and Terrell (1980) as well as Rausina (no 

date) have serious limitations, they clearly indicate no mortality at the nominal concentrations. It 

is likely, however, that the actual concentrations would not have exceeded the water solubility of 

disparlure – i.e., 0.0019 to 0.0028 mg/L (Table 2-1). The simple interpretation is that the water 

solubility of disparlure is so low that the maximum possible concentration in water is below the 

estimated toxicity values by a factor of about 43 [0.12 mg/L ÷ 0.0028 mg/L] to 74 [0.14 mg/L ÷ 

0.0019 mg/L]. This is the basis for asserting that disparlure is not likely to pose a risk of toxicity 

to fish. 

Thwaits and Sorensen (2005) have recently submitted a brief summary of a study using rainbow 

trout in which disparlure was assayed for olfactory stimulation. At nominal concentrations of 

either 0.028 mg/L or 0.28 mg/L, with or without the presence of methanol (used to enhance the 

solubility of disparlure in water), disparlure evidenced no activity relative to negative controls 

(well water or well water with methanol) or L-serine as a positive control. 

4.1.3.2. Amphibians– Neither the published literature nor the U.S. EPA files include data 

regarding the toxicity of disparlure to amphibian species. 

4.1.3.3. Aquatic Invertebrates – As with fish, the data on the toxicity of disparlure itself to 

aquatic invertebrates is relatively old (LeBlanc et al. 1980; Ward 1981) and these studies would 

not meet the current requirements of the U.S. EPA (e.g., U.S. EPA/OPPTS 2006) because of the 

same limitations discussed in Section 4.1.3.1 (Fish). The acute toxicity of disparlure to Daphnia 

was evaluated in a single study (LeBlanc et al. 1980). Details of this study are provided in 

Appendix 3. A dose-related increase in mortality was observed following 48 hours of exposure, 

with 7% mortality at 0.028 mg/L and 100% mortality at a 0.22 mg/L. The LC50 value was 

calculated at 0.098 mg/L and the NOEC for mortality was 0.017 mg/L. In Eastern oysters 

exposed to 1.25 to 20 mg/L disparlure for 96 hours, there was no effect on new shell growth 

(Ward 1981). Again, all of these toxicity values refer to nominal concentrations rather than 

4-4 
measured concentrations and all of these toxicity values exceed the plausible range of the 

solubility of disparlure in water – i.e., 0.0019 to 0.0028 mg/L (Table 2-1). 

The major difference, however, between the data on fish and data on daphnids involves the 

mortality. As detailed in Appendix 3, LeBlanc et al. (1980) report a clear dose-response 

relationship for daphnids. The important detail, however, is that this mortality was associated 

with organisms being trapped at the air-water interface. While not discussed by LeBlanc et al. 

(1980), it is likely that the entrapment of the daphnids at the air-water interface was attributable 

to the undissolved disparlure in the test solution. Based on the highest estimate of the solubility 

of disparlure in water (i.e., 0.0028 mg/L) the nominal test concentrations used by LeBlanc et al. 

(1980) exceed the solubility of disparlure in water by factors of 10 [0.028 mg/L ÷ 0.0028 mg/L] 

to about 78 [0.22 mg/L ÷ 0.0028 mg/L]. 

The supposition that daphnid mortality in the study by LeBlanc et al. (1980) is due to the 

physical trapping of the organisms at the water surface by undissolved disparlure is supported by 

the more recent studies by Palmer and Krueger (2006a,b) on various formulations of Disrupt II 

flakes. The studies were sponsored by the Forest Service because of concerns with the quality of 

the data on disparlure, the preliminary risk assessment on disparlure (SERA 2004), as well as a 

desire to better characterize the potential hazards of the inerts used in Disrupt II formulations. 

The studies by Palmer and Krueger (2006a,b) involved Disrupt II formulations that were 

designated as Standard Flakes and Modified Flakes. This nomenclature is somewhat awkward 

but will be maintained because these terms are used in the reports by Palmer and Krueger 

(2006a,b) and these terms are also used (at least currently) by individuals in the USDA who are 

involved in applications of Disrupt II (e.g., Leonard 2006b). Standard flakes refer to an older 

formulation that was the only formulation used operationally in USDA programs up through 

2003. Hercon modified their Disrupt II formulation by changing one of the inert ingredients and 

these modified flakes were first tested by USDA in 2002. By 2004 the modified formulation of 

Disrupt II had replaced the standard formulation in most operational applications (Leonard 

2006d). As noted in Section 2, the USDA has been involved in the refinement of various 

formulations of disparlure for many years and it seems likely that new formulations will be 

developed in the future. 

Standard Flakes were tested in the study by Palmer and Krueger (2006a) and Modified Flakes 

were tested in the study by Palmer and Krueger (2006b). Both of these studies involved identical 

experimental designs, the details of which are given in Appendix 3. Both studies involved three 

set of flakes: blank flakes that contained no disparlure (i.e., only the inerts), fully formulated 

flakes that were manufactured in 2003, and fully formulated flakes that were manufactured in 

2005. 

In each study, the daphnids were exposed to a series of six water accommodated fractions 

(WAF) at nominal concentrations of 0.18, 0.54, 1.8, 5.4, 18, and 54 mg a.i./L. The technique 

using water accommodated fractions is a method specifically designed for water insoluble 

compounds (e.g., French-McCay 2002; Pelletier et al. 1997). As implemented by Palmer and 

Krueger (2006a,b), the application of this method involved mixing the flakes (formulated or 

4-5 
blank) into 12 L of dilution water and stirring the mixture for approximately 24 hours. The test 

water (without flakes) was then decanted into the test chambers into which the daphnids were 

placed. 

As with the studies in fish and the earlier studies with invertebrates, the concentration of 

disparlure in the test water was not measured. Consequently, the “concentrations” of disparlure 

are reported as nominal concentrations rather than measured concentrations. As detailed in U.S. 

EPA guidelines for the conduct of acute bioassays in Daphnia (U.S. EPA 1996), the U.S. EPA 

guidelines for toxicity studies in Daphnia require measurements of the concentrations of the test 

substance in water. The rationale for this requirement is simple: if the concentration is not 

measured, there may be substantial uncertainty in attempting to characterize the exposure. The 

distinction between nominal concentrations and measured concentrations is particularly 

important for compounds such as disparlure which have a very low solubility in water. As 

detailed further below, the nominal concentrations of disparlure in the toxicity studies of 

disparlure and Disrupt II flakes substantially exceed the water solubility. This leads, in turn, to 

the development of a film on the surface of the water and this film traps the daphnids. Thus, the 

effect, while adverse, appears to be a physical rather than toxic effect. 

As detailed in Appendix 3, the blank flakes – i.e., the flakes without disparlure – did not result in 

any mortality in any of the test groups for either the Standard Flakes (Palmer and Krueger 

2006a) or the Modified Flakes (Palmer and Krueger 2006b). The flakes from 2003 – both 

standard and modified – resulted in very low rates of mortality and immobility and the estimated 

LC50 values in both of these bioassays were >300 mg formulation/L, equivalent to >53 mg a.i./L. 

The new flakes from 2005 – again both standard and modified – yielded much lower estimates of 

the 48 hour-LC50: 69 mg formulation/ L (12.3 mg a.i./L) for standard flakes (Palmer and Krueger 

2006a) and 48 mg formulation/L (8.6 mg a.i./L) for modified flakes (Palmer and Krueger 

2006b). The reason or reasons for the differences between the 2003 flakes and the 2005 flakes is 

unclear and this issue is not addressed in the report by Palmer and Krueger (2006a,b) other than 

to note the differences in toxicities. For the standard flakes, Palmer and Krueger (2006a) note 

only the following differences in physical appearance: 

The SF 2003 and SF 2005 test solutions and the blank solution 

appeared clear and colorless in the test chambers at test initiation. 

At test termination, all of the solutions, with the exception of the 

300 mg/L SF 2005 solution, appeared clear and colorless. The 300 

mg/L SF 2005 test solution appeared clear and colorless with 

white particulates on the bottom of the test chamber. (Palmer and 

Krueger (2006a, p. 12.) 

For the modified flakes, Palmer and Krueger (2006b) note differences in appearance between the 

2003 and 2005 flakes that are somewhat more striking than those for the standard flakes: 

4-6 
Prior to decanting, the MF 2003 and MF 2005 WAF solutions, and 

the blank solution, appeared clear and colorless, with white 

particles on the surface of the water and green and white particles 

settled on the bottom of the WAF bottles, increasing in amount 

with increasing concentration. The MF 2003 and MF 2005 test 

solutions and the blank solution appeared clear and colorless in 

the test chambers at test initiation and termination. (Palmer and 

Krueger (2006b, p. 12.) 

During the period when these bioassays were being conducted, the testing facility was visited by 

a toxicologist with the USDA Forest Service who reported striking differences in the appearance 

of the 2003 and 2005 flakes, both standard and modified, prior to mixing the flakes with water 

(Appleton 2006). 

As detailed in Appendix 3, the recent bioassays on the flake formulations using daphnids 

(Palmer and Krueger 2006a,b) are similar to the earlier bioassay on technical grade disparlure 

using daphnids (LeBlanc et al. 1980) in that all of these studies observed daphnids trapped at the 

surface of the water. While LeBlanc et al. (1980) did not report the numbers of daphnids that 

were trapped at various nominal concentrations, the data reported by Palmer and Krueger 

(2006a,b) clearly indicate an association between the nominal concentrations, number of 

organisms trapped at the water surface, and subsequent mortality or immobility. 

The observations in these studies and the QSAR estimate of the very low water solubility of 

disparlure (Table 2-1) suggest that the trapping of the daphnids at the surface of the water was 

due to a layer of insoluble disparlure at the surface of the test water. Because no daphnids were 

trapped at the water surface in the bioassays on the blank flakes, both standard and modified, it is 

not plausible to assert that any of the inerts in either the standard or modified flakes contributed 

to the entrapment of the organisms at the water surface. 

When daphnids are trapped at the surface of the water, the organisms are under substantial stress 

and, if they remain trapped for a prolonged period, the animals may die for reasons that are not 

directly related to the systemic toxicity of the disparlure – e.g., impaired respiration. This is 

noted by Palmer and Krueger (2006a,b) in both sets of bioassays: 

Due to the nature of the test substance, mortality/immobility 

among daphnids in the Disrupt II formulation treatment groups 

may have been due, in part, to a physical effect, rather than only to 

toxicity. (Palmer and Krueger (2006a,b p. 15) 

As with fish, the weight of the evidence suggest that disparlure will not pose any risk to daphnids 

in terms of toxicity. Unlike fish, however, the available data clearly indicated that disparlure 

could pose a physical hazard to daphnids and possibly other aquatic invertebrates if the amount 

of disparlure in the water is sufficient to create an insoluble film of disparlure on the surface of 

the water. 

4-7 
While the hazard during a laboratory bioassay is clearly documented, the likelihood of this 

physical hazard occurring in the field after a normal application of disparlure is more difficult to 

assess. Disrupt II is not intentionally applied to water. While no microcosm or mesocosm 

studies have been conducted, Disrupt II as well as other experimental formulations of disparlure 

have been used by the USDA for over a decade. In that period, no incidents or field observations 

have been made that would suggest any adverse effects on aquatic invertebrates (Leonard 

2006c). The potential for a physical hazard to aquatic invertebrates is considered further in 

Section 4.4.4 (risk characterization for aquatic invertebrates). 

4.1.3.4. Aquatic Plants– Neither the published literature nor the U.S. EPA files include data 

regarding the toxicity of disparlure to aquatic plants. 

4.1.3.5. Other Aquatic Microorganisms– Neither the published literature nor the U.S. EPA files 

include data regarding the toxicity of disparlure to aquatic microorganisms. 

4-8 
4.2. EXPOSURE ASSESSMENT 

4.2.1. Overview 

As discussed in Sections 3.1 and 4.1, disparlure appears to be essentially nontoxic to mammals 

and birds. While this assessment is limited by the lack of chronic toxicity data in terrestrial 

species, it is not expected that acute or chronic exposure of terrestrial mammals or birds to 

disparlure would result in the development of significant adverse effects. Given the low toxicity 

of disparlure and limited available data, an exposure assessment for terrestrial species would not 

add to the assessment of risk for terrestrial species. Thus, an exposure assessment for terrestrial 

species is not included in this risk assessment. For aquatic species, the range of plausible 

nominal concentrations of disparlure in water are calculated at 0.0015 mg/L to 0.0037 mg/L over 

the range of applications rates considered in this risk assessment – i.e., 6 g a.i./acre to 15 g 

a.i./acre. These concentrations apply to a 1 meter deep body of water. The lower end of this 

range is within the estimated solubility of disparlure in water – i.e., 0.0019 to 0.0028 mg/L – and 

the upper end of this range slightly exceeds the estimated solubility of disparlure in water. 

4.2.2. Exposure of Aquatic Animals 

Disparlure is not intentionally applied to bodies of water (Hercon 2006a; Leonard 2006b). Thus, 

under normal conditions, aquatic organisms are not likely to be exposed to substantial amounts 

of disparlure. Accidental applications to surface water have been reported (Leonard 2006c) and 

these can be considered. 

Disrupt II flakes could be accidentally applied to either standing bodies of water (e.g., ponds or 

lakes) or moving bodies of water (e.g., streams or rivers). As discussed in Section 4.1.3, there is 

no basis for asserting that disparlure will pose any risk of toxic effects to aquatic organisms at 

the limit of estimated solubility of disparlure in water. The only risk that can be identified is the 

entrapment of small aquatic invertebrates in a surface film of disparlure (Section 4.1.3.3). A 

surface film of disparlure could occur if Disrupt II flakes were accidentally applied to a standing 

body of water, such as a lake or pond, in a sufficient amount to exceed the solubility of 

disparlure in the water. The development of a film in a flowing body of water, such as a stream 

or river, does not appear to be plausible. Consequently, for this risk assessment, exposure 

scenarios are developed only for standing bodies of water and these scenarios are used to assess 

potential effects only on small aquatic invertebrates that might interact with the surface of the 

water – i.e., benthic species are not considered to be at any risk. 

If Disrupt II flakes are applied to a standing body of water, some disparlure will volatilize into 

the air and some disparlure will leach from the flakes into the water. The disparlure in the water 

will diffuse through the water and a film of disparlure on the surface of the water will form if the 

water becomes saturated. The film on the surface of the water with then volatilize over time. 

The kinetics of these processes cannot be characterized. Nonetheless, the bioassays conducted 

by Palmer and Krueger (2006a,b) suggest that this general scenario is plausible. Thus, in the 

exposure assessment for small aquatic invertebrates, instantaneous leaching will be assumed and 

the impact of volatilization will not be considered. These are conservative assumptions in that 

4-9 
 

they will tend to overestimate exposure. This is considered further in Section 4.4.4 (risk 

characterization for aquatic invertebrates). 

As discussed in Section 2.3, this risk assessment considers application rates in the range of 6 

grams a.i./acre to 15 grams a.i./acre. This range corresponds to application rates of about 1.5 

mg/m2 [6 grams a.i./acre × 1000 mg/g ×1 acre/4047 m2 = 1.4826 mg/m2] to 3.7 mg/m2 [15 grams 

a.i./acre × 1000 mg/g ×1 acre/4047 m2

 = 3.7064 mg/m2]. If these amounts of disparlure are 

applied accidentally to a 1 meter deep body of water, nominal concentrations – i.e., assuming 

complete mixing and ignoring solubility limitations – would be in the range of 0.0015 mg/L to 

0.0037 mg/L [1000 liters per m3

]. Details of these calculations are given in Worksheet A01of 

the EXCEL workbook that accompanies this risk assessment. 

As noted in Table 2-1 and discussed in Section 4.1.3, no measured values for the solubility of 

disparlure in water are available but estimates based on quantitative structure-activity 

relationships developed by the U.S. EPA (EPI Suite 2006) suggest that the solubility of 

disparlure in water is in the range of 0.0019 to 0.0028 mg/L. Thus, the nominal concentrations 

that might occur in a 1 meter deep body of water after an accidental direct application are within 

the estimated water solubility of disparlure at the lower bound of the application rate (i.e., an 

application rate of 6 g a.i./acre) [0.0015 mg/L < 0.0028 mg/L] but modestly exceed the estimates 

of the solubility of disparlure in water at the upper bound of the application rate by a factor of 

about 1.3 [0.0037 mg/L ÷ 0.0028 mg/L]. 

Deeper bodies of water will result in lower concentrations that are likely to be at or below the 

solubility of disparlure in water and shallower bodies of water would lead to nominal 

concentrations that would exceed the solubility of disparlure in water. This type of situational 

variability is difficult to encompass in a general risk assessment. As a tool for individuals who 

are involved in or wish to assess applications of disparlure under conditions other than those 

considered in this risk assessment, the workbook that accompanies this risk assessment includes 

a worksheet (named A02) that can be used to calculate nominal concentrations of disparlure 

based on specified application rates, fractional deposition (i.e., drift), and average depth of the 

water body. Worksheet A02 also calculates hazard quotients based on the dose-response 

assessment for daphnids (Section 4.3.3). 

Note that Worksheet A02 applies only to the accidental application of disparlure to a standing 

body of water. No exposure scenarios are developed for accidents that involve the dumping of 

large amounts of Disrupt II into a standing body of water. While such accidents are possible, 

none have been documented. In addition, the calculation of nominal concentrations is trivial 

under the assumption of instantaneous mixing – i.e., the amount of disparlure that is deposited in 

the water divided by the volume of the water. Given the available information on the toxicity of 

disparlure to aquatic species (Section 4.1.3), no further elaboration of this exposure assessment is 

warranted. Potential consequences for aquatic species are discussed in Section 4.4.3 (risk 

characterization for fish) and Section 4.4.4 (risk characterization for aquatic invertebrates). 

4-10 
4.3. DOSE-RESPONSE ASSESSMENT 

4.3.1 Overview 

Given the low toxicity of disparlure to terrestrial animals coupled with the limitations imposed 

by the lack of chronic toxicity data, no standard dose-response assessment can be made or is 

warranted for disparlure in terms of effects on terrestrial species. As reviewed in Section 4.1.2.3, 

disparlure is produced by other species of moths and has the ability to attract nun moths (Gries et 

al. 2001, Morewood et al. 1999, Morewood et al. 1999, Schaefer et al. 1999). However, since 

there are no quantitative data available regarding the efficacy of disparlure in nun moths, a doseresponse assessment for this effect in a nontarget species cannot be made. Similarly, no explicit 

dose-response relationship is proposed for fish. There is no basis for asserting that adverse 

effects in fish are plausible under any foreseeable conditions. For aquatic invertebrates, there is 

no basis for asserting that toxic effects are likely at the limit of the solubility of disparlure in 

water. At nominal concentrations that exceed the solubility of disparlure in water, small 

invertebrates that may interact with the water-surface interface could become trapped in this 

interface due to a layer of undissolved disparlure at the air-water interface. 

4.3.2. Fish 

As discussed in Section 4.1.3.1, the available information on the toxicity of disparlure to fish are 

extremely limited. Nonetheless, there is no basis for asserting that disparlure is likely to pose a 

risk to fish at the limits of water solubility – i.e., in the range of 0.0019 to 0.0028 mg/L (Table 2­

1) – or at nominal concentrations that are substantially in excess of the solubility of disparlure in 

water. Consequently, no formal dose-response relationship for fish is proposed. Nonetheless, it 

is noted that a nominal concentration of 10 mg/L from the study by Rausina (no date) is a clear 

NOEC – see Appendix 3 for details and the discussion in Section 4.1.3.1. This nominal 

concentration is a factor of about 3,500 to over 5,000 above the estimated values for the 

concentration of disparlure in water. The implications of this range of values are discussed 

further in Section 4.4.3. 

4.3.3. Aquatic invertebrates 

The risk characterization for aquatic invertebrates is somewhat more complicated than that for 

fish. As with fish, there is no basis for asserting that toxic effects are likely in daphnids at the 

limit of water solubility. However, as discussed in Section 4.1.3.3, information is available from 

toxicity tests with daphnids of both technical grade disparlure (LeBlanc et al. 1980) as well as 

Disrupt II formulations of disparlure (Palmer and Krueger 2006a,b) that exposures to disparlure 

that exceed the solubility of disparlure in water will result in a film (presumably composed of 

undissolved disparlure) at the water surface. While this may not pose a toxic risk to daphnids, 

the toxicity studies demonstrate that these organisms can become trapped at the water surface 

and this can result in the death of the animal. 

The nominal concentrations at which entrapment is pronounced is in the range of the three higher 

nominal concentrations in the studies by Palmer and Krueger (2006a,b) using the Disrupt II 

formulations – i.e., a range of about 5.4 mg a.i./L to 54 mg a.i./L. The utility of these values are 

limited because the amount of disparlure that leached from the flakes used in these bioassays was 

not determined. On the other hand, these nominal concentrations may better reflect conditions 

4-11 
that could occur in the field – i.e., the processes of leaching from flakes to water as well as 

volatilization from the water surface to air. 

Lower values can be identified from the earlier study by LeBlanc et al. (1980) using technical 

grade disparlure. As indicated in Appendix 3, the minimum nominal concentration from the 

LeBlanc et al. (1980) study at which any mortality was noted is 0.028 mg/L. At this 

concentration, mortality was 1/15. Using the Fischer Exact test (see Section 3.1.5.2. in SERA 

2006), this incidence is not statistically significant (p = 0.5) and this concentration could be 

regarded as a NOEC. A similar case could be made for regarding higher concentrations from 

LeBlanc et al. (1980) as NOEC values: 0.048 mg/L (1/15 mortality, p = 0.5) and 0.079 mg/L 

(2/15 mortality, p = 0.241379). The clear LOAEL from the study by LeBlanc et al. (1980) is 

0.13 mg/L (12/15 mortality, p = 0.00000526). The clear NOEC from this study is 0.01 mg/L at 

which no mortality was observed. The major limitation in the study by LeBlanc et al. (1980) is 

that trapping of the daphnids at the water surface is noted but details comparable to those given 

in Palmer and Krueger (2006a,b) are not provided. 

For the current risk assessment, the NOEC value of 0.01 mg/L (nominal concentration) from the 

study by LeBlanc et al. (1980) will be used for characterizing risk. This is substantially above 

the estimated water solubility of disparlure – i.e., 0.0019 to 0.0028 mg/L from Table 2-1. As 

discussed above, the mortality observed in both the study by LeBlanc et al. (1980) as well as the 

studies by Palmer and Krueger (2006a,b) are probably due to the formation of a slick of 

disparlure at the surface of the water. Thus, the use of a nominal concentration is simply an 

index of exposure intended to suggest a slick that would be sufficiently minimal to cause no 

adverse effect even to small aquatic invertebrates. 

No dose-response assessment is proposed for larger aquatic invertebrates or benthic 

invertebrates. These aquatic invertebrates would not likely be trapped in (large invertebrates) or 

interact with (benthic species) any slick of disparlure on the surface of the water that might be 

associated with the application of Disrupt II flakes for the control or eradication of the gypsy 

moth. 

While the studies by Palmer and Krueger (2006a,b) are more recent and contain much more 

detailed information than is presented in the earlier study by LeBlanc et al. (1980), the Palmer 

and Krueger (2006a,b) studies are not used explicitly to derive toxicity values. The rationale for 

this approach is that the study by LeBlanc et al. (1980) does involve the application of known 

amount of disparlure to the test water. In the studies by Palmer and Krueger (2006a,b), detailed 

in Section 4.1.3.3, a known amount of Disrupt II flakes was applied to water and a fixed amount 

of time was allowed for the disparlure to leach from the flakes into the water. The amount of 

disparlure that actually leached from the flakes into the water, however, was not measured. In 

addition, the treated water was then decanted to arrive at the test water. The proportion of any 

leached disparlure that was decanted, however, cannot be determined. Thus, while both the 

LeBlanc et al. (1980) study and the studies by Palmer and Krueger (2006a,b) involved nominal 

rather than measured concentrations, the uncertainties in the exposure to disparlure are greater in 

the studies by Palmer and Krueger (2006a,b). While it may be argued that the Palmer and 

Krueger (2006a,b) studies might better approximate the impact of an application of Disrupt II 

4-12 
flakes, the Palmer and Krueger (2006a,b) studies did not involve actual exposures to the flakes. 

Thus, while the Palmer and Krueger (2006a,b) studies were well-designed and provide useful 

information, the earlier study by LeBlanc et al. (1980) involves fewer uncertainties in terms of 

the exposure of the daphnids to disparlure. 

4-13 
4.4. RISK CHARACTERIZATION 

4.4.1. Overview 

As discussed in Section 4.3.1, there is little data available on terrestrial and aquatic animals to 

allow for a quantitative characterization of risk in species other than rainbow trout and Daphnia. 

Furthermore, the lack of chronic toxicity data in any species adds significant uncertainty to any 

risk characterization. Thus, for both terrestrial and aquatic species, the potential for the 

development of toxicity from long-term exposure to disparlure cannot be assessed. Nonetheless, 

given the low toxicity of disparlure based on acute toxicity studies, it is unlikely that exposure to 

disparlure will result in the development of serious adverse effects in terrestrial and aquatic 

species. Regarding effects on terrestrial invertebrates, it is not likely that disparlure would 

disrupt mating of other species of moths that are native to North America (Section 4.1.2.3). 

Under normal conditions, aquatic species will not be exposed to substantial levels of disparlure. 

At the limit of the solubility of disparlure in water, there is no indication that toxic effects are 

likely in any aquatic species. If Disrupt II flakes are accidently applied over water, the amount 

of disparlure in the water could result in the formation of an insoluble layer of disparlure at the 

air-water interface. This would occur only in standing bodies of water (ponds or lakes) and not 

in flowing bodies of water such as streams or rivers. There is no indication that the formation of 

disparlure film in a standing body of water would impact fish. Based on toxicity studies 

conducted in the laboratory, small invertebrates that come into contact with the air-water 

interface might become trapped in this insoluble film. The likelihood of this occurring and the 

likelihood of this causing any detectable impact in a body of water is difficult to determine and 

would vary with the quantity of flakes applied to the body of water and the depth of the body of 

water. Based on variability in the experimental data as well as the range of application rates used 

in the USDA programs, hazard quotients would vary from about 0.15 to about 0.37, assuming a 1 

meter deep body of water, below the level of concern by factors of about 3 to 10. 

4.4.2. Terrestrial Species 

Based on the results of acute toxicity studies, the toxicity of disparlure to terrestrial mammals is 

very low (See Sections 3.1 and 4.1). However, the lack of chronic toxicity studies adds 

uncertainty to the risk characterization for all terrestrial species. Since results of acute toxicity 

studies in mammals and birds do not suggest that acute adverse effects are likely, it is not 

anticipated that exposure of these species to disparlure will results in the development of serious 

adverse effects in longer term exposures. However, since no chronic toxicity data are available, 

it is not possible to provide a characterization of risk for longer term exposure. 

For terrestrial invertebrates, specifically other species of moths, exposure to disparlure has the 

potential to disrupt mating. However, due to the lack of data, it is not possible to quantify this 

risk. 

4-14 
 

4.4.3. Fish 

As discussed in Section 4.1.3.1, the hazard identification for fish indicates that no toxic effects 

are plausible at the limit of the solubility of disparlure in water. In addition, toxicity studies in 

fish indicate no effects at nominal concentrations of disparlure in water that factors of about 

3,500 to over 5,000 above the estimated values for the concentration of disparlure in water 

(Section 4.3.2). The reciprocals of these ratios could be taken as approximate hazard indices – 

i.e., 0.0002 to 0.0003 – and these could be useful in comparing the risks posed by disparlure to 

risks posed by other agents. A somewhat clearer articulation of the risk characterization, 

however, is that no risks to fish can be identified under any foreseeable circumstances. 

4.4.4. Aquatic Invertebrates 

As with fish, there is no indication that disparlure will be toxic to aquatic invertebrates at the 

limit of the solubility of disparlure in water. Also as with fish, the probability of substantial 

exposure to disparlure is remote except in the case of accidental misapplication of Disrupt flakes 

directly to water. Thus, under normal conditions, no risks to aquatic invertebrates can be 

identified. 

The accidental application of Disrupt II flakes to water is plausible and, under some conditions, 

this could pose risks to aquatic invertebrates that interface with the water surface. This has been 

clearly demonstrated in laboratory studies with daphnids (Sections 4.1.3.3 and 4.3.3). As 

discussed in Section 4.2.2, accidental applications to surface water have been reported. If 

applied to rapidly moving water such as stream, there is no indication that adverse effects would 

be likely. If applied to standing water, however, concentrations calculated in Section 4.2.2 

modestly exceed the estimate of the solubility of disparlure in water at the upper range by a 

factor of about 3 – i.e., a nominal concentration of 0.0074 mg/L. If the amount of disparlure 

deposited on the surface of standing water exceeds the solubility of disparlure in water, a surface 

film could form and some small aquatic invertebrates could be trapped at the air-water interface. 

It seems unlikely, however, that this would lead to substantial or even detectable effects based on 

the clear NOEC value of 0.01 mg/L from the study by LeBlanc et al. (1980). As detailed in 

Worksheet A01 of the EXCEL workbook that accompanies this risk assessment, the highest 

calculated hazard quotient is 0.37 and is associated with the application of disparlure at a rate of 

15 g a.i./acre to a body of water that is 1 meter deep. The hazard quotient will vary directly with 

the depth of the water. Since the calculations are based on a 1 meter deep body of water, the 

hazard quotients would be a factor of 10 lower in a 10 meter deep body of water and a factor of 

10 higher in a 0.1 meter deep body of water. 

Whether or not the accidental application of disparlure flakes to any body of water would lead to 

a detectable effect is unclear. As noted in Section 4.1.3.3, no incidents or field observations have 

been made that would suggest any adverse effects on aquatic invertebrates (Leonard 2006c). 

However, the only report of an accidental application to water involves application to a river. As 

noted above, applications to flowing bodies of water would not be expected to result in any 

adverse effects. Nonetheless, based on the application rates used in vast majority of program 

activities (Section 2.3), hazard quotients for small aquatic invertebrates would exceed unity only 

in very shallow bodies of water. 

4-15 
The duration of any exposure to disparlure accidentally applied to water cannot be well 

characterized. As indicated in Appendix 4, the halftime of disparlure in water is estimated at 360 

hours (15 days) based on algorithms used in EPI Suite (Meylan and Howard 2000; U.S. 

EPA/OPPT 2000). These algorithms, however, rely on estimates of water solubility and Henrys 

Law constant. As also indicated in Appendix 4, experimental values for the water solubility and 

Henrys Law constant of disparlure are not available and are themselves estimated by EPI Suite 

based on molecular structure. This adds uncertainty to the estimated halftime in water. The 

halftime in water will also be influenced by site-specific conditions as well as the formulation of 

disparlure in the Disrupt II flakes, increasing the uncertainty in estimates from EPI Suite. 

4-16 
5. REFERENCES 

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D.C. Email: happleton@fs.fed.us. Personal communication via telephone with Patrick Durkin 

(SERA, Inc) on June 29, 2006. 

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with insect attractants. Toxicology and Applied Pharmacology 31: 421. 

Bintein S, Devillers J, and Karcher W. 1993. Nonlinear dependence of fish bioconcentration on 

n-octanol/water partition coefficient. SAR QSAR Environ Res. 1(1):29-39. 

Bierl BA. 1977?. Rate of emission of disparlure from laminated plastic dispensers as affected 

by temperature and air flow rate. (U.S. Dept. of Agriculture, Science and Education 

Administration, Agricultural Environmental Quality Institute, unpublished study; CDL:236537­

E). MRID 00047219 

Bierl-Leonhardt BA; DeVilbiss ED; Plimmer JR. 1979. Rate of release of disparlure form 

laminated plastic dispensers. J Econ Entomol 72:319-321. 

Cameron EA. 1981. On the persistence of disparlure in the human body. Journal of Chemical 

Ecology 7(2): 313-318. 

Cameron EA. 1983. Apparent long-term bodily contamination by disparlure, the gypsy moth 

(Lymantria dispar) attractant. Journal of Chemical Ecology 9(1): 33-37. 

Cameron EA. 1995. On the apparent persistence of disparlure in the human body. J Chem Ecol 

21(4):385-386. 

Campbell RW. 1983. Gypsy moth (Lepidoptera: Lymantriidae) control trials combining 

nucleopolyhedrosis virus, disparlure, and mechanical methods. Journal of Economic 

Entomology 76(3): 610. 

Caro JH; Bierl BA; Freeman HP; Glotfelty DE; Turner BC. 1977. Disparlure: Volatilization 

rates of two microencapsulated formulations from a grass field. Environ Entom 6(6):877-881. 

Caro JH; Freeman HP; Brower DL; Bierl-Leonhardt BA. 1981. Comparative distribution and 

persistence of disparlure in woodland air after aerial application of three controlled-released 

formulations (Lymantria dispar, gypsy moth sex pheromone). Journal of Chemical Ecology 

7(5): 867-880. 

5-1 
Clements RG, Nabholz JV, and Zeeman M. 1996. Estimating Toxicity of Industrial Chemicals 

to Aquatic Organisms Using Structure-activity Relationships. Environmental Effects Branch, 

Health and Environmental Review Division, Office of Pollution Prevention and Toxics, U.S. 

Environmental Protection Agency. Report dated August 30, 1996. 

Coleman D. 2000. Racemic disparlure Acute Oral Toxicity to the Rat (Acute Toxic Class 

Method). (Disparlure Technical): Lab Project Number: SHE 032/003649/AC. Unpublished 

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45529801 

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(safety) factors. Regulatory Toxicology and Pharmacology 3: 224-238. 

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Specific Commentary on Glyphosate, Triclopyr, and Hexazinone: Final Report. SERA TR 01­

43-08-04a dated January 30, 2002. Available at www.fs.fed.us/foresthealth/pesticide/risk.htm. 

EPI Suite. 2006. Estimates of Chemical and Physical Properties of Disparlure Based on EPI 

Suite Version 3.12. EPI Suite runs conducted by Patrick Durkin, SERA, Inc. on June 28, 2006. 

Copy of EPI Suite run include with this risk Assessment as Appendix 4. 

Fink R; Beavers JB; Joiner G; et al. 1980. Final Report: Acute Oral LD50--Bobwhite Quail: 

Project No. 173-102. (Unpublished study received Nov 19, 1980 under 36638-5; prepared by 

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Model, OilToxEx. Environmental Toxicology and Chemistry. 21(10): 2080–2094. 

Grapenthien JR. 1972. Report to United States Department of Agriculture: Acute Aerosol 

Inhalation Toxicity Study with Dispar- lure in Albino Rats: IBT No. A1958. (Unpublished study 

received Feb 21, 1980 under 36638-5; prepared by Industrial Bio-Test Lab- oratories, Inc., 

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MRID 00059821 

Gries G; Schaefer PW; Gries R; Liska J; Gotoh T. 2001. Reproductive character displacement 

in Lymantria monacha from northern japan? J Chem Ecol. 27(6): 1163-76. 

Hercon Environmental. 1993. Letter from Priscilla MacLean (Hercon) to Noel Schneeberger 

dated June 22, 1993. Includes Product Label and Material Safety Data Sheet for Disrupt II. 

Hercon Environmental. 2004. Hercon Disrupt II Fact Sheet. Copy courtesy of Donna Leonard, 

USDA Forest Service, Forest Health Protection, PO Box 2680, Asheville, NC 28802. e-mail: 

dleonard@fs.fed.us. Received July 19, 2004. 

5-2 
Hercon Environmental. 2006a. Hercon Disrupt II Product Label. Copy courtesy of Donna 

Leonard, USDA Forest Service, Forest Health Protection, PO Box 2680, Asheville, NC 28802. 

e-mail: dleonard@fs.fed.us. Received June 27, 2006. 

Hercon Environmental. 2006b. Hercon Disrupt II Material Safety Data Sheet. Copy courtesy of 

Priscilla MacLean, Product Development Manager, Hercon Environmental, P.O. Box 435, 

Emigsville PA, 17318. e-mail: pmaclean@herconenviron.com. Received June 27, 2006. 

Herculite Products Incorporated. 1978. Hercon Luretape with disparlure: Gypsy Moth 

Pheromone Dispensers. (Reports by various sources; unpublished study including published 

data, received Dec 1, 1978 under 8730-17; CDL:236537-L). MRID 00047223 

Hudson RH. 1975. Report on the Study To Determine the LC50 of disparlure for Mallards and 

Bobwhite Quail. (U.S. Fish and Wildlife Service, unpublished study; CDL:236537-N). MRID 

00047225 

Jacobson, M. 1976. Impact of natural plant protectants on the environment. In: Marini-Bettolo, 

G.B., ed. Natural products and the protection of plants: proceedings of a study week at the 

Pontifical Academy of Sciences, Oct. 18-23, 1976. Amsterdam: Elsevier Scientific Publishing 

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Jeppsson R. 1975. Parabolic Relationship between Lipophilicity and Biological Activity of 

Aliphatic Hydrocarbons, Ethers and Ketones after Intravenous Injections of Emulsion 

Formulations into Mice. Acta Pharmacol. Et Toxicol. 37: 56-64. 

Kolodny-Hirsch DM; Webb RE. 1993. Mating disruption of gypsy moth Lepidoptera 

Lymantriidae following ground application of high rates of racemic disparlure. Journal of 

Economic Entomology 86(3): 815-820. 

Knapp T; Terrell Y. 1980. Static 96-hour Toxicity Study of Neat Gypsy Moth Pheromone in 

Bluegill Sunfish: Laboratory No. OF- 7473. (Unpublished study received Nov 19, 1980 under 

36638-5; prepared by Cannon Laboratories, Inc. submitted by Albany International, Controlled 

Release Div., Needham Heights, MA; CDL:244731-A). MRID 00127869 

Kretchmar B. 1972. Report to United States Department of Agriculture: Acute Toxicity Studies 

with disparlure: IBT No. A1958. (Unpublished study received Nov 21, 1972 under 19750-1; 

prepared by Industrial Bio-Test Laboratories, Inc., submitted by American Can Co., Rahway, NJ; 

CDL:004647-B). MRID 00140660 

LeBlanc G; Surprenant D; Sleight B; et al. 1980. Acute Toxicity of a Gypsy Moth Mating 

Disruption Pheromone, Active Ingredient Cis-7, 8-epoxy-2-octadene to the Water Flea : Report 

#BW-80-8-715. (Unpublished study received Nov 19, 1980 under 36638-5; prepared by EG & 

G, Bionomics, submitted by Albany International, Controlled Release Div., Needham Heights, 

MA; CDL:244730-A). MRID 00127868. 

5-3 
 

 

 

Leonard D. 2004. Comments on SERA TR 04-43-05-04a, Control/Eradication Agents for the 

Gypsy Moth – Human Health and Ecological Risk Assessment for Disparlure – Peer Review 

Draft. Comments by Donna Leonard, USDA Forest Service, Forest Health Protection, 

Asheville, NC. Comments received via email from dleonard@fs.fed.us. 

Leonard D. 2006a. Comments on Application Rates for Disparlure in STS (Slow-The-Spread) 

Programs. Comments by Donna Leonard, USDA Forest Service, Forest Health Protection, 

Asheville, NC. Comments received via email from dleonard@fs.fed.us on June 27, 2006. 

Leonard D. 2006b. Comments on The Use of Disparlure in STS (Slow-The-Spread) Programs. 

Comments by Donna Leonard, USDA Forest Service, Forest Health Protection, Asheville, NC. 

Comments received via email from dleonard@fs.fed.us on June 27, 2006. 

Leonard D. 2006c. Field Observations in Applications of Disparlure. Comments by Donna 

Leonard, USDA Forest Service, Forest Health Protection, Asheville, NC. Comments received 

via email from dleonard@fs.fed.us on June 29, 2006. 

Leonard D. 2006d. Review of SERA TR 06-52-02-01a, Control/Eradication Agents for the 

Gypsy Moth - Human Health and Ecological Risk Assessment for Disparlure (a.i.) and Disrupt II 

formulation – REVISED FINAL REPORT. USDA Forest Service, Forest Health Protection, 

Asheville, NC. Comments received via email from dleonard@fs.fed.us on July 12, 2006. 

Leonard D. 2006e. Comments on Disrupt formulations by Donna Leonard, USDA Forest 

Service, Forest Health Protection, Asheville, NC. Comments received via email from 

dleonard@fs.fed.us on August 24, 2006. 

Leonhardt BA; Mastro VC; Paszek EC; Schwable CP.; DeVilbiss ED. 1990. Dependence of 

gypsy moth (Lepidoptera: Lymantriidae) capture on pheromone release rate form laminate and 

other dispensers. J Econ Entomol. 83(5):1977-1981. 

Leonhardt BA; Mastro VC; Devilbiss ED. 1993. New dispenser for the pheromone of the gypsy 

moth Lepidoptera Lymantriidae. Journal of Economic Entomology 86(3): 821-827. 

Leonhardt BA; Mastro VC; Leonard DS; McLane W; Reardon RC; Thrope DS. 1996. Control 

of low-density gypsy moth (Lepidoptera: Lymantriidae) populations by mating disruption with 

pheromone. J Chem Ecology 22:1255-1272. 

MacLean P. 2006. Comments on Inerts in Disrupt II, Product Development Manager, Hercon 

Environmental, P.O. Box 435, Emigsville PA, 17318. e-mail: pmaclean@herconenviron.com. 

Received June 27, 2006. 

Meylan W; Howard P. 2000. Estimation Program Interface, Version 3.12. Syracuse Research 

Corporation, Syracuse, N.Y. for U.S. Environmental Protection Agency, Office of Pollution, 

Prevention and Toxics, Washington D.C. Downloadable copy of EPI-SUITE computer program 

available at: http://www.epa.gov/opptintr/exposure/docs/episuite.htm 

5-4 
 

Moody DE; Montgomery KA; Ashour MB; Hammock BD. 1991. Effects of environmentally 

encountered epoxides on mouse liver epoxide-metabolizing enzymes. Biochem Pharmacol 1991 

Jun 1;41(11):1625-37. 

Morewood P; Gries G; Haubler D; Moller K; Liska J; Kapitola P; Bogenschutz H. 1999. 

Towards pheromone-based detection of Lymantria monarcha (Lepidoptera: Lymanstriidae) in 

North America. The Canadian Entomogolsit 131:687-694. 

Morewood P; Gries G; Haubler D; Moller K; Liska J; Kapitola P; Bogenschutz H. 2000. 

Towards pheromone-based monitoring of nun moth, Lymantria monacha (L.) (Lep., 

Lymantriidae) populations. J Appl Ent 124:77-85. 

MTM Chemicals. 1991. Disparlure Material Safety Data Sheet. Prepared by MTM Chemicals, 

Inc. 1970 Atlas Street, Columbus, OH. 

Murlis J; Willis MA; Carde RT. 2000. Spatial and temporal structures of pheromone plumes in 

fields and forests. Physiol Entomol 25:211-222. 

NRC (National Research Council). 1983. Risk assessment in the Federal government: 

managing the process. Washington, DC: National Academy Press; 176 p. + app. 

Oguma Y. 1998. Mutagenicity Testing of 7,8-Epoxy-2-Methyloctadecane in Bacterial Reverse 

Mutation Assays: Lab Project Number: 6128. Unpublished study prepared by BML, Inc. 19 p. 

{OPPTS 870.5265}. MRID 45309502 

Oguma Y. 2000. Product Chemistry of Gypsy Moth Pheromone. (disparlure). Unpublished 

study prepared by Shin-Etsu Chemical Company. 22 p. MRID 45309501 

Palmer SJ; Krueger HO. 2006a. SF 2003 and SF 2005: A 48-Hour Static-Renewal Acute 

Toxicity Test with the Cladoceran (Daphnia magna). Wildlife International, Ltd. Project 

Number: 6 L4a- 102. Study completion date: Jan. 12, 2006. Copy courtesy of Paul Mistretta, 

USDA/FS. 

Palmer SJ; Krueger HO. 2006b. MF 2003 and MF 2005: A 48-Hour Static-Renewal Acute 

Toxicity Test with the Cladoceran (Daphnia magna). Wildlife International, Ltd. Project 

Number: 6 L4a- 101. Study completion date: Jan. 12, 2006. Copy courtesy of Paul Mistretta, 

USDA/FS. 

Pelletier MC; Burgess RM; Ho KT; Kuhn; McKinney RA; Ryba SA. 1997. Phototoxicity of 

Individual Polycyclic Aromatic Hydrocarbons and Petroleum to Marine Invertebrate Larvae and 

Juveniles. Environmental Toxicology and Chemistry. 16(10): 2190-2199. 

5-5 
 

 

Plettner E; Lazar J; Prestwich EG; Prestwich GD. 2000. Discrimination of pH enantiomers by 

two pH binding proteins from the gypsy moth Lymantria dispar. Biochemistry 2000 Aug 

1;39(30):8953-62. 

Plimmer JR; Schwalbe CP; Paszek EC; et al. 1977. Contrasting effectiveness of (+) and (-) 

enantiomers of disparlure for trapping native populations of gypsy moth in Massachusetts. 

Environmental Entomology 6(4):518-522. (Also~In~unpublished submission received Aug 12, 

1981 under 8730-31; submitted by Herculite Products, Inc., New York, N.Y.; CDL:245766-D). 

MRID 00080044 

Rausina G. No Date. Report to United States Department of Agriculture, Agricultural Research 

Service, Agricultural Environmental Quality Institute: Results of Four-Day Static Fish Toxicity 

Studies: Rainbow Trout and Bluegills: IBT No. A-1958. (Unpublished study received Jan 22, 

1975 under 11312-7; prepared by Industrial Bio-Test Laboratories, Inc., submitted by U.S. Dept. 

of Agriculture, Washington, D.C.; CDL:228392-B). MRID 00059735. [Note: This study is 

catalogued by the U.S. EPA with a date of 1949 but disparlure had not been identified in 1949. 

The 1949 date used by U.S. EPA appears to be associated with a citation to statistical methods 

used by Rausina rather than the date of the study. The submission by Rausina does not indicate 

the date for the conduct of the study.] 

Schaefer PW; Gries G; Gries R; Holden D. 1999. Pheromone components and diel periodicity 

of pheromonal communication in Lymantria funida. J Chem Ent 25(10):2305-2312. 

Schwalbe C; Paszek E; Webb R; et al. 1978. Field Evaluation of Controlled Release 

Formulations of disparlure for Gypsy Moth Mating Disruption. (Unpublished study received 

Apr 27, 1979 under 36638-EX-2; prepared by U.S. Dept. of Agriculture, Animal and Plant 

Health Inspection Service, Gypsy Moth Methods Development Center and Others, submitted by 

Albany International, Controlled Release Div., 

Schwalbe C; Paszek E; Webb R; Bierl-Leionhardt BA; Plimmer JR; McComb CW; Dull CW. 

1979. Field evaluation of controlled release formulations of disparlure for gypsy moth mating 

disruption. J Econ Entomol 72:322-236. 

SERA (Syracuse Environmental Research Associates, Inc.). 2004. Control/Eradication Agents 

for the Gypsy Moth - Human Health and Ecological Risk Assessment for Disparlure (a.i.) - 

FINAL REPORT, SERA TR 04-43-05-04b, reported dated August 27, 2004. Syracuse 

Environmental Research Associates, Inc., Fayetteville, NY. 

SERA (Syracuse Environmental Research Associates, Inc.). 2006. Preparation of Environmental 

Documentation and Risk Assessments, SERA MD 2006-01a, draft dated March 3, 2006. 

Syracuse Environmental Research Associates, Inc., Fayetteville, NY. Available at www.serainc.com 

Shin-Etsu Chemical Co Ltd. 2002. Submission of Product Chemistry Data in Support of the 

Application for Registration of disparlure Technical. MRID 45810500 

5-6 
 

 

Taylor AW. 1982. Field measurements of pheromone vapor distribution. In: Leonhardt, B.A.; 

Beroza, M., eds. Insect pheromone technology: chemistry and applications. Washington, DC: 

American Chemical Society. ACS Symposium Series 190; 193-207. 

Thwaits BF; Sorensen PW. 2005. Olfactory sensitivity of rainbow trout to racemic disparlure. 

Unpublished synopsis dated April 1, 2005. Copy courtesy of Donna Leonard, USDA/Forest 

Service. 2 pp. 

Thorpe WE; Ridgway RL; Leonhardt BA. 1993. Relationship between gypsy moth Lepidoptera 

Lymantriidae pheromone trap catch and population density comparison of traps baited with 1 

and 500 Micro Dextro disparlure lures. Journal of Economic Entomology 86(1): 86-92. 

Thorpe WE; Leonhardt BA; Mastro VC; Reardon RC; Sellers P; Webb RE; Talley SE. 1999. 

Effectiveness of gypsy moth mating disruption from aerial applications of plastic laminate flakes 

with and without a sticking agent. TEKTRAN, www.nal.usda.gov/ttic/tektran/data/ 

000010/74/0000107476.html. 

Tobin PC; Leonard DS. 2006. Estimating Pheromone Released by Female Gypsy Moths During 

an Outbreak and Comparing this with Racemic Disparlure Released after an Application of 

Disrupt II. Unpublished analysis dated August 2, 2006. Copy courtesy of Donna Leonard, 

USDA Forest Service, Forest Health Protection, PO Box 2680, Asheville, NC 28802. e-mail: 

dleonard@fs.fed.us. Received August 7, 2006. 

USDA (U.S. Department of Agriculture). 1995. Gypsy Moth Management in the United States: 

A Cooperative Approach. Final Environmental Impact Statement. Appendix F (Human Health 

Risk Assessment) and Appendix G (Risk Assessment). 

USDA/FS (U.S. Department of Agriculture/Forest Service). 2005. Gypsy Moth Digest. 

Available at: http://na.fs.fed.us/fhp/gm/index.shtm 

U.S. EPA (U.S. Environmental Protection Agency). 1994. Arthropod pheromones in solid 

matrix dispensers; experimental use permits. Federal Register 59(17): 3681-3684. 

U.S. EPA (U.S. Environmental Protection Agency). 1996. Ecological Effects Test 

Guidelines OPPTS 850.1010: Aquatic Invertebrate Acute Toxicity Test, Freshwater Daphnids. 

EPA 712–C–96–114 dated April 1996. U.S. EPA Office of Prevention, Pesticides and Toxic 

Substances. Available at: http://www.epa.gov/opptsfrs/publications/OPPTS_Harmonized/ 

850_Ecological_ Effects_Test_Guidelines/Drafts/850-1010.pdf. 

U.S. EPA (U.S. Environmental Protection Agency). 2003. Toxicity Categories and Pesticide 

Label Statements. Available at: http://www.epa.gov/pesticides/health/tox_categories.htm 

5-7 
 

U.S. EPA/OPPTS (U.S. Environmental Protection Agency/Office of Prevention, Pesticides, and 

Toxic Substances). 1996. Ecological Effects Test Guidelines. OPPTS 850.1075: Fish Acute 

Toxicity Test, Freshwater and Marine. Available at: 

http://www.epa.gov/opptsfrs/home/guidelin.htm 

U.S. EPA/OPPT (U.S. Environmental Protection Agency/Office of Pollution Prevention and 

Toxics). 2000. On-Line EPI Suite User's Guide, Version 3.12. Developed by the EPA's Office 

of Pollution Prevention Toxics and Syracuse Research Corporation (SRC). Available at: 

http://www.epa.gov/opptintr/exposure/docs/episuite.htm 

U.S. EPA/OPPTS (U.S. Environmental Protection Agency/Office of Prevention, Pesticides, and 

Toxic Substances). 2006. Harmonized Test Guidelines. Available at 

http://www.epa.gov/opptsfrs/home/guidelin.htm 

Ward GS. 1981. Acute Toxicity of a Synthetic Gypsy Moth Pheromone to Eastern Oysters 

(Crassostrea virginica): Report No. BP-81-3-31. (Unpublished study received Apr 17, 1981 

under 36638-5; prepared by EG & G Bionomics, submitted by Conrel, an Albany International 

Co., Needham Heights, Mass.; CDL:244882-A). MRID 00074291 

5-8 
 

 

 

 

Table 2-1. Identification and Physical/Chemical Properties of Disparlure. 

Property Value a	 Reference 

CAS Number 029804-22-6 EPI Suite (2006) 

Smiles Notation O(C1CCCCCCCCCC)C1CCCCC(C)C EPI Suite (2006) 

U.S. EPA Registration Number 8730-55 Hercon Environmental, 2004 

MW 282.51 EPI Suite (2006)
 

Henry’s Law Constant (atm 0.015 to 0.061 EPI Suite (2006)
 

m3

/mole) 

Vapor pressure (mm Hg) 0.00021 to 0.00034 EPI Suite (2006)
 

Water solubility (mg/L) 0.0019 to 0.0028 EPI Suite (2006)
 

log Ko/w 8.08 EPI Suite (2006)
 

Ko/c (acid, ml/g) 3.44 × 104 EPI Suite (2006)
 

Halftimes in water (days) 0.074 (river) EPI Suite (2006)
 

6.9 (lake) 

Halftimes in other media (days) 0.5 (air) EPI Suite (2006) 

15 (water) 

30 (soil) 

135 (sediment) 

a

 For many estimates, EPI Suite provides more than one estimate based on different estimation methods. When 

more than one estimate is provided, the range of values are given. Estimates from EPI Suite are often present out 

to several decimal places. Except for molecular weight, all values in this table are rounded to two significant 

places. 

Tables - 1 
Table 2-2: Use of Disparlure by the USDA to control the North American Gypsy Moth from 

1995 to 2005 by Type of Use (USDA/FS 2005) 

Year Acres Treated for Eradication Acres Treated to Slow the Spread 

1995 0 2,448 

1996 5,352 16,621 

1997 0 10,808 

1998 7,120 21,418 

1999 38,980 19,360 

2000 7,988 93,625 

2001 0 212,925 

2002 650 542,600 

2003 0 647,394 

2004 250 588,256 

2005 0 287,890 

Tables - 2 
 

 

 

 

 

Table 3-1: Summary of acute toxicity data of Disparlure in mammals (all values are expressed in 

terms of a.i.) 

Species Exposure/Dose Effect Reference 

rat single oral doses ranging from 

10,250 – 34,600 mg/kg 

LD50 > 34,600 mg/kg 

NOAEL (mortality) = 34,600 

mg/kg 

Kretchmar 1972 

rat single oral dose of 5000 mg/kg LD50 > 5,000 mg/kg 

NOAEL (mortality) = 5,000 mg/kg 

Coleman 2000 

rat inhalation exposure, 5.0 mg/L in 

air for 1 hour 

LD50 > 5 mg/L air 

NOAEL (mortality) = 5.0 mg/L air 

Grapenthien 1972 

rabbit dermal toxicity testing a single 

dose of 2,025 mg/kg 

LD50 > 5,000 mg /kg 

NOAEL (mortality) = 5,000 mg/kg 

Kretchmar 1972 

rabbit primary skin irritation testing a 

single dose of 0.5 g 

Not a skin irritant (only very mild 

skin irritation) 

Kretchmar 1972 

rabbit primary eye irritation testing a 

single dose of 0.1 g/eye 

not an eye irritant Kretchmar 1972 

Tables - 3 
 

 

 

 

 

 

Table 4-1: Summary of acute toxicity data of Disparlure in avian and aquatic species (all values 

are expressed in terms of a.i.) 

Species Exposure/Dose Effect Reference 

bobwhite quail single oral doses ranging from 

398 to 2510 mg/kg (by gavage) 

LD50 > 2510 mg/kg Fink et al. 1980 

bobwhite quail 

chicks 

313 to 5000 in diet for 5 days LD50 > 5000 ppm Hudson 1975 

mallard ducklings 313 to 5000 in diet for 5 days LD50 > 5000 ppm Hudson 1975 

bluegill sunfish a 300 mg/L for 96 hours LC50 > 300 mg/L Knapp and Terrell 

1980 

bluegill sunfish a 0.1 to 100 pm for 96 hours LC50 > 100 mg/L Rausina No Date 

rainbow trout a 0.1 to 100 pm for 96 hours LC50 > 100 mg/L 

NOEC = 10 mg/L 

Rausina No Date 

Daphnia a, b 0.01 to 0.22 mg/L for 96 hours LC50 > 0.098 mg/L 

NOEC = 0.017 mg/L 

LeBlanc et al. 1980 

Eastern oysters a 1.25 to 20 mg/L for 96 hours NOEC (new shell growth) = 

20 mg/L 

Ward 1981 

a

 All values expressed a nominal rather than measured concentrations. See Section 4.1.3.3 for a discussion of the 

significance of nominal versus measured concentrations. 

b

 Additional studies in Daphnia using water accommodated fractions of Disrupt II formulations have been 

conducted by Palmer and Krueger (2006a,b). The nominal concentrations reported in this study are not 

comparable to those reported above. See Section 4.3.3 for a more detailed discussion. 

Tables - 4 
 

 

 

 

 

Table 4-2. Summary of QSAR Toxicity Estimates for Disparlure to Aquatic Species and 

Algorithms for Estimating the Toxicity of Mono-Epoxide Compounds to Aquatic Species 

Developed by Clements et Al. (1996). 

Type of Estimate 

(Species) 

Slope Intercept r

2

 (n) a Limiting 

Log10 Kow b 

Estimated 

LC50 

mg/L 

Freshwater Acute 

Fish, 96h-LC50 (Fathead minnow) 0.382 -0.29 0.92 (4) 5 0.119 

Fish, 16 day (Guppy) 0.246 -0.5 0.87 (9) 5 0.144 

Invertebrate, 48h-LC50 (Daphnia) -0.567 0.036 1.0 (2) 5 0.008 

a

 Squared correlation coefficient and number of data points in analysis. b

 These values are reported in the output of EPI Suite Version 3.12. Slightly different values are reported in
 

Clements et al. (1996). 

Tables - 5 
LIST OF APPENDICES 

Appendix 1: Acute toxicity of disparlure to experimental mammals 

Appendix 2: Toxicity of disparlure to birds 

Appendix 3: Toxicity of disparlure aquatic species 

Appendix 4: EPI Suite Output for Disparlure 
 

 

Appendix 1: Toxicity of disparlure to experimental mammals (Unless otherwise specified, all concentrations 

are expressed as a.i.) 

Animal	 Dose/Exposure Response Reference 

ORAL - ACUTE 

rats, Sprague-Dawley	 single dose of 5000 mg 

5 males, 5 females a.i./kg (racemic 

preparation) by gavage. 

Animals observed for 15 

days. 

No control group. 

rats, Sprague-Dawley	 single dose of test 

albino material administered at 

several dose levels 

(10250, 15380, 23070, 

34600 mg/kg) by gavage. 

Rats observed for 14 

days following 

administration. No 

control group. 

DERMAL 

rabbits, New Zealand 2025 mg/kg test material 

applied to shaved skin 

and occluded for 24 

hours. Animals observed 

for 14 days for systemic 

toxicity 

No mortalities. No 

microscopic abnormalities 

observed. 

Clinical signs of toxicity were 

piloerection, hunched posture 

and ungroomed appearance 

appearing on Day 1 of 

treatment. All signs were 

resolved by Day 4 of the 

observation period. 

LD50 > 5000 mg a.i./kg 

No mortality at any dose 

level. 

No gross pathological lesions 

at any dose level. 

At all dose levels, 

hypoactivity, ruffed fur, and 

diuresis were observed, 

LD50 > 34600 mg a.i./kg 

No mortalities. No gross 

pathologic lesions on 

necropsy. 

Local skin irritation after 24 

hours (erythema and edema). 

7 days after dosing, 

escharosis, desquamation, 

hemorrhaging and fissures. 

After 14 days, desquamation, 

fissures and pustules 

LD50 > 2025 mg a.i./kg 

Coleman 2000 

MRID 45529801 

Beroza et al. 1975 

Hercon 1978 

Kretchmar 1972 

MRID 00128026 

Beroza et al. 1975 

Hercon 1978 

Kretchman 1972 

MRID 00128026 

Appendix 1-1 
 

 

 

Appendix 1: Toxicity of disparlure to experimental mammals (Unless otherwise specified, all concentrations 

are expressed as a.i.) 

Animal	 Dose/Exposure Response Reference 

0.5 mL of undiluted test 

rabbits, New Zealand material (0.5 g) applied 

to shaved skin and 

occluded for 24 hours. 

Animals were observed 

for 72 hours 

EYES 

6 young rabbits, New	 0.1 mL undiluted sample 

Zealand (0.1 g) applied to 

conjunctival sac. Eye 

was not washed. 

Severity of ocular lesions 

was monitored at 

intervals of 24, 48, and 

72 hours. Rabbits 

observed for 7 days. 

INHALATION 

Albino rats (10) Inhalation chamber 

study. Disparlure 

concentration 5.0 mg/L 

in air for 1 hour 

Primary dermal irritation 

study. 

Mild skin irritation (erythema 

and edema) was noted at 24 

and 72 hours after application 

of test material 

3/6 rabbits had conjunctival 

redness at 24 hours. 

No effects observed in any 

rabbits at later times of the 

observation period 

No deaths were observed in 

this study. No assessment of 

sublethal toxicity was made 

LC50>5.0 mg a.i./L air 

Beroza et al. 1975 

Hercon 1978 

Kretchman 1972 

MRID 00128026 

Beroza et al.1975 

Hercon 1978 

Kretchman 1972 

MRID 00128026 

Grapenthien 1972 

MRID 00059821 

Appendix 1-2 
 

 

 

Appendix 2: Toxicity of disparlure to birds (unless otherwise specified, all doses and 

concentrations are expressed in terms of a.i.) 

Animal Dose/Exposure Response Reference 

bobwhite quail (5 months 

old)
 

bobwhite quail (12 day old
 

chicks) 

mallard ducks (15 day old 

ducklings) 

Single oral doses of 398, 

631, 1590, and 2510 

mg/kg bw. Birds 

observed for 7 days after 

dosing 

Dietary exposure to 313, 

625, 1250,2500, 5000 

ppm for 5 days. Birds 

observed for 3 days after 

end of dosing period 

No mortalities at any dose 

level. No signs of toxicity 

associated with test material. 

At the highest dose, lethargy 

was observed in 3/10 birds on 

days 1-2 after dosing. 

Unclear if lethargy was 

related to test material. 

LD50 > 2510 mg/kg 

No mortalities in at any dose 

level for either species 

No signs of toxicity reported 

LC50 > 5000 ppm in diet for 

both quail and ducks 

Fink et al. 1980 

MRID 00083102 

Hudson 1975 

MRID 00105981 

same data reported 

in MRID 00047225 

Appendix 2-1 
 

 

 

 

 

 

 

 

 

Appendix 3: Toxicity of disparlure to aquatic species (unless otherwise specified, all 

concentrations are expressed in terms of a.i.) 

Animal Dose/Exposure Response Reference 

FISH 

Rainbow trout 

Bluegills, 10 fish 

per concentration 

Bluegill sunfish, 

30 fish in each 

group 

0.1, 1.0, 10.0, 100.0 ppm (mg 

a.i./L) for 96 hours. Survival 

assessed at 1-6, 24, 48, 72, and 96
 

hours. 

Note: Very poor quality fiche. 

Dissolved oxygen was measured in 

the test water only when mortality 

was observed. The measurement 

itself cannot be read from the fiche. 

Nominal concentration of 0 ppm 

(untreated control) and 300 ppm for 

96 hours. No aeration during the 

study. 

No description of how the test 

water was prepared. No discussion 

of any observations concerning a 

surface film on the water. 

No effect on dissolved oxygen. 

In bluegills, no affect on survivors 

at any concentration up to 96 hr 

exposure. 

LC50>100 ppm 

In Rainbow trout, for all 

concentrations, no affect on 

survivors up to 48 hours. At the 

100 ppm concentration, the number 

of survivors decreased to 8/10 after 

72 hours of exposure. 

LC50>100 ppm 

No mortalities observed and no 

signs of altered behavior. 

Dissolved oxygen in test water and 

control water were comparable: 

Day 1 11.0 ppm (control) 

10.4 ppm (test water) 

Day 4: 3.4 ppm (control) 

 3.4 ppm (test water) 

pH constant in test and control 

water (pH 6.4) of the duration of 

testing. 

LC50>300 ppm 

AQUATIC INVERTEBRATES
 

Technical Grade Disparlure 

Eastern oysters 96 hour exposure to concentrations No affect on new shell growth at 

(Crassostrea ranging from 1.25 to 20 ppm 92% any concentration 

virginica) disparlure 

NOEC > 20 ppm 

Acetone concentrations ranged up 

to 10% 

Rausina No 

Date 

MRID 

00059735 

Knapp and 

Terrell 1980 

MRID 

00127869 

Ward 1981
 

MRID
 

00074291
 

Appendix 3-1 
 

 

 

 

Appendix 3: Toxicity of disparlure to aquatic species (unless otherwise specified, all 

concentrations are expressed in terms of a.i.) 

Animal Dose/Exposure Response Reference 

Daphnia magna, Disparlure TGAI No mortalities or sublethal effects LeBlanc et 

<24 hours old, 15 48-hour exposure to 0.010 - 0.22 occurred at concentrations of 0.010 al. 1980 

daphnids/concentra mg/L [0.22, 0.13, 0.079, 0.048, and 0.017 mg/L. MRID 

tion. 0.028, 0.017, and 0.01 mg/L Mortality rates at higher doses: 00127868 

nominal]. The concentration of 0.22 mg/L 15/15 

disparlure in the test media was not 0.13 mg/L 12/15 

measured. Static conditions in 500 0.079 mg/L 2/15 

mL test solution. Mortalities were 0.048 mg/L 1/15 

recorded after 24 and 48 hours. 0.028 mg/L 1/15 

Additional notes on LeBlanc et al. 1980: Some organisms (number not specified) were trapped in the air-water 

interface at concentrations of 0.028 mg/L and higher. EC50 = 0.098 (0.019-0.12) mg/L. 

NOEC = 0.017 mg/L 

Appendix 3-2 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

Standard Disrupt II Flakes (SF) – i.e., flakes previously used by FS 

Daphnia magna, Disrupt II, SF (blank standard No mortality or immobility. Palmer and 

<24 hours old, 20 flakes, no disparlure) Krueger 

daphnids 300 mg/L for 48 hours. 2006a 

200 ml test solution volume 

Daphnia magna, 

<24 hours old, 20 

daphnids per 

concentration in 2 

replicates with 10 

organisms/replicat 

e. 

Daphnia magna, 

<24 hours old, 20 

daphnids per 

concentration in 2 

replicates with 10 

organisms/replicat 

e. 

Disrupt II, SF 2003 (standard 

flakes from 2003, 17.9% 

disparlure) 0, 1, 3, 10, 30, 100, and 

300 mg preparation/L. 

Preparations based on flakes mixed 

in water for 24 hours prior to the 

preparation of filtered test solutions 

(i.e., no flakes in the test solutions). 

Disparlure concentrations not 

monitored. 

The nominal formulation 

concentrations correspond to 

nominal concentrations of 

disparlure of: 0, 0.18, 0.54, 1.8, 5.4, 

18, and 54 mg a.i./L. 

Disrupt II, SF 2005 (standard 

flakes from 2005, 17.9% 

disparlure) 0, 1, 3, 10, 30, 100, and 

300 mg preparation/L. 

Preparations based on flakes mixed 

in water for 24 hours prior to the 

preparation of filtered test solutions 

(i.e., no flakes in the test solutions). 

Disparlure concentrations not 

monitored. 

The nominal formulation 

concentrations correspond to 

nominal concentrations of 

disparlure of: 0, 0.18, 0.54, 1.8, 5.4, 

18, and 54 mg a.i./L. 

No effects at any concentrations 

after 4 or 24 hours. 

At 48 hours, no effects at 1, 3, 30, 

and 100 mg formulation/L. 

At 10 mg/L, 1/20 organisms 

appeared lethargic. 

At 300 mg/L, 3/10 organisms in one 

replicate were trapped at the water 

surface but appeared normal after 

gentle submersion. 1/10 organisms 

did not appear normal (NOS) after 

being trapped on the water surface. 

EC50: > 300 mg/L (53.7 mg a.i./L 

based on nominal concentrations) 

No effects at any concentrations 

after 4 hours. 

At 24 hours, 20 of 20 daphnids 

were either dead (n=3) or immobile 

(n=17) in the 300 mg/L group. No 

effects at lower concentrations. 

At 48-hours, no effects in the 1, 3, 

or 10 mg/L groups. At 30 mg/L, 

9/20 organisms appeared to be 

lethargic. At 100 mg/L, 16/20 

organisms were immobile. At 300 

mg/L, 14/20 organisms were dead 

and the remaining 4 were immobile. 

Palmer and 

Krueger 

2006a 

Palmer and 

Krueger 

2006a 

Additional Notes on Palmer and Krueger 2006a, (standard flakes from 2005): At 48 hours, no effects at 1, 3, 

30, and 100 mg formulation/L. At 10 mg/L, 1/20 organisms appeared lethargic. At 300 mg/L, 3/10 organisms in 

one replicate were trapped at the water surface but appeared normal after gentle submersion. 1/10 organisms did 

not appear normal (NOS) after being trapped on the water surface. 

24 hr LC50: 173 (100-300 mg/L) 

48 hr LC50: 69 (30-100 mg/L) 

Appendix 3-3 
 

 

 

 

 

 

 

Modified Disrupt II Flakes – i.e., flakes currently used by FS 

Daphnia magna, 

<24 hours old, 20 

daphnids 

Daphnia magna, 

<24 hours old, 20 

daphnids per 

concentration in 2 

replicates with 10 

organisms/replicat 

e. 

Daphnia magna, 

<24 hours old, 20 

daphnids per 

concentration in 2 

replicates with 10 

organisms/replicat 

e. 

Disrupt II, MF (blank modified 

flakes, no disparlure) 

300 mg/L for 48 hours. 

200 ml test solution volume. 

Disrupt II, MF 2003 (modified
 

flakes from 2003, 17.9% 

disparlure) 

0, 1, 3, 10, 30, 100, and 300 mg
 

preparation/L.
 

Preparations based on flakes mixed 

in water for 24 hours prior to the 

preparation of filtered test solutions 

(i.e., no flakes in the test solutions). 

Disparlure concentrations not 

monitored. 

The nominal formulation 

concentrations correspond to 

nominal disparlure concentrations
 

of disparlure of: 0, 0.18, 0.54, 1.8,
 

5.4, 18, and 54 mg a.i./L. 

Disrupt II, MF 2005 (modified
 

flakes from 2005, 17.9% 

disparlure) 

0, 1, 3, 10, 30, 100, and 300 mg
 

preparation/L.
 

Preparations based on flakes mixed 

in water for 24 hours prior to the 

preparation of filtered test solutions 

(i.e., no flakes in the test solutions). 

Disparlure concentrations not 

monitored. 

The nominal formulation 

concentrations correspond to 

nominal concentrations of
 

disparlure of: 0, 0.18, 0.54, 1.8, 5.4,
 

18, and 54 mg a.i./L. 

No mortality or immobility. 

At 4 hours, 1/20 daphnids in the 1 

mg/L group trapped on the water 

surface but normal after gentle 

submersion. 

At 24 hours, no effects at any 

concentrations. 

At 48 hours, no effects at 3, 10, 30, 

and 100 mg formulation/L. At 1 

mg/L and 300 mg/L, 2/20 daphnids 

in each group were trapped at the 

water surface but normal after 

gentle submersion. 

EC50: > 300 mg/L 

At 4 hours, 17/20 daphnids in the 

300 mg/L group trapped on the 

water surface but normal after 

gentle submersion. No effects at 

lower concentrations. 

At 24 hours: 

No effects in the 1, 3, 10, and 30 

mg/L groups. 

At 100 mg/L, 14/20 dead and 6/20
 

trapped on the water surface. 

At 300 mg/L, 14/20 trapped on the 

water surface and lethargic after 

gentle submersion. 

Palmer and 

Krueger 

2006b 

Palmer and 

Krueger 

2006b 

Palmer and 

Krueger 

2006a 

Appendix 3-4 
 

 

Modified Disrupt II Flakes – i.e., flakes currently used by FS 

Daphnia magna, Disrupt II, MF (blank modified No mortality or immobility. Palmer and 

<24 hours old, 20 flakes, no disparlure) Krueger 

daphnids 300 mg/L for 48 hours. 2006b 

200 ml test solution volume. 

Additional Notes, Palmer and Krueger 2006a. Modified flakes, 2005: At 48-hours, no effects in 

the 1, 3, or 10 mg/L groups. At 30 mg/L, 1/20 organisms appeared to be lethargic and 1/20 

trapped on the water surface. At 100 mg/L, 20/20 organisms were dead. At 300 mg/L, 13/20 

organisms were dead, 1/20 was lethargic, 2 were trapped on the water surface. 

24 hr LC50: > 30 mg/L 

48 hr LC50: 48 (30-100 mg/L) 

Appendix 3-5 
 

 

 

 

 

 

 

 

Appendix 4: EPI Suite Output for Disparlure 

Run conducted on June 28, 2006 by Patrick Durkin using EPI-Suite Version 3.12. 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisCAS NUM: 029804-22-6 

MOL FOR: C19 H38 O1 

MOL WT : 282.51 

------------------------------ EPI SUMMARY (v3.12) --------------------------

Physical Property Inputs:

Water Solubility (mg/L): ------

Vapor Pressure (mm Hg) : ------

Henry LC (atm-m3/mole) : ------

Log Kow (octanol-water): ------

Boiling Point (deg C) : ------

Melting Point (deg C) : ------

KOWWIN Program (v1.67) Results:

=============================== 

Log Kow(version 1.67 estimate): 8.08 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMOL FOR: C19 H38 O1 

MOL WT : 282.51 

-------+-----+--------------------------------------------+---------+--------

TYPE | NUM | LOGKOW FRAGMENT DESCRIPTION | COEFF | VALUE 

-------+-----+--------------------------------------------+---------+--------

Frag | 3 | -CH3 [aliphatic carbon] | 0.5473 | 1.6419 

Frag | 13 | -CH2- [aliphatic carbon] | 0.4911 | 6.3843 

Frag | 3 | -CH [aliphatic carbon] | 0.3614 | 1.0842 

Frag | 1 | -O- [oxygen, aliphatic attach] |-1.2566 | -1.2566

Const | | Equation Constant | | 0.2290 

-------+-----+--------------------------------------------+---------+--------

Log Kow = 8.0828 

MPBPWIN (v1.41) Program Results:

=============================== 

Experimental Database Structure Match: no data 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMOL FOR: C19 H38 O1 

MOL WT : 282.51 

------------------------ SUMMARY MPBPWIN v1.41 --------------------

Boiling Point: 328.27 deg C (Adapted Stein and Brown Method) 

Melting Point: 56.00 deg C (Adapted Joback Method) 

Appendix 4-1 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

------------------------------------------------------- 

------------------------------------------------------- 

Melting Point: 78.02 deg C (Gold and Ogle Method)

Mean Melt Pt : 67.01 deg C (Joback; Gold,Ogle Methods)

Selected MP: 67.01 deg C (Mean Value) 

Vapor Pressure Estimations (25 deg C):

(Using BP: 328.27 deg C (estimated))

(Using MP: 67.01 deg C (estimated))

VP: 0.00021 mm Hg (Antoine Method)

VP: 0.000342 mm Hg (Modified Grain Method)

VP: 0.000321 mm Hg (Mackay Method)

Selected VP: 0.000342 mm Hg (Modified Grain Method) 

-------+-----+--------------------+----------+---------

TYPE | NUM | BOIL DESCRIPTION | COEFF | VALUE 

-------+-----+--------------------+----------+---------

Group | 3 | -CH3 | 21.98 | 65.94 

Group | 13 | -CH2- | 24.22 | 314.86 

Group | 1 | >CH- | 11.86 | 11.86 

Group | 2 | >CH- (ring) | 21.66 | 43.32 

Group | 1 | -O- (ring) | 32.98 | 32.98 

* | | Equation Constant | | 198.18 

=============+====================+==========+========= 

RESULT-uncorr| BOILING POINT in deg Kelvin | 667.14 

RESULT- corr | BOILING POINT in deg Kelvin | 601.43 

| BOILING POINT in deg C | 328.27 

-------+-----+--------------------+----------+---------

TYPE | NUM | MELT DESCRIPTION | COEFF | VALUE 

-------+-----+--------------------+----------+---------

Group | 3 | -CH3 | -5.10 | -15.30 

Group | 13 | -CH2- | 11.27 | 146.51 

Group | 1 | >CH- | 12.64 | 12.64 

Group | 2 | >CH- (ring) | 19.88 | 39.76 

Group | 1 | -O- (ring) | 23.05 | 23.05 

* | | Equation Constant | | 122.50 

=============+====================+==========+========= 

RESULT | MELTING POINT in deg Kelvin | 329.16 

| MELTING POINT in deg C | 56.00 

Water Sol from Kow (WSKOW v1.41) Results:

======================================== 

Water Sol: 0.001939 mg/L 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMOL FOR: C19 H38 O1 

MOL WT : 282.51 

Appendix 4-2 
 

 

 

 

 

 

 

-- 

-- 

-- 

-- 

---------------------------------- WSKOW v1.41 Results ----------------------

Log Kow (estimated) : 8.08 

Log Kow (experimental): not available from database 

Log Kow used by Water solubility estimates: 8.08 

Equation Used to Make Water Sol estimate:

Log S (mol/L) = 0.796 - 0.854 log Kow - 0.00728 MW + Correction

(used when Melting Point NOT available) 

Correction(s): Value 

-------------------- -----

No Applicable Correction Factors

Log Water Solubility (in moles/L) : -8.163 

Water Solubility at 25 deg C (mg/L): 0.001939 

WATERNT Program (v1.01) Results:

=============================== 

Water Sol (v1.01 est): 0.0027812 mg/L 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMOL FOR: C19 H38 O1 

MOL WT : 282.51 

-------+-----+--------------------------------------------+----------+-------

TYPE | NUM | WATER SOLUBILITY FRAGMENT DESCRIPTION | COEFF | VALUE 

-------+-----+--------------------------------------------+----------+-------

Frag | 3 | -CH3 [aliphatic carbon] |-0.3213 | -

0.9638 

Frag | 13 | -CH2- [aliphatic carbon] |-0.5370 | -

6.9812 

Frag | 3 | -CH [aliphatic carbon] |-0.5285 | -

1.5856 

Frag | 1 | -O- [oxygen, aliphatic attach] | 1.2746 |

1.2746 

Const | | Equation Constant | |

0.2492 

-------+-----+--------------------------------------------+----------+-------

Log Water Sol (moles/L) at 25 dec C = -

8.0068 

Water Solubility (mg/L) at 25 dec C =0.0027812 

Appendix 4-3 
 

 

 

 

 

 

 

------------------------------ 

ECOSAR Program (v0.99h) Results:

=============================== 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisCAS Num: 

ChemID1: 

ChemID2: 

ChemID3: 

MOL FOR: C19 H38 O1 

MOL WT : 282.51 

Log Kow: 8.08 (KowWin estimate)

Melt Pt: 

Wat Sol: 0.0007897 mg/L (calculated)

ECOSAR v0.99h Class(es) Found 

Epoxides 

Predicted 

ECOSAR Class Organism Duration End Pt mg/L

(ppm)

=========================== ================== ======== ====== 

========== 

Neutral Organic SAR : Fish 14-day LC50 0.00192 

* 

(Baseline Toxicity) 

Epoxides : Fish 96-hr LC50 0.119 

* 

Epoxides : Fish 14-day LC50 0.144 

* 

Epoxides : Daphnid 48-hr LC50 0.008 

* 

Note: * = asterisk designates: Chemical may not be soluble

enough to measure this predicted effect.

Fish and daphnid acute toxicity log Kow cutoff: 5.0

Green algal EC50 toxicity log Kow cutoff: 6.4

Chronic toxicity log Kow cutoff: 8.0

MW cutoff: 1000 

HENRY (v3.10) Program Results:

============================= 

Bond Est : 1.49E-002 atm-m3/mole

Group Est: 6.14E-002 atm-m3/mole 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMOL FOR: C19 H38 O1 

MOL WT : 282.51 

--------------------------- HENRYWIN v3.10 Results --------------------------

----------+---------------------------------------------+---------+----------

CLASS | BOND CONTRIBUTION DESCRIPTION | COMMENT | VALUE 

----------+---------------------------------------------+---------+----------

HYDROGEN | 38 Hydrogen to Carbon (aliphatic) Bonds | | -4.5477 

Appendix 4-4 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 FRAGMENT | 18 C-C | | 2.0935 

FRAGMENT | 2 C-O | | 2.1709 

FACTOR | * Epoxide | | .5000 

----------+---------------------------------------------+---------+----------

RESULT | BOND ESTIMATION METHOD for LWAPC VALUE | TOTAL | 0.217 

----------+---------------------------------------------+---------+----------

HENRYs LAW CONSTANT at 25 deg C = 1.49E-002 atm-m3/mole

= 6.07E-001 unitless 

--------+-----------------------------------------------+------------+-------

-

| GROUP CONTRIBUTION DESCRIPTION | COMMENT | VALUE 

--------+-----------------------------------------------+------------+-------

-

| 3 CH3 (X) | | -1.86

| 13 CH2 (C)(C) | | -1.95

| 1 CH (C)(C)(C) | | 0.24 

| 2 CH (C)(C)(O) | | 0.24 

| 1 O (C)(C) | | 2.93 

--------+-----------------------------------------------+------------+-------

-

RESULT | GROUP ESTIMATION METHOD for LOG GAMMA VALUE | TOTAL | -0.40

--------+-----------------------------------------------+------------+-------

-

HENRYs LAW CONSTANT at 25 deg C = 6.14E-002 atm-m3/mole

= 2.51E+000 unitless 

Henrys LC [VP/WSol estimate using EPI values]:

HLC: 6.556E-002 atm-m3/mole

VP: 0.000342 mm Hg

WS: 0.00194 mg/L 

BIOWIN (v4.02) Program Results:

============================== 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMOL FOR: C19 H38 O1 

MOL WT : 282.51 

--------------------------- BIOWIN v4.02 Results ----------------------------

Biowin1 (Linear Model Prediction) : Does Not Biodegrade Fast

Biowin2 (Non-Linear Model Prediction): Does Not Biodegrade Fast

Biowin3 (Ultimate Biodegradation Timeframe): Weeks 

Biowin4 (Primary Biodegradation Timeframe): Days-Weeks

Biowin5 (MITI Linear Model Prediction) : Does Not Biodegrade Fast

Biowin6 (MITI Non-Linear Model Prediction): Does Not Biodegrade Fast

Ready Biodegradability Prediction: NO 

------+-----+--------------------------------------------+---------+---------

TYPE | NUM | Biowin1 FRAGMENT DESCRIPTION | COEFF | VALUE 

------+-----+--------------------------------------------+---------+---------

Frag | 1 | Linear C4 terminal chain [CCC-CH3] | 0.1084 | 0.1084 

Appendix 4-5 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 Frag | 1 | Aliphatic ether [C-O-C] | -0.3474 | -0.3474

MolWt| * | Molecular Weight Parameter | | -0.1345

Const| * | Equation Constant | | 0.7475 

============+============================================+=========+========= 

RESULT | Biowin1 (Linear Biodeg Probability) | | 0.3741 

============+============================================+=========+========= 

------+-----+--------------------------------------------+---------+---------

TYPE | NUM | Biowin2 FRAGMENT DESCRIPTION | COEFF | VALUE 

------+-----+--------------------------------------------+---------+---------

Frag | 1 | Linear C4 terminal chain [CCC-CH3] | 1.8437 | 1.8437 

Frag | 1 | Aliphatic ether [C-O-C] | -3.4294 | -3.4294

MolWt| * | Molecular Weight Parameter | | -4.0117

============+============================================+=========+========= 

RESULT | Biowin2 (Non-Linear Biodeg Probability) | | 0.0699 

============+============================================+=========+========= 

A Probability Greater Than or Equal to 0.5 indicates --> Biodegrades Fast

A Probability Less Than 0.5 indicates --> Does NOT Biodegrade Fast 

------+-----+--------------------------------------------+---------+---------

TYPE | NUM | Biowin3 FRAGMENT DESCRIPTION | COEFF | VALUE 

------+-----+--------------------------------------------+---------+---------

Frag | 1 | Linear C4 terminal chain [CCC-CH3] | 0.2983 | 0.2983 

Frag | 1 | Aliphatic ether [C-O-C] | -0.0087 | -0.0087

MolWt| * | Molecular Weight Parameter | | -0.6243

Const| * | Equation Constant | | 3.1992 

============+============================================+=========+========= 

RESULT | Biowin3 (Survey Model - Ultimate Biodeg) | | 2.8645 

============+============================================+=========+========= 

------+-----+--------------------------------------------+---------+---------

TYPE | NUM | Biowin4 FRAGMENT DESCRIPTION | COEFF | VALUE 

------+-----+--------------------------------------------+---------+---------

Frag | 1 | Linear C4 terminal chain [CCC-CH3] | 0.2691 | 0.2691 

Frag | 1 | Aliphatic ether [C-O-C] | -0.0097 | -0.0097

MolWt| * | Molecular Weight Parameter | | -0.4076

Const| * | Equation Constant | | 3.8477 

============+============================================+=========+========= 

RESULT | Biowin4 (Survey Model - Primary Biodeg) | | 3.6995 

============+============================================+=========+========= 

Result Classification: 5.00 -> hours 4.00 -> days 3.00 -> weeks 

(Primary & Ultimate) 2.00 -> months 1.00 -> longer 

------+-----+--------------------------------------------+---------+---------

TYPE | NUM | Biowin5 FRAGMENT DESCRIPTION | COEFF | VALUE 

------+-----+--------------------------------------------+---------+---------

Frag | 1 | Aliphatic ether [C-O-C] | 0.0015 | 0.0015 

Frag | 3 | Methyl [-CH3] | 0.0004 | 0.0012 

Frag | 13 | -CH2- [linear] | 0.0494 | 0.6424 

Frag | 1 | -CH- [linear] | -0.0507 | -0.0507

Frag | 2 | -CH - [cyclic] | 0.0124 | 0.0249 

MolWt| * | Molecular Weight Parameter | | -0.8405

Const| * | Equation Constant | | 0.7121 

Appendix 4-6 
 

 

 

 

 

 

 

 

 

 

 

 

============+============================================+=========+========= 

RESULT | Biowin5 (MITI Linear Biodeg Probability) | | 0.4910 

============+============================================+=========+========= 

------+-----+--------------------------------------------+---------+---------

TYPE | NUM | Biowin6 FRAGMENT DESCRIPTION | COEFF | VALUE 

------+-----+--------------------------------------------+---------+---------

Frag | 1 | Aliphatic ether [C-O-C] | -0.1071 | -0.1071

Frag | 3 | Methyl [-CH3] | 0.0194 | 0.0583 

Frag | 13 | -CH2- [linear] | 0.4295 | 5.5834 

Frag | 1 | -CH- [linear] | -0.0998 | -0.0998

Frag | 2 | -CH - [cyclic] | -0.1295 | -0.2589

MolWt| * | Molecular Weight Parameter | | -8.1558

============+============================================+=========+========= 

RESULT |Biowin6 (MITI Non-Linear Biodeg Probability)| | 0.3883 

============+============================================+=========+========= 

A Probability Greater Than or Equal to 0.5 indicates --> Biodegrades Fast

A Probability Less Than 0.5 indicates --> Does NOT Biodegrade Fast 

AOP Program (v1.91) Results:

=========================== 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMOL FOR: C19 H38 O1 

MOL WT : 282.51 

------------------- SUMMARY (AOP v1.91): HYDROXYL RADICALS ------------------

-

Hydrogen Abstraction = 21.7096 E-12 cm3/molecule-sec

Reaction with N, S and -OH = 0.0000 E-12 cm3/molecule-sec

Addition to Triple Bonds = 0.0000 E-12 cm3/molecule-sec

Addition to Olefinic Bonds = 0.0000 E-12 cm3/molecule-sec

Addition to Aromatic Rings = 0.0000 E-12 cm3/molecule-sec

Addition to Fused Rings = 0.0000 E-12 cm3/molecule-sec 

OVERALL OH Rate Constant = 21.7096 E-12 cm3/molecule-sec

HALF-LIFE = 0.493 Days (12-hr day; 1.5E6 OH/cm3)

HALF-LIFE = 5.912 Hrs 

------------------- SUMMARY (AOP v1.91): OZONE REACTION ---------------------

-

****** NO OZONE REACTION ESTIMATION ****** 

(ONLY Olefins and Acetylenes are Estimated) 

Experimental Database: NO Structure Matches 

PCKOC Program (v1.66) Results:

============================= 

Appendix 4-7 
 

 

 

 

 

 Koc (estimated): 3.44e+004 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMOL FOR: C19 H38 O1 

MOL WT : 282.51 

--------------------------- PCKOCWIN v1.66 Results --------------------------

-

First Order Molecular Connectivity Index ........... : 9.736 

Non-Corrected Log Koc .............................. : 5.8004 

Fragment Correction(s):

1 Ether, aliphatic (-C-O-C-) .......... : -1.2643 

Corrected Log Koc .................................. : 4.5361 

Estimated Koc: 3.437e+004 

HYDROWIN Program (v1.67) Results:

================================ 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMOL FOR: C19 H38 O1 

MOL WT : 282.51 

--------------------------- HYDROWIN v1.67 Results --------------------------

-

NOTE: Fragment(s) on this compound are NOT available from the fragment

library. Substitute(s) have been used!!! Substitute R1, R2, R3,

or R4 fragments are marked with double astericks "**". 

O 

R1 / \ R3 

EPOXIDE: >C - C< 

R2 R4 

** R1: n-Octyl- ** R3: n-ButylR2: -H R4: -H 

Ka hydrolysis at (epoxy O) atom # 1: 4.271E-001 L/mol-sec 

Total Ka (acid-catalyzed) at 25 deg C : 4.271E-001 L/mol-sec

Ka Half-Life at pH 7: 187.803 days 

The rate constant estimated for the EPOXIDE DOES NOT 

include the neutral hydrolysis rate constant!!

For some epoxides, the neutral rate constant is the

dominant hydrolysis rate at environmental pHs!

If the neutral rate constant is important, the HYDRO

estimated rate will under-estimate the actual rate! 

Appendix 4-8 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

-- 

------------- 

BCF Program (v2.15) Results:

=========================== 

SMILES : O(C1CCCCCCCCCC)C1CCCCC(C)C

CHEM : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMOL FOR: C19 H38 O1 

MOL WT : 282.51 

--------------------------------- Bcfwin v2.15 ------------------------------

Log Kow (estimated) : 8.08 

Log Kow (experimental): not available from database 

Log Kow used by BCF estimates: 8.08 

Equation Used to Make BCF estimate:

Log BCF = -1.37 log Kow + 14.4 + Correction 

Correction(s): Value 

Alkyl chains (8+ -CH2- groups) -1.500 

Estimated Log BCF = 1.827 (BCF = 67.08) 

Volatilization From Water 

========================= 

Chemical Name: Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMolecular Weight : 282.51 g/mole

Water Solubility : -----

Vapor Pressure : -----

Henry's Law Constant: 0.0149 atm-m3/mole (estimated by Bond SAR Method) 

RIVER LAKE 

--------- ---------

Water Depth (meters):

Wind Velocity (m/sec):

Current Velocity (m/sec): 

1 

5 

1 

1 

0.5 

0.05 

HALF-LIFE (hours) :

HALF-LIFE (days ) : 

1.781 

0.07422 

160.4 

6.682 

STP Fugacity Model: Predicted Fate in a Wastewater Treatment Facility

====================================================================== 

(using 10000 hr Bio P,A,S)

PROPERTIES OF: Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMolecular weight (g/mol) 282.51 

Aqueous solubility (mg/l) 0 

Vapour pressure (Pa) 0 

(atm) 0 

(mm Hg) 0 

Henry 's law constant (Atm-m3/mol) 0.0149 

Appendix 4-9 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

------------- 

Air-water partition coefficient 0.609366 

Octanol-water partition coefficient (Kow) 1.20226E+008 

Log Kow 8.08 

Biomass to water partition coefficient 2.40453E+007 

Temperature [deg C] 25 

Biodeg rate constants (h^-1),half life in biomass (h) and in 2000 mg/L MLSS

(h): 

-Primary tank

-Aeration tank 

0.00 

0.00 

9999.79 

9999.79 

10000.00 

10000.00 

-Settling tank 0.00 9999.79 10000.00 

STP Overall Chemical Mass Balance: 

---------------------------------

g/h mol/h percent 

Influent 1.00E+001 3.5E-002 100.00 

Primary sludge

Waste sludge

Primary volatilization

Settling volatilization

Aeration off gas 

5.99E+000 

3.33E+000 

2.72E-005 

6.01E-005 

9.17E-003 

2.1E-002 

1.2E-002 

9.6E-008 

2.1E-007 

3.2E-005 

59.88 

33.28 

0.00 

0.00 

0.09 

Primary biodegradation

Settling biodegradation

Aeration biodegradation 

1.75E-002 

4.25E-003 

5.60E-002 

6.2E-005 

1.5E-005 

2.0E-004 

0.18 

0.04 

0.56 

Final water effluent 5.97E-001 2.1E-003 5.97 

Total removal 9.40E+000 3.3E-002 94.03 

Total biodegradation 7.77E-002 2.8E-004 0.78 

STP Fugacity Model: Predicted Fate in a Wastewater Treatment Facility

====================================================================== 

(using Biowin/EPA draft method)

PROPERTIES OF: Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMolecular weight (g/mol) 282.51 

Aqueous solubility (mg/l) 0 

Vapour pressure (Pa) 0 

(atm) 0 

(mm Hg) 0 

Henry 's law constant (Atm-m3/mol) 0.0149 

Air-water partition coefficient 0.609366 

Octanol-water partition coefficient (Kow) 1.20226E+008 

Log Kow 8.08 

Biomass to water partition coefficient 2.40453E+007 

Temperature [deg C] 25 

Biodeg rate constants (h^-1),half life in biomass (h) and in 2000 mg/L MLSS

(h): 

-Primary tank 0.02 30.00 30.00 

-Aeration tank 0.23 3.00 3.00 

-Settling tank 0.23 3.00 3.00 

Appendix 4-10 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 STP Overall Chemical Mass Balance: 

---------------------------------

g/h mol/h percent 

Influent 1.00E+001 3.5E-002 100.00 

Primary sludge 3.78E+000 1.3E-002 37.84 

Waste sludge 3.83E-002 1.4E-004 0.38 

Primary volatilization 1.72E-005 6.1E-008 0.00 

Settling volatilization 6.92E-007 2.4E-009 0.00 

Aeration off gas 1.14E-004 4.0E-007 0.00 

Primary biodegradation 3.69E+000 1.3E-002 36.91 

Settling biodegradation 1.63E-001 5.8E-004 1.63 

Aeration biodegradation 2.32E+000 8.2E-003 23.16 

Final water effluent 6.87E-003 2.4E-005 0.07 

Total removal 9.99E+000 3.5E-002 99.93 

Total biodegradation 6.17E+000 2.2E-002 61.70 

(** Total removal recommended maximum is 99 percent) 

Level III Fugacity Model (Full-Output):

======================================= 

Chem Name : Oxirane, 2-decyl-3-(5-methylhexyl)-, cisMolecular Wt: 282.51 

Henry's LC : 0.0149 atm-m3/mole (Henrywin program)

Vapor Press : 0.000342 mm Hg (Mpbpwin program)

Liquid VP : 0.00089 mm Hg (super-cooled)

Melting Pt : 67 deg C (Mpbpwin program)

Log Kow : 8.08 (Kowwin program)

Soil Koc : 4.93e+007 (calc by model) 

Mass Amount Half-Life Emissions 

(percent) (hr) (kg/hr)

Air 0.395 11.8 1000 

Water 3.77 360 1000 

Soil 28.1 720 1000 

Sediment 67.8 3.24e+003 0 

Fugacity Reaction Advection Reaction Advection 

(atm) (kg/hr) (kg/hr) (percent) (percent)

Air 1.26e-011 857 146 28.6 4.88 

Water 4.55e-010 269 140 8.96 4.66 

Soil 2.57e-012 1e+003 0 33.4 0 

Sediment 2.8e-010 537 50.2 17.9 1.67 

Persistence Time: 1.24e+003 hr 

Reaction Time: 1.39e+003 hr 

Advection Time: 1.1e+004 hr 

Percent Reacted: 88.8 

Percent Advected: 11.2 

Appendix 4-11 
 

 

 

 

 

 

-----------------------------------------------------------------------------

 Half-Lives (hr), (based upon Biowin (Ultimate) and Aopwin):

Air: 11.82 

Water: 360 

Soil: 720 

Sediment: 3240 

Biowin estimate: 2.865 (weeks ) 

Advection Times (hr):

Air: 100 

Water: 1000 

Sediment: 5e+004 

-

Appendix 4-12 
Appendix I 

Diflubenzuron

Risk Assessment
 

Figure I-1. The first power spraying apparatus was used in gypsy moth control 

operations before 1900. 

SERA TR 04-43-05-03b 

Control/Eradication Agents for the
 

Gypsy Moth ­

Human Health and Ecological Risk Assessment
 

for Diflubenzuron (Dimilin)
 

FINAL REPORT
 

Prepared for: 

USDA, Forest Service
 

Forest Health Protection 

GSA Contract No. GS-10F-0082F
 

USDA Forest Service BPA: WO-01-3187-0150
 

Requisition No.: 43-3187-1-0269
 

Task No. 5
 

Submitted to:
 

Dave Thomas, COTR
 

Forest Health Protection Staff
 

USDA Forest Service
 

Rosslyn Plaza Building C, Room 7129C
 

1601 North Kent Street
 

Arlington, VA 22209
 

Submitted by:
 

Patrick R. Durkin
 

Syracuse Environmental Research Associates, Inc.
 

5100 Highbridge St., 42C
 

Fayetteville, New York 13066-0950
 

Telephone: (315) 637-9560
 

Fax: (315) 637-0445
 

E-Mail: SERA_INC@msn.com
 

Home Page: www.sera-inc.com
 

July 30, 2004 
 

 

TABLE OF CONTENTS
 

TABLE OF CONTENTS ....................................................... ii
 

LIST OF APPENDICES ........................................................ v
 

LIST OF TABLES ............................................................ vi
 

LIST OF FIGURES .......................................................... vii
 

LIST OF WORKSHEETS ..................................................... vii
 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . viii
 

COMMON UNIT CONVERSIONS AND ABBREVIATIONS . . . . . . . . . . . . . . . . . . . . . . . . . x
 

CONVERSION OF SCIENTIFIC NOTATION ..................................... xi
 

EXECUTIVE SUMMARY .................................................... xii
 

1. INTRODUCTION ......................................................... 1-1
 

2. PROGRAM DESCRIPTION ................................................ 2-1
 

2.1. OVERVIEW ..................................................... 2-1
 

2.2. CHEMICAL DESCRIPTION AND COMMERCIAL FORMULATIONS . . . . . 2-1
 

2.3. APPLICATION METHODS AND RATES .......................... 2-2
 

2.4. USE STATISTICS ...... ........... .......................... 2-4
 

3. HUMAN HEALTH RISK ASSESSMENT ..................................... 3-1
 

3.1. HAZARD IDENTIFICATION ....................................... 3-1
 

3.1.1. Overview. ................................................ 3-1
 

3.1.2. Mechanisms of Action .................................... 3-1
 

3.1.3. Kinetics and Metabolism. .................................... 3-2
 

3.1.3.1. Oral Absorption .................................... 3-2
 

3.1.3.2. Dermal Absorption .................................. 3-3
 

3.1.3.3. Metabolism ....................................... 3-4
 

3.1.4. Acute Oral Toxicity. ........................................ 3-4
 

3.1.5. Subchronic and Chronic Toxicity. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-4
 

3.1.6. Effects on Nervous System. .................................. 3-5
 

3.1.7. Effects on Immune System. .................................. 3-6
 

3.1.8. Effects on Endocrine Function. ............................ 3-6
 

3.1.9. Reproductive and Teratogenic Effects. . . . . . . . . . . . . . . . . . . . . . . . . . . 3-7
 

ii 
TABLE OF CONTENTS (continued) 

3.1.10. Carcinogenicity and Mutagenicity. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-8
 

3.1.11. Irritation and Sensitization (Effects on the Skin and Eyes). . . . . . . . . . 3-8
 

3.1.12. Systemic Toxic Effects from Dermal Exposure. . . . . . . . . . . . . . . . . . 3-8
 

3.1.13. Inhalation Exposure. ....................................... 3-9
 

3.1.14. Inerts and Adjuvants. ...................................... 3-9
 

3.1.15. Impurities and Metabolites.................................. 3-10
 

3.1.16. Toxicologic Interactions. .................................. 3-10
 

3.2. EXPOSURE ASSESSMENT ....................................... 3-11
 

3.2.1. Overview. ............................................... 3-11
 

3.2.2. Workers. ................................................ 3-12
 

3.2.2.1. General Exposures ................................. 3-12
 

3.2.2.2. Accidental Exposures ............................... 3-14
 

3.2.3. General Public. ........................................... 3-15
 

3.2.3.1. General Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-15
 

3.2.3.2. Direct Spray ...................................... 3-15
 

3.2.3.3. Dermal Exposure from Contaminated Vegetation . . . . . . . . . 3-16
 

3.2.3.4. Contaminated Water ............................... 3-16
 

3.2.3.5. Oral Exposure from Contaminated Fish . . . . . . . . . . . . . . . . 3-23
 

3.2.3.6. Oral Exposure from Contaminated Vegetation . . . . . . . . . . . 3-24
 

3.3. DOSE-RESPONSE ASSESSMENT .................................. 3-26
 

3.3.1. Overview ................................................ 3-26
 

3.3.2. Diflubenzuron ............................................ 3-26
 

3.3.3. 4-Chloroaniline ........................................... 3-28
 

3.4. RISK CHARACTERIZATION ...................................... 3-31
 

3.4.1. Overview ................................................ 3-31
 

3.4.2. Workers ................................................. 3-31
 

3.4.3. General Public. ........................................... 3-32
 

3.4.4. Sensitive Subgroups........................................ 3-33
 

3.4.5. Connected Actions......................................... 3-33
 

3.4.6. Cumulative Effects......................................... 3-34
 

4. ECOLOGICAL RISK ASSESSMENT ......................................... 4-1
 

4.1. HAZARD IDENTIFICATION ....................................... 4-1
 

4.1.1. Overview. ................................................ 4-1
 

4.1.2. Toxicity to Terrestrial Organisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-2
 

4.1.2.1. Mammals ......................................... 4-2
 

4.1.2.2. Birds ............................................. 4-3
 

4.1.2.3. Terrestrial Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-3
 

4.1.2.4. Terrestrial Plants (Macrophytes) . . . . . . . . . . . . . . . . . . . . . . . 4-6
 

4.1.2.5. Terrestrial Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-6
 

iii 
 

TABLE OF CONTENTS (continued) 

4.1.3. Aquatic Organisms.......................................... 4-7
 

4.1.3.1. Fish .............................................. 4-7
 

4.1.3.2. Amphibians ....................................... 4-8
 

4.1.3.3. Aquatic Invertebrates ................................ 4-8
 

4.1.3.4. Aquatic Plants ..................................... 4-9
 

4.1.3.5. Aquatic Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-10
 

4.2. EXPOSURE ASSESSMENT ....................................... 4-11
 

4.2.1. Overview. ............................................... 4-11
 

4.2.2. Terrestrial Animals. ....................................... 4-12
 

4.2.2.1. Direct Spray ...................................... 4-13
 

4.2.2.2. Indirect Contact ................................... 4-14
 

4.2.2.3. Ingestion of Contaminated Vegetation or Prey . . . . . . . . . . . 4-14
 

4.2.2.4. Ingestion of Contaminated Water . . . . . . . . . . . . . . . . . . . . . 4-16
 

4.2.3. Terrestrial Plants. ......................................... 4-17
 

4.2.4. Soil Organisms............................................ 4-17
 

4.2.5. Aquatic Organisms......................................... 4-17
 

4.3. DOSE-RESPONSE ASSESSMENT .................................. 4-18
 

4.3.1. Overview. ............................................... 4-18
 

4.3.2. Toxicity to Terrestrial Organisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-19
 

4.3.2.1. Mammals ........................................ 4-19
 

4.3.2.2. Birds ............................................ 4-20
 

4.3.2.3. Terrestrial Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-21
 

4.3.2.4. Terrestrial Plants (Macrophytes) . . . . . . . . . . . . . . . . . . . . . . 4-23
 

4.3.2.5. Soil Microorganisms ............................... 4-23
 

4.3.3. Aquatic Organisms......................................... 4-23
 

4.3.3.1. Fish ............................................. 4-23
 

4.3.3.2. Amphibians ...................................... 4-24
 

4.3.3.3. Invertebrates .............................. . . . . . 4-24
 

4.3.3.4. Aquatic Plants .................................... 4-26
 

4.3.3.5. Microorganisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-27
 

4.4. RISK CHARACTERIZATION ...................................... 4-28
 

4.4.1. Overview. ............................................... 4-28
 

4.4.2. Terrestrial Organisms....................................... 4-28
 

4.4.2.1. Terrestrial Vertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-28
 

4.4.2.2. Terrestrial Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-29
 

4.4.2.3. Terrestrial Plants and Microorganisms . . . . . . . . . . . . . . . . 4-30
 

4.4.3. Aquatic Organisms......................................... 4-30
 

4.4.3.1. Aquatic Vertebrates ................................ 4-30
 

4.4.3.2. Aquatic Invertebrates ............................... 4-31
 

4.4.3.3. Aquatic Plants and Microorganisms . . . . . . . . . . . . . . . . . . . 4-32
 

5. REFERENCES ........................................................... 5-1
 

iv 
TABLE OF CONTENTS (continued) 

LIST OF APPENDICES 

Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Appendix 2: Laboratory and simulation studies on environmental fate of diflubenzuron and 

its metabolites 

Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its 

formulations 

Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its 

formulations 

Appendix 4: Toxicity of diflubenzuron and diflubenzuron formulations to birds 

Appendix 5: Toxicity to non-target terrestrial insects 

Appendix 6: Toxicity to diflubenzuron to fish 

Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Appendix 8: Toxicity of diflubenzuron to aquatic plants 

v 
LIST OF TABLES
 

Table 2-1: Selected physical and chemical properties of diflubenzuron . . . . . . . . . . . . . . Tables-1 

Table 2-2: Commercial formulations of diflubenzuron . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-2 

Table 2-3: Use of diflubenzuron by USDA from 1995 to 2002 for Suppression, 

Eradication, and Slow the Spread . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-3 

Table 3-1: Chemical and site parameters used in GLEAMS modeling for 

diflubenzuron . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-4 

Table 3-2: Summary of modeled concentrations of diflubenzuron in streams . . . . . . . . . Tables-5 

Table 3-3: Summary of modeled concentrations of diflubenzuron in ponds . . . . . . . . . . . Tables-6 

Table 3-4: Chemical and site parameters used in GLEAMS modeling for 

4-chloroaniline . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-7 

Table 3-5: Summary of modeled concentrations of 4-chloroaniline in streams . . . . . . . . Tables-8 

Table 3-6: Summary of modeled concentrations of 4-chloroaniline in ponds . . . . . . . . . . Tables-9 

Table 3-7: Estimated Environmental Concentrations of diflubenzuron in ponds 

and streams . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-10 

Table 3-8: Concentrations of diflubenzuron in surface water used in this 

risk assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-11 

Table 3-9: Concentrations of 4-chloroaniline in surface water used in this 

risk assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-12 

Table 4-1: Summary of field studies on the effects of diflubenzuron on 

terrestrial invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-13 

Table 4-2: Summary of modeled concentrations of diflubenzuron in soil . . . . . . . . . . . . Tables-14 

Table 4-3: Summary of modeled concentrations of 4-chloroaniline in soil . . . . . . . . . . . Tables-15 

Table 4-4: Summary of diflubenzuron toxicity values used in ecological 

risk assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-16
 

vi 
LIST OF TABLES (continued) 

Table 4-5: Diflubenzuron toxicity values used in risk assessment for 

terrestrial arthropods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-17 

Table 4-6: Diflubenzuron toxicity values used in risk assessment for 

aquatic invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-18 

Table 4-7: Summary of 4-chloroaniline toxicity values used in ecological 

risk assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-19 

Table 4-8: Acute toxicity of diflubenzuron in aquatic invertebrates. . . . . . . . . . . . . . . . Tables-20 

Table 4-9: Chronic toxicity of diflubenzuron in aquatic invertebrates . . . . . . . . . . . . . . Tables-21 

Table 4-10: Summary of field studies on the effects of diflubenzuron on 

aquatic invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-22
 

LIST OF FIGURES 

Figure 3-1: Overview of in vivo and environmental metabolism of diflubenzuron . . . . . Figures-1 

Figure 3-2: Modeled concentrations of diflubenzuron and 4-chloroaniline in ponds 

at an annual rainfall rate of 150 inches . . . . . . . . . . . . . . . . . . . . . . . . . Figures-2 

NOTE: Tables followed by figures are placed after Section 5 (References). 

LIST OF WORKSHEETS 

Supplement 1:	 Diflubenzuron – Worksheets for Human Health and Ecological Risk 

Assessments, SERA EXWS 04-43-05-03b1, Version 3.01. 

Supplement 2:	 4-Chloroaniline as an Environmental Metabolite of Diflubenzuron -Worksheets 

for Human Health and Ecological Risk Assessments, SERA EXWS 

04-43-05-03b2, Version 3.01. 

Located at: http://www.fs.fed.us/foresthealth/pesticide/risk.shtml 

vii 
AEL 

a.i. 

ATSDR 

BCF 

bw 

CBI 

CI 

cm 

CNS 

CPU 

DAA 

DAT 

d.f. 

ECx

EC25 

EC50 

EXAMS 

ExToxNet 

F 

FH 

FIFRA 

FOIA 

FQPA 

g 

GLEAMS 

ha 

Hb 

HQ 

IAA 

IARC 

IRIS 

ka

ke

kg 

Ko/c 

Ko/w 

Kp

L 

lb 

LC50 

LD50 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

adverse-effect level 

active ingredient 

Agency for Toxic Substances and Disease Registry 

bioconcentration factor 

body weight 

confidential business information 

confidence interval 

centimeter 

central nervous system 

chlorophenylurea 

days after application 

days after treatment 

degrees of freedom 

concentration causing X% inhibition of a process 

concentration causing 25% inhibition of a process 

concentration causing 50% inhibition of a process 

Exposure Analysis Modeling System 

Extension Toxicology Network 

female 

Forest Health 

Federal Insecticide, Fungicide and Rodenticide Act 

Freedom of Information Act 

Food Quality Protection Act 

gram 

Groundwater Loading Effects of Agricultural Management Systems 

hectare 

hemoglobin 

hazard quotient 

indole-3-acetic acid 

International Agency for Research on Cancer 

Integrated Risk Information System 

absorption coefficient 

elimination coefficient 

kilogram 

organic carbon partition coefficient 

octanol-water partition coefficient 

skin permeability coefficient 

liter 

pound 

lethal concentration, 50% kill 

lethal dose, 50% kill 

viii 
ACRONYMS, ABBREVIATIONS, AND SYMBOLS (continued) 

LOAEL lowest-observed-adverse-effect level 

LOC level of concern 

m meter 

M male 

MetHb methemoglobinemia 

mg milligram 

mg/kg/day milligrams of agent per kilogram of body weight per day 

mL milliliter 

mM millimole 

MOS margin of safety 

MRID Master Record Identification Number 

MSDS material safety data sheet 

MW molecular weight 

NCAP Northwest Coalition for Alternatives to Pesticides 

NCI National Cancer Institute 

NOAEL no-observed-adverse-effect level 

NOEC no-observed-effect concentration 

NOEL no-observed-effect level 

NOS not otherwise specified 

NRC National Research Council 

NTP National Toxicology Program 

OM organic matter 

OPP Office of Pesticide Programs 

OPPTS Office of Pesticide Planning and Toxic Substances 

OSHA Occupational Safety and Health Administration 

ppb parts per billion 

ppm parts per million 

ppt parts per trillion 

PRZM Pesticide Root Zone Model 

RBC red blood cells 

RED re-registration eligibility decision 

RfD reference dose 

SERA Syracuse Environmental Research Associates 

SRC Syracuse Research Corporation 

UF uncertainty factor 

U.S. United States 

USDA U.S. Department of Agriculture 

U.S. EPA U.S. Environmental Protection Agency 

USGS U.S. Geological Survey 

WCR water contamination rate 

WHO World Health Organization 

WP wettable powder 

: micron or micro4CA 4-chloroaniline 

ix 
COMMON UNIT CONVERSIONS AND ABBREVIATIONS
 

To convert ... Into ... Multiply by ... 

acres hectares (ha) 0.4047 

acres square meters (m ) 2 4,047 

atmospheres millimeters of mercury 760 

centigrade Fahrenheit 1.8 °C+32 

centimeters inches 0.3937 

cubic meters (m ) 3 liters (L) 1,000 

Fahrenheit centigrade 0.556 °F-17.8 

feet per second (ft/sec) miles/hour (mi/hr) 0.6818 

gallons (gal) liters (L) 3.785 

gallons per acre (gal/acre) liters per hectare (L/ha) 9.34 

grams (g) ounces, (oz) 0.03527 

grams (g) pounds, (oz) 0.002205 

hectares (ha) acres 2.471 

inches (in) centimeters (cm) 2.540 

kilograms (kg) ounces, (oz) 35.274 

kilograms (kg) pounds, (lb) 2.2046 

kilograms per hectare (hg/ha) pounds per acre (lb/acre) 0.892 

kilometers (km) miles (mi) 0.6214 

liters (L) cubic centimeters (cm ) 3 1,000 

liters (L) gallons (gal) 0.2642 

liters (L) ounces, fluid (oz) 33.814 

miles (mi) kilometers (km) 1.609 

miles per hour (mi/hr) cm/sec 44.70 

milligrams (mg) ounces (oz) 0.000035 

meters (m) feet 3.281 

ounces (oz) grams (g) 28.3495 

ounces per acre (oz/acre) grams per hectare (g/ha) 70.1 

ounces per acre (oz/acre) kilograms per hectare (kg/ha) 0.0701 

ounces fluid cubic centimeters (cm ) 3 29.5735 

pounds (lb) grams (g) 453.6 

pounds (lb) kilograms (kg) 0.4536 

pounds per acre (lb/acre) kilograms per hectare (kg/ha) 1.121 

pounds per acre (lb/acre) mg/square meter (mg/m ) 2 112.1 

pounds per acre (lb/acre) :g/square centimeter (:g/cm ) 2 11.21 

pounds per gallon (lb/gal) grams per liter (g/L) 119.8 

square centimeters (cm ) 2 square inches (in ) 2 0.155 

square centimeters (cm ) 2 square meters (m ) 2 0.0001 

square meters (m ) 2 square centimeters (cm ) 2 10,000 

yards meters 0.9144 

Note: All references to pounds and ounces refer to avoirdupois weights unless otherwise specified. 

x 
CONVERSION OF SCIENTIFIC NOTATION
 

Scientific Decimal Verbal 

Notation Equivalent Expression 

1 @ 10-10 0.0000000001 One in ten billion 

1 @ 10-9 0.000000001 One in one billion 

1 @ 10-8 0.00000001 One in one hundred million 

1 @ 10-7 0.0000001 One in ten million 

1 @ 10-6 0.000001 One in one million 

1 @ 10-5 0.00001 One in one hundred thousand 

1 @ 10-4 0.0001 One in ten thousand 

1 @ 10-3 0.001 One in one thousand 

1 @ 10-2 0.01 One in one hundred 

1 @ 10-1 0.1 One in ten 

1 @ 100 1 One 

1 @ 101 10 Ten 

1 @ 102 100 One hundred 

1 @ 103 1,000 One thousand 

1 @ 104 10,000 Ten thousand 

1 @ 105 100,000 One hundred thousand 

1 @ 106 1,000,000 One million 

1 @ 107 10,000,000 Ten million 

1 @ 108 100,000,000 One hundred million 

1 @ 109 1,000,000,000 One billion 

1 @ 1010 10,000,000,000 Ten billion 

xi 
EXECUTIVE SUMMARY
 

OVERVIEW 

While the data base supporting the risk assessment of diflubenzuron is large and somewhat 

complex, the risk characterization is relatively simple. Diflubenzuron is an effective insecticide. 

Consequently, application rates used to control the gypsy moth are likely to have effects on some 

nontarget terrestrial insects. Species at greatest risk include grasshoppers, various 

macrolepidoptera (including the gypsy moth), other herbivorous insects, and some beneficial 

predators of the gypsy moth. Some aquatic invertebrates may also be at risk; however, the risks 

appear to be less severe than risks to terrestrial insects. The risk characterization for aquatic 

invertebrates is highly dependant on site-specific conditions. In areas subject to minimal water 

contamination, the effects of diflubenzuron are expected to be marginally adverse or nonexistent. 

If diflubenzuron is applied when drift or direct deposition in water is not controlled well or in 

areas where soil losses from runoff and sediment to water are likely to occur, certain aquatic 

invertebrates are at risk of acute adverse effects, and exposure could cause longer-term effects on 

more sensitive species. Direct effects of diflubenzuron on humans and other groups of 

organisms—wildlife mammals, birds, amphibians, fish, terrestrial and aquatic plants, 

microorganisms, and non-arthropod invertebrates—do not appear to be plausible. Nontarget 

species that consume the gypsy moth or other invertebrates adversely affected by diflubenzuron 

may be at risk of secondary effects of exposure (for example, a change in the availability of 

prey). There is no indication that 4-chloroaniline formed from the degradation of diflubenzuron 

will have an adverse effect on any species. 

PROGRAM DESCRIPTION 

Diflubenzuron is an insecticide that inhibits chitin deposition in arthropods and is effective either 

as a stomach or contact insecticide. Two formulations of diflubenzuron are labeled for control of 

the gypsy moth: Dimilin 4L and Dimilin 25W. Other formulations of diflubenzuron are available 

but these are registered for agricultural uses which account for about 94% of the total amount of 

diflubenzuron applied each year. Both ground and aerial applications of Dimilin 4L and Dimilin 

25W are permitted. The current risk assessment concerns the range of labeled application 

rates—i.e., 0.0078-0.0624 lbs a.i./acre. Virtually all use of diflubenzuron in USDA programs 

occurs in suppression programs (about 99% of the treated acres) with only about 1% of the use 

in slow the spread programs. The use of diflubenzuron in eradication programs is less than 

0.001% of the total use. 

HUMAN HEALTH RISK ASSESSMENT 

Hazard Identification – There is no information regarding effects in humans exposed to 

diflubenzuron; however, the toxicity of this compound is well characterized in experimental 

mammals. In mammals, the most sensitive effect involves damage to hemoglobin, a component 

of blood involved in the transport of oxygen. Diflubenzuron causes the formation of 

methemoglobin, a form of hemoglobin that is not able to transport oxygen. Methemoglobinemia, 

an excessive formation of methemoglobin, is the primary toxic effect of diflubenzuron regardless 

of the route or duration of exposure in every species of animal tested. Diflubenzuron causes 

xii 
other effects on the blood; however, methemoglobinemia is the most sensitive effect—that is, the 

effect that occurs at the lowest dose. While effects on the blood are well documented, there is 

little indication that diflubenzuron causes other specific forms of toxicity. Diflubenzuron does 

not appear to be neurotoxic or immunotoxic, does not appear to affect endocrine function in 

laboratory mammals, and is not a carcinogen. In addition, diflubenzuron does not appear to 

cause birth defects or reproductive effects. Diflubenzuron is relatively nontoxic by oral 

administration, with reported single-dose LD50 values ranging from greater than 4640 to greater 

than10,000 mg/kg. There are numerous studies regarding the subchronic and chronic toxicity of 

diflubenzuron in laboratory animals, and these studies indicate that methemoglobinemia is the 

most consistent and sensitive sign of toxicity. Diflubenzuron can be absorbed from the skin in 

sufficient amounts to cause hematological effects—that is, methemoglobinemia and 

sulfhemoglobinemia. Nonetheless, the dermal exposure concentrations that are necessary to 

cause these hematological effects are higher than the oral exposure doses that are necessary to 

cause the same effects. 

Exposure Assessment – Exposure assessments are conducted for both diflubenzuron and 

4-chloroaniline. For diflubenzuron, a standard set of exposure scenarios are presented for both 

workers and members of the general public. Concern for 4-chloroaniline arises because it is an 

environmental metabolite of diflubenzuron and is classified as a carcinogen. 4-Chloroaniline is 

not a concern in worker exposure assessments because 4-chloroaniline will not be present at the 

time that diflubenzuron is applied. Also, 4-chloroaniline is not a concern in some acute exposure 

scenarios for the general public such as direct spray during the application of diflubenzuron. 

Consequently, only a subset of the standard exposure scenarios—those associated with exposure 

to vegetation or water contaminated with diflubenzuron—are presented for 4-chloroaniline. 

These scenarios, however, include all standard chronic exposure scenarios, which are of greatest 

concern because of the potential carcinogenicity of 4-chloroaniline. 

All exposure assessments are conducted at the maximum single application rate for 

diflubenzuron of 0.0625 lb/acre (equivalent to 70 g/ha). This is also the maximum application 

rate for a single season. Assuming that diflubenzuron is applied in a single application at the 

maximum rate leads to the highest estimates of peak as well as longer-term exposures. The 

consequences of using lower application rates are discussed in the risk characterization. 

For workers applying diflubenzuron, three types of application methods are considered: directed 

ground spray, broadcast ground spray, and aerial spray. Central estimates of exposure for 

workers are approximately 0.0009 mg/kg/day for aerial workers, 0.0008 mg/kg/day for backpack 

workers, and about 0.001 mg/kg/day for broadcast ground spray workers. Upper ranges of 

exposures are approximately 0.009 mg/kg/day for broadcast ground spray workers and 0.005 

mg/kg/day for backpack and aerial workers. All of the accidental exposure scenarios for workers 

involve dermal exposures, and most of these accidental exposures lead to dose estimates that are 

either in the range of or substantially below the general exposure estimates for workers. The one 

exception involves wearing contaminated gloves for 1hour. The upper range of exposure for this 

scenario is about 0.4 mg/kg/day. 

xiii 
For the general public, estimates of acute exposure range from approximately 0.0000005 mg/kg, 

which is the lower range estimate for the consumption by a child of water from a stream 

contaminated by diflubenzuron, to 1.5 mg/kg, which represents the upper range for consumption 

of contaminated fish by subsistence populations—individuals who consume free-caught fish as a 

major proportion of their diet. Relatively high dose estimates are also associated with the 

consumption of contaminated water after an accidental spill (about 0.13 mg/kg at the upper range 

of exposure) and for the consumption of fish by members of the general public (0.3 mg/kg). 

Other acute exposures are lower by about an order of magnitude or more. For chronic or longerterm exposures, the modeled exposures are much lower than for acute exposures, ranging from 

approximately 0.00000002 mg/kg/day (2 in 10 millionths of a mg/kg/day), which is the lower 

range estimate for the consumption of contaminated water, to approximately 0.002 mg/kg/day, 

which is the upper range for consumption of contaminated fruit. 

Estimates of exposure to 4-chloroaniline from contaminated vegetation are likely to be about 

0.02 times less than corresponding estimates of exposure to diflubenzuron. The lower estimate 

of exposure to 4-chloroaniline is due to its expected rapid dissipation from diflubenzuron 

deposited on vegetation. In water, however, estimated concentrations of 4-chloroaniline are 

likely to be equal to or greater than anticipated water concentrations of diflubenzuron under 

certain circumstances. Finally, peak exposures to 4-chloroaniline differ from peak exposures to 

diflubenzuron in the environment, usually occurring at different times (later after the application 

of diflubenzuron) and under different conditions of precipitation. These differences are due to 

the relatively slow rate in the formation of 4-chloroaniline from diflubenzuron in soil. 

Dose-Response Assessment – The dose-response assessment considers both diflubenzuron itself 

as well as 4-chloroaniline as an environmental metabolite of diflubenzuron. For systemic 

toxicity, the dose-response assessment involves the adoption or derivation of acute and chronic 

RfDs, doses that are considered to produce no adverse effects, even in sensitive individuals. 

RfDs are presented for both diflubenzuron and 4-chloroaniline. Cancer risk is considered 

quantitatively for 4-chloroaniline and is expressed as a dose associated with a risk of 1 in 

1million. Following standard practices for USDA risk assessments, risk assessment values 

available from U.S. EPA are adopted directly unless there is a compelling basis for doing 

otherwise. When risk values are not available from U.S. EPA, the methods used by U.S. EPA 

are employed to derive surrogate values. 

U.S. EPA derived a chronic RfD for diflubenzuron of 0.02 mg/kg/day. This chronic RfD is well 

documented and is used directly for all longer-term exposures to diflubenzuron. This value is 

based on a NOAEL in dogs and an uncertainty factor of 100. Because of the low acute toxicity 

of diflubenzuron, the U.S. EPA did not derive an acute RfD but identified an acute NOAEL of 

10,000 mg/kg. While this NOAEL could be used to derive a surrogate acute RfD of 100 mg/kg, 

a more conservative approach is taken and a surrogate acute RfD of 11 mg/kg is derived based on 

a NOAEL of 1118 mg/kg from a study using a petroleum-based formulation of diflubenzuron. 

Since diflubenzuron is classified as a non-carcinogen by both U.S. EPA and WHO, there is no 

reason to conduct a quantitative cancer risk assessment for exposure to diflubenzuron. 

xiv 
The U.S. EPA derived a chronic RfD for 4-chloroaniline of 0.004 mg/kg/day, and this value is 

used in the current risk assessment to characterize risks from 4-chloroaniline for longer-term 

exposures. This RfD is based on a chronic oral LOAEL of 12.5 mg/kg/day using an uncertainty 

factor of 3000—three factors of 10 each for intraspecies extrapolation, sensitive subgroups, and 

the use of a LOAEL with an additional factor of 3 due to the lack of data reproductive toxicity 

data. As with diflubenzuron, the U.S. EPA did not derived an acute RfD for 4-chloroaniline. For 

this risk assessment a conservative approach is taken in which a surrogate acute RfD of 0.03 

mg/kg is based on a subchronic (90-day) NOAEL of 8 mg/kg/day. Consistent with the approach 

taken by U.S. EPA for the chronic RfD, an uncertainty factor of 300 is used—a factor of 10 for 

interspecies extrapolation, 10 for intraspecies extrapolation, and 3 for the lack of data on 

reproductive toxicity. For cancer risk, the U.S. EPA proposes a human cancer potency factor for 

4-chloroaniline of 0.0638 (mg/kg/day) -1. This potency factor is used to calculate a dose of 

1.6×10-5 mg/kg/day that would be associated with a cancer risk of 1 in 1million. 

Risk Characterization – The risk characterization for potential human health effects associated 

with the use of diflubenzuron in USDA programs to control the gypsy moth is relatively 

unambiguous: none of the hazard quotients reach a level of concern at the highest application rate 

that could be used in USDA programs. In that many of the exposure assessments involve very 

conservative assumptions—that is, assumptions that tend to overestimate exposure—and because 

the dose-response assessment is based on similarly protective assumptions, there is no basis for 

asserting that this use of diflubenzuron poses a hazard to human health. 

Notwithstanding the above assertion, it is worth noting that the greatest relative risk concerns the 

contamination of water with 4-chloroaniline rather than exposure to diflubenzuron itself. The 

highest hazard quotient for diflubenzuron is 0.1, a factor of 10 below a level of concern. Since 

this hazard quotient is based on toxicity, an endpoint that is considered to have a population 

threshold, the assertion can be made that risk associated with exposure to diflubenzuron is 

essentially zero. 

This is not the case with 4-chloroaniline, which is classified as a probable human carcinogen and 

is an environmental metabolite of diflubenzuron. For 4-chloroaniline, the highest hazard 

quotient is 0.4, below the level of concern by a factor of only 2.5. The scenario of greatest 

concern involves cancer risk from drinking contaminated water. This risk would be most 

plausible in areas with sandy soil and annual rainfall rates ranging from about 50 to 250 inches. 

The central estimate of the hazard quotient for the consumption of water contaminated with 

4-chloroaniline and based on a cancer risk of 1 in 1million is 0.09, which is 10 times lower than 

the level of concern. 

ECOLOGICAL RISK ASSESSMENT 

Hazard Identification – The toxicity of diflubenzuron is well characterized in most groups of 

animals, including mammals, birds, terrestrial invertebrates, fish, and aquatic invertebrates. In 

general, diflubenzuron is much more toxic to some invertebrates, specifically arthropods, than 

vertebrates or other groups of invertebrates. This differential toxicity appears to involve 

xv 
fundamentally different mechanisms of action. Toxicity to sensitive invertebrate species is based 

on the inhibition of chitin synthesis. In the more tolerant vertebrate species, the mechanism of 

action appears to be a specific effect on the blood that inhibits oxygen transport. 

The species most sensitive to diflubenzuron are arthropods, a large group of invertebrates, 

including insects, crustaceans, spiders, mites, and centipedes. Most of these organisms use 

chitin, a polymer (repeating series of connected chemical subunits) of a glucose-based molecule, 

as a major component of their exoskeleton—that is, outer body shell. Diflubenzuron is an 

effective insecticide because it inhibits the the formation of chitin. This effect disrupts the 

normal growth and development of insects and other arthropods. Both terrestrial and aquatic 

arthropods are affected but some substantial differences in sensitivity are apparent. In terrestrial 

organisms, the most sensitive species include lepidopteran and beetle larvae, grasshoppers and 

other herbivorous insects. More tolerant species include bees, flies, parasitic wasps, adult 

beetles, and sucking insects. In aquatic organisms, small crustaceans that consume algae and 

serve as a food source for fish (e.g., Daphnia species) appear to be the most sensitive to 

diflubenzuron, while larger insect species such as backswimmers and scavenger beetles are much 

less sensitive. A wide range of other aquatic invertebrates, other crustaceans ,and small to 

medium sized aquatic insect larvae, appear to have intermediate sensitivities. Not all 

invertebrates use chitin and these invertebrates are much less sensitive to diflubenzuron than the 

arthropods. For terrestrial invertebrates, relatively tolerant species include earthworms and 

snails. For aquatic species, tolerant species include ostracods and non-arthropods such as 

rotifers, bivalves (clams), aquatic worms, and snails. 

The most sensitive effect in vertebrate species concerns damage to blood cells involved in the 

transport of oxygen. This effect was demonstrated in laboratory mammals used in toxicity studies 

(for example, rats and mice) as well as in domestic animals and livestock. Although the effect 

was not studied in wildlife mammals, birds, or fish, it is reasonable to assume that hemoglobin in 

all vertebrate species could be affected by exposure to diflubenzuron. Acute exposures to 

diflubenzuron are relatively non-toxic to mammals and birds. The U.S. EPA places 

diflubenzuron in low toxicity categories (III or IV) for mammals and considers diflubenzuron to 

be virtually non-toxic to birds in acute exposures and only slightly toxic to birds in subchronic 

exposures. This assessment is supported by a numerous field studies in which no direct toxic 

effects in mammals or birds is reported. Effects, if any, on terrestrial vertebrates from the 

application of diflubenzuron are likely to be secondary to changes in food availability—that is, 

reduced numbers of insects—or changes in habitat— for example, the loss of protective 

vegetation, relative to areas not treated with diflubenzuron. Aquatic vertebrates also appear to be 

relatively tolerant to diflubenzuron ,and this compound is classified by U.S. EPA as practically 

non-toxic to fish. This classification appears to be appropriate and is supported by several 

longer-term toxicity studies and field studies. Changes in fish populations are reported in some 

studies; however, the changes appear to be secondary to changes in food supply. Although the 

data on amphibians is much more limited than the data onfish, a similar pattern is apparent—that 

is, although there are no direct toxic effects from exposure, changes in food consumption patterns 

appear secondary to direct effects on invertebrate species. 

xvi 
Data on plants and microorganisms are more limited than the data on invertebrates or vertebrates. 

Nonetheless, there does not appear to be any basis for asserting that diflubenzuron will have a 

substantial effect on these organisms. 

Exposure Assessment – As in the human health risk assessment, exposures are estimated for 

both diflubenzuron and 4-chloroaniline. A full set of exposure assessments are developed for 

diflubenzuron but only a subset of exposure assessments are developed for 4-chloroaniline. This 

approach is taken, again as in the human health risk assessment, because 4-chloroaniline is 

assessed as an environmental metabolite of diflubenzuron. Thus, immediately after application, 

the amount of 4-chloroaniline as an environmental metabolite will be negligible. Consequently, 

the direct spray scenarios as well as the consumption of insects and the consumption of small 

mammals after a direct spray are not included for 4-chloroaniline. Also as in the human health 

risk assessment, all standard chronic exposure scenarios are included for 4-chloroaniline. 

Terrestrial animals might be exposed to any applied pesticide from direct spray, the ingestion of 

contaminated media (vegetation, prey species, or water), grooming activities, or indirect contact 

with contaminated vegetation. For diflubenzuron, the highest acute exposures for small 

terrestrial vertebrates will occur after a direct spray and could reach up to about 10 mg/kg at an 

application rate of 70 g/ha. Exposures anticipated from the consumption of contaminated 

vegetation by terrestrial animals range from central estimates of about 0.08 mg/kg for a small 

mammal to 2 mg/kg for a large bird with upper ranges of about 0.2 mg/kg for a small mammal 

and 5 mg/kg for a large bird. The consumption of contaminated water leads to much lower levels 

of exposure. A similar pattern is seen for chronic exposures. Estimated longer-term daily doses 

for the a small mammal from the consumption of contaminated vegetation at the application site 

range from approximately 0.001 to 0.005 mg/kg. Large birds feeding on contaminated 

vegetation at the application site could be exposed to much higher concentrations, ranging from 

about 0.08 to 0.7 mg/kg/day. The upper ranges of exposure from contaminated vegetation far 

exceed doses anticipated from the consumption of contaminated water, which range from about 

0.0000001 to 0.00001 mg/kg/day for a small mammal. 

Exposures of terrestrial organisms to 4-chloroaniline tend to be much lower than those for 

diflubenzuron. The highest acute exposure is about 0.2 mg/kg, the approximate dose for the 

consumption of contaminated water by a small mammal and the consumption of contaminated 

fish by a predatory bird. The highest longer term exposure is 0.0002 mg/kg/day, the dose 

associated with the consumption of contaminated vegetation by a large bird. 

Exposures to aquatic organisms are based on essentially the same information used to assess the 

exposure to terrestrial species from contaminated water. At the maximum application rate of 70 

g/ha, the upper range of the expected peak concentration of diflubenzuron in surface water is 

taken as 16 μg/L. The lower range of the concentration in ambient water is estimated at 0.01 

μg/L. The central estimate of concentration of diflubenzuron in surface water is taken as 

0.4 μg/L. 

xvii 
Dose-Response Assessment – Diflubenzuron is relatively non-toxic to mammals and birds. For 

mammals, the toxicity values used in the ecological risk assessment are identical to those used in 

the human health risk assessments: an acute NOAEL of 1118 mg/kg and a chronic NOAEL of 2 

mg/kg/day. A similar approach is taken for 4-chloroaniline for which an acute NOAEL is 8 

mg/kg is used based on a subchronic study and a chronic NOAEL is estimated at 1.25 mg/kg/day 

based on the chronic LOAEL of 12.5 mg/kg/day. For birds, the acute NOAEL for diflubenzuron 

is taken as 2500 mg/kg from an acute gavage study and the longer-term NOAEL is taken as 110 

mg/kg/day from a reproduction study. No data are available regarding the toxicity of 

4-chloroaniline in birds and the available toxicity values for mammals are used as a surrogate. 

For terrestrial invertebrates two general types of data could be used to assess dose-response 

relationships: laboratory toxicity studies and field studies. Field studies are used in the current 

risk assessment because the standard toxicity studies are extremely diverse and many are not 

directly applicable to a risk assessment. Despite the difficulty and uncertainty in interpreting 

some of the field studies, the relatively large number of field studies on diflubenzuron appear to 

present a reasonably coherent pattern that is at least qualitatively consistent with the available 

toxicity data and probably a more realistic basis on which to assess risk to nontarget species. The 

most sensitive species appear to be grasshoppers which may be adversely affected at an 

application rate of 22 g/ha. Somewhat high application rates—in the range of 30 to 35 

g/ha—will adversely effect macrolepidoptera and some beneficial parasitic wasps. At the 

maximum application rate considered in this risk assessment— 70 g/ha—some herbivorous 

insects are likely to be affected. No adverse effects in several other groups of insects are 

expected at this or much higher application rates. Honeybees are among the most tolerant 

species and are not likely to be adversely affected at application rates of up to 400 g/ha. 

Invertebrates that do not synthesize chitin are also relatively tolerant to diflubenzuron. The 

NOEC for a species of earthworm (Eisenia fetida) is 780 mg/kg soil and is used to represent 

tolerant species of soil invertebrates. Very little information is available on the toxicity of 

4-chloroaniline to terrestrial invertebrates. As with diflubenzuron, the earthworm appears to be 

relatively tolerant to 4-chloroaniline with a reported LC50 value of 540 mg/kg dry soil. The 

toxicity of both diflubenzuron and 4-chloroaniline to soil microorganisms is also relatively low. 

Toxicity values for aquatic species follow a pattern similar to that for terrestrial species: 

arthropods appear to be much more sensitive than fish or non-arthropod invertebrates. For 

diflubenzuron, LC50 values of 25-500 mg/L are used to characterize risks for sensitive and 

tolerant species of fish, respectively. 4-Chloroaniline appears to be more toxic to fish and an 

LC50 of 2.4 mg/L is used to characterize risks of peak exposures, while an LC50 of 0.2 mg/L is 

used to characterize risks of longer-term exposures. 

There is substantial variability in the response of different groups of aquatic invertebrates to 

diflubenzuron. Very small arthropods appear to be among the most sensitive species—with 

acute NOEC values ranging from 0.3 to about 1 ppb (μg/L) and chronic NOEC values ranging 

from 0.04 to 0.25 ppb. Based on acute NOEC values, larger arthropods, including crabs and 

xviii 
larger insects, appear to be more tolerant, with acute NOEC values ranging from 2 to 2000 ppb. 

For chronic effects, the differences between small and larger arthropods are less remarkable with 

stoneflies and mayflies (relatively large insects) having an NOEC value of 0.1 ppb, intermediate 

between Daphnia (0.04 ppb) and Ceriodaphnia (0.25 ppb). Molluscs (invertebrates including 

clams and snails) and worms (oligochaetes) appear to be much less sensitive to diflubenzuron. 

The data on the toxicity of 4-chloroaniline to aquatic invertebrates is sparse. An acute NOEC of 

0.013 mg/L is used to characterize acute risks associated with peak exposures in aquatic 

invertebrates, and an NOEC of 0.01 mg/L from a reproduction study is used to characterize 

longer-term risks to aquatic invertebrates. 

Risk Characterization – While the data base supporting the risk assessment of diflubenzuron is 

large and somewhat complex, the risk characterization is relatively simple. Diflubenzuron is an 

effective insecticide. Consequently, application rates used to control the gypsy moth are likely to 

have effects on some nontarget terrestrial insects. Species at greatest risk include grasshoppers, 

various macrolepidoptera (including the gypsy moth), other herbivorous insects, and some 

beneficial predators to the gypsy moth. These species are at risk because of the mode of action of 

diflubenzuron (i.e., inhibition of chitin) and the behavior of the sensitive insects (the 

consumption of contaminated vegetation or predation on the gypsy moth). Some aquatic 

invertebrates may also be at risk but the risks appear to be less than risks to terrestrial insects. 

The risk characterization for aquatic invertebrates is highly dependant on site-specific conditions. 

If diflubenzuron is applied when drift or direct deposition in water is not controlled well or in 

areas where soil losses from runoff and sediment to water are likely to occur, certain aquatic 

invertebrates are at risk of acute adverse effects, and exposure could cause longer-term effects on 

more sensitive species. 

Direct effects of diflubenzuron on other groups of organisms—that is, mammals, birds, 

amphibians, fish, terrestrial and aquatic plants, microorganisms, and non-arthropod 

invertebrates—do not appear to be plausible. Nontarget species that consume the gypsy moth or 

other invertebrates adversely affected by diflubenzuron may be at risk of secondary effects of 

exposure (for example, a change in the availability of prey). There is no indication that 

4-chloroaniline formed from the degradation of diflubenzuron will have an adverse effect on any 

species 

There is no indication that 4-chloroaniline formed from the degradation of diflubenzuron will 

have an adverse effects on any species. 

xix 
1. INTRODUCTION
 

This document provides updated risk assessments for human health effects and ecological effects 

to support an assessment of the environmental consequences of using diflubenzuron for the 

control or eradication of the gypsy moth (Lymantria dispar) in USDA/Forest Service and 

USDA/APHIS programs. This risk assessment is an update to the human health and ecological 

risk assessments prepared for the 1995 Final Environmental Impact Statement (FEIS) for the 

Cooperative Gypsy Moth Management Program (USDA 1995). 

In the preparation of this risk assessment, literature searches on diflubenzuron were conducted 

using PubMed, TOXLINE, AGRICOLA, as well as the U.S. EPA CBI files. There is a very large 

body of literature on the environmental fate and toxicology of diflubenzuron. In addition to the 

previous risk assessments (USDA 1995), the toxicology, environmental fate, and other aspects 

associated with the use of diflubenzuron are the subject of relatively comprehensive reviews of 

human health and ecological effects by the World Health Organization (WHO 1996; WHO 

2001). Several other reviews of various topics involving diflubenzuron have been published in 

the open literature (e.g. Cunningham 1986; Eisler 1992; Fisher and Hall 1992; Wilson 1997) and 

in materials submitted to U.S. EPA (Cardona 1999; Hobson 2001; Lengen, 1999; Wilcox and 

Coffey 1978). 

In addition, a large number of studies have been submitted to the U.S. EPA/OPP in support of the 

registration of diflubenzuron and most of these studies have been reviewed by U.S. EPA (U.S. 

EPA/OPP 1997a, 1997b, 2000) and the derivation of food tolerances (EPA/OPP 1999, 2002a, 

2003). The U.S. EPA (1997a) re-registration eligibility decision (RED) document and other 

reviews by U.S. EPA include summaries of the product chemistry, mammalian toxicology, and 

ecotoxicology studies that were submitted by industry to the U.S. EPA. Full text copies of the 

studies most relevant to this risk assessment (n=118) were kindly provided by the U.S. EPA 

Office of Pesticide Programs. The CBI studies were reviewed, and synopses of the information 

that can be disclosed from these studies are included in this document. 

While this document discusses the studies required to support the risk assessments, it makes no 

attempt to re-summarize all of the information cited in the existing reviews. This is a general 

approach in all Forest Service risk assessments. For diflubenzuron in particular, an attempt to resummarize all of the available information would tend to obscure rather than clarify the key 

studies that should and do impact the risk assessment. 

The Forest Service will update this and other similar risk assessments on a periodic basis and 

welcomes input from the general public on the selection of studies included in the risk 

assessment. This input is helpful, however, only if recommendations for including additional 

studies specify why and/or how the new or not previously included information would be likely 

to alter the conclusions reached in the risk assessments. 

1-1
 
For the most part, the risk assessment methods used in this document are similar to those used in 

risk assessments previously conducted for the Forest Service as well as risk assessments 

conducted by other government agencies. Details regarding the specific methods used to prepare 

the human health risk assessment are provided in SERA (2001). This document has four 

chapters, including the introduction, program description, risk assessment for human health 

effects, and risk assessment for ecological effects or effects on wildlife species. Each of the two 

risk assessment chapters has four major sections, including an identification of the hazards 

associated with diflubenzuron and its commercial formulations, an assessment of potential 

exposure to the product, an assessment of the dose-response relationships, and a characterization 

of the risks associated with plausible levels of exposure. These are the basic steps recommended 

by the National Research Council of the National Academy of Sciences (NRC 1983) for 

conducting and organizing risk assessments. 

Risk assessments are usually expressed with numbers; however, the numbers are far from exact. 

Variability and uncertainty may be dominant factors in any risk assessment, and these factors 

should be expressed. Within the context of a risk assessment, the terms variability and 

uncertainty signify different conditions. 

Variability reflects the knowledge of how things may change. Variability may take several 

forms. For this risk assessment, three types of variability are distinguished: statistical, 

situational, and arbitrary. Statistical variability reflects, at least, apparently random patterns in 

data. For example, various types of estimates used in this risk assessment involve relationships 

of certain physical properties to certain biological properties. In such cases, best or maximum 

likelihood estimates can be calculated as well as upper and lower confidence intervals that reflect 

the statistical variability in the relationships. Situational variability describes variations 

depending on known circumstances. For example, the application rate or the applied 

concentration of a herbicide will vary according to local conditions and goals. As discussed in 

the following section, the limits on this variability are known and there is some information to 

indicate what the variations are. In other words, situational variability is not random. Arbitrary 

variability, as the name implies, represents an attempt to describe changes that cannot be 

characterized statistically or by a given set of conditions that cannot be well defined. This type 

of variability dominates some spill scenarios involving either a spill of a chemical on to the 

surface of the skin or a spill of a chemical into water. In either case, exposure depends on the 

amount of chemical spilled and the area of skin or volume of water that is contaminated. 

Variability reflects a knowledge or at least an explicit assumption about how things may change, 

while uncertainty reflects a lack of knowledge. For example, the focus of the human health 

dose-response assessment is an estimation of an ‘acceptable’ or ‘no adverse effect’ dose that will 

not be associated with adverse human health effects. For diflubenzuron and for most other 

chemicals, however, this estimation regarding human health must be based on data from 

experimental animal studies, which cover only a limited number of effects. Generally, judgment 

is the basis for the methods used to make the assessment. Although the judgments may reflect a 

consensus (i.e., be used by many groups in a reasonably consistent manner), the resulting 

1-2
 
estimations of risk cannot be proven analytically. In other words, the estimates regarding risk 

involve uncertainty. The primary functional distinction between variability and uncertainty is 

that variability is expressed quantitatively, while uncertainty is generally expressed qualitatively. 

In considering different forms of variability, almost no risk estimate presented in this document 

is given as a single number. Usually, risk is expressed as a central estimate and a range, which is 

sometimes very large. Because of the need to encompass many different types of exposure as 

well as the need to express the uncertainties in the assessment, this risk assessment involves 

numerous calculations. Some of the calculations are relatively simple are included in the body of 

the document. Some sets of the calculations, however, are cumbersome. For those calculations, 

worksheets are included with this risk assessment. The worksheets provide the detail for the 

estimates cited in the body of the document. Documentation for these worksheets is provided in 

a separate document (SERA 2003). A set of worksheets is provided for diflubenzuron 

(Supplement 1) as well as 4-chloroaniline (Supplement 2). As discussed in this risk assessment, 

4-chloroaniline is a metabolite of diflubenzuron that is quantitatively considered in this risk 

assessment. Both sets of worksheets are provided with the hard-text copy of this risk assessment 

as well as with the electronic version of the risk assessment. 

This is a technical support document and it addresses some specialized technical areas. 

Nevertheless, an effort was made to ensure that the document can be understood by individuals 

who do not have specialized training in the chemical and biological sciences. Certain technical 

concepts, methods, and terms common to all parts of the risk assessment are described in plain 

language in a separate document (SERA 2001). General glossaries of environmental terms are 

widely available and a custom glossary designed to be used in conjunction with USDA risk 

assessments is available at www.sera-inc.com. Some of the more complicated terms that are 

specific to diflubenzuron are defined in the text of this risk assessment. 

1-3
 
2. PROGRAM DESCRIPTION
 

2.1. Overview 

Diflubenzuron is an insecticide that inhibits chitin deposition in arthropods and is effective either 

as a stomach or contact insecticide. Two formulations of diflubenzuron are labeled for control of 

the gypsy moth: Dimilin 4L and Dimilin 25W. Other formulations of diflubenzuron are available 

but these are registered for agricultural uses, which account for about 94% of the total amount of 

diflubenzuron applied each year. Both ground and aerial applications of Dimilin 4L and Dimilin 

25W are permitted. For the current risk assessment, the range of labeled application rates – i.e., 

0.0078 lb a.i./acre to 0.0624 lbs a.i./acre – are considered. Virtually all use of diflubenzuron in 

USDA programs occurs in suppression programs (about 99% of treated acres) with only about 

1% of the use in slow the spread programs. The use of diflubenzuron in eradication programs is 

less than 0.001% of the total use. 

2.2. Chemical Description and Commercial Formulations 

Diflubenzuron is the common name for [1-(4-chlorophenyl) 3-(2,6-difluorobenzoyl)urea]: 

Structurally, diflubenzuron consists of p-chloroanaline (the moiety on the left) linked to a 2,6­

difluorobenzoic acid (the moiety on the right) by a ureido (carbon-nitrogen) bridge. Other 

synonyms for diflubenzuron as well as selected chemical and physical properties of 

diflubenzuron are summarized in Table 2-1. Additional information on the environmental fate 

and transport of diflubenzuron is summarized in the exposure assessments for the human health 

risk assessment (Section 3.2) and ecological risk assessment (Section 4.2). 

Diflubenzuron is an insecticide that inhibits chitin deposition in arthropods and is effective either 

as a stomach or contact insecticide (Mabury and Crosby 1996). Chitin is a polymer (repeating 

series of connected chemical subunits) of a glucose-based molecule and comprises a substantial 

proportion of the exoskeleton (outer-shell) of arthropods. Consequently, the inhibition of chitin 

synthesis disrupts the growth and development (Baishya and Hazarika 1996; DeCleraq et al. 

1995a,b; Griffith et al. 1996; Post and others 1974; Wright et al. 1996). Thus, diflubenzuron is 

not specific to the gypsy moth (Griffith et al. 1996; Horst and Walker 1995; Kadam et al. 1995) 

and is used to control a variety of pests on a variety of vegetation (Booth Riedl 1996; Boyle et al. 

1996; McCasland et al. 1998). Because diflubenzuron can impact a number of invertebrate 

species, particularly aquatic species (e.g., Liber et al. 1996; O’Halloran et al. 1996), this 

compound is a restricted use pesticide that may only be applied by licenced applicators (C&P 

Press 2004). 

2-1
 
Various formulations of diflubenzuron are labeled for forestry applications as well as other 

applications. All formulations of diflubenzuron are currently registered to Uniroyal Chemical 

(Table 2-2). Two formulations of diflubenzuron are labeled for control of the gypsy moth: 

Dimilin 4L and Dimilin 25W. As indicated in Table 2-2, an additional formulation, Micromite 

25W, had been registered for gypsy moth but this formulation has been discontinued and the 

registration for this product has been canceled (U.S. EPA/OPP 2002b). Micromite 25WS and 

Micromite 25WGS are still available but these formulations are not used in USDA programs for 

the control of the gypsy moth. 

Information on the impurities in and composition of these and other formulations of 

diflubenzuron have been submitted to U.S. EPA/OPP and this information (i.e., Drozdick 

1998a,b,c,d,e; Van Kampen and Thus 1996; Vanstone 1998a,b,c; White 1998) has been reviewed 

as part of the current risk assessment. Specific information on inerts and contaminants in the 

diflubenzuron formulations is classified as CBI (confidential business Information) under 

Section 7(d) and Section (10) of FIFRA. This information cannot be specifically disclosed in 

this risk assessment. WHO (1996) has reported in the open literature that at least some processes 

in the synthesis diflubenzuron involve the reaction of 2,6-difluoro-benzamide with 

4-chlorophenylisocyanate. Some inerts, however, must be disclosed on the material safety data 

sheet. Dimilin 4L contains petroleum oil [CAS No. 64742-46-7] and Dimilin 25W contains 

kaolin clay (C&P Press 2004). WHO (1996) indicated that kaolin is the only inert in some 

formulations of diflubenzuron. The potential risks associated with these inerts in the 

diflubenzuron formulations are discussed in Section 3.1.14. 

2.3. Application Methods and Rates 

Both ground and aerial applications of Dimilin 4L and Dimilin 25W are permitted (C&P Press 

2004) and both methods are used in USDA programs. The most common methods for ground 

applications of diflubenzuron are hydraulic sprayers, mist blowers, or air blast sprayers 

(broadcast foliar). The spray equipment is typically mounted on tractors or trucks used to apply 

the insecticide on either side of the roadway. Usually, about 8 acres are treated in a 45-minute 

period (approximately 11 acres/hour). Special truck-mounted spray systems may be used to treat 

up to 12 acres in a 35-minute period with approximately 300 gallons of insecticide mixture 

(approximately 21 acres/hour and 510 gallons/hour) (USDA/FS89b, p 2-9 to 2-10). 

In some instances, directed foliar applications may be used. In selective foliar applications, 

backpack applicators are used and the insecticide is applied to target vegetation. Application 

crews may treat up to shoulder high brush, which means that chemical contact with the arms, 

hands, or face is plausible. To reduce the likelihood of significant exposure, application crews 

are directed not to walk through treated vegetation. Usually, a worker treats approximately 0.5 

acres/hour with a plausible range of 0.25-1.0 acre/hour. 

2-2
 
In aerial applications, diflubenzuron formulations are applied under pressure through specially 

designed spray nozzles and booms. The nozzles are designed to minimize turbulence and 

maintain a large droplet size, both of which contribute to a reduction in spray drift. In aerial 

applications, approximately 40 to 100 acres may be treated per hour (USDA/FS89b, p 2-11). For 

Dimilin 25W, recommended droplet sizes are in the range of 150 to 200 microns (C&P Press 

2004). 

As indicated in Table 2-2, the application rate for Dimilin 4L ranges from 0.5 fluids ounces to 2 

fluid ounces per acre. This corresponds to about 0.0039 to 0.0156 gallons [128 ounces per 

gallon] of Dimilin 4L per acre, which in turn corresponds to about 0.0156 to 0.0624 lbs 

diflubenzuron per acre [4 lbs diflubenzuron per gallon × 0.0039 to 0.0156 ] and 17 to 70 

grams/ha. While multiple applications are permitted, the maximum single application rate is 

equal to the maximum annual application rate. 

For Dimilin 25W, the range of labeled application rates is 0.5 ounces (avoirdupois) to 2 ounces 

per acre or 0.03125 to 0.125 pounds of Dimilin 25W per acre [i.e., 16 avoirdupois ounces per 

pound]. Since Dimilin 25W consists of 25% diflubenzuron, this range of application rates is 

equivalent to about 0.0078 to 0.03125 lb diflubenzuron per acre and 9 to 35 grams/ha. These 

rates for Dimilin 25W are about a factor of two below the corresponding rates for Dimilin 4L. 

The maximum application rate for Dimilin 25W in a single application is equivalent to the 

maximum annual application rate – i.e., multiple applications are allowed each year but the total 

amount applied in a single year cannot exceed 0.03125 lb a.i./acre [35 g/ha]. 

For the current risk assessment, the range of labeled application rates – i.e., 0.0078 lb a.i./acre to 

0.0624 lbs a.i./acre – are considered. As calculated above, these rates are equivalent to 9 g/ha to 

70 g/ha. All exposure assessments will be conducted at the maximum application rate. The 

consequences of using lesser rates are considered further in the risk characterization for human 

health (Section 3.4) and ecological effects (Section 4.4). These application rates are essentially 

the same as those used in the previous risk assessment (USDA 1995). 

Recommended high volume ground sprays of Dimilin 4L and Dimilin 25W typically involve 100 

to 400 gallons per acre but much concentrated solutions – i.e., 5 to 30 gallons per acre – are used 

in aerial applications. For the current risk assessment, the central value is taken as 30 gallons per 

acre and the range is taken as 5 to 400 gallons per acre. It should be noted that the selection of 

application rates and dilution volumes in this risk assessment is intended to simply reflect typical 

or central estimates as well as lower and upper ranges. In the assessment of specific program 

activities, the Forest Service will use program specific application rates in the worksheets that are 

included with this report to assess any potential risks for a proposed application. 

The product label for Dimilin 25W specifically requires a 25 foot buffer for ground applications 

and a 150 foot buffer for aerial applications. These buffers indicate an area between the treated 

area and open bodies of water that may not be treated with diflubenzuron. The product label for 

Dimilin 4L does not specify a buffer but does indicate that the formulation cannot be applied to 

2-3
 
water or “...to areas where surface water is present” (C&P Press 2004). In the aerial or ground 

applications, the USDA will use at least a 100 foot buffer and will extend the buffer up to 500 

feet in some instances (Cook 2004). 

2.4. Use Statistics 

In order to minimize the ecological effects and human health effects of gypsy moth infestations, 

the USDA has adopted various intervention strategies that are roughly categorized as 

suppression, eradication, and slow the spread (USDA 1995). Suppression efforts are conducted 

by the USDA Forest Service in areas of well established gypsy moth infestations to combat or 

interdict periodic gypsy moth population outbreaks. Eradication efforts are conducted by 

USDA/APHIS to completely eliminate gypsy moth populations in areas where new populations 

of the gypsy moth are found. Slow the spread, as the name implies, is a program to reduce the 

expansion of gypsy moth populations from areas of established populations to adjacent noninfested areas. 

As indicated in Table 2-3, a total of 664,560 acres were treated with diflubenzuron formulations 

between 1995 and 2003, for an average annual treatment of about 73,840 acres per year. 

Virtually all (about 99%) of this use occurred in suppression programs with only about 1% of the 

use slow the spread programs. Very little diflubenzuron has been used in eradication programs – 

i.e., only 6 acres were treated in eradication programs accounting for <0.001% of the total acres 

treated for suppression, eradication, and slow the spread combined. Complete statistics for the 

amount of diflubenzuron applied in these applications has not been encountered. At the 

maximum labeled rate of 0.0624 lbs a.i./acre, the average annual treatment of about 73,840 acres 

per year would correspond to about 4608 pounds per year. 

By comparison, the annual use of diflubenzuron on cotton for 1992 (the most recent year for 

which statistics are available) was 78,013 lbs (USGS 1998) or about a factor of 17 above the 

estimated average annul use by the Forest Service. The low use of the diflubenzuron by the 

USDA relative to agricultural applications – i.e., about 5.6% [4608 ÷ (78,013 + 4608) = 0.0558] 

– indicates that the use of diflubenzuron by the USDA will not contribute substantially to general 

levels of diflubenzuron in the environment. This 5.6% figure probably overestimates the use of 

diflubenzuron by the USDA relative to agricultural applications because USGS (1998) reports 

only use on cotton. Diflubenzuron is registered for application to a number of other agricultural 

crops. Nonetheless, localized release of diflubenzuron will occur and the consequences of this 

release is considered in the remainder of this risk assessment. 

2-4
 
3. HUMAN RISK ASSESSMENT
 

3.1. HAZARD IDENTIFICATION 

3.1.1. Overview 

No information is available on the effects of diflubenzuron on humans but the toxicity of this 

compound has been well characterized in experimental mammals. In mammals, the most 

sensitive effect involves damage to hemoglobin, a component of blood involved in the transport 

of oxygen. Diflubenzuron causes the formation of methemoglobin, a form of hemoglobin that is 

not able to transport oxygen. Methemoglobinemia, an excessive formation of methemoglobin, is 

the primary toxic effect of diflubenzuron by all routes of exposure and for all durations of 

exposure in all species of animals that have been tested. Diflubenzuron causes other effects on 

the blood but methemoglobinemia is the most sensitive effect – i.e., the effect that occurs at the 

lowest dose. While effects on the blood are well documented, there is little indication that 

diflubenzuron causes other specific forms of toxicity. Diflubenzuron does not appear to be 

neurotoxic or immunotoxic, does not appear to affect endocrine function in laboratory mammals, 

and is not a carcinogen. In addition, diflubenzuron does not appear to cause birth defects or 

reproductive effects. Diflubenzuron is relatively nontoxic by oral administration, with 

single-dose LD50 values reported as > 4640 mg/kg to >10,000 mg/kg. A large number of studies 

on the subchronic and chronic toxicity of diflubenzuron are available. As with acute toxicity, 

methemoglobinemia is the most consistent and sensitive sign of toxicity in laboratory mammals. 

Diflubenzuron can be absorbed from the skin in sufficient amounts to cause hematologic effects 

– e.g., methemoglobinemia and sulfhemoglobinemia. Nonetheless, these effects occur at higher 

doses after dermal administration than after oral administration. 

3.1.2. Mechanisms of Action 

Some specific mechanisms of action for diflubenzuron are well understood in both mammals and 

invertebrates. As discussed in Section 4.1, diflubenzuron inhibits chitin synthesis in 

invertebrates and this in turn disrupts normal growth and development and can lead to death. 

Mammals, including humans, do not produce chitin and this mechanism thus has no relevance to 

the human health risk assessment. Another mechanism of diflubenzuron involves damage to 

hemoglobin, a key component of blood, through the development of methemoglobin and 

sulfhemoglobin. This is highly relevant to the human health risk assessment and the formation of 

methemoglobin is the basis for the U.S. EPA RfD for diflubenzuron (Section 3.3). 

Hemoglobin is the component in red blood cells that is responsible for transporting oxygen 

throughout the body. If this function is impaired, either because of damage to hemoglobin (Hb) 

or lack of oxygen in the air, serious adverse effects (i.e., equivalent to suffocation) can occur. 

The formation of both methemoglobin and sulfhemoglobin can cause such impairment and lead 

to the formation of methemoglobinemia and sulfhemoglobinemia, respectively. Methemoglobin 

is formed by the oxidation of the heme iron in hemoglobin from the ferrous to the ferric state 

(Bradberry 2003; Smith 1996). Heme group oxidation occurs spontaneously and accounts for 

approximately 2% of the hemoglobin in normal individuals. Methemoglobin is reduced (restored 

to its natural state) by a set of enzymes referred to as methemoglobin reductases. The most 

3-1
 
common methemoglobin reductase is dependent on NADH, a molecule that is common in all 

living systems and is necessary for the proper function of many enzymes (Lo and Agar 1986). 

Some individuals are deficient in NADH-dependent methemoglobin reductase, in which case as 

much as 50% of their blood pigment may exist as methemoglobin. Newborns are also deficient 

in NADH-methemoglobin reductase. As discussed further in Section 3.1.15 (Impurities and 

Metabolites), 4-chloroaniline, a metabolite of diflubenzuron, has also been shown to induce 

methemoglobinemia (WHO 2003). 

Sulfhemoglobinemia is characterized by the presence of abnormal pigments, other than 

methemoglobin, in red cells and can be regarded as a form of nonspecific oxidative damage 

(Smith 1996) and, in some cases, the differential diagnosis of sulfhemoglobinemia and 

methemoglobinemia may be difficult (Demedts et al. 1997). As with methemoglobinemia, 

sulfhemoglobinemia can be induced by aromatic amines and hydroxyamines. Unlike 

methemoglobinemia, sulfhemoglobinemia is irreversible. Sulfhemoglobinemia is associated 

with the formation of Heinz bodies, dark-staining granules found in red blood cells. The 

formation of Heinz bodies can lead to red cell dysfunction and hemolysis (breakdown of the cell 

membrane). The damaged cells are in turn captured by the spleen, which can lead to spleen 

enlargement. In general, cats, mice, dogs, and humans are more susceptible to Heinz body 

formation compared with rabbits, monkeys, chickens, and guinea pigs (Smith 1996). Studies on 

the effects of diflubenzuron on methemoglobin, sulfhemoglobin, Heinz body formation, and the 

spleen are summarized in Appendix 1. These data are discussed in further detail in Section 3.3 

(Dose-Response Assessment). 

While diflubenzuron displays other types of toxicity, as discussed in the following subsections, 

the formation of methemoglobin and sulfhemoglobin are the only mechanisms of toxicity that 

have been clearly identified. 

3.1.3. Kinetics and Metabolism 

3.1.3.1. Oral Absorption – Diflubenzuron appears to be readily absorbed after oral 

administration but the extent of absorption is dose-dependant. Cameron et al. (1990) conducted 

a standard pharmacokinetic study on diflubenzuron in rats. Diflubenzuron was rapidly absorbed 

and excreted in both the urine and feces. Urine showed significant levels of 2,6-difluorobenzoic 

acid, 2,6-difluorophippuric acid, 2,6-difluorobenzaimide, 4-chlorophenyl urea, and 2'­

hydroxydiflubenzuron. Fecal excretion contained mostly unchanged parent compound. 

4-Chloroaniline was not detected in urine or bile (limit of detection = 7.5 ng/mL). As discussed 

further below, 4-chloroaniline is a metabolite of diflubenzuron in some species (Section 3.1.3.3) 

and is an environmental metabolite of diflubenzuron formed by biodegradation in soil. The oral 

absorption of diflubenzuron appears to be dependent on dose (e.g., Willems et al. 1980). At 

relatively low doses, in the range of 1 mg/kg/day, a substantial fraction of administered 

diflubenzuron (about 50%) is absorbed. At much higher doses, in the range of 1000 mg/kg/day, 

much less diflubenzuron is absorbed (about 5%) (WHO 1996, 2001). While studies on the basis 

for this dose-dependent absorption have not been located for diflubenzuron, this is a relatively 

common pattern in many compounds that are highly lipophilic – i.e., tend to concentrate in fat 

3-2
 
tissue – and probably involves saturable transport by the lymphatic system (e.g., Rozman et al. 

1979). 

3.1.3.2. Dermal Absorption – No studies have been found on the dermal absorption of 

diflubenzuron in humans. Dermal absorption in rats has been studied by Andre (1996) and this 

study is summarized in Appendix 1. The dermal absorption of diflubenzuron appeared to be 

linear for doses of 0.005 or 0.05 mg/cm . 2 This is unlike the pattern with oral absorption, as noted 

above, but the dermal doses are very low. In addition and unlike the case with oral absorption, 

there is no basis for asserting that dermal absorption is saturable. Andre (1996) does not provide 

a kinetic analysis of the absorption data. Andre (1996) does note that about 6% of the dose was 

bound to skin and that less than 1% of the dose was absorbed systemically over a 10 hour period. 

Taking 1% as an approximate measure of absorbed dose, the dermal absorption coefficient would 

-1 be about 0.001 hour-1 [k = -ln(1-0.01)/10 hour = 0.001 hour ].

While several additional studies are available on the toxicity of diflubenzuron after dermal 

administration (Section 3.1.12.), these studies do not address the kinetics of dermal absorption. 

WHO (1996, 2001) summarizes an unpublished study conducted in the Netherlands indicating 

that 0.2% of a dermal dose of 150 mg/kg was absorbed by rabbits over a 6 hour exposure period. 

This corresponds to a dermal absorption rate of about 0.04 hour-1 [k = -ln(1-0.002)/6 hours = 

-1 0.000358 hour ], substantially less than the estimate in rats from the study by Andre (1996).

Estimates of first-order dermal absorption rates can also be made from structure activity 

relationships (SERA 2001). Based on these relationships, the estimated first-order dermal 

absorption rate for diflubenzuron is 0.0044 hour-1 with a 95% confidence interval of 0.0019 

-1 -1 hour to 0.01 hour (Worksheet A09). These estimate first-order dermal absorption rates are 

somewhat higher than those based on experimental measurements. The higher dermal 

absorption rates from Worksheet A09 are used in the current risk assessment. While this is a 

somewhat conservative or protective approach, it has little impact on the risk characterization 

(Section 3.4) because none of the exposures based on these conservative estimates approach a 

level on concern. 

Dermal exposure scenarios involving immersion or prolonged contact with chemical solutions 

use Fick's first law and require an estimate of the permeability coefficient, K , expressed in p

cm/hour (SERA 2001). Using the method recommended by U.S. EPA/ORD (1992), the 

estimated dermal permeability coefficient for diflubenzuron is 0.012 cm/hour with a 95% 

confidence interval of 0.0066 to 0.021 cm/hour. The application of this method to diflubenzuron 

is given in Worksheet A10. 

Note that the first-order and zero-order absorption coefficients are summarized in Worksheet 03 

but are rounded to two significant places. Links to these values are used in all of the exposure 

worksheets involving dermal absorption. 

3-3
 
3.1.3.3. Metabolism – Two types of metabolites are considered in this risk assessment: 

metabolites that are formed in vivo by an animal after diflubenzuron has been absorbed and 

metabolites that the formed in the environment through the degradation of diflubenzuron in 

environmental media – i.e., soil, air, and water. The in vivo metabolism of diflubenzuron has 

been reviewed by WHO (1996, 2001) and additional unpublished studies have been submitted to 

the U.S. EPA on the metabolism of diflubenzuron in rats (Cameron et al. 1990; Gay et al. 1999) 

as well as the environmental metabolism of diflubenzuron (Dzialo and Maynard 1999; Thus et al. 

1991; Walstra and Joustra, 1990). 

An overview of the in vivo and environmental metabolism of diflubenzuron is given in 

Figure 3-1. Two basic pathways exist for the metabolism of diflubenzuron. In the environment 

as well as in sheep, pigs, and chickens, the major route of metabolism involves cleavage of the 

ureido bridge with the formation of 2,6-difluorobenzoic acid and 4-chlorophenyl urea. The latter 

compound is then metabolized to 4-chloroaniline. As discussed further in Section 3.1.15, the 

formation of 4-chloroaniline is important to the human health risk assessment because this 

compound is classified as a carcinogen. The other pathway for the metabolism of diflubenzuron 

predominates in rats and cows and involves hydroxylation rather than cleavage of the ureido 

bridge. Hydroxylation of the aromatic rings involves the addition of a hydrogen-oxygen or 

hydroxy (OH) group to one of the rings. Hydroxylation increases the water solubility of aromatic 

compounds. Particularly when followed by conjugation with other water soluble compounds in 

the body, such as sugars or amino acids, hydroxylation greatly facilitates the elimination of the 

compound in the urine or bile. As detailed further by WHO (2001), the ureido bridge may also 

be cleaved in rats but 4-chloroaniline does not appear to be a major metabolite. No information 

has been located on the metabolism of diflubenzuron in humans. 

3.1.4. Acute Oral Toxicity 

No information has been found on the acute toxicity of diflubenzuron in humans. Information 

regarding the acute toxicity of diflubenzuron and diflubenzuron formulations in laboratory 

mammals is summarized in Appendix 1. These data indicate that diflubenzuron is relatively 

nontoxic by oral administration, with single dose LD50 values in mice and rats reported as > 4640 

mg/kg to >10,000 mg/kg. In other words, less than half of the animals died at these doses. Many 

of the exposure scenarios considered in the current risk assessment for the use of diflubenzuron 

for the control of the gypsy moth do involve very short term acute exposures and the use of acute 

oral toxicity values is considered further in Section 3.3.3. 

3.1.5. Subchronic and Chronic Toxicity 

No information has been found on the subchronic or chronic toxicity of diflubenzuron in 

humans. A large number of studies using experimental mammals are available on the subchronic 

and chronic toxicity of diflubenzuron. Studies most relevant to the current risk assessment as 

summarized in Appendix 1 and additional information, including unpublished studies conducted 

in Europe, are summarized by WHO (1996, 2000). 

3-4
 
As with acute toxicity, methemoglobinemia is the most consistent and sensitive sign of toxicity 

in laboratory mammals and has been observed in all mammalian species on which bioassays have 

been conducted: cats (Keet et al. 1982), dogs (Chesterman et al. 1974; Keet et al. 1982; 

Greenough et al. 1985), mice (Colley et al. 1981; Colley et al. 1984; Keet et al. 1984b), rats 

(Berberian and Enan 1989; Burdock et al. 1980; Burdock 1984; Keet et al. 1984a), and sheep 

(Keet et al. 1982). 

For the current risk assessment, the most relevant longer-term toxicity study is the one-year oral 

toxicity study in which dogs were administered diflubenzuron in gelatin capsules at doses of 0, 2, 

10, 50, or 250 mg/kg/bw (Greenough et al. 1985). As indicated in Appendix 1 and discussed 

further in Section 3.3.2, this is the study on which the U.S. EPA (1988; 1997a; 2000) has based 

the chronic RfD. In this study, no clinical signs of toxicity or pathology attributable to treatment 

were observed. The only adverse effects that were observed included dose-related increases in 

methemoglobin and sulfhemoglobin accompanied by an increase in spleen weight. As noted in 

the previous section, the increased spleen weight is probably secondary to the hematologic 

effects of diflubenzuron. This study is also important in that a clear duration-response 

relationship is apparent, with no significant changes in methemoglobin and sulfhemoglobin 

concentrations at four weeks after the start of dosing. 

3.1.6. Effects on Nervous System 

As discussed in Durkin and Diamond (2002), a neurotoxicant is a chemical that disrupts the 

function of nerves, either by interacting with nerves directly or by interacting with supporting 

cells in the nervous system. This definition of neurotoxicant distinguishes agents that act directly 

on the nervous system (direct neurotoxicants) from those agents that might produce neurologic 

effects that are secondary to other forms of toxicity (indirect neurotoxicants). Virtually any 

chemical will cause signs of neurotoxicity in severely poisoned animals and, thus, can be 

classified as an indirect neurotoxicant. 

Diflubenzuron, however, evidences few characteristics of a neurotoxicant even in terms of 

indirect effects. In an acute inhalation study involving a diflubenzuron formulation not used by 

the USDA (i.e, Dimilin 2L), excessive salivation and labored breathing were observed both 

during and after exposure (Hoffman 1997). While these can be signs of neurologic effects, they 

can be secondary to general irritation as well as other toxic effects. The only study on 

diflubenzuron that specifically assayed for neurotoxicity is the inhalation study by Newton 

(1999) in rats (details in Appendix 1). The neuro-behavioral battery included assays for 

autonomic effects, central nervous system effects (e.g., tremors and convulsions), general motor 

activity, movement and posture, reactivity to handling or sensory stimuli, grip strength, and 

observations for atypical behavior. Newton (1999) noted no treatment related effects of any 

endpoints assayed. The review of this study by WHO (2001) indicates that: “A reduction in ‘grid 

count’ was evident in the neuro-functional assessment of males and females exposed to 110 

3 mg/m .” Here, grid count refers to the number of grids that both front feet simultaneously 

touched during a fixed observations period. Based on the data reported in Newton (1999) for 

males (summary in Table 3, p. 44 and individual data in Appendix pp. 150-151 in Newton 1999) 

3-5
 
and females (summary Table 3, p. 47 and individual data in Appendix pp. 168-169 in Newton 

1999), a slight reduction in mean grid count is apparent for this response in study weeks 1, 2, and 

3 but not in study week 4. There is, however, substantial scatter in the individual data in terms of 

the relationship of the response to concentration. The significance of the changes in grid count in 

the absence of any other sign of neurotoxicity is questionable. 

3.1.7. Effects on Immune System 

Immunotoxicants are chemical agents that disrupt the function of the immune system. Two 

general types of effects, suppression and enhancement, may be seen and both of these are 

generally regarded as adverse. Agents that impair immune responses (immune suppression) 

enhance susceptibility to infectious diseases or cancer. Enhancement or hyperreactivity can give 

rise to allergy or hypersensitivity, in which the immune system of genetically predisposed 

individuals inappropriately responds to chemical or biological agents (e.g., plant pollen, cat 

dander, flour gluten) that pose no threat to other individuals or autoimmunity, in which the 

immune system produces antibodies to self components leading to destruction of the organ or 

tissue involved. 

There is very little direct information on which to assess the immunotoxic potential of 

diflubenzuron. The only studies specifically related to the effects of diflubenzuron on immune 

function are skin sensitization studies (Section 3.1.11). While the study by Blaszcak (1997e) 

indicates that diflubenzuron is not a skin sensitizer, this provides no information useful for 

directly assessing the potential for diflubenzuron to disrupt immune function. 

Nonetheless, the toxicity of diflubenzuron has been examined in numerous acute, subchronic, 

and chronic bioassays. Although many of these studies did not focus on the immune system, 

changes in the immune system (which could potentially be manifest as increased susceptibility to 

infection compared to controls) were not observed in any of the available long-term animal 

studies (Appendix 1). Typical subchronic or chronic animal bioassays conduct morphological 

assessments of the major lymphoid tissues, including bone marrow, major lymph nodes, spleen 

and thymus (thymus weight is usually measured at autopsy as well), and blood leukocyte counts. 

These assessments can detect signs of inflammation or injury indicative of a direct toxic effect of 

the chemical on the lymphoid tissue. Changes in cellularity of lymphoid tissue and blood, 

indicative of a possible immune system stimulation or suppression, can also be detected (Durkin 

and Diamond 2002). None of these effects have been noted in any of the longer term toxicity 

studies on diflubenzuron (Appendix 1). 

3.1.8. Effects on Endocrine Function 

The endocrine system participates in the control of metabolism and body composition, growth 

and development, reproduction, and many of the numerous physiological adjustments needed to 

maintain constancy of the internal environment (homeostasis). The endocrine system consists of 

endocrine glands, hormones, and hormone receptors. Endocrine glands are specialized tissues 

that produce and export (secrete) hormones to the bloodstream and other tissues. The major 

endocrine glands in the body include the adrenal, hypothalamus, pancreas, parathyroid, pituitary, 

3-6
 
thyroid, ovary, and testis. Hormones are also produced in the gastrointestinal tract, kidney, liver, 

and placenta. Hormones are chemicals produced in endocrine glands that bind to hormone 

receptors in target tissues. Binding of a hormone to its receptor results in a process known as 

postreceptor activation which gives rise to a hormone response in the target tissue, usually an 

adjustment in metabolism or growth of the target tissue. Examples include the release of the 

hormone testosterone from the male testis, or estrogen from the female ovary, which act on 

receptors in various tissues to stimulate growth of sexual organs and development of male and 

female sexual characteristics. The target of a hormone can also be an endocrine gland, in which 

case, receptor binding may stimulate or inhibit hormone production and secretion. Adverse 

effects on the endocrine system can result in abnormalities in growth and development, 

reproduction, body composition, homeostasis (the ability to tolerate various types of stress), and 

behavior. 

There is no indication that diflubenzuron causes endocrine disruption in experimental mammals. 

Standard subchronic, chronic and reproductive toxicity studies provide no basis for asserting that 

any signs of overt toxicity are related to changes in endocrine function. As discussed further in 

Section 4, however, a few studies do indicate a potential endocrine effects in sheep (Section 

4.1.2.1), birds (Section 4.1.2.2) and terrestrial insects (Section 4.1.2.3) but the strength of the 

association is limited. 

3.1.9. Reproductive and Teratogenic Effects 

Diflubenzuron has been tested for its ability to cause birth defects (i.e., teratogenicity) as well as 

its ability to cause reproductive and developmental impairment. Teratogenicity studies typically 

entail gavage administration to pregnant rats or rabbits on specific days of gestation. Two such 

studies (each of which is detailed in Appendix 1) were conducted on diflubenzuron: one in rats 

(Kavanagh 1988a) and one in rabbits (Kavanagh 1988a). As discussed by U.S. EPA/OPP 

(1997a), both of these were screening studies conducted at one very high dose, 1000 mg/kg bw. 

Since no signs of maternal or fetal toxicity were observed, no additional testing was required. 

Another type of reproduction study involves exposing more than one generation of the test 

animal to the compound. One such study has been conducted on diflubenzuron (Brooker 1995). 

As detailed in Appendix 1, this study involved dietary exposures at concentrations of 0, 500, 

5000, or 50,000 ppm over two generations in rats. No effects on reproductive performance were 

noted even though effects were seen on body weight (F0 only) and increases were noted in 

methemoglobin and spleen weight – i.e., effects that may be attributable to diflubenzuron. 

3-7
 
3.1.10 Carcinogenicity and Mutagenicity 

There are no epidemiology studies or case reports that demonstrate or suggest that exposure to 

diflubenzuron leads to cancer in humans. 

The carcinogenicity of diflubenzuron has been tested in rats and mice and these studies are 

detailed in Appendix 1. No carcinogenic effects were observed in rats exposed to diflubenzuron 

in a 2-year feeding study (Keet et al. 1984a). Neither treated nor control rats had cancers of any 

type, although pathological changes were observed in the spleen of both male and female rats. In 

mice, no carcinogenic effects or changes in spleen pathology were observed in males or females 

in a 2-year feeding study (Colley et al. 1984). 

In addition to its lack of carcinogenic activity in in vivo bioassays, several bioassays of 

diflubenzuron for mutagenicity or other damage to DNA have failed to detect adverse effects. A 

lack of mutagenic activity has been reported in a dominant lethal study in mice (Arnold 1974), 

cell transformation assays using BALB/3T3 cells (Brusick and Weir 1977a), the induction of 

unscheduled DNA synthesis (Brusick and Weir 1977b), transplacental transformation assays 

using hamster cells (Quarles et al. 1980), and Ames assays using various strains of Salmonella 

typhimurium with and without metabolic activation (Brusick and Weir 1977c). Diflubenzuron 

did induce cell transformations in BALB/c 3T3 cells in the absence of metabolic activation; 

however, the effect was not observed with metabolic activation (Perocco and others 1993). 

Diflubenzuron has been shown to inhibit the uptake of uridine, adenosine, and cytidine in 

cultured melanoma cells (Mayer et al. 1984) and inhibit the in vivo growth of melanomas in mice 

(Jenkins et al. 1986). Since the inhibition was enhanced by mixed function oxidase induction 

with 3-methylcholanthrene or beta-napthaflavone, aromatic hydroxylation was suggested as a 

requisite to tumor inhibition. 

Both the U.S. EPA/OPP (1996a) and the WHO (1996, 2001) have concluded that diflubenzuron 

is not a carcinogen. This is detailed further in Section 3.3.2.3. However, the potential 

carcinogenicity of 4-chloroaniline, an environmental metabolite of diflubenzuron, is of concern 

and this is discussed further in Section 3.1.15 (Impurities and Metabolites) and in the doseresponse assessment (Section 3.3.3.3). 

3.1.11. Irritation and Sensitization (Effects on the Skin and Eyes) 

As summarized in Appendix 1, diflubenzuron and formulations of diflubenzuron do not appear 

to be skin irritants (Blaszcak 1997d; ) or sensitizers (Blaszcak 1997e). When instilled directly 

into the eye, however, diflubenzuron does cause slight to moderate conjunctival irritation 

(Blaszcak 1997c). 

3.1.12. Systemic Toxic Effects from Dermal Exposure 

As noted in Section 3.1.3.2, diflubenzuron can be absorbed from the skin and many of the 

exposure scenarios considered in this risk assessment involve dermal contact (Section 3.2). The 

available toxicity studies clearly indicate that diflubenzuron can be absorbed in sufficient 

3-8
 
amounts to cause hematologic effects – e.g., methemoglobinemia and sulfhemoglobinemia 

(Goldenthal 1996). Nonetheless, these effects occur only at higher doses after dermal 

administration (i.e., 1000 mg/kg/day) than after oral administration (i.e., about 100 to 250 

mg/kg/day). As with oral toxicity, severe signs of dermal toxicity are not observed even at doses 

that will induce methemoglobinemia and sulfhemoglobinemia (Blaszcak 1997b; Goldenthal 

1996). This is an important relationship that impacts that characterization of risk, as detailed 

further in Section 3.4. 

3.1.13. Inhalation Exposure 

As with oral and dermal exposure, inhalation exposures appear to primarily effect the blood, 

causing increases in methemoglobin and sulfhemoglobin (Eyal 1999; Hoffman 1997; Berczy et 

al. 1975; Newton 1999). The threshold for these effects appears to be lower in nose only 

3 3 exposures – i.e., an NOEC of 30 mg/m with an effect level of 100 mg/m in the study by Eyal

(1999) – compared to whole body exposures – i.e., an NOEC of 500 mg/m . 3 It is unclear why 

this would be the case. In any event, as discussed further in Section 3.2, inhalation is not likely 

to be a significant route of exposure because of the low vapor pressure of diflubenzuron (Table 

2-1) and ambient air will contain concentrations of diflubenzuron that are far below the NOEC 

values for nose-only exposure. 

3.1.14. Inerts and Adjuvants 

As noted in Section 2.2, Dimilin 4L contains petroleum oil [CAS No. 64742-46-7] and Dimilin 

25W contains kaolin clay [CAS No. 1332-58-7] (C&P Press 2004). Kaolin clay is classified as a 

List 4a inert by the U.S. EPA (2004). This classification indicates that the product is considered 

as “Minimal risk inert ingredient”. Petroleum oil with the CAS No. 64742-46-7 designation is 

classified as a List 2 inert which indicates a “Potentially Toxic Inert Ingredients/High Priority for 

Testing inerts”. Details of these classifications may be found at: 

http://www.epa.gov/opprd001/inerts/lists.html. The toxicology of petroleum oil has been 

reviewed in some detail by ATSDR (2003). At sufficiently high doses, some petroleum oils can 

cause gastrointestinal, central nervous system (CNS), and renal effects. Petroleum oils however 

are highly variable and it is difficult to assess the potential contribution of the petroleum oil in 

Dimilin 4L to the overall toxicity of the formulation. No information on the toxicity of Dimilin 

4L is included in the MSDS for this formulation (C&P Press 2004) or in the U.S. EPA RED 

(U.S. EPA/OPP 1997a) and no information on the toxicity of Dimilin 4L was encountered in the 

search of the U.S. EPA CBI files. The toxicity of Dimilin 2L (Blaszcak 1997a summarized in 

Appendix 1) appears to be comparable to that of Dimilin 25W (Koopman, 1977) as well as 

technical grade diflubenzuron (WHO 1996). 

The identity of all inerts in both diflubenzuron formulations has been disclosed to the U.S. EPA 

(i.e., Drozdick 1998b,d; Vanstone 1998a,b,c) and this information has been reviewed as part of 

this risk assessment. This information, however, is protected under FIFRA (Section 10). Other 

than to state that no apparently hazardous materials have been identified, which is consistent with 

the MSDS for both Dimilin 4L and Dimilin 25W (C&P Press 2004), the information on the 

inerts in these formulations cannot be detailed. 

3-9
 
3.1.15. Impurities and Metabolites 

As with inerts, the impurities in formulations of diflubenzuron have been identified and disclosed 

to U.S. EPA (Drozdick 1998a,c,e; Van Kampen and Thus 1996; Vanstone 1998a,b,c; White 

1998) and this information has been reviewed as part of this risk assessment. Again, this 

information is protected under FIFRA (Section 10) and cannot be disclosed in this risk 

assessment. Notwithstanding this limitation, the impurities that may be in diflubenzuron or 

formulations of diflubenzuron add relatively little uncertainty to this risk assessment. All 

toxicity studies summarized in Appendix 1 involved either technical grade diflubenzuron – i.e., 

diflubenzuron with any impurities – or the formulations which also contain the impurities. Thus, 

the available toxicity data should encompass the potential toxic effects of the impurities. 

In terms of metabolites, the toxicity of most in vivo metabolites, as defined in Section 3.1.3.3, 

should also be encompassed by the available in vivo toxicity studies because these metabolites 

will be formed during the course of a standard in vivo toxicity study. This argument, however, 

does not hold for 4-chloroaniline for two reasons. First, as noted in Section 3.1.3.3, 

4-chloroaniline does not appear to be metabolite in rodents, the species on which most toxicity 

studies have been conducted. Secondly, 4-chloroaniline is an environmental metabolite and is 

classified as a Group B2 carcinogen – i.e., indicating a probable human carcinogen following the 

classification of the U.S. EPA/OPP (1997a, 2000a) or a possible human carcinogen following the 

classification of the International Agency for Research on Cancer (IARC 1997). The 

carcinogenic activity of 4-chloroaniline has also been noted by WHO (2003). Consequently, 

potential exposures to 4-chloroaniline are quantitatively considered in the exposure assessment 

(Section 3.2), dose-response assessment (Section 3.3), and risk characterization (Section 3.4), 

3.1.16. Toxicologic Interactions 

There is no information on the interactions of diflubenzuron with other agents. Deleschuse et al. 

(1998) have investigated the cytotoxicity and induction of cytochromes P450 1A1/2 by 

insecticides in hepatic and epidermal cells. Diflubenzuron was one of the six pesticides studied 

and one of two that did not exert a cytotoxic effect in hepatocytes. In addition, de Sousa et al. 

(1997) noted a strong, dose-dependent, significant (p<0.001) induction of ethoxyresorufin 

O-deethylase (EROD) activity and or CYP1A1 mRNAs (5- to 7-fold greater than controls in 

human hepatocytes and approximately 7-fold greater than controls in rat hepatocytes). Any effect 

on hepatocytes and/or cytochrome P450 could impact how an organism would metabolize (either 

to toxicity or detoxify) a very large number of other compounds. The net effect of such 

interactions could be to enhance or inhibit toxicity and a more specific assessment would require 

data on specific combinations of other chemicals with diflubenzuron. 

3-10
 
3.2. EXPOSURE ASSESSMENT 

3.2.1. Overview. 

Exposure assessments are conducted for both diflubenzuron and 4-chloroaniline. For 

diflubenzuron, a standard set of exposure scenarios are presented for both workers and members 

of the general public. As discussed in the hazard identification, concern for 4-chloroaniline 

arises because it is an environmental metabolite of diflubenzuron and is classified as a 

carcinogen. Thus, 4-chloroaniline is not a concern in worker exposure assessments because 

4-chloroaniline will not be present at the time that diflubenzuron is applied. Nor is 

4-chloroaniline a concern in some acute exposure scenarios for the general public such as direct 

spray during the application of diflubenzuron. Consequently, only a subset of the standard 

exposure scenarios – those associated with contaminated vegetation and contaminated water – 

are presented for 4-chloroaniline but these do include all standard chronic exposure scenarios, 

which are of greatest concern because of the potential carcinogenicity of 4-chloroaniline. 

All exposure assessments are based on the maximum single application rate for diflubenzuron of 

0.0625 lb/acre. This is also the maximum application rate for a single season. Assuming that 

diflubenzuron is applied in a single application at the maximum rate leads to the highest 

estimates of peak as well as longer term exposures. The consequences of using lower application 

rates are discussed in the risk characterization. 

For workers applying diflubenzuron, three types of application methods are considered: directed 

ground spray, broadcast ground spray, and aerial spray. Central estimates of exposure for 

workers are approximately 0.0009 mg/kg/day for aerial workers, 0.0008 mg/kg/day for backpack 

workers and about 0.001 mg/kg/day for broadcast ground spray workers. Upper range of 

exposures are approximately 0.009 mg/kg/day for broadcast ground spray workers and 0.005 

mg/kg/day for backpack and aerial workers. All of the accidental exposure scenarios for workers 

involve dermal exposures and most of these accidental exposures lead to estimates of dose that 

are either in the range of or substantially below the general exposure estimates for workers. The 

one exception involves wearing contaminated gloves for one-hour where the upper range of 

exposure is about 0.4 mg/kg/day. 

For the general public, the range of acute exposures is from approximately 0.0000005 mg/kg 

associated with the lower range for the consumption of contaminated water from a stream by a 

child to 1.5 mg/kg associated with the upper range for consumption of contaminated fish by 

subsistence populations – individuals who consume free-caught fish as a major proportion of 

their diet. Relatively high dose estimates are also associated with the consumption of 

contaminated water after an accidental spill (about 0.13 mg/kg at the upper range of exposure) 

and for the consumption of fish by members of the general public (0.3 mg/kg). Other acute 

exposures are lower by about an order of magnitude or greater. For chronic or longer term 

exposures, the modeled exposures are much lower than for acute exposures, ranging from 

approximately 0.00000002 mg/kg/day (2 in 10 millionths of a mg/kg/day) associated with the 

lower range for the consumption of contaminated water to approximately 0.002 mg/kg/day 

associated with the upper range for consumption of contaminated fruit. 

3-11
 
 

Exposures to 4-chloroaniline from contaminated vegetation are likely to be below corresponding 

exposures to diflubenzuron by a factor of about 0.02. This follows from the expected rapid 

dissipation of 4-chloroaniline that is derived from diflubenzuron which has been deposited on 

vegetation. Estimated concentrations of 4-chloroaniline in water, however, are likely to equal or 

exceed anticipated concentrations of diflubenzuron under some circumstances. The peak 

exposures to 4-chloroaniline will occur at different times (later after the application of 

diflubenzuron) and under different conditions of precipitation than those of diflubenzuron. 

These differences are due to the relatively slow rate in the formation of 4-chloroaniline from 

diflubenzuron in soil. 

3.2.2. Workers. 

The Forest Service uses a standard set of exposure assessments in all risk assessment documents. 

All of the exposure assessments for workers as well as members of the general public are 

detailed in the worksheets on diflubenzuron that accompany this risk assessment (Supplement 1) 

and documentation for these worksheets is given in SERA (2003). A copy of this documentation 

is available at www.sera-inc.com. This section on workers and the following section on the 

general public provide plain verbal descriptions of the worksheets and discuss diflubenzuron 

specific data that are used in the worksheets. 

A summary of the exposure assessments for workers is presented in Worksheet E02 of the 

worksheets for diflubenzuron that accompany this risk assessment. Two types of exposure 

assessments are considered: general and accidental/incidental. The term general exposure 

assessment is used to designate those exposures that involve estimates of absorbed dose based on 

the handling of a specified amount of a chemical during specific types of applications. The 

accidental/incidental exposure scenarios involve specific types of events that could occur during 

any type of application. The exposure assessments developed in this section as well as other 

similar assessments for the general public (Section 3.2.3) are based on the maximum single and 

maximum annual application rate of 0.0624 lb/acre (Section 2). The consequences of using 

lower application rates are discussed further in the risk characterization (Section 3.4). 

3.2.2.1. General Exposures – As described in SERA (2001), worker exposure rates are 

expressed in units of mg of absorbed dose per kilogram of body weight per pound of chemical 

handled. Based on analyses of several different pesticides using a variety of application methods, 

default exposure rates are estimated for three different types of applications: directed foliar 

(backpack), boom spray (hydraulic ground spray), and aerial. 

The specific assumptions used for each application method are detailed in Worksheets C01a 

(directed foliar), C01b (broadcast foliar), and C01c (aerial). In the worksheets, the central 

estimate of the amount handled per day is calculated as the product of the central estimates of the 

acres treated per day and the application rate. 

As described in SERA (2001), worker exposure rates are expressed in units of mg of absorbed 

dose per kilogram of body weight per pound of chemical handled. These exposure rates are 

3-12
 
based on worker exposure studies on nine different pesticides with molecular weights ranging 

from 221 to 416 and log Kow values ranging from -0.75 to 6.50. The estimated exposure rates are 

based on estimated absorbed doses in workers as well as the amounts of the chemical handled by 

the workers. As summarized in Table 2-1 of this risk assessment, the molecular weight of 

diflubenzuron is 320 and the log Kow is about 3.9. These values are within the range of the 

pesticides used in SERA (2001). As described in SERA (2001), the ranges of estimated 

occupational exposure rates vary substantially among individuals and groups, (i.e., by a factor of 

50 for backpack applicators and a factor of 100 for mechanical ground sprayers). It seems that 

much of the variability can be attributed to the hygienic measures taken by individual workers 

(i.e., how careful the workers are to avoid unnecessary exposure); however, pharmacokinetic 

differences among individuals (i.e., how fast individuals absorb and excrete the compound) also 

may be important. 

The number of acres treated per hour is taken from previous USDA risk assessments (USDA/FS 

1989a,b,c). The number of hours worked per day is expressed as a range, the lower end of which 

is 6 hours based on an 8-hour work day with 1 hour at each end of the work day spent in 

activities that do not involve exposure to the compound. The upper end of the range, 8 hours per 

day, is based on an extended (10-hour) work day, allowing for 1 hour at each end of the work day 

to be spent in activities that do not involve exposure to the chemical. 

It is recognized that the use of 6 hours as the lower range of time spent per day applying 

herbicides is not a true lower limit. It is conceivable and perhaps common for workers to spend 

much less time in the actual application of a herbicide if they are engaged in other 

activities. Thus, using 6 hours may overestimate exposure. In the absence of any published or 

otherwise documented work practice statistics to support the use of a lower limit, this approach is 

used as a protective assumption. 

The range of acres treated per hour and hours worked per day is used to calculate a range for the 

number of acres treated per day. For this calculation as well as others in this section involving 

the multiplication of ranges, the lower end of the resulting range is the product of the lower end 

of one range and the lower end of the other range. Similarly, the upper end of the resulting range 

is the product of the upper end of one range and the upper end of the other range. This approach 

is taken to encompass as broadly as possible the range of potential exposures. 

The central estimate of the acres treated per day is taken as the arithmetic average of the range. 

Because of the relatively narrow limits of the ranges for backpack and boom spray workers, the 

use of the arithmetic mean rather than some other measure of central tendency, like the geometric 

mean, has no marked effect on the risk assessment. 

3-13
 
3.2.2.2. Accidental Exposures – Typical occupational exposures may involve multiple routes of 

exposure (i.e., oral, dermal, and inhalation); nonetheless, dermal exposure is generally the 

predominant route for herbicide applicators (Ecobichon 1998; van Hemmen 1992). Typical 

multi-route exposures are encompassed by the methods used in Section 3.2.2.1 on general 

exposures. Accidental exposures, on the other hand, are most likely to involve splashing a 

solution of herbicides into the eyes or various dermal exposure scenarios. 

Diflubenzuron can cause slight to moderate eye irritation (Section 3.1.11). The available 

literature does not include quantitative methods for characterizing exposure or responses 

associated with splashing a solution of a chemical into the eyes; furthermore, there appear to be 

no reasonable approaches to modeling this type of exposure scenario quantitatively. 

Consequently, accidental exposure scenarios of this type are considered qualitatively in the risk 

characterization (section 3.4). 

As detailed in Section 3.1.3, there are various methods for estimating absorbed doses associated 

with accidental dermal exposure (U.S. EPA 1992a, SERA 2001). Two general types of exposure 

are modeled: those involving direct contact with a solution of the herbicide and those associated 

with accidental spills of the herbicide onto the surface of the skin. Any number of specific 

exposure scenarios could be developed for direct contact or accidental spills by varying the 

amount or concentration of the chemical on or in contact with the surface of the skin and by 

varying the surface area of the skin that is contaminated. 

For this risk assessment, two exposure scenarios are developed for each of the two types of 

dermal exposure, and the estimated absorbed dose for each scenario is expressed in units of mg 

chemical/kg body weight. Both sets of exposure scenarios are summarize in Worksheet E01, 

which references other worksheets in which the specific calculations are detailed. 

Exposure scenarios involving direct contact with solutions of the chemical are characterized by 

immersion of the hands for 1 minute or wearing contaminated gloves for 1 hour. Generally, it is 

not reasonable to assume or postulate that the hands or any other part of a worker will be 

immersed in a solution of a herbicide for any period of time. On the other hand, contamination 

of gloves or other clothing is quite plausible. For these exposure scenarios, the key element is 

the assumption that wearing gloves grossly contaminated with a chemical solution is equivalent 

to immersing the hands in a solution. In either case, the concentration of the chemical in solution 

that is in contact with the surface of the skin and the resulting dermal absorption rate are 

essentially constant. 

For both scenarios (the hand immersion and the contaminated glove), the assumption of 

zero-order absorption kinetics is appropriate. Following the general recommendations of U.S. 

EPA/ORD (1992), Fick's first law is used to estimate dermal exposure. As discussed in Section 

3.1.3, an experimental dermal permeability coefficient (Kp) for diflubenzuron is not available. 

Thus, the Kp for diflubenzuron is estimated using the algorithm from U.S. EPA (1992a), which 

is detailed in Worksheet A10. 

3-14
 
Exposure scenarios involving chemical spills onto the skin are characterized by a spill on to the 

lower legs as well as a spill on to the hands. In these scenarios, it is assumed that a solution of 

the chemical is spilled on to a given surface area of skin and that a certain amount of the 

chemical adheres to the skin. The absorbed dose is then calculated as the product of the amount 

of the chemical on the surface of the skin (i.e., the amount of liquid per unit surface area 

multiplied by the surface area of the skin over which the spill occurs and the concentration of the 

chemical in the liquid) the first-order absorption rate, and the duration of exposure. 

For both scenarios, it is assumed that the contaminated skin is effectively cleaned after 1 hour. 

As with the exposure assessments based on Fick's first law, this product (mg of absorbed dose) is 

divided by body weight (kg) to yield an estimated dose in units of mg chemical/kg body weight. 

3.2.3. General Public. 

3.2.3.1. General Considerations – Although some applications of diflubenzuron may be made 

in relatively remote areas involving limited exposure to the general public, both aerial and 

ground applications may be made in residential areas. In residential applications, members of the 

general public are more likely to be exposed to diflubenzuron. Any number of exposure 

scenarios can be constructed for the general public, depending on various assumptions regarding 

application rates, dispersion, canopy interception, and human activity. Several scenarios are 

developed for this risk assessment which should tend to over-estimate exposures in general. 

The two types of exposure scenarios developed for the general public include acute exposure and 

longer-term or chronic exposure. All of the acute exposure scenarios are primarily accidental. 

They assume that an individual is exposed to the compound either during or shortly after its 

application. Specific scenarios are developed for direct spray, dermal contact with contaminated 

vegetation, as well as the consumption of contaminated fruit, water, and fish. Most of these 

scenarios should be regarded as extreme, some to the point of limited plausibility. The 

longer-term or chronic exposure scenarios parallel the acute exposure scenarios for the 

consumption of contaminated fruit, water, and fish but are based on estimated levels of exposure 

for longer periods after application. 

The exposure scenarios developed for the general public are summarized in Worksheet E03. As 

with the worker exposure scenarios, details of the assumptions and calculations involved in these 

exposure assessments are given in the worksheets that accompany this risk assessment 

(Worksheets D01a to D09b). The remainder of this section focuses on a qualitative description 

of the rationale for and quality of the data supporting each of the assessments. 

3.2.3.2. Direct Spray – Direct sprays involving ground applications are modeled in a manner 

similar to accidental spills for workers (Section 3.2.2.2). In other words, it is assumed that the 

individual is sprayed with a solution containing the compound and that an amount of the 

compound remains on the skin and is absorbed by first-order kinetics. For these exposure 

scenarios, it is assumed that during a ground application, a naked child is sprayed directly with 

diflubenzuron. These scenarios also assume that the child is completely covered with 

3-15
 
diflubenzuron (that is, 100% of the surface area of the body is exposed and contaminated). 

These exposure scenarios are likely to represent upper limits of plausible exposure. An 

additional set of scenarios are included involving a young woman who is accidentally sprayed 

over the feet and legs. For each of these scenarios, some assumptions are made regarding the 

surface area of the skin and body weight, as detailed in the Series B Worksheets. 

3.2.3.3. Dermal Exposure from Contaminated Vegetation – In this exposure scenario, it is 

assumed that the herbicide is sprayed at a given application rate and that an individual comes in 

contact with sprayed vegetation or other contaminated surfaces at some period after the spray 

operation. For these exposure scenarios, some estimates of dislodgeable residue and the rate of 

transfer from the contaminated vegetation to the surface of the skin must be available. No such 

data are available on dermal transfer rates for diflubenzuron and the estimation methods of 

Durkin et al. (1995) are used as defined in Worksheet D02. The exposure scenario assumes a 

contact period of one hour and assumes that the chemical is not effectively removed by washing 

for 24 hours. Other estimates used in this exposure scenario involve estimates of body weight, 

skin surface area, and first-order dermal absorption rates, as discussed in the previous section. 

3.2.3.4. Contaminated Water – Water can be contaminated from runoff, as a result of leaching 

from contaminated soil, from a direct spill, or from unintentional contamination from aerial 

applications. For this risk assessment, three exposure scenarios are considered for the acute 

consumption of contaminated water: an accidental spill into a small pond (0.25 acres in surface 

area and 1 meter deep), accidental direct spray of or incidental drift into a pond and stream, and 

the contamination of a small stream and pond by runoff or percolation. In addition, longer-term 

estimates of concentrations in water are based on a combination of modeling and monitoring 

data. Each of these scenarios are considered in the following subsections. 

3.2.3.4.1. Accidental Spill – The accidental spill scenario assumes that a young child 

consumes contaminated water shortly after an accidental spill into a small pond. The specifics of 

this scenarios are given in Worksheet D05. Because this scenario is based on the assumption that 

exposure occurs shortly after the spill, no dissipation or degradation of diflubenzuron is 

considered. This scenario is dominated by arbitrary variability and the specific assumptions used 

will generally overestimate exposure. The actual concentrations in the water would depend 

heavily on the amount of compound spilled, the size of the water body into which it is spilled, the 

time at which water consumption occurs relative to the time of the spill, and the amount of 

contaminated water that is consumed. Based on the spill scenario used in this risk assessment, 

the concentration of diflubenzuron in a small pond is estimated to range from about 0.014 mg/L 

to 1.1 mg/L with a central estimate of about 0.2 mg/L (Worksheet D05). This is and is intended 

to be an extreme accidental exposure scenario. The purpose of this scenario is simply to suggest 

the intensity of measures that would need to be taken in the event of a relatively large spill of 

diflubenzuron into a relatively small body of water. 

3-16
 
3.2.3.4.2. Accidental Direct Spray/drift for a Pond or Stream – These scenarios are less 

severe but more plausible than the accidental spill scenario described above. The U.S. EPA 

typically uses a two meter deep pond to develop exposure assessments (SERA 2004). If such a 

pond is directly sprayed with diflubenzuron at the nominal application rate of 0.0624 lb/acre, the 

peak concentration in the pond would be about 0.0035 mg/L (3.5 μg/L or 3.5 ppb) (Worksheet 

D10a). This concentration is a factor of about 300 below the peak concentration of 1.1 mg/L 

after the accidental spill. Because the USDA will not directly spray open bodies of water but will 

use buffers of 100 to 500 feet (Section 2.3), the concentration of 0.0035 mg/L from direct spray 

would be an accidental exposure. Using the 100 to 500 foot buffers, drift of diflubenzuron from 

aerial applications would result in water concentrations between about 7.7×10-06 mg/L (about 

0.008 ppb or 8 ppt – parts per trillion) to about 6.8×10-05 mg/L (0.07 ppb or 70 ppt) (Worksheet 

10a). 

Similar calculations can be made for the direct spray of a stream and the resulting water 

concentrations will be dependant on the surface area of the stream that is sprayed and the rate of 

water flow in the stream. The stream modeled using GLEAMS (see below) is about 6 feet wide 

and it is assumed that the pesticide is applied along a 1038 foot length of the stream with a flow 

rate of 710,000 L/day. The length of 1038 feet is based on the length of a side of a square 10 ha 

treatment plot. At an application rate of 0.0624 lb/acre, accidental direct spray onto the surface 

of the stream would deposit about 4047 mg and this would result in a downstream concentration 

of about 0.0057 mg/L. Using a buffer of 100 feet, the drift would be a fraction of 0.0195 of the 

application rate (Worksheet B24) and the concentration in the stream would be about 0.00011 

mg/L. Details of these and additional calculations for concentrations in stream water are given in 

Worksheet 10b. 

3.2.3.4.3. Gleams Modeling – For compounds such as diflubenzuron, which may be 

applied to an entire watershed, drift and even direct spray are not the only and may not be the 

greatest source of contamination of surface water. Water contamination may also occur from soil 

runoff or percolation and, depending on local conditions, can lead to substantial contamination of 

ponds or streams. Estimates of these concentrations can be based both on modeling and 

monitoring data. 

Modeling of concentrations in stream water conducted for this risk assessment are based on 

GLEAMS (Groundwater Loading Effects of Agricultural Management Systems) modeling. 

GLEAMS is a root zone model that can be used to examine the fate of chemicals in various types 

of soils under different meteorological and hydrogeological conditions (Knisel and Davis 2000). 

As with many environmental fate and transport models, the input and output files for GLEAMS 

can be complex. The general application of the GLEAMS model and the use of the output from 

this model to estimate concentrations in ambient water are detailed in SERA (2004). 

For the current risk assessment, the application site was assumed to consist of a 10 hectare square 

area that drained directly into a small pond or stream. The chemical specific values as well as 

the details of the pond and stream scenarios used in the GLEAMS modeling are summarized in 

3-17
 
Table 3-1. The GLEAMS modeling yielded estimates of runoff, sediment and percolation that 

were used to calculate concentrations in the stream adjacent to a treated plot, as detailed in 

Section 6.4 of SERA (2004). The results of the GLEAMS modeling for the small stream are 

summarized in Table 3-2 and the corresponding values for the small pond are summarized in 

Table 3-3. These estimates are expressed as both average and maximum concentrations in water. 

The top section of each table gives the contamination rates (WCR) – i.e., the concentration of 

the compound in water in units of ppb (μg/L) normalized for an application rate of 1 lb/acre. The 

bottom section of each table gives the estimated maximum and average concentrations adjusted 

for the application rate of 0.0624 lb/acre (Section 2.3). 

As indicated in Table 3-2, no stream contamination is estimated in very arid regions – i.e., annual 

rainfall of 10 inches of less. For regions with annual rainfall rates of 15 inches or more, the 

modeled peak concentrations in streams range from less than 0.01 μg/L (sandy soil) to about 

15 μg/L (clay soil at an annual rainfall rate of 250 inches per year). While not detailed in 

Table 3-2, the losses from clay are associated almost exclusively with sediment loss (about 94%), 

with the remaining amount due to runoff. No water contamination due to percolation is modeled. 

This is consistent with a large body of literature on diflubenzuron indicating that downward 

movement in the soil horizon is extremely limited (e.g., Sundaram and Nott 1989; WHO 1996). 

Even in sandy soils, where very little water contamination is anticipated, percolation accounts for 

only about 3% of the total loss at an annual rainfall rate of 250 inches. 

Modeled concentrations in a small pond (Table 3-3) are lower than those modeled in the stream. 

As discussed further below, this is consistent with similar modeling conducted by Schocken et al. 

(2001) using PRZM/EXAMS. As with the stream modeling, no surface water contamination is 

expected in very arid regions. For regions with annual rainfall rates of 15 inches or more, the 

modeled peak concentrations in ponds range from less than 0.004 μg/L (sandy soil) to about 

3 μg/L (clay soil at an annual rainfall rate of 250 inches per year). 

The GLEAMS scenarios do not specifically consider the effects of accidental direct spray. As 

discussed above and detailed in Worksheet B06b, direct spray of a standard pond could result in 

peak concentrations of about 3.5 μg/L, comparable to the peak levels modeled in ponds adjacent 

to fields with clay soil. 

As discussed in Section 3.1.15, this risk assessment is also concerned with concentrations of 

4-chloroaniline that could occur in water after the application of diflubenzuron. This process 

was also modeled using GLEAMS as described above for diflubenzuron. As illustrated in 

Figure 3-1, diflubenzuron does not degrade directly to 4-chloroaniline. It is first degraded to 

4-chlorophenylurea which is in turn degraded to 4-chloroaniline. For the GLEAMS modeling, 

however, the degradation was modeled as a one-step process, disregarding the formation of 

4-chlorophenylurea. This is a conservative approach in that the formation of 4-chlorophenylurea 

will attenuate the formation of 4-chloroaniline. As discussed further in the risk characterization 

(Section 3.4), this conservative approach has no impact on the risk assessment. 

3-18
 
The chemical specific properties for 4-chloroaniline used in the GLEAMS modeling are given in 

Table 3-4 and the results for the stream and pond are summarized in Tables 3-5 and 3-6, 

respectively. The pattern seen with 4-chloroaniline is somewhat more complex than that seen 

with the parent compound. For example, the average and peak concentrations of 4-chloroaniline 

in streams is not directly related to rainfall rates (Table 3-5). The highest peak concentration, 

about 2 μg/L, occurs at a rainfall rate of 100 inches per year. At a rainfall rate of 250 inches per 

year, the modeled peak concentration is only about 0.36 μg/L. This pattern occurs because the 

formation of 4-chloroaniline is more rapid in soil than in water – i.e., great microbial activity in 

soil. Thus, at higher rainfall rates, diflubenzuron is washed rapidly from soil and lesser amounts 

of 4-chloroaniline are formed. A similar pattern with respect to rainfall rates is seen in the 

modeling results for the pond (Table 3-6). 

The temporal exposures to 4-chloroaniline will also differ from those of diflubenzuron. This is 

illustrated in Figure 3-2 for concentrations of diflubenzuron and 4-chloroaniline in ponds at an 

annual rainfall rate of 150 inches. In clay and loam soils, diflubenzuron concentrations peak 

after the first rainfall and then steadily decline. Concentrations of 4-chloroaniline, however, peak 

after about 30 to 70 days. While diflubenzuron concentrations are much higher from clay than 

loam because of higher runoff from clay, the peak concentrations for 4-chloroaniline are similar 

for both clay (0.42 μg/L) and loam (0.35 μg/L), with the peak concentration in loam soil 

occurring somewhat later than that in clay soil. The greatest difference between diflubenzuron 

and 4-chloroaniline occurs in sand. As discussed above, virtually no diflubenzuron is expected 

to occur in ponds with very sandy soils. This is illustrated in Figure 3-2 for an annual rainfall of 

150 inches, in which the concentration of diflubenzuron in water for sand is estimated at zero 

over the one-year model run. Nonetheless, 4-chloroaniline as a breakdown product from 

diflubenzuron will form and will rapidly leach through sand. Thus, for 4-chloroaniline, the peak 

concentrations in the pond with sandy soil, about 1.4 μg/L, are substantially higher than the peak 

concentrations associated with either clay or loam soils. 

3.2.3.4.4. Other Modeling Efforts – A summary of the GLEAMS modeling discussed 

above as well as modeling of diflubenzuron conducted for other analyses is given in Table 3-7. 

While some of these modeling efforts involved assumptions substantially different from the 

GLEAMS modeling (i.e., application rates, soil types, and rainfall patterns), the results are 

reasonably consistent with the above estimates of concentrations in surface waters based on 

GLEAMS. All of these modeling efforts used PRZM/EXAMS. As discussed in SERA (2004), 

PRZM (Pesticide Root Zone Model) is model used by U.S. EPA that is comparable to GLEAMS. 

PRZM is often linked with EXAMS (Exposure Analysis Modeling System) to estimate 

concentrations of pesticides in water (U.S. EPA/OPPTS 2004). 

In the previous diflubenzuron risk assessment for the gypsy moth program (USDA 1995), 

maximum modeled concentrations at an application rate of 0.0624 lb/acre, identical to the rate 

used in the GLEAMS modeling, maximum concentrations in streams after direct spray were 

estimated at 16 ppb, very close to the estimate of 22 ppb made in the current risk assessment. 

Concentrations of diflubenzuron in streams associated with runoff were in the range of about 2 

3-19
 
 

ppb to 13 ppb. These are very similar to the central and upper range of concentrations in streams 

based on the GLEAM modeling (2 ppb to 16 ppb). For open water, USDA (1995) estimated a 

maximum concentration of 1.22 ppb, which is only somewhat below the maximum of 3 ppb 

based on GLEAMS. 

In the reregistration eligibility decision for diflubenzuron, U.S. EPA (1997a) modeled 

concentrations of diflubenzuron in surface water using Tier 2 computer models. These models 

are not otherwise specified in U.S. EPA (1997a). Typically, Tier 2 modeling by U.S. EPA 

involves PRZM/EXAMS. The U.S. EPA estimates much higher concentrations in water but this 

is largely due to differences in application rates. For example, at an application rate of 0.67 

lb/acre, about a factor of 10 higher than the rate used with GLEAMS (0.0624 lb/acre), the U.S. 

EPA estimates a peak concentration of about 8.1 μg/L. Adjusting for the differences in 

application rate, the EPA estimate would be 0.8 μg/L [8.1 μg/L × 0.0624 lb/acre ÷ 0.67 lb/acre = 

0.754 μg/L], similar to the estimates using GLEAMS with clay soil at rainfall rates of 100 to 150 

inches. While the U.S. EPA (1997a) does not specify rainfall rates or soil types, Tier 2 modeling 

generally involves “worse case” assumptions which, in this case, would be based on high runoff 

soils (i.e., clay) and relatively high rainfall rates. The U.S. EPA (1997a) modeling for “Forestry” 

applications are specified as direct application. U.S. EPA (1997a) does not indicate the nature of 

the forestry application but direct spray of water does not correspond to applications for the 

control the gypsy moth. The concentrations modeled by U.S. EPA (1997a) of about 23 μg/L at 

an application rate of 0.07 lb/acre is consistent with the direct spray of a small stream modeled in 

this risk assessment (i.e., 22μg/L) but substantially higher than the direct spray of a pond (i.e., 

3μg/L). In direct applications to shallow (1.3 to 1.7 m) ponds, Sundarum et al. (1991) monitored 

average day 1 concentrations in ponds of about 4 μg/L at an application rate of 70 g/ha (0.062 

lb/acre), consistent with the peak concentrations in ponds discussed above (Section 3.2.3.4.3). 

Harned and Relyea (1997) modeled diflubenzuron applications to a 10 ha plot after the 

application diflubenzuron at 350 g/ha, about a factor of 5 higher than the application rate used in 

the GLEAMS modeling. As with the EPA, Harned and Relyea (1997) used PRZM/EXAMS but 

combined these models with AgDrift. Harned and Relyea (1997) employed variable rainfall rates 

rather than fixed rates but the individual rainfall events varied from about 2.4 to 7.2 cm or about 

1 to 2.8 inches. Based on their modeling, peak concentrations in the pond were estimated at 

about 1 μg/L. Correcting for the difference in application rates, their estimate of 1 μg/L would 

correspond to 0.2 μg/L in the GLEAMS modeling – i.e., higher by a factor of 5. As indicated in 

Table 3-3, concentrations estimated using GLEAMS at comparable daily rainfall events ranged 

from 0.2 to about 0.8 μg/L. 

Schocken et al. (2001) also used AgDrift with PRZM/EXAMS to model diflubenzuron in 

streams and ponds beneath and adjacent to forests after an application of 0.125 lb/acre, about 

twice the application rate used in the GLEAMS modeling. Modeled estimates indicated that the 

initial concentration immediately after application should not exceed 0.255 μg/L in ponds and 

0.938 μg/L in streams under the canopy. In adjacent areas, modeled estimates indicated that 

concentrations in ponds and streams should not exceed 0.260 μg/L and 0.856 μg/L, respectively. 

3-20
 
The higher concentrations of diflubenzuron in streams compared to ponds is consistent with the 

GLEAMS modeling (Tables 3-2 and 3-3). The stream concentrations modeled by Schocken et 

al. (2001) of 1 μg/L are about a factor of 2 below the central estimates from GLEAMS – i.e., 

about 2 μg/L. This is probably due to the higher stream flow rate used by Schocken et al. (2001) 

– i.e., 58,320,000 L/day compared to 710,000 L/day used in the GLEAMS modeling. The peak 

concentrations in ponds modeled by Schocken et al. (2001), about 0.2 μg/L to 0.3 μg/L are very 

similar to the estimates from GLEAMS at rainfall rates of about 50 inches per year. 

3.2.3.4.5. Monitoring Data – Several monitoring studies (Carr et al. 1991; Nigg and 

Stamper 1987; Van Den Berg 1986) are available that can be used to assess the plausibility of the 

modeling estimates summarized in Table 3-7. The common feature in each of these studies is 

that concentrations in pond and/or stream water are reported and these concentrations can be 

associated with a defined application rate. The study by Van Den Berg (1986) is probably the 

most directly relevant to this risk assessment. In this study, diflubenzuron was applied to a 10­

acre mixed hardwood-conifer forested plot at an application rate of 0.0625 lb/acre. Initial 

concentrations of diflubenzuron in surface water (streams and stream pools) in treatment area 

ranged from 0.127-0.203 ppb and declined to 0.029-0.045 ppb after one day. These 

concentrations are in the range of concentrations modeled using GLEAMS for ponds (central 

range) and streams (lower range). Similar results are reported by Carr et al. (1991) who 

monitored concentrations in streams below 0.5 ppb after the application of diflubenzuron at rates 

of 13 g/ha or 26 g/ha. Adjusted for an application rate of 0.0624 lb/acre (70 g/ha), the 

concentration of 0.5 ppb would correspond to about 2.5 to 5 ppb, very close to the upper range of 

stream concentrations modeled using GLEAMS. The study by Nigg and Stamper (1987) 

involved a very high application rate, 560 g/ha (226 g/ac or 0.5 lb/acre) in a citrus grove. The 

maximum monitored concentration in an adjacent pond was 0.197 ppb. Adjusted to an 

application rate of 0.0624 lb/acre (70 g/ha), this corresponds to a concentration of about 0.02 

ppb, in the lower range of pond concentrations modeled using GLEAMS. 

This discussion of the monitoring data is not intended to imply a validation of the GLEAMS 

modeling or other modeling efforts. Model validation or calibration can only be done on a sitespecific basis. Nonetheless, the monitoring data do suggest that estimates from GLEAMS as 

well as other comparable modeling efforts are at least plausible and may reasonably reflect the 

highly variable concentrations of diflubenzuron that may occur in surface water over a wide 

range of site-specific conditions. 

3.2.3.4.6. Concentrations of Diflubenzuron in Water Used for Risk Assessment – A 

summary of the concentrations of diflubenzuron in water that are used for the current risk 

assessment is given in Table 3-8. The upper range of the expected peak concentration of 

diflubenzuron in surface water will be taken as 16 μg/L for an application rate of 0.0624 lb/acre. 

This is based on the upper range of concentrations estimated in streams from the GLEAMS 

modeling. This estimate is consistent with both the available monitoring data (Section 3.2.3.4.6) 

and other comparable modeling efforts (Section 3.2.3.4.5). This concentration also encompasses 

accidental direct sprays of both a small stream and small pond (Table 3-7). In most instances, 

3-21
 
concentrations in surface water are likely to be much lower. At the lower extreme, an argument 

may be made that concentrations of diflubenzuron are likely to be essentially zero – i.e., 

applications made at sites that are distant from open bodies of water and in areas in which runoff 

or percolation are not likely to occur. For this risk assessment, the lower concentration in 

ambient water will be set at 0.01 μg/L. This is in the lower range of non-zero concentrations 

modeled in streams and ponds in relatively arid regions. The central estimate of the 

concentration of diflubenzuron in surface water will be taken as 0.4 μg/L. This is the geometric 

mean of the range of 0.01 μg/L to 16 μg/L. 

Longer term concentrations of diflubenzuron in surface water will be much lower than peak 

concentrations. At an application rate of 0.0624 lb/acre, the highest longer term concentration 

will be taken as 0.1 μg/L. This is near the maximum longer term concentration given by U.S. 

EPA (1997a) after adjusting for differences in application rate – i.e., 0.74 μg/L ÷ 6 applications 

at 0.06 lb/acre. This longer term maximum concentration is also near the upper range of the 

longer term concentrations modeled using GLEAMS – i.e., 0.06 μg/L in streams at an 

application rate of 0.0624 lb/acre. As with peak concentrations, the lower range of longer term 

concentrations will approach zero. For this risk assessment, the lower range of longer term 

concentrations is taken as 0.001 μg/L, the lowest non-zero value modeled for diflubenzuron in 

ponds. This lower range is somewhat arbitrary but has no impact on the risk assessment. The 

central value for longer term concentrations of diflubenzuron in water will be taken as 0.02 μg/L. 

This is adapted from the longer term concentrations modeled by Harned and Relyea (1997) but 

adjusted for differences in the application rate – i.e., 0.1 μg/L × (70 g/ha ÷ 350 g/ha) = 

0.02 μg/L. This value is similar to the central estimates of longer term concentrations in streams 

modeled using GLEAMS – i.e., 0.01 μg/L in Table 3-7 – but is near the upper range of 

concentrations that would be expected in ponds – i.e., 0.06 μg/L in Table 3-7. 

3.2.3.4.7. Concentrations of 4-Chloroaniline in Water Used for Risk Assessment – A 

summary of the concentrations of 4-chloroaniline in water that are used for the current risk 

assessment is given in Table 3-9. The upper range of the expected peak concentration of 4­

chloroaniline in surface water will be taken as 3 μg/L for an application rate of 0.0624 lb/acre. 

This is based on the upper range of concentrations estimated in streams near application sites 

with sandy soil over a range of annual rainfall rates from about 25 to 250 inches (Table 3-5). 

This concentration is higher than concentrations that might be expected in ponds by about a 

factor of 3 (Table 3-6). As with diflubenzuron, the lower range of concentrations of 4­

chloroaniline in water will approach zero. For this risk assessment, the lower range is taken as 

0.00003 μg/L, the lowest non-zero concentration modeled in ponds (i.e., Table 3-6, peak 

concentration for loam at an annual rainfall rate of 15 inches). The central estimate is taken as 

0.5 μg/L. This is about the concentration modeled in stream with loam soil over a range of 

annual rainfall rates of 100 to 250 inches. 

Longer term concentrations of 4-chloroaniline are taken as 0.05 μg/L with a range of 0.0002 

μg/L to 0.2 μg/L at an application rate of 0.0624 lb/acre. The lower range is based on the lowest 

non-zero concentration modeled in ponds – i.e., loam soil at an annual rainfall rate of 15 inches. 

3-22
 
The upper range is taken as the highest concentration modeled in ponds – i.e., sandy soil at 

annual rainfall rate of about 25 to 100 inches. The central estimate is based on the relatively 

narrow range of concentrations modeled in ponds with loam soil over rainfall rates of 50 to 250 

inches per year – i.e., about 0.04 to 0.06 μg/L in Table 3-6. Much lower concentrations are likely 

to be seen in streams. 

3.2.3.5. Oral Exposure from Contaminated Fish – Many chemicals may be concentrated or 

partitioned from water into the tissues of animals or plants in the water. This process is referred 

to as bioconcentration. Generally, bioconcentration is measured as the ratio of the concentration 

in the organism to the concentration in the water. For example, if the concentration in the 

organism is 5 mg/kg and the concentration in the water is 1 mg/L, the bioconcentration factor 

(BCF) is 5 L/kg [5 mg/kg ÷ 1 mg/L]. As with most absorption processes, bioconcentration 

depends initially on the duration of exposure but eventually reaches steady state. Details 

regarding the relationship of bioconcentration factor to standard pharmacokinetic principles are 

provided in Calabrese and Baldwin (1993). 

Burgess (1989) assayed the bioconcentration of diflubenzuron in Bluegill sunfish, Lepomis 

macrochirus, over a 28 day exposure using C 14-labeled diflubenzuron. In this study, 

concentrations in water, whole fish, fillet (muscle), and viscera were measured at day 0.17 (4 

hours), as well as on days 1, 3, 7, 14, 21, and 28. In fillet, the fish muscle ,the BCF was 120 after 

1 day and 170 after 28 days with a peak of 200 after 7 days. In whole fish, the BCF was 260 

after 1 day and 350 after 28 days with a peak of 360 after 7 days. Similar BCF values have been 

noted for diflubenzuron by Schaefer et al. (1979, 1980). 

For the human health risk assessment of diflubenzuron, the BCF in fillet of 120 after 1 day will 

be used for acute exposures and the maximum BCF in fillet of 200 will be used for longer term 

exposures. This approach is taken under the assumption that humans will consume only the fish 

muscle. In the ecological risk assessment, however, the assumption will be made a predatory 

consumes the entire fish. Thus, for the ecological risk assessment, the whole body BCF values 

will be used, 260 for acute exposures and 360 for longer term exposures. These values are 

entered into Worksheet A02 for diflubenzuron and used in the subsequent worksheets involving 

exposures to contaminated fish. 

Less detailed information is available on the bioconcentration of 4-chloroaniline. Because 

4-chloroaniline is much more water soluble than diflubenzuron and has a much lower octanolwater partition coefficient, very little bioconcentration is expected in fillet or whole fish (WHO 

2003). In a 14-day exposure of carp to two concentrations of 4-chloroaniline, Tsuda et al. (1993) 

noted essentially no bioconcentration – i.e., the concentrations in water were essentially identical 

to those in the fish. Thus, in Worksheet A02 for 4-chloroaniline, values of 1 are used for all BCF 

values – acute and chronic, whole fish and muscle. 

For all scenarios involving the consumption of contaminated fish, concentrations of 

diflubenzuron or 4-chloroaniline in water are identical to the concentrations used in the 

3-23
 
contaminated water scenarios (see Section 3.2.3.4). The acute exposure scenario is based on the 

assumption that an adult angler consumes fish taken from contaminated water shortly after an 

accidental spill of 200 gallons of a field solution into a pond that has an average depth of 1 m and 

2 a surface area of 1000 m or about one-quarter acre. No dissipation or degradation is considered. 

Because of the available and well-documented information and substantial differences in the 

amount of caught fish consumed by the general public and native American subsistence 

populations, separate exposure estimates are made for these two groups (Worksheets D08a and 

D08b). The chronic exposure scenario is constructed in a similar way, as detailed in Worksheets 

D09a and D09b, except that estimates of concentrations in ambient water are based on the 

longer-term estimates summarized in Table 3-8 for diflubenzuron and Table 3-9 for 

4-chloroaniline. 

3.2.3.6. Oral Exposure from Contaminated Vegetation – Although Forest Service 

applications of diflubenzuron will not involve the intentional treatment of food crops, incidental 

exposure to vegetation that may be consumed by members of the general public is plausible 

during broadcast applications. Any number of scenarios could be developed involving either 

accidental spraying of crops or the spraying of edible wild vegetation, like berries. The two 

exposure scenarios developed for this exposure assessment include one scenario for acute 

exposure, as defined in Worksheet D03 and one scenario for longer-term exposure, as defined in 

Worksheet D04. In both scenarios, the concentration of diflubenzuron on contaminated 

vegetation is estimated using the empirical relationships between application rate and 

concentration on vegetation developed by Fletcher et al. (1994) which is in turn based on a 

re-analysis of data from Hoerger and Kenaga (1972). These relationships are defined in 

Worksheet B21. For the acute exposure scenario, the estimated residue level is taken as the 

product of the application rate and the residue rate (Worksheet D03). 

For the longer-term exposure scenario (Worksheet D04), a duration of 90 days is used. The rate 

of decrease in the residues over time is taken from the vegetation half-time of 9.3 days (Table 2­

1). Although the duration of exposure of 90 days is somewhat arbitrary, this duration is intended 

to represent the consumption of contaminated fruit that might be available over one season. 

Longer durations could be used for certain kinds of vegetation but would lower the estimated 

dose (i.e., would reduce the estimate of risk). 

For the longer-term exposure scenarios, the time-weighted average concentration on fruit is 

calculated from the equation for first-order dissipation. Assuming a first-order decrease in 

concentrations in contaminated vegetation, the concentration in the vegetation at time t after 

spray, C , can be calculated based on the initial concentration, C , as: t 0

C = C × e-kt 

t 0

where k is the first-order decay coefficient [k=ln(2)÷t ]. Time-weighted average concentration 50

(CTWA) over time t can be calculated as the integral of Ct (De Sapio 1976, p. p. 97 ff) divided by 

the duration (t): 

3-24
 
-k t C = C (1 - e ) ÷ (k t). TWA 0 

A somewhat different approach is required to assess exposures to 4-chloroaniline. Immediately 

after application, residues on vegetation will be comprised solely of diflubenzuron. As 

diflubenzuron degrades, 4-chloroaniline may be formed. Field studies, however, have indicated 

no residues of 4-chloroaniline on vegetation treated with diflubenzuron (Schroeder 1980). This 

may be due to the rapid atmospheric degradation of 4-chloroaniline in air – i.e., an estimated 

halftime of 3.9 hours or about 0.16 days. This is much less than the estimated vegetation 

halftime for diflubenzuron – i.e., 9.3 days (Sundaram 1986, 1996). Thus, the rate limiting step in 

the residues of 4-chloroaniline on vegetation will be the formation of 4-chloroaniline. 

The approach for estimating concentrations of 4-chloroaniline on vegetation is conceptually 

similar to the approach taken with estimating concentrations in water. As a simplifying 

assumption, 4-chloroaniline generation will be estimated from the halftime of 9.3 days of 

diflubenzuron – i.e., direct breakdown from diflubenzuron to 4-chloroaniline. In addition, the 

dissipation of 4-chloroaniline from vegetation will be taken as the atmospheric halftime of 0.16 

days, from WHO (2003). Under these conditions and at steady state, the ratio of 4-chloroaniline 

to diflubenzuron will be ratio of the these halftimes – i.e., 0.16 days ÷ 9.3 days = 0.017. In the 

scenario specific worksheets for 4-chloroaniline, all specific worksheets modeling exposure to 

contaminated vegetation are based on concentrations of diflubenzuron. The adjustment factor of 

0.017 for 4-chloroaniline is incorporated into all worksheets involving exposure to contaminated 

vegetation (Worksheets D03, D04, F04a, F04b, F10, F11a, F11b, F12, F13a, F13b, F14a, and 

F14b). 

3-25
 
3.3. DOSE-RESPONSE ASSESSMENT 

3.3.1. Overview 

The dose-response assessment considers both diflubenzuron itself as well as 4-chloroaniline as 

an environmental metabolite of diflubenzuron. For systemic toxicity, the dose-response 

assessment involves the adoption or derivation of acute and chronic RfDs, doses that are 

considered to produce no adverse effects, even in sensitive individuals. RfDs are presented for 

both diflubenzuron and 4-chloroaniline. Cancer risk is considered quantitatively for 

4-chloroaniline and is expressed as a dose associated with a risk of 1 in 1-million. Following 

standard practices for USDA risk assessments, risk assessment values available from U.S. EPA 

are adopted directly unless there is a compelling basis for doing otherwise. When risk values are 

not available from U.S. EPA, the methods used by U.S. EPA are employed to derive surrogate 

values. 

U.S. EPA has derived a chronic RfD for diflubenzuron of 0.02 mg/kg/day. This chronic RfD is 

well-documented and is used directly for all longer term exposures to diflubenzuron. This value 

is based on a NOAEL of 2 mg/kg/day in dogs and an uncertainty factor of 100 – a factor of 10 for 

interspecies differences and a factor of 10 for sensitive subgroups. Because of the low acute 

toxicity of diflubenzuron, the U.S. EPA has not derived an acute RfD but has identified an acute 

NOAEL of 10,000 mg/kg. While this NOAEL could be used to derive a surrogate acute RfD of 

100 mg/kg, a more conservative approach is taken and a surrogate acute RfD of 11 mg/kg is 

derived based on a NOAEL of 1118 mg/kg from a study using a petroleum-based formulation of 

diflubenzuron. Diflubenzuron has been classified as a non-carcinogen by both U.S. EPA and 

WHO and no quantitative cancer risk assessment for exposures to diflubenzuron is conducted. 

The U.S. EPA has derived a chronic RfD for 4-chloroaniline of 0.004 mg/kg/day and this value is 

used in the current risk assessment to characterize risks from 4-chloroaniline for longer term 

exposures. This RfD is based on a chronic oral LOAEL of 12.5 mg/kg/day using an uncertainty 

factor of 3000, three factors of 10 for interspecies extrapolation, sensitive subgroups, and the use 

of a LOAEL with an additional factor of 3 due to the lack of data reproductive toxicity data. As 

with diflubenzuron, the U.S. EPA has not derived an acute RfD for 4-chloroaniline. For this risk 

assessment a conservative approach is taken in which a surrogate acute RfD of 0.03 mg/kg is 

based on a subchronic (90-day) NOAEL of 8 mg/kg/day. Consistent with the approach taken by 

U.S. EPA for the chronic RfD, an uncertainty factor of 300 is used – a factor of 10 for 

interspecies extrapolation, 10 for intraspecies extrapolation, and 3 for the lack of data on 

reproductive toxicity. For cancer risk, the U.S. EPA proposes a human cancer potency factor for 

4-chloroaniline of 0.0638 (mg/kg/day) -1. This potency factor is used to calculate a dose of 

1.6×10-5 mg/kg/day that would be associated with a cancer risk of 1 in 1-million. 

3.3.2. Diflubenzuron 

3.3.2.1. Chronic RfD – The most recent RfD for diflubenzuron is 0.02 mg/kg/day. This value 

is given on the U.S. EPA’s agency-wide list of approved RfDs (i.e., IRIS) (U.S. EPA 1990) and 

has been adopted by the U.S. EPA’s Office of Pesticides (U.S. EPA/OPP 1997a,b, 2001a). 

3-26
 
The chronic RfD is based on a study by Greenough et al. (1985) in which technical grade 

diflubenzuron was administered daily in gelatin capsules to dogs at doses of 0, 2, 10, 50, or 250 

mg/kg/day, 7 days/week, for 52 consecutive weeks. At the lowest dose, 2 mg/kg/day, no effects 

were noted on methemoglobin formation or other standard endpoints. This study is detailed 

further in Appendix 1. The RfD was calculated by dividing the NOAEL of 2 mg/kg/day by an 

uncertainty factor of 100, a factor of 10 for interspecies differences – i.e., extrapolation of animal 

data to humans – and a factor of 10 for intraspecies variability – i.e., individuals who might be 

most sensitive to diflubenzuron. 

Under the Food Quality Protection Act (FQPA), the U.S. EPA is required to consider an 

additional uncertainty factor of 10 for the protection of infants and children. For diflubenzuron, 

the U.S. EPA (1997a) determined that the additional uncertainty factor is not required because of 

the information on the reproductive toxicity of diflubenzuron is adequate. As discussed in 

Section 3.1.9, diflubenzuron has been tested for and does not appear to cause birth defects or 

reproductive and developmental impairment. 

For this risk assessment, the chronic RfD of 0.02 mg/kg/day is used to characterize risks for the 

general public as well as workers in longer term exposures. Because the RfD is intended to 

protect for lifetime exposures, it provides a conservative basis for comparing estimated exposure 

levels to an index of acceptable exposure. 

3.3.2.2. Acute RfD – The U.S. EPA (1997a) did not specifically derive an acute RfD for 

diflubenzuron. In discussing the acute oral toxicity of diflubenzuron and referring specifically to 

the NOAEL of 10,000 mg diflubenzuron/kg bw from the single dose study in rats and mice by 

Koopman (1977) – i.e., a dose of 40,000 mg Dimilim/kg bw – the U.S. EPA/OPP (1996) 

concludes that: 

One day single dose oral studies in rats and mice indicated 

only marginal effects on methemoglobin levels at a dose level 

of 10,000 mg/kg of a 25% wettable powder formulation. 

Sulfhemoglobin levels and Heinz bodies were not affected. 

Therefore, there is no acute dietary endpoint and a risk 

assessment for acute dietary exposure (1 day) is not necessary. 

(U.S. EPA/OPP, 1996a, p. 16). 

While this is a reasonable position, the current risk assessment is concerned with characterizing 

the risks of acute exposures as well as comparing the risks of acute exposures to diflubenzuron 

with risks associated with acute exposures other agents used to control the gypsy moth. A 

surrogate acute RfD of 100 mg/kg could be derived using the NOAEL of 10,000 mg/kg identified 

by U.S. EPA/OPP (1996a) and the uncertainty factor of 100 used by U.S. EPA/OPP (1996a) in 

deriving the chronic RfD (Section 3.3.2.1). 

3-27
 
A more conservative approach, however, is taken for the current risk assessment. As noted in the 

hazard identification (Section 3.1.14), Dimilin 4L contains petroleum oil, a substance that is 

considered potentially toxic. While no acute toxicity studies have been encountered on Dimilin 

4L, Blaszcak (1997a) has conducted a single dose gavage study in rats with Dimilin 2L, another 

petroleum based formulation of diflubenzuron. In this study, no signs of toxicity associated with 

treatment were noted at a dose of 5000 mg/kg as Dimilin 2L, equivalent to 1118 mg/kg as 

diflubenzuron. Thus, 1118 mg/kg rather than 10,000 mg/kg will be taken as the acute NOAEL. 

This value is used to calculate an acute RfD of 11 mg/kg by applying an uncertainty factor of 

100, as in the chronic RfD, and rounding to the nearest integer. 

3.3.2.3. Cancer Potency – The U.S. EPA/OPP (1996a) has determined that diflubenzuron itself 

does not pose a carcinogenic risk. Specifically, the U.S. EPA/OPP (1997a) has concluded that: 

Based on the available evidence, which included adequate 

carcinogenicity studies in rats and mice and a battery of 

negative mutagenicity studies, diflubenzuron per se is classified 

as Group E (evidence of non-carcinogenicity for humans). – 

(U.S. EPA 1997a, p. 18) 

Thus, there is no basis for identifying carcinogenicity as and endpoint of concern and this effect 

is not treated quantitatively in the current risk assessment. This is consistent with the evaluation 

of the available data on carcinogenicity by WHO (1996, 2001). 

3.3.3. 4-Chloroaniline 

3.3.3.1. Chronic RfD – The chronic RfD for 4-chloroaniline is 0.004 mg/kg/day (U.S. EPA 

1995). This RfD is based on a 2-year feeding study using rats in which the formation of 

non-neoplastic lesions of the splenic capsule was observed at 250 ppm in the diet (12.5 

mg/kg/day) (NCI 1979). This dose is classified as a LOAEL and is divided by an uncertainty 

factor of 3,000 to derive the RfD. This uncertainty factor is intended to account for intra- and 

interspecies differences and the extrapolation from a LOAEL to a NOAEL. A value of ten is 

used for each of these three uncertainty factors is given – i.e., 10 × 10 × 10. An additional factor 

of 3 was incorporated into the uncertainty factor because of the lack of supporting reproductive 

toxicity data. This data gap has also been noted by WHO (2003). Confidence in the principal 

study, the database for toxic effects, and the RfD itself is low (U.S. EPA 1995). 

For this risk assessment, the chronic RfD derived by U.S. EPA (1995) is used for characterizing 

longer-term risks for the general public. As with the RfD for diflubenzuron, this provides a 

conservative basis for assessing the risks of longer term exposures, which are typically over 

periods far less than lifetime. 

3.3.3.2. Acute RfD – As with diflubenzuron, the U.S. EPA has not proposed an acute RfD for 

4-chloroaniline. As noted in Section 3.1, acute exposures to 4-chloroaniline are likely to be 

minimal immediately after the application of diflubenzuron – i.e., prior to the environmental 

3-28
 
 

metabolism of diflubenzuron to 4-chloroaniline. Nonetheless, as detailed in Section 3.2.3.4 and 

illustrated in Figure 3-2, peak exposures to 4-chloroaniline in water may be higher than peak 

exposures to diflubenzuron in water, although the peak 4-chloroaniline exposures may occur 

weeks to months after the application of diflubenzuron. Consequently, this risk assessment will 

derive a surrogate acute RfD for 4-chloroaniline. 

The toxicology of 4-chloroaniline has been reviewed in detail by WHO (2003) and the most 

relevant studies for the current risk assessment as summarized in Appendix 1. As a conservative 

approach, the surrogate acute RfD is based on the subchronic study by Scott and Eccleston 

(1967) in which rats were dosed daily with 4-chloroaniline at 0, 8.0, 20.0, or 50.0 mg/kg for 3 

months. No hematologic or other adverse effects were observed at the lowest dose, 8 mg/kg/day. 

For the surrogate acute RfD, an uncertainty factor of 300 is used – a factor of 10 for interspecies 

extrapolation, 10 for intraspecies extrapolation, and 3 for the lack of data on reproductive 

toxicity. Thus, the surrogate acute RfD is taken as 0.03 mg/kg/day [8 mg/kg/day ÷ 300 = 

0.02666 mg/kg/day which rounds to 0.03 mg/kg/day using one significant figure]. 

3.3.3.3. Cancer Potency – In the previous risk assessment for the use of diflubenzuron in gypsy 

moth programs (USDA 1995), a cancer potency factor of 0.013 (mg/kg/day)-1 was used in the 

human health risk assessment. This was based on the NCI (1979) using the linearized multistage model. More recently, the U.S. EPA/OPP (1999, 2000a) has calculated a human cancer 

-1 potency factor for 4-chloroaniline of 0.0638 (mg/kg/day) , about a factor of 5 greater than the

previous value used by USDA (1995). 

In implementing the dietary risk assessment for the formation 4-chloroaniline from 

diflubenzuron, the U.S. EPA (2000a) has noted a potential cancer risk from 4-chlorophenylurea. 

As noted in Figure 3-1 and discussed in Section 3.1.3.3, 4-chlorophenylurea is structurally 

similar to 4-chloroaniline and is formed as an intermediate in the environmental breakdown of 

diflubenzuron to 4-chloroaniline. No specific information is available on the carcinogenicity of 

4-chlorophenylurea. As a conservative approach in their dietary risk assessment of the 

degradation products of diflubenzuron, the U.S. EPA (2000a) elected to treat 4-chlorophenylurea 

as if it were a carcinogen with the same potency as 4-chloroaniline. This approach has been 

criticized by Cardona (1999, 2001) both because of the lack of information indicating that 

4-chlorophenylurea is carcinogenic and because 4-chloroaniline does not appear to be an in vivo 

metabolite of 4-chlorophenylurea in rodents. 

As detailed in Section 3.2.3.4.3 for drinking water and Section 3.2.3.6 for contaminated 

vegetation, the current risk assessment takes a somewhat different approach to the risks posed by 

4-chlorophenylurea. There is no doubt that 4-chlorophenylurea is metabolized to 4-chloroaniline 

in the environment. Because the toxicity data on 4-chlorophenylurea are limited, the current risk 

assessment models the degradation of diflubenzuron to 4-chloroaniline as a one-step process, 

omitting the formation of 4-chlorophenylurea. While this is conceptually different from the 

equal potency assumption used by U.S. EPA (2000a), it is a conservative approach but avoids the 

3-29
 
use of a surrogate potency parameter for a compound, 4-chlorophenylurea, for which there is no 

evidence of carcinogenicity. 

For this risk assessment, the human cancer potency factor for 4-chloroaniline of 

-1 0.0638 (mg/kg/day) proposed by U.S. EPA/OPP (1999, 2000a) is used to assess cancer risks for

all longer term exposure scenarios. This potency factor is not applied directly to any acute 

exposure assessments. Nonetheless, it is worth noting that all of the longer term estimates of 

exposure are based on average values that include short-term peak exposures. Thus, these higher 

but transient acute exposures are incorporated into the cancer risk assessment. 

In the risk characterization worksheet for 4-chloroaniline (Worksheet E04 in Supplement 2), 

cancer risk is expressed as the ratio of exposure (dose in mg/kg/day) to a dose with a risk of 1 in 

1-million. In a linear cancer model, such as that used by U.S. EPA, risk is assumed to be linearly 

related to dose: 

Risk = dose × potency 

-6 Thus, taking the potency factor of 0.0638 (mg/kg/day)-1 and a risk level of 1 in 1-million (1×10 ),

the dose associated with a risk of 1 in 1-million can be calculated as: 

-6 -1 -5 dose = 1×10 ÷ 0.0638 (mg/kg/day) = 0.000015673 . 1.6×10 mg/kg/day 

This dose is used in the Worksheet E04 for the risk characterization of cancer risks associated 

with exposure to 4-chloroaniline. 

3-30
 
 

3.4. RISK CHARACTERIZATION 

3.4.1. Overview 

The risk characterization for potential human health effects associated with the use of 

diflubenzuron in USDA programs to control the gypsy moth is relatively unambiguous: none of 

the hazard quotients reach a level of concern at the highest application rate that could be used in 

USDA programs. In that many of the exposure assessments involve very conservative 

assumptions – i.e., assumptions that will tend to overestimate exposure – and because the doseresponse assessment is based on similarly protective assumptions, there is no basis for asserting 

that this use of diflubenzuron poses a hazard to human health. 

Notwithstanding the above assertion, it is worth noting that the greatest relative concern is with 

the contamination of water with 4-chloroaniline rather than with any exposures to diflubenzuron 

itself. The highest hazard quotient for diflubenzuron is 0.1, a factor of 10 below a level of 

concern. Since this hazard quotient is based on toxicity, an endpoint that is considered to have a 

population threshold, the assertion can be made that risk associated with exposure to 

diflubenzuron is essentially zero. 

This is not the case with 4-chloroaniline, which is classified as a probable human carcinogen and 

is an environmental metabolite of diflubenzuron. For 4-chloroaniline, the highest hazard 

quotient is 0.4, below the level of concern by a factor of only 2.5. The scenario of greatest 

concern involves cancer risk from drinking contaminated water. This risk would be most 

plausible in areas with sandy soil and annual rainfall rates of about 50 to 250 inches. The central 

estimate of the hazard quotient for the consumption of water contaminated with 4-chloroaniline 

and based on a cancer risk of 1 in 1-million is 0.09, below the level of concern by a factor of 10. 

3.4.2. Workers 

A quantitative summary of the risk characterization for workers is presented in Worksheet E02 of 

the diflubenzuron worksheets (Supplement 1). The quantitative risk characterization is 

expressed as the hazard quotient, which is the ratio of the estimated exposure from Worksheet 

E01 to the RfD. For acute accidental/incidental exposures, the surrogate acute RfD of 11 mg/kg 

is used (Section 3.3.3.2). For longer term general exposures – i.e., exposures that could occur 

over the course of several days, weeks, or months during an application season – the chronic RfD 

of 0.02 mg/kg/day is used (Section 3.3.3.1). 

The qualitative risk characterization for workers is reasonably unequivocal. None of the acute or 

longer term hazard quotients exceed 1, the level of concern. In the normal application of 

diflubenzuron over the course of a season or even several years, the hazard quotients range from 

0.04 to 0.07 – i.e., below the level of concern by factors of about 14 to 25. At the upper ranges of 

exposure for workers, the hazard quotients approach but do not exceed a level of concern – i.e., 

0.2 to 0.5. Similarly, the upper range of hazard quotients for accidental/incidental exposures 

range from 0.0001 to 0.03, below the level of concern by factors of about 33 to 10,000. As noted 

in Section 3.2.2.2, the only accidental/incidental exposure that exceeds general exposures 

involves wearing contaminated gloves for 1 hour. While the hazard quotient of 0.03 is 

3-31
 
 

substantially below a level of concern, the use of contaminated gloves appears to be the greatest 

source of concern in the handling of diflubenzuron. 

Diflubenzuron can cause slight irritation to the eyes (section 3.1.11). Quantitative risk 

assessments for irritation are not derived; however, from a practical perspective, eye irritation is 

likely to be the only overt effect as a consequence of mishandling diflubenzuron. This effect can 

be minimized or avoided by prudent industrial hygiene practices during the handling of the 

compound. 

3.4.3. General Public 

3.4.3.1. Diflubenzuron – A quantitative summary of the risk characterization for members of the 

general public is presented in Worksheet E04 of the diflubenzuron worksheets (Supplement 1). 

As with the risk characterization for workers, risk is expressed quantitatively as the hazard 

quotient using the surrogate acute RfD of 11 mg/kg (Section 3.3.3.2) and the chronic RfD of 

0.02 mg/kg/day is used (Section 3.3.3.1). 

Also as with workers, the qualitative risk characterization for members of the general public is 

unambiguous, with none of the acute or longer term hazard quotients exceeding 1 even at the 

upper ranges of plausible exposure. The highest hazard quotient is 0.1, the upper range of risk 

for the consumption of contaminated fish by subsistence populations. Nonetheless, this extreme 

acute scenario is below the level of concern by a factor of 10. No other acute exposure scenarios, 

many of which involve extremely conservative assumptions, approach a level of concern at the 

upper range of exposure. Based on central estimates of acute exposure, which involve somewhat 

less conservative assumptions, the acute hazard quotients range from 0.000003 to 0.02 – i.e., 

below the level of concern by factors of 50 to over 300,000. 

3.4.3.2. 4-Chloroaniline – A quantitative summary of the risk characterization for members of 

the general public is presented in Worksheet E04 of the 4-chloroaniline worksheets 

(Supplement 2). Risk is expressed quantitatively as the hazard quotient using the surrogate acute 

RfD of 0.03 mg/kg (Section 3.3.3.2) and the chronic RfD of 0.004 mg/kg/day is used (Section 

3.3.3.1). 

In terms of both toxicity and carcinogenicity, the hazard quotients for members of the general 

public are comparable to but somewhat higher than the corresponding hazard quotients for 

diflubenzuron – a maximum hazard of 0.4 for 4-chloroaniline compared to a maximum hazard 

quotient of 0.1 for diflubenzuron. 

The hazard quotient of 0.4 for 4-chloroaniline is associated with contamination of water, the 

hazard quotient for toxicity for the consumption of contaminated fish by subsistence populations 

and the hazard quotient for the dose associated with a cancer risk of 1 in 1-million for the longer 

term consumption of contaminated water. As detailed in Section 3.2.3.4 and illustrated in 

Figure 3-2, these risks are associated with the application of diflubenzuron to sandy soils in areas 

with annual rainfall rates of about 50 to 250 inches. In areas with predominantly clay or loam 

3-32
 
soils, risks will be less by factors of about 3 to 10 (Table 3-6). Also, the relatively high hazard 

quotient of 0.4 is associated with standing bodies of water – i.e., ponds or lakes. Concentrations 

of 4-chloroaniline in streams even with sandy soil will be much less (Table 3-5). 

Based on central estimates of exposure, acute hazard quotients range from 0.0004 to 0.01, below 

the level of concern by factors of 100 to 2500. Most chronic hazard quotients are in the range of 

0.000002 to 0.0005, far below a level of concern. The only exception is the central estimate of 

the hazard quotient for the consumption of contaminated water based on a cancer risk of 1 in 1­

million. This hazard quotient is 0.09, below the level of concern by a about a factor of 10. 

Nonetheless, the consumption of water that is contaminated with 4-chloroaniline as the greatest 

source of concern for members of the general public in the application of diflubenzuron to 

control the gypsy moth. 

3.4.4. Sensitive Subgroups 

Some individuals are born with a form of congenital methemoglobinemia and may be at 

increased risk of adverse effects to compounds that induce methemoglobinemia (Barretto et al. 

1984). Infants less than 3 months old have lower levels of methemoglobin (cytochrome b5) 

reductase and higher levels of methemoglobin (1.32%), compared with older children or adults 

(Centa et al. 1985; Khakoo et al. 1993; Nilsson et al. 1990). A similar pattern is seen in many 

species of mammals (Lo and Agar 1986). Some infants with an intolerance to cow's milk or soy 

protein exhibit methemoglobinemia (Murray and Christie 1993; Wirth and Vogel 1988). These 

infants would be at increased risk if exposed to any materials contaminated with diflubenzuron or 

any compound that induces methemoglobinemia. 

Individuals with poor diets may be at increased risk to some chemicals. Based on a study in rats 

(Hagler et al. 1981), iron deficiency leads to anemia but does not influence methemoglobin 

reductase activity. Thus, although individuals with poor nutritional status are generally a group 

for which there is particular concern, the available information does not support an increased 

concern for these individuals with respect to diflubenzuron exposure. 

The RfDs used in the current risk assessment quantitatively consider sensitive subgroups. As 

noted in Section 3.3.2, the chronic RfD derived by U.S. EPA (1997a) incorporates a factor of 10 

into overall uncertainty factor of 100 used for diflubenzuron to account for sensitive subgroups. 

Based on differences in methemoglobin reductase activity, a recovery mechanism for 

methemoglobinemia (Section 3.1.2), among different species, the factor of 10 for intraspecies 

variability appears adequate. The activity of this enzyme in humans appears to be about half of 

that in dogs (Calabrese 1991). 

3.4.5. Connected Actions 

The most sensitive effect for diflubenzuron, methemoglobinemia, is associated tebufenozide, 

another agent used for gypsy moth control. These two agents are likely to have an additive effect 

on methemoglobinemia but these agents are not used together. Thus, simultaneous exposures are 

unlikely. Exposure to other compounds in the environment that induce methemoglobinemia may 

3-33
 
also lead to an additive effect. Individuals exposed to combustion smoke or carbon monoxide 

(that is, agents that do oxidative damage to blood) may be at increased risk of developing 

methemoglobinemia (Hoffman and Sauter 1989; Laney and Hoffman 1992). In addition, 

individuals exposed to high levels of nitrates, either in air or in water, will have increased levels 

of methemoglobin (Woebkenberg et al. 1981) and may be at increased risks of exposure to 

compounds such as diflubenzuron. 

3.4.6. Cumulative Effects 

This risk assessment is based on single applications at the maximum allowable rate, 70 g/ha. 

This is also the maximum rate that can be applied in a single season. This approach is used to 

estimate maximum daily exposure and daily absorbed dose. Because the dispersal rate for 

diflubenzuron in the environment is relatively fast, multiple applications at lower rates per 

application will result in risks that are less than those associated with a single application at the 

maximum approved rate. Given the narrow range of application rates compared with the 

variability and uncertainties in the exposure assessments, the risks of toxic effects associated 

with a single application at less than the maximum rate will be related directly to the application 

rate. Thus, an application at 35 g/ha will entail risks that are approximately one half of those 

expected at the maximum application rate. 

3-34
 
4. ECOLOGICAL RISK ASSESSMENT
 

4.1. HAZARD IDENTIFICATION 

4.1.1. Overview 

The toxicity of diflubenzuron is well characterized in most groups of animals including 

mammals, birds, terrestrial invertebrates, fish and aquatic invertebrates. In general, 

diflubenzuron is much more toxic to some invertebrates, specifically arthropods, than vertebrates 

or other groups of invertebrates. This differential toxicity appears to involve fundamentally 

different mechanisms of action. Toxicity to sensitive invertebrate species is based on the 

inhibition of chitin synthesis. In the more tolerant vertebrate species, the mechanism of action 

appears to be a specific effect on the blood that inhibits oxygen transport. 

The species most sensitive to diflubenzuron are arthropods, a large group of invertebrates 

including insects, crustaceans, spiders, mites, and centipedes. Most of these organisms use 

chitin, a polymer (repeating series of connected chemical subunits) of a glucose-based molecule, 

as a major component of their exoskeleton – i.e., outer body shell. Diflubenzuron is an effective 

insecticide because it inhibits the the formation of chitin. This effect disrupts the normal growth 

and development of insects and other arthropods. Both terrestrial and aquatic arthropods are 

affected but some substantial differences in sensitivity are apparent. In terrestrial organisms, the 

most sensitive species include lepidopteran and beetle larvae, grasshoppers and other 

herbivorous insects. More tolerant species include bees, flies, parasitic wasps, adult beetles, and 

sucking insects. In aquatic organisms, small crustaceans that consume algae and serve as a food 

source for fish (e.g., Daphnia species) appear to be the most sensitive to diflubenzuron while 

larger insect species such as backswimmers and scavenger beetles are much less sensitive. A 

wide range of other aquatic invertebrates, other crustaceans and small to medium sized aquatic 

insect larvae, appear to have intermediate sensitivities. Not all invertebrates utilize chitin and 

these invertebrates are much less sensitive to diflubenzuron than the arthropods. For terrestrial 

invertebrates, relatively tolerant species include earthworms and snails. For aquatic species, 

tolerant species include ostracods (an arthropod) and non-arthropods such as rotifers, bivalves 

(clams), aquatic worms, and snails. 

As detailed in the human health risk assessment, the most sensitive effect in vertebrate species 

appears to involve damage to blood cells involved in the transport of oxygen. This effect has 

been demonstrated in mammals that are often employed in toxicity studies (e.g., rats and mice) as 

well as domestic animals and livestock. The effect has not been demonstrated in wildlife 

mammals, birds, or fish but it seems reasonable to assume that hemoglobin in all vertebrate 

species could be affected by exposure to diflubenzuron. Acute exposures to diflubenzuron are 

relatively non-toxic to mammals and birds. The U.S. EPA places diflubenzuron in low toxicity 

categories (III or IV) for mammals and considers diflubenzuron to be virtually non-toxic to birds 

in acute exposures and only slightly toxic to birds in subchronic exposures. This assessment is 

supported by a large number of field studies in which no direct toxic effects in mammals or birds 

have been reported. Effects, if any, on terrestrial vertebrates from the application of 

diflubenzuron are likely to be secondary to changes in food availability (i.e., reduced numbers of 

4-1
 
insects) or changes in habitat (i.e., the protection of vegetation relative to untreated areas). 

Aquatic vertebrates also appear to be relatively tolerant to diflubenzuron and this compound is 

classified by U.S. EPA as practically non-toxic to fish. This classification appears to be 

appropriate and is supported by a relatively large number of longer term toxicity studies as well 

as field studies. Changes in fish populations have been noted in some studies but the changes 

appear to be secondary to changes in food supply. Although the data on amphibians are much 

more limited than the data in fish, a similar pattern is apparent – i.e., no direct toxic effects but 

changes in food consumption patterns secondary to effects on invertebrate species. 

Data on plants and microorganisms are more limited than the data on invertebrates or vertebrates. 

Nonetheless, there does not appear to any basis for asserting that diflubenzuron will have a 

substantial effect on these organisms. 

4.1.2. Toxicity to Terrestrial Organisms 

4.1.2.1. Mammals – As summarized in Appendix 1 and discussed in the human health risk 

assessment (Section 3.1), there are a large number of toxicity studies on diflubenzuron in 

experimental mammals and these studies are relevant to the risk assessment for terrestrial 

mammals. Potential hazard to all wildlife mammals, however, may not be encompassed by the 

available data on experimental mammals – i.e., rats, mice, and dogs. As discussed in Section 

3.1.3.1 and illustrated in Figure 3-1, some mammals such as sheep and pigs will metabolize 

diflubenzuron differently from rats. Specifically, metabolism in sheep, pigs, and perhaps other 

mammalian species, will result in cleavage of the ureido bridge with the formation of metabolites 

that are different from those seen in rats. There is little indication, however, that this difference 

in metabolism will lead to marked differences in toxicity. As summarized in Appendix 1, 

substantial differences in sensitivity among different species of mammals are not apparent. One 

possibly noteworthy difference, however, is a reduction in thyroid weight in sheep (Ross et al. 

1977). As discussed in Section 3.1.8, the thyroid is an important organ in endocrine function. 

This effect, however, occurred in the absence of any signs of toxicity or changes in growth and 

may have been incidental. 

The available field studies do not indicate any substantial impacts on mammalian wildlife from 

applications of diflubenzuron (Appendix 3a). As summarized in USDA (1995), applications of 

60 to 280 g a.i./ha (0.85 to 4 oz a.i./ac) had no detectable adverse effects on the abundance of 

or reproduction in voles, field mice, and shrews (O'Connor and Moore 1975; Henderson et al. 

1977). Small mammals increased in abundance on a plot receiving 280 g a.i./ha compared with a 

control plot (Henderson et al. 1977). The adverse effect that diflubenzuron might have on bot 

flies, a parasite of small as well as large mammals, was suggested as a possible explanation. 

A more recent published field study by Seidel and Whitmore (1995) reports no effects on body 

measurements, weight, or fat content in populations of mice in areas treated with Dimilin 25 WP 

at a rate of rate of 140 g formulation/ha (35 g a.i./ha). Mice in the treated areas did consume less 

lepidopteran prey, secondary to the toxicity of diflubenzuron to lepidoptera, but total food 

consumption was not significantly different in treated and untreated plots. 

4-2
 
4.1.2.2. Birds – A relatively large number of acute and subchronic toxicity studies are available 

in standard test species – i.e., mallard ducks and bobwhite quail – as well as other less 

commonly tested species – i.e., domestic hens and red-winged blackbirds (Appendix 4). Most of 

these studies were submitted to the U.S. EPA for the registration diflubenzuron (specified in 

Appendix 4 by MRID numbers) but some have been published in the open literature (e.g., 

Kubena 1981,1982, Kubena and Witzel 1980). 

The acute toxicity of diflubenzuron to birds appears generally to be low and consistent with the 

gavage studies in rats in which gavage oral LD50 values are greater than 5000 mg/kg (Section 3.1 

and Appendix 1). As summarized in Appendix 4, red-winged blackbirds appear to be somewhat 

more sensitive than mallard ducks – i.e., a gavage NOEL for red-winged blackbirds of 2500 

mg/kg compared to a gavage NOEL for mallards of 5000 mg/kg. Nonetheless, diflubenzuron is 

classified a “virtually non-toxic” to both species as well as to bobwhite quail (U.S. EPA 1997a, 

p. 44). Based on the results of several standard reproduction studies, the the chronic dietary 

NOEC in birds is 500 ppm (U.S. EPA/OPP 1997a). 

There is one atypical report of adverse reproductive effects in birds. Smalley (1976) reports that 

Dimilin (NOS), incorporated into the feed (dose not specified) of chicks (presumably chickens) 

for 13 weeks, resulted in an increased incidence of fat deposition in female chicks. The treated 

chicks weighed 6 1⁄2 lbs, compared to normal weight of 3 lbs for controls (broilers) and males. In 

addition, Smalley (1976) reports a dose-related decrease in testosterone in treated males resulting 

in undeveloped combs, wattles, feathers, and voice. Very few experimental details are included 

in this study. Given the large number of other studies in birds in which no effects on 

reproduction were apparent, the report by Smalley (1976) appears to be an aberration. 

The lack of direct effects on birds is supported by several field studies summarized in 

Appendix 3a. Some effects secondary to reduced lepidoptera prey may include increased 

foraging range (Cooper et al. 1990), relocation (Sample et al. 1993a,b) and lower body fat 

(Whitmore 1993). 

4.1.2.3. Terrestrial Invertebrates – A large and relatively complex body of information is 

available on the toxicity of diflubenzuron to both target and non-target invertebrates. This 

information consists of both laboratory studies in which exposures are relatively well defined and 

controlled (Appendix 5) as well as field studies in which exposures are typically characterized as 

application rates (Appendix 3a). 

A synopsis of the field studies in which exposures can be expressed in units of application rate 

(g/ha) are presented in Table 4-1. The first column in this table gives ranges of application rates 

spanning over an order of magnitude. The second and third columns provide species or groups 

of species in which no adverse effects (column 2) or adverse effects (column 3) were noted 

within the corresponding range of application rates. For each species or group the reference is 

given to a field study summarized in Appendix 3a. A similar summary table is not provided for 

the laboratory toxicity studies. As discussed further in the dose-response assessment 

4-3
 
(Section 4.3.2.3), these studies were conducted using highly variable experimental designs and 

meaningful comparisons among the various toxicity assays summarized in Appendix 5 are 

difficult. Additional details of the comparisons among the various field studies are also provided 

in the dose-response assessment (see discussion of Table 4-5 in Section 4.3.2.3). 

The insecticidal action of diflubenzuron is based on the inhibition of chitin synthesis. Chitin is a 

polymer (repeating series of connected chemical subunits) of a glucose-based molecule and 

comprises a substantial proportion of the exoskeleton (outer-shell) of insects. Consequently, the 

inhibition of chitin synthesis disrupts the growth and development of insects. Chitin is also 

contained in other arthropods (i.e., crustaceans, spiders, and centipedes) as well as some fungi. 

Thus, the mode of action of diflubenzuron as a insecticide to target species is also relevant to 

effects on non-target insects as well as other arthropods (Cardona 1999; Cunningham 1986; 

Eisler 1992; Fisher and Hall 1992; Hobson 2001; Lengen, 1999; Wilson 1997; Wilcox and 

Coffey 1978). Diflubenzuron also exerts ovicidal effects in several species (Ables et al. 1977; 

Büchi and Jossi, 1979; Kumar et al. 1994; ) and has been shown to inhibit egg production in 

some species (Rumpf et al. 1998; Medina et al. 2002; Medina et al. 2003). 

While the mechanism of action of diflubenzuron is not specific to target insects, there is ample 

data indicating substantial differences in sensitivity among various groups of terrestrial 

invertebrates. Invertebrates without exoskeletons, such as earthworms and snails, do not utilize 

chitin and diflubenzuron is relatively non-toxic to these species (Berends and Thus 1992; 

Berends et al. 1992). Even among different groups of arthropods, however, differences in 

sensitivity to diflubenzuron seem apparent. Species that are most sensitive to diflubenzuron 

include lepidopteran and beetle larvae, grasshoppers and other chewing herbivorous insects 

(Berry et al. 1993; Butler 1993; Butler et al. 1997; Elliott and Iyer 1982; Jepson and Yemane 

1991; Jepson and Martinat et al. 1998, 1993; Kumar et al. 1994; McWhorter and Shapard 1971; 

Sample et al.1993b; Sinha et al. 1990; Redfern et al. 1980; Yemane 1991). Other species are 

relatively tolerant to diflubenzuron. These include flies, wasps that are parasites on insect eggs, 

adult beetles, and sucking insects (Ables et al. 1975; Broadbent and Pree, 1984a; Brown and 

Respicio, 1981; Bull and Coleman, 1985; De Clercq et al. 1995b; Deakle and Bradley 1981; 

Delbeke et al. 1997; Gordon and Cornect, 1986; Keever et al. 1977; Martinat et al., 1988; Webb 

et al. 1989; Zacarias et al. 1998; Zungoli et al. 1983). 

The honey bee is a standard test species used by U.S. EPA to classify the toxicity of pesticides to 

non-target invertebrates. Based on early acute oral and contact toxicity studies in honey bees 

with LD50 values of >30 μg/bee and >114.8 μg/bee (Atkins et al. 1974; Stevenson 1978), the U.S. 

EPA (1997a) has classified diflubenzuron as “practically non-toxic to honey bees” (U.S. EPA 

1997a, p. 81). As discussed further in the dose-response assessment (Section 4.3.2.3), several 

other laboratory toxicity studies also indicate that diflubenzuron is not highly toxic to bees 

(Chandel and Gupta 1992; Elliott and Iyer, 1982; Gijswijt, 1978; Kuijpers, 1989; Nation et al. 

1986; Yu et al. 1984) and this is supported for several field studies conducted at application rates 

comparable to or substantially higher than those used to control the gypsy moth (Buckner et al. 

1975; Emmett and Archer 1980; Matthenius, 1975; Schroeder 1978a; Schroeder 1980). In 

4-4
 
addition, no detectable amounts of diflubenzuron were found in honey bees in areas treated with 

diflubenzuron (Cochran and Poling 1995). Some studies have noted adverse effects in bees. As 

summarized in Appendix 5, Stoner and Wilson (1982) and (Thompson and Wilkins 2003) noted 

transient decreases in brood production at relatively high concentrations (10 ppm) in longer term 

exposures. At 1 ppm or less, however, no effects were noted. Barrows (1995) noted a decrease 

in the mean number of pollinating insects in watersheds during a year in which diflubenzuron 

was applied but not in the following year. 

In addition to the acute toxic effects of diflubenzuron, mediated primarily through inhibition of 

chitin, adverse reproductive effects have been reported in several different orders of insects 

including moths (Beevi and Dale 1984; Tembhare and Shinde 1998), beetles (Büchi and Jossi 

1979; Khebbeb et al. 1997; Mani et al. 1997; Soltani and Soltani-Mazouni 

1994a,b,1995a,b,1997), grasshoppers (Mathur 1998), lacewings (Medina et al. 2002; Medina et 

al. 2003; Rumpf et al. 1998), and true bugs – i.e., Order Hemiptera including the suborder 

Heteroptera (Redfern et al. 1980; Sindhu and Muraleedharan 1997). 

In Lepidoptera, reproductive effects were reported by Beevi and Dale (1984), who noted a high 

incidence of sterility in the rice swarming caterpillar (Spodoptera mauritania) after exposures to 

relatively high concentrations of Dimilin – 10 ppm and higher. The mechanism of this 

reproductive effect is unclear but may involve the endocrine system – i.e., hormone release by 

neurosecretory cells. This has been noted in larvae of the fruit-sucking moth, Othreis materna 

(Tembhare and Shinde 1998) and in the cotton bug (Dysdercus cingzrlattis) (Sindhu and 

Muraleedharan 1997). In some other species of Lepidoptera – i.e., tufted apple bud moth – pupae 

are sensitive to diflubenzuron but no effects are apparent on reproduction (Biddinger and Hull 

1999). 

In beetles (Coleoptera), effects on larvae, eggs, and reproductive performance have been noted 

(Büchi and Jossi1979; Mani et al. 1997). In the mealworm, diflubenzuron impacts lipid 

metabolism in fat bodies and ovaries (Khebbeb et al. 1997). A series of studies in this species 

(Soltani and Soltani-Mazouni 1997; Soltani-Mazouni and Soltani1994a,b, 1995b) suggest that 

the decreased fecundity observed in this and other insect species may be associated with the 

effect of diflubenzuron on oogenesis, possibly due to changes vitellogenic precursors, the 

production of ecdysteroid by follicle cells, and/or the inhibition of ovarian DNA synthesis. 

Direct damage to ovary tissue has also been observed in one species of Orthoptera, a 

grasshopper, but the mechanism of action in this species has not been studied (Mathur 1998). 

Reproductive effects in lacewings (Neuroptera) have been noted by Rumpf et al. (1998) and 

Medina et al. (2002, 2003). As detailed in Appendix 5, contact exposures to diflubenzuron at 

2 0.07 μg/cm resulted in a substantial decrease in egg production and complete infertility in 13%

of the exposed animals. No effects on egg production or hatching in this species have been 

observed after direct topical applications at doses as low as 0.5 ng/insect. At a substantially 

higher dose, 75 ng/insect, egg hatching was reduced by 32%. (Medina et al. 2002, 2003). 

4-5
 
4.1.2.4. Terrestrial Plants (Macrophytes) – As noted in U.S. EPA/OPP (1997a), no terrestrial 

plant toxicity studies had been submitted to the U.S. EPA at the time of the reregistration of 

diflubenzuron. In the literature search conducted for the current risk assessment, no bioassays for 

herbicidal activity of diflubenzuron were encountered in either the published literature or in the 

more recent U.S. EPA/OPP files. 

There are a large number of terrestrial field studies regarding the efficacy of diflubenzuron 

applied to terrestrial vegetation for the control of various insect pests including the gypsy moth 

(Appendix 3a). If diflubenzuron were toxic to terrestrial plants at application rates that are used 

in the field, it is plausible that adverse effects would have been reported in this literature. No 

such reports were encountered. Thus, there is no basis for asserting that diflubenzuron will cause 

adverse effects in terrestrial plants and such effects will not be considered quantitatively in this 

risk assessment. 

4.1.2.5. Terrestrial Microorganisms – As discussed in Section 3.2 and summarized in 

Appendix 2 (Environmental Fate) and Appendix 3a (Terrestrial Field Studies), diflubenzuron is 

readily degraded by terrestrial microorganisms. The degradation of diflubenzuron by soil 

microorganisms suggests that this compound is not toxic to soil microorganisms and this 

presumption may account for the relatively few studies on microbial toxicity. Fungi, however, 

do contain chitin in cell walls and thus could be a potential target. Booth (1978) found no 

inhibition of fungal growth in several species of fungi (Aspergillus, Fusarium, Rhizopus, 

Trichoderma) at concentrations of up to 100 ppm in growth media – i.e. mg diflubenzuron per kg 

of soil. Some growth inhibition, however, was noted in a species of Pythium at a concentration 

of 50 ppm. Inhibition of Rhizoctonia solani, another terrestrial fungus, has been noted at 300 

ppm (Townshend et al. 1983). 

The lack of microbial toxicity was also specifically noted in one field study in which no effects 

on soil or litter populations of bacteria, actynomycetes or fungi were noted after applications of 

diflubenzuron at a rate of 67.26 g/ha (Kurczewski et al. 1975; Wang 1975), field and laboratory 

studies on molds and leaf litter or soil bacteria (Landolt and Stephenson 1995), and studies on 

mycorrhizal or debris decomposing fungi (Iskra et al. 1995; Gundrum et al. 1995). 

One study has noted minor and transient changes in microbial activity. Sexstone (1995) 

2 conducted a laboratory study in which soil cores were treated at 4.418μg/44.2 cm , roughly

2 2 2 2 equivalent to an application rate of 10 g/ha [4.418μg/44.2 cm × 10,000 cm /m × 10,000 m /ha =

9,995,475 μg/ha . 10 g/ha]. Only transient and sporadic decreases were noted in microbial 

biomass [Figure 14-1 in Sexton 1995]. These changes in microbial activity were apparent up to 

day 35 after treatment but there were no changes by 64 days after treatment. Changes in 

respiration [Figure 14-2 in Sexton 1995] and nitrification [Figures 14-3 to 14-6 in Sexton 1995] 

and appear to insubstantial. While some of the differences were statistically significant at some 

time points, Sexstone (1995) characterizes the effects a “minor” and this assessment appears 

reasonable. 

4-6
 
4.1.3. Aquatic Organisms. 

4.1.3.1. Fish – The toxicity of diflubenzuron to fish is well characterized in terms of both acute 

and chronic toxicity and one mesocosm study is available (Appendix 6). In addition, several of 

the aquatic field studies (Appendix 3b) involve observations on fish populations. Diflubenzuron 

has a low order of acute toxicity to fish, with 96-hour LC50 values in the range of over 25 

mg/L(the value for yellow perch reported by Johnson and Finley 1980) to over 500 mg/L (the 

value for fathead minnow reported by Reiner and Parke 1975). In addition to data on technical 

grade diflubenzuron, some studies have also been conducted on Dimilin 25W (Julin and Sanders 

1978 with additional studies summarized in U.S. EPA 1997a) and these studies indicate that the 

toxicity of Dimilin 25W is not greater than the toxicity of technical grade diflubenzuron. No 

studies have been encountered on the acute toxicity of Dimilin 4L to fish. Based on the available 

information, the U.S. EPA (1997a, p. 47) has classified diflubenzuron as “practically non-toxic” 

to fish in terms of risks from acute exposures. 

Diflubenzuron also appears to be relatively non-toxic to fish in longer term exposures. One 

standard assay for longer term toxicity in fish involves exposing fish eggs to a compound and 

maintaining the exposure through to the fry stage. In this type of assay, concentrations up to 45 

ppb has no effect on egg or fry of steelhead trout, fathead minnows, or guppies (Hansen and 

Garton 1982a). In addition, no effects were seen in longer-term studies at concentrations up to 

100 ppb (Cannon and Krize 1976) or in 2-generation reproduction studies at concentrations of up 

to 50 ppb (Livingston and Koenig 1977). 

As discussed in Section 4.1.3.2, diflubenzuron is much more toxic to invertebrates than to fish 

and indirect effects on fish are plausible based on a decrease in invertebrate populations. Such 

effects have been demonstrated in mesocosm studies (Moffett and Tanner 1995; Tanner and 

Moffett 1995) in which concentrations as low as 2.5 ppb resulted in decreased growth of fish in 

littoral enclosures – i.e., populations of fish placed and monitored in enclosures along the shore 

of a body of water. The reduced growth observed in these studies was attributed to a reduction in 

macroinvertebrates that serve as a food source for the fish. 

It is unclear, however, that secondary effects on fish growth or populations will be observed in 

the field. None of the field studies summarized in Appendix 3b note any adverse effects on fish 

in applications comparable to or greater than those used in the control of the gypsy moth. For 

example, Farlow et al. (1978) conducted a relatively large field study in a marsh area treated with 

six applications of diflubenzuron at 28 g a.i./ha – i.e., a cumulative application of 168 g/ha. 

While substantial shifts were noted in various invertebrates (Appendix 3a and Section 4.1.3.2), 

populations of mosquito fish (Gambusia affinis) and American flag fish (Jordanella floridae) 

increased. Similarly, no effects on the growth of fish were noted in ponds directly treated with 

diflubenzuron at a concentration of 5 ppb (Apperson et al. 1977, 1978) or 13 ppb (Colwell and 

Schaefer 1980). The study by Colwell and Schaefer (1980) did note a shift in diet of fish 

(secondary to changes in food availability) but no effect on growth rates or general condition of 

the fish. 

4-7
 
4.1.3.2. Amphibians – Amphibians are not standard test organisms for toxicity studies and no 

standard bioassays on amphibians have been encountered in the open literature or U.S. EPA/OPP 

files. Two field studies (Pauley 1995a,b), however, are available on salamanders. Both of these 

studies were conducted as part of a large study on the effects of spraying diflubenzuron in the 

northeast for control of the gypsy moth (Reardon 1995a). In this study, two watersheds were 

treated with Dimilin 4L in 1992 at a rate of 80g/ha (0.03 lb/acre) (Reardon 1995b). Pauley 

(1995a,b) conducted field studies to assess effects on both aquatic (Pauley 1995a) and terrestrial 

salamanders (Pauley 1995b). While all salamanders are amphibians, some species spend most of 

their time on land while others spend most of their time in water. In aquatic salamanders, 

diflubenzuron treatment was associated with a shift in dietary consumption to more hard-bodied 

prey secondary to a reduction in the availability of soft-bodied prey. This is similar to the pattern 

with fish as noted above. No effects in salamanders, however, were noted based on body size or 

population (Pauley 1995). In terrestrial salamanders, similar results were observed with no 

change in body size or body fat associated with treatment but a shift was seen in food 

consumption to hard-bodied prey (Pauley 1995b). 

4.1.3.3. Aquatic Invertebrates – As summarized in Appendix 7, there is a very large and diverse 

body of literature indicating that diflubenzuron is highly toxic to many aquatic invertebrates. 

Because diflubenzuron inhibits the synthesis of chitin, crustaceans (arthropods which rely on 

chitin synthesis for the formation of the exoskeleton) are the aquatic invertebrates that are most 

sensitive to diflubenzuron. 

One of the most common crustacean species used in freshwater invertebrate toxicity studies is 

Daphnia magna, a member of Daphnidae in the order Cladocera. These and other zooplankton 

feed on aquatic algae and are a source of food for fish. Many bioassays, both acute and chronic, 

have been conducted on Daphnia magna (Hansen and Garton 1982a; Kuijpers 1988; Majori et al. 

1984; Surprenant 1988) as well as a related species, Ceriodaphnia dubia (Hall 1986). As 

detailed further in the dose-response assessment, these organisms are among the most sensitive to 

diflubenzuron, with acute LC50 values of about 2 μg/L (Hall 1986; Hansen and Garton 1982a). 

Several other crustacean species appear to be about as sensitive or only somewhat less sensitive 

to diflubenzuron as daphnids (Appendix 7). 

Broad generalizations are somewhat difficult to make, however, because of the diversity of the 

studies that have been conducted. Nonetheless, large insects appear to be much more tolerant to 

diflubenzuron than crustaceans, with acute LC50 values on the order of 2123 μg/L for 

backswimmers (Lahr et al. 2001) and an NOEC of 250 μg/L for scavenger beetles (Miura and 

Takahashi 1974). 

Organisms that do not rely on chitin for an exoskeleton are much less sensitive to diflubenzuron. 

In the microcosm study by Corry et al. (1995) concentrations of diflubenzuron that caused 

adverse effects in cladocerans caused no adverse effects in rotifers – an aquatic invertebrate that 

lacks an exoskeleton. Similar tolerance in rotifers have been observed in littoral enclosure 

studies at diflubenzuron concentrations of up to 30 μg/L (Liber and O’Halloran 1995). At about 

4-8
 
the same concentration, 30 μg/L, two species of snails and aquatic worms were not affected by 

exposures to diflubenzuron (Hansen and Garton 1982a,b). One common genus of snail, Physa, 

had a reported LC50 value of greater than 125 mg/L – i.e., 125,000 μg/L. Ostracods (small 

bivalve crustaceans) were not affected by diflubenzuron at concentrations up to 2.5 μg/L (Liber 

and O’Halloran 1995) and much larger Quahog clams (Mercinaria mercinaria) were unaffected 

at concentrations up to 320 μg/L (Surprenant 1989). 

As with fish, no data have been located on the toxicity of Dimilin 4L. Lahr (2000, 2001) used a 

“solvent based” formulation of diflubenzuron but did not specify the formulation as Dimilin 4L. 

The 48-hour EC50 of 0.74 μg/L (0.60-0.88 μg/L) of the solvent based formulation in fairy shrimp, 

Streptocephalus sudanicus reported by Lahr (2001) is comparable to EC50 value of 0.65 μg/L for 

technical grade diflubenzuron reported in grass shrimp, Palaemonetes pugio (Tourat and Rao 

1987). Toxicity studies are available on Dimilin 25W and, as with fish, the toxicity of Dimilin 

25W appears to be the same as technical grade diflubenzuron when exposures are expressed in 

units of active ingredient (Wilson and Costlow 1986). Thus, there does not appear to be a basis 

for asserting that the formulated products containing diflubenzuron are more hazardous than 

diflubenzuron itself. 

The available field studies on the effects of diflubenzuron on aquatic invertebrates reenforce the 

standard toxicity studies, indicating that diflubenzuron will impact invertebrate populations. 

Several of these studies, however, were conducted at application rates substantially higher than 

those used to control the gypsy moth. As noted in the program description (Section 2), the 

maximum application rate that will be used in USDA programs is about 70 g/ha. Many of the 

studies in which severe adverse effects were observed in aquatic invertebrate populations 

involved multiple applications at rates between about 110 g/ha and 560 g/ha (e.g., Ali and Mulla 

1978a,b; Ali et al. 1988; McAlonan 1975). Similarly, other field studies involve direct 

applications to open water, a treatment method that is not part of USDA program activities, and 

which resulted in water concentrations that are in the range of 10 ppb (e.g., Apperson et al. 1977; 

Boyle et al. 1996; Colwell and Schaefer 1980; Lahr et al. 2000; Sundaram et al. 1991). As 

discussed further in Section 4.2, concentrations of 10 ppb or greater are in the range of peak 

concentrations that are likely to be encountered in USDA programs. Concentrations in the range 

of 10 ppb, however, are substantially higher than average concentrations of diflubenzuron in 

water that are likely to be encountered in USDA programs. 

Those field studies that used lower application rates more typical of USDA programs (e.g., 

Farlow 1976; Griffith et al. 1996; Griffith et al. 2000; Hurd et al. 1996; Jones and Kochenderfer 

1987; Reardon 1995a) have noted some effects on freshwater invertebrates, particularly smaller 

crustaceans, but the effects were much less severe than those seen in the higher application rate 

studies. This is discussed further in Section 4.4 (Risk Characterization). 

4.1.3.4. Aquatic Plants – Data on the toxicity of diflubenzuron to aquatic plants is summarized 

in Appendix 8. Most studies report no direct toxic effects of diflubenzuron on aquatic plants 

(algae or macrophytes) at concentrations of 100 μg/L or higher (Booth and Ferrell 1977; 

4-9
 
Thompson and Swigert 1993a,b,c) and no indirect effects on aquatic macrophytes (Moffett 

1995). A decrease in periphyton in littoral enclosures, however, was noted by Moffett (1995) at 

7.0, or 30 μg/L but not at 0.7 or 2.5 μg/L. This effect was attributed not to a direct toxic effect 

on the periphyton but to the loss of grazers (e.g., cladocera) that may have induced premature 

senescence in periphyton secondary to a decrement in water quality. 

4.1.3.5. Aquatic Microorganisms – There is very little information suggesting that 

diflubenzuron will adversely affect aquatic microorganisms. No marked differences in numbers 

of fungal taxa in treated and untreated watersheds were noted by Dubey (1995) in a survey of 

watersheds treated with diflubenzuron for the control of the gypsy moth. In an aquatic 

mesocosm, Kreutzweiser et al. (2001) did note a slight but significant effect of diflubenzuron (50 

μg/L and 50,000 μg/L) on microbial decomposition and respiration. Changes at 50 μg/L, 

however, were only marginally significant and variable over the 21-day period. 

In the Kreutzweiser et al. (2001) study, Dimilin 4L was used. This is the only laboratory study 

involving Dimilin 4L. Because no corresponding studies are available on Dimilin 25W or 

technical grade diflubenzuron, inferences concerning the potential effect of the petroleum solvent 

in Dimilin 4L cannot be made. 

4-10
 
4.2. EXPOSURE ASSESSMENT 

4.2.1. Overview 

As in the human health risk assessment (Section 3.2), exposures are estimated for both 

diflubenzuron and 4-chloroaniline. A full set of exposure assessments are developed for 

diflubenzuron but only a subset of exposure assessments are developed for 4-chloroaniline. This 

approach is taken, again as in the human health risk assessment, because 4-chloroaniline is 

assessed as an environmental metabolite of diflubenzuron. Thus, immediately after application, 

the amount of 4-chloroaniline as an environmental metabolite will be negligible. Consequently, 

the direct spray scenarios as well as the consumption of insects and the consumption of small 

mammals after a direct spray are not included for 4-chloroaniline. Also as in the human health 

risk assessment, all standard chronic exposure scenarios are included for 4-chloroaniline. 

Details of the exposure assessments for diflubenzuron and 4-chloroaniline are given in the two 

sets of worksheets that accompany this risk assessment: Supplement 1for diflubenzuron and 

Supplement 2 for 4-chloroaniline. All exposure assessments are based on the maximum 

application rate of 70 g/ha. 

Terrestrial animals might be exposed to any applied pesticide from direct spray, the ingestion of 

contaminated media (vegetation, prey species, or water), grooming activities, or indirect contact 

with contaminated vegetation. For diflubenzuron, the highest acute exposures for small 

terrestrial vertebrates will occur after a direct spray and could reach up to about 10 mg/kg at an 

application rate of 70 g/ha. Exposures anticipated from the consumption of contaminated 

vegetation by terrestrial animals range from central estimates of about 0.08 mg/kg for a small 

mammal to 2 mg/kg for a large bird with upper ranges of about 0.2 mg/kg for a small mammal 

and 5 mg/kg for a large bird. The consumption of contaminated water leads to much lower levels 

of exposure. A similar pattern is seen for chronic exposures. Estimated longer-term daily doses 

for a small mammal from the consumption of contaminated vegetation at the application site are 

in the range of about 0.001 mg/kg to 0.005 mg/kg. Large birds feeding on contaminated 

vegetation at the application site could be exposed to much higher concentrations, ranging from 

about 0.08 mg/kg/day to 0.7 mg/kg/day. The upper ranges of exposure from contaminated 

vegetation far exceed doses that are anticipated from the consumption of contaminated water, 

which range from about 0.0000001 mg/kg/day to 0.00001 mg/kg/day for a small mammal. 

Exposures of terrestrial organisms to 4-chloroaniline tend to be much lower than those for 

diflubenzuron. The highest acute exposure is about 0.2 mg/kg, the approximate dose for the 

consumption of contaminated water by a small mammal and the consumption of contaminated 

fish by a predatory bird. The highest longer term exposure is 0.0002 mg/kg/day, the dose 

associated with the consumption of contaminated vegetation by a large bird. 

Exposures to aquatic organisms are based on the same information used to assess the exposures 

of terrestrial species from contaminated water. At the maximum application rate of 70 g/ha, the 

upper range of the expected peak concentration of diflubenzuron in surface water is taken as 16 

μg/L. The lower range of the concentration in ambient water is estimated at 0.01 μg/L. The 

central estimate of concentration of diflubenzuron in surface water is taken as 0.4 μg/L. 

4-11
 
4.2.2. Terrestrial Animals 

Terrestrial animals might be exposed to any applied insecticide from direct spray, the ingestion 

of contaminated media (vegetation, prey species, or water), grooming activities, or indirect 

contact with contaminated vegetation. 

In this exposure assessment, estimates of oral exposure are expressed in the same units as the 

available toxicity data. As in the human health risk assessment, these units are usually expressed 

as mg of agent per kg of body weight and abbreviated as mg/kg for terrestrial animals. One 

exception in this risk assessment involves terrestrial invertebrates. As detailed in the doseresponse assessment (Section 4.3), toxicity data in units of mg/kg bw are available for some 

terrestrial invertebrates and these data are used in a manner similar to that for terrestrial 

vertebrates. For other species, however, standard toxicity studies report units that are not directly 

useful in a quantitative risk assessments – e.g., contact toxicity based on petri dish exposures. As 

an alternative, some dose response assessments are based on field studies in which the dose 

metameter is simply the application rate in units of mass per area such as g a.i./ha. 

For dermal exposures to terrestrial animals, the units of measure usually are expressed in mg of 

agent per cm2 of surface area of the organism and abbreviated as mg/cm . 2 In estimating dose, 

however, a distinction is made between the exposure dose and the absorbed dose. The exposure 

dose is the amount of material on the organism (i.e., the product of the residue level in mg/cm2 

and the amount of surface area exposed), which can be expressed either as mg/organism or 

mg/kg body weight. The absorbed dose is the proportion of the exposure dose that is actually 

taken in or absorbed by the animal. 

The exposure assessments for terrestrial animals are summarized in Worksheet G01. As with the 

human health exposure assessment, the computational details for each exposure assessment 

presented in this section are provided scenario specific worksheets (Worksheets F01 through 

F16b). Given the large number of species that could be exposed to insecticides and the varied 

diets in each of these species, a very large number of different exposure scenarios could be 

generated. For this generic risk assessment, an attempt is made to limit the number of exposure 

scenarios. 

Because of the relationship of body weight to surface area as well as the consumption of food 

and water, small animals will generally receive a higher dose, in terms of mg/kg body weight, 

than large animals will receive for a given type of exposure. Consequently, most general 

exposure scenarios for mammals and birds are based on a small mammal or bird. For mammals, 

the body weight is taken as 20 grams, typical of mice, and exposure assessments are conducted 

for direct spray (F01 and F02a), consumption of contaminated fruit (F03, F04a, F04b), and 

contaminated water (F05, F06, F07). Grasses will generally have higher concentrations of 

insecticides than fruits and other types of vegetation (Fletcher et al. 1994; Hoerger and Kenaga 

1972). Because small mammals do not generally consume large amounts of grass, the scenario 

for the assessment of contaminated grass is based on a large mammal (Worksheets F10, F11a, 

and F11b). Other exposure scenarios for mammals involve the consumption of contaminated 

4-12
 
insects by a small mammal (Worksheet F14a) and the consumption by a large mammalian 

carnivore of small mammals contaminated by direct spray (Worksheet F16a). Exposure 

scenarios for birds involve the consumption of contaminated insects by a small bird (Worksheet 

F14b), the consumption of contaminated fish by a predatory bird (Worksheets F08 and F09), the 

consumption by a predatory bird of small mammals contaminated by direct spray, and the 

consumption of contaminated grasses by a large bird (F12, F13a, and F13b). 

While a very large number of other exposure scenarios could be generated, the specific exposure 

scenarios developed in this section are designed as conservative screening scenarios that may 

serve as guides for more detailed site-specific assessments by identifying the groups of organisms 

and routes of exposure that are of greatest concern. 

4.2.2.1. Direct Spray – In the broadcast application of any insecticide, wildlife species may be 

sprayed directly. This scenario is similar to the accidental exposure scenarios for the general 

public discussed in Section 3.2.3.2. In a scenario involving exposure to direct spray, the amount 

absorbed depends on the application rate, the surface area of the organism, and the rate of 

absorption. 

For this risk assessment, three groups of direct spray exposure assessments are conducted. The 

first, which is defined in Worksheet F01, involves a 20 g mammal that is sprayed directly over 

one half of the body surface as the chemical is being applied. The range of application rates as 

well as the typical application rate is used to define the amount deposited on the organism. The 

absorbed dose over the first day (i.e., a 24-hour period) is estimated using the assumption of firstorder dermal absorption. An empirical relationship between body weight and surface area 

(Boxenbaum and D’Souza 1990) is used to estimate the surface area of the animal. The 

estimates of absorbed doses in this scenario may bracket plausible levels of exposure for small 

mammals based on uncertainties in the dermal absorption rate. 

Other, perhaps more substantial, uncertainties affect the estimates for absorbed dose. For 

example, the estimate based on first-order dermal absorption does not consider fugitive losses 

from the surface of the animal and may overestimate the absorbed dose. Conversely, some 

animals, particularly birds and mammals, groom frequently, and grooming may contribute to the 

total absorbed dose by direct ingestion of the compound residing on fur or feathers. Furthermore, 

other vertebrates, particularly amphibians, may have skin that is far more permeable than the skin 

of most mammals. Quantitative methods for considering the effects of grooming or increased 

dermal permeability are not available. As a conservative upper limit, the second exposure 

scenario, detailed in Worksheet F02a, is developed in which complete absorption over day 1 of 

exposure is assumed. 

Because of the relationship of body size to surface area, very small organisms, like bees and 

other terrestrial invertebrates, might be exposed to much greater amounts of a pesticide per unit 

body weight compared with small mammals. Consequently, a third exposure assessment is 

developed using a body weight of 0.093 g for the honey bee (USDA/APHIS 1993) and the 

4-13
 
equation above for body surface area proposed by Boxenbaum and D’Souza (1990). Because 

there is no information regarding the dermal absorption rate of diflubenzuron by bees or other 

invertebrates, this exposure scenario, detailed in Worksheet F02b, also assumes complete 

absorption over the first day of exposure. As noted above, exposures for other terrestrial 

invertebrates are based on field studies in which application rate is the most relevant expression 

of exposure. This is discussed further in Section 3.3 (Dose-Response Assessment) and Section 

3.4 (Risk Characterization). 

Direct spray scenarios are not given for large mammals. As noted above, allometric relationships 

dictate that large mammals will be exposed to lesser amounts of a compound in any direct spray 

scenario than smaller mammals. 

4.2.2.2. Indirect Contact – As in the human health risk assessment (see Section 3.2.3.3), the 

only approach for estimating the potential significance of indirect dermal contact is to assume a 

relationship between the application rate and dislodgeable foliar residue. Unlike the human 

health risk assessment in which transfer rates for humans are available, there are no transfer rates 

available for wildlife species. As discussed in Durkin et al. (1995), the transfer rates for humans 

are based on brief (e.g., 0.5 to 1-hour) exposures that measure the transfer from contaminated soil 

to uncontaminated skin. Wildlife, compared with humans, are likely to spend longer periods of 

time in contact with contaminated vegetation. It is reasonable to assume that for prolonged 

exposures a steady state may be reached between levels on the skin, rates of absorption, and 

levels on contaminated vegetation, although there are no data regarding the kinetics of such a 

process. The bioconcentration data on diflubenzuron indicates that this compound will 

accumulate in the tissue of the fish. Thus, it is plausible that absorbed dose resulting from 

contact with contaminated vegetation will be as great as those associated with comparable direct 

spray scenarios. 

4.2.2.3. Ingestion of Contaminated Vegetation or Prey – Since diflubenzuron will be applied to 

vegetation, the consumption of contaminated vegetation is an obvious concern and separate 

exposure scenarios are developed for acute and chronic exposure scenarios for a small mammal 

(Worksheets F04a and F04b) and large mammal (Worksheets F10, F11a, and F11b) as well as 

large birds (Worksheets F12, F13a, and F13b). 

For the consumption of contaminated vegetation, a small mammal is used because allometric 

relationships indicate that small mammals will ingest greater amounts of food per unit body 

weight, compared with large mammals. The amount of food consumed per day by a small 

mammal (i.e., an animal weighing approximately 20 g) is equal to about 15% of the mammal's 

total body weight (U.S. EPA/ORD 1989). When applied generally, this value may overestimate 

or underestimate exposure in some circumstances. For example, a 20 g herbivore has a caloric 

requirement of about 13.5 kcal/day. If the diet of the herbivore consists largely of seeds (4.92 

kcal/g), the animal would have to consume a daily amount of food equivalent to approximately 

14% of its body weight [(13.5 kcal/day ÷ 4.92 kcal/g)÷20g = 0.137]. Conversely, if the diet of 

the herbivore consists largely of vegetation (2.46 kcal/g), the animal would have to consume a 

4-14
 
daily amount of food equivalent to approximately 27% of its body weight [(13.5 kcal/day ÷ 2.46 

kcal/g)÷20g = 0.274] (U.S. EPA/ORD 1993, pp.3-5 to 3-6). For this exposure assessment 

(Worksheet F03), the amount of food consumed per day by a small mammal weighing 20 g is 

estimated at about 3.6 g/day or about 18% of body weight per day from the general allometric 

relationship for food consumption in rodents (U.S. EPA/ORD 1993, p. 3-6). 

A large herbivorous mammal is included because empirical relationships of concentrations of 

pesticides in vegetation, discussed below, indicate that grasses may have substantially higher 

pesticide residues than other types of vegetation such as forage crops or fruits (Worksheet B21). 

Grasses are an important part of the diet for some large herbivores, but most small mammals do 

not consume grasses as a substantial proportion of their diet. Thus, even though using residues 

from grass to model exposure for a small mammal is the most conservative approach, it is not 

generally applicable to the assessment of potential adverse effects. Hence, in the exposure 

scenarios for large mammals, the consumption of contaminated range grass is modeled for a 70 

kg herbivore, such as a deer. Caloric requirements for herbivores and the caloric content of 

vegetation are used to estimate food consumption based on data from U.S. EPA/ORD (1993). 

Details of these exposure scenarios are given in Worksheet F10 for acute exposures as well as 

Worksheets F11a and F11b for longer-term exposures. 

For the acute exposures, the assumption is made that the vegetation is sprayed directly – i.e., the 

animal grazes on site – and that100% of the animals diet is contaminated. While appropriately 

conservative for acute exposures, neither of these assumptions are plausible for longer-term 

exposures. Thus, for the longer-term exposure scenarios for the large mammal, two subscenarios are given. The first is an on-site scenario that assumes that a 70 kg herbivore consumes 

short grass for a 90 day period after application of the chemical. In the worksheets, the 

contaminated vegetation is assumed to account for 30% of the diet with a range of 10% to 100% 

of the diet. These are essentially arbitrary assumptions reflecting grazing time at the application 

site by the animal. Because the animal is assumed to be feeding at the application site, drift is set 

to unity - i.e., direct spray. This scenario is detailed in Worksheet 11a. The second sub-scenario 

is similar except the assumption is made that the animal is grazing at distances of 25 to 100 feet 

from the application site (lowering risk) but that the animal consumes 100% of the diet from the 

contaminated area (increasing risk). For this scenario, detailed in Worksheet F12b, AgDRIFT is 

used to estimate deposition on the off-site vegetation. Drift estimates from AgDrift are 

summarized in Worksheet B24 and this model is discussed further in Section 4.2.3.2. 

The consumption of contaminated vegetation is also modeled for a large bird. For these 

exposure scenarios, the consumption of range grass by a 4 kg herbivorous bird, like a Canada 

Goose, is modeled for both acute (Worksheet F12) and chronic exposures (Worksheets F13a and 

F13b). As with the large mammal, the two chronic exposure scenarios involve sub-scenarios for 

on-site as well as off-site exposure. 

For this component of the exposure assessment, the estimated amounts of pesticide residue on 

vegetation are based on the relationship between application rate and residue rates on different 

4-15
 
types of vegetation. As summarized in Worksheet B21, these residue rates are based on 

estimated residue rates from Fletcher et al. (1994). 

Similarly, the consumption of contaminated insects is modeled for a small (10g) bird and a small 

(20g) mammal. No monitoring data have been encountered on the concentrations of 

diflubenzuron in insects after applications of diflubenzuron. The empirical relationships 

recommended by Fletcher et al. (1994) are used as surrogates as detailed in Worksheets F14a and 

F14b. To be conservative, the residue rates from small insects are used – i.e., 45 to 135 ppm per 

lb/ac – rather than the residue rates from large insects – i.e., 7 to 15 ppm per lb/ac. 

A similar set of scenarios is provided for the consumption of small mammals by either a 

predatory mammal (Worksheet F16a) or a predatory bird (Worksheet F16b). Each of these 

scenarios assumes that the small mammal is directly sprayed at the specified application rate and 

the concentration of the compound in the small mammal is taken from the worksheet for direct 

spray of a small mammal under the assumption of 100% absorption (Worksheet F02a). 

In addition to the consumption of contaminated vegetation and insects, diflubenzuron may reach 

ambient water and fish. Thus, a separate exposure scenario is developed for the consumption of 

contaminated fish by a predatory bird in both acute (Worksheet F08) and chronic (Worksheet 

F09) exposures. Because predatory birds usually consume more food per unit body weight than 

do predatory mammals (U.S. EPA 1993, pp. 3-4 to 3-6), separate exposure scenarios for the 

consumption of contaminated fish by predatory mammals are not developed. 

4.2.2.4. Ingestion of Contaminated Water – Estimated concentrations of diflubenzuron in 

water are identical to those used in the human health risk assessment (Worksheet A04). The only 

major differences involve the weight of the animal and the amount of water consumed. There are 

well-established relationships between body weight and water consumption across a wide range 

of mammalian species (e.g., U.S. EPA 1989). Mice, weighing about 0.02 kg, consume 

approximately 0.005 L of water/day (i.e., 0.25 L/kg body weight/day). These values are used in 

the exposure assessment for the small (20 g) mammal. Unlike the human health risk assessment, 

estimates of the variability of water consumption are not available. Thus, for the acute scenario, 

the only factors affecting the variability of the ingested dose estimates include the field dilution 

rates (i.e., the concentration of the chemical in the solution that is spilled) and the amount of 

solution that is spilled. As in the acute exposure scenario for the human health risk assessment, 

the amount of the spilled solution is taken as 200 gallons. In the exposure scenario involving 

contaminated ponds or streams due to contamination by runoff or percolation, the factors that 

affect the variability are the water contamination rate, (see Section 3.2.3.4.2) and the application 

rate. Details regarding these calculations are summarized in Worksheets F06 and Worksheet 

F07. 

4-16
 
4.2.3. Terrestrial Plants 

Terrestrial plants will certainly be exposed to diflubenzuron. A large number of different 

exposure assessments could be made for terrestrial plants – i.e., direct spray, spray drift, runoff, 

wind erosion and the use of contaminated irrigation water. Such exposure assessments are 

typically conducted for herbicides. For diflubenzuron, however, the development of such 

exposure assessments would serve no purpose. As discussed in Section 4.1.2.4 (Hazard 

Identification for Terrestrial Plants), there is no basis for asserting that diflubenzuron will cause 

adverse effects in terrestrial plants. Thus, no formal exposure assessment is conducted for 

terrestrial plants. 

4.2.4. Soil Organisms 

For both soil microorganisms and soil invertebrates, the toxicity data are typically expressed in 

units of soil concentration – i.e., mg agent/kg soil which is equivalent to parts per million (ppm) 

concentrations in soil. The GLEAMS modeling, discussed in Section 3.2.3.4, provides estimates 

of concentration in soil as well as estimates of off-site movement (runoff, sediment, and 

percolation). Based on the GLEAMS modeling, concentrations in clay, loam, and sand over a 

wide range of rainfall rates are summarized in Table 4-2. As indicated in this table, peak soil 

concentrations at an application rate of 70 g/ha are in a relatively narrow range: about 0.003 to 

0.009 mg/kg (ppm) over all soil types and rainfall rates. Longer term concentrations in soil are 

all low and are on the order of 0.00005 to 0.0005 mg/kg – i.e., 0.05 ppb to 0.5 ppb. Modeled 

concentrations of 4-chloroaniline in soil are summarized in Table 4-3. As would be expected of 

any environmental metabolite, peak concentrations are lower than those of the parent compound. 

For 4-chloroaniline these range from about 0.0007 to 0.003 mg/kg, about a factor of three lower 

than the corresponding concentrations of diflubenzuron. 

4.2.5. Aquatic Organisms 

The potential for effects on aquatic species are based on estimated concentrations of 

diflubenzuron and 4-chloroaniline in water that are identical to those used in the human health 

risk assessment. As summarized in Table 3-8, the peak estimated concentration of diflubenzuron 

in ambient water is 0.4 (0.01 to 16) μg/L at an application rate of 70 g/ha. For longer-term 

exposures, the corresponding longer term concentrations in ambient water are estimated at 0.02 

(0.001 to 0.1) μg/L. The corresponding estimates for 4-chloroaniline are summarized in 

Table 3-9: 0.5 (0.00003 to 2) μg/L for acute exposures and 0.05 (0.0002 to 0.2) μg/L for longer 

term exposures. 

4-17
 
4.3. DOSE-RESPONSE ASSESSMENT 

4.3.1. Overview 

As in the human health risk assessment, toxicity values are derived for both diflubenzuron and 4­

chloroaniline. Several of the toxicity values used in the ecological risk assessment for 

diflubenzuron are summarized in Table 4-4. For two groups of organisms, terrestrial arthropods 

and aquatic invertebrates, detailed dose-response assessments can be made for several different 

subgroups. These toxicity values are summarized in Table 4-5 for terrestrial arthropods and 

Table 4-6 for aquatic invertebrates. The values for 4-chloroaniline are summarized in Table 4-7. 

Diflubenzuron is relatively non-toxic to mammals and birds. For mammals, the toxicity values 

used in the ecological risk assessment are identical to those used in the human health risk 

assessments: an acute NOAEL of 1118 mg/kg and a chronic NOAEL of 2 mg/kg/day. A similar 

approach is taken for 4-chloroaniline for which an acute NOAEL of 8 mg/kg is used based on a 

subchronic study and a chronic NOAEL is estimated at 1.25 mg/kg/day based on the chronic 

LOAEL of 12.5 mg/kg/day. For birds, the acute NOAEL for diflubenzuron is taken as 2500 

mg/kg from an acute gavage study and the longer term NOAEL is taken as 110 mg/kg/day from a 

reproduction study. No data are available on toxicity of 4-chloroaniline in birds and the available 

toxicity values for mammals are used as a surrogate. 

For terrestrial invertebrates two general types of data could be used to assess dose-response 

relationships: laboratory toxicity studies and field studies. Field studies are used in the current 

risk assessment because the standard toxicity studies are extremely diverse and many are not 

directly applicable to a risk assessment. Despite the difficulty and uncertainty in interpreting 

some of the field studies, the relatively large number of field studies on diflubenzuron appear to 

present a reasonably coherent pattern that is at least qualitatively consistent with the available 

toxicity data and probably a more realistic basis on which to assess risk to nontarget species. The 

most sensitive species appear to be grasshoppers which may be adversely affected at an 

application rate of 22 g/ha. Somewhat high application rates – in the range of 30 to 35 g/ha – 

will adversely affect macrolepidoptera and some beneficial parasitic wasps. At the maximum 

application rate considered in this risk assessment – i.e., 70 g/ha – some herbivorous insects are 

likely to be affected. No adverse effects in several other groups of insects are expected at this or 

much higher application rates. Honeybees are among the most tolerant species and are not likely 

to be adversely affected at application rates of up to 400 g/ha. 

Invertebrates that do not utilize chitin are also relatively insensitive to diflubenzuron. The 

NOEC for a species of earthworm (Eisenia fetida) is 780 mg/kg soil and is used to represent 

tolerant species of soil invertebrates. Very little information is available on the toxicity of 4­

chloroaniline to terrestrial invertebrates. As with diflubenzuron, the earthworm appears to be 

relatively tolerant to 4-chloroaniline with a reported LC50 value of 540 mg/kg dry soil. The 

toxicity of both diflubenzuron and 4-chloroaniline to soil microorganisms is also relatively low. 

Toxicity values for aquatic species follow a pattern similar to that for terrestrial species: 

arthropods appear to be much more sensitive than fish or non-arthropod invertebrates. For 

4-18
 
diflubenzuron, LC50 values of 25 mg/L to 500 mg/L are used to characterize risks for sensitive 

and tolerant species of fish, respectively. 4-Chloroaniline appears to be more toxic to fish and an 

LC50 value of 2.4 mg/L is used to characterize risks of peak exposures and 0.2 mg/L is used to 

characterize risks of longer term exposures. 

Substantial variability in the response of different groups of aquatic invertebrates to 

diflubenzuron is apparent. Very small arthropods appear to be among the most sensitive species 

– with acute NOEC values in the range of 0.3 to about 1 ppb (μg/L) and chronic NOEC values in 

the range of 0.04 to 0.25 ppb. Based on acute NOEC values, larger arthropods, including crabs 

and larger insects, appear to be more tolerant, with acute NOEC values in the range of 2 to 2000 

ppb. For chronic effects, the differences between small and larger arthropods are less 

remarkable, a stoneflies and mayflies (relatively large insects) having an NOEC value of 0.1 ppb, 

intermediate between Daphnia (0.04 ppb) and Ceriodaphnia (0.25 ppb). Molluscs (invertebrates 

including clams and snails) and worms (oligochaetes) appear to be much less sensitive to 

diflubenzuron. 

The data on the toxicity of 4-chloroaniline to aquatic invertebrates is sparse. An acute NOEC of 

0.013 mg/L is used to characterize acute risks associated with peak exposures in aquatic 

invertebrates and an NOEC of 0.01 mg/L from a reproduction study is used to characterize longer 

term risks to aquatic invertebrates. 

4.3.2. Toxicity to Terrestrial Organisms 

4.3.2.1. Mammals – The dose-response assessment for mammalian wildlife species is based on 

the same set of studies used in the human health risk assessment for diflubenzuron (Section 

3.3.2) and 4-chloroaniline (Section 3.3.3). 

For diflubenzuron, the most sensitive effect in experimental mammals involves toxic effects in 

red blood cells. The NOAEL for this endpoint in experimental mammals is 2 mg/kg/day (U.S. 

EPA 1997a) and is based on a study in which dogs were administered doses of 0, 2, 10, 50, or 

250 mg/kg/day, 7 days/week, for 52 consecutive weeks in gelatin capsules (Greenough et al. 

1985). No adverse effects, including changes in methemoglobin formation, were noted at 2 

mg/kg/day. This dose will be used to characterize longer term risks to mammals. For acute 

exposures, the acute NOAEL of 1118 mg/kg is used. As discussed in Section 3.3.2.2, this is 

based on a study using a petroleum based formulation of diflubenzuron, Dimilin 2L. Because 

none of the estimated exposures approach a level of concern, no elaboration of the dose-response 

assessment is needed. 

A similar approach is taken for 4-chloroaniline. The acute NOAEL is taken as 8 mg/kg. This is 

a very conservative approach – i.e., likely to be overly protective – because this NOAEL is from 

a 90 day study (Scott and Eccleston 1967). The chronic value is based on a LOAEL of 12.5 

mg/kg/day from a 2-year feeding study using rats (NCI 1979). Because a NOAEL was not 

identified in this study, the LOAEL of 12. 5 mg/kg/day is divided by 10 to estimate a chronic 

4-19
 
NOAEL of 1.25 mg/kg/day. This is essentially the same estimate used by U.S. EPA (1997a) in 

the derivation of the RfD based on the LOAEL of 12.5 mg/kg/day (Section 3.3.3.1). 

4.3.2.2. Birds 

4.3.2.2.1. Diflubenzuron – There appears to be relatively little difference in the acute 

toxicity of diflubenzuron to birds and mammals. As summarized above, the lowest acute 

NOAEL for mammals is 1118 mg/kg (rats dosed with Dimilin 2L in the study by Blaszcak 

(1997a). For birds, the lowest acute NOAEL is 2500 mg/kg from the study by Alsager and Cook 

(1975) in red-winged blackbirds. As detailed in Appendix 1 for mammals and Appendix 8 for 

birds, higher NOAEL values have been reported in other studies – i.e., up to 10,000 mg/kg for 

mammals (rats and mice in the study by Koopman 1977) and 5,000 mg/kg for birds (mallard 

ducks in the study by Roberts and Parke 1976). Analogous to the approach taken with rats, the 

lowest NOAEL is taken as the toxicity value for acute exposures in bird – i.e., the NOAEL of 

2500 mg/kg in red-winged blackbirds from the study by Alsager and Cook 1975. 

It should be noted that the variability in the acute NOAEL values does not imply any systematic 

differences among species but simply reflects the highest dose tested in the different experiments. 

Thus, the use of the lowest NOAEL rather than the highest NOAEL may be viewed as somewhat 

conservative. As discussed in Section 4.3.2.1, the use of the 1118 mg/kg dose for mammals is 

justified based on the use of a petroleum based formulation in the study by Blaszcak (1997a). 

The use of the lowest NOAEL for birds based on the conservative assumption that somewhat 

higher doses in the study by Alsager and Cook (1975) could have resulted in effects. 

Notwithstanding this assumption, the data are not sufficient to derive separate NOAEL values for 

tolerant and sensitive species because none of the available data actually demonstrated 

differences in sensitivity – i.e., differences in LOAEL values. 

In terms of chronic toxicity, however, birds appear to be somewhat more tolerant to 

diflubenzuron than mammals. Based on reproduction studies, the NOEC for reproductive 

toxicity in birds is greater than 500 ppm – i.e., at the highest dietary concentration, no effects 

were noted – in mallard ducks (Beavers et al. 1990a) and bobwhite quail (Beavers et al. 1990b). 

Based on differences in food consumption (Appendix 4), the lowest dose in terms of mg/kg 

bw/day is 110 mg/kg/day from the study in quail (Beavers et al. 1990b). This is substantially 

above for the mammalian NOAEL of 2 mg/kg/day and the corresponding mammalian LOAEL of 

10 mg/kg/day. While this suggests a difference in sensitivity between mammals and birds, the 

toxicity endpoints are different – i.e., effects on blood from chronic exposure in mammals and 

reproductive effects in birds. As noted in Appendix 1, doses as high as about 4000 mg/kg/day 

were not associated with reproductive effects in rats (Brooker 1995). In any event, the chronic 

NOAEL of 110 mg/kg/day in quail from the study by Beavers et al. (1990b) is used to 

characterize the risks associated with longer term exposures of birds to diflubenzuron. 

4.3.2.2.2. 4-Chloroaniline – No data have been encountered on the toxicity of 

4-chloroaniline to birds. For the current risk assessment, the toxicity values for 4-chloroaniline 

4-20
 
in mammals are used as surrogates for birds. This adds uncertainty to the risk assessment for 

birds and this is discussed further in Section 4.4 (Risk Characterization). 

4.3.2.3. Terrestrial Invertebrates 

4.3.2.3.1. Diflubenzuron – Two general types of data could be used to assess doseresponse relationships for terrestrial invertebrates: laboratory toxicity studies (Appendix 5) and 

field studies (Appendix 3a). In most risk assessments conducted by U.S. EPA (e.g. U.S. 

EPA/OPP 1997a) as well as risk assessments conducted for the USDA/Forest Service, doseresponse assessments for terrestrial invertebrates are based on controlled laboratory studies that 

are commonly conducted on the honey bee using relatively standard protocols. As indicated in 

Table 4-5, a different approach is used in the current risk assessment: the large number of field 

studies on diflubenzuron that report either effect or no effect levels are used directly for 

characterizing risk with exposures expressed in units of application rate. 

One reason for this approach involves the disparity in experimental designs among the toxicity 

studies that are available which confounds quantitative comparisons of relative sensitivities 

among species. As discussed in Section 4.1.2.3, there is an apparently wide range of 

sensitivities to diflubenzuron among different invertebrate species. Based on standard toxicity 

tests, the honey bee is among the more tolerant species. The U.S. EPA used an LD50 of greater 

than 30 μg/bee to classify diflubenzuron as practically non-toxic to the honey bee. Taking an 

average weight of 0.093 g/bee or 0.000093 kg/bee (USDA/APHIS 1993) and making the very 

conservative assumption of 100% absorption, this would correspond to an LD50 greater than 322 

mg/kg bw [0.03 mg/bee ÷ 0.000093 kg bw/bee = 322.58 mg/kg]. As summarized in Appendix 

5, somewhat lower LD50 values have been reported by Chandel and Gupta (1992) – i.e., about 22 

mg/kg for pupae and 53 mg/kg for third instar larvae. The gypsy moth is obviously a sensitive 

species, with a topical LD50 value of about 4 to 9 mg/kg, based on residues on vegetation (Berry 

et al. 1993), about a factor of 2 to 5 below the lowest LD50 value for the honey bee. A similar 

topical LD50 of 1.07 mg/kg has been reported by Sinha et al. (1990) for the butterfly, Pieris 

brassicae. Somewhat lower LD50 values have been reported for an orthopteran – i.e., 0.31 mg/kg 

in Oxya japonica from the study by Lim and Lee (1982). Based on topical LD50 values, the most 

sensitive species appears to be lacewing, Chysoperla carnea, with a reported topical LD50 values 

of 2.26 ng/insect or about 0.00226 μg/insect (Medina et al. 2003). Based on a mean body weight 

of 7.53 mg reported by Medina et al. (2003), this corresponds to a dose of 0.0003 μg/mg, which 

in turn corresponds to a dose of 0.0003 mg/g or 0.0000003 mg/kg bw. Thus, based on this LD 50, 

the lacewing would appear to be more sensitive than the gypsy moth by a factor of 13 to 30 

million [4 to 9 mg/kg ÷ 0.0000003 mg/kg]. The LD50 value from Medina et al. (2003), however, 

is not really comparable to the value for the gypsy moth because the topical application to the 

lacewing involved direct application of diflubenzuron (in acetone) rather than a spray or contact 

with a contaminated surface. Thus, while the various laboratory toxicity studies could be used to 

construct a standard dose-response assessment for tolerant and sensitive species, there would be 

substantial uncertainty in the comparisons because of the diversity in experimental designs. 

4-21
 
An alternative approach may be based on the available field studies. A summary of these studies 

is presented in Table 4-1 and additional details are provided in Appendix 3a. Field studies, like 

epidemiology studies, can be difficult to interpret because of differences in the treated site versus 

the control site. For example, the study by Van Den Berg (1986) on mites and collembolans is 

noted in Table 4-1 as providing a NOAEL in which transient or equivocal effects were noted. As 

detailed in Appendix 3a, Van Den Berg (1986) concluded that the effects on the mites and 

collembolans were insubstantial. The data, however, indicate generally fewer species over time 

in the treated site versus the untreated site. The author’s conclusion that the effects were 

insubstantial is based on the fact that the populations of mites and collembolans were different at 

the control and treated sites prior to treatment and that the capture patterns over time for mites 

were highly erratic. In other words, compared to pre-treatment populations as well as the time 

course of population changes, the effect of diflubenzuron in this study appeared to be marginal 

and insubstantial. An examination of the data presented by Van Den Berg (1986) supports the 

conclusion that the application of diflubenzuron in this study should be classified as a NOAEL. 

A similar assessment may be made of the study by Martinat et al. (1993) in which changes in 

populations of spiders and orthopteroids (i.e., cockroaches, mantises, locusts, and crickets) were 

only sporadically noted over time and no consistent effect is apparent. 

Despite the difficulty and uncertainty in interpreting some of the fields, the relatively large 

number of field studies on diflubenzuron appear to present a reasonably coherent pattern that is at 

least qualitatively consistent with the available toxicity data and probably a more realistic basis 

on which to assess risk to nontarget species. Consistent with the laboratory studies, the field 

studies clearly indicate that honey bees are relatively insensitive to diflubenzuron: application 

rates of up to 400 g/ha are not likely to affect honeybees (Table 4-5). The most sensitive species 

appear to be grasshoppers which may be adversely affected at an application rate of 22 g/ha. 

Somewhat high application rates – in the range of 30 to 35 g/ha – will adversely effect 

macrolepidoptera and some beneficial parasitic wasps. At the maximum application rate of 

considered in this risk assessment – i.e., 70 g/ha – some herbivorous insects are likely to be 

affected. No adverse effects in several other groups of insects are expected at this or much 

higher application rates, as detailed in Table 4-5. 

As also noted in Section 4.1.2.3, invertebrates that do not utilize chitin are relatively insensitive 

to diflubenzuron. Based on soil toxicity studies, the NOEC 780 mg/kg soil for the earthworm 

(Eisenia fetida) from the study by Berends et al. (1992) is used to represent tolerant species of 

soil invertebrates. 

4.3.2.3.2. 4-Chloroaniline – Very little information is available on the toxicity of 4­

chloroaniline to terrestrial invertebrates (WHO 2003). This is not uncommon for compounds 

that are not used or registered as insecticides. WHO (2003) summarizes a standard OECD study 

on earthworms in which the 28-day LC50 value was 540 mg/kg dry soil. As noted in Section 3.2, 

this is far higher than any concentrations of 4-chloroaniline that are likely to be found in soil. 

4-22
 
4.3.2.4. Terrestrial Plants (Macrophytes) – As discussed in 4.1.2.4 (Hazard Identification for 

Terrestrial Plants), no toxicity studies have been conducted on terrestrial plants and there is no 

basis for asserting that adverse effects on terrestrial plants are likely from exposures to either 

diflubenzuron or 4-chloroaniline. Consequently, no dose-response assessments for terrestrial 

plants are presented in this risk assessment. 

4.3.2.5. Soil Microorganisms 

4.3.2.5.1. Diflubenzuron – Diflubenzuron does not appear to be very toxic to soil 

microorganisms (Section 4.1.2.5). While one study (Sexstone 1995) has noted transient changes 

in gross microbial biomass and activity at one exposure rate (roughly equivalent to 10 g/ha), no 

dose-response relationship is demonstrated and the effects, if any, appear to be very minor. 

Consequently, this study is not used quantitatively in the dose-response assessment for soil 

microorganisms. For the current risk assessment, bioassays on fungi are used to identify tolerant 

and sensitive species – a LOEC of 50 ppm in Pythium for sensitive species and an NOEC of 100 

ppm for tolerant species (Aspergillus, Fusarium, Rhizopus, Trichoderma) from the study by 

(Townshend et al. 1983). If any species of microorganisms are at risk from exposure to 

diflubenzuron, fungi might be considered the most likely to be susceptible because some fungi 

utilize chitin in their cell walls. As summarized in Table 4-2, however, the NOEC and LOEC 

values are several orders of magnitude higher than any plausible soil exposures. 

4.3.2.5.2. 4-Chloroaniline – The only information encountered on the microbial toxicity 

of 4-chloroaniline is an ED10 of 1000 ppm for Fe(III) reductions by upper soil (Horizon A) 

microorganisms (Welp and Brummer 1999). As with diflubenzuron, this concentration is far 

above plausible levels of soil exposure. 

4.3.3. Aquatic Organisms 

4.3.3.1. Fish 

4.3.3.1.1. Diflubenzuron – The toxicity data on diflubenzuron are sufficient to identify 

sensitive and tolerant species for both acute and chronic exposures (Table 4-4). For acute 

toxicity, the lowest and highest LC50 values will be used consistent with the data in the risk 

assessment presented by U.S. EPA/OPP (1997a). The LC50 value for sensitive fish species will 

be taken as 25 mg/L from the study by Johnson and Finley (1980) in yellow perch and the LC50 

value for tolerant fish species will be taken as 500 mg/L from the study by Reiner and Parke 

(1975) in fathead minnow. Both of these are very protective values in that both concentrations 

are actually the highest concentration tested and less than 50% mortality was observed. As 

discussed further in Section 4.4, this protective approach has no impact on the risk assessment 

because the anticipated peak exposures to diflubenzuron are far below these concentrations. For 

longer term exposures, reproductive NOEC values will be used. The range of reported values is 

relatively narrow: 0.05 mg/L for mummichogs from the study by Livingston and Koenig (1977) 

to 0.1 mg/L for fathead minnows from the study by Cannon and Krize (1976). 

4.3.3.1.2. 4-Chloroaniline – Very little information is available on the toxicity of 4­

chloroaniline to fish. As reviewed by WHO (2003), an LC50 value of 2.4 mg/L is reported in 

4-23
 
bluegills and a reproductive NOEC of 0.2 mg/L in zebra fish is reported in Bresch et al. (1990). 

These values are used in the current risk assessment for characterizing risks to fish associated 

with exposures to 4-chloroaniline (Table 4-7). 

4.3.3.2. Amphibians – The only information on the toxicity of diflubenzuron to amphibians 

comes from two field studies conducted by Pauley (1995a,b). As discussed in Section 4.1.3.2, 

these studies indicate a change in the diet of both terrestrial and aquatic salamanders following an 

application of diflubenzuron at 80g/ha. This change was secondary to changes in available food 

items. No data are available on the toxicity of 4-chloroaniline to amphibians. Because of the 

very low apparent risks to fish (Section 4.4), the limited data on effects of diflubenzuron to 

amphibians, and the lack of data on the effects of 4-chloroaniline to amphibians, a quantitative 

dose-response assessment for this group of organisms is not proposed. 

4.3.3.3. Invertebrates 

4.3.3.3.1. Diflubenzuron – The toxicity values used in this risk assessment for aquatic 

invertebrates are summarized in Table 4-6, with the top section of this table summarizing acute 

toxicity values that are used to characterize risks associated with peak exposures and the bottom 

section of the table summarizing toxicity values used to characterize risks associated with longer 

term exposures. In all cases, the toxicity values are based on no-observed-effect concentrations 

(NOECs). This approach is somewhat different from the approach taken by U.S. EPA (1997a), 

in which toxicity values are based on LC50 values but the studies used and basic conclusions of 

the current risk assessment are similar to those of U.S. EPA (1997a). Diflubenzuron is very 

highly toxic to some aquatic invertebrates. 

As with the acute toxicity to terrestrial invertebrates, the dose-response assessment can be 

elaborated to include several groups of invertebrates rather than simply sensitive and tolerant 

species. Supporting information for the acute and chronic toxicity values are given in Table 4-8 

and Table 4-9, respectively, and additional information from field studies is summarized in 

Table 4-10. More detailed summarizes of the acute and chronic toxicity studies are given in 

Appendix 7 and details of a large number of field studies are given in Appendix 3b. 

As summarized in Table 4-6, there is a substantial variability in the response of different groups 

of aquatic invertebrates to diflubenzuron. Very small arthropods – i.e, cladocerans (Daphnia and 

Ceriodaphnia) as well as copepods – appear to be among the most sensitive aquatic species – 

with acute NOEC values in the range of 0.3 to about 1 ppb (μg/L) and chronic NOEC values in 

the range of 0.04 to 0.25 ppb. Based on acute NOEC values, larger arthropods, including crabs 

and larger insects, appear to be more tolerant, with acute NOEC values in the range of 2 to 2000 

ppb. In some of these assays of larger invertebrates, the short duration of the assay may be a 

factor in the apparently greater tolerance of larger invertebrates compared to small invertebrates. 

For example, Lahr et al. (2001) note that the backswimmers tested in their bioassay evidenced a 

NOEC of 2000 ppb but that lower NOEC values could have been evident if the organisms had 

been in a molting stage. This supposition is supported by chronic toxicity data (Table 4-9) in 

which differences between small and larger arthropods are less remarkable, with stoneflies and 

4-24
 
mayflies (relatively large insects) having an NOEC value of 0.1 ppb, intermediate between 

Daphnia (0.04 ppb) and Ceriodaphnia (0.25 ppb). In the tests using stonefly and mayflies, 

response was characterized as an inhibition of emergence rather than pre-emergent mortality. 

Again, this probably relates to the inhibition of chitin synthesis by diflubenzuron. Molluscs 

(invertebrates including clams and snails) and worms (oligochaetes) appear to be much less 

sensitive to diflubenzuron. 

Based on acute NOEC values, the range of sensitivities among aquatic invertebrates appears to 

span a factor of over 400,000 [125,000 ppb in molluscs ÷ 0.3 in Daphnia = 416,667] based on 

acute NOEC values and a factor of 8,000 [320 ppb in molluscs ÷ 0.04 in Daphnia] based on 

longer term NOEC values. These ratios are, at least to some extent, artifacts of experimental 

design. As summarized in Tables 4-8 and 4-9, acute and chronic NOEC and LOEC values are 

available for sensitive species such as daphnids. For molluscs, however, only NOEC values are 

available – i.e., no effects have been demonstrated in these species at the highest concentration 

tested. 

Although there is a large number of field studies available on effects of diflubenzuron on aquatic 

invertebrates (Appendix 3b), these studies are not directly used in the dose-response assessments. 

Unlike the case with terrestrial invertebrates, application rates (e.g., g/ha) in aquatic field studies 

do not provide a uniform basis for comparing exposures among the different studies because the 

amount of diflubenzuron entering the water may and probably did vary remarkably among the 

different field studies based on site-specific and meteorological differences among the studies. 

The magnitude of possible differences is illustrated in Tables 3-2 and 3-3. 

Nonetheless, some studies provide information on both application and concentrations in 

ambient water. An overview of these studies, summarized from Appendix 3b, is given in 

Table 4-10. As in the tables for standard toxicity studies, Tables 4-8 and 4-9, concentrations are 

given in braces [] between the species and the citation. Even these concentrations, however, are 

not readily comparable among studies, with some reported as peak concentrations and others as 

nominal or average concentrations over a given period. For example, Apperson et al. (1977) 

conducted a field study in which populations of cladocerans and copepods declined after an 

application of diflubenzuron to ponds and lakes at nominal concentrations of 2.5, 5, and 10 ppb. 

Actual monitored concentrations peaked at up to 32.2 ppb, however, and declined rapidly to less 

than 1 ppb. This type of pattern is typical in field studies in which concentrations will vary 

substantially both among different studies as well as over time within a single study. This 

probably accounts for the general pattern of field studies suggesting a higher tolerance in terms of 

reported concentrations than laboratory studies in which concentrations are better defined and 

less variable. The field studies summarized in Table 4-10, however, do support the general 

pattern of species sensitivity noted in the laboratory toxicity studies – i.e., small arthropods are 

more sensitive than larger arthropods and non-arthropod invertebrates. 

Notwithstanding the limitations inherent in field studies in terms of actual exposures and 

temporal variations, the field studies are directly useful in risk characterization and are discussed 

4-25
 
 

further in Section 4.4. One very important feature of field studies is ability to assess population 

recovery, which is not typically assayed in laboratory studies. As summarized in Table 4-10, 

most field studies that detect adverse effects also find evidence of population recovery after 

application so long as the duration of the study is sufficiently long to permit the detection of 

recovery. This is also discussed further in the risk characterization (Section 4.4). 

4.3.3.3.2. 4-Chloroaniline – The data on the toxicity of 4-chloroaniline to aquatic 

invertebrates is sparse, particularly when compared to the very rich data base on diflubenzuron. 

Notwithstanding this limitation, 4-chloroaniline appears to be much less toxic to aquatic 

invertebrates than diflubenzuron and the magnitude of the difference in potency can be 

quantified. In terms of acute toxicity to Daphnia magna, the 48-hour LC50 value for 4­

chloroaniline has been reported as 0.31 mg/L (Kuhn et al 1989a), 400 times higher than the LC50 

values of 0.0007 mg/L to 0.00075 mg/L for diflubenzuron (Corry et al. 1995; Kuijpers 1988; 

Majori et al. 1984). The corresponding NOEC for 4-chloroaniline is 0.013 mg/L (Kuhn et al 

1989a), 40 times higher than the acute NOEC of 0.0003 mg/L for diflubenzuron (Corry et al. 

1995). 

Similarly, the chronic NOEC in Daphnia magna for 4-chloroaniline in a standard reproduction 

study is 0.01 mg/L (Kuhn et al 1989b). This is a factor of 250 times higher than the 

corresponding value of 0.00004 mg/L in Daphnia magna reported by Surprenant (1988). 

As summarized in Table 4-7 (toxicity values for 4-chloroaniline), the acute NOEC of 0.013 mg/L 

(Kuhn et al 1989a) is used to characterize acute risks to aquatic invertebrates and the NOEC of 

0.01 mg/L for reproductive effects (Kuhn et al 1989b) is used to characterize longer term risks to 

aquatic invertebrates. 

4.3.3.4. Aquatic Plants 

4.3.3.4.1. Diflubenzuron – Compared to aquatic invertebrates, relatively little 

information is available on the toxicity of diflubenzuron to aquatic plants (Section 4.1.3.4 and 

Appendix 8). The lowest reported effect is a decrease in periphyton at a concentration 7.0 μg/L 

in littoral enclosures (Moffett 1995). As noted in Section 4.1.3.4 and Appendix 8, Moffett 

(1995) attributed this change to a decrease in the population density of zooplankton grazers. This 

conclusion seems reasonable and is supported by standard plant toxicity studies reporting no 

effects at concentrations of up to 380 μg/L (Booth and Ferrell 1977; Thompson and Swigert 

1993a,b,c). For assessing the risks of direct toxic effects on terrestrial plants, a NOEC of 45 

μg/L will be used for possibly sensitive species (Selenastrum capricornutum in the study by 

Hansen and Garton 1982a) and a NOEC of 380 μg/L (Navicula pelliculosa in the study by 

Thompson and Swigert 1993c) will be used for apparently tolerant species. Since no LOEC 

values are available for any species of aquatic plants, these different NOEC values may simply 

reflect differences in the highest dose tested in the respective experiments rather than true 

differences in species sensitivity to diflubenzuron. 

4-26
 
 

4.3.3.4.2. 4-Chloroaniline – The only information encountered on the toxicity of 

4-chloroaniline is summarized in WHO (2003) from two publications in the German literature 

(Schmidt 1989; Schmidt and Schnabl 1988). Based on this information, 4-chloroaniline appears 

to be somewhat more toxic to aquatic plants than diflubenzuron. While WHO (2003) does not 

report NOEC values for 4-chloroaniline, an EC10 of 0.02 mg/L for cell multiplication in 

Scenedesmus subspicatus, a species of green algae, will be used as surrogate NOEC. 

4.3.3.5. Microorganisms (excluding algae) 

4.3.3.5.1. Diflubenzuron – Very little information is available on the toxicity of either 

diflubenzuron or 4-chloroaniline to aquatic microorganisms. As summarized in Section 4.1.3.5, 

marginal and transient effects on microbial decomposition and respiration have been noted at 50 

μg/L and 50,000 μg/L (Kreutzweiser et al. 2001). Because of the insubstantial nature of the 

effects and the lack of a marked dose-response relationship, the concentration of 50 μg/L is used 

as a NOEC for aquatic microorganisms in Table 4-4. 

4.3.3.5.2. 4-Chloroaniline – The only information on 4-chloroaniline is the results of a 

assay for bioluminescence with Photobacterium phosphoreum in which the 30-minute EC50 for 

the inhibition of bioluminescence was 5.1 mg/L (Ribo and Kaiser 1984). While the utility of this 

type of assay for risk characterization may be marginal, it is the only information available and is 

included in Table 4-7 and used for the risk characterization of 4-chloroaniline. 

4-27
 
4.4. RISK CHARACTERIZATION 

4.4.1. Overview 

While the data base supporting the ecological risk assessment of diflubenzuron is large and 

complex, the risk characterization is relatively simple. Diflubenzuron is an effective insecticide 

and effects on some nontarget terrestrial insects are likely at application rates that are used to 

control the gypsy moth. Species at greatest risk include grasshoppers, various macrolepidoptera 

(including the gypsy moth), other herbivorous insects, and some beneficial predators of the gypsy 

moth. These species are at risk because of the mode of action of diflubenzuron (i.e., inhibition of 

chitin) and the behavior of the sensitive insects (the consumption of contaminated vegetation or 

predation on the gypsy moth). Some aquatic invertebrates may also be at risk but the risks 

appear to be less than risks to terrestrial insects. The risk characterization for aquatic 

invertebrates is highly dependant on site-specific conditions. In areas in which water 

contamination is likely to be minimal, no or only marginal effects are expected. During 

applications in which drift or direct deposition is not controlled well or in areas in which soil 

losses from runoff and sediment are likely, acute effects on some aquatic invertebrates are 

plausible and longer term effects on sensitive species could occur. 

Direct effects of diflubenzuron on other groups of organisms – i.e., mammals, birds, amphibians, 

fish, terrestrial and aquatic plants, microorganisms, and non-arthropod invertebrates – do not 

appear to be plausible. Secondary effects in some nontarget species could occur. The most 

common secondary effects will be seen in and associated with animals that consume either the 

the gypsy moth or other invertebrates that may be adversely affected by diflubenzuron. The most 

common secondary effect will be a change in prey items that are consumed. Changes in feeding 

territory and prey items as well as reductions in body fat are likely to be transient. 

There is no indication that 4-chloroaniline formed from the degradation of diflubenzuron will 

have an adverse effects on any species. 

4.4.2. Terrestrial Organisms 

4.4.2.1. Terrestrial Vertebrates – The risk characterizations for terrestrial vertebrates are 

essentially identical for both diflubenzuron and 4-chloroaniline. At the highest application rate 

of diflubenzuron that would be used in USDA programs, risks to mammals and birds are far 

below a level of concern. The quantitative risk characterization for terrestrial vertebrates 

(mammals and birds) is summarized in Worksheet G02a in the diflubenzuron worksheets 

(Supplement 1) and Worksheet G02 in the 4-chloroaniline worksheets (Supplement 2). The 

risk characterization is based on the estimates of exposure summarized in Section 4.2.3 and the 

toxicity values for diflubenzuron (Table 4-4) and 4-chloroaniline (Table 4-7) that were derived in 

Section 4.3.2. 

The highest hazard quotient (HQ) for diflubenzuron is 0.2, the value associated with the upper 

range of exposure from the longer term consumption of contaminated vegetation in the treated 

area by a large mammal. As discussed in Section 4.2.2, this exposure scenario is based on the 

consumption of contaminated grass by a large mammal. For the gypsy moth program, this is an 

4-28
 
extremely conservative scenario in that most large wildlife mammals will not consume grass as 

an exclusive or even predominant proportion of their diet (exceptions being elk and some 

livestock animals). In addition, this scenario assumes that the grass is directly sprayed. In the 

application of diflubenzuron, canopy interception would reduce residues on grass in most 

circumstances. Other hazard quotients for diflubenzuron are below a level of concern by factors 

of 50 (the upper range HQ of 0.02 for the consumption of contaminated fish by a predatory bird) 

to 1 in one billion (the lower range HQ for the consumption of contaminated water by a small 

mammal). 

The highest risk quotient for chloroaniline is 0.02, associated with the consumption of 

contaminated water by a small mammal. As discussed in Section 3.2.3.4, these peak exposures 

may occur months after the application of diflubenzuron and the concentrations of 

4-chloroaniline in water are likely to vary substantially with different soils as well as rainfall 

rates. The peak concentrations of 4-chloroaniline are based on very conservative and perhaps 

extreme assumptions and the very low of hazard quotient of 0.02 – i.e., below the level of 

concern by a factor of 50 – indicates that there is no plausible basis for asserting that such 

exposures would be hazardous. 

This risk characterization for terrestrial vertebrates is consistent with the risk characterization by 

U.S. EPA (1997a) as well as field studies which indicate a lack of adverse effects on terrestrial 

vertebrates after applications of diflubenzuron (Sections 4.1.2.1 and 4.1.2.2. and Appendix 3a). 

No toxic effects are likely to be seen in mammals or birds. 

The most common secondary effects will be seen in and associated with vertebrates that consume 

either the target species (the gypsy moth) or other invertebrates that may be adversely affected by 

diflubenzuron (see Section 4.4.2.2.1). For such vertebrates, the most common secondary effect 

will be a change in prey items that are consumed. 

4.4.2.2. Terrestrial Invertebrates 

4.4.2.2.1. Diflubenzuron – While risks to terrestrial vertebrates are implausible, risks to 

some terrestrial invertebrates are virtually certain (Worksheet G02b, Supplement 1). At an 

application rate of 70 g/ha, adverse effects – i.e., mortality and decreases in populations – have 

been demonstrated in field studies for grasshoppers, various macrolepidoptera (including the 

gypsy moth), some mandibulate herbivores, and some beneficial predators to the gypsy moth. 

Effects on some beneficial predators may be secondary but at least in one species, Apanteles 

melanoscelus, a wasp that is a parasite on the gypsy moth, the effect appears to be due to direct 

toxicity (Madrid and Stewart1981). Effects in the same species are likely to be seen at lower 

application rates that may be used in USDA programs – i.e., 35 g/ha. For effects in these 

sensitive groups to be avoided, the application rate would need to be below about 2 g/ha [70 g/ha 

from Worksheet G02b divided by the HQ of 32 for the grasshopper]. This damage to non-target 

species appears to be unavoidable given the mode of action of diflubenzuron (i.e., inhibition of 

chitin) and the behavior of the sensitive insects (the consumption of contaminated vegetation or 

predation on the gypsy moth). 

4-29
 
Most other insect groups are not likely to be affected at least directly. Some secondary effects 

associated with changes in available prey may be noted. As with most secondary effects, the 

changes in habitat or prey items are likely to be reversible. In other words, changes will be 

transient and populations will generally recover (e.g., Catangui et al. 1996). 

4.4.2.2.2. 4-Chloroaniline – Very little information is available on the toxicity of 

4-chloroaniline to invertebrates. One bioassay in earthworms reports an LC50 value of 540 mg/kg 

soil. The maximum concentration of 4-chloroaniline in soil is estimated at 0.0026 ppm 

-6 (Table 4-3). The resulting HQ is 4.8×10 , below the level of concern by over 200,000. No data 

are available on the toxicity of 4-chloroaniline to other terrestrial vertebrates and risks cannot be 

quantified. Given the relatively low risks of 4-chloroaniline in aquatic invertebrates (4.4.3.2.2) 

as well as other organisms, there is no basis for asserting that substantial risks are plausible, 

particularly when compared to clear risks associated with diflubenzuron. 

4.4.2.3. Terrestrial Plants and Microorganisms – No quantitative risk assessment to terrestrial 

plants is made for either diflubenzuron or 4-chloroaniline. As discussed in Section 4.1.2.4, there 

are no data on the phytotoxicity of either compound. This lack of data, however, adds no 

substantial uncertainty to this risk assessment. Diflubenzuron has been extensively tested in both 

the laboratory and field studies for efficacy in the protection of terrestrial plants from insect 

pests. If diflubenzuron were toxic to plants at applications at or substantially above those used to 

control the gypsy moth, it is likely that reports of such phytotoxicity would be noted. No such 

reports have been encountered (Appendix 3a and Appendix 8). 

Limited information is available on the toxicity of diflubenzuron and 4-chloroaniline to soil 

microorganisms. As summarized in Worksheet G02b for diflubenzuron (Supplement 1), 

exposures of soil microorganisms to diflubenzuron are likely to be below a level of concern for 

sensitive species by a factor of over 600 at the upper range of plausible exposure – i.e., an HQ of 

0.0016. For 4-chloroaniline, the toxicity value for microorganisms is 1000 ppm. As noted 

above, the highest estimated peak concentration of 4-chloroaniline in soil is 0.0026 ppm (Table 

-6 4-3). The resulting HQ is 2.6×10 , below the level of concern by over 350,000.

4.4.3. Aquatic Organisms 

4.4.3.1. Aquatic Vertebrates – As with terrestrial vertebrates, the risk assessment for fish is 

unequivocal. There is no indication that diflubenzuron or 4-chloroaniline associated with the 

degradation of diflubenzuron will approach a level of concern. 

The highest hazard quotient for diflubenzuron is 0.002 – i.e., longer term exposures to sensitive 

fish species (Worksheet G03b in Supplement 1). This is below the level of concern by a factor of 

500. The toxicity of diflubenzuron has been assayed in relatively few fish species and it is likely 

that the most sensitive species of fish has not been identified. Nonetheless, there is no basis for 

asserting that species variability will encompass the factor of 500 associated with the highest HQ 

for diflubenzuron. 

4-30
 
The risk characterization for 4-chloroaniline is virtually identical. The highest hazard quotient is 

0.001. Below the level of concern by a factor of 1000 (Worksheet G03, Supplement 2). 

4.4.3.2. Aquatic Invertebrates 

4.4.3.2.1. Diflubenzuron – As noted by U.S. EPA (1997a), risks to aquatic invertebrates 

in some applications of diflubenzuron may be substantial – i.e., direct applications to standing 

bodies of water for mosquito control and forestry uses involving direct applications to bogs, 

swamps or other standing bodies of water (U.S. EPA 1997a, p. 64). These types of applications, 

however, are not used in and are thus not relevant to USDA programs for the control of the gypsy 

moth. 

In USDA programs for control of the gypsy moth, risks to aquatic invertebrates appears to be 

substantially less than risks to terrestrial invertebrates. As noted in Section 2.3, USDA will use a 

100 to 500 foot buffer between the application site of diflubenzuron and bodies of open water. 

While it is possible that small streams could be over-sprayed in aerial applications if the stream is 

not visible from the air, the covering foliar canopy would intercept some of the diflubenzuron 

which would in turn reduce the initial concentrations in stream water. 

Based on the exposure assessments conducted in this risk assessment, which are consistent with 

several other exposure assessments as well as a number of relevant monitoring studies 

(Table 3-7), only the most sensitive species of aquatic invertebrates are likely to be adversely 

affected based on central estimates of plausible peak exposures. The central estimate of the 

hazard quotient for sensitive daphnids is only 1.3 (Worksheet G03a, Supplement 1). Typically, 

hazard quotients are rounded to a single significant digit. Thus, this hazard quotient reaches but 

does not exceed a level of concern. Based on central estimates of longer term exposures, all 

hazard quotients are less than 1 (Worksheets G03b, Supplement 1). 

At the upper ranges of plausible peak exposures, the level of concern is reached for crabs 

(HQ=1), modestly exceeded for Ceriodaphnia and copepods (HQ=2), and exceeded by a factor 

of 5 for Daphnia. For Daphnia, LC50 values are only modestly above the NOEC (Table 4-8) and 

substantial mortality in these species would be plausible. At the upper range of longer term 

exposures, the hazard quotient exceeds a value of 1 only for Daphnia – i.e., HQ=3. This is in the 

range in which longer term effects on Daphnia productivity would be expected and such effects 

have been observed in field studies (Ali and Mulla 1978b). 

Thus, based on the available toxicity data and dose response assessment, the risk characterization 

for aquatic invertebrates is highly dependant on site-specific conditions. In areas in which water 

contamination is likely to be minimal – i.e., areas with relatively low rainfall and areas in which 

drift can be controlled and runoff is limited – it is likely that no or only minimal effects would be 

observed (e.g., the field study by Ali et al. 1988). During applications in which drift or direct 

deposition is not controlled well or in areas in which soil losses from runoff and sediment are 

likely, acute effects on some aquatic invertebrates are plausible and longer term effects on 

sensitive species could occur. 

4-31
 
That any of these effects would result in substantial secondary effects does not seem likely. A 

large number of field studies are available on diflubenzuron (Appendix 3b) that indicate direct 

effects on several species of invertebrates at concentrations in water that are above those that 

would be encountered in many applications for the control of the gypsy moth (see Section 4.1.3.3 

for discussion). In addition, the only studies that suggest substantial secondary effects – such as 

decreased growth in fish – are litoral enclosure studies (Moffett and Tanner 1995; Tanner and 

Moffett 1995) in which fish were limited in their ability to seek prey. None of the field studies 

involving free-ranging fish have reported secondary effects other than a change in prey that are 

consumed. 

4.4.3.2.2. 4-Chloroaniline – The risks to aquatic invertebrates associated with 

4-chloroaniline are insubstantial relative to the risks associated with diflubenzuron. The highest 

hazard quotient is 0.2, associated with peak exposures to 4-chloroaniline in water. 

4.4.3.3. Aquatic Plants and Microorganisms – Risks to aquatic plants and microorganisms 

appear to be low. There is essentially no identifiable risk associated with diflubenzuron. The 

highest hazard quotient is 0.04 and is associated with peak exposures to sensitive aquatic plants 

(Worksheet G03a, Supplement 1). Peak risks associated with 4-chloroaniline are somewhat 

higher, 0.2, the HQ associated with peak exposures to aquatic plants (Worksheet G03, 

Supplement 2). 

A more plausible risk to aquatic plants may involve secondary effects – increased algal 

populations – associated with mortality in aquatic grazers such as Cladocerans. This effect has 

been noted in the mesocosm study by Boyle et al. (1996) . Apperson et al. (1977) noted a 

decrease in the concentration of a blue-green algae (Anabaena species) but no effect on diatoms 

or green algae. It is unclear if the effect was a primary, secondary, or incidental effect. 

4-32
 
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5 - 46
 
Table 2-1. Selected physical and chemical properties of diflubenzuron 1 

Synonyms and trade names	 DFB; Difluron; Dimilin; Duphacid; DU 112307; ENT 29054; Micromite; 

OMS 1804; PH 60-40; TH-6040 

U.S. EPA Reg. No. 400-465 and 400-474 (C&P Press, 2003)
 

CAS number 35367-38-5 (USDA/ARS 1995)
 

Molecular weight 310.69 (USDA/ARS 1995; Meylan and Howard 1995)
 

Molecular formula C H ClF N O (USDA/ARS 1995; Budavari 1989)
 

SMILES Notation O=C(NC(=O)c(c(F)ccc1)c1F)Nc(ccc(c2)Cl)c2
 

Appearance/state, ambient Solid (USDA/ARS 1995)
 

Melting point 230 to 232 °C (USDA/ARS 1995)
 

Vapor pressure 0.00012 mPa (USDA/ARS 1995)
 

Water solubility (mg/L) �0.3 (Budavari 1989)
 

14 9 2 2 2 

0.08 at 25°C (USDA/ARS 1995; Knisel et al. 1992)
 

0.0888 mg/L in deionized, 0.0926 mg/L in field water (Mabury and Crosby 1996)
 

log K	 3.89 (USDA/ARS 1995) [i.e., K = 103.89 = 7762] ow	 ow 

3.59 (estimated) (Meylan and Howard 1995) 

3.88 (experimental) (Meylan and Howard 1995) 

3.83 ±0.02 (Marsella et al. 2000) 

Koc 135.3 (organic soil) (Sundaram et al. 1997) 

332.0 (silty clay loam) (Sundaram et al. 1997) 

8700 (NOS) (USDA/ARS 1995) 

10000 (Knisel and Davis 2000) 

Kd 17.59 (organic soil) (Sundaram et al. 1997) 

16.42 (silty clay loam) (Sundaram et al. 1997) 

Foliar halftimes 9.3 days (Sundaram 1986, 1996) 2 8 days, 20-80% loss (Wimmer et al. 1993 )

29 days (hardwood, van den Berg 1986) 

36 days (conifer, van den Berg 1986) 

Foliar washoff 50% to 100% depending on formulation, intensity of rainfall, and time of rain after 

application (Sundaram and Sundaram 1994) 

Litter halftimes 8.36 days (Sundaram 1986, 1996) 

Soil halftimes sterile: 346 days in sand and muck (NOS)(Chapman et al. 1985) 

natural: 18.7 days in sand and muck (NOS)(Chapman et al. 1985) 

7.49 days (field study, Sundaram 1986, 1996) 

Water photolysis halftime 17±4 hours at pH 7 in distilled water (Marsella et al. 2000) 

8±2 hours at pH 9 in distilled water (Marsella et al. 2000) 

12.3±0.7 hours at pH 9 in stream water (Marsella et al. 2000) 

Aerobic microbial halftime 25.7 days for DFB; 39.7 days for 4-chlorophenylurea (Dzialo and Maynard 1999) 

(soil/water) 50 hours [2.1 days] (Walstra and Joustra 1990) 

5.4 days in water, 8.6 days in sediment (Willard 2000a) 

Anaerobic microbial 34 days (Thus et al. 1991) 

halftime (soil/water) 

Water halftime (NOS) 0.97 (0.77-1.16) days without aeration (Anton et al. 1993) 

3 Henry’s law constant	 0.00047 Pa m /mol at 25°C (USDA/ARS 1995) 3 0.234 ±0.002 Pa×m /mole at 20°C (Mabury and Crosby 1996)

1

 Specific environmental fate parameters used in modeling are discussed in Section 3.2.

2 Reflects initial losses. Remaining DFB much more persistent. 

Tables - 1 
Table 2-2: Commercial formulations of diflubenzuron 1 

Formulation 

(Supplier) 

Type of 

formulation 

%DFB (w/w) 2 

(Concentration) 

Application Rates 3 

Single Total for year 

Uses 

Adept (Uniroyal) Water Soluble 25% N/A N/A Ornamentals 

Bags 

Dimilin 2L Aqueous 22% 2-16 fl oz/acre 24 fl oz/acre Trees and 

(Uniroyal) flowable (2 lbs/gallon) various crops 

Dimilin 4L Liquid 40.4 % 0.5-2 fl oz/acre 2 fl oz/acre Forests, 

(Uniroyal) (4 lbs/gallon) ground or 

aerial. 

Dimilin 25W 4 Wettable 25% 1-4 oz/acre 4 oz/acre 

(Uniroyal) powder 

Dimilin SC Liquid 40.4 % N/A N/A Mushrooms 

(Uniroyal) (4 lbs/gallon) and ornaments 

Micromite 25W 5 Wettable 25% 1-4 oz/acre 4 oz/acre Forests, 

(Uniroyal) powder ground or 

aerial. 

Micromite 25WS Water Soluble 25% 1.25 lbs/acre 3.75 lbs/acre Citrus crops, 

(Uniroyal) Bags ground or 

aerial 

Micromite 25WGS Water 80% 6.25 oz/acre 18.75 oz/acre Citrus crops, 

(Uniroyal) Dispersible ground or 

Granules aerial

1

 Source: Specimen labels from C&P Press, 2004. Only products in bold font are labeled for gypsy moth.
 

2 The remainder of the product formulation is classified as inerts. See text for discussion.
 3

 All application rates are expressed in amount (lb or oz) of formulation not amounts of active ingredient per acre. 


N/A indicated that the product is not labeled for broadcast applications. For products labeled for gypsy moth, the
 

range of application rates are those that apply to the gypsy moth.
 

4

 A separate formulation is available for mushrooms and ornamentals.
 

5

 The registration for this formulation has been canceled (U.S. EPA/OPP 2002b)
 

Tables - 2 
757 

TABLE 2-3: Use of diflubenzuron by USDA from 1995 to 2002 for 

Suppression, Eradication, and Slow the Spread 1 

Year Suppression Eradication Slow the Spread 

1995 161,231 

1996 111,362 6 1,248 

1997 16,447 

1998 757 

1999 5,275 1,047 

2000 18,090 

2001 187,784 650 

2002 131,601 3,938 

2003 25,124 

Total Acres 657,671 6 6,883 

% of Total 98.96% 0.001% 1.04%

Total 

161,231 

112,616 

16,447 

6,322 

18,090 

188,434 

135,539 

25,124 

664,560 

1 Source: GMDigest, Morgantown, WV (http://na.fs.fed.us/wv/gmdigest/) 

Tables - 3 
Table 3-1: Chemical and site parameters used in GLEAMS modeling for diflubenzuron. 

Chemical Specific Parameters 

Parameter Clay Loam Sand Comment/ 

Reference 

Halftimes (days)

 Aquatic Sediment 34 34 34 Thus et al. 1991

 Foliar 9.3 9.3 9.3 Sundaram 1986, 1996

 Soil 10 1.1 2.1 Note 1

 Water 5.4 Note 2 

Ko/c, mL/g	 8700 Note 3 

K , mL/g 261 130 26.1 Note 4 

Water Solubility, mg/L 0.0926 Mabury and Crosby 1996, field sample 

Foliar wash-off fraction 0.5 Note 5 

Fraction applied to foliage 0.8 

Fraction applied to soil 0.2 

d

Note 1	 Value for sand taken as reported half-time of 50 hours (2.0833 days) taken from Walstra and Joustra 

1990. Value for loam taken as reported half-time in silt-loam from Thus and van der Laan-Straathof 

1994. No studies on aerobic soil metabolism in clay were found. The value of 10 days is taken from 

Knisel and Davis (2000) as an upper range. 

Note 2	 Value for microbial halftime in water from Willard 2000a. Halftimes may be substantially less under 

conditions where photolysis is the principal route of degradation. See Table 2-1. 

Note 3	 A very wide range of Koc values (about 135 to 10,000) have been reported (see Table 2-1). The value 

of 8700 is recommended by USDA/ARS (1995) and is close to the value of 10,000 recommended by 

Knisel and Davis (2000). 

Note 4	 Based on the general relationship: Kd = Koc × OC using OC values of 0.003 for sand, 0.015 for loam, 

and 0.030 for clay (SERA 2003b). 

Note 5	 This is highly variable. Knisel and Davis (2000) recommend 0.05. The higher value of 0.5 is 

consistent with the field studies by Sundaram and Sundaram (1994) and Wimmer et al. (1993). 

Site Parameters 

(see SERA 2004, TD 2004-02.04a dated February 8, 2004 for details) 

Pond	 1 hectare pond, 2 meters deep, with a 0.01 sediment fraction. 10 hectare square field (1093' by 

1093') with a root zone of 12 inches. 

Stream	 Base flow rate of 710,000 L/day with a flow velocity of 0.08 m/second or 6912 meters/day. 

10 hectare square field (1093' by 1093') with a root zone of 12 inches. 

Tables - 4 
Table 3-2: Summary of modeled concentrations of diflubenzuron in streams (all units are 

μg/L or ppb). 

Annual Rainfall Clay Loam Sand 

Rainfall per Event 

(inches) (inches)1 Average Maximum Average Maximum Average Maximum 

Concentration per lb/acre applied (from GLEAMS) 

5 0.14 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

10 0.28 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

15 0.42 0.04113 5.17705 0.00000 0.00000 0.00000 0.00000 

20 0.56 0.11543 14.59505 0.00000 0.00000 0.00000 0.00000 

25 0.69 0.20602 26.22114 0.00000 0.00000 0.00000 0.00000 

50 1.39 0.60485 81.46441 0.00000 0.00000 0.00000 0.00002 

100 2.78 1.02559 156.23308 0.03588 11.68278 0.00000 0.00028 

150 4.17 1.04171 199.48431 0.09107 29.67516 0.00000 0.00105 

200 5.56 0.97117 229.82322 0.15544 50.70660 0.00001 0.00258 

250 6.94 0.88544 253.52663 0.22002 71.88424 0.00045 0.13780 

Application rate: 0.0624 lbs/acre 

Concentration at above application rate 

5 0.14 0 0 0 0 0 0 

10 0.28 0 0 0 0 0 0 

15 0.42 0.003 0.32305 0 0 0 0 

20 0.56 0.007 0.91073 0 0 0 0 

25 0.69 0.0129 1.6362 0 0 0 0 

50 1.39 0.0377 5.08338 0 0 0 0 

100 2.78 0.064 9.74894 0.002 0.72901 0 0 

150 4.17 0.065 12.4478 0.006 1.85173 0 0 

200 5.56 0.0606 14.341 0.01 3.16409 0 0 

250 6.94 0.0553 15.8201 0.0137 4.48558 0 0.009

1 Rain is assumed to occur at the same rate every 10th day – i.e., 36 rainfall events per year. 

Tables - 5 
Table 3-3: Summary of modeled concentrations of diflubenzuron in ponds (all units are μg/L 

or ppb) 

Annual Rainfall Clay Loam Sand
 

Rainfall per Event
 

(inches) (inches)1
 Average Maximum Average Maximum Average Maximum 

Concentration per lb/acre applied (from GLEAMS) 

5 0.14 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

10 0.28 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

15 0.42 0.00704 0.07849 0.00000 0.00000 0.00000 0.00000 

20 0.56 0.01700 0.26465 0.00000 0.00000 0.00000 0.00000 

25 0.69 0.02989 0.56583 0.00000 0.00000 0.00000 0.00000 

50 1.39 0.11171 3.32693 0.00000 0.00000 0.00000 0.00000 

100 2.78 0.29257 12.37300 0.01577 1.63558 0.00000 0.00007 

150 4.17 0.39616 23.59907 0.04933 5.81660 0.00000 0.00033 

200 5.56 0.45379 35.86106 0.09695 12.41986 0.00001 0.00096 

250 6.94 0.48619 48.35946 0.15210 20.70574 0.00035 0.05865 

Application rate: 0.0624 lbs/acre 

Concentration at above application rate 

5 0.14 0 0 0 0 0 0 

10 0.28 0 0 0 0 0 0 

15 0.42 0.0004 0.0049 0 0 0 0 

20 0.56 0.00106 0.016514 0 0 0 0 

25 0.69 0.00187 0.035308 0 0 0 0 

50 1.39 0.00697 0.2076004 0 0 0 0 

100 2.78 0.018256 0.7720752 0.001 0.1020602 0 0 

150 4.17 0.02472 1.472582 0.00308 0.3629558 0 0 

200 5.56 0.028317 2.2377301 0.00605 0.7749993 0 0 

250 6.94 0.030338 3.0176303 0.00949 1.2920382 0 0.00366

1 Rain is assumed to occur at the same rate every 10th day – i.e., 36 rainfall events per year. 

Tables - 6 
Table 3-4: Chemical and site parameters used in GLEAMS modeling for 4-chloroaniline. 

Chemical Specific Parameters 

Parameter Clay Loam Sand Comment/ 

Reference 

Halftimes (days)

 Aquatic Sediment 150 Note 2

 Foliar 0.16 Note 2

 Soil 37.5 Note1

 Water 151 Note 2 

Ko/c, mL/g 72 Note 1 

K , mL/g d 2.2 1.1 0.22 Note 3 

Water Solubility, mg/L 3900 Note 1 

Foliar wash-off fraction 0.5 

Coefficient of transformation 0.41 Note 4 

Note 1	 Estimated from EPI-Suite (Meylan and Howard 1998, 2000) 

Note 2	 WHO 2003. Foliar halftime is not given explicitly in WHO (2003) and is estimated here based on the 

atmospheric halftime of 3.9 hours. 

Note 3	 Based on Kd = Ko/c × OC, where OC is the proportion of organic carbon. The OC in sand, loam, and 

clay is taken as 0.003 for sand, 0.015 for loam, and 0.030 for clay (SERA 2004). 

Note 4	 This is the ratio of the molecular weight of chloroaniline (127.57) to that of diflubenzuron (310.69). 

See discussion by Knisel and Davis (2000, p. 110). 

Tables - 7 
Table 3-5: Summary of modeled concentrations of 4-chloroaniline in streams (all units are 

μg/L or ppb) 

Annual Rainfall Clay Loam Sand 

Rainfall per Event 

(inches) (inches)1 Average Maximum Average Maximum Average Maximum 

Concentration per lb/acre applied (from GLEAMS) 

5 0.14 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

10 0.28 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

15 0.42 0.06559 4.19361 0.00048 0.01145 0.11234 2.57651 

20 0.56 0.15452 10.45786 0.01616 0.32734 0.36403 10.48046 

25 0.69 0.22436 15.84683 0.03969 0.85101 0.55917 19.16073 

50 1.39 0.31156 27.90970 0.16080 4.59647 0.77622 44.23856 

100 2.78 0.29226 30.80407 0.22906 9.17859 0.59128 52.72812 

150 4.17 0.13293 24.52481 0.20128 9.67567 0.45074 51.02312 

200 5.56 0.06009 14.09093 0.16267 8.73307 0.36145 49.79360 

250 6.94 0.01924 5.74944 0.12680 7.21420 0.30139 47.06395 

Application rate: 0.0624 lbs/acre 

Concentration at above application rate 

5 0.14 0 0 0 0 0 0 

10 0.28 0 0 0 0 0 0 

15 0.42 0.00409 0.2616813 0 0.0007 0.00701 0.1607742 

20 0.56 0.00964 0.6525705 0.00101 0.020426 0.022715 0.6539807 

25 0.69 0.014 0.9888422 0.00248 0.053103 0.034892 1.1956296 

50 1.39 0.019441 1.7415653 0.010034 0.2868197 0.048436 2.7604861 

100 2.78 0.018237 1.922174 0.014293 0.572744 0.036896 3.2902347 

150 4.17 0.00829 1.5303481 0.01256 0.6037618 0.028126 3.1838427 

200 5.56 0.00375 0.879274 0.010151 0.5449436 0.022554 3.1071206 

250 6.94 0.0012 0.3587651 0.00791 0.4501661 0.018807 2.9367905

1 Rain is assumed to occur at the same rate every 10th day – i.e., 36 rainfall events per year. 

Tables - 8 
Table 3-6: Summary of modeled concentrations of 4-chloroaniline in ponds (all units are μg/L 

or ppb) 

Annual Rainfall Clay Loam Sand
 

Rainfall per Event
 

(inches) (inches)1
 Average Maximum Average Maximum Average Maximum 

Concentration per lb/acre applied (from GLEAMS) 

5 0.14 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

10 0.28 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

15 0.42 0.31929 0.69851 0.00288 0.00477 0.65741 1.11311 

20 0.56 0.56688 1.80242 0.07465 0.15746 1.72523 4.20894 

25 0.69 0.74573 2.90175 0.16734 0.40004 2.48876 7.61750 

50 1.39 1.04158 6.43073 0.63473 2.28508 3.46266 18.05225 

100 2.78 1.01591 8.41740 0.97319 5.00787 2.89735 23.03849 

150 4.17 0.60259 6.77759 0.90309 5.52346 2.34727 22.92303 

200 5.56 0.29679 4.08394 0.75792 5.16526 1.96069 22.29465 

250 6.94 0.10055 1.77278 0.60774 4.47424 1.68309 21.01092 

Application rate: 0.0624 lbs/acre 

Concentration at above application rate 

5 0.14 0 0 0 0 0 0 

10 0.28 0 0 0 0 0 0 

15 0.42 0.019924 0.043587 0.0002 0.0003 0.041022 0.069458 

20 0.56 0.035373 0.112471 0.00466 0.00983 0.1076544 0.2626379 

25 0.69 0.046534 0.1810692 0.010442 0.024963 0.1552986 0.475332 

50 1.39 0.064995 0.4012776 0.039607 0.142589 0.21607 1.1264604 

100 2.78 0.063393 0.5252458 0.060727 0.3124911 0.1807946 1.4376018 

150 4.17 0.037602 0.4229216 0.056353 0.3446639 0.1464696 1.4303971 

200 5.56 0.01852 0.2548379 0.047294 0.3223122 0.1223471 1.3911862 

250 6.94 0.00627 0.1106215 0.037923 0.2791926 0.1050248 1.3110814

1 Rain is assumed to occur at the same rate every 10th day – i.e., 36 rainfall events per year. 

Tables - 9 
Table 3-7: Estimated Environmental Concentrations (μg/L or ppb) of diflubenzuron in ponds 

and streams. 

Scenario Peak Long-Term Average 

MODELING FOR THIS RISK ASSESSMENT (0.0624 lb/acre or 70 g/ha) 

Stream 

Direct Spray 1 5.7 N/A 

100 Foot buffer 1 0.11 N/A 

GLEAMS (Table 3-2) 2 (<0.01 to 16) 0.01 (0 to 0.06) 

Pond 

Direct Spray 2 3.5 N/A 

100 Foot buffer 2 0.07 N/A 

GLEAMS (Table 3-3) 0.2 (<0.005 to 3) at 0.06 lb/ac 0.007 (0 to 0.03) at 0.06 lb/ac 

OTHER MODELING 

USDA (1995) 16.01 (stream, direct spray) N/A 

2.76 to 13.14 (stream, runoff) 

1.22 (pond) 

U.S. EPA/OPP 1997a. Pond: citrus 3.4 ppb at 6x 0.06 lb/ac 0.74 ppb at 6x 0.06 lb/ac 

crops 8.1 ppb at 0.67 lb/ac 0.87 ppb at 0.67 lb/ac 

U.S. EPA/OPP 1997a. Pond: direct 11.7 ppb at 0.05 lb/ac N/A 

applications to water in forestry 22.8 ppb at 0.07 lb/ac 

46.2 ppb at 0.15 lb/ac 

91.8 ppb at 0.32 lb/ac 

Harned and Relyea 1997 Peak concentration of 1 ppb at an application rate of 350 g/ha. Longer 

term concentration of about 0.1 ppb. See text for discussion. 

Schocken et al. 2001 Peak concentrations of about 0.2 to 0.3 ppb in ponds and 0.9 ppb in 

steams at an application rate of 0.125 lb/acre. See text for discussion. 

1

 See Worksheet 10b 

2

 See Worksheet 10a 

Tables - 10 
Table 3-8: Concentrations of diflubenzuron in surface water used in this risk assessment (see 

Section 3.2.3.4.6 for discussion). 

At application rate: 0.0624 lb/acre 

Peak Concentration 

(ppb or μg/L) 

Longer Term Concentration 

(ppb or μg/L) 

Central 0.4 0.02 

Lower 0.01 0.001 

Upper 

Water contamination rate 1 

16 

mg/L per lb/acre applied. 

0.1 

Peak Concentration 

(mg/L per lb/acre) 

Longer Term Concentration 

(mg/L per lb/acre) 

Central 6.41e-03 3.21e-04 

Lower 1.60e-04 1.60e-05 

Upper 2.56e-01 1.60e-03

1 Water contamination rates – concentrations in units of mg/L expected at an application rate of 1 lb/acre. These 

values are entered into Worksheet A04 for diflubenzuron. This rate is adjusted to the program application rate in 

all worksheets involving exposure to contaminated water. 

Tables - 11 
Table 3-9: Concentrations of 4-chloroaniline in surface water used in this risk assessment (see 

Section 3.2.3.4.7 for discussion). 

At application rate: 0.0624 lb/acre 

Peak Concentration 

(ppb or μg/L) 

Longer Term Concentration 

(ppb or μg/L) 

Central 0.5 0.05 

Lower 0.00003 0.0002 

Upper 

Water contamination rate 1 

3 

mg/L per lb/acre applied. 

0.2 

Peak Concentration 

(mg/L per lb/acre) 

Longer Term Concentration 

(mg/L per lb/acre) 

Central 8.01e-03 8.01e-04 

Lower 4.81e-07 3.21e-06 

Upper 4.81e-02 3.21e-03

1 Water contamination rates – concentrations in units of mg/L expected at an application rate of 1 lb/acre. These 

values are entered into Worksheet A04 for 4-chloroaniline. This rate is adjusted to the program application rate 

in all worksheets involving exposure to contaminated water. 

Tables - 12 
 

 

 

 

 

Table 4-1: Summary of field studies on the effects of diflubenzuron on terrestrial invertebrates 1 

Range of 

Application 

Rates (g/ha) No Adverse Effects 

Species 

Adverse Effects 

<20 ants (Catangui et al. 1996) 

Cotesia melanoscelus (GM parasitic wasp)

 (Webb et al. 1989) 

grasshoppers (Jech et al. 1993) 

20 - <40 lacewing and beetles (Ables et al. 1977) 

carabids, crickets, lice (Butler et al. 1997) 

honey bee (Matthenius1975) 

honey bee [×8](Robinson 1978,1979) 

gypsy moth and macrolepidoptera (Butler et al. 1997) 

grasshopper (Everts 1990 ) 

Apanteles melanoscelus # (GM parasitic wasp) 

(Madrid and Stewart1981) 

40 - < 60 lacewing and beetles (Ables et al. 1977) 

60 - < 100 Ooencyrtus kuvanae (GM parasitic wasp)

 (Brown and Respicio 1981) 

lacewing and beetles (Deakle and Bradley1982) 

honey bee (Matthenius1975) 

sucking herbivorous insects, microlepidoptera, and

 predaceous arthropods(Martinat et al. 1988) 

spiders* and orthopteroid*(Martinat et al. 1993) 

mites and springtails (Perry et al. 1997) 

spiders** (Perry et al. 1997) 

non-lepidopteran insects (Sample et al. 1993a,b) 

mites* and collembolans* (Van Den Berg 1986) 

grasshopper (Everts 1990 ) 

grasshoppers, moths, carabid beetles (Butler 1993) 

lepidoptera (Sample et al. 1993a,b) 

macrolepidoptera and other herbivorous insects

 (Martinat et al. 1988) 

Yellow jacket wasp (Barrows et al. 1994) 

100 - < 150 ants (Weiland 2000) 

Psylla parasites and predators (Westigard 1979) 

lacewing and beetles (Ables et al. 1977) 

honey bee (Emmett and Archer 1980) 

honey bee [×8](Robinson 1978,1979) 

soil mites (Blumberg 1986) 

Yellow jacket wasp (Weiland 2000) 

150 - < 200 various arthropod predators (Keever et al. 1977) lepidopteran egg mortality (low) (Kumar et al. 1994) 

mites (Marshall 1979) 

200 - < 300 ants (Weiland 2000)
 

carabid beetles (Heinrichs et al. 1979)
 

lacewing and beetles (Ables et al. 1977)
 

mites (Marshall 1979)
 

borer weevil (Schroeder 1996)
 

predatory damsel bugs and sucking insects


 (Turnipseed et al. 1974) 

Yellow jacket wasp (Weiland 2000) 

Psylla parasites and predators (Westigard 1979) 

flying insects, esp. midges, gnats, and mosquitoes 

(Wilson and Wan 1977a) 

$ 300 honey bee (Buckner et al. 1975) lepidopteran egg mortality (high) (Kumar et al. 1994) 

honey bee (Emmett and Archer 1980) Psylla parasites and predators (Westigard 1979)

honey bee and other beneficial insects

 (Schroeder 1980) 

1 Studies summarized in Appendix 3a. See text for discussion. A single asterisk (*) indicates transient or equivocal effects. A 

double asterisk (**) indicates effects that were secondary to decrease in prey. The # symbol indicates an effect clearly due to 

toxicity. GM used as abbreviation for gypsy moth. Multiple applications are indicated in brackets with a × symbol followed by the 

number of applications. 

Tables - 13 
Table 4-2: Summary of modeled concentrations of diflubenzuron in soil (all units are mg/kg 

or ppm) 

Annual Rainfall Clay Loam Sand
 

Rainfall per Event
 

(inches) (inches)1
 Average Maximum Average Maximum Average Maximum 

Concentration per lb/acre applied (from GLEAMS) 

5 0.14 0.00841 0.14004 0.00092 0.11651 0.00169 0.12485 

10 0.28 0.00926 0.14004 0.00106 0.11652 0.00194 0.12484 

15 0.42 0.00924 0.13992 0.00106 0.11653 0.00193 0.12484 

20 0.56 0.00918 0.13962 0.00106 0.11653 0.00193 0.12484 

25 0.69 0.00910 0.13914 0.00106 0.11653 0.00193 0.12484 

50 1.39 0.00834 0.13431 0.00106 0.11653 0.00192 0.12484 

100 2.78 0.00650 0.11909 0.00104 0.11450 0.00190 0.12484 

150 4.17 0.00412 0.09305 0.00099 0.10879 0.00188 0.12484 

200 5.56 0.00234 0.06298 0.00091 0.09889 0.00186 0.12484 

250 6.94 0.00104 0.05236 0.00080 0.08527 0.00184 0.12478 

Application rate: 0.0624 lbs/acre 

Concentration at above application rate 

5 0.14 5.2e-04 0.00874 5.7e-05 0.00727 1.1e-04 0.00779 

10 0.28 5.8e-04 0.00874 6.6e-05 0.00727 1.2e-04 0.00779 

15 0.42 5.8e-04 0.00873 6.6e-05 0.00727 1.2e-04 0.00779 

20 0.56 5.7e-04 0.00871 6.6e-05 0.00727 1.2e-04 0.00779 

25 0.69 5.7e-04 0.00868 6.6e-05 0.00727 1.2e-04 0.00779 

50 1.39 5.2e-04 0.00838 6.6e-05 0.00727 1.2e-04 0.00779 

100 2.78 4.1e-04 0.00743 6.5e-05 0.00714 1.2e-04 0.00779 

150 4.17 2.6e-04 0.00581 6.2e-05 0.00679 1.2e-04 0.00779 

200 5.56 1.5e-04 0.00393 5.7e-05 0.00617 1.2e-04 0.00779 

250 6.94 6.5e-05 0.00327 5.0e-05 0.00532 1.1e-04 0.00779

1 Rain is assumed to occur at the same rate every 10th day – i.e., 36 rainfall events per year. 

Tables - 14 
Table 4-3: Summary of modeled concentrations of 4-chloroaniline in soil (all units are mg/kg 

or ppm) 

Annual Rainfall Clay Loam Sand
 

Rainfall per Event
 

(inches) (inches)1
 Average Maximum Average Maximum Average Maximum 

Concentration per lb/acre applied (from GLEAMS) 

5 0.14 0.00672 0.02893 0.00680 0.04917 0.00750 0.04216 

10 0.28 0.00655 0.02685 0.00626 0.04550 0.00666 0.04159 

15 0.42 0.00699 0.02697 0.00709 0.04556 0.00751 0.04167 

20 0.56 0.00691 0.02665 0.00734 0.04562 0.00728 0.04168 

25 0.69 0.00668 0.02618 0.00748 0.04566 0.00685 0.04157 

50 1.39 0.00360 0.02252 0.00737 0.04582 0.00493 0.04032 

100 2.78 0.00631 0.01739 0.00622 0.04519 0.00323 0.04015 

150 4.17 0.00307 0.01146 0.00529 0.04326 0.00254 0.04001 

200 5.56 0.00142 0.00759 0.00450 0.03994 0.00216 0.03997 

250 6.94 0.00050 0.00357 0.00375 0.03540 0.00193 0.03999 

Application rate: 0.0624 lbs/acre 

Concentration at above application rate 

5 0.14 4.2e-04 0.00181 4.2e-04 0.00307 4.7e-04 0.00263 

10 0.28 4.1e-04 0.00168 3.9e-04 0.00284 4.2e-04 0.0026 

15 0.42 4.4e-04 0.00168 4.4e-04 0.00284 4.7e-04 0.0026 

20 0.56 4.3e-04 0.00166 4.6e-04 0.00285 4.5e-04 0.0026 

25 0.69 4.2e-04 0.00163 4.7e-04 0.00285 4.3e-04 0.00259 

50 1.39 2.2e-04 0.00141 4.6e-04 0.00286 3.1e-04 0.00252 

100 2.78 3.9e-04 0.00109 3.9e-04 0.00282 2.0e-04 0.00251 

150 4.17 1.9e-04 0.0007 3.3e-04 0.0027 1.6e-04 0.0025 

200 5.56 8.9e-05 0.0005 2.8e-04 0.00249 1.3e-04 0.00249 

250 6.94 3.1e-05 0.0002 2.3e-04 0.00221 1.2e-04 0.0025

1 Rain is assumed to occur at the same rate every 10th day – i.e., 36 rainfall events per year. 

Tables - 15 
Table 4-4: Summary of diflubenzuron toxicity values used in ecological risk assessment 

Organism Endpoint Toxicity Value Reference, Species 

Mammals Acute NOAEL 1118 mg/kg Blaszcak 1997a, rats [Dimilin 2L] 

Chronic NOAEL 2 mg/kg/day Greenough et al. 1985, dogs 

Birds Acute NOAEL 2500 mg/kg Alsager and Cook 1975, blackbirds 

Chronic NOAEL 110 mg/kg Beavers et al. 1990b, quail 

Terrestrial arthropods See Table 4-5 for toxicity values 

Soil invertebrates 

Earthworm NOEC 780 mg/kg soil Berends et al. 1992 

Soil microorganisms 

Sensitive 50 ppm LOEC 50 ppm ÷ 10 Townshend et al. 1983 

Tolerant 100 pp NOEC 100 ppm Townshend et al. 1983 

Fish Acute 

Sensitive LC 50 25 mg/L Johnson and Finley 1980, yellow perch 

Tolerant LC 50 500 mg/L Reiner and Parke 1975, fathead minnow 

Fish Chronic 

Sensitive Reproductive NOEC 0.05 mg/L Livingston and Koenig 1977, 

mummichog 

Tolerant Reproductive NOEC 0.1 mg/L Cannon and Krize 1976, fathead minnow 

Aquatic Invertebrates See Table 4-6 for toxicity values 

Aquatic Plants 

Sensitive NOEC for growth 0.045 mg/L Hansen and Garton 1982a, Selenastrum 

capricornutum 

Tolerant NOEC for growth 0.38 mg/L Thompson and Swigert 1993c, Navicula 

pelliculosa 

Aquatic Microorganisms NOEC for respiration 0.05 mg/L Kreutzweiser et al. 2001 [4.3.3.4]

1 NOECs are used directly when available. When only a LOEC is available, the LOEC is divided by 10 to approximate the 

NOEC. This is indicated by the “÷10” following the LOEC. 

Tables - 16 
 

Table 4-5: Diflubenzuron toxicity values used in risk assessment for terrestrial arthropods (see 

Table 4-1 for additional details). 

Organism Endpoint Toxicity Reference 

Value1 

Grasshoppers Field LOAEL 22 g/ha ÷ 10 Jech et al. 1993 

Apanteles melanoscelus 2 Field LOAEL 30 g/ha ÷ 10 Madrid and Stewart1981 

Macrolepidoptera Field LOAEL 35 g/ha ÷ 10 Butler et al. 1997 

Mandibulate herb. insects Field LOAEL 70 g/ha ÷ 10 Martinat et al. 1988 

Ooencyrtus kuvanae 2 Field NOAEL 67 g/ha Brown and Respicio 1981 

 Microlepidoptera Field NOAEL 70 g/ha Martinat et al. 1988 

Predaceous arthropods Field NOAEL 70 g/ha Martinat et al. 1988 

Sucking herbaceous Field NOAEL/LOAEL 70/281 g/ha Martinat et al. 1988/Turnipseed et al. 1974 

insects 

Spiders Field NOAEL 70 g/ha Martinat et al. 1993 

Mites and collembolans Field NOAEL/LOAEL 70/140 g/ha Perry et al. 1997/Blumberg 1986 

ants Field NOAEL 280 Weiland 2000 

Lacewing Field NOAEL/LOAEL 140/280 g/ha Ables et al. 1977 

Honey bee Field NOAEL 400 g/ha Emmett and Archer 1980

1 Field NOAELs are used directly when available. When only a LOAEL is available, the LOAEL is divided by
 

10 to approximate the NOAEL. This is indicated by the “÷10” following the LOAEL.
 

2

 A parasitic wasp to the gypsy moth.
 

Tables - 17 
Table 4-6: Diflubenzuron toxicity values used in risk assessment for aquatic invertebrates. 

Organism Endpoint Toxicity Value Reference 

ppb or μg/L1 

ACUTE (see Table 4-8 for additional details) 

Daphnia NOEC 0.3 Corry et al. 1995 

Ceriodaphnia NOEC 0.75 Hall 1986 

Copepods NOEC 0.93 Savitz et al. 1994 

crabs NOEC 2 Cunningham and Meyers 1987 

rotifers NOEC 20 Corry et al. 1995 

large insects NOEC 2000 Lahr et al. 2001 

molluscs NOEC 125000 Wilcox and Coffey 1978 

LONGER TERM (see Table 4-9 for additional details) 

Daphnia NOEC 0.04 Surprenant 1988 

stoneflies and mayflies NOEC 0.1 Hansen and Garton 1982b 

Ceriodaphnia NOEC 0.25 Hall 1986 

dragonflies NOEC 0.7 O’Halloran and Liber 1995 

ostracods NOEC 2.5 Liber and O’Halloran 1995 

coleoptera and oligochaetes NOEC 50 Hansen and Garton 1982a 

molluscs NOEC 320 Surprenant 1989 

1

 In worksheets, all concentrations in ppb are divided by 1000 to convert to concentrations in ppm or mg/L. 

Tables - 18 
Table 4-7: Summary of 4-chloroaniline toxicity values used in ecological risk assessment 

Organism Duration/Endpoint Toxicity Value Reference, species 

Mammals Acute/Toxicity NOAEL 8 mg/kg/day Used in HHRA 

Chronic/Toxicity NOAEL 1.25 Estimated from LOAEL of 12.5 

mg/kg/day mg/kg/day 

Birds Acute/Toxicity NOAEL 8 mg/kg/day No data. Uses value for mammals 

Chronic/Toxicity NOAEL 1.25 No data. Uses value for mammals 

mg/kg/day 

Earthworms NOEC 540 mg/kg soil WHO 2003 

Soil Microorganisms NOEC 1000 ppm Welp and Brummer 1999 

Fish 

Acute LC 50 2.4 mg/L WHO 2003, Bluegill 

Chronic NOEC, reproduction 0.2 mg/L Bresch et al. 1990, Zebra fish 

Aquatic Invertebrates 

Acute NOEC, mortality 0.013 mg/L Kuhn et al 1989a 

Chronic NOEC, reproduction 0.01 mg/L Kuhn et al 1989a 

Aquatic plants EC 10 0.02 mg/L Schmidt and Schnabl 1988, green algae 

Aquatic NOEC (30 min) 5.1 mg/L Ribo and Kaiser 1984, photobacteria 

Microorganisms 

Tables - 19 
Table 4-8: Acute toxicity of diflubenzuron in aquatic invertebrates 

Concentrations No Effect Adverse Effect
 

(μg/L or ppb) Species/group [conc. ppb](Reference) Species/group [conc. ppb](Reference)
 

0.1 to <1	 mysid shrimp[0.12] (Breteler 1987) 

Daphnia [0.3](Corry et al. 1995) 

Daphnia [0.45](Kuijpers 1988) 

Ceriodaphnia [0.75](Hall 1986) 

copepods [0.93](Savitz et al. 1994) 

1 to <10	 fiddler crabs [2] (Cunningham and Meyers 1987) 4 Horseshoe crabs [5] (Weis and Ma 1987)

amphipods [7] (Corry et al. 1995) 

10 to <100	 rotifers[20] (Corry et al. 1995) 

snails [45](Hansen and Garton 1982a) 

100 to <1000 

>1000	 backswimmer2

 [2000] (Lahr et al. 2001) 

snail [125,000](Wilcox and Coffey 1978) 

Mosquito [0.5] (Miura and Takahashi 1974)
 

Daphnia [0.7](Corry et al. 1995)
 

Daphnia [0.7](Kuijpers 1988)
 

Daphnia [0.75, neonate](Majori et al. 1984)
 

fairy shrimp [0.74] (Lahr et al. 2001)
 

gammarids[1](Hansen and Garton 1982a)
 

Ceriodaphnia [1.7](Hall 1986)
 

copepods [1.7](Savitz et al. 1994)
 

midges[1.8](Hansen and Garton 1982a)
 

blue crab eggs [1.8] (Lee and Oshima 1998)
 

grass shrimp [3.4](Tourat and Rao 1987)
 

grass shrimp [2-3](Wilson and Costlow 1986)
 

mysid shrimp[2.1]Nimmo et al. 1979
 

Mayfly [10] (Miura and Takahashi 1974)
 

Amphipods [13](Corry et al. 1995)
 

Daphnia [23, adult](Majori et al. 1984)
 

Dragonfly [50] (Miura and Takahashi 1974)
 

Horseshoe crabs [50] (Weis and Ma 1987)
 

beetles [100] (Miura and Takahashi 1974)
 

fiddler crabs [200] (Cunningham and Meyers 1987)
 

tricoptera [250] (Bradt and Williams 1990)
 

grass shrimp[640] (Bionomics-EG&G 1975)
 

midge [560] (Julin and Sanders 1978)
 

1

 Macrocosm study 

2

 No molting during short term exposures 

3

 Litoral enclosures 

4

 Marginal signs of toxicity 

Tables - 20 
 

 

 

 

 

Table 4-9: Chronic toxicity of diflubenzuron in aquatic invertebrates 

Concentrations No Effect Adverse Effect 

(μg/L or ppb) Species/group [conc. ppb](Reference) Species/group [conc. ppb](Reference) 

>0.01 to 0.1 Daphnia[0.04]Surprenant 1988 

stream inverts1

 [0.1](Hansen and Garton 1982a )1 

stoneflies and mayflies[0.1] 

(Hansen and Garton 1982b )1 

Daphnia [0.06] U.S. EPA 1997a 5 

mysid shrimp[0.075]Nimmo et al. 1979 

Daphnia [0.09]LeBlanc (1975) 

Daphnia[0.093]Surprenant 1988 

>0.1 to 1 Ceriodaphnia [0.25](Hall 1986) 

mayflies, damselflies, and dragonflies[0.7]

 (O'Halloran and Liber 1995) 

mixed insects3

 [1](Liber 1995) 

Ceriodaphnia [0.5](Hall 1986) 

clodacera and copopods [0.7] 3

 (Liber and O'Halloran 1995) 

copepods [0.7-0.9](Wright et al. 1996) 

grass shrimp (Bionomics-EG&G 1975) 

grass shrimp [0.7](Tourat and Rao 1987) 

stream inverts 1

 (Hansen and Garton 1982a) 

stoneflies and mayflies[1] 

(Hansen and Garton 1982b ) 1 

>1 to 10 

Ostracoda [2.5](Liber and O'Halloran 1995) 3 

dipterans[10] (Hansen and Garton 1982a )1 

mixed insects3 [1.9](Liber 1995) 

Ostracoda [7](Liber and O'Halloran 1995) 3 

mayflies, damselflies, and dragonflies[2.5]

 (O'Halloran and Liber 1995) 

>10 to 100 coleoptera, oligochaetes, and gastropods [50] 1

 (Hansen and Garton 1982a) 

rotifers [30](Liber and O'Halloran 1995) 3 

>100 to 1000 clams [320](Surprenant 1989) 

1

 Macrocosm study 

2

 No molting during short term exposures 

3

 Litoral enclosures 

4

 Marginal signs of toxicity

5 Cited in U.S. EPA (1997a) as Beltsville Lab Test 2424. This study is not identified by MRID number or otherwise described. 

Tables - 21 
Table 4-10: Summary of field studies on the effects of diflubenzuron on aquatic invertebrates 1 

Range of Species [conc ppb](Reference) 

Application 

Rates (g/ha) No Adverse Effects Adverse Effects with 

Observed Recovery 

Adverse Effects with No 

Observed Recovery 

>0.1 to 1	 pond invertebrates [0.2] Ali et 

al. 1988 

>1 to 10	 shrimp, cyclops, and some 

cladocera (Bosmina), worms 

[3.7] (Ali and Mulla 1978a)

 worms [7.4] (Ali and Mulla 

1978a) 

>10 to 100	 rotifers [13](Colwell and 

Schaefer 1980) 

zooplankton mortality and 

insect emergence [1.8](Wan 

and Wilson 1977) 

daphnids and copepods [3.7] 

(Ali and Mulla 1978a) 

copepods, shrimp [7.4] (Ali 

and Mulla 1978a) 

cladocera, copepods [2.5 to 

10](Apperson et al. 1977) 

cladocerans, copepods and 

rotifers[10](Boyle et al. 

1996) 

cladocera [10.4](Lahr et al. 

2000) 

cladocera incl. Bosmina, 

copepods, [13](Colwell and 

Schaefer 1980) 

amphipods [3.7] (Ali and 

Mulla 1978a) 

amphipods, daphnids [7.4] 

(Ali and Mulla 1978a) 

shrimp [10.4](Lahr et al. 

2000)

1 The concentrations given in braces [] represent peak or typical concentrations shortly after exposure. In all 

cases, post-application concentrations will decline. See text for discussion. 

Tables - 22 
Figure 3-1: Overview of the In vivo and environmental metabolism of diflubenzuron (adapted 

from WHO 1996). 

Figure - 1 
Figure 3-2: Modeled concentrations of diflubenzuron (thick lines) and 4­

chloroaniline (thin lines) in ponds at an annual rainfall rate of 150 inches (see text for 

discussion). 

Figure - 2 
APPENDICES
 

Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Appendix 2: Laboratory and simulation studies on environmental fate of diflubenzuron and its 

metabolites 

Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its 

formulations 

Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its 

formulations 

Appendix 4: Toxicity of diflubenzuron and diflubenzuron formulations to birds 

Appendix 5: Toxicity of diflubenzuron to terrestrial invertebrates 

Appendix 6: Toxicity of diflubenzuron to fish 

Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Appendix 8: Toxicity of diflubenzuron to aquatic plants 
 

 

 

Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal	 Dose/Exposure Response Reference 

Oral 

Diflubenzuron 

Acute Oral 

Mouse and rat	 LD 50, technical grade > 4640 mg/kg WHO 1996 

Mouse and rat	 LD 50,90% concentrate > 5000 mg/kg WHO 1996; 

U.S. EPA 

1997a 

Mouse and rat	 LD 50, Du 112307 W.P. 25% > 40,000 mg Dimilim/kg Koopman 1977 

(Dimilin WP 25%)	 > 10,000 mg DFB/kg MRID 

A marginal effects on methemoglobin 00070025 

levels. 

Rat, Sprague- single gavage dose of 5000 No mortality. Except for moist rales in Blaszcak 1997a 

Dawley, 5 mg/kg Dimilin 2L (22.36% two treated rats on the day of dosing, no MRID 

males(290-330 pure) clinical signs of toxicity, all rats gained 44574504 

g) and 5 weight both 7 and 14 days after dosing, 

females (215- and no abnormalities observed during 

233 g), 9- to macroscopic postmortem evaluation. 

12-weeks old 

NOEC = 5000 mg/kg as Dimilin 2L 

1118 mg/kg as DFB 

Subchronic Oral 

Cat (NOS)	 0, 30, 100, 300, or 1000 NOEC (Hb) >1250 mg/kg/day Keet et al. 1982 

mg/kg/day diflubenzuron for NOEC (%PCV) not estimated 

3 weeks NOEC (RBC) not estimated 

NOEC (reticulocyte count) not estimated 

NOEC (MetHb) = 30 mg/kg/day 

NOEC (SulpHb) = 3 mg/kg/day 

(calculated with regression analysis) 

NOEC (spleen weight) >1000 mg/kg/day 

Appendix 1-1 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response	 Reference 

Dogs, beagle, dietary levels of 10, 20, 40 or 

pure-bred, 15 60 ppm (actual dosages 

males and 15 levels of 0.42, 0.84, 1.64, or 

females 6.24 mg/kg/day) Du 112307 

for 13 weeks 

Dog (NOS)	 0, 2, 10, 50, or 250 

mg/kg/day diflubenzuron for 

13 weeks 

No mortality; no clinical signs of toxicity, 

no adverse effects on food or water 

consumption, no ocular effects, no 

treatment-related macroscopic post 

mortem findings, no adverse effects on 

organ weights, and no morphological 

abnormalities considered to be treatment 

related. 

At 2 weeks, all laboratory tests were 

within normal limits; 

at 4 and 6 weeks, SAP and SGPT were 

increased among some dogs at 40 or 160 

ppm; 

after 6 weeks, the presence of 

methaemoglobin and other abnormal 

haemoglobin pigments was apparent in 

dogs at 160 ppm; 

after 12 weeks, one dog at 160 ppm had an 

elevated SGPT level and one dog at 160 

ppm and one dog had a greater 

methaemoglobin value than all the other 

dogs. 

NOEC = 20 ppm 

NOEC (Hb) = 10 mg/kg/day 

NOEC (%PCV) not estimated 

NOEC (RBC) >250 mg/kg/day 

NOEC (reticulocytes) = 50 mg/kg/day 

NOEC (MetHb) = 50 mg/kg/day 

NOEC (SulpHb) = 10 mg/kg/day 

NOEC (spleen weight) not estimated 

Chesterman et 

al. 1974 

MRID 

00038706 

Keet et al. 1982 

Appendix 1-2 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

Mice, 

40/sex/dose 

group 

Rats, Swissalbino, males, 

weighing 90 g, 

5/dose group 

Rats, SpragueDawley, 40/sex/ 

dose group 

in diet concentrations of 0, 

80, 400, 2000, 10,000, or 

50,000 ppm 97.2% pure, 

technical grade, air-milled 

diflubenzuron for 14 weeks 

with 7-week interim 

sacrifice. 

The calculated mean intake 

of diflubenzuron was 9.7, 

50.7, 240, 1174, or 6114 

mg/kg/day (males) and 11.1, 

54.9, 288, 1393, or 7506 

mg/kg/day (females) [cf page 

27] 

gavage doses of 96.7 mg/kg 

of Dimilin in corn oil 

solution each day for 48 days 

(i.e., total of 4640 mg/kg of 

Dimilin) 

in diet concentrations of 160, 

400, 2000, 10,000, or 50,000 

ppm technical grade 

diflubenzuron for 90 days 

No treatment-related mortality throughout 

the study; no significant, treatment-related 

changes in food consumption or body 

weight; numerous hematological effects, 

including statistically significant increases 

(see pg 29) in Met Hb% and Sulph Hb% in 

males and females at 400-50,000 ppm; 

statistically significant increase in spleen 

weight in males and females at 400-50,000 

ppm; statistically significant increase in 

liver weight of males and females at 2000­

50,000 ppm; 

Mean hemoglobin concentration (g/100 

mL blood) was significantly lower than 

that of controls; mean hematocrit percent 

of the Dimilin was significantly higher 

than that of controls. 

No mortality; no clinical signs of toxicity, 

no adverse effects on body weight or food 

consumption. 

Treatment-related adverse effects included 

a significant increase in methemoglobin at 

weeks 7 and 13 in males at 400, 2000, 

10,000, and 50,000 ppm and in females at 

all dose levels, as well as significant 

increases in sulfhemoglobin at week 7 in 

50,000 ppm males and 10,000 and 50,000 

ppm females, and at week 13 in males at 

10,000 and 50,000 ppm and in females at 

2000, 10,000, and 50,000 ppm. 

Other pathological, treatment-related 

changes included decreases in hematocrit 

and hemoglobin values and the erythrocyte 

count and an increase in the number of 

reticulocytes, increases in absolute liver 

weight and absolute and relative spleen 

weights, and enlargement of the spleen. 

NOEC (for males only) = 160 ppm 

Colley et al. 

1981 

MRID 

00114330 

Berberian and 

Enan 1989 

Burdock et al. 

1980 

MRID 

00064550 

Appendix 1-3 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response	 Reference 

Sheep (NOS)	 0, 25, 125, or 500 mg/kg/day 

diflubenzuron for 13 weeks. 

Sheep	 0, 500, 2500 and 10,000 

mg/kg in feed for 13 weeks. 

NOEC (Hb) >500 mg/kg/day 

NOEC (%PCV) >500 mg/kg/day 

NOEC (RBC) >500 mg/kg/day 

NOEC (reticulocyte count) not estimated 

NOEC (MetHb) = 25 mg/kg/day 

NOEC (SulpHb) = 3 mg/kg/day 

(calculated with regression analysis) 

NOEC (spleen weight) >500 mg/kg/day 

No treatment-related effects were observed 

on food consumption, body weight gain, 

hematological parameters or urinalysis. 

Increase in MetHb and SulfHb and a 

reduction in the weight of the thyroid. 

Keet et al. 1982 

Ross et al. 

1977 

Appendix 1-4 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

Chronic Oral 

Dogs, beagle, 

5/sex/dose 

group 

daily oral administration of 

0, 2, 10, 50, or 250 

mg/kg/bw technical grade, 

air-milled diflubenzuron via 

There were no clinical signs of toxicity, no 

treatment-related effects on body weight, 

food consumption, or water consumption; 

no ocular effects; there were treatmentGreenough et 

al. 1985 

MRID 

00146174 

gelatin capsules, 7 days/week 

for 52 consecutive weeks. 

related marginal but statistically 

significant increases in met Hb% and 

sulph Hb% (at $10 mg/kg/day bw) and in 

Heinz body counts (at 50 and 250 

mg/kg/day bw); there was a marginal but 

consistent compound-related decrease in 

MCHC (at $10 mg/kg/day bw); 

histopathological changes included 

increased spleen weight (statistically 

significant in males at $50 mg/kg/day bw), 

increased liver weight (significant at $50 

mg/kg/day bw in males and females) and 

hemosiderin deposition in the liver. 

[This study is 

the basis for 

the chronic 

RfD] 

The investigators conclude: the no effect 

level demonstrated...was 2 mg/kg/day. 

However, this level is based on minor 

hematological changes of no toxicological 

significance seen at 10 mg/kg/day. Hence 

it is more realistic to consider the no effect 

level based on organ weights and 

histopathology as being at least 10 

mg/kg/day. 

Mortality: 2 females dogs died during the 

study. One dog at 250 mg/kg/bw) was 

sacrifice in extremis at week 33 due to 

liver failure and the other dog (at 50 

mg/kg/day bw) died during week 40 due to 

bronchopneumonia. These effects were 

not attributable to treatment. 

Appendix 1-5 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

Mice, CFLP, 

approximately 

8 weeks old, 

36/sex/dose 

group 

Mice, 

88/sex/dose 

group 

Rats, SpragueDawley, 

50/sex/dose 

group 

In diet concentrations of 16, 

80, 400, 2000, or 10,000 

ppm (intake values = 1.24, 

6.40, 32. 16, 163.29, or 

835.55 mg/kg/day for males 

and 1.44, 7.26, 35.38, 

186.59, or 958.51 

mg/kg/day for females) 

technical grade DFB (97.6% 

pure) for 91 weeks. 

in diet concentrations of 0, 

16, 80, 400, 2000, or 10, 000 

ppm 97.6% pure 

difllubenzuron for 91 weeks. 

The calculated mean intake 

of diflubenzuron was 1.24, 

6.40, 32.16, 163.29, or 

835.55 mg/kg/day (males) 

and 1.44, 7.26, 35.38, 

186.59, or 958.51 mg/kg/day 

(females) [cf page 47] 

in diet concentrations of 0, 

156, 625, 2500, or 10,000 

ppm technical grade 

diflubenzuron (97.6% a.i.) 

for 104 weeks. 

Treatment-related clinical sign of toxicity 

was a blue/gray discoloration of the 

extremities and dark eyes in all mice at 

10,000 ppm, a majority of mice at 2000 or 

400 ppm, and in a number of mice at 80 

ppm. The NOEC for this effect =16 ppm. 

No obvious treatment-related effect on 

mortality was observe; no obvious 

treatment-related effect on food 

consumption, body weight, food 

efficiency, or water intake was observed; 

treatment-related changes were principally 

associated with oxidation of the 

haemoglobin or with hepatocyte changes. 

DFB is not carcinogenic to DFLP mice. 

No treatment-related mortality throughout 

the study, no evidence of tumorigenic 

effect; treatment-related effects were 

primarily associated with oxidation of 

haemoglobin (treatment-related 

increases in Met Hb% were recorded 

from week 26 onwards and in Sulph 

Hb% from week 52 onwards; these 

changes principally affected mice at 80­

10,000 ppm and were dose-related in 

degree) or with hepatocyte changes (an 

increased incidenc of hepatocyte 

enlargement was observed in males and 

females at 400-10,000 ppm). 

No treatment related effects with regard to 

mortality or clinical observations; no 

evidence of carcinogenicity after 2 years of 

dietary exposure to diflubenzuron; 

statistically significant dose-related 

increases in met Hb% and sulph Hb% in 

males and females; numerous 

hematological effects; histomorphological 

changes observed in sections of the spleen, 

liver, and bone marrow; in general adverse 

effects were most pronounced at the 2500 

and 10,000 dose levels. 

Keet et al. 

1984b 

Colley et al. 

1984 

MRID 

00142490 

Burdock 1984 

MRID 

00145467 

Appendix 1-6 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

Rats, SpragueDawley, 

approximately 

7 weeks old, 

50/sex/dose 

group 

In diet concentrations of 156, 

625, 2500, or 10,000 ppm 

(intake values =6.99, 28, 36, 

114.35 or 463.80 mg/kg/day 

for males and 9.23, 37.98, 

153.96, or 633.41 

mg/kg/day for females) 

technical grade DFB (97.6% 

pure) for 104 weeks. 

No treatment related clinical signs Keet et al. 

observed; no obvious treatment-related 1984a 

effect on mortality; no obvious treatmentrelated effect on food consumption or 

body weight, except in high dose females 

where terminal body weight was 

significantly less than controls; no 

evidence of tumorigenic effects, treatmentrelated changes were principally 

associated with oxidation of haemoglibin 

or with hepatocyte changes. 

DFB is not carcinogenic to SpragueDawley CR-CD rats. 

Appendix 1-7 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

Reproduction Studies 

Rats, 

Crl:CD(SD)BR 

Rabbits, New 

Zealand White 

Rats, Charles 

River 

32/sex/dose 

group 

0 or 1000 mg/kg bw per day 

on days 6–15 of gestation 

0 or 1000 mg/kg bw per day 

on days 7–19 of gestation 

in diet nominal 

concentrations of 0, 500, 

5000, or 50,000 ppm 

technical diflubenzuron 

through two consecutive 

generations. 

F generation mean intake 

values (weeks 1-10 premate) were 36.2, 360, or 

3755 mg/kg/day for males 

and 42.0, 427, or 4254 

mg/kg/day for females. 

0

F generation mean intake 

values (weeks 5-16 premate) were 39.2, 394, or 

4089 mg/kg/day for males 

and 44.9, 473, or 4611 

mg/kg/day for females 

1

Screening assay for teratogenicity. No 

signs of developmental toxicity, birth 

defects or maternal toxicity. 

Screening assay for teratogenicity. No 

signs of developmental toxicity, birth 

defects or maternal toxicity. 

No treatment-related morality; toxicity 

manifested as hematological effects 

characterized primarily by anemia and 

increases in MetHb% associated with 

increased spleen weight and pathological 

lesions of hemosiderosis of the spleen and 

brown pigmented Kupffer cells in the liver 

were observed all dose levels. Increases in 

MetHb ranged from about 115% in the low 

dose group to over 300% in the high dose 

group (see Section 3.3 for more complete 

discussion and details). 

Other treatment related effects on the 

parental rats included lower body weight 

gains of the F generation at 50,000 ppm, 

with higher food intake values in males; 

increased water consumption among males 

and females at 5000 or 50,000 ppm and 

among males at 500 ppm. 

0

No treatment-related effects on 

reproductive performance at any dose 

level. In the F generation, liter and mean 

pup weights of the offspring from parents 

in the 50,000 dose group were lower than 

controls. The effect was not observed in 

1 

the F offspring. 2

NOEL = 50,000 ppm for reproductive 

function 

NOEL = 5000 ppm for pre-weaning 

development of the offspring. 

NOEL = >500 ppm for MetHb 

Kavanagh 

1988a 

Kavanagh 

1988b 

Brooker 1995 

MRID 

43578301 

NOTE: U.S. 

EPA (1996) 

appears to 

classify the low 

dose group as 

the LEL for 

MetHb but 

specifies the 

dose as 25 

mg/kg/day. 

This error 

appears to be 

based on the 

use of default 

values for 

converting food 

concentrations 

to mg/kg/day 

doses. 

Appendix 1-8 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

DERMAL 

Rabbits, New 

Zealand white, 

5 males and 5 

females 

Rabbits, New 

Zealand white, 

4 males and 2 

females, young 

adults, 2.2-2.6 

g 

Guinea pigs, 

Dunken 

Hartley, 10/sex 

Rats. Charles 

River, 

10/sex/dose 

group, weight = 

284-314 g 

(males) and 

201-233 g 

(females) 

Dermal application of 5000 

mg/kg Dimilin 2L (22.36% 

pure) to closely clipped 

intact trunks (approximately 

10% of the body surface 

area). Treated area covered 

with gauze and occlusive 

wrap for 24 hours. 

Dermal application of 0.5 

mL Dimilin 2L (22.36% 

pure) to intact skin of backs 

(hair closely clipped). Test 

site was semi-occluded with 

gauze for 4 hours 

Induction dose of 

approximately 0.3 mL 

Dimilin 2L (22.36% pure) 

for 6 hours; challenge dose 

after 14 days with 100% test 

material 

Dermal application of 20, 

500, or 1000 mg/kg/day 

Dimilin (technical 

diflubenzuron) to shaved 

intact skin for 21 days. 

No mortality; no pharmacological or 

toxicological signs of toxicity; no severe 

dermal effects; no abnormalities observed 

during postmortem macroscopic 

evaluation. 

NOEC = 5000 mg/kg 

4/6 rabbits had slight, barely perceptible, 

erythema; 1/6 had slight erythema; 1/6 had 

no signs of dermal irritation. 

Dimilin 2L considered slightly irritating 

(FIFRA Primary Irritation Index = 0.5) 

No dermal sensitization responses during 

induction or challenge phase. 

No treatment-related effects on survival, 

clinical signs of toxicity, dermal 

observations, body weights, food 

consumption or macroscopic and 

microscopic pathology. 

Females in the 500 and 1000 mg/kg/day 

group had mild but statistically significant 

decreases in mean erythrocyte counts, 

hemoglobin, and hematocrit values; males 

in the 1000 mg/kg/day group had mild but 

statistically significant decreases in mean 

hemoglboin and hematocrit values. At 500 

and 1000 mg/kg/day, males and females 

had an increased incidence of 

polychromasia, hypochromasia, and 

anisocytosis. At 1000 mg/kg/day, males 

and females had mild but statistically 

significant increases in Met Hb values and 

males also had mildly increased Sulph Hb 

values. 

NOEL = 20 mg/kg/day. 

Blaszcak 1997b 

MRID 

44574505 

Blaszcak 1997d 

MRID 

44574508 

Blaszcak 1997e 

MRID 

44574509 

Goldenthal 

1996 

MRID 

43954100-01 

Appendix 1-9 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

Rats, SpragueDawley, males, 

12/dose group 

EYES 

Rabbits, New 

Zealand white, 

6 

INHALATION
 

Rats, SpragueDawley, 

approximately 

6-weeks old, 

10/sex/dose 

group 

single dermal applications of 

14C-diflubenzuron suspended 

in 0.25% (w/v) gum tragacan 

WLC-grade water at 0.005 or 

0.05 mg/cm2

 to shaved skin 

for periods of 1, 4, and 10 

hours. 

0.1 mL Dimilin 2L instilled 

in lower conjunctival sac of 

the right eye of each rabbit. 

Observations for ocular 

irritation made at 1, 24, 48, 

and 72 hours. 

Nose-only exposure to 0, 10, 3 30, or 100 mg/m Dimilin

technical 6 hours/day, 5 

days/week for 4 consecutive 

weeks. 

> 89% of the applied dose was removed by 

washing; 6% of the applied dose was 

found in the skin and increased exposure 

time did not increase the percent of dose 

found in the skin, although the amount of 

test material found in the skin was nearly 

proportional to dose; blood, carcass, and 

excreta accounted for negligible amounts 

of the applied dose; systemic absorption, 

excluding the skin was <1% of the total 

applied dose. These data indicate that 

the material that was absorbed was 

absorbed quickly. and the percent of 

applied dose that was absorbed appeared 

to be constant regardless of dose. 

Positive scores ( slight to moderate 

conjunctival irritation) in 3/6 rabbits 

within 24 hours of exposure with full 

recovery within 48 hours. No signs of 

iridial or corneal changes. The remaining 

3 rabbits did not have positive scores for 

ocular irritation at any time during the 

study. 

Study demonstrates that Dimilin 2L is an 

“eye irritant” based on the results of 

positive scores in 3/6 animals with all 

changes being reversible. 

Dimilin technical produced minimal 

toxicity, including a slight (5-7%) decrease 

in erythrocytes, slight statistically 

significant decreases in hemoglobin and 

hematocrit in males and females at 100 

mg/m3

 and an increase in bilirubin in 

males at 100 mg/m .3 No treatment-related 

effect observed on methemoglobin. 

NOEC = 30 mg/m3 

Andre 1996 

MRID 

44053101 

Blaszcak 1997c 

MRID 

44574507 

Eyal 1999 

MRID 

44950601 

Appendix 1-10 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

Rats, SpragueDawley, 9 

weeks old, 5 

males (323-335 

g) and 5 

females (234­

249 g) 

Rats, SpragueDawley, 20 

males and 20 

females, 

5/sex/dose 

group 

4-hour nose only exposure to 

2.0 mg/L Dimilin 2 L 

(22.36% pure) with 14-day 

post exposure observation 

period 

Whole body exposure to 

nominal concentrations of 

0.5, 5.0, or 50 mg/L air 5 

days/week for 3 weeks. 

Corresponds to 500, 5000, 

and 50,000 mg/m3

 – i.e., 

1000 L = 1 m .3 

No mortality; signs of toxicity during 

exposure included red nasal discharge and 

labored breathing; chromodacryorrhea, red 

nasal discharge, and excessive salivation, 

labored breathing, and moist rales were 

observed in some rats up to 1 day after 

exposure with complete recovery 

thereafter; slight weight loss was observed 

in some females during the first week after 

exposure followed by complete recovery 

during the second week; no abnormal 

macroscopic effects were observed during 

postmortem evaluation. 

LC50 >2.0 mg/L 

No signs of irritation at 0.5 mg/L; frequent 

blinking and occasional bouts of persistent 

sneezing and slightly labored breathing 

during exposures to 5.0 mg/L, followed by 

rapid recovery between exposures; at 50 

mg/L, the signs observed in the mid-dose 

group were more pronounced and more 

persistent but repeated exposure did not 

result in cumulative adverse effects and 

recovery was rapid after each exposure 

period. 

No changes in body weight, compared 

with controls and no effects on water or 

food consumption were observed. 

Post-exposure methaemoglobin levels 

were increased 0.2-0.5 g% over controls 

(0.1 g%). The increase was statistically 

significant in the mid and high-dose males 

and in all treated females. 

Hoffman 1997 

MRID 

44574506 

Berczy et al. 

1975 

MRID 

00044325 

Appendix 1-11 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

Rats, SpragueDawley, males 

and females, 6­

weeks-old, 

10/sex/dose 

group 

Rats, Wistar, 

SPF albino, 

males and 

females, 

10/sex/dose 

group 

Nose-only exposure to 0, 10, 

30, or 100 mg/m3

 (measured 

as 12, 34, or 109 mg/m3 

diflubenzuron technical 

(95.6% purity) 6 hours/day, 5 

days/week for 4 weeks. 

daily oral doses of 0, 8.0, 

20.0, or 50.0 mg/kg 

4-chloroaniline (4-CA) for 3 

months 

Minimal toxicity: slight decrease (5-7%) in 

erythrocytes, slight statistically significant 

decreases in hemoglobin and hematocrit in 3 males and females at 100 mg/m ; increase

in bilirubin males at 100 mg/m . 3 A 

reduction in ‘grid count’ was evident in a 

neuro-functional assessment at the highest 

concentration. 

No effect observed on methemoglobin. 

NOEC = 30 mg/m3 

4-chloroaniline 

All rats at 50 mg/kg had increased 

numbers of Heinz bodies (>20/100 RBC) 

and a reticulocyte response (>2%); 

however there was no evidence of a 

decrease in hemoglobin, packed cell 

volume, or RBC count. 

Histological changes were observed only 

in the high dose group and included 

increased extramedullary haematopoiesis 

in spleen and liver and occasionally in the 

lung; increased hemosiderin (from 

hemoglobin breakdown) in the liver and 

spleen and occasionally in the kidneys 

(epithelium of proximal convoluted 

tubules). 

NOEC = 8.0 mg/kg 

Newton 1999 

MRID 

44950601 

Scott and 

Eccleston 1967 

Appendix 1-12 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

Dog, Beagle, 

males and 

females, 

4/sex/dose 

group 

Rats, Fischer 

344, males, 

10/dose group 

daily oral doses of 0, 5, 10, 

or 15 mg/kg 4-chloroaniline 

(4-CA) for 3 months 

In diet concentrations of 

1240 ppm 4-chloroaniline or 

1240 or 4320 ppm pchlorophenylurea for 7 days 

One dog in the 15 mg/kg dose group died 

as a result of excessive haemolysis (after 

receiving 25 mg/kg 4-CA). 5/7 remaining 

dogs receiving 15 mg/kg 4-CA showed an 

early and marked decrease in RBC count 

(>1.5 M) and packed cell volume (>15%) 

with a concomitant decrease in 

hemoglobin levels. The same trend was 

observed in half the dogs at 10 mg/kg and 

one of the dogs at 5 mg/kg. 

Lowest levels of RBC and hemoglobin 

were reached at approx 3-4 weeks, after 

which time, there was a slow but steady 

improvement in all values, despite the 

persistence of increased numbers of Heinz 

bodies. A reticulocyte response and an in 

crease in Heinz bodies were observed in 

all dogs at 15 mg/kg, most dogs at 10 

mg/kg, and three dogs at 5 mg/kg, while 

the control group remained normal. 

All treated dogs showed histological 

changes, including evidence of 

hematopoietic response in extramedullary 

activity in spleen and liver at all doses 

(The marrows showed hyperplasia of the 

erythroid phase) and marked evidence of 

RBC destruction in the spleen, and liver. 

1240 ppm 4-chloroaniline caused 

statistically significant increases in 

methemoglobin values at all intervals of 

analysis 

No treatment related effects on 

methmoglobin values in rats treated with 

1240 or 4320 ppm p-chlorophenylurea. 

The only macroscopic change observed 

was enlargement of the spleen in rats from 

the 1240 ppm 4-chloroaniline group. 

No mortality. 

Scott and 

Eccleston 1967 

Goldenthal 

1999b 

MRID 

44871303 

Appendix 1-13 
Appendix 1: Toxicity of diflubenzuron and 4-chloroaniline to experimental mammals 

Animal Dose/Exposure Response Reference 

Rats, Fischer 

344, approx 6­

weeks old, 

males and 

females, 

25/sex/dose 

group 

Mice B63CF1, 

approx 6­

weeks-old, 

males and 

females, 

25/sex/dose 

group 

In diet concentrations of 250 

or 500 ppm 4-chloroaniline 

for 78 weeks with a 24-week 

observation period. 

In diet concentrations of 

2500 or 5000 ppm 

4-chloroaniline for 78 weeks 

with a 24-week observation 

period 

Mean body weight depression associated 

with treatment was observed in high dose 

females, compared with controls. 

No significant treatment-related mortality 

among females; however, there was a 

significant (p=0.0294) correlation between 

dose and mortality in males rats. 

In the high dose male rates, the incidence 

of unusual splenic neoplasms (i.e., 

fibroma, fibrosarcoma, sarcoma, 

hemangiosarcoma, and osteosarcoma) was 

increased (0/20 controls; 0/49 low dose, 

10/49 high dose). This finding was 

considered strongly suggestive of 

carcinogenicity because of the rarity of the 

rumors in the spleens of controls rats. 

Formation of non-neoplastic lesions of the 

splenic capsule in rats in all dose groups. 

Mean body weight depression associated 

with treatment was observed in all mice, 

compared with controls. 

No significant treatment-related mortality 

in mice of either sex. 

NCI 1979 

NCI 1979 

Appendix 1-14 
Appendix 2: Laboratory and simulation studies on environmental fate of diflubenzuron and its metabolites. 

Data Summary Reference 

Aquatic Sediments 

anaerobic aquatic metabolism of 1.3 mg/kg 14-C diflubenzuron in silt loam/water 

system. 

DT50 = 34 days for total hydrosoil/water system and 18 days for water-phase only. 

2,6-difluorobenzoic acid, and 4-chlorophenylurea were main metabolites that
 

accumulated in the anaerobic water phase; hardly any bound residue detected.
 

two model ditch (water/sediment) systems (sandy loam or silt loam covered with
 

surface water) with addition of 0.94 ppm. diflubenzuron with continuous flow through
 

upper layer of surface water. Incubation at 20±1° w/12 hour photo period.
 

Results indicate rapid disappearance of compound from model ditch systems due to
 

rapid metabolism and adsorption to sediment.
 

Water phase DT50 = 1.1 day (silt loam) and 1.9 days (sandy loam).
 

Complete sediment/water systems DT50 = 10 days (silt loam/surface) and 25 days
 

(sandy loam/surface).
 

Only metabolites were DFBA and CPU
 

0.013 ppm DFB in a microbially viable soil/water test system 

DFB was readily degradable under aerobic aquatic conditions 2 half-life (first-order kinetics) = 25.7 days (r =0.709)

DT50 = 5.3 days 

major metabolite formed, 4-chlorophenylurea 2 half-life (first-order kinetic) = 39.7 days (r =0.671)

degradation rate of 50 μg/L diflubenzuron in seawater in the presence of salmon feces 

and sediment is temperature dependent: at 15°, DT50 = 3 1⁄2 weeks (anaerobic) or 4 1⁄2 

weeks (aerobic); however at 5°C, there was no significant difference between the 

anaerobic (DT50 = 99 days) or the aerobic (DT50 = 100 days) test conditions. 

The metabolites included 4-chlorophenylurea, 2,6-difluorobenzoic acid, 2,6­

difluorobenzamide, and CO2 

Thus et al. 1991 

MRID 41837601 

Thus and van der LaanStraathof 1994 

MRID 44399307 

Dzialo and Maynard 

1999 

MRID 44895001 

van der Laan 1995 

In: Technology 

Sciences Group 1998 

MRID 44399307 

Appendix 2 - 1 
Appendix 2: Laboratory and simulation studies on environmental fate of diflubenzuron and its metabolites. 

Data Summary Reference 

laboratory microcosm study using 10 μg/L DFB in seawater with or without sediment. Wilson et al. 1995 

half-life of DFB in seawater without sediment = 18.7 days 

half-life of DFB in seawater with sediment = 5.2 days 

presence of organic sediment in DFB-treated microcosm significantly reduced the 

efficacy of DFB in seawater as measured by toxicity of aged DFB (initial nominal 

concentration of 10 μg/L) to 5-day old grass shrimp embryos. By day 30, embryos 

reared in seawater from DFB-sediment microcosm produced larvae with no signficant 

morphological abnormality and larval viability was comparable to controls; embryos 

reared in DFB-treated seawater without sediment produced larvae with severe 

abnormalities and very low viability even after the seawater aged for 65 days. 

persistence of diflubenzuron (Dimilin) in sod-lined water pools after repeated Madder and Lockhart 

applications: 1980 

Bioassay data indicate toxicity greatest during the first 24 hours; DFB fell below 

detection limits (1μg/L) within 24 hours, whereas chlorophenylurea concentration 

increased for several days after treatment. 

Bioconcentration 

Channel catfish, Ictolurus punctatus: No bioconcentration. In 0.01 ppm tanks, Booth and Ferrell 1977 

concentration in muscle was below 0.002 ppm and concentration in viscera peaked at 

about 0.003 ppm (Figure 2). Similar pattern in 0.5 ppm tanks (Figure 3). 

Algae: BCFs of 2412 at hour 1 to 109 at day 4. Probably reflects degradation – i.e., 

algae degraded 80% of the DFB in a 1-hour incubation period. 

Laboratory algae culture system of Scenedesmus subspicatus exposed at an initial Yu-Yn et al. 1993 

concentration of 200 μg/L DFB for 7 days 

no growth inhibition; half-life = 3 days 

DFB radioactivity in algae increased steadily and leveled off at approx. 60% after 5 

days 

BCF values decreased from 4310 to 889 during the exposure period 

elimination was rapid during the first hours. 

Hydrolysis 

rapid decrease in of residue levels. Anton et al. 1993 

Half-life w/aeration = 0.41899 days (tap water and natural sunlight) 

Half-life wo/aeration = 0.96685 days (tap water and natural sunlight) 

Appendix 2 - 2 
Appendix 2: Laboratory and simulation studies on environmental fate of diflubenzuron and its metabolites. 

Data Summary Reference 

Two applications of Dimilin 25W (25% a.i. by weight) to surface of littoral enclosures Knuth 1995 

using portable hand sprayer at rates of 4-210 g/ha. MRID 44386201 

(This is chapter 2 of 

Maximum residues in water column measured withing first 24 hours after application, Moffet 1995) 

Half-lives ranged from 3.28 to 8.23 days with a mean of 4.28 days. By 14-35 days (or 

a mean of 18.5 days), 95% of the diflubenzuron dissipated. Principal loss from water 

column early in the study probably due to adsorptive processes because temperature 

and pH were not favorable for rapid aqueous hydrolysis. 

11 μg/L 14C-diflubenzuron in a CO 2-evolution test (concentration below aqueous van der Laan and Thus 

solubility). 1993 

In: Technology 

DT50 = approximately 2.5 days; hydrolysis products are DFBA, CPU, and CO 2 Sciences Group 1998 

MRID 44399307 

High temperature (121/C) increased the degradation of diflubenzuron in aqueous Ivie et al. 1980 

media at levels greatly above its solubility in water and resulted in its rapid 

degradation to as many as seven identified products: 4-CPU, 2,6-DFBA, 2,6-difluorobenzamide,

 4-chloroaniline, N,N'-bis (4-chlorophenyl) urea, 1-(4-chlorophenyl)- 5-fluoro-2,4 

(1H,3H)-quinazolinedione and 2-[(4-chlorophenyl) amino]- 6-fluorobenzoic acid. 

4-Chloroaniline, N,N'-bis (4-chlorophenyl) urea and 2[(4-chlorophenyl) 

amino]-6-fluorobenzoic acid were not detected at lower temperatures (0.1 mg [14C]­

diflubenzuron/L water or buffer at 36/C). 4-Chloroaniline was a major degradation 

product of diflubenzuron in heat-treated samples, but it was not seen at lower 

temperatures 

Photolysis 

Photodegradation half-lives of diflubenzuron in deionized water (pH 7) = 17 hours; in Marsella et al. 2000 

deionized water (pH 9) = 8 hours; and in river water (pH 9) = 12.3 hours. 

In a solar simulator using river water buffered to pH 9.0, the half-life for diflubenzuron 

=12 hours; dark controls showed no loss of parent compound over similar time 

periods. 

Log Kow = 3.8 (determined using reverse phase HPLC) 

Appendix 2 - 3 
 

Appendix 2: Laboratory and simulation studies on environmental fate of diflubenzuron and its metabolites. 

Data Summary Reference 

Residues on Plants 

persistence of diflubenzuron (commercial grade 25% WP) on Appalachian forest 

leaves. 

Leaves sprayed in spring and left to weather during growing season. 

white oak leaves collected in July and August and placed in headwater stream to 

monitor residual diflubenzuron showed rapid decrease in residue (36% in July and 

23% in August) within the first 48 hours of stream incubation, reaching less than 10% 

of the original concentration within 3 weeks. 

Yellow poplar, read maple, and white oak leaves collected in December and place in 

headwater stream showed a much slower rate of loss. After 54 days in the stream, 

yellow poplar and red maple leaves retained 45 and 40%, respectively of the original 

diflubenzuron concentration and white oak showed no significant loss. 

Stream water temperatures averaged 17°C lower in December than in August 

(temperature readings were not made in July). 

Soil Degradation/Transport 

fate of 4-chloroaniline in nonautoclaved and autoclaved soil.
 

in soil treated with 4-chloroanline and incubated for 6 weeks, no CO2 evolution in
 

occurred in autoclaved soil; in nonautoclaved samples, CO2 was determined as 7.5% of
 

the originally applied radioactivity.
 

Cell suspension of 0.04 g Pseudomonas putida in 2 mL of 0.05 M phosphate buffer
 

(pH 7.0) incubated with 10 μg 14C-PH-6040 (DFB) (both A and B labels) for 6 hours
 

produced no evidence of degradation upon extraction. Both labeled preparation were
 

recovered intact as 99.9+% of total 14C 


10 ppm 14C-PH-6040 (DFB) added to fresh, air-dried Drummer soil (17.4% moisture)
 

and incubated at 80°F for 1, 2, or 4 weeks.
 

Compound appeared to be very stable, with degradation products comprising only
 

0.7% of total extracted radioactivity after 4 weeks. 


aerobic soil metabolism of 0.69 mg/kg 14-C diflubenzuron in sandy loam 

DT50 = 50 hours; DT90 = 181 hours 

CO , 2,6-difluorobenzoic acid, and 4-chlorophenylurea were main metabolites; 2,6­

difluorobenzamide and 4-chloroaniline were minor metabolites. 

2

Harrahy et al. 1993 

Bollag et al. 1978 

Metcalf et al. 1975 

Metcalf et al. 1975 

Walstra and Joustra 

1990 

MRID 41722801 

Appendix 2 - 4 
Appendix 2: Laboratory and simulation studies on environmental fate of diflubenzuron and its metabolites. 

Data Summary Reference 

10 ppm technical DFB applied on quartz sand to natural sandy loam and much soils Chapman et al. 1985 

at 12 weeks: 2-12% remaining (compared with 80-87% remaining in sterilized soil), 

indicating that soil microorganisms play a major role in the degradation of DFB. 

Kinetic analysis based on 1st order dependence indicates that the rate constants for 

disappearance reactions decreased with time. 

breakdown of DFB by soil isolates: Seuffer et al. 1979 

Rhodotorula sp. half-life of detectable DFB = 18 days (carbon source: acetone) 

Fusarium sp. half-life of detectable DFB = 7 days (carbon source: DFB/acetone) 

Penicillium sp.half-life of detectable DFB = 14 days (carbon source: acetone) 

Cephalosporium sp.half-life of detectable DFB = 13 days (carbon source: acetone) 

Control half-life of detectable DFB = 27 days. 

14C-DFB readily degraded in various agricultural soils and in hydrosoil: 50% of Nimmo et al. 1984 

applied dose (1 mg/kg) metabolized in #2 days. Chief products of hydrolysis were 4­

chlorophenylurea and 2,6-difluorobenzoic acid. 

initial dose of 1 mg/kg 4-chlorophenylurea in decreased to 50% in about 5 weeks in Nimmo et al. 1986 

aerobic sandy clay and in about 16 weeks in anaerobic hydrosoil 

Investigators assume that two sorts of bound residues are formed from 4-chlorophenylurea: one is fairly stable and might consist of bound 4-chloroaniline or its 

transformation products and the other is presumed to be a degradable derivatie of 4­

chlorophenylurea. 

2-6-difluorobenzoic acid is rapidly and completely degraded in soil: time to 50% Nimmo et al. 1990 

disappearance in 9 days in humus sand and after 12 days in sandy clay. DFBA 

completely disappeared in the humus sand after 32 days. 

DFB (technical), Dimilin WP-25, and Dimilin SC-48 were applied separately at 70, Sundaram and Nott 

210, or 630 g ai./ha (corresponding to 17.23, 51.69, or 155.07 μg a.i.) To top layer of 1989 

columns (30x5.6 cm id) packed either with sandy or clay loam forest soils. 

Mobility of DFB was low and did not increase with dosage. At deposit rate equivalent 

to 70 g a.i./ha, nearly all the residues were found with 2.5 cm of the top of the column. 

At 630 g a.i./ha, only about 9% of the technical DFB, 7% of Dimilin SC-48, and 4% of 

Dimilin WP-25 moved below the 2.5 cm level in sandy loam. 

No residues were found below the 10 cm level or in the leachates in either soil type at 

all dosage levels. 

In addition to soil type, mobility of DFB was also influenced by the additives present 

in the formulations with technical DFB > Dimilin SC-48 > Dimilin WP-25. 

Appendix 2 - 5 
Appendix 2: Laboratory and simulation studies on environmental fate of diflubenzuron and its metabolites. 

Data Summary Reference 

Organic soil and silty clay loam soil collected from a boreal forest in northern Ontario, Sundaram et al. 1997 

Canada 

maximum amount adsorbed: 88 μg/g (organic soil); 73 μg/g (silty clay loam) 

time required for maximum adsorption: 18 h (organic soil); 24 h (silty clay loam) 

Organic soil characterized as about equal parts sand, silt, and clay and 21% OM and 

13% OC. 

Silty clay loam characterized as 22% sand, 49% silt, and 29% clay, and 8.2% OM and 

5.1% OC. 

KD = 17.59 (organic soil) 

KD = 16.42 (silty clay loam) 

K = 135.3 (organic soil) oc

K = 332.0 (silty clay loam) oc

oc calculated K = 144.4 (organic soil) 

oc calculated K = 345.3 (silty clay loam) 

Appendix 2 - 6 
Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

33, 66, and 140 g a.i./ha (0.5, 1, and 

2 oz a.i./ac) 

No evidence of negative effects on 

predators/parasites – lacewing (Chrysopa carnea), 

ladybird beetle (Hippodamia convergens), Wasp 

parasite Trichograma pretiosum of bollworm 

(Heliothis). Immigration from untreated fields could 

mask negative effects on beetles seen in lab (see 

Appendix 5). 

Ables et al. 1977 

280 g a.i./ha (4 oz a.i./ac) Caged lacewing suffered increased mortal. eating 

treated eggs. No effect on parasitic wasp through 2­

3 generations; wasp developed in treated eggs and in 

eggs produced by treated adults, and direct 

exposure to adults was not toxic. 

Ables et al. 1977 

187 ppm spray to apple orchards 

(NOS) 

Application (spray) of Dimilin WP, 

0.6 kg in 600 L/ha to a 2.4 ha apple 

orchard (integrated pest management 

program). 250 g Dimilin/ha, 

62.5 g a.i./he 

No adverse effects in Phytoseiid and stigmaeid mites 

No population increases following treatment in 

European red and rust mites 

DFB persisted on foliage until leaf-fall and was 

detected on the peel of harvested fruit. Mean 

residue on harvested Worcester fruit = 0.05 mg/kg 

fresh weight and on harvested Cox fruit, mean 

residue = 0.02 mg/kg fresh weight. 

Anderson and 

Elliott1982 

Austin and 

Carter 1986 

4-year field study (1992-1995) in 

apple orchards in a codling moth 

control program based on 4 seasonal 

sprays/year. Diflubenzuron at 3-12 

g/100 L. Application rate in g/ha not 

specified.

Dimilin 4 liquid applied at 70 g 

a.i./ha to watersheds in a central 

Appalachian broadleaf forest 

 Spider fauna (26 genera and 30 identifiable spider 

species) in apple orchards of Western Oregon. DFB 

was harmless to spider species tested (p>0.05) 

Yellowjackets, (10 species of wasps, Family 

Vespidae): Diflubenzuron decreased worker number 

during application year but not in post application 

year. There was no effect of trap site on worker 

sample size. 

Bajwa and 

AliNiazee 2001 

Barrows et al. 

1994 

140 g a.i./ha (2 oz a.i./ac). 4.05 ha in 

41 ha woods. 

Some species of soil mites were adversely affected. 

Half the number in treated v. untreated samples. 

Blumberg 1986 

67 g a.i./ha (0.96 oz a.i./ac) Wasp parasite on Gypsy moth eggs (Ooencyrtus 

kuvanae) on gypsy moth. Egg masses in treated 

plots were parasitized as heavily as egg masses in 

control plots. Lab data showed no effect on 

development and emergence from treated eggs or 

from eggs laid by treated adults. 

Brown and 

Respicio 1981 

Appendix 3a - 1 
Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

350 g a.i./ha (5 oz a.i./ac) 5 hives (Honey bee, Apis mellifera.) in treated and 

untreated sites. No effects on egg hatch, brood 

production, numbers of adults, and honey 

production. 

Buckner et al. 

1975 

70 g a.i./ha (1 oz a.i./ac) Under tree bands, Carabidae (beetles), 

Gryllacrididae (grasshoppers), and two families of 

moths were significantly reduced in total taxa 

richness and abundance on treated sites. 

Butler 1993 

Additional Notes on Butler 1993: Foliage sampling found reduced abundance and richness in the following 

groups: Lepidoptera, Symphyta (sawflies, horntails), some herbivorous Coleoptera (beetles), Psocoptera (book 

lice, wood lice), predatory Thysanoptera (thrips), some Homoptera (leaf hoppers, aphids, cicadas), Diptera (flies), 

Orthoptera (grasshoppers), and Arachnida (spiders). Some affected by direct toxicity and others (predators) 

indirectly through prey reduction. 

Aerial application of Dimilin 4L Gypsy moth larvae decreased in number on the Butler et al. 

(35.1 g a.i./ha) to two watersheds in a treated watersheds, especially during the treatment 1997 

Central Appalachian forest; two and post-treatment year. Macro lepidoptera larvae 

untreated watersheds served as also decreased in number during the treatment year. Butler 1995 

controls. 

Additional Notes on Butler et al. 1997: In treated watersheds, there was an overall reduction in arthropod family 

diversity and abundance on foliage and a significant reduction in the number of macro Lepidoptera and beetles. 

27 months after treatment, total arthropod abundance and macro lepidoptera abundance on foliage remained 

significantly reduced. Decreases in the numbers of Carabidae (ground beetles), Gryllacrididae (crickets), 

Psocoptera (booklice/barklice), Phlaeothripidae (alligatorweed thrips), and some sapfeeders were observed but 

reductions were not significant. 

Aerial application of 0.0084, 0.0168, 

or 0.0336 kg a.i./ha Dimlin 2F [8.4, 

16.8, 33.6 g/ha] or 0.0168 kg a.i./ha 

Dimilin 25W [16.8 g/ha] to mixedgrass rangelands near Amidon, ND 

(experimental plots were 0.4x0.4 

km). 

Abundance of ants was not significantly reduced by 

treatment at any levels. Ant diversity declined 

temporarily (13-19 days) after treatment with 

Dimilin 25W, but recovered immediately the 

following week and no further declines were 

observed. Twenty species of ants were encountered 

in the experimental site. 

Catangui et al. 

1996 

Aerial application of diflubenzuron 

(25% WP) to treatment plots at a rate 

of 70.75 g/ha on May 8, 1985 and 

May 9, 1986 as part of Gypsy moth 

suppression program in WV. Plots 

were located in an 8000 ha oakhickory forest. Untreated plots 

served as controls. 

Abundance: No significant differences were 

observed (p<0.10) in abundance of 21 species of 

birds examined between treated and control plots. 

Diets: All species in untreated plots ate more 

Lepidoptera larvae than species on treated plot; 

difference was significant (p<0.10) in 5 of 7 species. 

Foraging: Vireo foraging areas were 3.1 and 2 times 

larger on treated areas, compared with untreated 

areas. 

Cooper et al. 

1990 

Appendix 3a - 2 
Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

70 g a.i./ha (1 oz a.i./ac) to cotton, 

applied in paraffinic crop oil 

(Dimoil) and water. Sampling took 

place 1 week after each treatment. 

Fields 15 ha each. 

Assay of populations of predators of bollworms 

(Heliothis): lacewings (Chrysopa spp.), ladybird 

beetle (Hippodamia convergens), Coleomegilla 

maculata big-eyed bug (Geocoris punctipes), Nabis 

spp., Orius insidiosus. Numbers of predators 

unaffected by 4 treatments 1 week apart. The study 

did not look at parasite numbers. The authors note 

that crop oil could have affected some species. 

Deakle and 

Bradley1982 

0.3 to 3.3 kg Dimilin 25W/ha [75 g 

a.i. to 825 g a.i./ha] 

No effect on breeding success or growth of nestlings 

for tree sparrows (Passer montanus) or two species 

of tits (Parus major and Parus caeruleus). 

Endpoints examined included number of occupied 

nest boxes, mean number of offspring, nesting 

period, mortality of nestlings, and breeding success. 

De Reed 1982 

110 to 400 g a.i./ha Honey bee, Apis mellifera. No effect from spray on 

trees on adults or larvae. 

Emmett and 

Archer 1980 

38 and 83 g a.i./ha Nearly 90% reduction in grasshoppers (nymphs and Everts 1990 

applied in diesel oil adults) 7 d. after treatment at higher rate. Low rate 

(0.54 and 1.19 oz a.i./ac) had minimal effects on larval grasshoppers. At least 

one taxon of beetle showed reductions of 50% at 

highest dose. Possible reduction in trap catches of 

members of 1 of 3 families (the Gnaphosidae) of 

ground spiders, at highest dose, 4 weeks after 

treatment. Reduced populations of Ichneumonids 

and Braconids in sprayed plots for at least 3 weeks. 

Possibly due to effects on host species rather than 

direct toxicity. Tiphiids unaffected by treatments. 

Predatory wasp reduced in treated plot, possibly a 

response to prey reduction (grasshoppers). 

Brazil: 250 g a.i./ha (3.6 oz a.i./ac). 

Applied 3x by mistblower. 

No effect on adult levels of predator Calosoma, nor 

on nabids or geocorids. 

Heinrichs et al. 

1979 

70 g a.i./ha (1 oz a.i./ac) in 4.7 l/ha 

crop oil (Savol) + H2O, applied 6x at 

5 d. intervals 

Treatments reduced parasitism by Trichogramma 

pretiosum to Heliothis spp. by 44% after spray. 

House et al. 

1980 

Apple orchard in Union, CT. 57 g 

a.i./10 gal water with spreader 

sticker. Applied with backpack 

sprayer. 

Parasitic wasp Apanteles melanoscelus Parasitism 

rate on treated vs. control trees roughly equal before 

spray, but lower on treated trees 7 d. after spray 

(1.81% v. 0.67%). Some adult wasps developed 

successfully, perhaps those in later stages of 

development. 

Granett and 

Dunbar 1975 

Appendix 3a - 3 
 

Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Apple orchard in Brooklyn, CT. 3.5 

g a.i./10 gal. water with spreader 

sticker. Applied w/ backpack 

sprayer. 

Parasitic wasp Apanteles melanoscelus 1st 

application of spray decreased parasitism rate. 2nd 

and 3rd applications did not. 

Granett and 

others1976 

About 11 and 22 g a.i./ha (0.75% and 

1.0% a.i./kg. At 1.1 and 2.2 kg/ha.) 

Treated bran bait. 

30+ spp. of grasshoppers, counted on treated and 

control fields. Total populations were reduced 28 

days after treatment by 60 and 70% at highest rates 

of application (0.75 and 1.0% a.i./kg; 2.2 kg/ha). 

Populations reduced <20% at half that rate. Greater 

effects early instars. 

Jech et al. 1993 

Cotton fields treated with nine 

applications of 2 oz a.i./acre (140 

g/ha) diflubenzuron (NOS) from 

June17-Aug12 

Monitoring of arthropod predator populations: 

Geocoris punctipes, Nabis spp., Hippodamia 

convergens, Coleomegilla maculata, Orius 

insidiosus, Chrysopa spp. Diflubenzuron treatment 

did not skew the relative abundance of the predators 

sampled. For 6 days after collection, egg hatch in 

the laboratory held H. convergens was significantly 

lower in females collected from treated cotton fields, 

compared with those form untreated cotton fields. 

Keever et al. 

1977 

Backpack application of 8 oz Dimilin 

25W or 0.5 pints Dimilin 2L (0.125 

lbs a.i./acre in each case) to maturing 

cotton foliage in Fresno, CA or East 

Bernard TX 

Over 5 weeks, dislodgeable foliar residue ranged 

from 0.40 μg/cm down to 0.01 μg/cm (limit of 2 2 

quantitation). Regression analysis predicted mean 

dislodge able residues on cotton leaves of 0.189 

μg/cm at 4 hours and 0.180 μg/cm at 24 hours at 2 2 

both locations. 

Korpalski 1996a 

MRID 

44081401 

Three applications of Micromite 

25W via calibrated airblast sprayer 

to orange trees at a rate of 1.25 lbs 

(0.3125 lbs a.i./acre) in LaBelle, FL. 

[0.35 g/ha × 3] 

Diflubenzuron at 150, 450, or 750 g 

a.i/ha. 

Over 5 weeks, dislodgeable foliar residue ranged 

from approximately 0.8 to 1.0 μg/cm2

 shortly after 

the last application and down to 0.22 to 0.48 μg/cm2 

at 35 days post application. Regression analysis 

predicted mean dislodgeable residues on orange tree 

leaves of approximately 0.59-0.82 μg/cm2

 at 4 hours 

and approximately 0l58-0.81 μg/cm2

 at 24 hours at 

both locations. 

Gram pod borer, Helicoverpa armigera (Lepidotera: 

Noctuidae) [crop pest] field collected eggs on gram 

plants in sprayed and unsprayed plots. % egg 

mortality: Controls = 13.0% ; 150 g/ha. =39.0%; 450 

g/ha. = 61.0%; 750 g/ha. = 100.0 % 

Korpalski 1996b 

MRID 

440814012 

Kumar et al. 

1994 

Appendix 3a - 4 
Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Site 1: 140 g a.i./ha (2 oz a.i./ac) Mites counted in the top 6 cm of soil. About half of Marshall1979 

Site 2: 280 g a.i./ha (4 oz a.i./ac) the taxa showed significant decreases in abundance 

both in Kamloops, British Columbia from diflubenzuron applications. Overall population 

unaffected by spraying; increases in some species 

compensated for decreases in others. Mites in upper 

3 cm of soil more severely affected than mites 

below. Some predators decreased and some 

increased (trophic level not predictive of 

susceptibility). 4 species apparently eliminated from 

site 2, after a year; other species persisted at low 

levels a year after spray. 

34 and 68 g a.i./ha (0.5 and 0.97 oz 

a.i./ac) 

Honey bee, Apis mellifera. Hives placed in gypsy 

moth treatment blocks. No effects from applications 

on numbers of adults, larvae, or honey production. 

Matthenius1975 

Aerial application of Dimilin 25-W 

at a rate of 70.75 g/ha (2 oz/acre) to 

770x770 m (60 ha) plots on May 8, 

1985. Plots were separated by at 

least 150 m to minimize the effects 

of spray drift. The study area 

(Morgan Co, WV) was characterized 

by mature oak-pine and oak-hickory 

forests. Gypsy moths were mostly 1st 

and 2nd instars and foliage was not 

fully expanded at the time of 

treatment. 

Foliage residues: 

1 day after treatment = 0.45±0.25 ppm 

3 days after treatment =0.31±0.16 ppm 

10 days after treatment =0.10±0.06 ppm 

21 days after treatment =0.18±0.16 ppm 

Martinat et al. 

1987 

Aerial application of Dimilin 25-W 

at a rate of 70.75 g/ha (2 oz/acre) to 

770x770 m (60 ha) plots on May 8, 

1985 and May 9, 1986 Plots were 

separated by at least 150 m to 

minimize the effects of spray drift. 

The study area (Morgan Co, WV) 

was characterized by mature oakpine and oak-hickory forests. 

Significant, treatment-related reductions were 

observed primarily in canopy macrolepidoptera and 

non-lepidopteran mandibulate herbivores. Sucking 

herbivorous insects, microlepidoptera, and 

predaceous arthropods were not affected. 

Martinat et al. 

1988 

70 g a.i./ha (1 oz a.i./ac)applied to 120 species of spiders (Araneae) and orthopteroid Martinat et al. 

oak-pine and oak-hickory hardwood. (Orthoptera and Dictyoptera). Significant effects 

from treatments noted on spider on 1 of 10 sampling 

dates, and on orthopteroid abundance on 2 of 10 

sampling dates. Trend in expected direction on 

other dates. No change in diversity of these groups. 

Effect on spiders could be from loss of prey or direct 

toxicity. Orthropoids picking up from litter that they 

ingest. 

1993 

Appendix 3a - 5 
Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Application of 280 g/ha a.i. Dimilin Half-life (calculated from first order kinetics) = 27 Mabury and 

WP-25 via backpack sprayer to rice hours; residues were below detection limit after 96 Crosby 1996 

field 5 days after emergence of rice hours. 

leaves out of the water. 

Additional notes on Maybury and Crosby 1996: Sensitized photolysis was the primary route of degradation, 

although partitioning to sediment and volatilization may have played minor roles in the fate of the compound. 

Rapid photolysis of DFB to CPU and DFBA. This mixture was as toxic to daphnids as DFB. Photolysis in 

distilled water is slow. Halftime of alkaline (pH 8.8) photodegradation was 157 hours (2.4 days) and filtered field 

water (pH 7.4) was 32 hours (1.3 days). Slower rates of photodegradation for CPU and DFBA. Field dissipation 

halftime for DBF of 27.3 hours (1.1 days) with typical initial increase in concentrations of CPU. 

30 g a.i./ha, in 4.78 l water (0.43 oz Wasp parasite on gypsy moth larvae (Apanteles Madrid and 

a.i./ac) melanoscelus)Parasitic fly in family Tachinidae. Stewart1981 

Wasp mortality 80% in 2 weeks from field spray. 

Development halted in most cases, failed to spin 

cocoons upon emergence, etc. 100% mortality in 

tachinid parasite. Gypsy moths in 2nd, 3rd, and 4th 

instar. 

Aerial (fixed-wing aircraft) Residues in run-off water decreased from 5 :g/L one Mutanen et al. 

application of Dimilin WP-25 at a day after spraying to 0.1 :g/L after 2 months. The 1988 

rate of 75 g a.i. in 50 L water/ha in A concentration in water in open pits was 0.1 :g/L 1 

total of 1160 ha of insect-infested and 7 days after application and 0.2 :g/L 1 month 

forest in Finland in August 1984 in after application. After 2 months no residues were 

an effort to control the pine looper, detected. All water samples taken from outside the 

Bupalus piniarius.A solution of treated area contained < 0.1 :g/L (the limit of 

hydroxyethyl cellulose and 15 g sensitivity). No DFB was detected in the treated area 

sodium bicarbonate in 50 L water/ha the year following application or outside the treated 

was added to formulation to area. Neither 4-chloroaniline nor 4-chlorophenylurea 

minimize drift, especially near the was detected in the water at any time. Residue data 

borders of the treated area. for the litter layer, humus layer, pine needles, wild 

mushrooms, boletus samples, and bilberries are 

provided. 

DFB (25% WP) via handgun to four­ Half-lives of DFB surface residues (Exp 1 cool-dry Nigg et al. 1986 

tree Valencia orange blocks at a rate period: March to April): leaves =essentially none 

of 10 oz a.i./acre. Trees were fruit =118±100 days; soil (middle) = 19±11 days; 

sprayed to runoff to control citrus soil (dripline) = 21±10 days; 

rust mite. Half-lives of DFB surface residues (Exp 1 hot-wet 

period: March to April): leaves + 27±8 days; fruit 

18±2days; soil (middle) = levels too low to be 

detected; soil (dripline) = levels too low to be 

detected 

Appendix 3a - 6 
Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Aerial application of 70 g a.i./ha Throughout the study, mites (49%) and springrtails Perry et al. 1997 

Dimilin to experimental watershed (28%) dominated the soil core sample. A total of 19 Perry 1995b 

(two treated; two controls) in the taxonomic groups were suitable for statistical 

Fernow Experimental Forest, WV. analysis. No significant treatment effects were 

Soil and leaf litter arthropods were observed, based on total organism counts or counts 

monitored before and after by trophic categories (p<0.05). 

application for a total of 36 months. 

Additional Notes on Perry et al. 1997:No significant treatment-related effects for populations of major 

taxonomic groups, except for Araneae (spiders) were observed. Analysis of leaf-litter bags also indicated no 

significantly differences in total numbers of invertebrates or in trophic categories between treated and untreated 

watersheds during the 12-month post treatment study. There appeared to be an indirect effect on spiders as a 

taxon, which may have resulted from changes in prey populations. 

Aerial application of Dimilin 25W at 

a rate of 33.23 g a.i./ha in 9.4 L/ha to 

a 20-ha forest block in central PA. 

Leaf samples were collected from the 

upper and lower canopies of 27 oaks 

and understory within the block on 

the day of application, May 29, 

1991. Canopy leaves were also 

collected on May 31, June 10, July 

29, and September 26, 1991. 

Three cover sprays of diflubenzuron 

(NOS) at 3.7 or 7.4 g a.i./100 L in a 

pear orchard in northern CA. 

[Data to calculate application rate in 

g/ha not given] 

0.5 and 2 oz a.i./ac, w/ crop oil, 

sprayed 8 times on cotton. [35 to 140 

g/ha × 8] 

Aerial application of oil formulation 

of DFB (Dimilin 45 ODC) on August 

31st in a conifer forest in the north of 

Spain at a dose of 56.3 g a.i./ha 

a(125 cm3

 Dimilin in 5 L diesel oil) 

(volume rate of application = 5 

L/ha). The day of application was 

clear with no rainfall in the previous 

48 hours. 

On the day of application, DFB residues on the 

upper canopy, lower canopy, and understory 

averaged 81.18, 39.65, and 8.35 ng/cm .2 

DFB residues on canopy leaf residues were: 

14.83 ng/cm2

 (day 2 post spray) 

16.75 ng/cm2

 (day12 post spray) 

12.84 ng/cm2

 (day 61 post spray) 

11.20 ng/cm2

 (day 120 post spray) 

DFB residues on litter-leaf sample collected after 

leaf senescence 169 and 323 days after treatment 

contained measurable amounts of DFB in 51 and 

59% of the samples, respectively. 

DFB treatment had no direct effect on pear psylla 

(pest species), did not induce phytophagous mites, 

and was weak, compared with the synthetic 

pyrethroid, fenvalerate against the codling moth. 

Direct spray of bee hives. No effects noted on adult 

mortality, rate of larval growth, brood production, or 

honey or wax production. No residues in wax or 

honey. Not caged study, so bees could have foraged 

outside of spray area. 

DFB persisted for 10-12 weeks on the foliage of the 

conifer forest; 55-80% of the insecticide was 

removed from the foliage within 22-30 days after 

treatment; aerial application resulted in residue 

levels of 867.5-1824.4 ng/g, depending on the forest 

characteristics. 

2,6-difluorobenzmide was the only metabolite 

detected and persisted only until the first rainfall. 

Prendergast et 

al. 1995 

Riedl and 

Hoying 1980 

Robinson 

1978,1979 

Rodriguez et al. 

2001 

Appendix 3a - 7 
Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Aerial application of Dimilin 25-W 

at a rate of 70.75 g/ha (2 oz/acre) to 

770x770 m (59.2 ha) plots on May 9, 

1986 Plots were separated by at 

least 150 m to minimize the effects 

of spray drift. The study area 

(Morgan Co, WV) was characterized 

by mature oak-pine and oak-hickory 

forests. 

Diets of five species of forest birds were 

significantly different between treated and untreated 

plots. Treatment generally decreased the biomass of 

Lepidoptera larvae and increased the biomass of 

other orders (Homoptera, Diptera, Coleoptera, 

etc.).Two species of birds in treated sites had 

decreased total gut biomass. 

The investigators conclude that DFB has an indirect 

adverse effect on forest birds by reducing the 

availability of Lepidoptera larvae. 

Sample et al. 

1993a 

Aerial application of Dimilin 25-W 

at a rate of 70.75 g/ha as part of a 

gypsy moth suppression program in 

WV. 

Treatment adversely affected Lepidoptera resulting 

in decreased abundance and species richness; no 

effects were observed among Coleoptera, Diptera, or 

Hymenoptera. Trap catches of 3 families of 

Hymenoptera were unaffected, including two 

parasitic families, Ichneumonidae and Braconidae. 

Sample et al. 

1993b 

Application of Dimilin on a regular 

basis (i.e., 8 applications between 

May 16 and December 14 1977) to th th 

a small citrus grove in which there 

were two bee hives. 

The hives remained in the same location throughout 

the study and were covered with plastic as a means 

of protection. There were no adverse effects on 

brood development of honey bees. 

Schroeder 1978a 

MRID 

00099731 

Appendix 3a - 8 
Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Aerial application of 350 g a.i./ha Residues in ppm on fruit harvested 27 days after the Schroeder 1980 

diflubenzuron to a commercial citrus 6th application were: 0.34 on unwashed fruit; 0.11 on 

grove to control Diaprepes washed fruit; 0.26 on dried pulp; 0.31 on peel fruit; 

abbreviatus 0.12 on chopped peel; and 20.55 in oil. 

Additional Notes on Schroeder 1980: No detectable residue (<0.05) of DFB was found in the finisher pulp, fruit 

juice, pressed liquor, molasses, prewash or afterwash water, and emulsion water fractions. No detectable residue 

(<0.05) of 4-chlorophenylurea or 4-chloroaniline was found in the citrus fractions or in the prewash or afterwash 

water. The total sealed brood in honey bee (Apis mellifera) was not significantly different from control at 7 

months and there was no detectable residue (<0.05 ppm) of DFB,CPU, or 4-chloroaniline was found in the honey 

obtained after 8 aerial sprays. Populations of non-target citrus pests and beneficial species were not affected by 

the spray program. 

Sour orange (Citrus aurantium) trees 

sprayed to runoff with Micromite 

25W at 149 or 298 g a.i./1000 liters 

Efficacy study. 

Diflubenzuron, formulated as Micromite 25W, 

significantly affected the reproductive potential of 

the sugarcane rootstock borer weevil, Diaprepes 

abbreviatus (pest of sugarcane and citrus). 

Schroeder 1996 

Aerial application of Dimilin 

formulated as 25% wettable powder 

at the rate of 140g/ha to 770 m 

square plots with a buffer strip of at 

least 150 m between adjacent plots in 

May 1985 and 1986. 

Estimates of density of white-footed mouse, 

Peromyscus leucopus) did not differ significantly 

(p>0.05) between treated and untreated areas. 

Juvenile/adult female ratios on untreated areas were 

significantly higher (p<0.05), compared with those 

on treated sites. Mice on treated sites consumed less 

Lepidotera prey, compare with controls (p<0.05); 

however, the total amount of food consumed per 

mouse did not differ significantly between treated 

and untreated areas (p>0.05). There were no 

treatment-related adverse effects on body 

measurements, weight, or fat content. 

Seidel and 

Whitmore 1995 

Aerial application of Dimilin (NOS) 

at a rate of 140 kg/ha. The 

application rate is presumably a.i. but 

this is not specified in the 

publication. 

No effect on bird populations that could be 

attributed to diflubenzuron. Various changes in the 

populations of different bird species are discussed 

but detailed data are not reported in the publication. 

Stribling and 

Smith 1987 

Simulated aerial application of The residue levels 1 hour after application varied, Sundaram 1986 

diflubenzuron in acetone or in fuel respectively, from 23.8 to 30.6 μg/g in foliage and MRID 

oil each at 90 g a.i. in 18 L/ha to from 3.08 to 4.60 μg/g in litter. Forty-five days after 00161955 

spruce foliage (Picea glauca). spraying the residue levels in foliage were 0.80 and 

3.9 μg/g, respectively, for acetone and fuel-oil 

formulations. 

Sundaram 1986 

Appendix 3a - 9 
 

Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Simulated aerial spray application of 

technical grade DFB in acetone 

formulation with tracer dye or in fuel 

oil with tracer dye at a rate of 90 g 

a.i. in 18 L/ha to white spruce foliage 

of uneven height. The forest floor 

was flat and covered with grass and 

moss patches. 

The half-lives for DFB in foliage, litter, and soil for 

the acetone-based formulation were 9.30, 8.36, and 

7.49, respectively. 45 days after application, the 

residue levels in foliage were 0.80 μg/g (fresh 

weight) for the acetone-based formulation. There 

were no detectable residues in litter or soil on the 

45th day post application of the acetone-based 

formulations. 

Sundaram 1996 

Soybeans in S.C. treated with 281 or 

562 g a.i./ha (4 or 8 oz a.i./ac) 

Significantly fewer nabids and geocorids on treated 

v. control sites. 

Turnipseed et al. 

1974 

Aerial application of 8 oz Dimilin 

WP-25/acre (equivalent to 0.0625 

lb/acre) to 10-acre mixed hardwoodconifer forested plot near Boone N. 

Carolina, which consisted of a 

stream, two stream pools, and a 

stream-fed pond outside the treated 

area. Sandy loam soil. Cumulative 

rainfall of 43.1 cm (16.9 inches) over 

a 1 year period. Daily rainfall and 

temperature data are given. 

Initial concentration on leaves in canopy of 13 ppm 

on hardwood and 5.9 ppm on conifer. 

Initial concentrations on understory vegetation of 

about 0.13 ppm that increased initially as with litter. 

Diflubenzuron was rather persistent on leaf litter. 

Initial residues of 0.07 ppm. This increased over a 

60 day period, probably due to drying of litter, 

washoff of DFB, and leaf fall from canopy. 

Van Den Berg 

1986 

MRID 

00163853 

Additional Notes on Van Den Berg 1986: A single application resulted in initial water concentrations in 

treatment area of 0.127-0.203 ppb. Declined to 0.029-0.045 ppb after one day. No detectable contamination in 

an adjacent pond after heavy rains. Initial soil concentrations of 0.02 ppm and 0.03 ppm after a 6.5 cm rain 

(probably washoff). No DFB in 3"-6" soil samples. The study authors conclude that the effects on the mites and 

collembolans present at the time of application were insubstantial. In general, fewer of each group on treated 

than untreated sites. The data are somewhat difficult to interpret because of erratic capture patterns over time the 

populations of collembolans were different at the control and treated sites prior to treatment. [NOTE: Data on 

other species presented in Tables 10 and 11 but the numbers of insects are too small for analysis. Species list in 

Table 11 cut off on fiche ] 

28 g a.i./ha (0.4 oz a.i./ac) Wasp parasite on Gypsy moth larvae (Cotesia Webb et al. 

melanoscelus) Pathogen: gypsy moth nuclear 

polyhedrosis virus (NPV). Numbers of the wasp no 

different on Control v. treated plots. Incidence of 

NPV significantly lower in treated plots. Late instar 

spraying may preserve larvae long enough for 

parasitoid to complete development. Earlier 

spraying kills host too quickly, hence parasitoid as 

well. NPV lower in treated plots because fewer 

Gypsy moths to transmit virus. 

1989 

Appendix 3a - 10 
Appendix 3a: Summary of terrestrial field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

application of 2 or 4 oz a.i./acre [140 Yellowjackets (Vespula or Dolichovespula). Weiland 2000 

to 280 g/ha] to plots with large, Treatment decreased populations and the effect was MRID 

active nests of yellowjackets readily observed during the following year. 

No effects observed on Mound-building ants 

(Formica) 

45245403 

560, 280, and 140 g a.i./ha (8, 4, and Nearly twice as many Psylla predators and parasites Westigard 1979 

2 oz a.i./ac). 2 and 3 treatments. per season in the lowest application rate. Higher 

Handgun and air-carrier sprayer. rates resulted in higher populations of the pear 

psylla. 

Aerial application of 0.03lbs a.i./acre 

Dimilin 25WP to Appalachian forest 

ecosystem during 1991 season (20 

trees representing 7 species) in WV 

Univ. Experimental Forest. 

Residue on leaves: significant loss of DFB from 

foliage ranging from 20 to 80% within the first 8 

days after application; remaining DFB generally 

persisted for the rest of the growing season until leaf 

fall, at which time 13/20 treated trees retained more 

than 20% of the original pesticide applied. 

Wimmer et al. 

1993 

Dimilin (TH-6040) formulated as 

dispersable powder (a.i. 25% by 

weight) applied aerially at the rate of 

0.28 kg a.i./ha (0.25 lbs a.i./acre) to a 

Douglas-fir forest ecosystem in 

British Columbia 

Treatment decreased the total number of flying 

insects and the effect was sustained throughout the 

study period, with the greatest impact observed on 

midges and gall gnats. Mosquitoes were completely 

wiped out as a result of treatment. 

Wilson and Wan 

1977a MRID 

00095419 

Wilson and Wan 

1977b MRID 

00129973 

Appendix 3a - 11 
Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Two applications (NOS) of granular 

diflubenzuron at 0.11 kg a.i./ha 

(about 3.7 μg/L) or 0.22 kg a.i./ha 

(about 7.4 μg/L) to residentialrecreational lakes in San Bernadino 

County (June 1967 - January 1977) 

At 0.11 kg a.i./ha application: 

Daphnia pulex and Daphnia galeata: 62-75% 

decrease in population during 7 days after treatment; 

populations recovered in the second week after 

treatment. 

Ali and Mulla 

1978a 

Additional Notes on Ali and Mulla 1978a: 0.11 kg/ha continued. Diaptomus spp. (copepods): 30% decrease in 

population observed 2 days after treatment. Hyalella azteca (amphipods): 97% decrease in population observed 3 

weeks after treatment; populations remained below pretreatment levels throughout 8-9 week evaluation. 

Treatment had no detectable effects on Cyprihnotus sp.(seed shrimp), Cyclops, or Bosmina longirostris 

(Cladocera). 

At 0.22 kg a.i./ha application: Daphnia pulex and Daphnia galeata: completely eliminated for 3 weeks after 

treatment. Diaptomus spp (copepods): populations decreased to 0 within 7 days after treatment, but recovered 

completely soon thereafter. Hyalella azteca (amphipods): 30-100% decrease in population during 2 1⁄2 months 

after treatment. Cyprihnotus sp.(seed shrimp): population stressed for only 2 weeks. Oligochaete (mostly 

Naididae found in marine, brackish, and freshwater habitats): no significant effects observed at either treatment 

level. 

Two spray application of 

diflubenzuron (25% WP) to entire 

surface of residential-recreational 

lake in Riverside County at a rate of 

156 g a.i./ha-surface (about 0.012 

ppm) in April and August 1977. 

Ali and Mulla 

1978b 

Daphnia leavis and Ceriodaphnia sp: population 

eliminated within 1 week with no recovery 6 

months after treatment. 

First application (April) 

Additional Notes on Ali and Mulla 1978b: Bosmina longirostris (cladocerans): population eliminated within 1 

week with recovery after 11 weeks. Cyclops sp. (crustaceans): population eliminated within 1 week with recovery 

within 6-7 weeks. Diaptomus spp. (copepods): population eliminated within 1 weeks with recovery after 4 

months. Hyalella azteca (amphipods): population eliminated within 4 weeks with no recovery 6 months after 

treatment. Caenis sp. [Hemeroptera (mayflies, immature)]: elimination within 3 weeks with recovery within 6-7 

weeks. Physa sp. (sinistral snails, referred to as pond snails or pouch snails): no adverse effects. Cypridopsis 

sp.(bivalve): no adverse effects. Second application (August) Bosmina longirostris (cladocerans): population 

eliminated after 1 week; reappearance in small numbers 8-9 weeks after treatment. Cyclops sp. (crustaceans): 

population eliminated within 1-2 weeks with recovery after 4 weeks. Diaptomus spp. (copepods): population 

absent prior to treatment; reappearance in small numbers 1-2 months later. Caenis sp. [Hemeroptera (mayflies, 

immature)]: elimination within 2-3 weeks with recovery within 4-5 weeks. 

Study does not provide monitoring data. See Ali et al. 1988 below. 

Appendix 3b - 1 
Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Application via airblast sprayer of No apparent adverse effects on zooplankton and Ali et al. 1988 

Dimilin 25 WP at a rate of 0.56 kg benthic invertebrates in treated pond. Minor 

a.i./ha to 0.8 ha of citrus immediately reductions of copepods and cladocerans during postsurrounding a pond located in Winter treatment period most likely due to short life cycle, 

Garden, FL. The pond was exposed seasonal population changes, and possible sampling 

to air-drift diflubenzuron from deficiencies. 

surrounding citrus area commercially 

treated for the control of citrus rust Largest detected diflubenzuron residue = 197 ppt, 2 

mite. The control pond was located days after application with levels returning to trace 

0.4 km NE of the exposed pond. amounts (<27 ppt) by day 14 after application. 

Specifics on the pond: circular,2 ha at the surface; 

3/4 of its border was lined by citrus trees. 

One surface application (via rowboat 

hand sprayer)of Dimilin (25% WP in 

20.5 L water) to each of three ponds 

(0.6-0.2 hectares) at rates of 2.5, 5, 

or 10 ppb a.i. in California to control 

gnats (Chaoborus astictopus)and one 

application to a large lake at a rate of 

5 ppb a.i. 

Surface area of ponds ranged from 

0.06-0.2 ha; ponds were rectangular 

in shape with steep sides and flat 

bottoms. 

Treatment was effective against gnats, decreasing 

larval abundance by 99%. Crustacean zooplankton 

populations declined precipitously at all application 

rates, but the effects were not permanent. 

Cladocerans were more susceptible than copepods 

and required longer recovery period. Anabaena sp 

(blue-green algae) decreased by approximately 70% 

within 2 weeks after treatment and remained at low 

levels throughout the study period; treatment seemed 

to have no effect on diatoms or green algae. The 

bioaccumulation of diflubenzuron in bluegill sunfish 

diminished rapidly as the residues in water 

decreased. No effect on growth of bluegills. 

Apperson et al. 

1977 

MRID 

00099897 

Apperson et al. 

1978 

Additional Notes on Apperson et al. 1977, 1978: The investigators indicate that no severe or permanent 

nontarget effects were observed in this study. Residues: In pond water, residues in the 10 and 5 ppb ponds 1 

hour after treatment ranged from non-detectable to 23.6 and 32.2 ppb and averaged 9.8 and 4.6 ppb, respectively 

and residues levels in the 2.5 ppb pond at 4 hours after treatment ranged from N.D. to 8.3 ppm with an average of 

1.9 ppb. Maximum values in bottom water samples in the 5 and 2.5 ppb ponds occurred at 4 hours and 14 days 

and averaged 5.3 and .5 ppb, respectively. The DFB residues declined steadily soon after treatment and at the end 

of the study, levels averaged 0.2, 0.3, and 0.5 ppb for the 10, 5, and 2.5 ppb ponds, respectively. No residues 

were found in the sediment samples. 

Applications to test ponds at 1X and 

4X of the typical application rate. 

No effects on invertebrates or fish. [This study is 

poorly documented and should be given minimal 

weight.] 

Birdsong 1965 

Four applications of Dimilin W25 to 

ponds located in Salt Lake County 

Utah between 7/14/15 and 10/7/75 

Algae (Plectonema) degraded 80% of the TH-6040 

in a 1-hour incubation period. Degradation products 

were primarily p-chlorophenyl urea and pchloroaniline. 

Booth and 

Ferrell 1977 

MRID 

00099884 

Additional Notes on Booth and Ferrell 1977: Bacteria (Pseudomonas sp.) accumulated “rather large amounts” of 

TH-6040 from the incubation media when used as the sole carbon source. No degradation products were 

observed in the media. Channel catfish id not bioaccumulate DFB residues from treated soil in a simulated lake 

ecosystem constructed in the laboratory. 

Appendix 3b - 2 
Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Repeated, pulsed exposures of Within 4 weeks after monthly and biweekly Boyle et al. 

diflubenzuron on twelve outdoor treatment, direct effects on Cladocerans, Copepods 1996 

aquatic mesocosms (0.1 ha each). and Rotifers included 5-fold decrease in total 

Random assignment of mesocosms numbers, 2-fold decrease in species richness, and 2­

(four/treatment) to either monthly fold increase in zooplankton. Direct reductions in 

(five total 10 μg/L applications) or the numbers of invertebrate grazers caused indirect 

biweekly (nine total 10 μg/L increases in algal biomass. Decreased invertebrate 

applications). Direct and indirect numbers resulted in decreases in invertebrate food 

impacts on mesocosms were resources that resulted in a 50% reduction in both 

measured over 16 weeks after biomass and individual weights of juvenile bluegills. 

treatment. There were no statistically significant impacts 

observed on adult bluegills or largemouth bass for 

the duration of the study. 

Additional Notes on Boyle et al. 1996: DFB concentrations averaged 9.9 μg/L 24 hours after chemical 

application. The half-life of disappearance of DFB from water, calculated across all ponds and dates using a 

negative exponential decay model was 2.33 days (range = 1.76-2.96 days). There were no significant differences 

in DFB dissipation rate between treatment type (monthly or weekly; p$0.5815) or season (early or late in the 

study; p$0.4728. 

aerial application of 35 g/ha in 

Canada 

No toxic effect on bullheads or sunfish. Buckner et al. 

1975 

Two ground spray applications (at 2­

week intervals) to each of two CA 

sites (one in Tiburon, Marin County 

and one in Roseville, Placer County). 

The first Tiburon application = 13 

g/ha (0.19 oz/acre) and the second 

Tiburon application = 35g/ha (0.5 

oz/acre);both Roseville applications 

= 26.25 g/ha (0.38 oz/acre) of 

Dimilin 25W (diflubenzuron 25% 

a.i.). Foliage was sprayed to the 

point of drip. Each site was 

approximately 0.8 ha. The 

applications were made in MarchApril 1990. 

Foliage: DFB concentrations from 0 (not detected) 

to 18.31 μg/g immediately after the second 

application; and from 0(not detected) for 

background to 0.252 μg/cm2

 leaf area immediately 

after the second application. 28 days after the second 

application, the DFB concentration decreased 

sharply suggesting possible degradation during that 

period, but no samples were collected during the 28 

days to document a degradation trend. 

Air: During 3 of the 4 applications, DFB 

concentrations in air ranged from 0.0106 to 0.0187 

μg/m . DFB was not detected in any background air 3 

samples or in any 1 day post application air samples 

(i.e., DFB was detected in air only during 

application periods). 

Water: Samples collected from streams and water 

bodies in and near the treated areas on the day prior 

to application, immediately after each application, 

and 7 days after each application showed no 

detectable levels of DFB (minimum detection limit = 

0.5 ppb). 

Carr et al. 1991 

Appendix 3b - 3 
Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Application (NOS) of diflubenzuron 

to five experimental, rectangular 

ponds in Lakeport, CA, yielding a 

mean concentration of 13 μg/L DFB. 

Each pond had a surface area of 

about 0.01 ha (1 ha =10,000 m ) and 3 

a depth of 1.2 m. 

Residues in water decreased below detectable 

limits (0.2 μg/L) by 14 days after treatment; at one 

hour after treatment, the mean concentration of DFB 

in water was 13.2 μg/L. 

Colwell and 

Schaefer 1980 

Additional Notes on Colwell and Schaefer 1980: Cladocerans: most abundant species included Ceriodaphnia, 

Diaphanosoma, Chydorus, Bosmina, and Daphnia, all of which showed population reductions in all treated ponds 

within a few days of DFB application. Copepods: abundance of naupli decreased in all ponds after treatment and 

returned to pretreatment levels from 7 days to >4 weeks after treatment. Diaptomus (filter feeders) and Cyclops 

were similar in their susceptibilities to DFB, although in most of the treated ponds, Diaptomus populations 

recovered more rapidly than Cyclops populations. Rotifers: Brachionus, Keratella, and Hexartha populations 

increased in treated and control ponds during the first 8 days after treatment. Asplanchna, which are mostly 

predatory increased from 0.18 to 0.43 organisms/L after treatment. Fish: Young-of-the year black crappie, 

Pomoxis nigromaculatus, and brown bullhead, Ictalurus nebulosus, accumulated DFB and then eliminated all 

residues by day 7 after treatment. No fish mortalities occurred after treatment. For 1 month after treatment, the 

stomach content analyses of exposed fish indicated major alterations in diet. Neither growth rates or general 

condition of the fish 3 months after treatment differed from those of controls. 

Six aerial applications of 28 g/ha of Treatment resulted in statistically significant Farlow 1976 

diflubenzuron over 18 months (June differences in population density of non-target MRID 

1974 through Sept 1975) to a aquatic organisms (target organism - mosquito), 00099678 

Louisiana intermediate marsh compared with controls, but none of the affected 

organisms were completely eliminated from the 

ecosystem. The investigators speculate that the 

untreated marsh areas would provide populations of 

aquatic organisms that could repopulate the treated 

areas. 

[Also published 

as Farlow et al. 

1978] 

Six applications of diflubenzuron (28 Statistically significant differences in the population Farlow et al. 

g a.i./ha) in a Louisiana coastal density of aquatic organisms; however, none of the 1978 

marsh over an 18-month period. organisms affected were completely eliminated from 

the ecosystem. 

Additional Notes on Farlow et al. 1978: Significant populations decreases observed in five taxa: nymphs of 

Trichocorixa louisianae (water boatman) and Buenoa spp.(backswimmers), Coenagrionidae naiad 

spp.(damselflies), Berosus infuscatus adults (water beetles), and Hyalella azteca (amphipods). Significant 

increases were observed in populations of 15 taxa exposed to diflubenzuron, i.e., Physa sp. (snails), Ceanis sp. 

and Callibaetis sp. naiads (mayflies), Noteridae larvae (water beetles), Hydrovatus cuspidatus, adults (water 

beetles), Hydrovatus sp. larvae (water beetles), Dytiscidae larvae (great diving beetle), Mesovelia mulsanti adults 

(water treaders), Trichocorixa louisiana adults (water boatman), larvae of Chironomidae (non-biting or true 

midges), Ephydridae (shore flies), Dolichopodidae (long-legged flies) and Tabanidae (horseflies), as well as 

mosquito fish (Gambusia affinis) and American flag fish (Jordanella floridae). The 27 remaining aquatic 

organisms (members of the Hemiptera, Coleoptera, Mysidacea, Decapoda, Diptera and Odonata) showed no 

statistically significant differences, compared with untreated populations. 

Appendix 3b - 4 
Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Aerial application of Dimilin 4 L at a Treatment decreased the adult emergence of Griffith et al. 

rate of 35.1 g a.i./ha to two stream stoneflies, Peltroperla arcuata, during the first 4 1996 

catchments in the Fernow months after treatment, compared with untreated 

Experimental Forest, WV in May catchments. Adults populations of other species did 

1992. not decrease in the treatment catchments during the 

period of study. 

Additional Notes on Griffith et al. 1996: The investigators speculate that additional detritivourous species 

might have shown an adverse effect if the monitoring were extended through the period after treated leaves 

entered the streams. Stoneflies are considered to be obligate large-particulate organic matter feeders and like 

ingested diflubenzuron from leaves that fell earlier in the year, thus ingesting diflubenzuron. Diflubenzuron was 

not detected in water samples taken from the streams following treatment, perhaps due rainfall just prior to 

treatment. 

Aerial application of Dimilin 4 L at a 

rate of 35.1 g a.i./ha to two stream 

catchments in the Fernow 

Experimental Forest, WV in May 

1992. During 1993, no additional 

diflubenzuron was applied to any of 

the watersheds. 

The investigators tested the hypothesis that 

diflubenzuron affected adult flight following 

emergence during the year following abscission and 

possible ingestion of the treated leaves. The flight of 

the stonefly, Leuctra ferruginea, was reduced in the 

treatment watersheds, compared with the reference 

watersheds during the year following abscission of 

the treated leaves. Adult flight of other species did 

not decrease in the treatment watersheds during 

1993. 

Griffith et al. 

2000 

Aerial application of Dimlin 4L at a 

rate of 70 g a.i./ha to two of four 

watersheds in the Fernow 

Experimental Forest, WV. 

Spray application (via portable 

garden sprayer) of Dimilin (25% 

wettable powder) at recommended 

rate of 0.03 lbs a.i./acre or 4X 

application rate to each of two 10­

acre earthen ponds (avg depth of 3 

ft). 4X applications were made 

biweekly beginning in early Feb. 

Stream macroinvertebrate taxa that had reduced 

mean densities in treated watersheds (% = 0.05) 

included the stoneflies, Leuctra sp. and Isoperla sp., 

mayflies, Paraleptophlebiaspia sp., and cran flies, 

Hexatoma sp. Shredders, the dominant functional 

feeding group also had reduced mean densities in 

treated watersheds. Densities of Oligochaeta (aquatic 

worms) and Turbellaria (flat worms) increased in 

treated watersheds. 

No appreciable mortality of fish or clams in any of 

the ponds. Treatment significantly decreased 

Daphnia spp. populations and virtually eliminated 

dipterans. Olgochaete populations, which increased 

in the control pond during the study, decreased in 

response to treatment. 

Hurd et al. 1996 

Jackson 1976 

MRID 

00099891 

Appendix 3b - 5 
Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Aerial application of 0.06 lbs 

a.i./acre (67.26 g/ha) Dimilin to 75­

acre watershed containing small, first 

order stream. 

Spray application of 60 g a.i./ha 

diflubenzuron to five Sahelian 

temporary ponds (surface areas 0.36­

0.65 ha) conducted in midSeptember (half-way through rainy 

season) in vast savannah-type 

cultivated region in Senegal’s 

ground-nut producing area. Table 1 

provides a summary of wind speed, 

surface area treated, quantity of 

formulation applied in mL and 

calculated application rates at each of 

the 5 treated ponds. 

Dimilin reached the stream channel during aerial 

application and as a result of wash-off from the 

foliage during several subsequent rainfall events. 

DFB levels (measured) exceed the acute (1.0-1.8 

ppb) and chronic (60 ppt) toxicity doses for tolerant 

taxa, like Ephemeroptera (mayflies) and Plecoptera 

(stone flies). The residence time for Dimilin in these 

high-gradient streams was very short, and as a result 

of the short residence time or low concentrations, 

toxic effects were not evident. 

Average initial concentrations in water = 10.4 μg/L, 

with an estimated half-life of <24 hours. 

DFB only affected crustaceans (i.e., cladocerans and 

fairy shrimp) in the treated ponds. DFB virtually 

eradicated the abundant fairy shrimps, 

Streptocephalus spp., and the populations did not 

recover despite the rapid disappearance of DFB. In 

general, cladocerans populations were initially 

wiped out (densities dropped to 0) after DFB 

treatment but returned to normal values in 3-4 weeks 

(M micrura), 4-6 weeks (D senegal), or 6-7 weeks C 

quadrangula). 

Jones and 

Kochenderfer 

1987 

Lahr et al. 2000 

Application (via backpack sprayer) 

of Dimilin WP-25 at 280 g/ha a.i. 5 

days after emergence of rice leaves 

out of the water to sic 20 m flooded 2 

plots in June 1991 and 1992. 

Field dissipation rates were similar for the six 

replicate plates with a half-life (1st order) of 27 

hours; residues dropped to below detection limit 

after 96 hours. 

Residues in sediment were 0.16 μg/g (after 24 

hours), 0.10 μg/g (after 48 hours) and 0.08 μg/g 

(after 72 hours); residues were below detection limit 

after 4 days. 

Mabury and 

Crosby 1996 

Spray application (via hand 

sprayer)of Dimilin 25% WP 

(TH6040) to semi-natural pools at 

the Univ. Delaware Experimental 

Farm to study the cumulative toxicity 

to killifish (3 applications over 29 

days) and crustaceans (one 13-day 

test and one 15-day test). 

Applications were made at the rate of 

0.01, 0.04, 0.10, and 0.20 lbs 

a.i./acre – i.e., up to 224 g/ha. 

There was no significant mortality in killifish after 

three successive applications of Dimilin at 0.01-0.20 

lbs a.i./acre. Behavioral responses were similar to 

those of controls. 

In the first test involving crustaceans, grass shrimp 

mortality was 83.3% (p<0.01) after the first 

application of 0.20 lbs a.i./acre. After two 

applications the average mortality (p<0.01) was 

86.6% at 0.4 lbs a.i./acre and 100% at 0.10 and 0.20 

lbs a.i./acre. 

McAlonan 1975 

MRID 

00099895 

Appendix 3b - 6 
Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Additional Notes on McAlonan 1975: In the second test involving crustaceans, grass shrimp average mortality 

(p<0.01) was 91.6% at 0.4 lbs a.i./acre, 96.6% at 0.10 lbs a.i./acre, and 98.3% at 0.20 lbs a.i./acre. In the first 

test involving crustaceans, fiddler crab average mortality was 60.0% and 46.6% (p<0.01) after one application of 

0.10 or 0.20 lbs a.i./acre, respectively. After two applications of 0.04 and 0.10 lbs a.i./acre the average mortality 

(p<0.01) was 53.3% and 66.6%, respectively. In the second test involving crustaceans, fiddler crab average 

mortality (p<0.05) was 46.6% at 0.4 lbs a.i./acre,60.0% at 0.10 lbs a.i./acre, and 66.6% at 0.20 lbs a.i./acre. 

Aerial application of 0.56 kg a.i./ha 

(8 oz a.i./acre) Dimilin 25 WP to a 

citrus grove in Florida with an 

experimental pond 

DFB was not observed in water samples at 

quantitative methods 1 hour post application; 

maximum levels occurred at 1 and 2 days post 

application, primarily along the line of drift. Pad 

data indicate that the pesticide drift deposited along 

a small portion of the shoreline at a rate 7% of the 

theoretical application rate (38÷104÷5.6) and the 

drift continuing out into the pond was as much as 

0.8% the application rate (4.4÷104÷5.6). 

Nigg and 

Stamper 1987 

MRID 

40197002 

Dimilin 4L at a rate of 80g/ha (0.03 Decreased populations of stoneflies in treated areas. Perry 1995a 

lb/acre) in two forest watersheds In untreated areas, the populations of stoneflies 

increased. After treatment, populations of 

roundworms, flatworms, and segmented worms were 

higher in treated areas. 

Aerial application of 0.0624, 0.125, 

or 0.25 lbs/acre Dimilin to plots in 

Oxbow, Maine that included four 

streams. [up to 280 g/ha] 

Effects of a single application (to control spruce 

budworm) on stream invertebrate fauna 

(Trichloptera, Plecoptera, Ephemeroptera, Diptera, 

Odonata, and Coleptera). No pattern of decrease in 

any individual genus; no treatment-related increase 

in drift among samples; no treatment related changes 

in the number of dead drift when collections were 

made 1-2 days after treatment. 

Rabeni and 

Gibbs 1975 

MRID 

00159905 

Application (NOS) of 1.25 ppm 

Dimilin 25WPfor 1 hour on July 13, 

1984 to four points of the Kokawa 

River in the Izu Peninsula to control 

blackflies. The gradient of the river 

was approx. 2% and sampling 

stations are located between 50 and 

250 m above sea level. 

Most invertebrates were eliminated within 2 weeks, 

while Hydropsycidae (caddisfly) died out gradually. 

Adults of Elmidae (Riffle beetles), previously 

absent, appeared 1 week after treatment in large 

numbers at the uppermost of the treated region. No 

fish mortality was observed. 

Satake and 

Yasuno 1987 

Appendix 3b - 7 
Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Aerial application of Dimilin WP-25 

at a rate of 70 g a.i. in 10, 5, and 

2.5/H to three spray blocks in a 

mixed boreal forest near Kaladar 

Ontario Canada. Water, sediment 

and aquatic plants were analyzed for 

DFB residues. Ponds appear to have 

been directly sprayed. 

The duration of detectable DFB residues in water, 

sediment, and aquatic plants differed for each 

substrate but in all cases was less than 2 weeks. 

There was significant mortality in two groups of 

caged pond invertebrates (amphipods and corixidae 

[water boatman]) 1-6 days after treatment. Three 

taxa of littoral insects (mayflies, dragonflies, and 

damselflies) were significantly reduced in abundance 

in treated ponds 21-34 days post treatment but 

recovered to pre-treatment levels by the end of the 

season. Cladoceran and copepod populations were 

reduced 3 days after treatment and remained 

suppressed for 2-3 months. 

Sundaram et al. 

1991 

5 monthly surface applications of No adverse effects on population growth of fish. Takahashio and 

0.05 lbs a.i./acre Dimilin (25% WP) Miura 1975 

[56 g/ha] to artificial pond containing MRID 

mosquito fish (Gambusia affinis) 00016545 

2x application of Dimlin W-25 at a 

rate of 0.03 lbs a.i./acre at 14-day 

interval to an outdoor 750 gallon 

aquarium containing pond water and 

sediment, bluegill sunfish, clams, and 

crayfish; fate of diflubenzuron in all 

elements of the simulated ecosystem 

was monitored for 42 days from 

initial treatment. 

Rapid dissipation of DFB (half-life < 12 hours); 

rapid accumulation of compound by fish and clams 

with rapid elimination (plateau of approx. 55 ppb by 

day 27 which was maintained for the duration of the 

experiment); fish samples contained several 

degradation products (CPU and DFB represent the 

only organo-extractable residues; clam samples 

contained only DFB; crayfish did not accumulate 

any of the compound during the week after the initial 

treatment. 

ThompsonHayward 

Chemical Co 

1979 

In: Technology 

Sciences Group 

Inc. 1998 

MRID 

44460702 

Aerial application of Dimilin at a rate 

of 4.5 kg/ha (4 lbs granules/acre) to a 

tidal flood plain of the Fraser River 

in British Columbia in June 1976 . 

The organisms in the tidal flats of the 

Fraser River at the time of the study 

included crustaceans (zooplankton), 

insects, water mites and bugs, snails, 

and clams. 

Dimilin forestry spray at 67 g 

DFB/ha 

Residue: Dimilin, which was detected in the water 

up to 71 days after treatment, peaked at 1.8 ppb 8 

days after application and decreased slowly to a 

minimum level of 0.24 ppb at 2 months after 

application. In mud, Dimilin peaked at 5.66 ppb 4 

hours after application and decreased to a minimum 

level of 1.3 ppb by 2 months after treatment. 

Biological effects: Treatment arrested mosquito 

development but also decreased the population of 

zooplankton and suppressed the emergence of nontarget insects of the same order as the mosquitoes. 

No effect on aged brown trout in stream from day -7 

to day +6. Observations along length of stream 

revealed no indication of fish mortality. Based on 

population estimates 6 weeks following application, 

no delayed effects on fish populations. 

Wan and Wilson 

1977 

MRID 

00095416 

White 1975 

Appendix 3b - 8 
Appendix 3b: Summary of aquatic field or field simulation studies on diflubenzuron and its formulations. 

Application Observations Reference 

Broadcast foliar spray at rate of 0.25 

lbs a.i./acre of Dimilin 2L to rice 

paddy test plots in Arkansas and 

California 40 days after rice planting. 

DFB and its metabolites (DFBA and CPU) 

dissipated rapidly in the aquatic environment and 

there was no downward movement of DFB or its 

degradation products in aquatic soil/sediment. 

Willard 1999 

MRID 

45009601 

Broadcast spray application of 

Dimilin 25W to entire surface area of 

pond (containing fish) at a rate of 

0.36 lbs a.i./acre. 

calculated half-life for DFB in water = 5.4 days 

calculated half-life for DFB in soil/sediment = 8.6 

days. 

Willard 2000a 

MRID 

45191001 

Benthic communities in outdoor 

experimental streams , 

concentrations of 1 or 10 mg/L 

diflubenzuron for 30 minutes 

No drift of macrobenthos was Yasuno and 

induced at the time of application. However, Satake 1990 

diflubenzuron affected 

the emergence of all species examined. High larval 

mortality for a species of chironomid was observed 

directly in the stream treated with diflubenzuron, 

where numbers of mayfly nymphs and caddisfly 

larvae were also decreased 

Appendix 3b - 9 
Appendix 4: Toxicity of diflubenzuron and diflubenzuron formulations to birds 

Nature of Exposure Exposure 

Species Time Effects Reference 

Single Dose 

Mallard ducks, 

males and 

females, 10 

birds/dose 

group 

Bobwhite quail 

single gavage doses 

ranging from 1000 to 

5000 mg/kg bw TH6040 (99.4% pure) 

5000 mg/kg single 

gavage dose 

single dose 

single dose 

No mortality, no signs of Roberts and 

abnormal behavior or toxicity, Parke 1976 

and no gross pathological MRID 

changes to organs. 00073936 

NOEC = 5000 mg/kg bw 

LD50 >5000 mg/kg bw	 U.S. 

EPA/OPP 

1997a 

Note on above study: U.S. EPA/OPP 1997a attributes this study to Roberts and Parke 1976. Roberts and Parke 

1976, however, only assayed mallard ducks. A review of the CBI files did not identify an acute oral study in 

bobwhite quail. The above entry is included in the peer review draft but should be deleted in the final report 

unless the value can be verified. 

Red-winged single gavage dose of 

black birds, 1000, 2500, 3000, 

Agelaius 4000, or 5000 mg/kg 

phoeniceus, 5 bw technical grade 

or 6/dose group (99%) TH 6040; 

observation period of 

14 days 

Acute Dietary 

Mallard ducks in diet concentrations 

#4640 ppm technical 

grade TH-6040 (purity 

assumed to be 100%) 

dissolved in corn oil 

single dose	 Mortality: 

1/6 at 1000 mg/kg (considered 

unrelated to treatment); 

0/5 at 2500 mg/kg 

1/6 at 3000 mg/kg following 

signs of piloerection, asthenia, 

and ataxia; 

4/6 at 4000 mg/kg 

5/6 at 5000 mg/kg 

NOEC = 2500 mg/kg bw 

8 days	 NOEC =4640 ppm; no mortality 

and no observable signs of 

toxicity. 

Alsager and 

Cook 1975 

MRID 

00038614 

Fink and 

Petrocelli 

1973 

MRID 

00038613 

Appendix 4 - 1 
Appendix 4: Toxicity of diflubenzuron and diflubenzuron formulations to birds 

Nature of Exposure Exposure 

Species Time Effects Reference 

Reproduction Studies 

Mallard ducks, 

Anas 

platyrhynchos, 

young adults, 

16/sex/dose 

group 

dietary nominal 

concentrations of 0, 

250, 500, or 1000 

ppm. Based on mean 

body weights (about 

1.25 kg) and mean 

food consumption 

(about 160 g/day), the 

dietary concentrations 

correspond to about 0, 

32, 64, and 128 mg/kg 

bw/day. 

20 weeks No treatment-related mortality; 

no overt signs of toxicity; no 

treatment-related effects on body 

weight or feed consumption; no 

treatment-related effects of 

reproduction; and no treatmentrelated effects on body weights 

of hatchlings or 14-day old 

survivors.. 

At 1000 ppm, there was slight, 

but statistically significant 

decrease in mean egg shell 

thickness. 

Beavers et al. 

1990a 

MRID 

41668001 

NOEC = 500 ppm 

Bobwhite quail, 

Colinus 

virginianus, 

young adults, 

16/sex/dose 

group 

dietary nominal 

concentrations of 0, 

250, 500, or 1000 

ppm. Based on mean 

body weights (about 

200 g) and mean food 

consumption (about 

22 g/day), the dietary 

concentrations 

correspond to about 0, 

27.5, 55, and 110 

mg/kg bw/day. 

21 weeks 

(1-generation) 

No treatment-related mortality, 

overt signs of toxicity, or effects 

on body weight or food 

consumption during 

experimental period. 

At 1000 ppm, there was a 

marginal decrease in the number 

of eggs laid. 

NOEC (based on possible effect 

on egg production at 1000 ppm) 

=500 ppm. 

Beavers et al. 

1990b 

MRID 

41668002 

Beavers et al. 

1990c 

Appendix 4 - 2 
Appendix 4: Toxicity of diflubenzuron and diflubenzuron formulations to birds 

Nature of Exposure Exposure 

Species Time Effects Reference 

Bobwhite quail, dietary nominal 12 weeks No adverse effects on the Booth et al. 

Colinus concentrations of 2.5, reproductive parameters 1977 

virginianus, 25, or 250 ppm air- measured, including eggs laid, MRID 

adults milled (99.9% pure) cracked eggs, eggs set, fertile 00099719 

diflubenzuron eggs, hatched eggs, egg shell 

thickness, feed consumption, 

adult deaths, or chick survival. 

NOEC = 250 ppm based on 

review by U.S. EPA/OPP 1997a. 

The study authors attribute some 

observed differences between 

treated groups and controls to 

random variation and the large 

sample size (i.e, 500 eggs). 

Chickens, dietary nominal 8 weeks No adverse effects on food Cecil et al. 

White leghorn concentrations of 0, consumption, body weight, egg 1981 

laying hens, 10, 50, 100, or 500 production, egg weight, egg MRID 

27-weeks old ppm diflubenzuron shell thickness, fertility, 00156781 

10/dose group hatchability, or progeny 

development. Cecil et al. 

1981 

Diflubenzuron accumulated in [published in 

eggs and body tissues; 5 weeks the open 

after treatment, diflubenzuron literature] 

was not delectable in the egg, 

liver, fat, or muscle tissues of 

hens fed any of the dose levels 

of the compound. 

Growing male Diflubenzuron at from 1 day of No consistent differences over Kubena 1981 

broiler and dietary concentrations age to 98 days time on body weight, food 

layer chickens of up to 250 mg/kg consumption, or testes, liver, 

feed comb and feet weights. 

Layer-breed diflubenzuron was fed from 1 day of No effects on egg production, Kubena 1982 

chickens, males at levels of 0, 2.5, 25 age through a egg weight, eggshell weight, 

and females and 250 mg/kg feed laying cycle fertility, hatchability or progeny. 

NOS = Not otherwise specified. 

Appendix 4 - 3 
Appendix 5: Toxicity of diflubenzuron to terrestrial invertebrates 

Species 

House fly( Musca 

domestica) and parasitoid 

Muscidifurax raptor 

Gypsy moth predators: 

lacewing (Chrysopa 

carnea), ladybird beetle 

(Hippodamia 

convergens), Wasp 

parasite Trichograma 

pretiosum of bollworm 

(Heliothis) 

Honey bees, Apis 

mellifera L. 

Honey bees, Apis 

mellifera L. 

Rice swarming caterpillar 

adult Spodoptera 

mauritania 

Gypsy moth Lymantria 

dispar 

Gypsy moth Lymantria 

dispar 

earthworm (Eisenia 

fetida) 

earthworm (Eisenia 

fetida) 

Exposure 

Dimilin, topical 

exposure 

10 mg on 9-cm filter 

paper (contact); and 5 

ppm sugar-water fed to 

host. 

Dietary exposure at 

concentrations of 0.59, 

5.9, and 59 mg/kg diet 

for 10 days. Vehicle: 

Sugar syrup. 

Diflubenzuron (25% 

WP) formulation (100 

ppm a.i.) supplied in 

water and 60 ppm 

supplied in sucrose 

syrup to colonies of 

honey bees in outdoor 

cages. 

Dimilin 25-WP, dietary 

exposure 

topical exposure 

acute oral exposure 

soil exposure 

soil exposure 

Effects 

No effect to eggs or pupae at 

10,000 ppm. > 90% mortality to 

intermediate to late stage larvae at 

1.25 to 10 ppm. No effects to 

parasitoid. 

Lab rearing of hosts on 

diflubenzuron diets and raising 

parasites on those eggs. And raised 

lacewings from topically treated 

eggs and adults. Negative effects 

on lacewing and ladybird beetle in 

lab; egg hatch of beetle returned to 

normal after 30-40 d. 

Reduced brood production at the 

highest concentration. No effect at 

two lower concentrations. 

Brood production almost 

eliminated; treated bees consumed 

significantly less water and pollen 

cake and produced significantly less 

comb, brood, and new workers. 

Number of eggs increased in treated 

colonies. No significant differences 

in survival of treated bees, 

compared with controls and both 

treated and untreated colonies built 

queen cells when the original queen 

was removed. 

60-64% sterility at 10 ppm, 100% 

sterility at 100-1,000 ppm 

LD50 = 3.58 mg/kg (alder) 

LD50 = 8.96 mg/kg (douglas fir) 

LC50 = 0.06 ppm diet (alder) 

LC50 = 0.45 ppm diet (douglas fir) 

NOEC = 1 g Dimilin WP-25 per kg 

dry soil 

NOEC = 780 mg diflubenzuron per 

kg dry soil 

Reference 

Ables et al. 1975 

Ables et al. 1977 

Barker and 

Taber 1977 

Barker and 

Waller 1978 

Beevi and 

Dale1984 

Berry et al. 1993 

Berry et al. 1993 

Berends and 

Thus 1992 

Berends et al. 

1992 

Appendix 5 - 1 
Appendix 5: Toxicity of diflubenzuron to terrestrial invertebrates 

Species Exposure Effects Reference 

Nontarget insects Dimilin 25-WP - topical Considerable mortality and Broadbent and 

(lacewing Chrysopa exposure up to 300 ppm inhibition of molting to lacewing, Pree1984a 

oculata, braconid wasp and contact with treated but no effects to wasp or bug. 

Macrocentrus leaves 

ancylivorous, assassin 

bug Acholla 

multispinosa) 

consumption of treated 

host larvae 

reduced emergence of wasp, but no 

effect on lacewing. 

cockchafer Melolontha beech or sorrel leaves repellant effects and 100% ovicidal Büchi and 

melolontha, leaf beetle treated with 0.1% effect to chafer. Effective against Jossi1979 

Gastroidea viridula Dimilin 25-WP larvae and eggs of beetle. 

Honey bee Dimilin - topical LD50 = 52.9 mg/kg (3rd instar) Chandel and 

exposure LD50 = 45.51 mg/kg (4th instar) Gupta1992 

LD50 = 22.33 mg/kg (pupa) 

Bee Apis cerana indica Dimilin - topical LD50 = 56.15 mg/kg (3rd instar) Chandel and 

exposure LD50 = 49.13 mg/kg (4th instar) Gupta1992 

LD50 = 22.69 mg/kg (pupa) 

Spined soldier bug, Topical, residual, and Diflubenzuron harmless to De Clercq et al. 

Podisus maculiventris, oral exposure to predatory bug by direct and residual 1995b 

(predator) diflubenzuron 48% contact, but highly toxic when 

suspension concentrate. ingested via drinking water. Five 

days after adult emergence, LC50 

(for ingestion to 5th instar nymphs) 

= 7.20 μg/mL. 

Exposure of 5th instars to sublethal 

concentrations (around LC 10) had 

no adverse effects on reproduction 

of emerging adults. 

Flower bug, Orius 5th instar nymphs were LC50 (residual contact) = 391.1 mg Delbeke et al. 

laevigatus, predatory bug exposed to formulated a.i./L (95% CI = 140.5-825.6 mg 1997 

used as a biological diflubenzuron WP 25 a.i./L) 

control for thrips. N= 20 via ingestion of 

contaminated (saturated) LC50 (ingestion) = 229.9 mg a.i./L 

cotton wool plug and (95% CI = 108.0-397.3 mg a.i./L) 

residual contact for 3 

days. 

Migratory grasshopper Dimilin 25-WP, dietary LC50 = 0.08 ppm (lettuce diet) Elliott and 

Melanoplus sanguinipes exposure LC50 = 0.1 ppm (wheat seedling Iyer1982 

diet) 

Honey bee Dimilin - topical or LD50 > 30 :g/bee (topical) Gijswijt1978 

dietary exposure LD50 > 200 :g Dimilin WP-25 per 

bee (dietary). No adverse effects at 

5.9 ppm. 

Appendix 5 - 2 
Appendix 5: Toxicity of diflubenzuron to terrestrial invertebrates 

Species 

Rove beetle (Aleochara 

bilineata) and Cabbage 

maggot (target) 

Desert locust 

(Schistocerca gregaria) 

Mealworms, Tenebrio 

molitor, adults 

Gram pod borer, 

Helicoverpa armigera 

(Lepidotera: Noctuidae) 

[crop pest] eggs 0-24 and 

24-48 hours. 

Honey bee 

Honey bee 

Oxya japonica 

(Orthoptera ) 

Australian ladybird 

beetle, Cryptolaemus 

montrouzierei , adults 

(excellent predator of 

mealybug species) 

Gypsy moth 

Exposure 

Consumption of cabbage 

maggot treated with 

Dimilin 25-WP 

Dietary exposure 

10 mg/g technical 

Diflubenzuron 

incorporated into the 

diet (wheat flour) for 

period of ecysis to 9 

days 

eggs dipped for two 

minutes in different 

concentrations (NOS) of 

a suspension of 

diflubenzuron in 

distilled water. 

acute topical exposure 

acute oral exposure 

Dimilin 25-WP, topical 

exposure 

200 ppm diflubenzuron 

on treated surface 

Dimilin 25-WP, dietary 

exposure at 0.1 mg/kg 

Effects 

No adverse effects on rove beetle. 

Suppression of egg hatching and 

larva development of the cabbage 

maggot Delia radicum 

LD50 = 886.7 :g AI (2nd instar) 

LD50 = 207.4 :g AI (4th instar) 

LD50 = 325.2 :g AI (5th instar) 

Treatment quantitatively and 

qualitatively altered the lipid 

metabolism during sexual 

maturation. Fatty acid composition 

of the ovaries was not affected. 

IC50 (0-24 hours) = 0.0055 ppm 

(fiducial limits= 0.007-0.004 ppm) 

IC50 (24-48 hours) = 0.0061 ppm 

(fiducial limits= 0.01-0.0034 ppm) 

LD50 > 100 :g/bee (adult) 

LD50 >0.0125 :g/bee (larva) 

LD50 > 100 :g/bee (adult) 

LD50 > 0.030 :g/bee (larva) 

LD50 = 0.06 :g per insect or 0.31 

mg/kg 

No adverse effects on longevity or 

feeding; however treatment had 

effects on adult females, yielding 

only 278 progeny, compared with 

419 yielded by controls. 

100% lethal to larvae 

Reference 

Gordon and 

Cornect1986 

Jepson and 

Yemane1991 

Khebbeb et al. 

1997 

Kumar et al. 

1994 

Kuijpers1989 

Kuijpers1989 

Lim and 

Lee1982 

Mani et al. 1997 

Martinat et al. 

1988 

Appendix 5 - 3 
Appendix 5: Toxicity of diflubenzuron to terrestrial invertebrates 

Species 

Grasshopper, 

Poekilocerus pictus, 2­

day-old, virgin females 

Mexican bean beetle 

Lacewing, Chysoperla 

carnea, adults <24 hours 

old 

Lacewing, Chysoperla 

carnea, adults <24 hours 

old 

Honey bees, caged 

colonies 

Cotton leafworm 

Spodoptera littoralis 

Predacious phytoseiid 

mite, Amblyseius 

womersleyi, adult females 

Oncopeltus fasciatus, 

Large milkweed bug 

Exposure 

20 μg/insect 

Diflubenzuron dissolved 

in acetone applied on the 

ventral side of the 

abdomen. 

Dimilin 25-WP, dietary 

exposure 

topical application 

topical application 

10 mg/kg diflubenzuron 

for 10 weeks 

Dietary exposure 

Diflubenzuron (Dimilin) 

(25% pure) at field rate 

of 100 ppm on bean leaf 

disks dipped in test 

substance 

Topical exposure to 1 

μg/insect 

Effects	 Reference 

In few treated females, the abdomen Mathur 1998 

could not come out of the sand after 

egg laying and mortality occurred in 

the same position. When the 

abdomen was stretched back, the 

normal position was not attained 

again, which may be attributed to 

the chitin synthesis inhibiting 

activity of diflubenzuron. 

Ovaries of treated females were 

adversely affected by treatment, 

which probably accounts for the 

decrease in reproduction. 

LC50 = 3.4 ppm (3rd instar)	 McWhorter and 

Shepard 1977 

At a diflubenzuron at dose of 7,000 Medina et al.
 

ng/insect, no mortality among 2002
 

adults; 100% inhibition of egg
 

hatching due to death embryo. At
 

the lowest dose, 75 ng/insect), 32%
 

reduction in egg hatch.
 

LD50 = 2.26 ng/insect Medina et al.
 

LD10 = 0.74 ng/insect 2003
 

LD90 = 6.87 ng/insect
 

No effect on reproduction at a dose
 

of 0.5 ng/insect.
 

No adverse effects on pollen Nation et al.
 

consumption or brood production; 1986
 

however treatment resulted in a
 

50% decrease in the amount of
 

syrup stored.
 

LC50 = 1 mg/kg	 Neumann and 

Guyer1987 

No mortality 3 days after treatment.	 Park et al. 1996 

Inhibition of reproduction	 Redfern et al. 

1980 

Appendix 5 - 4 
Appendix 5: Toxicity of diflubenzuron to terrestrial invertebrates 

Species 

Brown lacewing, 

Micromus tasmaniae 

(beneficial predator) 

Brown lacewing, 

Micromus tasmaniae 

(beneficial predator) 

European earwig 

Forficularia auricularia 

Pieris brassicae (Large 

White Butterfly) 

Mealworms, Tenebrio 

molitor, adults 

Mealworms, Tenebrio 

molitor, adults 

Mealybug ladybird beetle, 

Crptolaemus 

montrouzieri, predator of 

mealybugs 

Honey bee 

Exposure 

contact exposure: 0.07 

μg/cm2

 a.i. as Dimilin 25 

WP sprayed on petri 

dishes 

contact exposure: 

Dimilin 25 WP sprayed 

on petri dishes 32 hours 

after the 2nd larval molt 

12.5 g a.i./ha 

Topical exposure 

5 or 10 mg/g 

Diflubenzuron (NOS) 

incorporated into diet for 

3 or 6 days post 

emergence. 

5 or 10 mg/g 

Diflubenzuron (NOS) 

incorporated into diet . 

Duration of exposure not 

clear. 

freshly emerged final 

instar nymphs were fed 

with mealy bugs treated 

with 0.153 ppm 

Diflubenzuron and 

sacrificed after 24, 48, 

72, or 96 hours. 

oral and contact LD 50 

values 

Effects 

Treatment caused a strong trend 

toward decrease in fertility where 

13% of all pairs did not lay any 

eggs; total numbers of eggs 

produced per females were reduced 

by approx. 50%; treated females 

deposited significantly fewer eggs 

per day than the control females 

(p<0.01). 

120 hour LC50 = 0.069% a.i. 

(95% CI: = 0.049-0.107% a.i.) 

360 hour LC50 = 0.009% a.i. 

(95% CI: = 0.003-0.012% a.i.) 

growth and mobility adversely 

affected 

LD50 = 2.5 :g/insect or 1.07 mg/kg 

Diflubenzuron had no significant 

effect on fat body protein. 

treatment caused a decrease in both 

the cell density of germarium and 

the thickness of chorion. 

There was a significant reduction in 

protein content after 2 hours; 

however, with prolonged exposure, 

the insect was found to adapt itself 

to the toxic stress and the adverse 

effect was much less pronounced 

after 96 hours. 

>30 :g/bee 

Reference 

Rumpf et al. 

1998 

Rumpf et al. 

1997 

Sauphanor et al. 

1993 

Sinha et al. 1990 

Soltani-Mazouni 

and Soltani 

1995a 

Soltani and 

Soltani-Mazouni 

1997 

Sundari et al. 

1998 

Stevenson 1978 

Appendix 5 - 5 
Appendix 5: Toxicity of diflubenzuron to terrestrial invertebrates 

Species 

Honey bee, Apis mellifera 

Fruit-sucking moth, 

Othreis materna, 5th 

instar larvae 

Honey bee colonies 

Nematodes 

German cockroach 

Blatella germanica 

Codling moth (Cydia 

pomonella), neonates of 

field-collected and 

laboratory strains 

Exposure 

0.1, 1, & 10 ppm in 

Sugar-cake for 12 wks. 

0.01, 0.1, & 1.0 ppm in 

sucrose syrup next year 

for 10 weeks. 

topical application of 0 

or 0.025 μL Dimilin (25 

WP) in 5 μL acetone to 

ventral region of the 

abdomen. Larvae were 

sacrificed 24, 48, or 72 

hours after exposure. 

Diflubenzuron diluted 

with sucrose to a rate 

equivalent to maximum 

application rate on 

flowering crops. 

10 day dietary exposure 

to Dimilin at 10 ppm 

Dimilin 25W® - contact 

with spray of treated 

cage plywood panels 

Dimilin WP 

Effects 

At 10 ppm diflubenzuron in sugarcake, significantly fewer sealed 

brood were produced, and colony 

size was reduced significantly 

compared to control and lower 

dosed colonies. No effects on 

brood production, colony size or 

adult bee mortality were seen the 

following year, when lower doses in 

a fluid solution was used. 

Degradation in sucrose solution 

might have reduced the potential for 

adverse effects. 

Inhibition of molting in larvae 

seems to occur due to 

neuroendocrine failure. See Section 

4.1.2.3. for discussion. 

Treatment with diflubenzuron 

resulted in short-term decrease in 

the numbers of adult bees and 

brood, compared with controls. No 

significant effect on development of 

brood during the following spring; 

however, there appeared to be a 

slower expansion, compared with 

controls. No adverse effects on 

queen viability. 

Adults unaffected but reproduction 

hindered and egg hatch prevented. 

Population reductions of 5% for 

Pelodera sp., 47% for Panagrellus 

redivivus, and 94% for 

Acrobeloides sp. 

population reduction of 67.3% at 30 2 2 mg/m , 93% at 60 mg/m , and

98.2% at 120 mg/m .2 egg hatch 

unaffected, but high first instar 

mortality. 

5-day LC50 = 13.9 mg/L

 (95% CI = 10.7-18.2 mg/L) 

Reference 

Stoner and 

Wilson 1982 

Tembhare and 

Shinde 1998 

Thompson and 

Wilkins 2003 

Veech 1978 

Wadleigh et 

al.1991 

Weiland 2000 

MRID 45245403 

Appendix 5 - 6 
Appendix 5: Toxicity of diflubenzuron to terrestrial invertebrates 

Species Exposure Effects Reference 

Honey bee Dimilin 25-WP, dietary	 LC 50 = 3.7 ppm Wittmann1982 

Honey bee Diflubenzuron dietary	 No toxicity at concentrations up to Yu et al. 1984 

1000 mg/kg in the diet. 

Stinkbug, Podisus Diflubenzuron sprayed No effect on egg viability. Zacarias et al. 

nigrispinus, eggs and on eggs and nymphs. 1998 

nymphs 

Host: Mexican bean 100, 1,000, and 10,000 Topical application to adults did not Zungoli et al. 

beetle (Epilachna ppm affect survival or reproduction, nor 1983 

varivestis). that of their progeny. Emergence of 

Parasite: wasp (Pediobius parasite from larvae treated after 

foveolatus). parasitism and before was 0 or 

nearly 0. 

Appendix 5 - 7 
Appendix 6: Toxicity of diflubenzuron to fish 

Species Nature of 

Exposure 

Exposure 

Time Effects a Reference 

Diflubenzuron 

Acute 

Bluegill sunfish, 

Lepomis 

macrochirus 

static renewal 

bioassay 

96 hours LC = 135 mg/L 50 Marshall and 

Hieb 1973 

MRID 00056150 

Fathead minnow static 96 hours LC > 500 mg/L 50 Reiner and Parke 

1975 

MRID 00060376 

Mummichog, 

Fundulus 

heteroclitus 

static renewal 

bioassay 

96 hours NOEC = 29.86 mg/L 

LC50 = 32.99 (CL = 29.01-37.52 

mg/L) 

Lee and Scott 

1989 

Rainbow trout, 

Salmo gairdneri 

static renewal 

bioassay 

96 hours LC = 140 mg/L 50 Marshall and 

Hieb 1973 

MRID 00056150 

Rainbow trout, 

Channel Catfish, 

and Bluegills 

static 96 hours LC > 100 mg/L 50 Johnson and 

Finley 1980 

Brook trout static 96 hours LC > 50 mg/L 50 Johnson and 

Finley 1980 

Yellow perch static 96 hours LC = 25 mg/L 50 Johnson and 

Finley 1980 

Rainbow trout static 96 hours LC = 240 mg/L as Dimilin 50

25-WP 

Julin and Sanders 

1978 

Channel catfish static 96 hours LC = 370 mg/L as Dimilin 50

25-WP 

Julin and Sanders 

1978 

Fathead minnow static 96 hours LC = 430 mg/L as Dimilin 50

25-WP 

Julin and Sanders 

1978 

Bluegill sunfish static 96 hours LC = 660 mg/L as Dimilin 50

25-WP 

Julin and Sanders 

1978 

Yellow perch static 96 hours LC > 50 mg/L 50 Mayer and 

Ellersieck, 1986 

Brook trout static 96 hours LC > 50 mg/L 50 Mayer and 

Ellersieck, 1986 

Cutthroat trout static 96 hours LC > 60 mg/L 50 Mayer and 

Ellersieck, 1986 

Atlantic salmon static 96 hours LC > 50 mg/L 50 Mayer and 

Ellersieck, 1986 

Appendix 6 -1 
Appendix 6: Toxicity of diflubenzuron to fish 

Nature of Exposure a Species Effects Reference Exposure Time 

Longer Term 

Fathead continuous exposure 10 months No effects on survival, growth, Cannon and Krize 

minnows to concentrations of behavior or reproduction, 1976 

0, 0.00625, 0.0125, compared with controls; no MRID 00099755 

0.025, 0.05, or 0.10 observable effects on 

ppm 99.4% pure hatchability of eggs spawned by 

TH-6040 (air fish. 

milled) 

Fry, hatched from eggs spawned 

by treated fish showed no 

appreciable differences, 

compared with controls after 60 

days of exposure to TH-6040, 

under same conditions as 

parental fish. 

Salmonids Diflubenzuron 96 hours or No effects at any concentration. Hansen and 

(steelhead trout) under flow-through 30 days Garton 1982a 

and non- conditions at (survival NOEC >45 μg/L (highest 

salmonids concentrations up to and growth concentration tested) 

(fathead 45 μg/L. in early life 

minnows and stages 

guppies) fish 

species 

Mummichug, Life cycle involving life cycle (2­ No significant dose-response Livingston and 

Fundulus continuous (flow generations) relationships. Koenig 1977 

heteroclitus through) exposure MRID 

(marine species) to TH-6040 014402120 

dissolved in acetone 

to deliver Livingston and 

concentrations of Koenig 1977 

0.003, 0.006, MRID 00099722 

0.0125, 0.025, or 

0.05 ppm 

Mesocosm 

Bluegill sunfish, Dimilin at nominal 70 days NOEC = 0.7 μg/L Moffett and 

Lepomis treatment levels of LOEC = 2.5 μg/L Tanner 1995 

macrochirus, 0.7, 2.5, 7.0, or 30 Secondary effects on endpoints In: Moffett 1995 

“young-of-the μg/L to littoral based on growth (individual fish MRID 44386201 

year” enclosures size). See additional notes 

below. 

Appendix 6 -2 
Appendix 6: Toxicity of diflubenzuron to fish 

Nature of Exposure a Species Effects Reference Exposure Time 

Additional Notes on Moffett and Tanner 1995: In indigenous fish species, mean fish size, population numbers, 

and biomass were not affected by exposure to diflubenzuron (#30 μg/L). Indigenous species included brook 

stickleback, northern redbelly dace, and central mudminnows. Young-of-the-year bluegill growth rates were 

directly correlated to the density of several invertebrates (cladoceran and copepods) in the enclosures and 

inversely correlated to the measured concentration of diflubenzuron. The results indicate that the indirect effects 

of diflubenzuron on bluegill sunfish were caused by a reduction in food resources due to the direct toxicity of the 

pesticide on the chitinous invertebrates preferred by the bluegill. 

Bluegill sunfish, Dimilin 25 W in reproductive Treatment adversely affected Tanner and 

Lepomis littoral enclosures at cycle reproductive success by Moffett 1995 

macrochirus nominal decreasing growth of young of In: Moffett 1995 

concentrations of the year bluegills at 2.5 and 3.0 MRID 44386201 

2.5 or 30 μg/L μg/L by eliminating or reducing 

preferred bluegill food choices 

(cladocerans and copepods). 

Additional Notes on Tanner and Moffett 1995: No behavioral effects related to reproduction of adult bluegills 

were observed in the enclosures. There was no clearly determined effect on spawning; however it appeared by 

spawning was influenced more by water temperature than by diflubenzuron. No direct effects on larvae prior to 

swim-up; however secondary effects on growth were evident following swim-up, apparently due to the precipitous 

decrease of zooplankton and the decline of chironomids and other macroinvertebrates. 

Bioconcentration 

Bluegill sunfish, dynamic 42-day 28 days In fillet, the BCF was 120 after 1 Burgess 1989 

Lepomis study to evaluate under flow- day and 170 after 28 days with a MRID 42258401 

macrochirus bioconcentration of through peak of 200 after 7 days. In 

C 14-diflubenzuron conditions, whole fish, the BCF was 260 

with 14 day after 1 day and 350 after 28 days 

depuration with a peak of 360 after 7 days. 

period 

White crappies 10 ppb DFB 24 hours BCF = 82.2 Schaefer et al. 

1979 

Bluegill sunfish 10 ppb DFB 24 hours Residues of approximately 848 Schaefer et al. 

ppb; 218 ppb in skin and 232 1979 

ppb in inner tissues (NOS); 

residues decreased rapidly when 

fish were transferred to the rinse 

tank for $48 hours. 

Appendix 6 -3 
Appendix 6: Toxicity of diflubenzuron to fish 

Nature of	 Exposure a Species	 Effects Reference Exposure Time 

Acute 

Bluegill Lepomis Static 

macrochirus 

Longer Term 

Medaka, Oryzias Larval growth; 

latipes flow-through 

Zebra fish growth and 

Brachydanio reproduction at 

rerio 0.04, 0.2, and 1 

mg/liter 

Zebra fish	 Flow-through 

Brachydanio 

rerio 

Bioconcentration 

Medaka, Oryzias	 Static aqueous 

latipes (Killifish)	 exposures to [14C]­

chloroaniline (8.9­

17 mCi/ 

mmol; >98% pure) 

for up to 320 

minutes 

Carp, Cyprinus	 continuous flowcarpio	 through exposure to 

0.30±0.07 or 

10.4±0.4 μg/L pchloroaniline 

p-Chloroaniline 

96 hour	 LC 50 value = 2.4 mg/L 

28 days	 MATC <2.25 mg/L 

5 weeks	 Adverse effects at 1 mg/L: 

abdominal swelling, spinal 

deformations, reduced number 

of eggs, and reduced fertilization 

in the F1 and F2 generations. 

3 weeks	 NOEC for Mortality and other 

effects = 1.8 mg/L 

up to 320	 Due to low elimination rates, 

minutes	 20% of the absorbed dose 

remained within the fish through 

330 minutes after exposure. Nacetylation was the dominant 

route of in vivo metabolism, with 

no indication of ring 

hydroxylation. 

up to 335 average BCF in whole body 

hours (about were 1.7 (low concentration) and 

14 days) 0.8 (high concentration). 

a

 Values in parentheses are the 95% confidence limits. 

WHO 2003 

WHO 2003 

Bresch et al. 1990 

WHO 2003 

Bradbury et al. 

1993 

Tsuda et al. 1993 

Appendix 6 -4 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species 

Grass shrimp, 

Palaemonetes pugio 

Grass shrimp, 

Palaemonetes pugio 

Hydropsychidae 

(Trichoptera) 

Mysid shrimp, 

Mysidopsis bahia, F1 

second generation 

Mysid 

shrimp,Mysidopsis 

bahia, juvenile 

Exposure Time 

Subchronic exposure 

to measured 

concentrations of 

0.70, 1.73, 5.51, 6.79, 

or 16.4 μg/L for 35 

days in flowing 

seawater 

Acute exposure to 

nominal 

concentrations of #1.0 

mg/L TH-6040 in 

static seawater 

Dimilin 25-WP, 15 

days at 0.0025 to 0.25 

mg/L 

mean measured 

concentration of 123 

ng/L (0.123 μg/L) 

diflubenzuron (97.6% 

pure) for up to 5 days 

Continuous exposure 

to mean measured 

concentrations of 29, 

45, 86, 140 or 210 

ng/L diflubenzuron 

through entire life 

cycle over a 28-day 

test period. 

Juvenile mysids 

produced during the 

test at the lowest four 

test concentrations 

(29-140 ng/L) were 

continuously exposed 

for the 8 days of the 

28-day test. 

Effects a 

No survival to day 7 among zoea 

exposed to initial measured 

concentrations of 5.5, 6.8, or 16.4 μg/L; 

survival among shrimp exposed to 0.70 

or 1.73 μg/L was significantly less than 

survival among controls; no significant 

difference in size of shrimp exposed to 

0.70 or 1.73 μg/L, compared with 

controls. 

96-hour LC50 = 0.64 mg/L 

(0.13-3.1 mg/L) 

No adult emergence from treated tanks 

and only 31.6% emergence from control 

tanks 

upon removal of treated water, juvenile 

second generation mysids completely 

recovered and had survival and 

reproductive success similar to that of 

the controls. 

F0 survival at 86, 140, and 210 ng/L was 

significantly reduced (p#0.05) compared 

with controls; treatment caused 

significant reduction in growth and 

development (as measured by dry 

weight) in F0 males (210 ng/L) and F 0 

females (140 and 210 ng/L); 

reproduction of F mysids was 

significantly reduced at 86, 140, and 210 

ng/L. 

0

The NOEC = 86 ng/L for growth LOEC 

= 140 ng/L for growth. 

Survival of the second generation (F ) 

mysids was not affected by continuous 

exposure to any of the mean measured 

concentrations tested (21, 33, 83, or 123 

ng/L). The NOEC after 8 days of 

1

exposure of F generation mysids was 

>83 ng/L. 

1

Reference 

BionomicsEG&G 1975 

MRID 00038612 

BionomicsEG&G 1975 

MRID 00038612 

Bradt and 

Williams 1990 

Breteler 1987 

MRID 40237501 

Breteler 1987 

MRID 40237501 

Note: This 

summary is of the 

primary study on 

which the studies 

discussed below 

are based. 

Appendix 7 - 1 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species 

Mysid shrimp, 

Mysidopsis bahia 

Marine crabs, 

Pontonia 

pinnophylax, larvae 

Mixed aquatic 

invertebrates (i.e., 

cladocerans, rotifers, 

and adult amphipods) 

Exposure Time 

24-hour exposure to 

mean concentration of 

298 ng/L 

diflubenzuron (97.6% 

pure), followed by 

transfer to clean 

control water for 27 

days. 

#10 ppb 

diflubenzuron 

Microcosm 1: nominal 

concentrations of 0.3, 

0.7, 1.4, 3.4, 6.8, or 

13.6 μg/L Dimilin 

25W 

Microcosm 2: nominal 

concentrations of 1.4. 

3.4. 6.8, or 20.0 μg/L 

Dimilin 25W 

Effects a 

Survival, growth, and reproductive 

success similar to that of controls. 

larvae of four different crab species 

appeared normal during inter-molt 

periods and adverse effects were 

apparent until molting (similar to effect 

of DFB on insect larvae). 

Treatment deformed both the exocuticle 

and the endoculticle and was lethal to all 

four species of marine crabs. 

Major effect of diflubenzuron in the 

microcosms was on the cladocerans. 

Population density was decreased within 

3-4 days after treatment at $0.7 μg/L 

and remained consistently low, 

compared with controls throughout the 

study duration. Statistically significant 

(p#0.05) differences in population 

density at $1.4 μg/L in Microcosm 1 

between days 3 and 10 and at $0.7 μg/L 

in Microcosm 2 between days 4 and 14. 

Cladoceran population densities did not 

generally increase in either microcosm 

at $0.7 μg/L. 

Rotifers were no adversely affected by 

treatment at any concentration. 

The numbers of adult amphiphods 

(Hyalella azteca) were significantly 

different from controls (p#0.05) at 13.6 

μg/L (Microcosm 1) and 20 μg/L 

(Microcosm 2). Amphipods exposed to 

concentrations <13.6 μg/L were not 

different (p#0.05) from controls in 

either experiment. 

NOEC for cladocerans = 0.3 μg/L 

LOEC for cladocerans = 0.7 μg/L 

Reference 

Breteler 1987 

MRID 40237501 

Christiansen 1987 

Corry et al. 1995 

In: Moffett 1995 

MRID 44386201 

Appendix 7 - 2 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species Exposure Time 

Fiddler crabs, Uca repetitive 24-hours 

pugilator, juveniles weekly exposures to 

0.2, 2, 20, or 200 μg/L 

Dimilin in static 

seawater systems for 

10 weeks. 

Barnacles, Balanus Exposure to1-1000 

eburneus, Cirripede μg/L technical grade, 

crustaceans. air-milled 

diflubenzuron 

w/acetone as carrier 

solvent (preliminary 

studies showed no 

mortality in acetone 

controls) for 28 days 

Ceriodaphnia dubia, Exposure to 0.50, 

neonates, <12 hours 0.75, 1.0, 1.5, 2.0, 3.0, 

old or 4.0 ng/mL Dimilin 

for 48 hours. 

Ceriodaphnia dubia Chronic exposure to 0, 

0.05, 0.1, 0.25, 0.5, 

0.75, or 1.0 ng/mL 

(μg/L). Used 

methanol carrier with 

carrier control. 

Effects a 

NOEC (time to first molt) =20μg/L 

NOEC (survival) = 2 μg/L 

NOEC (ability to escape from test 

container) = 0.2 μg/L 

Behavioral effect caused by DFB 

exposure ($2 μg/L) was most sensitive 

indicator of DFB toxicity. 

Investigators conclude that survival, 

molting, and behavior of juvenile fiddler 

crabs are significantly affected by 

exposure to repetitive applications of 

DFB. 

Dose-dependent mortality, with drastic 

mortality observed during the second 

week of exposure. Lethal and sublethal 

effects were observed at 

concentrations as low as 50 μg/L 

Disruption of the exoskeleton caused by 

diflubenzuron was similar to that 

observed in insects. 

Development of barnacles exposed to 

diflubenzuron for 10 days or more at 

750 and 1000 μg/L was delayed in the 

pre-molt phase of cuticle secretion 

48-hr NOEC = 0.75 ng/mL [0.75 μg/L] 

48-hr LC50 =1.7 ng/mL (95% CI = 1.36­

2.02 ng/mL) [1.7 μg/L] 

NOEC = 0.25μg/L 

At $0.5 μg/L, significant decrease in 

numbers of neonates produced, 

compared with controls; at 0.75 and 1.0 

μg/L, adults produced no viable young; 

mortality increased at exposures to >0.1 

μg/L. 

No carrier effect: 31.7 (28.4-34.9) 

neonates/female with 20% mortality in 

adults in untreated control and 30.9 

(26.9-35) in carrier control with 10% 

mortality in adults. 

Reference 

Cunningham and 

Meyers 1987 

Gulka et al. 1980 

Hall 1986 

MRID 40130601 

Hall 1986 

MRID 40130601 

Appendix 7 - 3 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species Exposure Time Effects a Reference 

CRITICAL NOTE on HALL 1986: Hall (1986) reports concentrations as nanograms/mL. These are converted 

above to μg/L. 

Daphnia magna Diflubenzuron under 

static conditions for 

48 hours 

Midges, Tanytarsus	 Diflubenzuron under 

dissimilis, 2nd to 3rd	 flow-through 

larval instar	 conditions for 5 days; 

effect criteria = 

molting success 

Midges, Cricotopus,	 Diflubenzuron under 

sp, 4th larval instar to	 flow-through 

pupae	 conditions for 7 days; 

effect criteria = 

molting success 

Daphnia magna Survival and 

reproduction in full 

life cycle after 

exposure to 

diflubenzuron 

(conditions not 

specified) 

Freshwater molluscs	 Diflubenzuron under 

(two species of snails)	 flow-through 

conditions for 96 

hours; effect criteria 

for chronic exposure 

(3 weeks) = survival, 

growth and 

reproduction 

Stream invertebrates	 Technical 

(most abundant),	 diflubenzuron in 

including	 dimethlformamide at 

Ephemeroptera,	 0.1, 1, 10, and 50 

Plecoptera, Diptera,	 μg/L added 

Tricoptera, and	 continuously to 

Coleoptera.	 complex laboratory 

stream channels 

supplied periodically 

with field-collected 

microorganisms for 5 

months 

LC50 = 1.84 μg/L 

(95% CI = 0.05-3.71 μg/L) 

LC50 = 1.02 μg/L 

(95% CI = 0.56-1.47 μg/L) 

LC50 = 1.79 μg/L 

(95% CI = 1.48-2.13 μg/L) 

LC50 = 0.062 μg/L 

(95% CI = 0.051-0.071 μg/L) 

NOEC 45 μg/L (highest concentration 

tested) 

Invertebrates were most adversely 

affected undergoing rapid and 

permanent reductions in biomass and 

diversity at diflubenzuron concentrations 

of $1.0 μg/L. These effects were the 

results of major in reductions in many of 

the aquatic insect populations, primarily 

among mayflies, stoneflies and diptera. 

Hansen and 

Garton 1982a 

Hansen and 

Garton 1982a 

Hansen and 

Garton 1982a 

Hansen and 

Garton 1982a 

Hansen and 

Garton 1982a 

Hansen and 

Garton 1982a 

Appendix 7 - 4 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species Exposure Time Effects a Reference 

Additional Notes on Hansen and Garton 1982a: Diversity in all groups of stream invertebrates was clearly 

dose-related with little or no reductions observed at 0.1 μg/L, intermediate reductions observed at 1.0 μg/L (some 

dipteran tax were relatively insensitive at this concentration but eliminated at higher concentrations), and maximal 

reductions observed at $10.0 μg/L. 

Algal, fungal, and bacterial functional groups were also adversely affected by exposure to diflubenzuron. 

Generally the adverse effects observed among these organisms was variable and transient alterations in biomass 

and diversity with algae and bacteria affected at 1.0 μg/L and fungi affected at as little as 0.1 μg/L. 

Total biological 8- month continuous Insects were directly affected at $1.0 Hansen and 

community in 8 exposure to 0.1, 1.0, μg/L (stoneflies and mayflies were the Garton 1982b 

stream microcosms 10, or 50 μg/L most sensitive with adverse effects 

diflubenzuron apparent at 1.0 μg/L, dipterans affected 

dissolved in at 10.0 μg/L, and coelopterans were not 

dimethylformamide affected at any test concentrations); 

Algae and fungi were mildly affected at 

$1.0 μg/L, but the effects were 

considered indirect in response to the 

decreases in herbivore and shredder 

components of the insects; 

No effects were observed in bacteria, 

oligochaetes, or gastropods at any test 

concentration. 

Gammarid, Hyallela Diflubenzuron under LC50 = 1.84 μg/L Hansen and 

azteca (Benthic flow-through (95% CI = 0.05-3.71 μg/L) Garton 1982a 

crustacea) conditions for 96 

hours 

Stoneflies, Peltoperla DFB-treated yellow Peltoperla: survival significantly Harrahy et al. 

arcuata and poplar leaves via different from controls at day 60; 1994 

Pteronarcys proteus ingestion for 24-hours however survival of Pteronarcys was 

with 60- and 90-day not significantly different from controls 

observation periods. at 90 days, although the low number of 

molts that occurred during that time may 

have influenced the results. 

Appendix 7 - 5 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species 

Stoneflies, Peltoperla 

arcuata 

Mayflies, Cyngmula 

subaequalis, 

Stenacron 

interpunctatum, 

Stenonema 

meririvulanum, and S. 

femaratum 

Daphnids, Daphnia 

magna 

Fairy shrimp, 

Streptocephalus 

sudanicus, females 

Backswimmer, 

Anisops sardeus, 

females 

Daphnia magna 

Exposure Time 

nominal 

concentrations of 0, 

1.0, 10, 100, or 1000 

ppb DFB in 

dechlorinated tap 

water for 96 hours and 

then transferred to 

glass chambers 

containing pesticidefree water and fed 

stream conditioned red 

maple and white oak 

leaves. 

0, 0.6, 5.6, 55.7, or 

557.2 ppb DFB 

(Dimilin 25% WP) in 

water for 96 hours 

then placed in 

pesticide-free water 

for 36-day observation 

period 

48-hour exposure to 

diflubenzuron (97.6% 

pure) 

Dimilin (solventbased, liquid ULV 

formulation) for 24 or 

48 hours under static 

conditions 

Dimilin (solventbased, liquid ULV 

formulation) for 24 or 

48 hours under static 

conditions 

Technical grade 

diflubenzuron 

(TH-6040) 

Effects a 

Survival at 10 and 1000 ppb was 

significantly different from controls; 

however, survival at 100 ppb was not 

significantly different from survival of 

controls. No behavioral changes were 

observed. 

after 4 days of exposure, mayflies were 

significantly lower than controls al all 

concentrations tested. At the lowest 

concentration, only about 45% survived 

to day 36. Many of the treated mayflies 

died while molting, while others died 

from incomplete hardening of the new 

cuticle. 

Behavioral changes observed included 

decreased swimming speed at higher 

concentrations, and no avoidance of 

pipet or hands during water replacement 

activities. Some mayflies were observed 

to shake sporadically before dying. 

48-hour NOEC = 0.45 μg/L 

48-hour EC50 = 7.1 μg/L (95% CI = 5.0­

1.0 μg/L) 

24-hour EC50 =13.3 μg/L 

(range = 12.8-14.0 μg/L) 

48-hour EC50 =0.74 μg/L 

(range = 0.60-0.88 μg/L) 

24-hour EC50 =2123 μg/L 

(range = μg/L) 

48-hour EC50 =1937 μg/L 

(range = 1800-2020 μg/L) 

LOEC for reproduction: 0.09 ppb 

Reference 

Harrahy et al. 

1994 

Harrahy et al. 

1994 

Kuijpers 1988 

MRID 40840502 

Lahr et al. 2001 

Lahr et al. 2001 

LeBlanc 1975 

Appendix 7 - 6 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species 

Blue crabs, 

Callinectes sapdidus, 

embryos 

Littoral enclosure 

community of mixed 

insects 

Littoral zooplankton 

community dominated 

by cladocera, 

copepoda, rotifera, 

and ostracoda. 

Chironomus 

plumosus, 4th instar 

larvae 

Exposure Time 

acute toxicity; 

diflubenzuron 

exposure in culture 

plates 

Dimilin at nominal 

treatment levels of 

0.7, 2.5, 7.0, or 30 

μg/L to littoral 

enclosures 

Dimilin at nominal 

treatment levels of 

0.7, 2.5, 7.0, or 30 

μg/L to littoral 

enclosures. 

Dimilin 25-WP, 48 

hour exposure 

Effects a 

hatching EC50 =1.8 μg/L 

EC50 = 1.2 μg/L (measured 

concentration) 

NOEC = 1.0 μg/L (measured 

concentration) 

LOEC = 1.9 μg/L (measured 

concentration) 

Cladocera were extremely sensitive to 

treatment, with mean population 

abundances significantly reduced, 

compared with controls, at all four 

treatment levels. Mean population 

densities at $2.5 μg/L were 92 to >99% 

lower than mean control values by day 6 

and remained at those levels through day 

56. None of the decreased populations at 

$2.5 μg/L showed any sign of recovery 

throughout the study. 

Copepoda were adversely affected by 

treatment at all concentration levels. 

LOEC = 0.7 μg/L. The measured peak 

diflubenzuron concentration in water 

was 1.0 μg/L. Copepoda were 

significantly affected at this level, not 

unlike the Cladocera. The NOEC for 

both Claodcera and Copepoda was 

defined as <0.7 μg/L; however the 

effects at 0.7 μg/L appeared to be 

transistory with recovery after a single 

application observed within 12-29 days. 

Ostracoda densities were reduced at the 

two highest concentrations. 

NOEC = 2.5 μg/L 

Rotifera were not affected by treatment 

at any concentration level. 

NOEC = >30 μg/L. 

EC50 = 0.56 mg/L 

Reference 

Lee and Oshima 

1998 

Liber 1995 

In: Moffett 1995 

MRID 44386201 

Liber and 

O’Halloran 1995 

In: Moffett 1995 

MRID 44386201 

Published as 

Liber et al. 1996 

and as O’Halloran 

et al. 1996 

Julin and Sanders 

1978 

Appendix 7 - 7 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species 

Daphnia magna 

Dragonfly nymphs 

Orthemis spp., 

Pantala sp. 

Mayfly nymphs 

Callibaetis sp. 

Aedes 

nigromaculatum 

Water scavenger 

beetle larvae 

Hydrophilus 

triangularis 

Water scavenger 

beetle adults 

Laccophilus spp., 

Thermonectus 

basillaris, 

Tropisternus lateralis 

Mysid 

shrimp,Mysidopsis 

bahia 

Littoral enclosure 

community of mixed 

benthic 

marcroinvertebrates,pr 

edominantly, 

Chironomidae 

(midges), Oligochaeta 

(earthworms), and 

Mollusca 

Exposure Time 

Dimilin 25-WP® - 48 

hour exposure 

TH 6040 

(diflubenzuron) - 168 

hour exposure 

TH 6040 

(diflubenzuron) - 168 

hour exposure 

TH 6040® 

(diflubenzuron) - 48 

hour exposure 

TH 6040® 

(diflubenzuron) - 48 

hour exposure 

TH 6040® 

(diflubenzuron) 

concentrations as high 

as 250 :g/L 

life-cycle exposure 

under flow-through 

conditions 

Dimilin at nominal 

treatment levels of 

0.7, 2.5, 7.0, or 30 

μg/L to littoral 

enclosures. Study 

duration = 71 days. 

Effects a	 Reference 

LC50 = 0.00075 mg/L (neonate) Majori et al. 1984 

LC50 = 0.02345 mg/L (adult) 

LC50 = 50 :g/L	 Miura and 

Takahashi 1974 

LC90 = 10 :g/L	 Miura and 

Takahashi 1974 

LC50 = 0.5 :g/L	 Miura and 

Takahashi 1974 

LC50 = 100 :g/L	 Miura and 

Takahashi 1974 

no mortality	 Miura and 

Takahashi 1974 

96-hour LC50 = 2.1 μg/L Nimmo et al. 

21-day LC50 = 1.24 μg/L 1979 

direct adverse effect on reproduction: 

the numbers of juveniles/female were 

significantly depressed at all nominal 

concentrations (0.075-0.75 μg/L) 

Reductions in abundance of O’Halloran and 

Ephemeroptera (mayflies) and Odonata Liber 1995 

(damselflies and dragonflies) were In: Moffett 1995 

observed at all nominal concentrations MRID 44386201 

$2.5 μg/L. 

No adverse effects were observed on 

molluscs or earthworms at any of the 

four diflubenzuron test concentrations. 

Overall, the only benthic 

macroinvertebrate group that appeared 

to have been adversely affected by 

exposure to diflubenzuron was the 

Insecta. 

Appendix 7 - 8 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species 

Stoneflies 

(Pteronarcys proteus 

and Pteronarcys 

arcuata) 

Blue crabs, 

Callinectes sapidus, 

juveniles 

Copepods, 

Eurytemora affinis, 

naupli 

Copepods, 

Eurytemora affinis, 

naupli 

Daphnids, Daphnia 

magna 

Quahog clams, 

Mercenaria 

mercenaria 

Exposure Time 

fed leaves from 

treated poplar after 

conditioning in stream 

Dimilin WP-25 in 

static renewal tests 

0.78 μg/L WP25 

commercial DFB 

(25% DFB, 75% 

kaolin) and filtered 

river water for 5 or 6 

days 

WP25 commercial 

DFB (25% DFB, 75% 

kaolin) and filtered 

river water. 

Continuous exposure 

to 14-C-diflubenzuron 

nominal 

concentrations of 6.3­

100 ng/L (mean 

measured 

concentrations of 5.6, 

14, 23, 40, or 93 ng/L) 

under flow-through 

conditions for 21 days 

(one generation) 

48-hour exposure to 

nominal 

concentrations of 100 

or 500 μg a.i./L (mean 

measured 

concentrations of 79, 

or 320 μg a.i./L) of 

diflubenzuron (97.6% 

pure) 

Effects a 

No effect on survival. 

both molt stage and renewal frequency 

affected toxicity: 

LC50 (random molt stages) = 3.5 mg/L 

LC50 (day of molt) = 300 μg/L 

LC50 (day of molt and repeated dosing) 

= 18.5 μg/L 

0% survival at >1.69 μg/L; at 0.93 μg/L 

survival did not differ significantly from 

controls. 

48-hour LC50 =2.2 μg/L 

50% survival at 93 ng/L[0.093 μg/L]; 

survival at the other test concentrations 

ranged from 93 to 98%, comparable to 

controls. 

significant reduction in reproduction and 

body length at 93 ng/L, compared with 

controls (p#0.05); at other test 

concentrations, reproduction and growth 

were comparable to controls. 

NOEC = 40 ng/L [0.04μg/L] 

No adverse effects on development of 

quahog embryos and larvae 

NOEC >320 μg a.i./L 

Reference 

Perry 1995a 

Rebach 1996 

Savitz et al. 1994 

Savitz et al. 1994 

Surprenant 1988 

MRID 40840501 

Surprenant 1989 

MRID 41392001 

Appendix 7 - 9 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Reference 

Tourat and Rao 

1987 In: 

Technology 

Sciences Group 

1998 

MRID 44399307 

Tourat and Rao 

1987 In: 

Technology 

Sciences Group 

1998 

MRID 44399307 

Tourat and Rao 

1987 In: 

Technology 

Sciences Group 

1998 

MRID 44399307 

Weis and Ma 

1987 

Wilcox and
 

Coffey 1978
 

Species 

Grass shrimp, 

Palaemonetes pugio 

Grass shrimp, 

Palaemonetes pugio 

Grass shrimp, 

Palaemonetes pugio, 

Horseshoe crabs, 

Limulus polyphemus, 

eggs 

snail Physa sp. 

Exposure Time 

continuous exposure 

to 1-10 μg/L from 

inter-molt to molt 

(normally 7-14 days) 

and transfer to filtered 

seawater 

24-hour pulsed 

exposure with transfer 

to DFB-free medium 

96 hours 

0, 5, or 50 μg/L DFB 

acute exposure 

Effects a 

Mortalities generally related to molt 

cycle with death occurring at the time of 

ecdysis or immediately after (LC50 = 

0.65 μg/L); at concentrations of 7.5-10 

μg/L, some shrimp did not die during 

the exposure period and displayed 

delayed progress in the molt cycle, and 

although these shrimp began progressing 

through the molt cycle when transferred 

to filtered seawater, they all failed to 

reach ecdysis and eventually died. 

Control shrimp were never observed in 

an arrested stage in the molt cycle 

during the experiment. 

LC50 =3.4 μg/L (premolt animals D1 ­

D )2

LC50 =1.1 μg/L (premolt animals D1 ­

D )2

very few or no mortalities among shrimp 

in very late premolt, early premolt, 

intermolt, or early postmolt stages 

during the 96-hour exposure. 

at 5 μg/L, crabs showed a slight, but 

significant (p<0.05) delay in molt at 14 

days, then molted at a rate comparable 

to controls and did not exhibit 

significant mortality. 

At 50 μg/L, molted at the same rate as 

controls but exhibited significant 

mortality immediately after ecdysis. 

Also, the prosomal width of the crabs in 

this group was smaller, compared with 

controls and crabs in the low dose 

group. 

LC50 > 125 ppm 

Appendix 7 - 10 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species Exposure Time Effects a Reference 

Grass shrimp, continuous exposure No correlation between age of the Wilson 1997b 

Palaemonetes pugio, for 4 days to 0.3-5.0 embryos at exposure and either 

ovigerous carrying μg/L DFB in static hatchability or duration of larval 

0.5-, 1-,3-, 6-, or 8­ system with transfer development; severity of abnormality 

day old embryos after exposure to did not vary with the age of the embryos 

DFB-free seawater for except at exposure concentration of 2.5 

rest of the embryonic μg/L. 

development. 

Larval viability was significantly 

(p<0.05) affected by the age of the 

embryos at the time of exposure to DFB, 

with older embryos more sensitive to 

sublethal effects of DFB. 

Grass shrimp, 96 hours under static larvae and post-larvae most sensitive to Wilson and 

Palaemonetes pugio at renewal conditions acute toxicity of DFB with LC50 values Costlow 1987 

different life stages of 1.44 and 1.62 μg/L, respectively; 

(embryos, larvae, ovigerous females (hence embryos) 

postlarvae male and appeared to be the most resistant to the 

female non-spawning acute toxicity of DFB with a mean LC50 

adults, and ovigerous of 6985 μg/L. 

females. 

Grass shrimp, chronic exposure to 72-hr and 96-hr calculated LC50 values Wilson and 

Palaemonetes pugio either technical grade were similar for the two formulations of Costlow 1986 

DFB (98.4% a.i.) Or DFB (WP-25 and TG): 

the wettable powder 

(WP-25) (25% a.i.) 72-hr LC50 = 2.95 μg/L (TG) 

72-hr LC50 = 2.83 μg/L (WP-25) 

96-hr LC50 = 1.84 μg/L (TG) 

96-hr LC50 = 1.39 μg/L (WP-25) 

The investigators conclude that results 

from studies using technical grade DFB 

are applicable to the WP-25 formulation 

without the need for a “correction 

factor.” 

Appendix 7 - 11 
Appendix 7: Toxicity of diflubenzuron to aquatic invertebrates 

Species Exposure Time 

Copepods, 0, 0.5, 0.78, or 0.93 

Eurytemora affinis, ppb DFB under pulse 

naupli, 24- to 48-hours (two 6.5 exposure 

old, initially periods) and 

continuous (14-day) 

exposure regimens. 

a

Values in parentheses are 95% confidence limits. 

Effects a Reference 

In pulse exposures, copepods exposed in Wright et al. 1996 

the first 6.5 days showed a significantly 

lower survival rate at 0.78 and 0.93 ppb; 

copepods exposed during the second 

half of the experiment showed no 

significant differences in mortality, 

compared with controls. 

In the 14-day continuous exposure, 

survival was significantly lower at 0.78 

and 0.93 ppb, but was significantly 

higher than that in the early pulse 

exposure to 0.78ppb. 

Effects on brood production were 

observed at 0.8 ppb in individuals 

exposed only during the copepodite 

stages. Significant effects on production 

of naupli were observed only in the first 

6.5 days of pulse exposure to 0.93 ppb. 

At salinities of 2, 10, and 15 ppt, 

survival from naupilar to adult stages 

was significantly reduced at 0.84 ppb 

and none survived to adulthood at 1.7 

ppb. 

Appendix 7 - 12 
Appendix 8: Toxicity of diflubenzuron to aquatic plants 

a Species Exposure Effects Reference 

Phytoplankton 

communities in 

littoral enclosures 

Periphyton 

communities in 

littoral enclosures 

Macrophyte 

populations in 

littoral enclosures 

Blue-green algae, 

Plectonema 

boryanum 

Dimilin at nominal 

treatment levels of 0.7, 

2.5, 7.0, or 30 μg/L to 

littoral enclosures 

Dimilin at nominal 

treatment levels of 0.7, 

2.5, 7.0, or 30 μg/L to 

littoral enclosures 

Dimilin at nominal 

treatment levels of 0.7, 

2.5, 7.0, or 30 μg/L to 

littoral enclosures 

0.1 ppm TH-6040 in 

pure culture for 4 days 

ALGAE 

Phytoplankton, as measured by cell 

size distributions and chlorophyll a 

in the enclosures, were not affected 

directly or indirectly by 

diflubenzuron treatment. No 

occasions of significant (p#0.05) 

linear correlations between the 

nominal concentrations of 

diflubenzuron and phytoplankton 

measures. These results were 

consistent with the idea that 

diflubenzuron does not directly 

inhibit non-chitinous biota due to 

the specificity of its mode of action. 

Late in the season (September), a 

80 and 90% reduction in periphyton 

dry weight and 75 and 80% 

reduction in chlorophyll a at 7.0 

and 30 μg/L treatment levels, 

respectively. Differences were 

statistically significant (p=0.01) on 

day 55 and nearly significant 

(p=0.07) on day 67. 

No adverse effects, direct or 

indirect, were observed on 

macrophyte species composition or 

total standing crop. There was no 

correlation between treatment 

concentrations and total macrophyte 

density throughout the study. 

The investigator indicates that 

direct effects were not anticipated 

because macrophytes do not have 

chitin. 

No growth inhibition, rapid 

metabolism of compound in water. 

Algae degraded 80T of compound 

in 1-hour incubation period to pchlorophenyl urea and pchloroaniline. 

Moffett 1995 

In: Moffett 1995 

MRID 44386201 

Moffett 1995 

In: Moffett 1995 

MRID 44386201 

Moffett 1995 

In: Moffett 1995 

MRID 44386201 

Booth and Ferrell 

1977 

Appendix 8 - 1 
Appendix 8: Toxicity of diflubenzuron to aquatic plants 

Species Exposure Effects a Reference 

Freshwater algae 300 μg/L diflubenzuron NOEC = 300 μg/L Thompson and 

Selenastrum for 5 days Swigert 1993b 

capricornutum MRID 42940104 

Freshwater algae, 120 hour exposures; NOEC 45 μg/L (highest Hansen and 

Selenastrum effect criteria = growth concentration tested) Garton 1982a 

capricornutum 

Freshwater 380 μg/L for 5 days NOEC = 380 μg/L Thompson and 

diatoms (Navicula Swigert 1993c 

pelliculosa) MRID 42940105 

Marine diatoms 270 μg/L for 5 days NOEC = 270 μg/L Thompson and 

(Skeletonema Swigert 1993d 

costatum) MRID 42940106 

MACROPHYTES 

Macrophyte Dimilin at nominal No adverse effects, direct or Moffett 1995 

populations in treatment levels of 0.7, indirect, were observed on In: Moffett 1995 

littoral enclosures 2.5, 7.0, or 30 μg/L to macrophyte species composition or MRID 44386201 

littoral enclosures total standing crop. There was no 

correlation between treatment 

concentrations and total macrophyte 

density throughout the study. 

Duckweed (Lemna 190 μg/L diflubenzuron NOEL = 190 μg/L Thompson and 

gibba) for 14 days Swigert 1993a 

MRID 42940103 

Appendix 8 - 2 
 

 

 

 

 

 

 

 

 

 

 

Pesticide Precautionary Statement 

This publication reports research involving pesticides. It does not contain recommendations for their 

use, nor does it imply that the uses discussed here have been registered. All uses of pesticides must 

be registered by appropriate State and/or Federal agencies before they can be recommended. 

CAUTION: Pesticides can be injurious to humans, domestic animals, desirable plants, and fish or 

other wildlife--if they are not handled or applied properly. Use all pesticides selectively and carefully. 

Follow recommended practices for the disposal of surplus pesticides and pesticide containers. 

The use of trade or firm names in this publication is for reader information and does not imply endorsement 

by the U.S. Department of Agriculture. 

The U.S. Department of Agriculture (USDA) prohibits discrimination in all its programs and activities on the basis of 

race, color, national origin, age, disability, and where applicable, sex, marital status, familial status, parental status, 

religion, sexual orientation, genetic information, political beliefs, reprisal, or because all or part of an individual’s 

income is derived from any public assistance program. (Not all prohibited bases apply to all programs.) Persons 

with disabilities who require alternative means for communication of program information (Braille, large print, 

audiotape, etc.) should contact USDA’s TARGET Center at (202) 720-2600 (voice and TDD). To file a complaint 

of discrimination, write to USDA, Director, Office of Civil Rights, 1400 Independence Avenue, S.W., Washington, 

D.C. 20250-9410, or call (800) 795-3272 (voice) or (202) 720-6382 (TDD). USDA is an equal opportunity provider 

and employer. 

Printed on recycled paper with soy-based ink. 
Gypsy Moth Management 

in the United States: 

a cooperative approach 

Draft 

Supplemental Environmental 

Impact Statement 

Risk Assessments and 

Risk Comparison 

Volume IV of IV 

Appendixes J-M 

United States 

Department of Agriculture 

Forest Service 

Animal and Plant Health 

Inspection Service 

Newtown Square, PA 

NA-MR-01-08 

June 2008 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Gypsy Moth Management in the United States: a cooperative approach 

Draft Supplemental Environmental Impact Statement 

The complete Draft Supplemental Environmental Impact Statement, Gypsy 

Moth Management in the United States: a cooperative approach, consists of 

four volumes: 

Volume I Summary 

Volume II Chapters 1-8 and Appendixes A, B, C, D, E 

Volume III Appendixes F, G, H, I 

Volume IV Appendixes J, K, L, M 

Lead Agency:	 Forest Service, U.S. Department of Agriculture 

(USDA) 

Responsible Official:		 James R. Hubbard, Deputy Chief for State and 

Private Forestry 

Sidney R. Yates Federal Building 

201 14th Street, SW 

Washington, DC 20250 

For More Information: Bill Oldland, SEIS Team Leader 

180 Canfield Street 

Morgantown, WV 26505 

Joint Lead Agency:	 Animal and Plant Health Inspection Service 

(APHIS), USDA 

Responsible Official:		 Richard L. Dunkle, Deputy Administrator for Plant 

Protection and Quarantine 

4700 River Road 

Riverdale, MD 20737-1236 

For More Information:	 Weyman P. Fussell, Ph.D., Gypsy Moth Program 

Coordinator 

Emergency and Domestic Programs, APHIS 

4700 River Road, Unit 134 

Riverdale, MD 20737-1236 
 

 

 

 

Abstract: The USDA Forest Service and Animal and Plant Health Inspection Service are proposing an addition 

to the gypsy moth management program that was described in the 1995 Environmental Impact Statement--Gypsy 

Moth Management in the United States: a cooperative approach--and chosen in the 1996 Record of Decision. 

The agencies are proposing these new treatment options: adding the insecticide tebufenozide, or adding the 

insecticide tebufenozide and other new treatment(s) that may become available in the future to manage gypsy 

moths, provided that the other treatment(s) poses no greater risk to human health and nontarget organisms than are 

disclosed in this Draft SEIS for the currently approved treatments and tebufenozide. 

Commenting on this Draft Supplemental Environmental Impact Statement: Reviewers should provide the 

Forest Service with their comments during the review period of this draft supplemental environmental impact 

statement. Timely comments will enable the Forest Service to analyze and respond to all of the comments at one 

time and to use information acquired in the preparation of the final supplemental environmental impact statement, 

thus avoiding undue delay in the decision making process. Furthermore, the more specific and substantive 

the comments, the better for reviewers and the agencies alike. Reviewers have an obligation to structure their 

participation in the National Environmental Policy Act process so that it is meaningful and alerts the agency to 

the reviewer’s position and contentions (Vermont Yankee Nuclear Power Corp. v. NRDC, 435 U.S. 519, 553, 

1978). Environmental objections that could have been raised at the draft stage may therefore be forfeited, if not 

raised until after completion of the final environmental impact statement (Department of Transportation v. Public 

Citizen, 541 U.S. 752, 764 (2004). Comments on this draft supplemental environmental impact statement should 

be specific and should address the adequacy of the statement and the merits of the alternatives discussed (40 CFR 

1503.3). 

Web Site for Draft SEIS: The Draft SEIS is available for viewing at www.na.fs.fed.us/wv/eis 

Send Comments to: Bill Oldland, SEIS Team Leader 

USDA Forest Service 

180 Canfield Street 

Morgantown, WV 26505
 

304-285-1585
 

Date Comments Must Be Received: See cover letter for the date that comments are due in Morgantown, WV. 
 

 

 

 

Contents 

Appendix J. Tebufenozide Risk Assessment 

Appendix K. DDVP (Dichlorvos) Risk Assessment 

Appendix L. Gypsy Moth Risk Assessment 

Appendix M. Risk Comparison 

Figures 

Figure J-1. DDT was applied using airplanes in the early years of gypsy moth control programs. 

Figure K-1. A sprayer unit mounted on a Model A Ford truck was used for gypsy moth control. 

Figure L-1. Gypsy moth caterpillars cluster at the base of a banded tree (Arlington, Virginia, 1905). 

Figure M-1. Ropes were used to climb trees, to treat them for gypsy moths in the 1930s. 

Photo Credits 

Figure J-1. (UGA1275035) USDA Forest Service Archives, www.forestryimages.org 

Figure K-1. (UGA1275022) USDA Forest Service Archives, www.forestryimages.org 

Figure L-1. (UGA1275074) USDA Forest Service Archives, www.forestryimages.org 

Figure M-1. (UGA1275057) USDA Forest Service Archives, www.forestryimages.org 
Appendix J 

Tebufenozide 

Risk Assessment 

Figure J-1. DDT was applied using airplanes in the early years of gypsy moth 

control programs. 

SERA TR 04-43-05-06c 

Control/Eradication Agents for the
 

Gypsy Moth ­

Human Health and Ecological Risk Assessment
 

for Tebufenozide (Mimic)
 

Final Report
 

Prepared for: 

USDA, Forest Service
 

Forest Health Protection 

GSA Contract No. GS-10F-0082F
 

USDA Forest Service BPA: WO-01-3187-0150
 

Requisition No.: 43-3187-1-0269
 

Task No. 5
 

Submitted to:
 

Dave Thomas, COTR
 

Forest Health Protection Staff
 

USDA Forest Service
 

Rosslyn Plaza Building C, Room 7129C
 

1601 North Kent Street
 

Arlington, VA 22209
 

Prepared by:
 

Patrick R. Durkin and Julie Klotzbach
 

Syracuse Environmental Research Associates, Inc. 

5100 Highbridge St., 42C
 

Fayetteville, New York 13066-0950
 

Telephone: (315) 637-9560
 

Fax: (315) 637-0445
 

E-Mail: SERA_INC@msn.com
 

Home Page: www.sera-inc.com
 

August 8, 2004 
TABLE OF CONTENTS
 

TABLE OF CONTENTS ....................................................... ii
 

LIST OF APPENDICES ...................................................... v i 
 

LIST OF TABLES ........................................................... v i 
 

LIST OF FIGURES .......................................................... v i 
 

LIST OF WORKSHEETS ..................................................... v i 
 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . vii
 

COMMON UNIT CONVERSIONS AND ABBREVIATIONS . . . . . . . . . . . . . . . . . . . . . . . . ix
 

CONVERSION OF SCIENTIFIC NOTATION .................................... x
 

EXECUTIVE SUMMARY .................................................... xi
 

1. INTRODUCTION ......................................................... 1-1
 

2. PROGRAM DESCRIPTION ................................................ 2-1
 

2.1. OVERVIEW ..................................................... 2-1
 

2.2. CHEMICAL DESCRIPTION AND COMMERCIAL FORMULATIONS . . . . . 2-1
 

2.3. APPLICATION METHODS ......................................... 2-2
 

2.4. MIXING AND APPLICATION RATES ................................ 2-3
 

2.5. USE STATISTICS ................................................. 2-4
 

3. HUMAN HEALTH RISK ASSESSMENT ..................................... 3-1
 

3.1. HAZARD IDENTIFICATION ....................................... 3-1
 

3.1.1. Overview. ................................................ 3-1
 

3.1.2. Mechanism of Action........................................ 3-1
 

3.1.3. Kinetics and Metabolism. .................................... 3-1
 

3.1.3.1. Pharmacokinetic Studies .............. ..... 3-2
 

3.1.3 2. Dermal Absorption Rates .......................... 3-2
 

3.1.4. Acute Oral Toxicity. ........................................ 3-3
 

3.1.5. Subchronic or Chronic Systemic Toxic Effects. . . . . . . . . . . . . . . . . . . . 3-3
 

3.1.6. Effects on Nervous System. .................................. 3-4
 

3.1.7. Effects on Immune System. .................................. 3-4
 

ii 
 

TABLE OF CONTENTS (continued) 

3.1.8. Effects on Endocrine System. ................................. 3-5
 

3.1.9. Reproductive and Teratogenic Effects. . . . . . . . . . . . . . . . . . . . . . . . . . . 3-6
 

3.1.10. Carcinogenicity and Mutagenicity. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-6
 

3.1.11. Irritation and Sensitization (Effects on the Skin and Eyes). . . . . . . . . . 3-7
 

3.1.12. Systemic Toxic Effects from Dermal Exposure. . . . . . . . . . . . . . . . . . 3-7
 

3.1.13. Inhalation Exposure. ..................................... .. 3-8
 

3.1.14. Inerts and Adjuvants. ..................................... . 3-8
 

3.1.15. Impurities and Metabolites................................. . . 3-9
 

3.1.16. Toxicologic Interactions. .................................. . 3-9
 

3.2. EXPOSURE ASSESSMENT ....................................... 3-10
 

3.2.1. Overview. ............................................... 3-10
 

3.2.2. Workers. ................................................ 3-10
 

3.2.2.1. General Exposures ................................. 3-11
 

3.2.2.2. Accidental Exposures ............................... 3-12
 

3.2.3. General Public. ........................................... 3-14
 

3.2.3.1. General Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-14
 

3.2.3.2. Direct Spray ...................................... 3-14
 

3.2.3.3. Dermal Exposure from Contaminated Vegetation . . . . . . . . . 3-14
 

3.2.3.4. Contaminated Water ............................... 3-15
 

3.2.3.4.1. Accidental Spill ........ . ............ 3-15
 

3.2.3.4.2. Accidental Direct Spray . . . . . . . . . . . . . . . . . . . . . 3-15
 

3.2.3.5. Oral Exposure from Contaminated Fish . . . . . . . . . . . . . . . . 3-19
 

3.2.3.6. Oral Exposure from Contaminated Vegetation . . . . . . . . . . . 3-20
 

3.3. DOSE-RESPONSE ASSESSMENT .................................. 3-23
 

3.3.1. Overview ................................................ 3-23
 

3.3.2. Chronic RfD ............................................. 3-23
 

3.3.3. Acute RfD ............................................... 3-24
 

3.4. RISK CHARACTERIZATION ...................................... 3-26
 

3.4.1. Overview ................................................ 3-26
 

3.4.2. Workers ................................................. 3-26
 

3.4.3. General Public. ........................................... 3-27
 

3.4.4. Sensitive Subgroups........................................ 3-28
 

3.4.5. Connected Actions......................................... 3-28
 

3.4.6. Cumulative Effects......................................... 3-28
 

iii 
TABLE OF CONTENTS (continued) 

4. ECOLOGICAL RISK ASSESSMENT ......................................... 4-1
 

4.1. HAZARD IDENTIFICATION ....................................... 4-1
 

4.1.1. Overview. ................................................ 4-1
 

4.1.2. Toxicity to Terrestrial Organisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-2
 

4.1.2.1. Mammals ......................................... 4-2
 

4.1.2.2. Birds ............................................. 4-2
 

4.1.2.3. Terrestrial Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-4
 

4.1.2.4. Terrestrial Plants (Macrophytes) . . . . . . . . . . . . . . . . . . . . . . . 4-5
 

4.1.2.5. Terrestrial Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-6
 

4.1.3. Aquatic Organisms.......................................... 4-6
 

4.1.3.1. Fish .............................................. 4-6
 

4.1.3.2. Amphibians ....................................... 4-6
 

4.1.3.3. Aquatic Invertebrates ................................ 4-6
 

4.1.3.4. Aquatic Plants ..................................... 4-8
 

4.1.3.5. Microorganisms (other than algae) . . . . . . . . . . . . . . . . . . . . . 4-8
 

4.2. EXPOSURE ASSESSMENT ........................................ 4-9
 

4.2.1. Overview. ................................................ 4-9
 

4.2.2. Terrestrial Animals. ....................................... . 4-9
 

4.2.2.1. Direct Spray ...................................... 4-11
 

4.2.2.2. Indirect Contact ................................... 4-12
 

4.2.2.3. Ingestion of Contaminated Vegetation or Prey . . . . . . . . . . . 4-12
 

4.2.2.4. Ingestion of Contaminated Water . . . . . . . . . . . . . . . . . . . . . 4-14
 

4.2.3. Terrestrial Plants. ......................................... 4-14
 

4.2.4. Soil Organisms............................................ 4-15
 

4.2.5. Aquatic Organisms......................................... 4-15
 

4.3. DOSE-RESPONSE ASSESSMENT .................................. 4-16
 

4.3.1. Overview. ............................................... 4-16
 

4.3.2. Toxicity to Terrestrial Organisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-17
 

4.3.2.1. Mammals ........................................ 4-17
 

4.3.2.2. Birds ............................................ 4-17
 

4.3.2.3. Terrestrial Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-19
 

4.3.2.4. Terrestrial Plants and Microorganisms . . . . . . . . . . . . . . . . . 4-20
 

iv 
 

 

TABLE OF CONTENTS (continued) 

4.3.3. Aquatic Organisms......................................... 4-20
 

4.3.3.1. Fish ............................................. 4-20
 

4.3.3.2. Aquatic Invetebrates .................. . .. . 4-21
 

4.3.3.3. Aquatic Plants . . ............................... 4-22
 

4.3.3.4. Aquatic Microorganisms .................. .. .. 4-22
 

4.4. RISK CHARACTERIZATION ...................................... 4-23
 

4.4.1. Overview. ............................................... 4-23
 

4.4.2. Terrestrial Organisms....................................... 4-23
 

4.4.2.1. Terrestrial Vertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-23
 

4.4.2.2. Terrestrial Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-24
 

4.4.2.3. Terrestrial Plants and Microorganism s . ...... 4-25
 

4.4.3. Aquatic Organisms......................................... 4-25
 

4.4.3.1. Aquatic Vertebrates ................................ 4-25
 

4.4. 3 . 2. Aquatic Invertebrates .............................. 4-26
 

4.4.3 . 3. Aquatic Plants ................................... 4-26
 

5. REFERENCES ........................................................... 5-1
 

v 
 

TABLE OF CONTENTS (continued) 

LIST OF APPENDICES 

Appendix 1: Estimates of dermal absorption rates for tebufenozide 

Appendix 2: Oral toxicity of tebufenozide to experimental mammals 

Appendix 3: Dermal, inhalation, and ocular effects of tebufenozide in experimental mammals 

Appendix 4: Toxicity of tebufenozide to birds after oral administration 

Appendix 5: Toxicity to non-target terrestrial invertebrates 

Appendix 6: Terrestrial field studies 

Appendix 7: Toxicity of tebufenozide to fish 

Appendix 8: Toxicity of tebufenozide to aquatic invertebrates and algae 

LIST OF TABLES 

Table 2-1: Selected physical and chemical properties of tebufenozide with selected additional 

properties for the commercial formulation Mimic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .Tables-1 

Table 3-1: Chemical and site parameters used in GLEAMS modeling for tebufenozide .. . . . . Tables-2 

Table 3-2: Summary of modeled concentrations of tebufenozide in streams . . . . . . . . . . . . . . .Tables-3 

Table 3-3: Summary of modeled concentrations of tebufenozide in ponds . . . . . . . . . . . . . . . . . Tables-4 

Table 3-4: Estimated environmental concentrations of tebufenozide in surface and groundwater 

at two applications of 0.12 lb a.i./acre, three days apart . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-5 

Table 3-5: Concentrations of tebufenozide in surface water used in risk assessment . . . . . . . . . Tables-6 

Table 4-1: Summary of modeled concentrations of tebufenozide in soil, two applications spaced 

three days apart. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .. . . Tables-7 

Table 4-2: Summary of tebufenozide toxicity values used in ecological risk assessment . . . . . Tables-8 

Table 4-3: Summary of field studies on the effects of tebufenozide of terrestrial invertebrates.. Tables-9 

LIST OF FIGURES 

Figure 2-1: Agricultural use of tebufenozide in the United States for 1992 . . . . . . . . . . . . . . . Figures-1 

NOTE: Tables followed by figures are placed after Section 5, References.
 

LIST OF WORKSHEETS
 

Supplement 1:	 Tebufenozide -EXCEL Worksheets for Human Health and Ecological 

Risk Assessments, SERA EXWS 04-43-05-06c, Version 3.01. 

Located at: http://www.fs.fed.us/foresthealth/pesticide/risk.shtml

 vi 
AEL 

a.i. 

BCF 

bw 

CBI 

CI 

cm 

CNS 

DAA 

DAT 

d.f. 

ECx

EC25 

EC50 

ExToxNet 

F 

FH 

FIFRA 

FOIA 

FQPA 

g 

ha 

HQ 

IAA 

IARC 

IRIS 

ka

ke

kg 

Ko/c 

Ko/w 

Kp

L

lb 

LC50 

LD50 

LOAEL 

LOC 

m 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

adverse-effect level 

active ingredient 

bioconcentration factor 

body weight 

confidential business information 

confidence interval 

centimeter 

central nervous system 

days after application 

days after treatment 

degrees of freedom 

concentration causing X% inhibition of a process 

concentration causing 25% inhibition of a process 

concentration causing 50% inhibition of a process 

Extension Toxicology Network 

female 

Forest Health 

Federal Insecticide, Fungicide and Rodenticide Act 

Freedom of Information Act 

Food Quality Protection Act 

gram 

hectare 

hazard quotient 

indole-3-acetic acid 

International Agency for Research on Cancer 

Integrated Risk Information System 

absorption coefficient 

elimination coefficient 

kilogram 

organic carbon partition coefficient 

octanol-water partition coefficient 

skin permeability coefficient 

liter 

pound 

lethal concentration, 50% kill 

lethal dose, 50% kill 

lowest-observed-adverse-effect level 

level of concern 

meter 

vii 
ACRONYMS, ABBREVIATIONS, AND SYMBOLS (continued) 

M male 

mg milligram 

mg/kg/day milligrams of agent per kilogram of body weight per day 

mL milliliter 

mM millimole 

MRID Master Record Identification Number 

MSDS material safety data sheet 

MW molecular weight 

NCAP Northwest Coalition for Alternatives to Pesticides 

NCI National Cancer Institute 

NIOSH National Institute for Occupational Safety and Health 

NOAEL no-observed-adverse-effect level 

NOEC no-observed-effect concentration 

NOEL no-observed-effect level 

NOS not otherwise specified 

NRC National Research Council 

NTP National Toxicology Program 

OM organic matter 

OPP Office of Pesticide Programs 

OPPTS Office of Pesticide Planning and Toxic Substances 

OSHA Occupational Safety and Health Administration 

ppm parts per million 

RBC red blood cells 

RED re-registration eligibility decision 

RfD reference dose 

SERA Syracuse Environmental Research Associates 

SGOT serum glutamic oxaloacetic transaminase 

SGPT serum glutamic pyruvic transaminase 

SRC Syracuse Research Corporation 

UF uncertainty factor 

U.S. United States 

USDA U.S. Department of Agriculture 

U.S. EPA U.S. Environmental Protection Agency 

USGS U.S. Geological Survey 

WCR water contamination rate 

WHO World Health Organization 

: micron 

viii 
COMMON UNIT CONVERSIONS AND ABBREVIATIONS
 

To convert ... Into ... Multiply by ... 

acres hectares (ha) 0.4047 

acres square meters (m ) 2 4,047 

atmospheres millimeters of mercury 760 

centigrade Fahrenheit 1.8 °C+32 

centimeters inches 0.3937 

cubic meters (m ) 3 liters (L) 1,000 

Fahrenheit centigrade 0.556 °F-17.8 

feet per second (ft/sec) miles/hour (mi/hr) 0.6818 

gallons (gal) liters (L) 3.785 

gallons per acre (gal/acre) liters per hectare (L/ha) 9.34 

grams (g) ounces, (oz) 0.03527 

grams (g) pounds, (oz) 0.002205 

hectares (ha) acres 2.471 

inches (in) centimeters (cm) 2.540 

kilograms (kg) ounces, (oz) 35.274 

kilograms (kg) pounds, (lb) 2.2046 

kilograms per hectare (hg/ha) pounds per acre (lb/acre) 0.892 

kilometers (km) miles (mi) 0.6214 

liters (L) cubic centimeters (cm ) 3 1,000 

liters (L) gallons (gal) 0.2642 

liters (L) ounces, fluid (oz) 33.814 

miles (mi) kilometers (km) 1.609 

miles per hour (mi/hr) cm/sec 44.70 

milligrams (mg) ounces (oz) 0.000035 

meters (m) feet 3.281 

ounces (oz) grams (g) 28.3495 

ounces per acre (oz/acre) grams per hectare (g/ha) 70.1 

ounces per acre (oz/acre) kilograms per hectare (kg/ha) 0.0701 

ounces fluid cubic centimeters (cm ) 3 29.5735 

pounds (lb) grams (g) 453.6 

pounds (lb) kilograms (kg) 0.4536 

pounds per acre (lb/acre) kilograms per hectare (kg/ha) 1.121 

pounds per acre (lb/acre) mg/square meter (mg/m ) 2 112.1 

pounds per acre (lb/acre) :g/square centimeter (:g/cm ) 2 11.21 

pounds per gallon (lb/gal) grams per liter (g/L) 119.8 

square centimeters (cm ) 2 square inches (in ) 2 0.155 

square centimeters (cm ) 2 square meters (m ) 2 0.0001 

square meters (m ) 2 square centimeters (cm ) 2 10,000 

yards meters 0.9144 

Note: All references to pounds and ounces refer to avoirdupois weights unless otherwise specified. 

ix 
CONVERSION OF SCIENTIFIC NOTATION
 

Scientific Decimal Verbal 

Notation Equivalent Expression 

1 @ 10-10 0.0000000001 One in ten billion 

1 @ 10-9 0.000000001 One in one billion 

1 @ 10-8 0.00000001 One in one hundred million 

1 @ 10-7 0.0000001 One in ten million 

1 @ 10-6 0.000001 One in one million 

1 @ 10-5 0.00001 One in one hundred thousand 

1 @ 10-4 0.0001 One in ten thousand 

1 @ 10-3 0.001 One in one thousand 

1 @ 10-2 0.01 One in one hundred 

1 @ 10-1 0.1 One in ten 

1 @ 100 1 One 

1 @ 101 10 Ten 

1 @ 102 100 One hundred 

1 @ 103 1,000 One thousand 

1 @ 104 10,000 Ten thousand 

1 @ 105 100,000 One hundred thousand 

1 @ 106 1,000,000 One million 

1 @ 107 10,000,000 Ten million 

1 @ 108 100,000,000 One hundred million 

1 @ 109 1,000,000,000 One billion 

1 @ 1010 10,000,000,000 Ten billion 

x 
EXECUTIVE SUMMARY
 

OVERVIEW
 

The use of tebufenozide to control the gypsy moth may result in adverse effects in nontarget 

Lepidoptera but there is little indication that humans or non-lepidopteran wildlife species will be 

impacted under normal conditions of use even at the highest application rate. 

The only hazard quotient for humans that exceeds the level of concern (HQ of 1.5) involves the 

longer term consumption of contaminated vegetation. While the longer-term consumption of 

contaminated vegetation is probably not a likely scenario, it is a standard exposure scenario used 

in Forest Service risk assessments to consider the longer term consumption of food items such as 

berries that might be sprayed during the broadcast application of a pesticide. This risk 

assessment suggests that two applications at 0.08 lb/acre or more should be avoided in areas 

where members of the general public might consume contaminated fruits or other contaminated 

vegetation. 

Tebufenozide is an insecticide that is effective in controlling populations of lepidopteran pests. 

No data, however, are available on toxicity to nontarget Lepidoptera. For this risk assessment, 

the assumption is made that nontarget Lepidoptera may be as sensitive to tebufenozide as target 

Lepidoptera. Thus, adverse effects in nontarget Lepidoptera would be expected after 

applications that are effective for the control of lepidopteran pest species. 

There is no indication that short term exposures to tebufenozide will cause adverse effects in any 

terrestrial vertebrates or non-lepidopteran invertebrates even at the upper range of plausible 

exposures as well as accidental exposures. Similarly, adverse effects from longer terms 

exposures in birds and mammals appears to be unlikely under most conditions. In some extreme 

cases, exposures in some large mammals could exceed the NOEC but the but exposures would 

be below levels that have been associated with frank signs of toxicity. Adverse effects in aquatic 

species are not expected under normal conditions of use. In the case of a large accidental spill 

into a relatively small body of water, however, adverse effects could be expected in aquatic 

vertebrates, invertebrates, and plants. 

PROGRAM DESCRIPTION 

Mimic is a commercial formulation of tebufenozide, a synthetic chemical that acts like an 

invertebrate hormone that controls molting in insects and various terrestrial and aquatic 

invertebrates. While Mimic is specifically used by the USDA for the control of the Gypsy moth, 

tebufenozide is also used in the control of other lepidopteran pest species. Mimic is comprised 

of 23-25% tebufenozide and 75-77% inert ingredients. The inert ingredients consist of glycerol, 

related reaction products, alkylaryl polyether alcohol, glyceridic and canola oils, and water. 

Tebufenozide is relatively persistent in the environment and may be subject to bioconcentration. 

Although the compound is not highly mobile in soil, it may be transported by percolation or 

runoff from soil to ambient water. Potential concentrations of tebufenozide in ambient water 

depend largely on site specific conditions. 

xi 
Ground and aerial applications of Mimic are permitted and both methods may be considered in 

USDA programs. The labeled application rates for tebufenozide range from 2 to 8 ounces of 

Mimic/acre, corresponding to tebufenozide application rates of 0.03 lbs/acre to 0.12 lbs/acre. 

Multiple applications of tebufenozide are permitted but the maximum annual application rate is 

16 fl ounces/acre or 0.24 lb a.i./acre. The application rates for Mimic may vary among USDA 

programs – i.e., suppression, eradication, and Slow-the-Spread. For the current risk assessment, 

a range of application rates – i.e., 0.015 lb a.i./acre to 0.12 lb a.i./acre – are considered. All 

exposure assessments are conducted at the maximum application rate of 0.12 lb/acre, assuming 

two applications with a 3 day interval. This is essentially a worse-case scenario using a shortest 

interval between applications and two applications that reach the maximum annual application 

rate of 0.24 lb/acre. 

HUMAN HEALTH RISK ASSESSMENT 

Hazard Identification – A relatively detailed and consistent series of studies in mice, rats, and 

dogs indicates that the primary mechanism of tebufenozide toxicity in mammals involves effects 

on the blood. Tebufenozide does not appear to be carcinogenic and does not appear to cause 

birth defects. Nonetheless, the compound is associated with adverse reproductive effects in 

experimental mammals. Tebufenozide itself does not seem to be irritating to the skin or eyes. 

Mimic, however, appears to contain other constituents (inerts or adjuvants) that may cause skin 

or eye irritation. 

As discussed in the exposure assessment, dermal absorption is the primary route of exposure for 

workers. Data regarding the dermal absorption kinetics of tebufenozide are not available in the 

published or unpublished literature. For this risk assessment, estimates of dermal absorption 

rates are based on quantitative structure-activity relationships. Although the lack of experimental 

data regarding dermal absorption of tebufenozide adds uncertainties to this risk assessment, the 

available data regarding the oral and dermal toxicity of tebufenozide are sufficient to suggest that 

the estimated dermal absorption rates are plausible. 

The inhalation toxicity of tebufenozide is not well documented in the literature. The available 

studies indicate that tebufenozide induces irritant effects at very high exposure levels. Because 

inhalation exposure involving high concentrations of tebufenozide is implausible, the potential 

inhalation toxicity of the compound is not of substantial concern to this risk assessment. 

Exposure Assessment – A standard set of exposure scenarios are presented for both workers and 

members of the general public. All exposure assessments are conducted at the maximum 

application rate for tebufenozide of 0.12 lb/acre using two applications with an application 

interval of three days. This cumulative application (0.24 lb a.i./acre) is the maximum application 

rate for a single season. This leads to the highest estimates of peak as well as longer term 

exposures. 

For workers applying tebufenozide, three types of application methods are modeled: directed 

ground spray, broadcast ground spray, and aerial spray. Central estimates of exposure for 

xii 
workers are approximately 0.002 mg/kg/day for aerial and backpack workers and about 0.003 

mg/kg/day for broadcast ground spray workers. Upper ranges of exposures are approximately 

0.02 mg/kg/day for broadcast ground spray workers and 0.01 mg/kg/day for backpack and aerial 

workers. All of the accidental exposure scenarios for workers involve dermal exposures and 

most of these accidental exposures lead to estimates of dose that are either in the range of or 

substantially below the general exposure estimates for workers. The one exception involves 

wearing contaminated gloves for one-hour. The upper range of exposure for this scenario is 

about 4 mg/kg/day. 

For the general public, the range of acute exposures is from approximately 0.0000002 mg/kg 

associated with the lower range for the consumption of contaminated water from a stream by a 

child to 1.2 mg/kg associated with the upper range for consumption of contaminated water by a 

child after an accidental spill. Relatively high dose estimates are also associated with the direct 

spray of a child (about 0.4 mg/kg at the upper range of exposure) and for the consumption of fish 

after an accidental spill by members of the general public (0.2 mg/kg) and subsistence 

populations (0.9 mg/kg). Other acute exposure scenarios are associated with doses that are lower 

by at least an order of magnitude. For chronic or longer term exposures, the modeled exposures 

are much lower than for acute exposures, ranging from approximately 0.000000002 mg/kg/day (2 

in 1 billionth of a mg/kg/day) associated with the lower range for the consumption of 

contaminated water to approximately 0.03 mg/kg/day associated with the upper range for 

consumption of contaminated fruit. 

Dose-Response Assessment – Acute and chronic risk values are derived for tebufenozide. 

Following standard practices for USDA risk assessments, risk assessment values available from 

U.S. EPA are adopted directly unless there is a compelling basis for doing otherwise. When risk 

values are not available from U.S. EPA, the methods used by U.S. EPA are employed to derive 

surrogate values. 

U.S. EPA has derived a chronic RfD for tebufenozide of 0.018 mg/kg/day. This chronic RfD is 

well-documented and is used directly for all longer term exposures to tebufenozide. This value is 

based on a NOAEL of 1.8 mg/kg/day in dogs and an uncertainty factor of 100 – two factors of 10 

for interspecies and intraspecies variability. Because of the low acute toxicity of tebufenozide, 

the U.S. EPA has not derived an acute RfD but has identified an acute NOAEL of 1000 

mg/kg/day from reproduction studies in both rats and rabbits involving 10 to 13 day exposure 

periods. This NOAEL is the basis for a surrogate acute RfD of 10 mg/kg using an uncertainty 

factor of 100 as in the chronic RfD. This surrogate acute RfD is applied to all incidental or 

accidental exposures that involve an exposure period of 1 day. 

Risk Characterization – At the maximum application rate considered in this risk assessment, 

two applications at 0.12 lb/acre spaced three day apart, there is little indication that adverse 

effects on human health are likely. Based on central estimates of exposure – those that might be 

considered typical and expected – hazard quotients including workers and members of the 

general public range from 0.00003 to 0.03, below a level of concern by factors of about 30 to 

xiii 
33,000. At the upper range of plausible exposures, the hazard quotient for ground spray workers 

reaches a level of concern – i.e., a hazard quotient of 1. For members of the general public, the 

upper range of exposure leads to a hazard quotient of 1.5 for the longer-term consumption of 

contaminated vegetation for two applications at 0.12 lb/acre. Because of the linear relationship 

between exposure and application rate, two applications at 0.08 lb/acre would reach but not 

exceed a level of concern. With a single application at the maximum rate of 0.12 lb/acre, the 

hazard index is 0.8, below the level of concern. While the longer-term consumption of 

contaminated vegetation is probably not a likely scenario, it is a standard exposure scenario used 

in Forest Service risk assessments to consider the longer term consumption of food items such as 

berries that might be sprayed during the broadcast application of a pesticide. This risk 

assessment suggests that two applications at 0.08 lb/acre or more should be avoided in areas 

where members of the general public might consume contaminated fruits or other contaminated 

vegetation. 

ECOLOGICAL RISK ASSESSMENT 

Hazard Identification – The toxicity of tebufenozide is well characterized in experimental 

mammals, birds, terrestrial invertebrates, and aquatic animals. Nonetheless, given the very large 

number of species in the environment which could be exposed to tebufenozide, toxicity data are 

available on relatively few species. 

The most sensitive effects in wildlife mammalian species will probably be the same as those in 

experimental mammals (i.e., effects on the blood). At higher doses, tebufenozide was associated 

with impaired reproductive performance in experimental mammals, and this effect is also 

considered quantitatively in this risk assessment. Potential reproductive effects are also of 

concern for birds, although there are inconsistencies in the available experimental data. The 

available literature includes a reproduction study investigating effects in mallard ducks and two 

reproduction studies investigating effects in bobwhite quail. In one of the quail studies, dietary 

concentrations of 300 and 1000 ppm caused reproductive effects. These effects were not 

observed in that study at 100 ppm or in the more recent quail study or in the study on mallard 

ducks. A field study on the effects of tebufenozide on reproductive performance in birds noted 

trends that were statistically insignificant but suggestive of adverse reproductive effects in a 

warbler species. Thus, consistent with the interpretation by the U.S. EPA, reproductive effects in 

both mammals and birds are considered endpoints of concern in this risk assessment. 

The mechanism of action of tebufenozide in target insects is relatively well understood. 

Tebufenozide mimics the action of the invertebrate hormone, 20-hydroxyecdysone, which 

controls molting. The effectiveness of tebufenozide in mimicking 20-hydroxyecdysone activity, 

however, appears to vary markedly among orders and species of invertebrates. In general, moths 

are sensitive to tebufenozide but other insects are much less sensitive. 

There are no bioassays regarding the toxicity of tebufenozide to terrestrial plants or terrestrial 

microorganisms in the literature. There are a number of field studies and field simulation studies 

available on tebufenozide and effects that might be associated with toxicity to plants or soil 

xiv 
microorganisms have not been noted. 

The acute toxicity of tebufenozide to aquatic animals is relatively low, with acute LC50 values 

ranging from 2.2 to 6.5 mg/L for fish and 0.3 to 3.8 mg/L for aquatic invertebrates. Nonetheless, 

much lower concentrations of tebufenozide may cause reproductive effects in fish (0.048 mg/L) 

and aquatic invertebrates (0.0053 mg/L). 

Exposure Assessment – As in the human health risk assessment, most exposure assessments 

used in the ecological risk assessment are based on two applications spaced 3 days apart at an 

application rate of 0.12 lb/acre. Two sets of exposure assessments are given for scenarios 

involving the longer-term consumption of contaminated vegetation: one for a single application 

at 0.12 lb/acre and another for two applications spaced 3 days apart at an application rate of 0.12 

lb/acre. 

Terrestrial animals might be exposed to any applied pesticide from direct spray, the ingestion of 

contaminated media (vegetation, prey species, or water), grooming activities, or indirect contact 

with contaminated vegetation. For tebufenozide, the highest acute exposure for a terrestrial 

vertebrate is associated with a fish-eating bird and could reach up to about 85 mg/kg. Exposures 

anticipated from the consumption of contaminated vegetation by terrestrial animals range from 

central estimates of about 0.15 mg/kg for a small mammal consuming fruit to about 3 mg/kg for 

a large bird with upper ranges of about 0.4 mg/kg for a small mammal and 9 mg/kg for a large 

bird. The consumption of contaminated water leads to much lower levels of exposure. A similar 

pattern is seen for chronic exposures. Estimated longer-term daily doses for the a small mammal 

from the consumption of contaminated vegetation at the application site are in the range of about 

0.000002 mg/kg/day to 0.08 mg/kg/day. Large birds feeding on contaminated vegetation at the 

application site could be exposed to much higher concentrations, ranging from about 0.015 

mg/kg/day to 11 mg/kg/day. The upper ranges of exposure from contaminated vegetation far 

exceed doses that are anticipated from the consumption of contaminated water, which range from 

about 0.0000003 mg/kg/day to 0.0002 mg/kg/day for a small mammal. 

Exposure to aquatic organisms is based on essentially the same information used to assess the 

exposure to terrestrial species from contaminated water. The peak estimated concentration of 

tebufenozide in ambient water is 10 (0.005 to 40) μg/L after two applications of 0.12 lb/acre 

spaced three days apart. For longer-term exposures, the corresponding longer term 

concentrations in ambient water are estimated at about 0.004 (0.00002 to 0.01) μg/L. 

Dose-Response Assessment – The available toxicity data support separate dose-response 

assessments in six classes of organisms: terrestrial mammals, birds, nontarget terrestrial 

invertebrates, fish, aquatic invertebrates, and aquatic algae. Different units of exposure are used 

for different groups of organisms depending on how exposures are likely to occur and how the 

available toxicity data are expressed. 

Tebufenozide is relatively non-toxic to mammals and birds. For mammals, the toxicity values 

xv 
used in the ecological risk assessment are identical to those used in the human health risk 

assessments: an acute NOAEL for reproductive toxicity of 1000 mg/kg and a chronic NOAEL of 

1.8 mg/kg/day based on effects on the blood. For birds, the acute NOAEL for tebufenozide is 

taken as 2150 mg/kg from an acute oral study in which the dose was administered in capsules for 

21-days. The longer term NOAEL is taken as 15 mg/kg/day from a standard reproduction study 

in bobwhite quail. 

For terrestrial invertebrates, three types of data are used to characterize risks: a contact bioassay 

in the honey bee, a soil bioassay in earthworms, and field studies in which population level 

effects were monitored in insects. The standard contact bioassay in honey bees indicates an 

NOEC of 2500 mg/kg bw, comparable to the acute toxicity values in mammals and birds. The 

earthworm bioassay indicates a NOEC of 1000 mg/kg soil. The available field studies indicate 

that tolerant insect species are not affected by application rates up to 0.24 lb/acre. The true 

NOEC may be higher – i.e., an LOEC has not been identified for tolerant species of terrestrial 

insects. Conversely, application rates as low as 0.03 lb/acre have been shown to adversely affect 

sensitive non-target insects, primarily Lepidoptera. A NOEC for sensitive species has not been 

identified. 

Acute toxicity values for aquatic species indicate relatively little difference between fish and 

aquatic invertebrates. For fish, the acute NOEC values are 0.39 mg/L and 1.9 mg/L for sensitive 

and tolerant species, respectively. For invertebrates, the corresponding acute NOEC values are 

0.12 mg/L and 0.82 mg/L. Differences between fish and invertebrates are difficult to assess in 

terms of longer-term toxicity. For fish, data are available on only a single species, the fathead 

minnow, and only a LOAEL of 0.048 mg/L is available. For invertebrates, longer-term NOEC 

values of 0.0035 mg/L and 0.029 mg/L are used for sensitive and tolerant species. Toxicity 

values for aquatic plants are taken as 0.077 mg/L for sensitive species and 0.64 mg/L for tolerant 

species, somewhat below the acute NOEC values in fish and aquatic invertebrates. Because of 

the short life-cycle of individual algal cells, the relatively short-term bioassays in algae (i.e., 96 to 

120 hours) are applied to both acute and longer-term concentrations for the characterization of 

risk. 

Risk Characterization – The use of tebufenozide to control the gypsy moth may result in adverse 

effects in nontarget Lepidoptera but there is little indication that other species will be impacted 

under normal conditions of use even at the highest application rate. Tebufenozide is an 

insecticide that is effective in controlling populations of lepidopteran pests. No data, however, 

are available on toxicity to nontarget Lepidoptera. For this risk assessment, the assumption is 

made that nontarget Lepidoptera may be as sensitive to tebufenozide as target Lepidoptera. 

Thus, adverse effects in nontarget Lepidoptera would be expected after applications that are 

effective for the control of lepidopteran pest species. 

There is no indication that short term exposures to tebufenozide will cause direct adverse effects 

in any terrestrial vertebrates or non-lepidopteran invertebrates even at the upper range of 

plausible exposures as well as accidental exposures. Similarly, direct adverse effects from longer 

xvi 
term exposures in birds and mammals appear to be unlikely under most conditions. Effects on 

birds due to a decrease in available prey – i.e., terrestrial invertebrates – may be plausible. In 

some extreme cases, exposures in some large mammals could exceed the NOEC but the 

exposures would be below levels that have been associated with frank signs of toxicity. Adverse 

effects in aquatic species are not expected under normal conditions of use. In the case of a large 

accidental spill into a relatively small body of water, however, adverse effects could be expected 

in aquatic vertebrates, invertebrates, and plants. 

xvii 
1. INTRODUCTION
 

The USDA uses Mimic, a commercial formulation of tebufenozide, to control infestations of the 

Gypsy Moth. This risk assessment is an update to a risk assessment prepared for the USDA 

Forest Service in 2000 (SERA 2000) and is intended to support an assessment of the 

environmental consequences of using Mimic in USDA programs for the control of the gypsy 

moth. 

For the most part, the risk assessment methods used in this document are similar to those used in 

risk assessments previously conducted for the Forest Service as well as risk assessments 

conducted by other government agencies. Four chapters, including the introduction, program 

description, risk assessment for human health effects, and risk assessment for ecological effects 

or effects on wildlife species comprise the main body of this document. Each of the two risk 

assessment chapters has four major sections, including an identification of the hazards associated 

with Mimic, an assessment of potential exposure to the product, an assessment of the doseresponse relationships, and a characterization of the risks associated with plausible levels of 

exposure. These sections incorporate the basic steps recommended by the National Research 

Council of the National Academy of Sciences (NRC 1983) for conducting and organizing risk 

assessments. 

This is a technical support document, and it addresses some specialized technical areas. 

Nevertheless, an effort was made to ensure that the document can be understood by individuals 

who do not have specialized training in the chemical and biological sciences. Certain technical 

concepts, methods, and terms common to all parts of the risk assessment are described in plain 

language in a separate document (SERA 2001). The general technical terms used in this 

document are defined in an environmental glossary available at www.sera-inc.com. Some of the 

more complicated terms and concepts are defined, as necessary, in the text. 

There are no detailed reviews regarding the toxicity of tebufenozide or Mimic in the published 

literature. Risk assessments for human health and ecological effects were conducted by the U.S. 

EPA (1999a,b,c,d,e). The registrant for Mimic at that time, Rohm and Haas, also prepared a 

series of risk assessments and other evaluations on Mimic (Hawkins 1998; Hazelton and Quinn 

1994; Kaminski 1997; Keller 1994, 1996a, 1998; Keller and Brown 1998a,b; Quinn and 

Hazelton 1997). These unpublished documents were obtained and reviewed in the preparation of 

this Forest Service risk assessment. 

Because of the preponderance of unpublished relevant data in U.S. EPA files, a complete search 

of the U.S. EPA files was conducted in the preparation of this risk assessment. Full text copies 

of the most relevant studies [n=107] were kindly provided by the U.S. EPA Office of Pesticide 

Programs. The studies were reviewed, and synopses of the most relevant studies are included in 

the appendices to this document. 

The information presented in the appendices and the discussions in chapters 2, 3, and 4 of the 

1-1
 
risk assessment are intended to be detailed enough to support a review of the risk analyses; 

however, they are not intended to be as detailed as the information generally presented in 

Chemical Background documents or other comprehensive reviews. Almost no risk estimates 

presented in this document are given as single numbers. Usually, risk is expressed as a central 

estimate and a range, which is sometimes very large. Because of the need to encompass many 

different types of exposure as well as the need to express the uncertainties in the assessment, this 

risk assessment involves numerous calculations. Most of the calculations are relatively simple, 

and the very simple calculations are included in the body of the document. Some of the 

calculations, however, are cumbersome. For those calculations, worksheets are included as an 

attachment to the risk assessment. The worksheets provide the detail for the estimates cited in 

the body of the document. The worksheets are divided into the following sections: general data 

and assumptions, chemical specific data and assumptions, exposure assessments for workers, 

exposure assessments for the general public, and exposure assessments for effects on nontarget 

organisms. The worksheets for tebufenozide are contained in an EXCEL workbook and are 

included as Supplement 1 to this risk assessment. SERA (2004a) contains documentation for the 

use of these worksheets. 

1-2
 
2. PROGRAM DESCRIPTION
 

2.1. OVERVIEW 

Mimic is a commercial formulation of tebufenozide, a synthetic chemical that acts like an 

invertebrate hormone that controls molting in insects and various terrestrial and aquatic 

invertebrates. While Mimic is specifically used by the USDA for the control of the Gypsy moth, 

tebufenozide is also used in the control of other lepidopteran pest species. Mimic is comprised 

of 23-25% tebufenozide and 75-77% inert ingredients. The inert ingredients consist of glycerol, 

related reaction products, alkylaryl polyether alcohols, glyceridic and canola oils, and water. 

Additional specific information on the inerts was reviewed in the preparation of this risk 

assessment. The specific chemical identity of these inerts cannot be provided in this public 

document. Tebufenozide is relatively persistent in the environment and may be subject to 

bioconcentration. Although the compound is not highly mobile in soil, it may be transported by 

percolation or runoff from soil to ambient water. Potential concentrations of tebufenozide in 

ambient water depend largely on site specific conditions. 

Ground and aerial applications of Mimic are permitted and both methods may be considered in 

USDA programs. The labeled application rates for tebufenozide range from 2 to 8 ounces of 

Mimic/acre, corresponding to tebufenozide application rates of 0.03 lbs/acre to 0.12 lbs/acre. 

Multiple applications of tebufenozide are permitted but the maximum allowable cumulative 

amount applied is is 16 fl ounces/acre or 0.24 lb a.i./acre. The application rates for Mimic may 

vary among these USDA programs – i.e., suppression, eradication, and slow the spread. For the 

current risk assessment, the range of labeled application rates – i.e., 0.015 lb a.i./acre to 0.12 lb 

a.i./acre – are considered. All exposure assessments are conducted at the maximum application 

rate of 0.12 lb/acre, assuming two applications with a 3 day interval. This is essentially a worsecase scenario using a shortest interval between applications and two applications that reach the 

maximum annual application rate of 0.24 lb/acre. The consequences of using lesser rates are 

considered in the risk characterization for human health (Section 3.4) and ecological effects 

(Section 4.4). 

2.2. CHEMICAL DESCRIPTION AND COMMERCIAL FORMULATIONS 

Mimic 2LV, hereafter referred to simply as Mimic, is an insecticide initially registered by Rohm 

and Haas and currently registered by Dow AgroSciences (C&P Press 2004). The active 

ingredient (a.i.) in Mimic is tebufenozide, the common name for 3,5-dimethyl-, (1,1­

dimethylethyl)-2-(4-ethylbenzoyl) hydrazide benzoic acid: 

2-1
 
As detailed in Section 4.1.2.3, tebufenozide mimics the action of the invertebrate hormone 20­

hydroxyecdysone. This hormone controls molting in insects and various terrestrial and aquatic 

invertebrates. While Mimic is specifically used by USDA for the control of the Gypsy moth, 

tebufenozide is effective in the control of other lepidopteran pest species. 

Selected chemical and physical properties of tebufenozide are summarized in Table 2-1, and the 

physical and chemical properties that are directly used in this risk assessment are presented in 

worksheet B03. Dow AgroSciences also provides two other formulations, Confirm 2F and 

Confirm TO, that contains tebufenozide as the active ingredient (C&P Press 2004). 

Mimic is comprised of 23-25% tebufenozide and 75-77% inert ingredients. The inert ingredients 

consist of glycerol, related reaction products, alkylaryl polyether alcohols, glyceridic and canola 

oils (not otherwise specified), and water. The specific identity of the alkylaryl polyether alcohols 

as well as the amounts of each of the other inert ingredients is considered a trade secret 

proprietary to Dow AgroSciences. Hence, this information is not identified on the product labels 

or material safety data sheets (C&P Press 1999). Information about the impurities in technical 

grade tebufenozide were submitted to the U.S. EPA by the initial registrant (Kelly 1992; Patel 

1998) and this information was reviewed in the preparation of this risk assessment. Although 

additional specific information on the inerts cannot be provided in this public document, the 

potential impact of inert ingredients and product impurities is considered in Section 3.1.9. Spray 

adjuvants are not recommended for use with Mimic and are not given further consideration in 

this risk assessment. 

The environmental fate and transport of tebufenozide is relatively well characterized in studies 

conducted as part of the registration process for this pesticide (Hawkins 1992, 1993, 1994, 1996, 

1998) as well as in series of studies conducted by the Canadian Forest Service (Sundaram 

1994a,b, 1995, 1996, 1997a, 1997b; Sundaram et al. 1996ab, 1997a, 1997b). Pertinent 

information about the environmental fate and transport of tebufenozide is provided in Table 2-1. 

Additional detailed on environmental fate and transport are discussed in the exposure 

assessments for human health effects (Section 3.2) as well as ecological effects (Section 4.2). 

Briefly, tebufenozide is relatively persistent in the environment and may be subject to 

bioconcentration. Although the compound is not highly mobile in soil, it may be transported by 

percolation, sediment, or runoff from soil to ambient water. Potential concentrations of 

tebufenozide in ambient water depend largely on site specific conditions. 

2.3. APPLICATION METHODS 

The product label for Mimic indicates that ground or aerial applications are permitted, and both 

methods may be considered for use by the USDA. Supplemental labels indicating further 

restrictions on ground or aerial applications were not located (C&P Press 1999). 

The most common method for ground application of Mimic is hydraulic sprayers, mist blowers, 

or air blast sprayers (broadcast foliar). The spray equipment is typically mounted on tractors or 

trucks used to apply the insecticide on either side of the roadway. Usually, about 8 acres are 

2-2
 
treated in a 45-minute period (approximately 11 acres/hour). Special truck-mounted spray 

systems may be used to treat up to 12 acres in a 35-minute period with approximately 300 

gallons of insecticide mixture (approximately 21 acres/hour and 510 gallons/hour) (USDA 

1989b, p 2-9 to 2-10). 

In some instances, directed foliar applications may be used. In selective foliar applications, the 

sprayer or container containing the pesticide is carried by backpack and is applied to selected 

target vegetation. Application crews may treat up to shoulder high brush, which means that 

chemical contact with the arms, hands, or face is plausible. To reduce the likelihood of 

significant exposure, application crews are directed not to walk through treated vegetation. 

Usually, a worker treats approximately 0.5 acres/hour with a plausible range of 0.25-1.0 

acre/hour. 

In aerial applications, Mimic is applied under pressure through specially designed spray nozzles 

and booms. The nozzles are designed to minimize turbulence and maintain a large droplet size, 

both of which contribute to a reduction in spray drift. In aerial applications, approximately 10 

acres may be treated per minute (Reardon 2000). 

2.4. MIXING AND APPLICATION RATES 

The labeled application rates for tebufenozide range from 2 to 8 ounces of Mimic/acre, 

corresponding to tebufenozide application rates of 0.03 lbs/acre to 0.12 lbs/acre. This range of 

application rates is recommended for the control of Gypsy moth and several other lepidopteran 

pest species. The highest recommended application rate for any species is 8 ounces of 

Mimic/acre or 0.12 lb tebufenozide per acre. This is the only application rate recommended for 

the control of the pine tip moth. Application rates from 4 to 8 ounces of Mimic per acre are 

recommended on the label for gypsy moth. The maximum amount of Mimic that may be applied 

per year is 16 fl ounces/acre or 0.24 lb a.i./acre (C&P Press 2004). 

Commercial formulations of tebufenozide are diluted with water prior to application. In ground 

applications, application volumes of 50 gallons per acre are recommended for hydraulic ground 

sprayers and a minimum of 10 gallons per acre is recommended for mist blowers or air blast 

sprayers. For aerial applications, a minimum of 0.5 gallon per acre is recommended. As 

specified on the product label, uniform coverage is essential for efficacy and higher spray 

volumes are recommended for large trees, dense stands, and/or heavy infestations (C&P Press 

2004). 

The USDA has adopted various intervention strategies that are roughly categorized as 

suppression, eradication, and Slow-the-Spread (Liebhold and McManus 1999). These programs 

may be conducted by either the USDA Forest Service or the Animal and Plant Health Inspection 

Service (APHIS). Suppression efforts are conducted in areas of well established gypsy moth 

infestations to combat or interdict periodic gypsy moth population outbreaks. Eradication efforts 

are intended to completely eliminate gypsy moth populations in areas where new populations of 

the gypsy moth are found. Slow-the-Spread, as the name implies, is a program to reduce the 

2-3
 
expansion of gypsy moth populations from areas of established populations to adjacent noninfested areas. 

The application rates for Mimic may vary among these USDA programs. For the USDA Forest 

Service, the typical application rates will range from 0.015 to 0.06 lb a.i. per acre. A single 

application is used in suppression programs and two to three applications may be made in 

eradication programs. Mimic as well as other formulations of tebufenozide may be reapplied. 

The interval between applications in Forest Service programs will generally be 3 to 10 days. The 

Forest Service may consider using the maximum application rate of 0.12 lb a.i./acre in some 

instances (Cook 2004). In eradication programs, APHIS will use an application rate of 0.06 lb 

a.i. per acre. Two applications may be made with an application interval of 7 to 10 days. 

For the current risk assessment, the range of labeled application rates – i.e., 0.015 lb a.i./acre to 

0.12 lb a.i./acre – are considered. All exposure assessments will be conducted at the maximum 

application rate of 0.12 lb/acre, assuming two applications with a 3 day interval. This is 

essentially a worst-case scenario using a shortest interval between applications and two 

applications that reach the maximum annual application rate of 0.24 lb/acre. The consequences 

of using lesser rates are considered further in the risk characterization for human health (Section 

3.4) and ecological effects (Section 4.4). 

Mimic is diluted prior to application. In this risk assessment, the extent to which Mimic is 

diluted prior to application primarily influences dermal and direct spray scenarios, both of which 

depend on the ‘field dilution’ (i.e., the concentration of tebufenozide in the applied spray). 

Invariably, the higher the concentration of tebufenozide, the greater the risk. For this risk 

assessment, the lowest dilution is taken at 0.5 gallon/acre, the minimum recommended for aerial 

applications. The highest dilution (i.e., that which results in the lowest risk) is based on 50 

gallons of water per acre, the highest application volume specifically recommended on the 

product label (C&P Press 2004). The central estimate is taken as 5 gallons of water per acre, the 

geometric mean of the range. Detailed calculations of field dilution rates are provided in 

worksheet B01, and the calculations following worksheet B01 and the values used in various 

exposure assessments are summarized in worksheet B02. 

2.5. USE STATISTICS 

Neither Mimic nor other pesticides containing tebufenozide have been used previously by the 

USDA in full scale control programs. Consequently past use statistics that might reflect the 

amounts of tebufenozide that may be used in USDA programs are not available. Experimental 

programs have been conducted by the USDA in the northeast and have involved the treatment of 

experimental plots ranging from 16 to 135 acres (Reardon 2000). 

Tebufenozide was used extensively as a pest control agent on cotton. In 1992, the most recent 

year for which data are available, 42,104 lbs were used for that purposes. As illustrated in Figure 

2-1, all of the tebufenozide applied to cotton in1992 was used in Texas and Mississippi (USGS 

1998). 

2-4
 
Tebufenozide is used in Canada at an application rate of 0.07 kg a.i./ha or 0.062 lb a.i./acre to 

control spruce budworms. In 1994, only 400 acres were treated; however, in 1997, 14,875 acres 

were treated (Canadian Council of Forest Ministers 1999), and the amount of tebufenozide used 

is calculated as 922.25 lbs [14,875 acres × 0.062 lb a.i./acre]. 

2-5
 
3. HUMAN HEALTH RISK ASSESSMENT
 

3.1. HAZARD IDENTIFICATION 

3.1.1. Overview 

A relatively detailed and consistent series of studies in mice, rats, and dogs indicates that the 

primary mechanism of tebufenozide toxicity in mammals involves hematological effects, 

specifically the formation of methemoglobin. Tebufenozide does not appear to be carcinogenic 

and does not appear to cause birth defects. Nonetheless, the compound is associated with 

adverse reproductive effects in experimental mammals. Tebufenozide itself does not seem to be 

irritating to the skin or eyes. Mimic, however, appears to contain other constituents (inerts or 

adjuvants) that may cause skin or eye irritation. 

As discussed in the exposure assessment, dermal absorption is the primary route of exposure for 

workers. Data regarding the dermal absorption kinetics of tebufenozide are not available in the 

published or unpublished literature. For this risk assessment, estimates of dermal absorption 

rates are based on quantitative structure-activity relationships. The estimated dermal absorption 

rates are used in turn to estimate the amounts of tebufenozide that might be absorbed by workers. 

Then, those estimates are used with the available dose-response data to characterize risk. 

Although the lack of experimental data regarding dermal absorption of tebufenozide adds 

uncertainties to this risk assessment, the available data regarding the oral and dermal toxicity of 

tebufenozide are sufficient to suggest that the estimated dermal absorption rates are plausible. 

The inhalation toxicity of tebufenozide is not well documented. Irritant effects have been noted 

in laboratory studies involving exposures to very high concentrations of tebufenozide in air. 

Because inhalation exposure involving high concentrations of tebufenozide is implausible under 

normal field conditions, the potential inhalation toxicity of the compound is not of substantial 

concern to this risk assessment. 

3.1.2. Mechanism of Action 

In mammals, tebufenozide is known to damage hemoglobin, a key component of blood, through 

the formation of methemoglobin. This is highly relevant to the human health risk assessment 

because effects on the blood are the basis for the U.S. EPA RfD for tebufenozide (Section 3.3). 

Hemoglobin is the component in red blood cells that is responsible for transporting oxygen 

throughout the body. If this function is impaired, either because of damage to hemoglobin or 

lack of oxygen in the air, serious adverse effects (i.e., equivalent to suffocation) can occur. The 

formation of both methemoglobin and sulfhemoglobin can cause such impairment and lead to the 

formation of methemoglobinemia and sulfhemoglobinemia, respectively. Methemoglobin is 

formed by the oxidation of the heme iron in hemoglobin from the ferrous (Hb++) to the ferric 

state (MetHb+++) (Bradberry 2003; Smith 1996). Heme group oxidation occurs spontaneously 

and accounts for approximately 2% of the hemoglobin in normal individuals. Methemoglobin is 

reduced (restored to its natural state) by a set of enzymes referred to as methemoglobin 

reductases. Some individuals are deficient in NADH-dependent methemoglobin reductase, in 

3-1
 
which case as much as 50% of their blood pigment may exist as methemoglobin. Newborns are 

also deficient in NADH-methemoglobin reductase. 

While tebufenozide displays other types of toxicity, as discussed in the following subsections, the 

formation of methemoglobin is the only mechanisms of toxicity that has been clearly identified. 

3.1.3. Kinetics and Metabolism 

3.1.3.1. Pharmacokinetic Studies – The pharmacokinetics of tebufenozide have been 

studied in rats after oral doses of 3 or 250 mg/kg of 14C-labeled tebufenozide (Struble and 

Hazelton 1992). Tebufenozide was rapidly absorbed and excreted. Concentrations of 

tebufenozide in blood were not linearly related to dose. Concentrations of tebufenozide in the 

blood were only about 4 to 6 times those in the low dose. While absorption rates are not 

calculated in Struble and Hazelton (1992), this pattern suggests a less rapid absorption rate in the 

high dosed animals or a saturation of critical pathways involving absorption. About 75% to 99% 

was excreted in the feces during the first 24 hours with virtually complete excretion by 48 hours 

after dosing. In the blood, most of the radioactivity was associated with blood cells rather than 

plasma – i.e., blood to plasma ratios of 10:1 to 15:1. 

3.1.3.2. Dermal Absorption Rates – As detailed further in Section 3.2.2.2, two types of 

dermal exposure scenarios are considered in this risk assessment: those involving direct contact 

with a solution of the herbicide (e.g., immersion) and those associated with accidental spills of 

the herbicide onto the surface of the skin. 

As detailed in SERA (2001), dermal exposure scenarios involving immersion or prolonged 

contact with chemical solutions use Fick's first law and require an estimate of the permeability 

coefficient, K , expressed in cm/hour. Using the method recommended by U.S. EPA (1992), the p

estimated dermal permeability coefficient for tebufenozide is 0.013 cm/hour with a 95% 

confidence interval of 0.0066-0.025 cm/hour. These estimates are used in all exposure 

assessments that are based on Fick’s first law. For exposure scenarios like direct sprays or 

accidental spills, which involve deposition of the compound on the skin’s surface, dermal 

absorption rates (proportion of the deposited dose per unit time) rather than dermal permeability 

rates are used in the exposure assessment. The estimated first-order dermal absorption 

-1 -1 coefficient is 0.0032 hour with 95% confidence intervals of 0.0012-0.0082 hour . The 

calculations for these estimates are presented in Appendix 1. Note that the values for both 

dermal permeability and the first order dermal absorption rates are rounded to two significant 

figure in Table A1-5 of Appendix 1 and these values are entered into Worksheet A03 and used in 

all scenarios involving dermal exposures for both workers (Worksheet Series C) and the general 

public (Worksheet Series D). 

There are no experimental data regarding the absorption of tebufenozide by humans. 

Wederbrand and Potter (1993) report that a proportion of 0.05 of a dermal dose of tebufenozide 

was absorbed by rats after 10 hours. The 14C-tebufenozide was dissolved in a solution that 

approximated the 2F formulation – i.e., Confirm. While the specific ingredients in the 

3-2
 
 

formulation are specified in a confidential appendix to this study, these ingredients (other than 

the general description given in Section 2) cannot be disclosed in this risk assessment. Taking 

0.05 as the absorbed dose, the first-order dermal absorption coefficient would be about [k = 

-ln(1-0.05)/10 hours = 0.005 per hour]. This is very close to the estimate of 0.0032 hour-1 given 

above. Thus, at least for short term exposures, the available data on absorption kinetics in rats 

are consistent with the estimate of the human first-order dermal absorption rate. Consequently, 

the lack of human data regarding the dermal absorption rate of tebufenozide adds relatively little 

uncertainty to this risk assessment. In addition, the available dermal toxicity data are adequate to 

address this uncertainty to some extent (Section 3.1.12.). 

3.1.4. Acute Toxicity 

Information regarding the acute oral toxicity of tebufenozide is summarized in Appendix 2. All 

of the available studies are standard bioassays conducted as part of the registration process for 

Mimic. Tebufenozide has a very low order of acute toxicity to mammals. Single oral gavage 

doses of 2000 mg/kg caused no observable signs of toxicity in mice or rats (Hazleton and Quinn 

1995b; Swenson et al. 1994). Mimic, the commercial formulation of tebufenozide covered in 

this risk assessment, caused no signs of toxicity at doses of up to 5 g/kg or 5000 mg/kg (Parno 

and Gingrich 1994b). Mimic contains 23-25% tebufenozide by weight (see section 2), which 

corresponds to tebufenozide doses of about 1250 mg/kg body weight. As discussed in section 

3.1.9.3, Mimic contains inert ingredients, the identity of which cannot be disclosed in this 

document. The lack of evidence that Mimic is toxic at a dose of 5000 mg/kg is consistent with 

the acute toxicity data on tebufenozide. Although this observation cannot be overly interpreted, 

it does at least suggest that the inerts in Mimic do not have a high order of acute oral toxicity. 

3.1.5. Subchronic or Chronic Systemic Toxic Effects 

Information on the subchronic and chronic oral toxicity of tebufenozide is summarized in 

Appendix 2. Like the acute studies, all of these studies were conducted as part of the registration 

process. 

Appendix 2 summarizes subchronic studies in mice, rats, and dogs, with exposure durations 

ranging from 2 weeks to 90 days. The most consistently observed effects are related to the 

formation of methemoglobin, which can lead to decreases in red blood cell volume due to the 

destruction of the red blood cells (i.e., hemolytic anemia). 

Methemoglobin induction involves the chemical oxidation of the heme iron in hemoglobin from 

the ferrous (Hb++) to the ferric state (MetHb+++), resulting in the inability of hemoglobin to 

combine reversibly with oxygen (Smith 1996). Heme group oxidation occurs spontaneously and 

accounts for approximately 2% of the hemoglobin in normal individuals. Methemoglobin is 

reduced (restored to its natural state) by a set of enzymes referred to as methemoglobin 

reductases. The most common methemoglobin reductase is dependent on NADH. Some 

individuals are deficient in NADH-dependent methemoglobin reductase, in which case, as much 

as 50% of their blood pigment may exist as methemoglobin. Newborns are also deficient in 

NADH-methemoglobin reductase. Aromatic amines are known to induce methemoglobinemia, 

3-3
 
most likely by the formation of N-hydroxy metabolites (Smith 1996). 

As discussed in section 3.3.2, methemoglobin formation and other effects on blood are the most 

sensitive endpoints for tebufenozide and is the basis for the U.S. EPA RfD for this compound. In 

test animals, specific changes in hematological parameters included decreases red blood cell 

count, mean cell volume, reticulocyte counts, methemoglobin, the incidence of Heinz bodies, and 

platelet counts as well as increases in spleen weight. The quantitative dose-response 

relationships for this effect are discussed further in section 3.3. Increased liver weight also was 

observed in three animal species [mice and rats (Osheroff 1991a,b), dogs (Clay 1992)]. This 

effect may be secondary to the formation of methemoglobin, which increases the destruction of 

red blood cells in the liver (Richards 1992a,b). Theoretically, increased liver weight may be 

observed as the result of enzyme induction in which a compound will induce enzymes that are 

associated with its own metabolism. This induction can lead to an increase in total liver weight 

and is often regarded as an adaptive rather than toxic response (Moslen 1996). 

The chronic toxicity of tebufenozide was assayed in dogs (Richards 1992a,b), mice (Trutter 

1992a,b) and rats (Trutter 1992c). As in the subchronic studies, signs of hemolytic anemia were 

observed in all three species. 

3.1.6. Effects on Nervous System 

As discussed in Durkin and Diamond (2002), a neurotoxicant is a chemical that disrupts the 

function of nerves, either by interacting with nerves directly or by interacting with supporting 

cells in the nervous system. This definition of neurotoxicant distinguishes agents that act directly 

on the nervous system (direct neurotoxicants) from those agents that might produce neurologic 

effects that are secondary to other forms of toxicity (indirect neurotoxicants). Virtually any 

chemical will cause signs of neurotoxicity in severely poisoned animals and, thus, can be 

classified as an indirect neurotoxicant. 

In a standard assay for neurotoxicity, no signs of toxicity were noted in rats after single oral doses 

up to 2000 mg/kg (Swanson et al. 1994). In addition, signs of neurotoxicity have not been noted 

in a large number of acute and chronic toxicity studies (Appendices 2 and 3). 

3.1.7. Effects on Immune System 

Immunotoxicants are chemical agents that disrupt the function of the immune system. Two 

general types of effects, suppression and enhancement, may be seen and both of these are 

generally regarded as adverse. Agents that impair immune responses (immune suppression) 

enhance susceptibility to infectious diseases or cancer. Enhancement or hyperreactivity can give 

rise to allergy or hypersensitivity, in which the immune system of genetically predisposed 

individuals inappropriately responds to chemical or biological agents (e.g., plant pollen, cat 

dander, flour gluten) that pose no threat to other individuals or autoimmunity, in which the 

immune system produces antibodies to self components leading to destruction of the organ or 

tissue involved. 

3-4
 
There is very little direct information on which to assess the immunotoxic potential of 

tebufenozide. The only studies specifically related to the effects of tebufenozide on immune 

function are skin sensitization studies (Section 3.1.11). While the studies by Anderson and 

Shuey (1994) and Glaza (1993) indicate that tebufenozide is not a skin sensitizer, this provides 

no information useful for directly assessing the potential for tebufenozide to suppress or 

otherwise disrupt immune function. 

Nonetheless, the toxicity of tebufenozide has been examined in numerous acute, subchronic, and 

chronic bioassays. Although many of these studies did not focus on the immune system, changes 

in the immune system (which could potentially be manifest as increased susceptibility to 

infection compared to controls) were not observed in any of the available long-term animal 

studies (Appendix 2). Typical subchronic or chronic animal bioassays conduct morphological 

assessments of the major lymphoid tissues, including bone marrow, major lymph nodes, spleen 

and thymus (thymus weight is usually measured as well), and blood leukocyte counts. These 

assessments can detect signs of inflammation or injury indicative of a direct toxic effect of the 

chemical on the lymphoid tissue. Changes in cellularity of lymphoid tissue and blood, indicative 

of a possible immune system stimulation or suppression, can also be detected (Durkin and 

Diamond 2002). None of these effects have been noted in any of the longer term toxicity studies 

on tebufenozide (Appendix 2). 

3.1.8. Effects on Endocrine System 

The endocrine system participates in the control of metabolism and body composition, growth 

and development, reproduction, and many of the numerous physiological adjustments needed to 

maintain constancy of the internal environment (homeostasis). The endocrine system consists of 

endocrine glands, hormones, and hormone receptors. Endocrine glands are specialized tissues 

that produce and export (secrete) hormones to the bloodstream and other tissues. The major 

endocrine glands in the body include the adrenal, hypothalamus, pancreas, parathyroid, pituitary, 

thyroid, ovary, and testis. Hormones are also produced in the gastrointestinal tract, kidney, liver, 

and placenta. Hormones are chemicals produced in endocrine glands that bind to hormone 

receptors in target tissues. Binding of a hormone to its receptor results in a process known as 

postreceptor activation which gives rise to a hormone response in the target tissue, usually an 

adjustment in metabolism or growth of the target tissue. Examples include the release of the 

hormone testosterone from the male testis, or estrogen from the female ovary, which act on 

receptors in various tissues to stimulate growth of sexual organs and development of male and 

female sexual characteristics. The target of a hormone can also be an endocrine gland, in which 

case, receptor binding may stimulate or inhibit hormone production and secretion. Adverse 

effects on the endocrine system can result in abnormalities in growth and development, 

reproduction, body composition, homeostasis (the ability to tolerate various types of stress), and 

behavior. 

There is no indication that tebufenozide causes endocrine disruption in experimental mammals. 

Tebufenozide showed no activity in an in vitro test system (human estrogen receptor cDNA in 

the yeast, Saccharomyces cerevisae) for the human estrogen receptor (Cress 1996). In addition, 

3-5
 
standard subchronic, chronic and reproductive toxicity studies (Section 3.1.9) provide no basis 

for asserting that any signs of overt toxicity are related to changes in endocrine function in 

mammals. 

3.1.9. Reproductive and Teratogenic Effects 

Tebufenozide was tested for its ability to cause birth defects (i.e., teratogenicity) as well as its 

ability to cause reproductive impairment. All of these studies are discussed in Appendix 2. Like 

the acute, subchronic, and chronic studies, all of the reproductive and developmental studies are 

unpublished and were conducted in support of the registration of this compound. 

Teratogenicity studies usually entail gavage administration to pregnant rats or rabbits on specific 

days of gestation. Two such studies were conducted on tebufenozide: one in rats (Hoberman 

1991) and one in rabbits (Swenson and Solomon 1992). No signs of teratogenicity or fetal 

toxicity were noted in either study. In the rat study, decreased weight gain was observed in dams 

treated with the highest dose (1000 mg/kg). Even at this dose, however, developmental effects 

were not observed. 

Another type of reproduction study involves exposing more than one generation of the test 

animal to the compound. In other words, both the parent animals and the offspring are exposed 

to the substance. Two such studies (Aso 1995; Danberry et al. 1993) were conducted on 

tebufenozide. In the study by Aso (1995), signs of toxicity to the blood were observed in both 

male and female adult rats at dietary concentrations of 200 and 2000 ppm but not at a dietary 

concentration of 25 ppm. For offspring, no effects were observed at dietary concentrations of 25 

or 200 ppm; however, treatment with 2000 ppm caused decreases in body weight. At the dietary 

concentration of 2000 ppm, the estimated dose levels were 126.0 mg/kg/day for males and 143.2 

mg/kg/day for females (U.S. EPA 1999b). In the rat study by Danberry et al. (1993), no 

reproductive effects were observed at a dietary concentration of 150 ppm (.12 mg/kg bw). At 

2000 ppm (.160 mg/kg bw), however, there was an increased incidence of mortality among 

females during delivery (P2), an increase in gestation length (P2), a decrease in the mean number 

of implantation sites per female (P2), and an increased incidence of pregnant females that did not 

deliver (P1 and P2). 

As discussed further in section 4, there is concern for potential reproductive effects in birds. 

Based on a dietary study in quail (Beavers et al. 1993b), dietary concentrations of 300 or 1000 

ppm, corresponding to estimated doses of 45 or 150 mg/kg bw, were associated with decreases in 

hatching and other indices of reproductive toxicity. 

3.1.10. Carcinogenicity and Mutagenicity 

Trutter (1992a,b,c) assayed the potential carcinogenicity of tebufenozide in an 18-month bioassay 

in mice and a 24-month bioassay in rats. Both studies, summarized in Appendix 2, were 

accepted by the U.S. EPA (1999b). Moreover, neither of the two studies shows evidence of 

carcinogenicity. 

3-6
 
Tebufenozide was assayed also for mutagenic activity in a number of test systems with uniformly 

negative results. At a maximum concentration of 5000 μg a.i./ plate, tebufenozide was not 

mutagenic to Salmonella typhimurium strains TA98, TA100, TA1535 and TA1537 with or 

without metabolic activation (S-9 liver fraction from Aroclor 1254 induced rats) (Black 1992; 

Sames and Elia 1993). In addition, tebufenozide did not induce gene mutations (HGPRT locus) 

in Chinese hamster ovary (CHO) cells with or without S-9 activation (Thilagar 1988, 1990a) and 

was also negative in an in vivo chromosome aberration assay in rat bone marrow cells (Gudi 

1992). Finally, tebufenozide failed to induce DNA damage in primary rat hepatocytes (Thilagar 

1990b). 

Based on the lack of carcinogenic activity from in vivo assays and the lack of mutagenic activity 

in several in vitro assays, tebufenozide is classified as a Group E chemical (i.e., no evidence of 

carcinogenicity for humans) (U.S. EPA 1999b). 

3.1.11. Irritation and Sensitization (Effects on the Skin and Eyes) 

Tebufenozide was tested for toxic effects after dermal exposure as well as irritant effects on the 

skin and eyes of rabbits (Appendix 3). Technical grade tebufenozide does not appear to be an 

eye irritant (Hazleton and Quinn 1995b); nevertheless, a commercial formulation was shown to 

cause moderate eye irritation in rabbits (Gingrich and Parno 1994). The available studies on 

Mimic suggest that the other components in the formulation can cause skin irritation in rats 

(Morrison et al. 1993) and rabbits (Parno 1997). Neither tebufenozide nor Mimic, however, 

appear to cause skin sensitization in guinea pigs (Anderson and Shuey 1994; Glaza 1993). 

The product label for Mimic advises that the formulation may cause moderate eye irritation and 

that contact with eyes, skin, or clothing should be avoided. This kind of advisory is, of course, 

standard and prudent practice for any chemical. 

3.1.12. Systemic Toxic Effects from Dermal Exposure 

Single dermal applications of technical grade tebufenozide are not toxic to rats at applied doses 

of up to 5000 mg/kg . These findings are consistent with the data indicating that tebufenozide 

has a low order of oral toxicity. Similarly, technical grade tebufenozide caused no signs of 

toxicity in rats and no hematological changes in rats when a dose of 1000 mg/kg was applied 

directly to the skin 5 days per week for 4 weeks (Hazleton and Quinn 1995b). 

As indicated in Appendix 3, technical grade tebufenozide caused no signs of toxicity in rats and 

no change in hematological parameters in rats when applied directly to the skin at a dose of 1000 

mg/kg, 6 hours per day, 5 days per week for 4 weeks (Hazleton and Quinn 1995b). Given the 

estimated first-order dermal absorption rate coefficient of 0.00317 hour-1 (Section 3.1.3.2), the 

absorbed dose from this exposure may be estimated at about 13.5 mg/kg/day: 

1000 mg/kg/day × (1-e-0.00317×6) × 5/7 = 13.45 mg/kg/day. 

As also summarized by Hazleton and Quinn (1995b) and detailed in Appendix 2, dietary 

3-7
 
 

concentrations of 1000 ppm tebufenozide for 2 weeks caused hematological effects in rats; 

however, the effects were not observed in rats exposed to 250 ppm. In this study, rats consumed 

food amounts equivalent to about 7% of their body weight per day. Thus, the dietary 

concentrations correspond to doses of 17.5 mg/kg/day (NOAEL of 250 ppm × 0.07 mg/kg per 

ppm) and 70 mg/kg/day (LOAEL of 1000 ppm × 0.07 mg/kg per ppm). Therefore, the estimate 

of the first-order dermal absorption rate is at least consistent with the comparable NOAEL values 

for oral and dermal exposures. 

3.1.13. Inhalation Exposure 

Acute inhalation studies are required for the registration of pesticides and three studies were 

submitted to U.S. EPA, one on technical grade tebufenozide, summarized by Hazleton and Quinn 

(1995b) and two conducted on wettable powder and LV Mimic formulations (Bemacki and 

Ferguson 1994a,b). At the highest technically achievable concentration of 0.43 mg/L, no 

mortality was observed in rats over a 2-week observation period after a single 4-hour exposure. 

At a concentration of 1.83 mg/L for 4 hours, the wettable formulation also caused no mortalities 

and no gross lesions (Bemacki and Ferguson 1994a). The liquid LV formulation, however, 

caused irritant changes in the respiratory tract after a single 4-hour exposure to 1.33 mg/L. Thus, 

as with dermal irritation, the liquid formulation of Mimic appears to be a greater irritant than 

tebufenozide. 

These limited data suggest that the liquid formulation, LV Mimic, can induce irritant effects at 

very high exposure levels. Since the wettable powder did not produce irritant effects, the 

observed effects after exposure to LV Mimic may have been due to the presence of different 

materials in the LV Mimic formulation or due to the differences in the physical form – i.e., liquid 

and solid. As discussed in section 3.3, this effect by LV Mimic is not directly relevant to this 

risk assessment because of the implausibility of exposure to high concentrations of the 

compound. 

3.1.14. Inerts and Adjuvants 

Mimic contains materials other than technical grade tebufenozide that are included as inerts or 

adjuvants to improve either efficacy or ease of handling and storage. The identity of these 

materials is confidential. The additives were disclosed to the U.S. EPA and were reviewed in the 

preparation of this risk assessment. All that can be disclosed explicitly is that none of the 

additives is classified by the U.S. EPA as toxic. 

Notwithstanding this assertion, it is apparent from a comparison of the acute dermal and 

inhalation data on technical grade tebufenozide and Mimic (see Sections 3.1.12 and 3.1.13) that 

Mimic contains materials that cause irritant effects not characteristic of technical grade 

tebufenozide. Thus, in terms of acute irritant effects that might be associated with the handling 

or application of Mimic, it is likely that the adjuvants or other inerts are of greater concern than 

tebufenozide. In terms of potential systemic toxic effects, however, there is no information to 

suggest that the adjuvants or inerts have an impact on the toxicity of this product. 

3-8
 
3.1.15. Impurities and Metabolites 

3.1.15.1. Impurities – There is no published information regarding the impurities in 

technical grade tebufenozide or any of its commercial formulations. Information on all of the 

impurities in technical grade tebufenozide were disclosed to the U.S. EPA, and the information 

was obtained and reviewed as part of this risk assessment (Kelly 1992). Because this 

information is classified as confidential business information, details about the impurities cannot 

be disclosed. Nonetheless, all of the toxicology studies on tebufenozide involve technical 

tebufenozide, which is presumed to be the same as or comparable to the active ingredient in the 

formulation used by the Forest Service. Thus, if toxic impurities are present in technical 

tebufenozide, they are likely to be encompassed by the available toxicity studies using technical 

grade tebufenozide. 

3.1.15.2. Metabolites – As reviewed by the U.S. EPA (1999b), tebufenozide is subject to 

metabolism in mammals and more than 10 metabolites have been identified. The metabolic 

pathway appears primarily to involve oxidation of aliphatic groups on the benzyl rings to 

alcohols, aldehydes, or acids. No cleavage of the aliphatic rings has been noted. Since all of the 

in vivo toxicology studies on tebufenozide involve the generation of metabolites, the potential 

toxicity of the metabolites should be encompassed by the available toxicity data on tebufenozide. 

Major metabolites of tebufenozide have a low order of acute oral toxicity (LD50 values >5000 

mg/k) and are inactive in bacterial mutagenicity assays (Quinn 1997). 

3.1.16. Toxicologic Interactions 

No information has been encountered on the toxicologic interactions of tebufenozide with other 

agents. As discussed in Section 3.1.2, tebufenozide causes methemoglobinemia in mammals. 

Many other chemicals may cause this effect and, as discussed in Section 3.4.5, interactions 

between tebufenozide and these agents are most likely to be additive rather than synergistic or 

antagonistic. 

3-9
 
 

3.2. EXPOSURE ASSESSMENT 

3.2.1. Overview. 

Standard sets of exposure scenarios are presented for both workers and members of the general 

public. The exposure assessments for these groups are summarized in Worksheet E01 (workers) 

and Worksheet E03 (general public). All exposure assessments are conducted at the maximum 

application rate for tebufenozide of 0.12 lb/acre using two applications with a minimum 

application interval of three days. This cumulative application (0.24 lb a.i./acre) is the maximum 

application rate for a single season. This leads to the highest estimates of peak as well as longer 

term exposures. The consequences of using lower application rates are discussed in the risk 

characterization (Section 3.4). 

For workers applying tebufenozide, three types of application methods are modeled: directed 

ground spray, broadcast ground spray, and aerial spray. Central estimates of exposure for 

workers are approximately 0.002 mg/kg/day for aerial and backpack workers and about 0.003 

mg/kg/day for broadcast ground spray workers. Upper ranges of exposures are approximately 

0.02 mg/kg/day for broadcast ground spray workers and 0.01 mg/kg/day for backpack and aerial 

workers. All of the accidental exposure scenarios for workers involve dermal exposures and 

most of these accidental exposures lead to estimates of dose that are either in the range of or 

substantially below the general exposure estimates for workers. The one exception involves 

wearing contaminated gloves for one-hour. The upper range of exposure for this scenario is 

about 4 mg/kg/day. 

For the general public, the range of acute exposures is from approximately 0.0000002 mg/kg 

associated with the lower range for the consumption of contaminated water from a stream by a 

child to 1.2 mg/kg associated with the upper range for consumption of contaminated water by a 

child after an accidental spill. Relatively high dose estimates are also associated with the direct 

spray of a child (about 0.4 mg/kg at the upper range of exposure) and for the consumption of fish 

after an accidental spill by members of the general public (0.2 mg/kg) and subsistence 

populations (0.9 mg/kg). Other acute exposure scenarios are associated with doses that are lower 

by at least an order of magnitude. For chronic or longer term exposures, the modeled exposures 

are much lower than for acute exposures, ranging from approximately 0.000000002 mg/kg/day (2 

in 1 billionth of a mg/kg/day) associated with the lower range for the consumption of 

contaminated water to approximately 0.03 mg/kg/day associated with the upper range for 

consumption of contaminated fruit. 

3.2.2. Workers. 

The Forest Service uses a standard set of exposure assessments in all risk assessment documents. 

While these exposure assessments vary depending on the characteristics of the specific chemical 

as well as the relevant data on the specific chemical, the organization and assumptions used in 

the exposure assessments are standard and consistent. All of the exposure assessments for 

workers as well as members of the general public are detailed in the worksheets on tebufenozide 

that accompany this risk assessment (Supplement 1) and documentation for these worksheets is 

given in SERA (2003). A copy of this documentation is available at www.sera-inc.com. This 

3-10
 
section on workers and the following section on the general public provides are plain verbal 

description of the worksheets and discuss tebufenozide specific data that are used in the 

worksheets. 

A summary of the exposure assessments for workers is presented in Worksheet E01 of the 

worksheets for tebufenozide that accompany this risk assessment. Two types of exposure 

assessments are considered: general and accidental/incidental. The term general exposure 

assessment is used to designate those exposures that involve estimates of absorbed dose based on 

the handling of a specified amount of a chemical during specific types of applications. The 

accidental/incidental exposure scenarios involve specific types of events that could occur during 

any type of application. The exposure assessments developed in this section as well as other 

similar assessments for the general public (Section 3.2.3) are based on two applications spaced 

three days apart at the maximum single application rate of 0.12 lb/acre (Section 2). The 

consequences of using lower application rates are discussed further in the risk characterization 

(Section 3.4). 

3.2.2.1. General Exposures – No studies on worker exposures to tebufenozide are available. 

As described in SERA (2001), worker exposure rates are expressed in units of mg of absorbed 

dose per kilogram of body weight per pound of chemical handled. Based on analyses of several 

different pesticides using a variety of application methods, default exposure rates are estimated 

for three different types of applications: directed foliar (backpack), boom spray (hydraulic ground 

spray), and aerial. 

The specific assumptions used for each application method are detailed in Worksheets C01a 

(directed foliar), C01b (broadcast foliar), and C01c (aerial). In the worksheets, the central 

estimate of the amount handled per day is calculated as the product of the central estimates of the 

acres treated per day and the application rate. 

Estimates of worker exposure rates are expressed in units of mg of absorbed dose per kilogram of 

body weight per pound of chemical handled. These estimates of exposure rates are based on 

worker exposure studies on nine different pesticides with molecular weights ranging from 221 to 

416 and log Kow values ranging from -0.75 to 6.50. The estimated exposure rates are based on 

estimated absorbed doses in workers as well as the amounts of the chemical handled by the 

workers. As summarized in Table 2-1 of this risk assessment, the molecular weight of 

tebufenozide is 352.48 and the log Kow is about 4.25. These values are within the range of the 

pesticides used in SERA (2001) to estimate worker exposures. As discussed in SERA (2001), 

the ranges of estimated occupational exposure rates vary substantially among individuals and 

groups, (i.e., by a factor of 50 for backpack applicators and a factor of 100 for mechanical ground 

sprayers). It seems that much of the variability can be attributed to the hygienic measures taken 

by individual workers (i.e., how careful the workers are to avoid unnecessary exposure); 

however, pharmacokinetic differences among individuals (i.e., how individuals absorb and 

excrete the compound) also may be important. 

3-11
 
The number of acres treated per hour is taken from previous USDA risk assessments (USDA 

1989a,b,c). The number of hours worked per day is expressed as a range, the lower end of which 

is based on an 8-hour work day with 1 hour at each end of the work day spent in activities that do 

not involve exposure to the compound. The upper end of the range, 8 hours per day, is based on 

an extended (10-hour) work day, allowing for 1 hour at each end of the work day to be spent in 

activities that do not involve exposure to the chemical. 

It is recognized that the use of 6 hours as the lower range of time spent per day applying 

herbicides is not a true lower limit. It is conceivable and perhaps common for workers to spend 

much less time in the actual application of a herbicide if they are engaged in other 

activities. Thus, using 6 hours may overestimate exposure. In the absence of any published or 

otherwise documented work practice statistics to support the use of a lower limit, this approach is 

used as a protective assumption. 

The range of acres treated per hour and hours worked per day is used to calculate a range for the 

number of acres treated per day. For this calculation as well as others in this section involving 

the multiplication of ranges, the lower end of the resulting range is the product of the lower end 

of one range and the lower end of the other range. Similarly, the upper end of the resulting range 

is the product of the upper end of one range and the upper end of the other range. This approach 

is taken to encompass as broadly as possible the range of potential exposures. 

The central estimate of the acres treated per day is taken as the arithmetic average of the range. 

Because of the relatively narrow limits of the ranges for backpack and boom spray workers, the 

use of the arithmetic mean rather than some other measure of central tendency, like the geometric 

mean, has no marked effect on the risk assessment. 

3.2.2.2. Accidental Exposures – Typical occupational exposures may involve multiple routes of 

exposure (i.e., oral, dermal, and inhalation); nonetheless, dermal exposure is generally the 

predominant route for herbicide applicators (Ecobichon 1998; van Hemmen 1992). Typical 

multi-route exposures are encompassed by the methods used in Section 3.2.2.1 on general 

exposures. Accidental exposures, on the other hand, are most likely to involve splashing a 

solution of herbicides into the eyes or various dermal exposure scenarios. 

Tebufenozide may cause eye irritation (Section 3.1.11). The available literature does not include 

quantitative methods for characterizing exposure or responses associated with splashing a 

solution of a chemical into the eyes; furthermore, there appear to be no reasonable approaches to 

modeling this type of exposure scenario quantitatively. Consequently, accidental exposure 

scenarios of this type are considered qualitatively in the risk characterization (section 3.4). 

As detailed in Section 3.1.3, there are various methods for estimating absorbed doses associated 

with accidental dermal exposure (U.S. EPA 1992; SERA 2001). Two general types of exposure 

are modeled: those involving direct contact with a solution of the herbicide and those associated 

with accidental spills of the herbicide onto the surface of the skin. Any number of specific 

3-12
 
exposure scenarios could be developed for direct contact or accidental spills by varying the 

amount or concentration of the chemical on or in contact with the surface of the skin and by 

varying the surface area of the skin that is contaminated. 

For this risk assessment, two exposure scenarios are developed for each of the two types of 

dermal exposure, and the estimated absorbed dose for each scenario is expressed in units of mg 

chemical/kg body weight. Both sets of exposure scenarios are summarize in Worksheet E01, 

which references other worksheets in which the specific calculations are detailed. 

Exposure scenarios involving direct contact with solutions of the chemical are characterized by 

immersion of the hands for 1 minute or wearing contaminated gloves for 1 hour. Generally, it is 

not reasonable to assume or postulate that the hands or any other part of a worker will be 

immersed in a solution of a herbicide for any period of time. On the other hand, contamination 

of gloves or other clothing is quite plausible. For these exposure scenarios, the key element is 

the assumption that wearing gloves grossly contaminated with a chemical solution is equivalent 

to immersing the hands in a solution. In either case, the concentration of the chemical in solution 

that is in contact with the surface of the skin and the resulting dermal absorption rate are 

essentially constant. 

For both scenarios (the hand immersion and the contaminated glove), the assumption of 

zero-order absorption kinetics is appropriate. Following the general recommendations of U.S. 

EPA/ORD (1992), Fick's first law is used to estimate dermal exposure. As discussed in Section 

3.1.3, an experimental dermal permeability coefficient (Kp) for tebufenozide is not available. 

Thus, the Kp for tebufenozide is estimated using the algorithm from U.S. EPA (1992a). 

Exposure scenarios involving chemical spills onto the skin are characterized by a spill on to the 

lower legs as well as a spill on to the hands. In these scenarios, it is assumed that a solution of 

the chemical is spilled on to a given surface area of skin and that a certain amount of the 

chemical adheres to the skin. The absorbed dose is then calculated as the product of the amount 

of the chemical on the surface of the skin (i.e., the amount of liquid per unit surface area 

multiplied by the surface area of the skin over which the spill occurs and the concentration of the 

chemical in the liquid), the first-order absorption rate, and the duration of exposure. 

For both scenarios, it is assumed that the contaminated skin is effectively cleaned after 1 hour. 

As with the exposure assessments based on Fick's first law, this product (mg of absorbed dose) is 

divided by body weight (kg) to yield an estimated dose in units of mg chemical/kg body weight. 

3-13
 
3.2.3. General Public. 

3.2.3.1. General Considerations – Although some applications of tebufenozide may be made in 

relatively remote areas involving limited exposure to the general public, both aerial and ground 

applications may be made in residential areas. In residential applications, members of the 

general public are likely to be exposed to tebufenozide. Any number of exposure scenarios can 

be constructed for the general public, depending on various assumptions regarding application 

rates, dispersion, canopy interception, and human activity. Several scenarios are developed for 

this risk assessment which should tend to over-estimate exposures in general. 

The two types of exposure scenarios developed for the general public include acute exposure and 

longer-term or chronic exposure. All of the acute exposure scenarios are primarily accidental. 

They assume that an individual is exposed to the compound either during or shortly after its 

application. Specific scenarios are developed for direct spray, dermal contact with contaminated 

vegetation, as well as the consumption of contaminated fruit, water, and fish. Most of these 

scenarios should be regarded as extreme, some to the point of limited plausibility. The 

longer-term or chronic exposure scenarios parallel the acute exposure scenarios for the 

consumption of contaminated fruit, water, and fish but are based on estimated levels of exposure 

for longer periods after application. 

The exposure scenarios developed for the general public are summarized in Worksheet E03. As 

with the worker exposure scenarios, details of the assumptions and calculations involved in these 

exposure assessments are given in the worksheets that accompany this risk assessment 

(Worksheets D01a to D09b). The remainder of this section focuses on a qualitative description 

of the rationale for and quality of the data supporting each of the assessments. 

3.2.3.2. Direct Spray – Direct sprays involving ground applications are modeled in a manner 

similar to accidental spills for workers (Section 3.2.2.2). In other words, it is assumed that the 

individual is sprayed with a solution containing the compound and that an amount of the 

compound remains on the skin and is absorbed by first-order kinetics. For these exposure 

scenarios, it is assumed that during a ground application, a naked child is sprayed directly with 

tebufenozide. These scenarios also assume that the child is completely covered with 

tebufenozide (that is, 100% of the surface area of the body is exposed and contaminated). These 

exposure scenarios are likely to represent upper limits of plausible exposure. An additional set of 

scenarios are included involving a young woman who is accidentally sprayed over the feet and 

legs. For each of these scenarios, some assumptions are made regarding the surface area of the 

skin and body weight. These are detailed in Worksheets B05, B06, and B07, for an adult male, 

and adult female, and a young child, respectively. 

3.2.3.3. Dermal Exposure from Contaminated Vegetation – In this exposure scenario, it is 

assumed that the herbicide is sprayed at a given application rate and that an individual comes in 

contact with sprayed vegetation or other contaminated surfaces at some period after the spray 

operation. For these exposure scenarios, some estimates of dislodgeable residue and the rate of 

transfer from the contaminated vegetation to the surface of the skin must be available. No such 

3-14
 
data are available on dermal transfer rates for tebufenozide and the estimation methods of Durkin 

et al. (1995) are used as defined in Worksheet D02. The exposure scenario assumes a contact 

period of one hour and assumes that the chemical is not effectively removed by washing until 24 

hours after exposure. Other estimates used in this exposure scenario involve estimates of body 

weight, skin surface area, and first-order dermal absorption rates, as discussed in the previous 

section. 

3.2.3.4. Contaminated Water – Water can be contaminated from runoff, as a result of leaching 

from contaminated soil, from a direct spill, or from unintentional contamination from aerial 

applications. For this risk assessment, three exposure scenarios are considered for the acute 

consumption of contaminated water: an accidental spill into a small pond (0.25 acres in surface 

area and 1 meter deep), accidental direct spray of or incidental drift into a pond and stream, and 

the contamination of a small stream and pond by runoff or percolation. In addition, longer-term 

estimates of concentrations in water are based on a combination of modeling and monitoring 

data. Each of these scenarios are considered in the following subsections. 

3.2.3.4.1. Accidental Spill – The accidental spill scenario assumes that a young child 

consumes contaminated water shortly after an accidental spill into a small pond. The specifics of 

this scenarios are given in Worksheet D05. Because this scenario is based on the assumption that 

exposure occurs shortly after the spill, no dissipation or degradation of tebufenozide is 

considered. This scenario is dominated by arbitrary variability and the specific assumptions used 

will generally overestimate exposure. The actual concentrations in the water would depend 

heavily on the amount of compound spilled, the size of the water body into which it is spilled, the 

time at which water consumption occurs relative to the time of the spill, and the amount of 

contaminated water that is consumed. Based on the spill scenario used in this risk assessment, 

the concentration of tebufenozide in a small pond is estimated to range from about 0.22 mg/L to 

11 mg/L with a central estimate of about 2.2 mg/L (Worksheet D05). This is and is intended to 

be an extreme accidental exposure scenario. The purpose of this scenario is simply to suggest the 

intensity of measures that would need to be taken in the event of a relatively large spill of 

tebufenozide into a relatively small body of water. 

3.2.3.4.2. Accidental Direct Spray/drift for a Pond or Stream – These scenarios are less 

severe but more plausible than the accidental spill scenario described above. The U.S. EPA 

typically uses a two meter deep pond to develop exposure assessments (SERA 2004b). If such a 

pond is directly sprayed with tebufenozide at the nominal application rate of 0.12 lb/acre, the 

peak concentration in the pond would be about 0.0067 mg/L, equivalent to 6.7 μg/L or 6.7 ppb 

(Worksheet D10a). This concentration is a factor of about 325 below central estimate of the peak 

concentration of 2.2 mg/L after the accidental spill (Worksheet D05). Because the USDA will 

not directly spray open bodies of water, the concentration of 0.0067 mg/L from direct spray 

would be an accidental exposure. At distances of 100 to 500 feet down wind, estimates of drift 

of tebufenozide from aerial applications would result in water concentrations between about 

0.000015 mg/L (500 feet) to about 0.00013 mg/L (100 feet) (Worksheet D10a). 

3-15
 
Similar calculations can be made for the direct spray of a stream and the resulting water 

concentrations will be dependant on the surface area of the stream that is sprayed and the rate of 

water flow in the stream. The stream modeled using GLEAMS (see below) is about 6 feet wide 

(1.82 meters) and it is assumed that the pesticide is applied along a 1038 foot (316.38 meters) 

length of the stream with a flow rate of 710,000 L/day. An application rate of 0.12 lb/acre, is 

2 2 equivalent to 13.45 mg/m [0.12 lb/acre × 112.1 mg/m per lb/acre]. Thus, a direct spray would 

2 be equivalent to about 7745 mg [1.82 meters × 316.38 meters × 13.45 mg/m ]. The daily average 

concentration in the stream segment would be about 0.011 mg/L [7745 mg ÷ 710,000 L/day]. 

Instantaneous concentrations would, of course, vary remarkably over time during and after drift. 

If the stream were 100 feet downwind of the application site, the drift would be a factor of 

0.0195 of the application rate (Worksheet B23). Thus, the average daily concentration in the 

stream would be about 0.2 μg/L [0.011 mg/L × 0.0195 = 0.00021 mg/L or 0.21 μg/L]. Similar 

calculations for other distances are summarized in Worksheet D10b. 

3.2.3.4.3. Gleams Modeling – For compounds such as tebufenozide, which may be 

applied over a large proportion of a watershed, drift and even direct spray are not the only and 

may not be the greatest source of contamination of surface water. Water contamination may also 

occur from soil runoff or percolation and, depending on local conditions, can lead to substantial 

contamination of ponds or streams. Estimates of these concentrations can be based both on 

modeling and monitoring data. 

Modeling of concentrations in stream water conducted for this risk assessment are based on 

GLEAMS (Groundwater Loading Effects of Agricultural Management Systems) modeling. 

GLEAMS is a root zone model that can be used to examine the fate of chemicals in various types 

of soils under different meteorological and hydrogeological conditions (Knisel and Davis 2000). 

As with many environmental fate and transport models, the input and output files for GLEAMS 

can be complex. The general application of the GLEAMS model and the use of the output from 

this model to estimate concentrations in ambient water are detailed in SERA (2004b). 

For the current risk assessment, the application site was assumed to consist of a 10 hectare square 

area that drained directly into a small pond or stream. The chemical specific values as well as 

the details of the pond and stream scenarios used in the GLEAMS modeling are summarized in 

Table 3-1. The GLEAMS modeling yielded estimates of runoff, sediment and percolation that 

were used to calculate concentrations in the stream adjacent to a treated plot, as detailed in 

Section 6.4 of SERA (2004b). The results of the GLEAMS modeling for the small stream are 

summarized in Table 3-2 and the corresponding values for the small pond are summarized in 

Table 3-3. These estimates are expressed as both average and maximum concentrations in water. 

The top section of each table gives the water contamination rates (WCR) – i.e., the concentration 

of the compound in water in units of ppb (μg/L) normalized for an application rate of 1 lb/acre. 

The bottom section of each table gives the estimated maximum and average concentrations 

adjusted for the two applications spaced three days apart at a rate of 0.12 lb/acre (Section 2.3). 

At the application rate of 0.12 lb/acre, no stream contamination is estimated in very arid regions 

3-16
 
– i.e., annual rainfall of 10 inches of less. At higher rainfall rates, the modeled peak 

concentrations in streams range from about 0.04 μg/L (loam at an annual rainfall rate of 15 

inches) to about 40 μg/L (clay soil at an annual rainfall rate of 150 inches per year) (Table 3-2). 

While not detailed in Table 3-2, the losses from clay are about equally divided between sediment 

loss (about 51%) and runoff loss (about 49%). Water contamination due to percolation is 

-9 negligible (a proportion of about 8×10 ). In sandy soils, however, percolation accounts for 

virtually all of the total loss at an annual rainfall rate of 250 inches. 

Modeled concentrations in a small pond (Table 3-3) are lower than those modeled in the stream. 

As with the stream modeling, no surface water contamination is expected in very arid regions. 

For regions with annual rainfall rates of 15 inches or more, the modeled peak concentrations in 

ponds range from less than 0.006 μg/L (loam) to about 20 μg/L (clay soil at an annual rainfall 

rate of 250 inches per year). The GLEAMS scenarios do not specifically consider the effects of 

accidental direct spray. As discussed above and detailed in Worksheet A04b, direct spray of a 

standard pond could result in peak concentrations of about 6.7 μg/L, somewhat less than the 

20 μg/L peak concentration modeled in ponds. 

3.2.3.4.4. Other Modeling Efforts – A summary of the GLEAMS modeling discussed 

above as well as modeling of tebufenozide conducted for other analyses is given in Table 3-4. In 

addition to GLEAMS, two other water contamination models were used: GENEEC and SciGrow. As discussed in SERA (2004b), these are Tier 1 screening models developed by the U.S. 

EPA that are intended to provide very conservative upper range estimates of concentrations of a 

compound in surface water (GENEEC) and groundwater (Sci-Grow) based on a given 

application rate, number of applications, the interval between applications, and standard 

environmental fate parameters for a specific compound (i.e., a subset of those summarized in 

Table 3-1). 

Estimates of peak concentrations from GENEEC, about 8μg/L, are similar to the central 

estimates from GLEAMS, 5 to 10 μg/L, but are somewhat less than the peak estimates from 

GLEAMS, 20 to 40 μg/L. This suggests that although GENEEC is designed as a very 

conservative model, the application of GLEAMS to the modeling for tebufenozide incorporated 

more extreme scenarios for contamination. As detailed in SERA (2004b), the application of 

GLEAMS is intended to encompass extreme situations which favor high runoff from clay and 

high percolation losses from sand. GENEEC does not provide direct estimates of annual average 

concentration but does provide 90-day average concentrations. Adjusting the GENEEC modeled 

90-day average of 6 μg/L over a one-year period, the concentration of 1.5 μg/L is very close to 

the upper range of the average concentration modeled using GLEAMS – i.e., 1.4 μg/L for the 

pond. Sci-Grow estimates a ground water concentration of about 0.09 μg/L. This is in the lower 

range of the estimates from GLEAMS. This is probably due to the very shallow root zone used 

in the GLEAMS modeling – i.e., 12 inches – compared to the 8 to 25 feet water table depth used 

in Sci-Grow 

(http://www.epa.gov/oppefed1/models/water/scigrow_description.htm#characteristics). 

3-17
 
The only other modeling effort encountered for tebufenozide is the use of PRZM/EXAMS by the 

U.S. EPA (1999e) for the reregistration of tebufenozide. As summarized in Table 3-4, the U.S. 

EPA (1999e) modeled the application of tebufenozide to an apple orchard (6 applications at 0.31 

lb/acre) and to a cotton field (4 applications at 0.25 lb/acre) for a pond. While this modeling 

effort used assumptions and weather data substantially different from the GLEAMS modeling 

(i.e., application rates, soil types, and rainfall patterns), the results are reasonably consistent with 

the above estimates of concentrations in surface waters based on GLEAMS correcting for 

differences in the total amount of tebufenozide applied. In the modeling of applications to cotton 

at a cumulative application rate of 1 lb/acre, for example, the peak concentration estimated by 

U.S. EPA (1999e) is 17 μg/L. The GLEAMS model was run at a cumulative application of 0.24 

lb/acre and the adjusted peak concentration for a pond from U.S. EPA (1999e) would be about 

4 μg/L [17 μg/L × 0.24 = 4.08 μg/L], very close to the central estimate of 5 μg/L modeled using 

GLEAMS. The average annual concentration modeled by U.S. EPA (1999e) was about 8.2 μg/L, 

which would correspond to 2 μg/L [8.2 μg/L × 0.24 = 1.96 μg/L] at an application rate of 0.24 

lb/acre. This is only modestly higher than the peak concentration from GLEAMS of 1.4 μg/L. 

3.2.3.4.5. Monitoring Data – Very little water monitoring data are available on 

tebufenozide. Although the USGS (1998) provides information on the agricultural uses of 

tebufenozide, no monitoring data on tebufenozide are available from the USGS National Water 

Quality Assessment (NAWQA). Sundaram et al. (1996a) published a monitoring study of 

concentrations of tebufenozide in water that might be associated with the application of this 

pesticide in a forest environment. In this study, tebufenozide was aerially applied at a rate of 70 

g/ha (0.07 kg/ha or 0.06244 lb/acre) to a 500 ha boreal forest. Two applications were made at 4 

days apart. Water concentrations were then monitored in a small pond and stream. The pond 

2 3 had a surface area of 500 m and an average depth of 0.6 m for a volume of 300 m or 300,000 L

3 [1,000 L/m ]. Water concentrations were monitored at 1 , 8, and 12 hours after application as 

well as 1, 2, 3, 4, 5, 8, 12, and 24 days after application. 

The peak concentration, 5.31 ppb (0.00531 mg/L) occurred 1 hour after the first application, 

clearly indicating that the water had been directly sprayed. Taking the water volume of 300,000 

L, the amount applied to the pond can be calculated as, 1,593 mg, 

0.00531 mg/L × 300,000 L. 

2 2 The nominal application rate of 0.07 kg/ha is equivalent to 70,000 mg/10,000 m or 7 mg/m . At 

this nominal application rate, the total amount applied to a 500 m2

 pond would be 3500 mg, 

2 2 7 mg/m × 500 m .

Thus, it appears that the initial concentrations of tebufenozide in water are consistent with the 

direct spray of about 50% [1,593 mg/3500 mg = 0.455 .50%] of the pond at the nominal 

application rate. 

3-18
 
3.2.3.4.6. Concentrations of Tebufenozide in Water Used for Risk Assessment – A 

summary of the concentrations of tebufenozide in water that are used for the current risk 

assessment is given in Table 3-5. The upper range of the expected peak concentration of 

tebufenozide in surface water will be taken as 40 μg/L. This is based on the upper range of 

concentrations estimated in streams from the GLEAMS modeling. This concentration also 

encompasses accidental direct sprays of both a small stream and small pond (Table 3-4). In most 

instances, concentrations in surface water are likely to be much lower. At the lower extreme, an 

argument may be made that concentrations of tebufenozide are likely to be essentially zero – i.e., 

applications at sites that are distant from open bodies of water and in areas in which runoff or 

percolation are not likely to occur. For this risk assessment, the lower range of the peak 

concentration in ambient water will be set at 0.005 μg/L. This is in the lower range of non-zero 

concentrations modeled in streams and ponds in relatively arid regions. The central estimate of 

concentration of tebufenozide in surface water will be taken as 10 μg/L. This is the central 

estimate of the concentrations modeled in ponds (Table 3-4). 

Longer term concentrations of tebufenozide in surface water will be much lower than peak 

concentrations. At an application rate of 0.12 lb/acre, the highest longer term concentration will 

be taken as 1.4 μg/L. This is the maximum longer term concentration modeled using GLEAMS 

and is near the maximum longer term concentration given by U.S. EPA (1999e) after adjusting 

for differences in application rate. As with peak concentrations, the lower range of longer term 

concentrations will approach zero. For this risk assessment, the lower range of longer term 

concentrations is taken as 0.002 μg/L, the lowest non-zero value modeled for tebufenozide in 

ponds at the application rate of 0.12 lb/acre. This lower range is somewhat arbitrary but has no 

impact on the risk assessment. The central value for longer term concentrations of tebufenozide 

in water will be taken as 0.5 μg/L. This is the central estimate of the longer term concentrations 

in ponds modeled using GLEAMS and is somewhat higher than the central estimate of the longer 

term concentration in streams (Table 3-4). 

3.2.3.5. Oral Exposure from Contaminated Fish – Many chemicals may be concentrated or 

partitioned from water into the tissues of animals or plants in the water. This process is referred 

to as bioconcentration. Generally, bioconcentration is measured as the ratio of the concentration 

in the organism to the concentration in the water. For example, if the concentration in the 

organism is 5 mg/kg and the concentration in the water is 1 mg/L, the bioconcentration factor 

(BCF) is 5 L/kg [5 mg/kg ÷ 1 mg/L]. As with most absorption processes, bioconcentration 

depends initially on the duration of exposure but eventually reaches steady state. Details 

regarding the relationship of bioconcentration factor to standard pharmacokinetic principles are 

provided in Calabrese and Baldwin (1993). 

The bioconcentration of tebufenozide was determined in fathead minnows (Rhodes and Leak 

1996) and bluegill sunfish (Dong and Hawkins, 1993). In fathead minnows, bioconcentration 

factors (BCF) range from about 17 in pre-spawn adults to greater than 100 in newly fertilized 

embryos (Rhodes and Leak 1996). In bluegills, Dong and Hawkins (1993) provide data on 

bioconcentration in the edible muscle (BCF=7.5) as well as viscera (BCF=106) and whole body 

3-19
 
(BCF=52). For the human health risk assessment, the bioconcentration factor of 7.5 from Dong 

and Hawkins (1993) is used. Taking the value for the edible portion of fish is not the most 

conservative approach but seems the most realistic approach because humans usually clean 

caught fish and consume only the fillet or muscle. For the ecological risk assessment, however, 

the higher BCF value of 52 (whole body) is used. 

For the acute and longer-term exposure scenarios involving the consumption of contaminated 

fish, the water concentrations of tebufenozide used are identical to the concentrations used in the 

contaminated water scenarios (Section 3.2.3.4.6). The acute exposure scenario is based on the 

assumption that an adult angler consumes fish taken from contaminated water shortly after an 

accidental spill of 200 gallons of a field solution into a pond that has an average depth of 1 m and 

2 a surface area of 1000 m or about one-quarter acre. No dissipation or degradation is considered. 

Bioconcentration is a dynamic process and for some compounds time to maximum steady state 

may be prolonged. For tebufenozide, Dong and Hawkins (1993) found that time to steady state 

was reached in about 1-day. Thus, the use of the experimental BCF for the acute accidental 

scenario is not overly conservative. Nonetheless, this scenario may somewhat overestimate 

exposure in that some degradation of tebufenozide could occur during the course of the acute 

spill scenario. 

Because of the available and well-documented information and substantial differences in the 

amount of caught fish consumed by the general public and native American subsistence 

populations (U.S. EPA 1996), separate exposure estimates are made for these two groups, as 

illustrated in Worksheet D08a and D08b. The chronic exposure scenario is constructed in a 

similar way, as detailed in Worksheets D09a and D09b. 

3.2.3.6. Oral Exposure from Contaminated Vegetation – Although Forest Service 

applications of tebufenozide will not involve the intentional treatment of food crops, incidental 

exposure to vegetation that may be consumed by members of the general public is plausible 

during broadcast applications. Any number of scenarios could be developed involving either 

accidental spraying of crops or the spraying of edible wild vegetation, like berries. The exposure 

scenarios developed for this exposure assessment include one scenario for acute exposure, as 

defined in Worksheet D03 and two scenarios for longer-term exposure, as defined in Worksheets 

D04a and D04b. In both acute and longer-term scenarios, the concentration of tebufenozide on 

contaminated vegetation is estimated using the empirical relationships between application rate 

and concentration on vegetation developed by Fletcher et al. (1994) which is in turn based on a 

re-analysis of data from Hoerger and Kenaga (1972). These relationships are defined in 

Worksheet B20. 

For the acute exposure scenario involving only a single application (Worksheet D03a), the 

estimated residue level is taken as the product of the application rate and the residue rate for 

contaminated fruit. For multiple applications, the peak concentration on fruit or other vegetation 

will occur immediately after the last application. This concentration can be calculated based on 

3-20
 
the initial concentration after the first application (C0), the number of applications (n), and the 

first-order decay coefficient (k), which can be calculated from the halftime (t50) [k=ln(2)÷t50]. 

Assuming a first-order decrease in concentrations in contaminated vegetation, the concentration 

in the vegetation at time t after the first application (Ct), can be calculated as: 

- t Ct = C0 × e k (Eq. 3-1) 

Using the plateau principle (e.g., Goldstein et al. 1974, p. 321) and defining )t as the interval -k ) t between applications and e as p to simplify notation, the concentration immediately after the 

nth application (Cn) can be calculated as: 

n Cn = C0 × (1- p ) ÷ (1- p). (Eq. 3-2) 

This algorithm is used in Worksheet D03b to calculate the maximum concentration on vegetation 

after multiple applications at the specified interval. 

For the longer-term exposure scenario (Worksheets D04a and D04b), a duration of 90 days is 

used. Although the duration of exposure of 90 days is somewhat arbitrarily, this duration is 

intended to represent the consumption of contaminated fruit that might be available over one 

season. Longer durations could be used for certain kinds of vegetation but would lower the 

estimated dose (i.e., would reduce the estimate of risk). 

The reported halftimes on vegetation are highly variable (Table 2-1), ranging from 2.8 days, the 

lower value of the range reported by Hawkins (1998) to 58.7 days, the upper value of the range 

reported by Sundaram et al. (1996a). This substantial variability is not uncommon in field 

measurements of halftimes of vegetation, which are substantially impacted by site and situational 

differences such as rainfall, temperature, wind velocity, and the type of vegetation. For this risk 

assessment, the range of vegetation halftimes will be taken as 3 to 60 days (the approximate 

range summarized in Table 2-1) and the central estimate will be taken as 13.4 days, the geometric 

mean of this range. 

For the longer-term exposure scenarios, the time-weighted average concentration on fruit is 

calculated from the equation for first-order dissipation. Assuming a first-order decrease in 

concentrations in contaminated vegetation, the concentration in the vegetation at time t after 

spray, C , can be calculated based on the initial concentration, C , as: t 0

C = C × e-kt 

t 0

where k is the first-order decay coefficient which can be calculated from the halftime (t ) 50

[k=ln(2)÷t 50]. For a single application, the time-weighted average concentration (C TWA) over time 

t can be calculated as the integral of Ct

 (De Sapio 1976, p. p. 97 ff) divided by the duration (t): 

-k t C = C (1 - e ) ÷ (k t). TWA 0 

3-21
 
This equation is used to estimate the time-weighted average concentration on vegetation after a 

single applications (Worksheet D04a). 

For two applications, such as those modeled in this risk assessment, the expression of the 

time-weighted average concentration is somewhat more complicated. Defining exp(x) as ex , 

where x is any number, the time-weighted average concentration over a period from the day of 

application to time t with a second application occurring on day t (where t # t ) is: 2 1 1 2
 

C = ( C (1-exp(-kt )) + [ {C + C exp(-kt )} × {1-exp(-k [t - t ])}] ) ÷ (k t ) TWA 0 1 0 0 1 2 1 2
 

This equation is used to estimate the time-weighted average concentration on vegetation after a 

single applications (Worksheet D04b). 

3-22
 
3.3. DOSE-RESPONSE ASSESSMENT 

3.3.1. Overview 

Acute and chronic risk values are derived for tebufenozide. Following standard practices for 

USDA risk assessments, risk assessment values available from U.S. EPA are adopted directly 

unless there is a compelling basis for doing otherwise. When risk values are not available from 

U.S. EPA, the methods used by U.S. EPA are employed to derive surrogate values. 

U.S. EPA has derived a chronic RfD for tebufenozide of 0.018 mg/kg/day. This chronic RfD is 

well-documented and is used directly for all longer term exposures to tebufenozide. This value is 

based on a NOAEL of 1.8 mg/kg/day in dogs and an uncertainty factor of 100 – two factors of 10 

for interspecies and intraspecies variability. Because of the low acute toxicity of tebufenozide, 

the U.S. EPA has not derived an acute RfD but has identified an acute NOAEL of 1000 

mg/kg/day from reproduction studies in both rats and rabbits involving 10 to 13 day exposure 

periods. This NOAEL is the basis for a surrogate acute RfD of 10 mg/kg using an uncertainty 

factor of 100 as in the chronic RfD. This surrogate acute RfD is applied to all incidental or 

accidental exposures that involve an exposure period of 1 day. 

3.3.2. Chronic RfD 

The most recent RfD for tebufenozide is 0.018 mg/kg/day, a value derived by the U.S. EPA’s 

Office of Pesticide Programs (U.S. EPA 1999b,e). This compound is not listed on the U.S. 

EPA’s agency-wide list of approved RfDs (i.e., IRIS) (U.S. EPA 2004). As noted in section 

3.1.2 and detailed in Appendix 2, the most sensitive endpoint for tebufenozide is hematological 

effects including methemoglobin formation and several other endpoints that are characteristic of 

hemolytic anemia. These effects were observed in mice, rats, and dogs, with the dog being the 

most sensitive species tested with tebufenozide. As reviewed by Calabrese (1991), this pattern is 

consistent with known differences in methemoglobin reductase activity which suggest that the cat 

may be the most sensitive species, followed by humans (half as susceptible as cats), dogs (half as 

susceptible as human), and rats (about one-tenth as susceptible as humans). 

The RfD derived by the U.S. EPA (1999b) is based on a study by Richards (1992a,b) in which a 

dietary concentration of 0, 15, 50, 250, or 1500 ppm technical grade tebufenozide was provided 

to male and female beagles for 52 weeks (Appendix 2). In the 250 and 1500 ppm groups, the 

primary hematological effects were increased concentrations of methemoglobin. The increases 

in methemoglobin concentrations were associated with increased breakdown of red blood cells in 

the liver and spleen, and decreases in red blood cell counts, hemoglobin concentrations, and 

packed red cell volume, along with several other associated hematological effects. None of these 

effects were observed in beagles exposed to a dietary concentration of 50 ppm technical grade 

tebufenozide, which corresponded to a daily dose of 1.5-2.4 mg/kg bw (based on measured food 

consumption). Taking 1.8 mg/kg bw/day as a central estimate of the NOAEL, the U.S. EPA 

(1999b) applied an uncertainty factor of 100, two factors of 10 for interspecies and intraspecies 

variability, to arrive at the chronic RfD of 0.018 mg/kg/day. 

Under the Food Quality Protection Act (FQPA), the U.S. EPA is required to consider an 

3-23
 
additional uncertainty factor of 10 for the protection of infants and children. For tebufenozide, 

the U.S. EPA (1999b) determined that the additional uncertainty factor is not required because of 

the information indicating that tebufenozide does not have developmental or reproductive effects 

at doses below those associated with hematological effects. Hence, because the RfD should 

protect against hematological effects, it should also protect against developmental or 

reproductive effects. As discussed in Section 3.4.4, infants less than three months old have lower 

levels of methemoglobin reductase than older children or adults and may be more sensitive 

tebufenozide and other agents that cause methemoglobinemia. While it may be argued that an 

uncertainty factor for very young children might be appropriate, this would not have an impact on 

the risk characterization because of the very low hazard quotients associated with various 

exposure scenarios for tebufenozide (Section 3.4.3). 

3.3.3. Acute RfD 

The U.S. EPA (1999b) considers the acute and intermediate risk from acute or intermediate 

exposure to tebufenozide negligible and does not propose short-term or intermediate-term criteria 

for exposure to tebufenozide. Specifically, the U.S. EPA (1999b) made the following judgement: 

1. Acute toxicity. Toxicity observed in oral toxicity studies were 

not attributable to a single dose (exposure). No neuro or systemic 

toxicity was observed in rats given a single oral administration of 

tebufenozide at 0, 500, 1,000, or 2,000 mg/kg. No maternal or 

developmental toxicity was observed following oral administration 

of tebufenozide at 1,000 mg/kg/day (Limit-Dose) during gestation 

to pregnant rats or rabbits. Thus, the risk from acute exposure is 

considered negligible. 

2. Short- and intermediate-term toxicity. No dermal or systemic 

toxicity was seen in rats receiving 15 repeated dermal applications 

of the technical (97.2%) product at 1,000 mg/kg/day (Limit-Dose) 

as well as a formulated (23% a.i.) product at 0, 62.5, 250, or 

1,000 mg/kg/day over a 21-day period. The Agency noted that in 

spite of the hematological effects seen in the dog study, similar 

effects were not seen in the rats receiving the compound via the 

dermal route indicating poor dermal absorption. Also, no 

developmental endpoints of concern were evident due to the lack 

of developmental toxicity in either rat or rabbit studies. This risk 

is considered to be negligible. -- U.S. EPA (1999b). 

In paragraph 1 above, the acute toxicity study with a single-dose NOAEL of 2000 mg/kg appears 

to refer to the study by Swenson et al. (1994) and the NOAEL of 1000 mg/kg/day for maternal 

toxicity and reproductive effects in rats and rabbits appears to refer to the studies by Hoberman 

(1991) and Swenson and Solomon (1992), respectively. In paragraph 2 above, the U.S. EPA 

(1999b) refers to a dermal study with a NOAEL of 1000 mg/kg/day. In this study, tebufenozide 

3-24
 
was applied 5 days per week for three weeks – i.e., 15 exposures over a 21 day period. Two 

repeated dermal dose studies have been identified with a NOAEL of 1000 mg/kg/day (Hazleton 

and Quinn 1995b; Morrison et al. 1993). As summarized in Appendix 3, both of these studies 

report exposure periods of 4 weeks rather than 3 weeks. 

While the decision of the U.S. EPA (1999b) to classify acute and short-term risks associated with 

tebufenozide appears reasonable, the failure of the U.S. EPA (1999b) to derive an acute RfD 

limits the ability to quantitatively characterize risks associated with acute exposures. As detailed 

in Section 3.2, the current risk assessment is concerned with characterizing the risks of several 

acute exposure scenarios. In addition, the current risk assessment is part of a series of risk 

assessments on different agents used to control the gypsy moth the estimates of risks from the 

various agents will be compared in a companion document. 

Consequently, this risk assessment will use a surrogate acute RfD. Typically, the U.S. EPA will 

base acute RfDs on reproduction studies, specifically teratology studies that involve multiple 

daily gavage doses to pregnant animals. For the current risk assessment, the NOAEL of 1000 

mg/kg/day in pregnant rats and rabbits identified by U.S. EPA (1999b) will be used As detailed 

in Appendix 2, the NOAEL in rabbits is from a study (Swenson and Solomon 1992) in which 

animals were dosed on Days 7-19 of gestation – i.e., repeated exposures over 13 days – and the 

NOAEL in rats is from a study (Hoberman 1991) in which animals were dosed on Days 6-15 of 

gestation – i.e., repeated exposures over 10 days. Dividing this NOAEL by an uncertainty factor 

of 100, identical to that used by U.S. EPA (1999b) in the chronic RfD, yields a surrogate acute 

RfD of 10 mg/kg/day. This value is used to characterize risks associated to incidents or accidents 

that involve an exposure period of 1 day. 

3-25
 
3.4. RISK CHARACTERIZATION 

3.4.1. Overview 

At the maximum application rate considered in this risk assessment, two applications at 0.12 

lb/acre spaced three day apart, there is little indication that adverse effects on human health are 

likely. Based on central estimates of exposure – those that might be considered typical and 

expected – hazard quotients including workers and members of the general public range from 

0.00003 to 0.03, below a level of concern by factors of about 30 to 33,000. At the upper range of 

plausible exposures, the hazard quotient for ground spray workers reaches a level of concern – 

i.e., a hazard quotient of 1. For members of the general public, the upper range of exposure leads 

to a hazard quotient of 1.5 for the longer-term consumption of contaminated vegetation for two 

applications at 0.12 lb/acre. Because of the linear relationship between exposure and application 

rate, two applications at 0.08 lb/acre would reach but not exceed a level of concern. With a 

single application at the maximum rate of 0.12 lb/acre, the hazard index is 0.8, below the level of 

concern. While the longer-term consumption of contaminated vegetation is probably not a likely 

scenario, it is a standard exposure scenario used in Forest Service risk assessments to consider 

the longer-term consumption of food items such as berries that might be sprayed during the 

broadcast application of a pesticide. This risk assessment suggests that two applications at 0.08 

lb/acre or more should be avoided in areas where members of the general public might consume 

contaminated fruits or other contaminated vegetation. 

3.4.2. Workers 

A quantitative summary of the risk characterization for workers is presented in Worksheet E02 

(Supplement 1). The quantitative risk characterization is expressed as the hazard quotient, which 

is the ratio of the estimated exposure from Worksheet E01 to the RfD. For acute 

accidental/incidental exposures, the surrogate acute RfD of 10 mg/kg is used (Section 3.3.3). For 

longer term general exposures – i.e., exposures that could occur over the course of several days, 

weeks, or months during an application season – the chronic RfD of 0.018 mg/kg/day is used 

(Section 3.3.2). 

At the maximum application rate considered in this risk assessment, 0.12 lb/acre, none of the 

acute hazard quotients exceed a level of concern – i.e., a hazard quotient of 1. The highest acute 

hazard quotient is 0.4, associated with wearing contaminated gloves for 1 hour. It should be 

noted, however, that the magnitude of the hazard quotient is linearly related to the duration of 

exposure. The 1-hour exposure period is simply a convention that is uniformly used in Forest 

Service risk assessments (SERA 2001). For tebufenozide, the estimated exposure would exceed 

the acute RfD – i.e., result in a hazard quotient greater than 1 – if a worker were to wear 

contaminated gloves for a period greater than 2.5 hours. Thus, the exposure involving 

contaminated gloves is of greatest concern and this concern would apply to wearing any clothing 

that is saturated with tebufenozide. 

For longer-term exposures, the highest hazard quotient is 1.008 and is associated with the upper 

range of exposure for ground spray workers at the maximum application rate of 0.12 lb/acre. In 

Worksheet E02, this value is presented as 1.0 – i.e., rounded to one significant place after the 

3-26
 
 

decimal. This very minor exceedence of the chronic RfD is interpreted as a hazard quotient of 

1.0 – i.e., the level of concern is not exceeded. All of the other hazard quotients are below a 

level of concern by a factor of at least 2 at the upper range of exposures and a factor of at least 10 

at the central estimates of exposure. It should be noted that multiple applications of 

tebufenozide, such as those covered in this risk assessment, have no effect on the hazard 

quotients for workers. This is because all worker exposure assessments are based on the 

assumption that the worker applies the compound daily, albeit at different sites, over the course 

of an application season. 

Mimic can cause eye irritation (section 3.1.11). Quantitative risk assessments for irritation are 

not derived; however, from a practical perspective, eye irritation is likely to be the only overt 

effect as a consequence of mishandling tebufenozide. This effect can be minimized or avoided 

by prudent industrial hygiene practices during the handling of the compound. 

3.4.3. General Public 

A quantitative summary of the risk characterization for members of the general public is 

presented in Worksheet E04 (Supplement 1). With the exception of the scenarios for the longerterm consumption of contaminated vegetation, all exposure scenarios are based on the highest 

application considered in this risk assessment – i.e., two applications at a rate of 0.12 lb/acre with 

an interval of 3 days between applications. Two scenarios are conducted for the longer-term 

consumption of contaminated vegetation, one involving two applications spaced three days apart 

and the other involving only a single application. Both are modeled at the maximum rate of 0.12 

lb/acre. As with the risk characterization for workers, risk is expressed quantitatively as the 

hazard quotient using the surrogate acute RfD of 10 mg/kg (Section 3.3.3) for acute exposures 

and the chronic RfD of 0.018 mg/kg/day (Section 3.3.2) for longer-term exposures. 

The only exposure scenario that leads to any unacceptable risk is the longer-tern consumption of 

contaminated vegetation. For two applications spaced three days apart at the maximum rate of 

0.12 lb/acre, the hazard quotient 1.5 for the longer-term consumption of contaminated vegetation 

– i.e., the exposure exceeds the RfD by a factor of 1.5. Because the exposure is linearly related 

to the application rate, two exposures at an application rate of 0.08 lb/acre [0.12 lb/acre ÷ 1.5] 

would reach but not exceed the level of concern. With a single application at the maximum rate 

of 0.12 lb/acre, the hazard index is 0.8, below the level of concern. As discussed in Section 

3.2.3.6, this exposure scenario assumes that an individual will consume over a 90 day period 

after that fruit had been directly sprayed. The probability of this occurring is unlikely because the 

USDA will not intentionally apply tebufenozide to crops or other food items. Nonetheless, this 

is a standard exposure scenario used in Forest Service risk assessments to consider the longerterm consumption of food items such as berries that might be sprayed during the broadcast 

application of a pesticide. This risk assessment suggests that two applications at 0.08 lb/acre or 

more should be avoided in areas where members of the general public might consume 

contaminated fruits. 

None of the acute or other longer-term hazard quotients exceed 1 even at the upper ranges of 

3-27
 
plausible exposure. The highest acute hazard quotient is 0.1, the upper range of risk for the 

consumption of contaminated water by child after an accidental spill. This extreme and 

accidental acute scenario is below the level of concern by a factor of 10. No other acute exposure 

scenarios, many of which involve extremely conservative assumptions, approach a level of 

concern at the upper range of exposure. Based on central estimates of exposure, which involve 

somewhat less conservative assumptions, the acute hazard quotients range from 0.00008 to 0.02 

– i.e., below the level of concern by factors of 50 to 12,500. Based on central estimates of 

longer-term exposures, the hazard quotients range from 0.00003 to 0.03, below the level of 

concern by factors of about 30 to over 33,000. 

3.4.4. Sensitive Subgroups 

Some individuals are born with a form of congenital methemoglobinemia and may be at 

increased risk of adverse effects to compounds that induce methemoglobinemia (Centa et al. 

1985; Das Gupta et al. 1980). Infants less than 3 months old have lower levels of 

methemoglobin (cytochrome b5) reductase and higher levels of methemoglobin (1.32%), 

compared with older children or adults (Centa et al. 1985; Smith 1996). A similar pattern is seen 

in many species of mammals (Lo and Agar 1986). Thus, it is possible that infants could be more 

sensitive to the effects of tebufenozide than adults. 

3.4.5. Connected Actions 

The most sensitive effect for tebufenozide, methemoglobinemia, is also associated with 

exposures to diflubenzuron, another agent used for gypsy moth control. These two agents are 

likely to have an additive effect on methemoglobinemia but these agents are not used together. 

Thus, simultaneous exposures are unlikely. Exposure to other compounds in the environment 

that induce methemoglobinemia may also lead to an additive effect. Any agent or condition that 

may reduce the oxygen carrying capacity of the blood could lead to increased risks from exposure 

to either tebufenozide or diflubenzuron. For example, individuals exposed to combustion smoke 

or carbon monoxide (that is, agents that do oxidative damage to blood) may be at increased risk 

of developing methemoglobinemia (Hoffman and Sauter 1989; Laney and Hoffman 1992). In 

addition, individuals exposed to high levels of nitrates, either in air or in water, will have 

increased levels of methemoglobin (Woebkenberg et al. 1981) and may be at increased risks of 

exposure to compounds such as tebufenozide. 

3.4.6. Cumulative Effects 

This risk assessment is based on two applications at the maximum allowable rate of 0.12 lb/acre. 

This approach is used to estimate maximum daily exposure and daily absorbed dose. In addition, 

this risk assessment specifically considers the effect of repeated exposure in that the chronic RfD 

is used as an index of acceptable longer-term exposures and an acute RfD based on an exposure 

period of 10 to 13 days is used for the risk characterization of single day exposures. 

Consequently, the risk characterizations presented in this risk assessment specifically addresses 

and encompasses the potential impact of long-term exposure and cumulative effects. 

3-28
 
4. ECOLOGICAL RISK ASSESSMENT
 

4.1. HAZARD IDENTIFICATION 

4.1.1. Overview. The toxicity of tebufenozide is well characterized in experimental mammals, 

birds, terrestrial invertebrates, and aquatic animals. Nonetheless, given the very large number of 

species in the environment which could be exposed to tebufenozide, toxicity data are available 

on relatively few species. 

It seems reasonable to assume the most sensitive effects in wildlife mammalian species will be 

the same as those in experimental mammals (i.e., effects on the blood, specifically the formation 

of methemoglobin, which leads to a spectrum of other effects in blood that can be characterized 

as hemolytic anemia). At higher doses, tebufenozide was associated with impaired reproductive 

performance in experimental mammals, and this effect is also considered quantitatively in this 

risk assessment. Potential reproductive effects are also of concern for birds, although there are 

inconsistencies in the available experimental data. The available literature includes a 

reproduction study investigating effects in mallard ducks and two reproduction studies 

investigating effects in bobwhite quail. In one of the quail studies, dietary concentrations of 300 

and 1000 ppm caused reproductive effects. These effects were not observed in that study at 100 

ppm or in the more recent quail study or in the study on mallard ducks. A field study on the 

effects of tebufenozide on reproductive performance in birds noted trends that were statistically 

insignificant but suggestive of adverse reproductive effects in a warbler species. Thus, consistent 

with the interpretation by the U.S. EPA, reproductive effects in both mammals and birds are 

considered endpoints of concern in this risk assessment. 

The mechanism of action of tebufenozide in target insects is relatively well understood. 

Tebufenozide mimics the action of the invertebrate hormone, 20-hydroxyecdysone, which 

controls molting. The effectiveness of tebufenozide in mimicking 20-hydroxyecdysone activity, 

however, appears to vary markedly among orders and species of invertebrates. In general, 

lepidopteran species are sensitive to tebufenozide but other insects are much less sensitive. 

There are no bioassays regarding the toxicity of tebufenozide to terrestrial plants or terrestrial 

microorganisms in the literature. There are a number of field studies and field simulation studies 

available on tebufenozide and effects that might be associated with toxicity to plants or soil 

microorganisms have not been noted. 

The acute toxicity of tebufenozide to aquatic animals is relatively low, with acute LC50 values 

ranging from 2.2 to 6.5 mg/L for fish and 0.3 to 3.8 mg/L for aquatic invertebrates. Nonetheless, 

much lower concentrations of tebufenozide may cause reproductive effects in fish (0.048 mg/L) 

and aquatic invertebrates (0.0053 mg/L). 

4-1
 
4.1.2. Toxicity to Terrestrial Organisms. 

4.1.2.1. Mammals– As summarized in the human health risk assessment (see Section 3.1), the 

mode of action of tebufenozide in mammals is relatively well characterized. Several standard 

toxicity studies in experimental mammals were conducted as part of the registration process 

(Appendix 2). The most sensitive effect in several species of experimental mammals involves 

effects on the blood, specifically the formation of methemoglobin, which leads to a spectrum of 

other effects in blood that can be characterized as hemolytic anemia. Since higher doses of 

tebufenozide were associated with impaired reproductive performance (see Section 3.1.4), both 

toxic and reproductive effects are considered in this risk assessment. 

The acute toxicity of tebufenozide is relatively low, with an oral LD50 greater than 5000 mg/kg. 

The subchronic and chronic toxicity studies on tebufenozide were conducted in dogs, mice, and 

rats. The most sensitive effects involve changes to blood. The most sensitive species is the dog, 

with a NOAEL of 50 ppm in the diet (1.8 mg/kg bw/day) and an effect level of 500 ppm (about 

20 mg/kg bw/day ) over an exposure period of 1 year. 

As discussed in Section 3.3.3, there is no apparent dose duration relationship for tebufenozide. 

In other words, short-term exposures are likely to lead to changes in the blood comparable to 

those observed after longer-term exposures. Thus, the chronic NOAEL of 1.8 mg/kg/day is used 

to characterize risks associated with both short- and long-term exposures. 

4.1.2.2. Birds– Toxicity studies have been conducted on the acute toxicity and reproductive 

effects of tebufenozide in birds and a field study is available on reproductive effects. 

Information regarding the laboratory tests on the toxicity of tebufenozide to birds is summarized 

in Appendix 4. The acute toxicity of tebufenozide is low for birds, as it is for mammals. When 

administered in gelatin capsules, the 21-day oral LD50 is greater than 2150 mg a.i./kg bw 

(Fletcher 1987). Similarly, in 5-day dietary studies, the dietary LC50 is greater than 5000 ppm 

(Fletcher 1990a,b). Hematological endpoints are not usually assayed in bioassays with birds, 

and there are no data regarding the hematological effects in birds after exposure to tebufenozide. 

Nevertheless, the most relevant and significant studies for this risk assessment involve the 

potential reproductive effects in birds exposed to tebufenozide. Reproduction studies were 

conducted in mallard ducks (Beavers et al. 1993a) and bobwhite quail (Beavers et al. 1993b; 

Reinert 1995a). As indicated in Appendix 4, dietary concentrations less than or equal to 1000 

ppm tebufenozide did not cause reproductive effects in mallard ducks In the quail studies, 

however, the results are inconsistent. In the earlier study by Beavers et al. (1993b), reproductive 

effects - including a reduced number of eggs laid, viable embryos and 14 day old survivors - were 

noted at dietary concentrations of 300 and 1000 ppm, but not at 100 ppm. In a similar study 

conducted later by Reinert (1995a), there were no substantial dose-related effects in quail 

exposed to dietary concentrations of up to 615 ppm. 

In terms of the hazard identification, the most important question involves the extent to which 

4-2
 
 

the Reinert (1995a) study reporting negative results for reproductive toxicity reduces the 

concerns raised by the Beavers et al. (1993b) study, which reports positive results. The earlier 

study was accepted by the U.S. EPA (1999e) and used in their ecological risk assessment of 

tebufenozide; however, the U.S. EPA (1999e) does not discuss the later negative study. The 

negative study is discussed in a review by Rohm and Haas (Keller and Brown 1998b), who 

question whether the NOAEL for the earlier study was 100 ppm or 300 ppm. 

Regardless of which dose is classified as a NOAEL in the Beavers et al. (1993b) study, there 

seems to be no evidence that the study is flawed in any way. The minor differences between the 

early study and the later study, as detailed in Appendix 4, relate primarily to how exposures were 

reported and how food consumption was measured. 

Notably, reproductive effects were observed also in mammals exposed to a dietary concentration 

of 2000 ppm (.160 mg/kg bw), with a NOAEL of 150 ppm (.12 mg/kg bw) (see Section 3.1.4). 

In the bobwhite quail study conducted by Beavers et al. (1993b), the dietary effect levels (AELs) 

of 300 and 1000 ppm correspond to estimated daily doses of 45 and 150 mg/kg/day, and the 

NOAEL of 100 ppm corresponds to an estimated daily dose of 15 mg/kg bw. Thus, the apparent 

NOAEL values and AEL values for mammals and birds are reasonably consistent. Finally, 

based on a metabolism study in hens (Sharma and Schuck 1996), the metabolic pathways for 

birds and mammals appear to be similar. 

In the absence of any basis for discounting the earlier study in bobwhite quail (Beavers et al. 

1993b) and given the reasonable consistency in dose levels associated with reproductive effects 

in mammals and birds as well as the similar metabolic pathways in mammals and birds, 

reproductive effects are considered an endpoint of concern in this risk assessment. 

A field study on the reproductive performance of Tennessee warblers (Vermivora peregrina) in 

forests treated with Mimic has been published (Holmes 1998). In this study, Mimic was applied 

at a rate of 0.07 a.i. kg/ha, approximately 0.06 lb a.i./acre, in a forest area in Ontario. Two 

applications were made at this rate with a 4 day interval between applications. A number of 

reproductive parameters were assayed including number of eggs laid, percent hatch and growth 

of the hatchlings. There parameters were compared to an untreated control plot. A total of six 

nests were observed in the control plot and 5 nests in the plot treated with Mimic. No 

statistically significant adverse effects were noted. However, there were decreases in both the 

average number of eggs per nest (6.3 in the control area and 5.8 in the treated area) as well as the 

percent hatch (97.4% in the control area and 89.7% in the treated area). As noted by Holmes 

(1998, p. 191), the small sample sizes result in a low statistical power and the results are 

“suggestive, although not necessarily compelling, that reproductive parameters were 

consistently lower in the treated blocks than in the control block.” Some differences in adult 

behavior were observed in the plot treated with Mimic – i.e., an increase in foraging time and an 

associated decrease in brooding time. This suggests that the primary effect on the birds may have 

been a decrease in food abundance. 

4-3
 
This field study by Holmes (1998) combined with bobwhite quail assay conducted by Beavers et 

al. (1993b) raise concern that tebufenozide could cause adverse reproductive effects in birds. 

This concern is addressed quantitatively in this risk assessment for exposures involving the 

consumption of contaminated vegetation, fish, and insects. 

4.1.2.3. Terrestrial Invertebrates – While Mimic is specifically used by the Forest Service for 

the control of the Gypsy moth, tebufenozide is effective in the control of other lepidopteran pest 

species, including the apple bud moth (Platynota idaeusalis, Biddinger et al. 1998), various 

species of spruce budworm (Cadogan et al. 1997; Payne et al. 1997; Retnakaran et al. 1997a,b), 

the tomato looper (Deixis chalcites, Smagghe et al. 1997), and the Indian-meal moth (Plodia 

interpunctella) (Oberlander et al. 1998). A complete list of the pest species for which 

tebufenozide is specified is provided in U.S. EPA (1999e). 

The toxicity of tebufenozide has been assayed in several species (Appendix 5). The mechanism 

of action of tebufenozide in target insects is relatively well understood. In sensitive species, 

tebufenozide mimics the action of the invertebrate hormone 20-hydroxyecdysone. This hormone 

controls molting in insects and various terrestrial and aquatic invertebrates, which is mediated 

through binding to species-specific ecdysone receptors present in the cytoplasm of epidermal 

cells (Addison 1996; Keller 1998; Smagghe and Degheele 1994a; U.S. EPA 1999e). 

While 20-hydroxyecdysone is a hormone common to many invertebrates, the effectiveness of 

tebufenozide in mimicking 20-hydroxyecdysone activity seems to vary markedly among orders 

and species of invertebrates. Although the specificity of tebufenozide is not addressed in detail 

in the recent U.S. EPA (1999e) ecological risk assessment, it was reviewed in detail by Rohm 

and Haas (Keller 1998). The review by Keller (1998) is consistent with publications in the open 

literature relating to species specificity of tebufenozide (Addison. 1996; Biddinger and Hull. 

1995; Biddinger et al. 1998; Brown. 1996; Butler et al. 1997; Dhadialla et al. 1998; Rumpf et 

al. 1998; Smagghe and Degheele 1994a,b, 1997; Smagghe et al. 1995, 1996a,b; Valentine et al. 

1996). In general, Lepidoptera are sensitive to tebufenozide but other insects are much less 

sensitive (Smagghe and Degheele 1994a). The differences in sensitivity appear to be related to 

differences in ecdysone receptor binding (Smagghe et al. 1996a) rather than differences in 

pharmacokinetics (Smagghe and Degheele 1994b). 

There are four studies regarding the effects of tebufenozide to terrestrial invertebrates under field 

or field simulation conditions (Appendix 6). Three of these studies are published in the open 

literature (Addison 1996; Butler et al. 1997; Valentine et al. 1996), and one unpublished study 

was conducted by Rohm and Haas (Walgenbach 1995). The studies by Addison (1996) and 

Butler et al. (1997) are most directly relevant to this risk assessment because they assayed the 

effects on nontarget invertebrates in the forest canopy (Butler et al. 1997) and forest soil 

(Addison 1996) after the application of tebufenozide. 

In the study by Addison (1996), tebufenozide was incorporated into forest soil at a concentration 

of 72.1 ppm. Based on a typical application rate of 70 g/ha and the assumption that tebufenozide 

4-4
 
will remain in the top 2 cm of soil, Addison (1996) estimated that the soil concentration of 72.1 

ppm is equivalent to a concentration that is 100 times greater than expected environmental 

concentrations. There were no adverse effects on one species of earthworm (Dendrobaena 

octaedra) or on four species of Colembola (Folsomia candida, Folsomia nivalis, Onychiurus 

parvicornis, and Hypogastrura pannosa), which are indigenous to forest soils in Canada and the 

northern United States. Consistent with results of the Addison (1996) study, a standard bioassay 

on earthworms (Eisenia foetida) noted no adverse effects at soil concentrations of up to 1000 

ppm over a 14-day exposure period (Garvey 1992). 

Butler et al. (1997) conducted a study on canopy arthropods in which Mimic 4F was applied at 

rates of 0.03 and 0.06 lb a.i./acre to a mixed oak plot in Ohio. The investigators examined 

Mimic’s efficacy against Gypsy moth larvae and its effects on nontarget arthropods. Population 

assays included measures of abundance and diversity in10 arthropod families and 15 lepidopteran 

species. No effects on abundance or richness were noted in any organisms other than 

lepidopteran species. A decrease in abundance was noted in some lepidopteran species. The 

study indicates that there were problems associated with the application of Mimic 4F that 

resulted in poorer than expected efficacy, and that consequently, effects in nontarget lepidopteran 

species may have been underestimated. 

The studies by Valentine et al. (1996) and Walgenbach (1995) involve the application of 

tebufenozide formulations to apple orchards. The study by Valentine et al. (1996) found no 

effects of tebufenozide on species of mites, spiders, various beetles (Coleoptera), and true bugs 

(Hemiptera) after Mimic was applied to apple orchards at rates that were effective in controlling 

lepidopteran pest species. Similarly, Walgenbach (1995) noted no effects on beneficial insect 

populations after Confirm was applied to apple plots. While not as directly relevant to this risk 

assessment as the forestry studies summarized above, these two studies support the general 

conclusion that tebufenozide is likely to have an adverse impact on Lepidoptera but not on nonlepidopteran species. 

In addition to the above studies, the standard bee toxicity assay was conducted on tebufenozide 

(Atkins. 1990; Chan 1995). In this study, no mortality was observed at doses of up to 233.98 μg 

a.i./bee. Using a body weight of 0.093 g for the honey bee (USDA/APHIS 1993), this 

corresponds to a dose of about 2500 mg/kg bw [0.23 mg/0.000093 kg = 2473 mg/kg bw]. 

4.1.2.4. Terrestrial Plants (Macrophytes)– Standard bioassays for toxicity to terrestrial plants are 

required by the U.S. EPA for the registration of herbicides but not insecticides. No bioassays for 

herbicidal activity of tebufenozide were encountered in the published literature or in the U.S. 

EPA/OPP files. Thus, the potential effects of tebufenozide on terrestrial plant species is not 

discussed in other reviews of this compound (U.S. EPA 1999d,e; Keller 1998). The implicit 

presumption is that plausible levels of exposure to tebufenozide will not adversely affect 

terrestrial plant species. 

There are several field studies regarding the efficacy of tebufenozide applied to terrestrial 

4-5
 
 

vegetation for the control of various insect pests (e.g., Biddinger et al. 1998; Cadogan et al. 1997; 

Oberlander et al. 1998; Payne et al. 1997; Retnakaran et al. 1997a,b; Valentine et al. 1996; West 

et al. 1997). If tebufenozide were toxic to terrestrial plants at application rates that are used in 

the field, it is plausible that adverse effects would be reported in this literature. No such reports 

were encountered. 

Because there is no basis for further evaluating the assumption that tebufenozide will not cause 

adverse effects in terrestrial plants, such effects will not be considered quantitatively in this risk 

assessment. 

4.1.2.5. Terrestrial Microorganisms– As indicated in U.S. EPA (1999e), microbial 

transformation is the predominant route of environmental degradation in soil and water. Data 

regarding the toxicity of tebufenozide to terrestrial microorganisms, as with terrestrial plants, is 

not available in the open literature or the U.S. EPA/OPP files. Tebufenozide is degraded in soil 

by some microorganisms (e.g., Sundaram 1996, 1997a). Nonetheless, given the diversity of soil 

microorganisms and soil environments, generalizations concerning the potential effects on soil 

microflora cannot be supported. 

4.1.3. Aquatic Organisms. 

4.1.3.1. Fish– Information on the toxicity of tebufenozide to fish is summarized in Appendix 7. 

All of the available studies were conducted in support of the registration of tebufenozide and 

submitted to U.S. EPA/OPP. The summaries of these studies given in Appendix 7 were taken 

from the full text copies of the studies submitted to U.S. EPA. 

The acute toxicity of tebufenozide to fish is relatively low – i.e., LC50 values of 3.0 mg a.i./L in 

Bluegill sunfish (Graves and Smith 1992b) and 5.7 mg a.i./L in Rainbow trout (Graves and Smith 

1992c). There is greater concern, however, regarding the potential chronic toxicity of 

tebufenozide to fish. The U.S. EPA evaluates all studies like those summarized in Appendix 7 to 

determine whether the conclusions from the studies are consistent with the data presented in the 

studies. In many instances, the U.S. EPA accepts the study conclusions. For tebufenozide, 

however, the U.S. EPA has disagreed with conclusions for a fathead minnow egg and fry study 

(Bettancourt 1992) as well as a fathead minnow full life cycle study (Rhodes and Leak 1996). 

This is discussed further in the dose-response assessment (Section 4.3.3.1). 

4.1.3.2. Amphibians– No information was encountered on the toxicity of tebufenozide to 

amphibians. 

4.1.3.3. Aquatic Invertebrates – Unpublished studies on the toxicity of tebufenozide to aquatic 

invertebrates that were submitted to the U.S. EPA in support of the registration of tebufenozide 

are summarized in Appendix 8. Some invertebrate assays were conducted in support of the 

registration of tebufenozide, and the summaries of these studies are based on the full text copies 

of the studies submitted to U.S. EPA. Additional studies published in the open literature are 

discussed below. Unlike some of the fish studies, the studies on aquatic invertebrates, 

4-6
 
summarized in Appendix 8, were accepted without exception by the U.S. EPA (1999e). 

In the studies submitted for registration, the acute toxicity of tebufenozide to daphnia 

(Crustacea) and midges (Insecta) is on the same order as that for fish, with a 48 hour LC50 value 

of 3.8 mg/L for daphnids (Graves and Smith 1992a) and a 96 hour LC50 value of 0.3 mg/L for 

midge larvae (van der Kolk 1997). Similarly, in a study published in the open literature and 

sponsored by the U.S. Geological survey, Song et al. (1997) report higher LC50 values for 

Crustacea (daphnia = 17.37 mg/L; Artemia = 5.53 mg/L) than for two species of mosquitoes 

(0.92 mg/L for Aedes aegypti and 0.15 mg/L for Aedes taeniorhynchus). All of these bioassay 

results from Song et al. (1997) involved exposures at 27°C. In similar bioassays conducted at 

20°C, tebufenozide was substantially less toxic to both daphnids and Aedes aegypti. This 

negative relationship between toxicity and temperature is common. 

As with fish, there is a concern for potential reproductive effects in both a free swimming species 

(Daphnia) as well as a sediment dwelling species (midge). In Daphnia magna, significant 

decreases in the number of offspring/female were noted at 0.12 mg/L and a significant decrease 

in the growth of offspring was noted at 0.059 mg/L (McNamara 1991). In midges (Chironomus 

riparius), a decrease in larval emergence was noted at a concentration of 0.0053 mg/L. At 

concentrations of 0.04 mg/L and higher, midge emergence was completely suppressed (van der 

Kolk 1997). 

Kreutzweiser and Thomas (1995) assayed the effects of tebufenozide on aquatic invertebrate 

communities in lake enclosures at nominal concentrations of 0.07, 0.13, 0.33, and 0.66 mg/L. A 

dose-related decrease in cladoceran abundance was noted and persisted for 1-2 months at the two 

lower concentrations and for 12-13 months at the two higher concentrations. The decrease in 

cladoceran abundance was accompanied by an increase in the abundance of rotifers, suggesting 

that the changes in community structure could be attributable to secondary or trophic effects 

rather than to toxicity. 

Rohm and Haas summarized the results of Kreutzweiser and Thomas (1995) along with several 

other field studies or field simulation studies (e.g. Kreutzweiser et al. 1994) regarding the effects 

of tebufenozide to aquatic invertebrates (Keller 1998). The most relevant study for this risk 

assessment is an unpublished report submitted to U.S. EPA (Russell et al. 1996). In this study, 

Mimic was applied at a rate of 70 g a.i./ha to a small forest pond. The application resulted in an 

initial concentration of 0.00837 mg/L which decreased to 0.00016 mg/L 1 month after spray. 

During the 1-month post-application observation period, no adverse effects were noted on 

invertebrate populations, compared with a control (untreated) pond. Notably, the maximum 

concentration of 0.00837 mg/L is very close to the effect level of 0.0053 mg/L for midge larvae; 

however, the average concentration during the 1-month study was probably substantially below 

the effect level in midges. Thus, although this study seems to support the assertion that 

tebufenozide can be applied without interfering with aquatic invertebrate communities, it is not 

in conflict with the available bioassay data. 

4-7
 
4.1.3.4. Aquatic Plants – The toxicity of tebufenozide was assayed in two species of freshwater 

green algae, and details of these studies are presented in Appendix 8 along with the studies on 

aquatic invertebrates. Selenastrum capricornutum appears to be relatively insensitive to 

tebufenozide, with a NOEC for reduced cell density of 0.64 mg/L (Reinert 1993b), which is 

greater than the effect levels in aquatic invertebrates by a factor of 10-100. 

Scenedesmus subspicatus appears to be much more sensitive than Selenastrum capricornutum 

although still much less sensitive than aquatic invertebrates, with a NOAEL and LOAEL for 

growth rate inhibition of 0.077 and 0.15 mg/L, respectively. Decreased cell density was a 

somewhat more sensitive effect with a NOAEL 0.046 mg/L and a LOAEL of 0.077 mg/L 

(Hoberg 1992a). 

In an aquatic microcosm study with mixed species of algae, Sundaram et al. (1997b) report that 

tebufenozide stimulated algal growth at concentrations of 0.25 and 0.75 mg/L. 

4.1.3.5. Aquatic Microorganisms (Other than algae) – Other than the effect in algae, 

summarized in the previous section, no studies regarding the toxicity of tebufenozide to aquatic 

microorganisms were encountered. 

4-8
 
4.2. EXPOSURE ASSESSMENT 

4.2.1. Overview 

Details of the exposure assessments for tebufenozide are given in the EXCEL workbook that 

accompany this risk assessment (Supplement 1). Most exposure assessments are based on two 

applications spaced 3 days apart at an application rate of 0.12 lb/acre. As in the human health 

risk assessment, two sets of exposure assessments are given for scenarios involving the longerterm consumption of contaminated vegetation: one for a single application at 0.12 lb/acre and 

another for two applications spaced 3 days apart at an application rate of 0.12 lb/acre. 

Terrestrial animals might be exposed to any applied pesticide from direct spray, the ingestion of 

contaminated media (vegetation, prey species, or water), grooming activities, or indirect contact 

with contaminated vegetation. For tebufenozide, the highest acute exposure for a terrestrial 

vertebrate is associated with a fish-eating bird and could reach up to about 85 mg/kg. Exposures 

anticipated from the consumption of contaminated vegetation by terrestrial animals range from 

central estimates of about 0.15 mg/kg for a small mammal consuming fruit to about 3 mg/kg for 

a large bird with upper ranges of about 0.4 mg/kg for a small mammal and 9 mg/kg for a large 

bird. The consumption of contaminated water leads to much lower levels of exposure. A similar 

pattern is seen for chronic exposures. Estimated longer-term daily doses for the a small mammal 

from the consumption of contaminated vegetation at the application site are in the range of about 

0.000002 mg/kg/day to 0.08 mg/kg/day. Large birds feeding on contaminated vegetation at the 

application site could be exposed to much higher concentrations, ranging from about 0.015 

mg/kg/day to 11 mg/kg/day. The upper ranges of exposure from contaminated vegetation far 

exceed doses that are anticipated from the consumption of contaminated water, which range from 

about 0.0000003 mg/kg/day to 0.0002 mg/kg/day for a small mammal. 

Exposure to aquatic organisms is based on essentially the same information used to assess the 

exposure to terrestrial species from contaminated water. The peak estimated concentration of 

tebufenozide in ambient water is 10 (0.005 to 40) μg/L after two applications of 0.12 lb/acre 

spaced three days apart. For longer-term exposures, the corresponding longer term 

concentrations in ambient water are estimated at about 0.004 (0.00002 to 0.01) μg/L. 

4.2.2. Terrestrial Animals 

Terrestrial animals might be exposed to any applied insecticide from direct spray, the ingestion 

of contaminated media (vegetation, prey species, or water), grooming activities, or indirect 

contact with contaminated vegetation. 

In this exposure assessment, estimates of oral exposure are expressed in the same units as the 

available toxicity data. As in the human health risk assessment, these units are usually expressed 

as mg of agent per kg of body weight and abbreviated as mg/kg for terrestrial animals. One 

exception in this risk assessment involves terrestrial invertebrates. As detailed in the doseresponse assessment (Section 4.3), toxicity data in units of mg/kg bw are available for some 

terrestrial invertebrates and these data are used in a manner similar to that for terrestrial 

vertebrates. For other species, however, standard toxicity studies report units that are not directly 

4-9
 
useful in a quantitative risk assessments – e.g., contact toxicity based on petri dish exposures. As 

an alternative, some dose response assessments are based on field studies in which the dose 

metameter is simply the application rate in units of mass per area such as g a.i./ha. 

For dermal exposures to terrestrial animals, the units of measure usually are expressed in mg of 

agent per cm2 of surface area of the organism and abbreviated as mg/cm . 2 In estimating dose, 

however, a distinction is made between the exposure dose and the absorbed dose. The exposure 

dose is the amount of material on the organism (i.e., the product of the residue level in mg/cm2 

and the amount of surface area exposed), which can be expressed either as mg/organism or 

mg/kg body weight. The absorbed dose is the proportion of the exposure dose that is actually 

taken in or absorbed by the animal. 

The exposure assessments for terrestrial animals are summarized in Worksheet G01. As with the 

human health exposure assessment, the computational details for each exposure assessment 

presented in this section are provided as scenario specific worksheets (Worksheets F01 through 

F16b). Given the large number of species that could be exposed to insecticides and the varied 

diets in each of these species, a very large number of different exposure scenarios could be 

generated. For this generic risk assessment, an attempt is made to limit the number of exposure 

scenarios. 

Because of the relationship of body weight to surface area as well as to the consumption of food 

and water, small animals will generally receive a higher dose, in terms of mg/kg body weight, 

than large animals will receive for a given type of exposure. Consequently, most general 

exposure scenarios for mammals and birds are based on a small mammal or bird. For mammals, 

the body weight is taken as 20 grams, typical of mice, and exposure assessments are conducted 

for direct spray (F01 and F02a), consumption of contaminated fruit (F03, F04a, F04b), and 

contaminated water (F05, F06, F07). Grasses will generally have higher concentrations of 

insecticides than fruits and other types of vegetation (Fletcher et al. 1994; Hoerger and Kenaga 

1972). Because small mammals do not generally consume large amounts of grass, the scenario 

for the assessment of contaminated grass is based on a large mammal (Worksheets F10, F11a, 

and F11b). Other exposure scenarios for a mammals involve the consumption of contaminated 

insects by a small mammal (Worksheet F14a) and the consumption of small mammals 

contaminated by direct spray by a large mammalian carnivore (Worksheet F16a). Exposure 

scenarios for birds involve the consumption of contaminated insects by a small bird (Worksheet 

F14b), the consumption of contaminated fish by a predatory bird (Worksheets F08 and F09), the 

consumption by a predatory bird of small mammals contaminated by direct spray and the 

consumption by a large bird of contaminated grasses (F12, F13a, and F13b). 

While a very large number of other exposure scenarios could be generated, the specific exposure 

scenarios developed in this section are designed as conservative screening scenarios that may 

serve as guides for more detailed site-specific assessments by identifying the groups of organisms 

and routes of exposure that are of greatest concern. 

4-10
 
4.2.2.1. Direct Spray – In the broadcast application of any insecticide, wildlife species may be 

sprayed directly. This scenario is similar to the accidental exposure scenarios for the general 

public discussed in Section 3.2.3.2. In a scenario involving exposure to direct spray, the amount 

absorbed depends on the application rate, the surface area of the organism, and the rate of 

absorption. 

For this risk assessment, three groups of direct spray exposure assessments are conducted. The 

first, which is defined in Worksheet F01, involves a 20 g mammal that is sprayed directly over 

one half of the body surface as the chemical is being applied. The range of application rates as 

well as the typical application rate is used to define the amount deposited on the organism. The 

absorbed dose over the first day (i.e., a 24-hour period) is estimated using the assumption of firstorder dermal absorption. An empirical relationship between body weight and surface area 

(Boxenbaum and D’Souza 1990) is used to estimate the surface area of the animal. The 

estimates of absorbed doses in this scenario may bracket plausible levels of exposure for small 

mammals based on uncertainties in the dermal absorption rate. 

Other, perhaps more substantial, uncertainties affect the estimates for absorbed dose. For 

example, the estimate based on first-order dermal absorption does not consider fugitive losses 

from the surface of the animal and may overestimate the absorbed dose. Conversely, some 

animals, particularly birds and mammals, groom frequently, and grooming may contribute to the 

total absorbed dose by direct ingestion of the compound residing on fur or feathers. Furthermore, 

other vertebrates, particularly amphibians, may have skin that is far more permeable than the skin 

of most mammals. Quantitative methods for considering the effects of grooming or increased 

dermal permeability are not available. As a conservative upper limit, the second exposure 

scenario, detailed in Worksheet F02a, is developed in which complete absorption over day 1 of 

exposure is assumed. 

Because of the relationship of body size to surface area, very small organisms, like bees and 

other terrestrial invertebrates, might be exposed to much greater amounts of a pesticide per unit 

body weight compared with small mammals. Consequently, a third exposure assessment is 

developed using a body weight of 0.093 g for the honey bee (USDA/APHIS 1993) and the 

equation above for body surface area proposed by Boxenbaum and D’Souza (1990). Because 

there is no information regarding the dermal absorption rate of tebufenozide by bees or other 

invertebrates, this exposure scenario, detailed in Worksheet F02b, also assumes complete 

absorption over the first day of exposure. As noted above, exposures for other terrestrial 

invertebrates are based on field studies in which application rate is the most relevant expression 

of exposure. This is discussed further in Section 3.3 (Dose-Response Assessment) and Section 

3.4 (Risk Characterization). 

Direct spray scenarios are not given for large mammals. As noted above, allometric relationships 

dictate that large mammals will be exposed to lesser amounts of a compound in any direct spray 

scenario than smaller mammals. 

4-11
 
4.2.2.2. Indirect Contact – As in the human health risk assessment (see Section 3.2.3.3), the 

only approach for estimating the potential significance of indirect dermal contact is to assume a 

relationship between the application rate and dislodgeable foliar residue. Unlike the human 

health risk assessment in which transfer rates for humans are available, there are no transfer rates 

available for wildlife species. As discussed in Durkin et al. (1995), the transfer rates for humans 

are based on brief (e.g., 0.5 to 1-hour) exposures that measure the transfer from contaminated soil 

to uncontaminated skin. Wildlife, compared with humans, are likely to spend longer periods of 

time in contact with contaminated vegetation. It is reasonable to assume that for prolonged 

exposures an equilibrium may be reached between levels on the skin, rates of absorption, and 

levels on contaminated vegetation, although there are no data regarding the kinetics of such a 

process. The bioconcentration data on tebufenozide indicates that this compound will 

accumulate in the tissue of the fish. Thus, it is plausible that the absorbed dose resulting from 

contact with contaminated vegetation will be as great as those associated with comparable direct 

spray scenarios and possibly larger than those associated with the consumption of contaminated 

vegetation. 

4.2.2.3. Ingestion of Contaminated Vegetation or Prey – Since tebufenozide will be applied to 

vegetation, the consumption of contaminated vegetation is an obvious concern and separate 

exposure scenarios are developed for acute and chronic exposure scenarios for a small mammal 

(Worksheets F04a and F04b) and large mammal (Worksheets F10, F11a, and F11b) as well as 

large birds (Worksheets F12, F13a, and F13b). 

As discussed in Section 2.4, tebufenozide may be applied once or twice per season at an 

application rate of up to 0.12 lb/acre per application. In order to encompass the effects of both a 

single application per season and two applications per season, two sets of exposure assessments 

are given for the all scenarios involving the longer-term consumption of contaminated 

vegetation: one for a single application at 0.12 lb/acre and another for two applications spaced 3 

days apart at an application rate of 0.12 lb/acre. For example, Worksheet 04bi presents the timeweighted average dose for a single application and Worksheet 04bii presents the time-weighted 

average dose for two applications spaced 3 days apart. This is also done for Worksheets F11a, 

F11b, F13a, and F13b. The calculation of the time-weighted average doses are identical to those 

used in the human health risk assessment (Section 3.2.3.6). 

For the consumption of contaminated vegetation, a small mammal is used because allometric 

relationships indicate that small mammals will ingest greater amounts of food per unit body 

weight, compared with large mammals. The amount of food consumed per day by a small 

mammal (i.e., an animal weighing approximately 20 g) is equal to about 15% of the mammal's 

total body weight (U.S. EPA/ORD 1989). When applied generally, this value may overestimate 

or underestimate exposure in some circumstances. For example, a 20 g herbivore has a caloric 

requirement of about 13.5 kcal/day. If the diet of the herbivore consists largely of seeds (4.92 

kcal/g), the animal would have to consume a daily amount of food equivalent to approximately 

14% of its body weight [(13.5 kcal/day ÷ 4.92 kcal/g)÷20g = 0.137]. Conversely, if the diet of 

the herbivore consists largely of vegetation (2.46 kcal/g), the animal would have to consume a 

4-12
 
daily amount of food equivalent to approximately 27% of its body weight [(13.5 kcal/day ÷ 2.46 

kcal/g)÷20g = 0.274] (U.S. EPA/ORD 1993, pp.3-5 to 3-6). For this exposure assessment 

(Worksheet F03), the amount of food consumed per day by a small mammal weighing 20 g is 

estimated at about 3.6 g/day or about 18% of body weight per day from the general allometric 

relationship for food consumption in rodents (U.S. EPA/ORD 1993, p. 3-6). 

A large herbivorous mammal is included because empirical relationships of concentrations of 

pesticides in vegetation, discussed below, indicate that grasses may have substantially higher 

pesticide residues than other types of vegetation such as forage crops or fruits (Worksheet A04). 

Grasses are an important part of the diet for some large herbivores, but most small mammals do 

not consume grasses as a substantial proportion of their diet. Thus, even though using residues 

from grass to model exposure for a small mammal is the most conservative approach, it is not 

generally applicable to the assessment of potential adverse effects. Hence, in the exposure 

scenarios for large mammals, the consumption of contaminated range grass is modeled for a 70 

kg herbivore. Caloric requirements for herbivores and the caloric content of vegetation are used 

to estimate food consumption based on data from U.S. EPA/ORD (1993). Details of these 

exposure scenarios are given in worksheets F10 for acute exposures as well as Worksheets F11a 

and F11b for longer-term exposures. 

For the acute exposures, the assumption is made that the vegetation is sprayed directly – i.e., the 

animal grazes on site – and that 100% of the animal’s diet is contaminated. While appropriately 

conservative for acute exposures, neither of these assumptions are plausible for longer-term 

exposures. Thus, for the longer-term exposure scenarios for the large mammal, two subscenarios are given. The first is an on-site scenario that assumes that a 70 kg herbivore consumes 

short grass for a 90 day period after application of the chemical. In the worksheets, the 

contaminated vegetation is assumed to account for 30% of the diet with a range of 10% to 100% 

of the diet. These are essentially arbitrary assumptions reflecting grazing time at the application 

site by the animal. Because the animal is assumed to be feeding at the application site, drift is set 

to unity - i.e., direct spray. This scenario is detailed in Worksheet 11a. The second sub-scenario 

is similar except the assumption is made that the animal is grazing at distances of 25 to 100 feet 

from the application site (lowing risk) but that the animal consumes 100% of the diet from the 

contaminated area (increasing risk). For this scenario, detailed in Worksheet F12b, AgDRIFT is 

used to estimate deposition on the off-site vegetation. Drift estimates from AgDrift are 

summarized in Worksheet A06 and this model is discussed further in Section 4.2.3.2. 

The consumption of contaminated vegetation is also modeled for a large bird. For these 

exposure scenarios, the consumption of range grass by a 4 kg herbivorous bird, like a Canada 

Goose, is modeled for both acute (Worksheet F12) and chronic exposures (Worksheets F13a and 

F13b). As with the large mammal, the two chronic exposure scenarios involve sub-scenarios for 

on-site as well as off-site exposure. 

For this component of the exposure assessment, the estimated amounts of pesticide residue in 

vegetation are based on the relationship between application rate and residue rates on different 

4-13
 
types of vegetation. As summarized in Worksheet A04, these residue rates are based on 

estimated residue rates from Fletcher et al. (1994). 

Similarly, the consumption of contaminated insects is modeled for a small (10g) bird and a small 

(20g) mammal. No monitoring data have been encountered on the concentrations of 

tebufenozide in insects after applications of tebufenozide. The empirical relationships 

recommended by Fletcher et al. (1994) are used as surrogates as detailed in Worksheets F14a and 

F14b. To be conservative, the residue rates from small insects are used – i.e., 45 to 135 ppm per 

lb/ac – rather than the residue rates from large insects – i.e., 7 to 15 ppm per lb/ac. 

A similar set of scenarios is provided for the consumption of small mammals by either a 

predatory mammal (Worksheet 16a) or a predatory bird (Worksheet 16a). Each of these 

scenarios assumes that the small mammal is directly sprayed at the specified application and the 

concentration of the compound in the small mammal is taken from the worksheet for direct spray 

of a small mammal under the assumption of 100% absorption (Worksheet F02a). 

In addition to the consumption of contaminated vegetation and insects, tebufenozide may reach 

ambient water and fish. Thus, a separate exposure scenario is developed for the consumption of 

contaminated fish by a predatory bird in both acute (Worksheet F08) and chronic (Worksheet 

F09) exposures. Because predatory birds usually consume more food per unit body weight than 

do predatory mammals (U.S. EPA 1993, pp. 3-4 to 3-6), separate exposure scenarios for the 

consumption of contaminated fish by predatory mammals are not developed. 

4.2.2.4. Ingestion of Contaminated Water – Estimated concentrations of tebufenozide in water 

are identical to those used in the human health risk assessment (Worksheet B06). The only major 

differences involve the weight of the animal and the amount of water consumed. There are 

well-established relationships between body weight and water consumption across a wide range 

of mammalian species (e.g., U.S. EPA 1989). Mice, weighing about 0.02 kg, consume 

approximately 0.005 L of water/day (i.e., 0.25 L/kg body weight/day). These values are used in 

the exposure assessment for the small (20 g) mammal. Unlike the human health risk assessment, 

estimates of the variability of water consumption are not available. Thus, for the acute scenario, 

the only factors affecting the estimate of the ingested dose include the field dilution rates (i.e., the 

concentration of the chemical in the solution that is spilled) and the amount of solution that is 

spilled. As in the acute exposure scenario for the human health risk assessment, the amount of 

the spilled solution is taken as 200 gallons. In the exposure scenario involving contaminated 

ponds or streams due to contamination by runoff or percolation, the factors that affect the 

variability are the water contamination rate, (see Section 3.2.3.4.2) and the application rate. 

Details regarding these calculations are summarized in Worksheets F06 and Worksheet F07. 

4.2.3. Terrestrial Plants 

Terrestrial plants will certainly be exposed to tebufenozide. A large number of different 

exposure assessments could be made for terrestrial plants – i.e., direct spray, spray drift, runoff, 

wind erosion and the use of contaminated irrigation water. Such exposure assessments are 

4-14
 
typically conducted for herbicides. For tebufenozide, however, the development of such 

exposure assessments would serve no purpose. As discussed in Section 4.1.2.4 (Hazard 

Identification for Terrestrial Plants), there is no basis for asserting that tebufenozide will cause 

adverse effects in terrestrial plants. Thus, no formal exposure assessment is conducted for 

terrestrial plants. 

4.2.4. Soil Organisms 

For both soil microorganisms and soil invertebrates, the toxicity data are typically expressed in 

units of soil concentration – i.e., mg agent/kg soil which is equivalent to parts per million (ppm) 

concentrations in soil. The GLEAMS modeling, discussed in Section 3.2.3.4.3, provides 

estimates of concentration in soil as well as estimates of off-site movement (runoff, sediment, 

and percolation). Based on the GLEAMS modeling, concentrations in clay, loam, and sand over 

a wide range of rainfall rates are summarized in Table 4-1. As indicated in this table, peak soil 

concentrations after two applications at an application rate of 0.12 lb/acre are in a relatively 

narrow range: about 0.02 to 0.1 mg/kg (ppm) over all soil types and rainfall rates. Longer term 

concentrations in soil are all low and are on the order of 0.003 to 0.05 mg/kg – i.e., 3 ppb to 

50 ppb. 

4.2.5. Aquatic Organisms 

The plausibility of effects on aquatic species is based on estimated concentrations of 

tebufenozide in water that are identical to those used in the human health risk assessment. As 

summarized in Table 3-5, the peak estimated concentration of tebufenozide in ambient water is 

10 (0.005 to 40) μg/L after two applications of 0.12 lb/acre spaced three days apart. For longerterm exposures, the corresponding longer term concentrations in ambient water are estimated at 

about 0.004 (0.00002 to 0.01) μg/L. 

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4.3. DOSE-RESPONSE ASSESSMENT 

4.3.1. Overview 

The specific toxicity values used in this risk assessment are summarized in Table 4-2, and the 

derivation of each of these values is discussed in the various subsections of this dose-response 

assessment. The first column in Table 4-2 specifies the organism to which the toxicity value 

applies. The available toxicity data support separate dose-response assessments in six classes of 

organisms: terrestrial mammals, birds, nontarget terrestrial invertebrates, fish, aquatic 

invertebrates, and aquatic algae. Different units of exposure are used for different groups of 

organisms depending on how exposures are likely to occur and how the available toxicity data 

are expressed. 

Tebufenozide is relatively non-toxic to mammals and birds. For mammals, the toxicity values 

used in the ecological risk assessment are identical to those used in the human health risk 

assessments: an acute NOAEL for reproductive toxicity of 1000 mg/kg and a chronic NOAEL of 

1.8 mg/kg/day based on effects on the blood. For birds, the acute NOAEL for tebufenozide is 

taken as 2150 mg/kg from an acute oral study in which the dose was administered in capsules for 

21-days. The longer term NOAEL is taken as 15 mg/kg/day from a standard reproduction study 

in bobwhite quail. 

For terrestrial invertebrates, three types of data are used to characterize risks: a contact bioassay 

in the honey bee, a soil bioassay in earthworms, and field studies in which population level 

effects were monitored in insects. The standard contact bioassay in honey bees indicates an 

NOEC of 2500 mg/kg bw, comparable to the acute toxicity values in mammals and birds. The 

earthworm bioassay indicates a NOEC of 1000 mg/kg soil. The available field studies indicate 

that tolerant insect species are not affected by application rates up to 0.24 lb/acre. The true 

NOEC may be higher – i.e., an LOEC has not been identified for tolerant species of terrestrial 

insects. Conversely, application rates as low as 0.03 lb/acre have been shown to adversely affect 

sensitive nontarget insects, primarily Lepidoptera and a NOEC for sensitive species has not been 

identified. 

Acute toxicity values for aquatic species indicate relatively little difference between fish and 

aquatic invertebrates. For fish, the acute NOEC values are 0.39 mg/L and 1.9 mg/L for sensitive 

and tolerant species, respectively. For invertebrates, the corresponding acute NOEC values are 

0.12 mg/L and 0.82 mg/L. Differences between fish and invertebrates are difficult to assess in 

terms of longer-term toxicity. For fish, data are available on only a single species, the fathead 

minnow, and only a LOAEL of 0.048 mg/L is available. For invertebrates, longer-term NOEC 

values of 0.0035 mg/L and 0.029 mg/L are used for sensitive and tolerant species. Toxicity 

values for aquatic plants are taken as 0.077 mg/L for sensitive species and 0.64 mg/L for tolerant 

species, somewhat below the acute NOEC values in fish and aquatic invertebrates. Because of 

the short life-cycle of individual algal cells, the relatively short-term bioassays in algae (i.e., 96 to 

120 hours) are applied to both acute and longer-term concentrations for the characterization of 

risk. 

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4.3.2. Toxicity to Terrestrial Organisms 

4.3.2.1. Mammals – As summarized in the dose-response assessment for the human health risk 

assessment (see Section 3.3.3.), the most sensitive effect in experimental mammals involves 

toxic effects in red blood cells. The chronic NOAEL for this endpoint in experimental mammals 

is 1.8 mg/kg/day (U.S. EPA 1999b) and is based on a dog study (Richards 1992a) in which 

beagles of either sex were provided with dietary concentrations of 0, 15, 50, 250, or 1500 ppm 

technical grade tebufenozide for 52 weeks (Appendix 2). No effects were seen in the 50 ppm 

exposure group which corresponded to an average dose of 1.8 mg/kg/day. At 250 ppm, which 

corresponded to an average dose of 20 mg/kg/day, a direct effect on red blood cells was indicated 

by increased concentrations of methemoglobin in the blood as well as changes in several other 

hematological parameters associated with toxic effects in red blood cells. Thus, for this risk 

assessment, 1.8 mg/kg/day is taken as the chronic NOAEL for general toxic effects. 

Tebufenozide is also associated with adverse reproductive effects in mammals in a 2-generation 

study (see Section 3.1.4). In the study by Danberry et al. (1993), reproductive effects were not 

observed in rats given a dietary concentration of 150 ppm (.12 mg/kg bw) tebufenozide; 

however, in the same study, rats given a dietary concentration of 2000 ppm (.160 mg/kg bw) 

demonstrated clearly adverse effects, including increased mortality in females during delivery 

and decreases in implantation. This endpoint, with a longer-term NOAEL of 12 mg/kg/day and a 

LOAEL of 160 mg/kg/day, is also used in the characterization of risk (Section 4.4.2) to help 

elaborate the potential effects of exposures that exceed the general NOAEL of 1.8 mg/kg/day. 

Consistent with the approach taken in the human health risk assessment (Section 3.3.4), acute 

(1-day) exposures will be based on the acute NOAEL of 1000 mg/kg/day from reproduction 

studies in both rats (Hoberman 1991) and rabbits (rabbits) involving 10 to 13 day exposure 

periods. 

4.3.2.2. Birds – As detailed in Appendix 4, adverse reproductive effects were observed in 

bobwhite quail provided with dietary concentrations of 300 or 1000 ppm (Beavers et al. 1993b). 

Similar effects were not observed in mallard ducks provided with dietary concentrations of up to 

1000 ppm in a study conducted by the same investigators (Beavers et al. 1993a) or in a follow-up 

study on bobwhite quail provided with dietary concentrations of up to 615 ppm (Reinert 1995a). 

As discussed in Section 4.1.2.2, the earlier study by Beavers et al. (1993b) is used to identify 

reproductive toxicity as an endpoint of concern in this risk assessment because there is no basis 

for discounting the study or explaining the discrepancies between the Beavers et al. (1993b) and 

Reinert (1995a) studies in bobwhite quail. In addition, reasonable consistency is apparent in the 

reported dose levels associated with reproductive effects in mammals and the reported dose 

levels in Beavers et al. (1993b) study. This approach is consistent with that taken by U.S. EPA 

(1999e). 

It is worth noting that the two quail studies use different methods to report the estimated dose 

(i.e., the dose as mg/kg bw/day based on dietary concentrations and food consumption). In the 

study by Beavers et al. (1993b), “No attempt was made to quantify the amount of feed wasted by 

4-17
 
the birds, as the wasted feed is normally scattered and mixed with water and excreta.” (Beavers et 

al. 1993b, p. 16). In the study by Reinert (1995a), food consumption estimates did explicitly 

consider measurements of food wastage (i.e., food scattered from the container and not 

consumed). Furthermore, the study by Beavers et al. (1993b) states explicitly that food was 

administered ad libitum—an excess of food was freely available to the animals. This protocol is 

not specified in the study by Reinert (1995a); however, it seems reasonable to assume that the 

food was available ad libitum because a restricted feeding protocol is atypical and would have 

been specified in the methods section of the study. These reporting differences are relatively 

inconsequential, assuming that both studies use ad libitum feeding. 

Of greater importance, however, is the exposure metameter (i.e., how the exposure is expressed 

in the dose-response and the exposure assessments). The U.S. EPA (1999e) uses reported dietary 

concentrations. This approach, however, may be under protective. Laboratory diets generally 

involve the use of dry food, and dry food is specified in all of the bird feeding studies on 

tebufenozide. Dry laboratory chow usually has a higher caloric content than food consumed in 

the wild, if only because most food consumed in the wild has a high water content. In addition, 

most reported concentrations of a pesticide in environmental samples are given on a wet (natural) 

weight rather than a dry (dedicated) weight basis. Consequently, animals tend to eat greater 

amounts of food in the wild than they do under laboratory conditions (U.S. EPA 1993). 

Consequently, for a fixed concentration in food, ingested doses expressed as mg/kg bw/day often 

will be higher in free living animals than in laboratory animals. 

Because of these relationships, Forest Service risk assessments use doses expressed as mg/kg 

body weight for both the exposure and dose-response assessments. As detailed in the 

worksheets, information on caloric requirements and caloric values of different foods are used to 

estimate the amount of a particular food that an animal will use. 

For this risk assessment, the food consumption values reported by Beavers et al. (1993b) are used 

to estimate a NOAEL and a LOAEL of 15 and 45 mg/kg bw/day, respectively. This is not the 

most conservative approach that could be taken, because Beavers et al. (1993b) did not consider 

wastage in their estimates of food consumption. By comparison with the study by Reinert 

(1995a), the food consumption and hence the ingested amounts of tebufenozide could have been 

lower by a factor of about 2 [i.e., food consumption rates of 30 g per bird in Beavers et al. 

(1993b) and 16 g per bird in Reinert (1995a)]. Compared with other uncertainties in this risk 

assessment, this difference is relatively modest. The dose adjustment is incorporated explicitly 

into the dose-response assessment, and given further consideration in the risk characterization. 

As with mammals, the acute toxicity of tebufenozide to birds appears to be very low. As 

indicated in Appendix 4, acute dietary LC50 values are greater than 5000 ppm (mg tebufenozide 

per kg diet) in both bobwhite quail and mallard ducks (Fletcher 1990a,b). In addition, 21 daily 

doses at both 1470 and 2150 mg a.i./kg bw, via gelatin capsule, caused no signs of toxicity in 

male or female bobwhite quail (Fletcher 1987). For this risk assessment, the 21-day exposure 

data from Fletcher (1987) will be used set an acute NOAEL of 2150 mg/kg bw for birds and this 

4-18
 
value will be applied to all short-term (1-day) exposure assessments. 

4.3.2.3. Terrestrial Invertebrates – As discussed in Section 4.1.2.3, tebufenozide mimics the 

invertebrate hormone 20-hydroxyecdysone and could cause adverse effects in a variety of 

terrestrial invertebrates. Notwithstanding this assertion, however, there are adequate field and 

field simulation studies clearly indicating that tebufenozide is much more toxic to Lepidoptera 

than to other insects. 

Dose-response assessments for the effects of tebufenozide on terrestrial invertebrates could be 

based on either laboratory toxicity studies (Appendix 5) or field studies (Appendix 6). Most of 

the laboratory studies are on target rather than nontarget invertebrates and many involve 

exposures that are not readily applied to risk assessment. Studies that do involve both target and 

nontarget insects indicate that tebufenozide is more toxic to Lepidoptera (target species) than 

non-lepidopteran arthropods (Medina et al. 2002, 2003; Pietrantionio and Benedict 1999). In 

addition, tebufenozide appears to be less toxic to one nontarget species (lacewing) than 

diflubenzuron, another agent used to control the gypsy moth (Medina et al. 2002, 2003; Rumph 

et al. 1998). 

The laboratory observations that non-lepidopteran arthropods are less sensitive to tebufenozide 

than Lepidoptera are supported by the field studies detailed in Appendix 6. A summary of the 

most relevant field studies is given in Table 4-3. In this table, efficacy studies summarized in 

Appendix 6 – i.e., those studies looking only at effects on target species, are omitted. Based on 

the study by Butler et al. (1997), both target and nontarget macrolepidoptera will be adversely 

affected at application rates as low as 0.03 lb/acre. Field studies at lower application rates have 

not been encountered and a NOAEL for nontarget macrolepidoptera cannot be identified. 

Similarly, a clear LOAEL for non-lepidopteran arthropods has not been identified. Mulder and 

Prescott (1999a) report a decrease in the numbers of beneficial arthropods on Day 3 after the 

application of tebufenozide at 0.125 lb a.i./acre but not at 0.24 lb a.i./acre. In addition, no effects 

on beneficial arthropods were seen at 0.125 lb/acre or 0.25 lb/acre on Day 5 to Day 15 after 

treatment. 

For this risk assessment, the assumption is made that effects on sensitive nontarget Lepidoptera 

are likely to be comparable to those seen in target species. This assumption is based on the field 

study by Butler et al. (1997) in which a decrease in abundance in some lepidopteran species was 

noted after the application of Mimic 4F at rates of 0.03 and 0.06 lb a.i./acre. This may be a 

conservative assumption because, as noted by Butler et al. (1997), not all nontarget lepidopteran 

species were affected. Conversely, these investigators also noted that problems were 

encountered in the application of Mimic 4F, which resulted in poorer than expected efficacy. 

Thus, effects in nontarget lepidopteran species also may have been underestimated. 

In the risk characterization, the minimum recommended application rate of 0.03 lb a.i./acre is 

taken as the exposure level that could be associated with adverse effects in some nontarget 

lepidopteran species. The true NOAEL in terms of application rate has not been defined for 

4-19
 
nontarget lepidopteran species. 

The potential for adverse effects on other nontarget insects is characterized quantitatively on the 

basis of the standard bioassay in the honey bee (Atkins. 1990; Chan 1995) in which no mortality 

was observed at doses of up to 233.98 μg a.i./bee or about 2500 mg/kg bw (see Section 4.1.2.3). 

As indicated in Table 4-2, this risk assessment also uses an application rate of 0.24 lb/acre as a 

functional NOEC for non-lepidopteran arthropods. This is based on the studies summarized in 

Table 4-3. As noted above, the application rate of 0.125 lb/acre from Mulder and Prescott 

(1999a) could be interpreted as a marginal LOEC. This interpretation would be grossly 

conservative because the effects seen at 0.125 lb/acre were transient and were not seen at 0.24 

lb/acre. 

Toxicity to soil invertebrates will be based on the standard toxicity bioassay in earthworms 

(Garvey 1992, discussed in Section 4.1.2.3) in which no effects were noted at soil concentrations 

of up to 1000 ppm (1000 mg/kg soil). 

4.3.2.4. Terrestrial Plants and Microorganisms – As discussed in Sections 4.1.2.4. and 4.1.2.5., 

there is no reason to assume that tebufenozide will cause adverse effects in terrestrial plants or 

terrestrial microorganisms. Nonetheless, no standard toxicity studies have been encountered that 

could be used to quantify risk in either terrestrial plants or soil microorganisms. Consequently, 

no dose-response assessment for these groups can be proposed. 

4.3.3. Aquatic Organisms. 

4.3.3.1. Fish – The acute bioassays on fish summarized in Appendix 7 provide estimates of 

exposures which might be associated acute effects in fish but only two species have been tested. 

The most sensitive species is the bluegill sunfish with a 96-hour LC50 of 3.0 (2.2 to 4.0) mg/L 

with an NOEC of 0.39 mg/L (Graves and Smith 1992b). Rainbow trout appear to be somewhat 

less sensitive, with an LC50 value of 5.7 mg/L (4.7 to 6.5 mg/L) and an NOEC of 1.9 mg/L 

(Graves and Smith 1992c). For this risk assessment, the NOEC values of 0.39 mg/L and 1.9 

mg/L are used to assess the consequences of short-term exposures for sensitive and tolerant 

species. 

The assessment of the effects of tebufenozide that might be associated with chronic exposure to 

contaminated ambient water from the normal use and application of this product is based on the 

full life cycle study in fathead minnows by Rhodes and Leak (1996) supported by the egg and fry 

study by Bettancourt (1992). 

In the egg and fry study (Bettancourt 1992), eggs were incubated at mean measured 

concentrations of 0, 0.084, 0.14, 0.22, 0.36, or 0.71 mg a.i./L by continuous exposure for 35 

days. Based on a comparison to pooled controls (i.e., untreated and solvent treated animals with 

a combined survival of 94% ), Bettancourt (1992) reports no effects on survival at any 

concentration level. The U.S. EPA (1999e), however, classified the 0.71 mg/L concentration as 

an effect level based on decreased survival (88%) relative to survival in the solvent control 

4-20
 
(98%). The U.S. EPA analysis was challenged by Rohm and Haas (Surprenant 1994). 

In the full life cycle study (Rhodes and Leak 1996), newly hatched eggs were exposed to mean 

measured concentrations of 0, 0.048, 0.090, 0.18, 0.35, or 0.72 mg a.i./L, again using both 

untreated and solvent (acetone) controls. The exposure was continued for 219 days which 

allowed for full development of the fish and reproduction. The most sensitive endpoint reported 

by Rhodes and Leak (1996) using pooled control data was survival with a LOAEL of 0.35 mg 

a.i./L and a NOAEL of 0.18 mg a.i./L. Again using solvent control rather than pooled control 

data, the U.S. EPA identified the most sensitive effect as decreased eggs/spawn and identified the 

LOAEL as 0.048 mg a.i./L, the lowest concentration tested. Because the U.S. EPA does not 

consider that this study identified a NOAEL, the U.S. EPA stated that the full life cycle study 

must be repeated (U.S. EPA 1999e). Again, the U.S. EPA analysis was contested by Rohm and 

Haas (Reinert et al. 1999). 

The decision to pool or not pool control data is both statistical and judgmental, and the 

discussion provided by Reinert et al. (1999) is reasonably complete and objective. It is worth 

noting, nonetheless, that the statistical re-analysis presented by Reinert et al. (1999) does indicate 

that the dose-response relationship for eggs/spawn has p values of 0.077 or 0.058, depending on 

whether standard or weighted regression is used. Although these values may be classified as 

‘insignificant’ using the standard cutoff p value of 0.05, the selection of this or any other p value 

is itself judgmental. 

The statistical analyses of these studies are open to reasonable debate; however, the Forest 

Service attempts to maintain a consistency with the U.S. EPA unless there is a compelling reason 

to do otherwise. For this risk assessment, there appears to be no compelling reason to deviate 

from the U.S. EPA assessment. Notwithstanding the reasonable arguments put forth by Reinert 

et al. (1999), the effect of tebufenozide on eggs/spawn is at least marginally significant. 

Furthermore, the use of solvent control data leads to more conservative assessments of risk in 

both the egg and fry study as well as the full life cycle study. While this may be coincidental, the 

consistency between the two studies suggests that the differences could be related to some factor 

that is not fully understood at this time. Consequently, this risk assessment treats 0.048 mg/L, 

the lowest concentration tested in the full life cycle study, as a LOAEL for fish reproduction. 

For this risk assessment, a LOAEL of 0.048 mg/L is adopted for chronic effects in fish. This 

interpretation of the study is identical to that of the U.S. EPA (1999e). The data are not sufficient 

to propose separate values for tolerant and sensitive species. 

4.3.3.2. Aquatic Invertebrates – Although data on the effects of tebufenozide on aquatic 

invertebrates is limited to three species (i.e, daphnids, midge larvae and lobsters as summarized 

in Appendix 8), variability is apparent regarding the acute toxicity of tebufenozide to aquatic 

invertebrates. Based on the available bioassays, the most sensitive species is the midge 

(Chironomus riparius) with an acute LC50 of 0.3 mg/L and an NOEC of 0.12 mg/L (van der Kolk 

1997). Daphnids appear to be much more tolerant, with an LC50 value of 3.8 mg/L and a 

4-21
 
corresponding NOEC of 0.82 mg/L (Graves and Smith 1992a). The apparent high sensitivity of 

midge relative to Daphnia may be related to differences in the types of bioassays that are run on 

midges (sediment assays) compared to those run on Daphnia (water only without sediment). The 

highest reported NOEC in lobsters is 0.1 mg/L (Dionne 1998). Because the study on lobsters 

was conducted at very low concentrations and no effects were seen at any concentration, there is 

no basis for asserting that lobsters are sensitive species. For this risk assessment, the acute 

NOEC values of 0.12 mg/L and 0.82 mg/L are used to assess the consequences of short-term 

exposures for sensitive and tolerant species of aquatic invertebrates. 

The midge is the most sensitive species for assessing the potential effects of chronic exposure. In 

the study by van der Kolk (1997), a concentration of 0.0053 mg/L caused a decrease in the larval 

emergence rate, and a concentration of 0.04 mg/L caused complete suppression of larval 

emergence. The NOAEL in this study is 0.0035 mg/L. Based on a standard 21-day reproductive 

study, Daphnia magna are substantially less sensitive with a reproductive NOEC of 0.029 mg/L 

and a corresponding LOEC of 0.059 mg/L (McNamara 1991). For this risk assessment, the 

longer-term NOEC values of 0.0035 mg/L and 0.029 mg/L are used to assess the consequences 

of longer-term exposures for sensitive and tolerant species of aquatic invertebrates. 

4.3.3.3. Aquatic Plants – As with fish and invertebrates, the available studies (Section 4.3.3.4 

and Appendix 8) suggest substantial differences in sensitivity among species of freshwater algae. 

For this risk assessment, risks to sensitive species are characterized using the lowest reported 

NOEC for algal growth of 0.077 mg/L in Scenedesmus subspicatus from the study by (Hoberg 

1992a). An over eight-fold higher NOEC of 0.64 mg/L has been reported for Selenastrum 

capricornutum (Reinert 1993b) and this value will be used to characterize risks in tolerant algal 

species. Although these tests are conducted for relatively short periods of time (i.e., 96 to 120 

hours), these NOEC values are applied to both acute and longer-term concentrations because of 

the short life-cycle of individual algal cells. 

4.3.3.4. Aquatic Microorganisms – Other than the information on algae provided above, there 

are no data regarding the toxicity of tebufenozide to aquatic microorganisms. Accordingly, no 

dose-response assessment is possible for this group. 

4-22
 
4.4. RISK CHARACTERIZATION 

4.4.1. Overview 

The use of tebufenozide to control the gypsy moth may result in adverse effects in nontarget 

Lepidoptera but there is little indication that other species will be impacted under normal 

conditions of use even at the highest application rate. Tebufenozide is an insecticide that is 

effective in controlling populations of lepidopteran pests. No data, however, are available on 

toxicity to nontarget Lepidoptera. For this risk assessment, the assumption is made that 

nontarget Lepidoptera may be as sensitive to tebufenozide as target Lepidoptera. Thus, adverse 

effects in nontarget Lepidoptera would be expected after applications that are effective for the 

control of lepidopteran pest species. 

There is no indication that short term exposures to tebufenozide will cause direct adverse effects 

in any terrestrial vertebrates or non-lepidopteran invertebrates even at the upper range of 

plausible exposures as well as accidental exposures. Similarly, direct adverse effects from longer 

term exposures in birds and mammals appear to be unlikely under most conditions. In some 

extreme cases, exposures in some large mammals could exceed the NOEC but the exposures 

would be below levels that have been associated with frank signs of toxicity. Effects on birds 

due to a decrease in available prey – i.e., terrestrial invertebrates – may be plausible. Adverse 

effects in aquatic species are not expected under normal conditions of use. In the case of a large 

accidental spill into a relatively small body of water, however, adverse effects could be expected 

in aquatic vertebrates, invertebrates, and plants. 

4.4.2. Terrestrial Organisms 

4.4.2.1. Terrestrial Vertebrates – The risk characterization for terrestrial vertebrates is 

summarized in Worksheet G02 for the maximum application rate of 0.12 lb/acre. The risk 

characterization is based on the estimates of exposure summarized in Section 4.2.3 and the 

toxicity values for diflubenzuron derived in Section 4.3.2.1 and summarized in Table 4.2. For 

most exposure scenarios, hazard quotients are included for both single applications and two 

applications spaced three days apart. For those exposure scenarios that do not include both 

single and double applications, the exposures are based on two applications 

None of the acute exposures result in hazard quotients that exceed the level of concern. The 

highest acute hazard quotient for any vertebrate is 0.04 – i.e., the consumption of contaminated 

fish by a fish-eating bird after an accidental spill – and this is below the level of concern by a 

factor of 20. Other more plausible exposure scenarios such as the consumption of contaminated 

vegetation and water are in the range of 0.000006 to 0.008, below the level of concern by factors 

of 125 to about 160,000. 

Similarly, for longer term exposures, central and lower estimates of hazard quotients are 

substantially below a level of concern. The highest central estimate for any hazard quotient is 0.1 

– i.e., below the level of concern by a factor of 10. At the upper ranges of exposure, however, 

the hazard quotient exceeds a level of concern for the consumption of contaminated vegetation 

on-site by a large mammal after either a single application (HQ=2) or two applications (HQ=4). 

4-23
 
As noted in the dose response assessment for mammals, the hazard quotients for mammals are 

based on a NOAEL of 1.8 mg/kg/day from the study by Richards (1992a) in which the 

corresponding LOAEL – based on toxic effects in the blood – of 20 mg/kg/day. Thus, a hazard 

quotient of 11 [20 mg/kg/day ÷ 1.8 mg/kg/day] would suggest a high likelihood of adverse 

effects in blood. The estimated hazard quotients of 2 to 4 are below this level where adverse 

effects would be expected but some changes in blood could occur although the toxicologic 

significance of these effects would most likely be marginal because the 20 mg/kg/day dose group 

in the study by Richards (1992a) did not display any overt signs of toxicity. Another factor to 

consider in interpreting these risk quotients is the proportion of the animal’s diet that is 

contaminated. The risk quotients for the consumption of contaminated vegetation that exceed the 

level of concern are all based on the assumption that 100% of the animal’s diet is contaminated. 

In other words, the animal consumes only vegetation that has been directly sprayed with 

tebufenozide. Thus, the potential impact of canopy interception is not considered. 

As discussed in Section 4.1.2.2 and detailed further in Appendix 6, the field study by Holmes 

(1998) noted suggestive effects on reproductive performance in Tennessee warblers – i.e., a 

decrease in the average number of eggs per nest and percent of eggs hatching. In addition, 

female warblers evidenced a decrease in brooding time and increase in foraging times, suggesting 

a decrease in prey availability. While the effects were not statistically significant, this study 

suggests that some birds may be impacted through a decrease in available prey secondary to the 

effects of tebufenozide on terrestrial invertebrates, as discussed further in Section 4.4.2.2. 

The verbal interpretation of these risk quotients is thus somewhat uncertain. There is no 

indication that short term exposures to tebufenozide will cause adverse effects in any terrestrial 

vertebrates even at the upper range of plausible exposures as well as accidental exposures. 

Similarly, adverse effects from longer terms exposures in birds and mammals appears to be 

unlikely under most conditions. In some extreme cases, exposures in some large mammals could 

exceed the NOEC but the exposures would be below the known LOEC. 

4.4.2.2. Terrestrial Invertebrates – Tebufenozide is an insecticide that is effective in 

controlling populations of lepidopteran pests. No data, however, are available on toxicity to 

nontarget Lepidoptera. For this risk assessment, the assumption is made that nontarget 

Lepidoptera may be as sensitive to tebufenozide as target Lepidoptera. Thus, applications of 

0.03 lb/acre are considered a LOEC based on the studies summarized in Table 4-3. As noted in 

Section 4.3.2.3, a NOEC for target and nontarget Lepidoptera cannot be identified. The USDA 

may use application rates as low as 0.015 lb/acre and these applications are presumably effective 

in the control of the gypsy moth. Under the assumption that nontarget Lepidoptera are as 

sensitive to tebufenozide as target species, adverse effects in nontarget Lepidoptera would be 

expected. 

Adverse effects in other insect species do not appear to be likely based on either the standard 

toxicity study in bees or the available field studies. As indicated in Worksheet G01, the hazard 

quotient for the direct spray of a bee is 0.08 at the maximum application rate of 0.12 lb/acre. 

4-24
 
Based on field studies, application rates of up to 0.24 lb/acre appear to have no adverse effect on 

beneficial arthropods. Using application rates, the highest hazard quotient would be 0.5 [0.12 

lb/acre ÷ 0.24 lb/acre]. Because effects on beneficial arthropods have not be examined at higher 

applications rates, the true NOEC for beneficial arthropods may be higher and perhaps 

substantially higher than 0.24 lb/acre. Consequently, the hazard quotient of 0.5 based on 

application rates is not inconsistent with the hazard quotient of 0.08 based on the honey bee 

toxicity bioassay. 

Toxicity data are also available on earthworms in which no effects were noted at soil 

concentrations of up to 1000 ppm (1000 mg/kg soil) (Section 4.3.2.3). As noted in Table 4-1, the 

peak concentration that would be expected in soil after two applications at a rate of 0.12 lb/acre 

is about 0.1 ppm, below the level of concern by a factor of 10,000. 

Thus, while the available data on nontarget terrestrial invertebrates are limited, it seems 

reasonable to assert that effected on nontarget lepidopterans are plausible at application rates that 

are effective in the control of target lepidopterans such as the gypsy moth. There is no basis for 

asserting that effects on other nontarget arthropods or other terrestrial invertebrates are plausible. 

4.4.2.3. Terrestrial Plants and Microorganisms – No quantitative risk assessment to terrestrial 

plants is made for tebufenozide. As discussed in Section 4.1.2.4, there are no data on the toxicity 

of this compound to either terrestrial plants or microorganisms. This lack of data, however, adds 

no substantial uncertainty to this risk assessment. Tebufenozide has been extensively tested in 

both the laboratory and field studies for efficacy in the protection of terrestrial plants from insect 

pests. If tebufenozide were toxic to plants at applications at or substantially above those used to 

control the gypsy moth, it is likely that reports of such phytotoxicity would be noted. No such 

reports have been encountered. 

4.4.3. Aquatic Organisms 

A summary of the risk quotients for aquatic organisms is presented in worksheet G03. Risk 

characterizations are presented for sensitive and tolerant species of aquatic organisms 

(vertebrates, invertebrates, and plants) for three exposure scenarios (an accidental spill, expected 

peak concentrations, and expected longer term concentrations of tebufenozide in water). The 

expected peak and longer term concentrations are summarized in Table 3-5 and discussed in 

Section 3.2.3.4.6. The concentrations associated with an accidental spill are calculated in 

Worksheet D05 and discussed in Section 3.2.3.4.1. The toxicity values used for each group of 

organisms are summarized in Table 4-2 and discussed in Section 4.3. 

The risk characterizations for each group of aquatic organisms are essentially identical. Under 

normal conditions of use at the highest anticipated application rate, no effects are expected in any 

group of organisms: vertebrates, invertebrates, or plants. In the case of an accidental spill, 

however, adverse effects would be expected in each group of organisms. 

4.4.3.1. Aquatic Vertebrates – Under normal conditions of use, the highest hazard quotient for 

4-25
 
sensitive species of fish is 0.1 – the hazard quotient associated with expected peak concentrations 

in water at the maximum anticipated application rate. The upper range of longer term 

concentrations in water are below a level of concern by a factor of about 33 (HQ=0.03). In the 

case of an accidental spill, however, the central estimate and the upper range of the hazard 

quotients exceeds a level of concern for both sensitive and tolerant species. As discussed in 

3.2.3.4.1, the accidental spill scenario is both extreme and arbitrary, involving the spill of a 

relatively large amount of chemical into a small body of water. 

4.4.3.2. Aquatic Invertebrates – Based on expected concentrations of tebufenozide in water 

under normal conditions of use, the upper ranges of the hazard quotients for sensitive aquatic 

invertebrates are 0.3 for short term peak concentrations and 0.4 for longer term concentrations. 

While these hazard quotients are somewhat higher than the corresponding hazard quotients for 

aquatic vertebrates, they are below a level of concern. In the case of an accidental spill, the 

concentrations in water exceed the level of concern for both sensitive and tolerant species of 

aquatic invertebrates. 

4.4.3.3. Aquatic Plants – The risk characterization for aquatic plants is based on bioassay data 

using algae. Because bioassay on algae are conducted only over relatively short periods of time – 

i.e., 96 to 120 hours – the toxicity values for both tolerant and sensitive species of algae are all 

essentially short term. As with both aquatic vertebrates and invertebrates, none of the expected 

concentrations in water exceed the level of concern for sensitive or tolerant species of algae even 

at the upper ranges of plausible exposures. Also as with aquatic vertebrates and invertebrates, 

the level of concern is exceed for both sensitive and tolerant species of algae in the case of an 

accidental spill. 

4-26
 
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registration of tebufenozide (Confirm 2F and 70WSP) for use on tree nuts and almond hull; DP 

Barcode: D255779. Memo from N.E. Federoff, M. Rexrode, and J. Cowles dated July 13, 1999. 

(Copy courtesy of Janet Bressant, Public Information and Records Integrity Branch, U.S. 

EPA/OPP). 

U.S. EPA (U.S. Environmental Protection Agency). 1999e. EFED risk assessment for Section 3 

registration of tebufenozide (Confirm 2F and 70WSP). Memo from M. Rexrode, and J. Cowles 

dated March 1, 1999. (Copy courtesy of Janet Bressant, Public Information and Records 

Integrity Branch, U.S. EPA/OPP). 

5-20
 
U.S. EPA (United States Environmental Protection Agency). 2004. IRIS Database. On-Line 

Search as of March 7, 2004. Address: http://www.epa.gov/iriswebp/iris/ index.html. 

USDA/APHIS. 1993. Nontarget Risk Assessment for the MEDFLY Cooperative Eradication 

Program. USDA Animal and Plant Health Inspection Service. February 1993. 

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Statement: Vegetation Management in the Coastal Plain/Piedmont, Management Bulletin R8­

MB-23, dated January, 1989. 1213 pp. 

USDA/FS (U.S. Department of Agriculture/Forest Service). 1989b. Draft Environmental Impact 

Statement: Vegetation Management in the Ozark/Ouachita Mountains, Management Bulletin R8­

MB-23, dated June, 1989. 499 pp. 

USDA/FS (U.S. Department of Agriculture/Forest Service). 1989c. Final Environmental Impact 

Statement: Vegetation Management in the Appalachian Mountains, Management Bulletin R8­

MB-38, dated July, 1989. 1104 pp. 

USDA (U.S. Department of Agriculture). 1995. Gypsy Moth Management in the United States: 

A Cooperative Approach. Final Environmental Impact Statement. Appendix F (Human Health 

Risk Assessment) and Appendix G (Ecological Risk Assessment). 

USGS (U.S. Geological Survey). 1998. Data on Pesticides in Surface and Ground Water of the 

United States., Results of the National Water Quality Assessment Program (NAWQA). Revised 

Oct. 23, 1998. http://water.wr.usgs.gov/pnsp/use92/tebufzd.html . 

U.S. Weather Service. 1998. NWS Climate Tables, 1961-1990. 

http://www.nws.noaa.gov/climatex.shtml. Files downloaded on January 14, 1999. 

van Hemmen JJ. 1992. Agricultural pesticide exposure data bases for risk assessment. Rev. 

Environ. Contam. Toxicol. 126: 1-85. 

Valentine BJ; Gurr GM; Thwaite WG 1996. Efficacy of the Insect Growth Regulators 

Tebufenozide and Fenoxycarb for Lepidopteran Pest Control in Apples and Their Compatibility 

with Biological Control for Integrated Pest Management. Aust. J. Exp. Agric. 36 (4):501-506. 

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Water/Sediment System: Final Report: Lab Project Number: 95RC-0196: 96-047-1007: 

1007.011.173. Unpublished study prepared by Spring. MRID No.44198301. 109 pp. 

van de Vrire M, Smagghe G, and Degheele D. 1996. Laboratory test method to evaluate the 

effect of 31 pesticides on the predatory bug Orius laevegatus (Het: Anthocoridae). 

Entomolphaga 41(2):235-243. Summarized in Keller and Brown 1998a. 

5-21
 
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in apples. Rohm and Hass Cooperator Report No. 2599521. Unpublished report summarized in 

Keller 1998. 

Webb RE; Peiffer R; Fuester RW; Thorpe KW; Calabrese L; Mclaughlin JM. 1998. An 

evaluation of the residual activity of traditional, safe, and biological insecticides against the 

gypsy moth. Journal of Arboriculture. 24(5): 286-292. 

Wederbrand K; Potter D. 1993. Carbon 14-RH-5992: Dermal Absorption in Male Rats: Final 

Report: Lab Project Number: 92R-183: 92P-183. Unpublished study prepared by Rohm and 

Haas Co. MRID No.42991510. 61 pp. 

West RJ; Thompson D; Sundaram KMS; Sundaram A; Retnakaran A; Mickle R 1997. Efficacy 

of aerial applications of Bacillus thuringiensis berliner and tebufenozide against Stern hemlock 

looper (Lepidoptera: Geometridae). Can. Entomol. 129(4): 613-626. 

Whiting DC; Jamieson LE; Connolly PG. 1999. Effect of sublethal tebufenozide applications on 

the mortality responses of Epiphyas postvittana (Lepidoptera: tortricidae) larvae exposed to a 

high-temperature controlled atmosphere. Journal of Economic Entomology. 92(2): 445-452. 

Wirth S; Vogel K. 1988. Cow's milk protein intolerance in infants with methemoglobinemia 

and diarrhea. Eur. J. Ped. (Berlin). 148(2): 172. 

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methemoglobin levels in residents living in industrial and nonindustrial communities. Environ. 

Res. 26(2): 347-352. 

5-22
 
Figure 2-1: Agricultural Use of Tebufenozide on Cotton in 1992 (USGS 1998). 

Figures-1 
Table 2-1. Selected physical and chemical properties of tebufenozide with selected additional 

properties for the commercial formulation Mimic. 

Appearance, ambient Mimic: off-white, cream color liquid. (C&P Press 2004) 

Tebufenozide, technical: while solid (Kelly 1992) 

Bioconcentration factor 151 in whole fish (Dong and Hawkins. 1993) 

16 in edible tissue (Dong and Hawkins. 1993) 

CAS number 112410-23-8 (C&P Press 2004; Kelly 1992) 

Commercial formulations Mimic 2LV; Confirm 2F 

EPA Registration Number 707-237 (Patel 1998) 

Foliar half-time (days) 2.8 to 13.3 days (Hawkins 1998) 

11.3 to 14 days (Kaminski 1997)
 

about 18.4 to 58.7 days (Sundaram et al. 1996a, Table 6, p. 725)
 

about 20 days (white spruce) (Sundaram et al. 1996b, )
 

Foliar wash-off fraction	 0.3 to 0.7 Sundaram et al. (1997b, Table 6, p. 514) 

0.2 to 0.8 Sundaram (1994b) 

log Ko/w 4.25 (Hawkins 1995) 

4.25 (SRC 1999)[Ko/w = 17,800] 

Molecular weight 352.48 (Patel 1998) 

pH 6.5-7.5 (C&P Press 2004) 

Photolysis (days) 98[soil surface] (Hawkins 1995) 

67[in aqueous solution] (Hawkins 1995) 

Soil half-time (days) 99 to 101[aerobic] (Hawkins 1995) 

66[aerobic] (Kaminski 1997 ) 

Soil sorption, Ko/c 572 (Hawkins 1995) 

Specific Gravity Mimic: 1.0 (C&P Press 2004) 

Synonyms 3,5-dimethyl-, 1-(1,1-dimethylethyl)-benzoic acid (C&P Press 2004) 

N-tert-butyl-N’-(4-ethylbenzoyl)-3,5-dimethylbenzol)hydrazide 

(Kaminski 1997) 

RH-5992 (Kelly 1992), Confirm 

Vapor pressure	 17 mm Hg @ 20°C/68°F (C&P Press 2004) 

2×10-8 torr at 25°C (Kaminski 1997) 

Volatility 60% (C&P Press 2004) 

Water solubility (mg/L) 0.83 (Kaminski 1997) 

Tables-1 
Table 3-1: Chemical and site parameters used in GLEAMS modeling for tebufenozide. 

Chemical Specific Parameters 

Parameter Clay Loam Sand Comment/ 

Reference 

Halftimes (days)

 Aquatic Sediment 179 U.S. EPA 1999e, p. 5

 Foliar 13.4 Note 1

 Soil 100 270 730 Note 2

 Water 67 Note 3 

Ko/c, mL/g 572 Note 4 

K , mL/g d 7.8 4.4 1.7 Note 5

Water Solubility, mg/L 0.83 Kaminski 1997 

Foliar wash-off fraction 0.5 Note 6 

Fraction applied to 0.8 

foliage 

Note 1	 Geometric mean of range of values from Table 2-1: 3 to 60 days. 

Note 2	 The soil half time for sand is taken as 730 days, the value used by U.S. EPA (1999e) in 

PRZM/EXAMS modeling. For clay, a soil halftime of 100 days is used (Hawkins 1995). As an 

intermediate value, the geometric mean of this range is used for loam. 

Note 3	 Photolysis halftime used by U.S. EPA 1999e from study by Hawkins 1995. 

Note 4	 This is taken from Hawkins (1995) and is identical to the value used by U.S. EPA (1999e) in the 

PRZM/EXAMS modeling 

Note 5	 Taken from U.S. EPA (1999e), Table 1, p. 6. 

Note 6	 Sundaram et al. (1997) have reported wash-off fractions 30% to 70% (Table 6, p. 514). Somewhat 

wider ranges, 20% to 80%, have been reported by Sundaram (1994b). For the GLEAMS modeling, a 

central value of 50% is used. 

Site Parameters 

(see SERA 2004b for details) 

Pond	 1 hectare pond, 2 meters deep, with a 0.01 sediment fraction. 10 hectare square field (1093' by 

1093') with a root zone of 12 inches. 

Stream	 Base flow rate of 710,000 L/day with a flow velocity of 0.08 m/second or 6912 meters/day. 

Stream width of 2 meters (about 6.6 feet'). 10 hectare square field (1093' by 1093') with a root 

zone of 12 inches. 

Tables-2 
Table 3-2: Summary of modeled concentrations of tebufenozide in streams (all units are μg/L 

or ppb) 

Annual Rainfall Clay Loam Sand 

Rainfall per Event 

(inches) (inches)1 Average Maximum Average Maximum Average Maximum 

Concentration per lb/acre applied (from GLEAMS) 

5 0.14 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

10 0.28 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

15 0.42 0.69713 19.95600 0.00878 0.29002 1.90923 52.54274 

20 0.56 1.68973 54.33504 0.06773 1.43491 5.30526 101.05556 

25 0.69 2.55255 91.00476 0.16814 3.12871 7.05234 111.28758 

50 1.39 4.09339 219.00699 0.77041 11.44738 6.85127 93.61309 

100 2.78 3.52070 317.12471 1.34698 30.36614 4.42689 88.43373 

150 4.17 2.70849 334.75298 1.35142 45.96028 3.16969 88.64864 

200 5.56 2.16187 320.13751 1.24326 55.46092 2.43988 87.51616 

250 6.94 1.78771 287.69153 1.12607 60.75455 1.97609 84.88519 

Application rate: 0.12 lbs/acre 

Concentration at above application rate 

5 0.14 0 0 0 0 0 0 

10 0.28 0 0 0 0 0 0 

15 0.42 0.083656 2.39472 0.00105 0.034802 0.2291076 6.3051288 

20 0.56 0.2027676 6.5202048 0.00813 0.1721892 0.6366312 12.126667 

25 0.69 0.306306 10.920571 0.020177 0.3754452 0.8462808 13.35451 

50 1.39 0.4912068 26.280839 0.092449 1.3736856 0.8221524 11.233571 

100 2.78 0.422484 38.054965 0.1616376 3.6439368 0.5312268 10.612048 

150 4.17 0.3250188 40.170358 0.1621704 5.5152336 0.3803628 10.637837 

200 5.56 0.2594244 38.416501 0.1491912 6.6553104 0.2927856 10.501939 

250 6.94 0.2145252 34.522984 0.1351284 7.290546 0.2371308 10.186223

1 Rain is assumed to occur at the same rate every 10th day – i.e., 36 rainfall events per year. 

Tables-3 
Table 3-3: Summary of modeled concentrations of tebufenozide in ponds (all units are μg/L 

or ppb) 

Annual Rainfall Clay Loam Sand
 

Rainfall per Event
 

(inches) (inches)1
 Average Maximum Average Maximum Average Maximum 

Concentration per lb/acre applied (from GLEAMS) 

5 0.14 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

10 0.28 0.00000 0.00000 0.00000 0.00000 0.00000 0.00000 

15 0.42 1.62583 3.41905 0.01831 0.04465 4.17974 8.26554 

20 0.56 3.01439 9.47016 0.10599 0.18515 8.82060 13.48834 

25 0.69 4.18885 16.64130 0.23102 0.36543 10.95654 15.44082 

50 1.39 7.25113 51.67100 0.93903 1.28274 11.29006 26.68412 

100 2.78 8.47509 103.59184 2.06369 6.79246 8.75309 39.33410 

150 4.17 7.95210 134.03042 2.47999 16.52847 7.16252 45.03134 

200 5.56 7.23386 157.87981 2.59791 25.60810 6.09099 47.50864 

250 6.94 6.58435 168.88316 2.59975 32.69145 5.32904 48.43668 

Application rate: 0.12 lbs/acre 

Concentration at above application rate 

5 0.14 0 0 0 0 0 0 

10 0.28 0 0 0 0 0 0 

15 0.42 0.1950996 0.410286 0.0022 0.00536 0.5015688 0.9918648 

20 0.56 0.3617268 1.1364192 0.012719 0.022218 1.058472 1.6186008 

25 0.69 0.502662 1.996956 0.027722 0.043852 1.3147848 1.8528984 

50 1.39 0.8701356 6.20052 0.1126836 0.1539288 1.3548072 3.2020944 

100 2.78 1.0170108 12.431021 0.2476428 0.8150952 1.0503708 4.720092 

150 4.17 0.954252 16.08365 0.2975988 1.9834164 0.8595024 5.4037608 

200 5.56 0.8680632 18.945577 0.3117492 3.072972 0.7309188 5.7010368 

250 6.94 0.790122 20.265979 0.31197 3.922974 0.6394848 5.8124016

1 Rain is assumed to occur at the same rate every 10th day – i.e., 36 rainfall events per year. 

Tables-4 
Table 3-4: Estimated Environmental Concentrations (μg/L or ppb) of tebufenozide in surface 

and groundwater at two applications of 0.12 lb a.i./acre (0.134 kg/ha), three days apart. 

Scenario	 Peak Long-Term Average 

MODELING FOR THIS RISK ASSESSMENT 

Direct Spray of Pond (Worksheet 6.73 N/A 

04b) 

Pond, drift at 100 feet (Worksheet 0.13 N/A 

04b) 

GLEAMS, Stream	 10 (0.03 to 40) 0.3 (0.001 to 0.8) 

GLEAMS, Pond	 5 (0.005 to 20) 0.5 (0.002 to 1.4) 

GENEEC Version 2, Pond 8.21 1.5 

[90 day value of 6.01 x 90/360] 

Sci-Grow 2.3, groundwater	 0.093 

OTHER MODELING 

U.S. EPA/OPP 8.7 ppb at 6x0.31 lb/ac 5.4 ppb at 6x0.31 lb/ac 

1999e.PRZM/EXAMS modeling of 

application to apples, Pond 

U.S. EPA/OPP 17 ppb at 4x0.25 lb/ac 8.2 ppb at 4x0.25 lb/ac 

1999e.PRZM/EXAMS modeling of 

application to cotton, Pond 

MONITORING STUDIES 

Sundarum et al. 1996a	 At an application rate of 2x0.070 kg/ha (0.062 lb/acre) with a 4 day 

interval. Peak stream concentrations of 1.32 ppb and peak pond 

concentrations of 5.31 ppb. Concentrations were below the limit of 

quantization limit of 0.04 μg/L by day 24 after application. 

Pond=300,000 liters in volume, 500 m2 surface area, 0.6 m deep. Stream 

width=2m, depth=20 cm, 7 m/min flow. 

Tables-5 
Table 3-5: Concentrations of tebufenozide in surface water used in this risk assessment (see 

Section 3.2.3.4.6 for discussion). 

At application rate: 0.12 lb/acre, 2 applications, 3 days apart 

Peak Concentration 

(ppb or μg/L) 

Longer Term Concentration 

(ppb or μg/L) 

Central 10 0.5 

Lower 0.005 0.002 

Upper 40 1.4 

Water contamination rate mg/L per lb/acre applied, 2 applications, 3 days apart 1 

Peak Concentration 

(mg/L per lb/acre) 

Longer Term Concentration 

(mg/L per lb/acre) 

Central 8.33e-02 4.17e-03 

Lower 4.17e-05 1.67e-05 

Upper 3.33e-01 1.17e-02

1 Water contamination rates – concentrations in units of mg/L expected at an application rate of 1 lb/acre. These 

values are entered into Worksheet B06a for diflubenzuron. This rate is adjusted to the program application rate in 

all worksheets involving exposure to contaminated water. 

Tables-6 
Table 4-1: Summary of modeled concentrations of tebufenozide in soil (all units are mg/kg or 

ppm), two applications spaced three days apart. 

Annual Rainfall Clay Loam Sand
 

Rainfall per Event
 

(inches) (inches)1
 Average Maximum Average Maximum Average Maximum 

Concentration per lb/acre applied (from GLEAMS) 

5 0.14 0.14894 0.33141 0.29680 0.49427 0.50678 0.84666 

10 0.28 0.15592 0.33655 0.31226 0.51438 0.51705 0.86709 

15 0.42 0.14905 0.33070 0.29440 0.48949 0.48422 0.79343 

20 0.56 0.14349 0.32703 0.29249 0.48803 0.43053 0.67757 

25 0.69 0.13746 0.32353 0.29102 0.48656 0.37176 0.57116 

50 1.39 0.10849 0.30803 0.27746 0.46370 0.20593 0.34765 

100 2.78 0.06705 0.27677 0.22935 0.39646 0.10536 0.28079 

150 4.17 0.04360 0.24522 0.19143 0.35247 0.07083 0.27603 

200 5.56 0.03094 0.21427 0.16493 0.32387 0.05381 0.27361 

250 6.94 0.02313 0.18274 0.14567 0.30341 0.04358 0.27084 

Application rate: 0.12 lbs/acre 

Concentration at above application rate 

5 0.14 0.017873 0.039769 0.035616 0.059312 0.060814 0.1015992 

10 0.28 0.01871 0.040386 0.037471 0.061726 0.062046 0.1040508 

15 0.42 0.017886 0.039684 0.035328 0.058739 0.058106 0.095212 

20 0.56 0.017219 0.039244 0.035099 0.058564 0.051664 0.081308 

25 0.69 0.016495 0.038824 0.034922 0.058387 0.044611 0.068539 

50 1.39 0.013019 0.036964 0.033295 0.055644 0.024712 0.041718 

100 2.78 0.00805 0.033212 0.027522 0.047575 0.012643 0.033695 

150 4.17 0.00523 0.029426 0.022972 0.042296 0.0085 0.033124 

200 5.56 0.00371 0.025712 0.019792 0.038864 0.00646 0.032833 

250 6.94 0.00278 0.021929 0.01748 0.036409 0.00523 0.032501

1 Rain is assumed to occur at the same rate every 10th day – i.e., 36 rainfall events per year. 

Tables-7 
Table 4-2: Summary of tebufenozide toxicity values used in ecological risk assessment 

Organism Endpoint Toxicity Value Reference, Species 

Mammals (Rats and Acute NOAEL, 1000 mg/kg Swenson and Solomon 1992 (rabbits) 

Rabbits) reproduction Hoberman 1991 (rats) 

Chronic NOAEL, 1.8 mg/kg/day Richards 1992a 

toxicity 

Birds (Bobwhite Quail) Acute NOAEL 2150 mg/kg Fletcher 1987 

Chronic NOAEL 15 mg/kg/day Beavers et al. 1993b 1 

Terrestrial Invertebrates 

Honey bee NOEC 2500 mg/kg Atkins (1990) and Chan (1995) 

Tolerant Insect Species NOEC 0.24 lb a..i. /acre Mulder and Prescott 1999a,b 

Sensitive Lepidoptera LOEC 0.03 lb Butler et al. (1997) 

a..i./acre 

Earthworm NOEC 1000 mg/kg Garvey (1992) 

soil 

Fish Acute 

Sensitive (Bluegills) NOEC 0.39 mg/L Graves and Smith (1992b) 

Tolerant (Trout) NOEC 1.9 mg/L Graves and Smith (1992c) 

Fish Chronic 

Sensitive/Tolerant LOEC, reproduction 0.048 mg/L Rhodes and Leak (1996) as interpreted by 

(Fathead Minnows) U.S. EPA (1999e) 3 

Aquatic Invertebrates, Acute 

Sensitive (Midge larvae) NOEC 0.12 mg/L van der Kolk (1997) 

Tolerant (Daphnids) NOEC 0.82 mg/L Graves and Smith (1992a) 

Aquatic Invertebrates, Chronic 

Sensitive (Midge larvae) NOEC, reproduction 0.0035 mg/L van der Kolk (1997) 

Tolerant (Daphnids) NOEC, reproduction 0.029 mg/L McNamara (1991) 

Aquatic Plants 

Sensitive (Scenedesmus NOEC for growth 0.077 mg/L Hoberg (1992a) 

subspicatus) 

Tolerant (Selenastrum NOEC for growth 0.64 mg/L Reinert (1993b)

capricornutum) 

1 Other studies are available indicating higher NOAELs. See 4.3.2.2 for discussion. 

2

 Other studies are available indicating no effects on tolerant invertebrates at application rates up to 0.25 lb/acre. 

See Table 4-3 and Section 4.3.2.3 for discussion. 3

 See Section 4.3.3.1 for a discussion of interpretation of studies. 

Tables-8 
Table 4-3: Summary of field studies on the effects of tebufenozide on terrestrial invertebrates 1 

Range of Species 

Application 

Rates (lb No Adverse Effects Adverse Effects a.i./acre) 

0.03 - <0.06 abundance of non-target arthropods other 

than macrolepidoptera (0.03 – Butler et al. 

1997) 

abundance of various macrolepidoptera (0.03 – 

Butler et al. 1997) 

0.06 - < 0.12 abundance of non-target arthropods other 

than macrolepidoptera (0.06 – Butler et al. 

1997) 

abundance of various macrolepidoptera (0.06 – 

Butler et al. 1997) 

spruce budworm (0.06 – Cadogan et al. 1997) 

0.12 - < 0.24 spiders, lacewings, and predatory mites 

(0.23 – Gurr et al. 1999) 

Mexican rice borer (0.12 and 0.18 – 

Legaspi et al. 1999) 

various beneficial arthropods* (0.125 – 

Mulder and Prescott 1999a) 

spruce budworm (0.12 – Cadogan et al. 1997) 

various lepidopteran pests (0.23 – Gurr et al. 

1999) 

beet armyworm (0.125 – Mulder and Prescott 

1999a) 

0.24 various beneficial arthropods (0.24 – 

Mulder and Prescott 1999a) 

beneficial arthropods (0.24 – Mulder and 

Prescott 1999b) 

beet armyworm (0.24 – Mulder and Prescott 

1999a) 

potato leafhopper (0.25 – Mulder and Prescott 

1999b)

1 Studies summarized in Appendix 6 with some efficacy studies omitted. The application rate in lb/acre and 

citation is given in parenthesis following the species or group. See text for discussion. A single asterisk (*) 

indicates transient or equivocal effects. 

Tables-9 
LIST OF APPENDICES
 

Appendix 1: Estimates of dermal absorption rates for tebufenozide 

Appendix 2: Oral toxicity of tebufenozide to experimental mammals 

Appendix 3: Dermal, inhalation, and ocular effects of tebufenozide in experimental mammals 

Appendix 4: Toxicity of tebufenozide to birds after oral administration 

Appendix 5: Toxicity to non-target terrestrial invertebrates 

Appendix 6: Terrestrial field studies 

Appendix 7: Toxicity of tebufenozide to fish 

Appendix 8: Toxicity of tebufenozide to aquatic invertebrates and algae 
Appendix 1: Estimates of dermal absorption rates for tebufenozide 

-1 Table A1-1: Estimate of first-order absorption rate (k in hours ) and 95% confidence a 

intervals. 

Model parameters 

Coefficient for ko/w 

Coefficient for MW 

ID 

C_KOW 

C_MW 

Value 

0.233255 

0.005657 

Model Constant C 1.49615 

Number of data points 

Degrees of Freedom (d.f.) 

Critical value of 0.025 t with 26 d.f.a 

DP 

DF 

CRIT 

29 

26 

2.056 

Standard error of the estimate SEE 16.1125 

Mean square error or model 

variance 

MDLV 0.619712 

Standard deviation of model (s) MSD 0.787218 MDLV0.5 

XNX, cross products matrix 0.307537 -0.00103089 0.00822769 

-0.00103089 0.000004377 -0.0000944359 

0.0082 -0.0000944359 0.0085286 

a

 Mendenhall and Scheaffer, 1973, Appendix 3, 4, p. A31. 

Central (maximum likelihood ) estimate: 

log k = 0.233255 log ( 10 a 10 k o/w ) - 0.005657 MW - 1.49615

95% Confidence intervals for log10 ka 

log k ± t × s × (aNXNX a)

0.5 

10 a 0.025 

where a is a column vector of {1, MW, log (k )}. 10 o/w 

NB: Although the equation for the central estimate is presented with ko/w appearing before MW 

to be consistent with the way a similar equation is presented by EPA, MW must appear first in 

column vector a because of the way the statistical analysis was conducted to derive XNX . 

See following page for details of calculating aNXNX a without using matrix arithmetic. 

Appendix 1-1 
Worksheet A07a (continued) 

Details of calculating aNXNX a 

The term a'A(X'X)

-1Aa requires matrix multiplication. While this is most easily accomplished 

using a program that does matrix arithmetic, the calculation can be done with a standard 

calculator. 

Letting 

a = {a_1, a_2, a_3} 

and

-1 (X'X) = {
 

{b_1, b_2, b_3},
 

{c_1, c_2, c_3},
 

{d_1, d_2, d_3}
 

},
 

a'A(X'X)

-1Aa is equal to 

Term 1: {a_1 ×([a_1×b_1] + [a_2×c_1] + [a_3×d_1])} + 

Term 2: {a_2 ×([a_1×b_2] + [a_2×c_2] + [a_3×d_2])} + 

Term 3: {a_3 ×([a_1×b_3] + [a_2×c_3] + [a_3×d_3])}. 

Appendix 1-2 
Appendix 1: Estimates of dermal absorption rates for tebufenozide (continued) 

Table A1-2: a Calculation of first-order dermal absorption rate (k ) for tebufenozide. 

Parameters Value Units Reference 

Molecular weight 352.48 g/mole Table 2-1 

Ko/w at pH 7 17,800 unitless Table 2-1 

10 o/w log K 4.25 

Column vector a for calculating confidence intervals (see Worksheet A07a for definitions.) 

a_1 1 

a_2 352.48 

a_3 4.25 

Calculation of a' A (X'X) A a - see Worksheet A07a for details of calculation. -1 

Term 1 -0.0209811072 

Term 2 0.0389710295 

Term 3 0.0475467644 

a' A (X'X) A a -1 0.0655 calculation verified in Mathematica 3.0.1.1 

10 a 10 o/w log k = 0.233255 log (k ) - 0.005657 MW - 1.49615 Worksheet A07a 

10 a log of first order absorption rate (k ) 

Central estimate -2.49869764236 ± 0.025 t × s × (a'A(X'X)

­

1

Aa)0.5 

Lower limit -2.91292499777 - 2.0560 × 0.787218 × 0.2559296778 

Upper limit -2.08447028695 � 2.0560 × 0.787218 × 0.2559296778 

First order absorption rates (i.e., antilog or 10 of above values). x 

Central estimate 0.003171775 hours-1 

Lower limit 0.001222011 hours-1 

Upper limit 0.008232462 hours-1 

Appendix 1-3 
Appendix 1: Estimates of dermal absorption rates for tebufenozide (continued)
 

Table A1-3: Estimate of dermal permeability (K in cm/hr) and 95% confidence
 p 

intervals. 

Model parameters 

Coefficient for ko/w 

Coefficient for MW 

ID 

C_KOW 

C_MW 

Value 

0.706648 

0.006151 

Model Constant C 2.72576 

Number of data points 

Degrees of Freedom (d.f.) 

Critical value of 0.025 t with 87 d.f.a 

DP 

DF 

CRIT 

90 

87 

1.96 

Standard error of the estimate SEE 45.9983 

Mean square error or model 

variance 

MDLV 0.528716 

Standard deviation of model (s) MSD 0.727129 MDLV0.5 

XNX, cross products matrix 0.0550931 -0.0000941546 -0.0103443 

-0.0000941546 0.0000005978 -0.0000222508 

-0.0103443 -0.0000222508 0.00740677 

a

Mendenhall and Scheaffer, 1973, Appendix 3, Table 4, p. A31. 

NOTE: The data for this analysis are taken from U.S. EPA (1992), Dermal Exposure 

Assessment: Principles and Applications, EPA/600/8-91/011B, Table 5-4, pp. 5-15 through 5-19. 

The U.S. EPA report does not provide sufficient information for the calculation of confidence 

intervals. The synopsis of the above analysis was conducted in STATGRAPHICS Plus for 

Windows, Version 3.1 (Manugistics, 1995) as well as Mathematica, Version 3.0.1.1 (Wolfram 

Research, 1997). Although not explicitly stated in the U.S. EPA report, 3 of the 93 data points 

are censored from the analysis because they are statistical outliers: [Hydrocortisone-21-yl]­

hemipimelate, n-nonanol, and n-propanol. The model parameters reported above are consistent 

with those reported by U.S. EPA but are carried out to a greater number of decimal places to 

reduce rounding errors when calculating the confidence intervals. See notes to Worksheet A07a 

for details of calculating maximum likelihood estimates and confidence intervals. 

Appendix 1-4 
Appendix 1: Estimates of dermal absorption rates for tebufenozide (continued) 

Table A1-4 p : Calculation of dermal permeability rate (K ) in cm/hour for tebufenozide. 

Parameters Value Units Reference 

Molecular weight 352.48 g/mole 

Ko/w at pH 7 17800 unitless 

10 o/w log K 4.25 

Column vector a for calculating confidence intervals (see Worksheet A07a for definitions.) 

a_1 1 

a_2 352.48 

a_3 4.25 

Calculation of a' A (X'X) A a - see Worksheet A07b for details of calculation. -1 

Term 1 -0.0220577884 

Term 2 0.007751756 

Term 3 0.0564889197 

a' A (X'X) A a -1 0.0422 calculation verified in Mathematica 3.0.1.1 

10 p 10 o/w log k = 0.706648 log (k ) - 0.006151 MW - 2.72576 Worksheet A07b 

log10 of dermal permeability 

Central estimate -1.89061048 ± 0.025 t × s × a'A(X'X) Aa -1 0.5 

Lower limit -2.18337858572 - 1.9600 × 0.727129 × 0.2054263858 

Upper limit -1.59784237428 � 1.9600 × 0.727129 × 0.2054263858 

Dermal permeability 

Central estimate 0.0128644 cm/hour 

Lower limit 0.0065557 cm/hour 

Upper limit 0.025244 cm/hour 

Appendix 1-5 
Table A1-5: Summary of chemical specific dermal absorption values used for tebufenozide 

dermal absorption. 

Description Code Value Units Reference/Source 

First-order absorption rates (k ) a

Central estimate AbsC 0.0032 hour-1 Table A1-2, values rounded 

to two significant figures Lower limit AbsL 0.0012 hour-1 

Upper limit AbsU 0.0082 hour-1 

Zero-order absorption (K ) p

Central estimate KpC 0.013 cm/hour Table A1-4, values rounded 

to two significant figures Lower limit KpL 0.0066 cm/hour 

Upper limit KpU 0.025 cm/hour 

Appendix 1-6 
Appendix 2: Oral toxicity of tebufenozide to experimental mammals (subdivided as acute, subchronic, chronic 

and reproductive toxicity and sorted by species/duration). 

Animal Dose/Exposure Response Reference 

ACUTE 

Mice (NOS) >5.0 g/kg technical, 

single oral dose 

(NOS) 

No treatment related mortalities or signs of 

toxicity at limit dose of 5.0 g/kg 

LD50>5.0 g/kg 

Hazleton and Quinn 

1995b 

MRID 43781708 

Rats, Crl:CD, 

29 to 34-days 

old, weighing 

73-101 g, 10 

males and 10 

0, 500, 1000, or 2000 

mg/kg bw by gavage 

(single dose) 

No treatment-related mortalities, clinical 

signs of toxicity, or effects on body weight at 

any dose level; no neurotoxic or 

neuropathological effects at any dose level. 

Swenson et al. 1994 

MRID 43781706 

females per 

dose group 

NOEL >2000 mg/kg bw (highest dose tested) 

Rats, CD, 

adults, 6 males 

and 6 females 

single gavage dose of 

5.0 g/kg bw Mimic® 

240 LV 

No mortalities, body weight effects, or 

clinical signs of toxicity. 

Parno and Gingrich. 

1994b 

MRID 44727702 

Acute oral LD50 >5.0 g/kg bw or 5000 mg/kg 

This study reveals the components of Mimic 

formulation. This information cannot be 

disclosed in this document. 

Rats (NOS) >5.0 g/kg technical, 

single oral dose 

(NOS) 

“practically non-toxic;” no treatment-related 

mortalities or signs of toxicity at the limit 

dose of 5.0 g/kg 

LD50>5.0 g/kg 

Hazleton and Quinn 

1995b 

MRID 43781708 

(This appears to be 

a summary of Parno 

and Gingrich 

1994b, detailed 

above) 

SUBCHRONIC 

Dogs, 4 males 

and 4 females 

per dose group 

(NOS) 

0, 150, 600, 2400, or 

9600 ppm ai in diet 

for 2 weeks 

No effects on body weight or food 

consumption and no clinical or gross 

observations of toxicity. 

No effects at 150 ppm ai (5.1 mg/kg bw/day) 

At $600 ppm ai, increased spleen weight was 

noted; at $2400 ppm ai, increased spleen-tobody weight ratio was noted; at 9600 ppm ai, 

additional adverse effects included decreased 

Hazleton and Quinn 

1995b 

MRID 43781708 

(Hazard 

evaluation/data 

summary) 

RBC, hemoglobin, and hematocrit values. 

Appendix 2-1 
Appendix 2: Oral toxicity of tebufenozide to experimental mammals (subdivided as acute, subchronic, chronic 

and reproductive toxicity and sorted by species/duration). 

Animal Dose/Exposure Response Reference 

Dogs, one male 

and one female 

per dose group 

(NOS) 

Dogs, males, 4 

per dose group 

(NOS) 

Dogs, beagles, 

purebred, -8­

months old, 4 

males and 4 

females per 

dose group 

limit dose of 30,000 

ppm ai (1000 mg/kg 

bw/day) in diet for 2 

weeks 

0 or 1500 ppm ai 

technical for 6 

weeks, followed by 

control diet (0 ppm) 

for additional 4 

weeks; hematological 

parameters were 

measured in controls 

and treated dogs 

prior to treatment, at 

6 weeks, at 8 weeks, 

and at 10 weeks 

oral administration 

by admixture of 0, 

50, 500, or 5000 ppm 

(active ingredient) 

for 90 days; group 

mean compound 

consumption in 

mg/kg/day for 13 

weeks was: 2.09, 

20.13, or 202.42 

mg/kg/day 

(FEMALES) and 

2.05, 21.42, or 

201.82 mg/kg/day 

(MALES) 

decrease in food consumption during week 1 

but not week 2 (both sexes); decreased body 

weight (male), hematological effects (both 

sexes) included decreased RBC, hemoglobin, 

and hematocrit values, increased 

methemoglobin (females), reticulocytes, 

Heinz bodies, platelets and white blood cells. 

Treatment-related effects included increased 

bilirubin and other changes in serum 

chemistry (NOS) and increased spleen 

weights above the upper limit expected for 

this species. 

Limit dose of 30,000 ppm was considered 

too high to be used in 13-week study. 

Study designed to examine reversibility of 

hematological effects after exposure to RH5992 technical. 

After 6 weeks, hematological effects in 

treated dogs included decreases in RBC, 

hemoglobin, and hematocrit values; increases 

in methemoglobin, mean corpuscular 

volume, reticulocytes, and platelets. 

Complete recovery (i.e., effects on 

hemopoietic system returned to control 

values) by the end of the 2- or 4-week 

recovery period. 

Dietary concentrations of 500 or 5000 ppm 

had a direct effect on red blood cells, leading 

to low grade hemolytic anemia. 

NOEL = 50 ppm 

No clinical signs of toxicity were attributed 

to treatment; high dose males gained slightly 

less weight than controls but the difference 

was not statistically significant; high dose 

males and females ate slightly less food than 

controls but the difference was not 

statistically significant; treatment had no 

effect on food conversion efficiency; and no 

ocular lesions resulted from treatment. 

Hazleton and Quinn 

1995b 

MRID 43781708 

(Hazard 

evaluation/data 

summary) 

Hazleton and Quinn 

1995b 

MRID 43781708 

(Hazard 

evaluation/data 

summary) 

Clay 1992 

MRID 42436223 

Appendix 2-2 
Appendix 2: Oral toxicity of tebufenozide to experimental mammals (subdivided as acute, subchronic, chronic 

and reproductive toxicity and sorted by species/duration). 

Animal Dose/Exposure Response Reference 

Additional Observations from Clay 1992 MRID 42436223: 

Hematology: there were several statistically significant effects on hematological parameters (e.g., red blood cell 

count, mean cell volume, reticulocyte counts, methemoglobin, incidence of Heinz bodies, and platelet counts) in 

males and females exposed to 500 or 5000 ppm. The presence of Heinz bodies is considered to represent a 

direct effect on the RBC and led to increased destruction of RBC in liver and spleen. 

Urinalysis: urine of treated males was darker than urine of controls in week 13; three high dose males had 

bilirubin present in their urine (consistent with destruction of red blood cells). 

Organ weights: in high dose males, mean absolute spleen weight was 30% greater than that of controls (p#

0.05) and relative spleen weight was 44% greater (p# 0.01); in females there was a significant dose response in 

relative spleen weight (p# 0.05); no statistically significant differences in relative liver weight among treated 

dogs; in high dose females, there was a statistically significant dose response with respect to increased liver 

weight. 

Various treatment-related effects indicative of low grade hemolytic anemia were observed in the liver (increased 

incidence of pigment in the Kupffer cells), spleen (increased hemopoiesis and increased sinusoidal engorement) 

and bone marrow (hyperplasia) of males and female exposed to 500 or 5000 ppm. 

Mice, males, 8 0, 60, 200, 600, 2000 No effects at #600 ppm; increased liver-to- Hazleton and Quinn 

per dose group or 6000 ppm ai body weight ratio at 2000 or 6000 ppm; 1995b 

(NOS) technical in diet for 2 increased liver weight at 6000 ppm (-1000 MRID 43781708 

weeks mg/kg bw/day); no adverse effects on (Hazard 

survival, clinical chemistry, body weight or evaluation/data 

food consumption. summary) 

Mice, Crl:CD­ 0, 20, 200, 2000 or No mortality; no treatment related clinical, Osheroff 1991a 

1, -4-weeks 20,000 ppm in the cageside, or ophthalmoscopic observations. MRID 42436221 

old, 10 males diet for 13 weeks Body weight: significantly decreased mean 

and 20 females body weight values at weeks 0-13 in males at 

per dose group 200 or 2000 ppm and at weeks 0-4 and 0-13 

in males at 20,000 ppm; no statistically 

significant differences in mean food 

consumption values among all dose groups. 

Appendix 2-3 
Appendix 2: Oral toxicity of tebufenozide to experimental mammals (subdivided as acute, subchronic, chronic 

and reproductive toxicity and sorted by species/duration). 

Animal Dose/Exposure Response Reference 

Additional Notes on Osheroff 1991a MRID 42436221 

Hematology: significant increases in reticulocyte and absolute reticulocyte counts (males and females at 2000 or 

20,000 ppm), mean cell volume (males at 2000 or 20,000 ppm), mean cell hemoglobin (males and females at 

2000 or 20,000 ppm), mean cell hemoglobin concentration (males at 2000 and males and females at 20,000 

ppm), white blood cell count, corrected white blood cell count, and lymphocyte counts (females at 2000 ppm 

and males and females at 20,000 ppm), heinz bodies (males at 2000 ppm and males and females at 20,000 ppm), 

and segmented neutrophils (males at 2000 ppm and males and females at 20,000 ppm). Decreased erythrocyte 

counts in males and female at 2000 or 20,000 ppm (significant only in males), decreased myeloid/erythroid 

ratios in males and female at 2000 or 20,000 ppm (significant only in females), significant increases in 

methemoglobin values in males and females at 2000 or 20,000 ppm, significant increased mean alkaline 

phosphatase and potassium values in males at 2000 or 20,000 ppm and significantly increased mean total protein 

and calcium values in males at 20,000 ppm. 

Organ weights: significant decrease in mean terminal body weight in males at 20,000 ppm, significantly 

increased mean absolute and relative liver and spleen weights in males and 2000 ppm and in males and females 

at 20,000 ppm. 

Gross necropsy: increased incidence in enlarged spleen males and females at 2000 or 20,000 ppm, increased 

incidence or severity of pigment accumulation in liver, spleen and kidney as well as increased extramedullary 

hematopoiesis in spleen of males and females at 2000 or 20,000 ppm. 

Rats, 6 males 0, 50, 250, 1000, No effects at 50 or 250 ppm Hazleton and Quinn 

and 6 females 2500, or 10,000 ppm target organ = hemopoietic system 1995b 

per dose group ai technical in diet MRID 43781708 

(NOS) for 2 weeks at 1000 ppm, observations included (Hazard evaluation) 

decreased RBC (females), hemoglobin 

(females), and hematocrit (both sexes); 

increased liver weight (females) and liver-tobody weight ratio (both sexes). 

at 2500 ppm, additional effects included 

increased spleen weight (females) and 

spleen-to-body weight ratio (females) 

at 10,000 ppm (-700 mg/kg/day), additional 

effects included decreased food 

consumption, body weight (males), RBC 

(males), and hemoglobin (males); increased 

spleen weight (males) and spleen-to-body 

weight ratio (males). 

Effects at higher doses generally more severe 

than those observed at lower doses; no 

effects on survival or body weight (females), 

and no clinical signs of toxicity or gross 

pathology 

Appendix 2-4 
Appendix 2: Oral toxicity of tebufenozide to experimental mammals (subdivided as acute, subchronic, chronic 

and reproductive toxicity and sorted by species/duration). 

Animal Dose/Exposure Response Reference 

Rats, 10 males 0 or 20,000 ppm ai in 

and 10 females diet for 4 weeks; 

(NOS) (20,000 ppm 

approximates limit 

dose of 1000 

mg/kg/day) 

Decreases observed in body weight, body 

weight gain, food consumption, RBC, 

hemoglobin, and hematocrit. Males showed 

increased liver and spleen weights (absolute 

and relative to body weight). There were no 

effects on survival and no clinical or gross 

signs of toxicity. 

This study together with the 2-week range 

finding test was used to select doses for the 

13-week study. 

Hazleton and Quinn 

1995b 

MRID 43781708 

(Hazard Evaluation 

and toxicity 

summary) 

Appendix 2-5 
Appendix 2: Oral toxicity of tebufenozide to experimental mammals (subdivided as acute, subchronic, chronic 

and reproductive toxicity and sorted by species/duration). 

Animal Dose/Exposure Response Reference 

Rats, CD, -4- 0, 20, 200, 2000, or 

weeks old, 10 20,000 ppm in diet 

males and 10 for 13 weeks 

females per 

dose group 

No mortality; no adverse neurobehavioral, 

clinical, ophthalmoscopic, or gross necropsy 

findings. 

Body weight: statistically significant 

decrease at weeks 4 and 13 in females at 

2000 ppm and in males and females at 

20,000 ppm; body weight gain values 

significantly decreased at weeks 0-4 and 0­

13 in males and females at 2000 or 20,000 

ppm; food consumption significantly 

decreased at weeks 1-4 in males and females 

at 2000 or 20,000 ppm. 

Hematology: significant decreases in mean 

erythrocyte count, hemoglobin , and mean 

cell hemoglobin values as well as significant 

increases in mean cell volumes in males and 

females at 2000 or 20,000 ppm; decreased 

hematocrit and platelet values and increased 

mean cell hemoglobin and reticulocyte 

values in 20,000 ppm females; decreased 

myeloid/erythroid ratio in 2000 ppm females 

(with slight but not significant decrease in 

males and females at 20,000 ppm); 

significant increases in mean glucose and 

globulin values in females at 20,000 ppm. 

Organ weights: significantly decreased 

terminal body weight value for females at 

2000 ppm and for males and females at 

20,000 ppm; increased absolute liver weight 

in females at 20,000 ppm; increased spleento-body weight values in males and females 

at 20,000 ppm; increased liver-to-body 

weight values in females at 2000 ppm and 

males and females at 20,000 ppm; increased 

liver-to-brain weight value in females at 

2000 or 20,000 ppm. 

Histomorphology: increased severity of 

splenic pigmentation in males and females at 

2000 or 20,000 ppm. 

NOEL (dietary administration for 13 

weeks) = 200 ppm 

Osheroff 1991b
 

MRID 42436219
 

MRID 43781708
 

(data summary)
 

Appendix 2-6 
Appendix 2: Oral toxicity of tebufenozide to experimental mammals (subdivided as acute, subchronic, chronic 

and reproductive toxicity and sorted by species/duration). 

Animal Dose/Exposure Response Reference 

CHRONIC 

Dogs, beagles, oral administration No clinical signs of toxicity associated with Richards 1992a,b 

purebred, 6- to by admixture of 0, treatment; no adverse effects at #50 ppm; MRID 42931203 

7-months old, 15, 50, 250, or 1500 slight reduction in body weight gain (in the MRID 42931204 

weighing: 7.00­ ppm for 52 weeks. absence of any effect on food consumption) 

10.55 kg in males at 1500 ppm. 

(males) and Based on measured 

5.75-9.05 kg food consumption, At 250 and 1500 ppm, a direct effect of 

(females), 4 these dietary treatment on red blood cells was indicated by 

males and 4 concentrations the presence of Heinz bodies and an increase 

females per corresponded to in levels of methemoglobin, which resulted 

dose group doses of 0.4 to 0.7 in the increased destruction of red blood 

mg/kg bw (15 ppm), cells in the liver (histologically associated 

1.5 to 2.4 mg/kg bw with an increase in Kupffer cell pigment) and 

(50 ppm), 6.4 to 11.3 spleen. The increased destruction of red 

mg/kg bw (250 ppm), blood cells most likely accounted for the 

and 42.8 to 71.1 statistically significant increase in liver/body 

mg/kg bw (1500 weight ratio in males at 1500 ppm and the 

ppm) increased spleen weights in dogs exposed to 

250 and 1500 ppm. Also consistent with the 

effect of increased red blood cell destruction 

is the increase in plasma bilirubin at 250 and 

1500 ppm. 

Additional Notes on Richards 1992a,b: 

Other adverse effects included decreases in red blood cell counts, hemoglobin concentrations, and packed cell 

volume, compensatory increased in red blood cell production, minimal hemopoiesis in the spleen and 

hyperplasia in the sternal and femoral bone marrow, and increases in platelet and reticulocyte counts. All of 

these effects, which were observed consistently at 1500 ppm and to a lesser extent at 250 ppm, are indicative of 

low grade hemolytic anemia. 

The increase in methemoglobin levels evidenced a statistically significant dose-response relationship at weeks 

13, 15, 21, 39, and 52. [Table 5.1, p. 86. Fiche of this table is very difficult to read. Durations are taken from 

section 3.7, p. 23.] Based on comparisons to the control group, however, only the high dose group male dogs 

had a statistically significant increase by the end of the study, 1.7% in exposed group compared to 0.9% in the 

control group. 

Appendix 2-7 
Appendix 2: Oral toxicity of tebufenozide to experimental mammals (subdivided as acute, subchronic, chronic 

and reproductive toxicity and sorted by species/duration). 

Animal Dose/Exposure Response Reference 

Mice, Crl:CD1, -6-weeks 

old, weighing 

23-33 g (males) 

and 17-26 g 

(females), 60 

males and 60 

females per 

dose group 

Rats, CRL:CD, 

-6-weeks old, 

70 males and 

70 females per 

dose group 

nominal dietary 

concentrations of 0, 

5, 50, 500, or 1000 

ppm ai for 18 

months, 

corresponding to 

overall compound 

consumption of 1, 8, 

78, or 155 mg/kg/day 

(males) or 1, 9, 94, or 

186 mg/kg/day 

(females). 

0, 10, 100, 1000, or 

2000 ppm in diet for 

24 months (interim 

sacrifice at 12 

months); overall 

compound 

consumption values 

for males: 0.5, 5, 48, 

or 97 mg/kg/day, and 

for females: 0.6, 6, 

61, or 125 mg/kg/day 

NOEL = 50 ppm 

[8 mg/kg/day (males) and 9 mg/kg/day 

(females). 

No oncogenic effects at dietary levels up to 

1000 ppm (equivalent to intake of 155 and 

186 mg/kg/day for males and females, 

respectively);

 no adverse effects on body weight, body 

weight gain, food consumption, or food 

efficiency; treatment related effects 

indicative of chronic toxicity included 

hematological changes and spleen 

histopathology at 500 or 1000 ppm. 

Decreased survival in males at 500 and 1000 

ppm and in females at 1000 ppm was judged 

to be an equivocal finding based on historical 

control data and lack of associated 

pathologies. 

no treatment related effect on survival; no 

oncogenic effects; treatment-related effects 

indicative of chronic toxicity at 1000 or 2000 

ppm included decreased mean body weight 

and body weight gains, hematological effects 

(e.g., decreases in mean erythrocyte count, 

hematocrit and hemoglobin counts), and 

spleen histopathology (e.g., statistically 

significant increase in spleen-to-body weight 

ratio in high dose females, likely related to 

hematology findings). 

NOEL = 100 ppm (5 and 6 mg/kg/day for 

males and females, respectively) 

Trutter 1992a 

MRID 42931205 

Trutter 1992b 

MRID 42931206 

Trutter 1992c 

MRID 42931208 

Appendix 2-8 
Appendix 2: Oral toxicity of tebufenozide to experimental mammals (subdivided as acute, subchronic, chronic 

and reproductive toxicity and sorted by species/duration). 

Animal Dose/Exposure Response Reference 

REPRODUCTION/TERATOLOGY
 

Rabbits, New 

Zealand white, 

pregnant 

females, 5.5- to 

6-months old, 

20 per dose 

group 

Rats, SpragueDawley, 

pregnant 

females, 25 per 

dose group. 

Rats, Crj:CD, 

-5-weeks old, 

24 males and 

24 females per 

dose group 

0, 50, 250, 1000 

mg/kg/day once daily 

by gavage on day 7­

19 of gestation; 

vehicle: aqueous 

0.5% (w/w) sodium 

carboxymethylcellulose 

0, 50, 250, or 1000 

mg/kg/day once daily 

by gavage on days 6­

15 of gestation; 

vehicle: aqueous 

0.5% (w/w) sodium 

carboxymethylcellulose 

0, 25, 200, or 2000 

ppm in diet for two 

consecutive 

generations 

No treatment-related deaths or clinical signs Swenson and 

of toxicity; no treatment-related effects on Solomon 1992 

maternal body weight or food consumption; MRID 42436227 

no signs of maternal or developmental 

toxicity at any dose level. 

NOEL = 1000 mg/kg/day (highest dose 

tested) 

No mortality; no clinical toxicity or adverse Hoberman 1991 

findings at necropsy. MRID 42436225 

At 1000 mg/kg/day: reduced maternal body 

weight gain on days 6-20 of gestation (after 

correction for gravid weight); decrease in 

relative food consumption on days 7-8 and 6­

9 of gestation, significantly reduced (p#0.05) 

on days 8-9 of gestation. 

No effects on litter averages for corpora 

lutea, implantations, sitter sizes, live fetuses, 

early and late resorptions, or the number of 

dams with any resorptions. No 

developmental effects occurred at the high 

(1000 mg/kg/day) dose. 

NOAEL = 250 mg/kg/day. 

no reproductive effects at concentrations Aso 1995 

#2000 ppm MRID 43797701 

systemic toxicity observed in parental rats Hazleton and Quinn 

(i.e., adverse effects on hemopoietic system 1995b 

and body weight effects) at concentrations MRID 43781708 

$200 ppm (Hazard evaluation 

and data summary) 

NOEL (for reproductive effects) = 2000 ppm 

ai (149-195 mg/kg/day in males and females, 

respectively) 

NOEL (for systemic toxicity) = 25 ppm ai 

(1.9-2.3 mg/kg/day for males and females, 

respectively) 

Appendix 2-9 
Appendix 2: Oral toxicity of tebufenozide to experimental mammals (subdivided as acute, subchronic, chronic 

and reproductive toxicity and sorted by species/duration). 

Animal Dose/Exposure Response Reference 

Rats, CRL:CD, 0, 10, 150, or 2000 NOEL (for reproductive effects) = 150 ppm Danberry et al. 

-6-weeks old, ppm in diet through (11.5-13.6 mg/kg/day for males and 12.8­ 1993 

25 males and two generations 14.5 mg/kg/day for females) MRID 42931207 

25 females per 

dose group Hazleton and Quinn 

1995a 

MRID 43781707 

Additional Details from Danberry et al. 1993: No treatment related mortality or clinical signs of toxicity in 

any generation at any dose level; #150 ppm did not cause effects on body weights or food consumption in any 

generation; 2000 ppm caused a decrease in body weight and food consumption in P1 and P2 males; 

histopathological changes in the spleen and toxicity of the hemopoietic system in rats of both sexes from both 

generations were consistent with the general pattern of toxicity observed in other non-developmental/nonreproductive studies 

There were no treatment-related effects on mating or fertility in either generation at any dose level; there were 

no treatment related effects on reproduction in either generation at 10 or 150 ppm; at 2000 ppm, there was an 

increased incidence of mortality of females during delivery (P ), an increase in gestation length (P ), a 2 2

decrease in the mean number of implantation sites per female (P ), and an increased incidence (equivocal) 2

of pregnant females that did not deliver (P and P ). 1 2 

There were no treatment related effects on any offspring with respect to body weights, viability, malformations, 

or variations. 

Hazleton and Quinn 1995a (MRID 43781707) conclude that dietary concentrations #2000 ppm tebufenozide do 

not cause reproductive effects in rats; NOEL = 149-195 mg/kg/day in males and females, respectively; NOEL 

for toxicity = 25 ppm (1.9-2.3 mg/kg/day in males and females, respectively 

Appendix 2-10 
Appendix 3: Dermal, inhalation, and ocular effects of tebufenozide in experimental mammals 

Animal Dose/Exposure Response Reference 

DERMAL 

Rats, CD, adults, 

6 males and 6 

females 

Rats (NOS) 

Rats, CD, adults, 

6 males and 6 

females 

Rats, 10 males 

and 10 females 

per dose group 

(NOS) 

2.0 g/kg bw 

undiluted 

Mimic®240 LV 

applied to shaved 

intact skin and 

occluded for 24 

hours, after which the 

application sites were 

wiped with paper 

towels saturated with 

tap water and blotted 

dry. 

5.0 g/kg technical, 

single dermal 

application 

5000 mg/kg bw 

undiluted 

Mimic®240 LV 

applied to shaved 

intact skin and 

occluded for 24 

hours, after which the 

application sites were 

wiped with paper 

towels saturated with 

tap water and blotted 

dry. 

0 or 1000 mg ai/kg 

bw/day semiocclusive 6-hour 

dermal exposure, 5 

days/week for 4 

weeks or 0, 62.5, 

250, or 1000 mg 

ai/kg bw/day. 

No mortalities, clinical signs of toxicity, or 

body weight effects. Red stains observed on 

the fur surrounding the eyes and muzzle of 

several animals were attributed to test 

methods and use of collars. Skin irritation, 

manifested as erythema, edema, dessication, 

and scabs, was observed; however, necropsy 

revealed no gross changes. 

Acute dermal LD50 >2.0 g/kg bw 

Rohm and Haas classifies the test formulation 

as “PRACTICALLY NON-TOXIC by single 

dermal exposure” 

This study reveals the components of in the 

formulation. This information cannot be 

released 

“practically non-toxic;” no treatment-related 

mortalities or signs of toxicity at limit dose of 

5.0 g/kg 

LD50>5.0 g/kg 

No mortalities, clinical signs of toxicity, or 

body weight effects. 

Dessication at the application site affected 

several of the animals beginning on day 3 and 

continuing until day 9; necropsy revealed no 

gross changes. 

Acute dermal LD50 >5000 mg/kg bw 

Rohm and Haas classifies the test formulation 

as “PRACTICALLY NON-TOXIC by single 

dermal exposure” 

This study reveals the components of in the 

formulation. This information cannot be 

released 

NOEL (dermal application for 4 weeks) = 

1000 mg ai/kg bw/day 

No treatment-related effects on hematology or 

clinical chemistry parameters, organ weights, 

gross pathology or histopathology at any dose 

level 

Parno and 

Gingrich 1994a 

MRID 44727703 

Hazleton and 

Quinn 1995b 

MRID 43781708 

(Hazard 

evaluation/Toxicity summary) 

Parno 1997 

MRID 44727704 

Hazleton and 

Quinn 1995b 

MRID 43781708 

(Hazard 

Evaluation/data 

summary) 

Appendix 3-1 
 

Appendix 3: Dermal, inhalation, and ocular effects of tebufenozide in experimental mammals 

Animal Dose/Exposure Response Reference 

Rats, Crl:CD, 

adults, 6 males 

and 6 females per 

dose group 

Rabbits, New 

Zealand white, 

adults, 6 males 

Guinea pigs, 

Hartley, young 

females, 20 

treated, 10 

positive controls, 

10 naive controls 

Daily dermal 

applications of RH75,992 2F 

formulation and 

RH75,992 technical 

or skin of rats for 4 

weeks at doses up to 

and including 1000 

mg ai/kg/day. 

0.5 mL undiluted 

Mimic®240 LV 

applied to shaved 

intact skin and sites 

were semi-occluded 

for 4 hours, after 

which the application 

sites were wiped with 

paper towels 

saturated with tap 

water and blotted 

dry. 

Skin sensitization 

protocol as detailed 

in the first row of the 

next page. 

NOEL = 1000 mg ai/kg 

No treatment-related systemic effects; minor 

dermal irritation observed in females were 

attributed to RH-75,992 2F formulation 

solvent and not the active ingredient. 

No mortalities or clinical signs of toxicity. At 

1 hour, well-defined erythema was observed 

in all rabbits (6/6). Observed erythema 

ranged from well-defined to none among 

rabbits at 24, 48, and 72 hours but was no 

longer evident by day 7. Edema was not 

observed during the study. 

Rohm and Haas classifies the test formulation 

as slightly irritating to skin. 

This study reveals the components of in the 

formulation. This information cannot be 

released 

No significant erythema observed in any of 

the guinea pigs induced with mimic 

formulation; 100% incidence of erythema in 

positive control group; no erythema in naive 

control group. 

Mimic did not produce delayed contact 

hypersensitivity in guinea pigs in this study. 

This study reveals the components of in the 

formulation. This information cannot be 

released 

Morrison et al. 

1993 

MRID 42991507 

Parno 1997 

MRID 44727704 

Anderson and 

Shuey 1994 

Anderson and Shuey 1994 Exposure details: 

Induction: treated guinea pigs received three 6-hour induction doses (1 dose/week for 3 consecutive weeks) of 

0.4 mL undiluted Mimic®240 LV to shaved skin; positive controls received three 6-hour induction doses (1 

dose/week for 3 consecutive weeks) of 0.4 mL DNCB (1600 ppm in 80% aqueous ethanol). Challenge dose: 2 

weeks after the last induction dose, treated pigs received 0.4 mL undiluted Mimic®240 LV and positive controls 

received 0.4 mL DNCB (800 ppm in acetone). Naive control group received 0.4 mL undiluted Mimic®240 LV 

to shaved skin at one site and 0.4 mL DNCB (800 ppm in acetone) at a separate site. 

Appendix 3-2 
Appendix 3: Dermal, inhalation, and ocular effects of tebufenozide in experimental mammals 

Animal Dose/Exposure Response Reference 

Guinea pigs, 

young adults, 

albino, 20 (test 

group), 10 

(control and 

positive control 

groups), 5 

(positive controlnaive control). 

INHALATION 

Rats, 5 males and 

5 females (NOS) 

Rats, Crl:CD, 6 

males and 6 

females 

Rats, CD, adults, 

6 males and 6 

females 

Test material 

administered as 5% 

w/w mixtures for 

intradermal injection 

and as 25% w/w 

mixture in petrolatum 

for topical induction 

and challenge 

applications 

4.3 mg/L aerosol dust 

for 4 hours (NOS) 

MIMIC wettable 

powder formulation. 

Mean aerosol 

concentration of 1.83 

mg/L, nose-only 

exposure for 4 hours, 

followed by 14-day 

observation period 

4-hour nose only 

exposure to measured 

concentration of 1.33 

mg/L Mimic®240 

LV (nominal 

concentration = 

178.2 mg/L 

The difference 

between the 

measured and 

nominal 

concentrations is 

attributed to the 

impaction of a 

portion of the aerosol 

on the interior 

surfaces of the 

exposure system. 

No skin sensitization in guinea pigs treated 

with test material; sulfathizole (used for 

positive control group) was shown to be an 

extreme sensitizer. 

LC50>4.3 mg/L (males) [0/5 deaths] 

LC50>4.5 mg/L (females) [0/5 deaths] 

These were highest technically achievable 

concentrations. 

No mortality; no treatment-related clinical 

signs of toxicity or body weight effects; no 

treatment-related gross lesions observed at 

necropsy. 

LC50 >1.83 mg/L 

This study reveals the components of in the 

formulation. This information cannot be 

released 

No mortalities or body weight effects. 

Clinical signs included wet fur immediately 

after exposure, respiratory noise (1/6 males 

and 1/6 females), red-stained fur around eyes 

(1/6 males and 1/6 females), red-stained 

muzzle (1/6 males), tan-stained muzzle (5/6 

males and 5/6 females). The tan stains 

(appearing to be test material) were attributed 

to poor positioning of the animals in the noseonly tubes. Tan stains, which appeared up to 

and including day 1 were not evident by day 

2. Necropsy revealed the following changes: 

red pinpoint foci in the lungs (5/6 males, 1/6 

females), slight to severe redness on all lobes 

of the lung (4/6 males and 6/6 females), which 

were considered to be consistent with 

irritation of the respiratory tract and judged to 

be treatment related. 

Combined male and female LC50 >1.33 mg/L 

Glaza 1993 

MRID 42991506 

Hazleton and 

Quinn 1995b 

MRID 43781708 

(hazard 

evaluation) 

Bemacki and 

Ferguson 1994a 

MRID 44200306 

Bemacki and 

Ferguson 1994b 

MRID 44727705 

This study reveals 

the components of 

in the 

formulation. This 

information 

cannot be 

released 

OCULAR 

Appendix 3-3 
Appendix 3: Dermal, inhalation, and ocular effects of tebufenozide in experimental mammals 

Animal Dose/Exposure Response	 Reference 

Rabbits (NOS)	 direct application to 

corneal surface of 

eye or into 

conjunctival sac 

(NOS) 

Rabbits, New	 0.1 mL undiluted 

Zealand white,	 Mimic®240 LV 

adults, 6 males	 applied to 

conjunctival sac of 

one eye; untreated 

eye served as control. 

After 24 hour 

observation period, 

eyes irrigated with 

saline for 

approximately 60 

seconds. 

Approximately 75% 

of test substance 

remained in contact 

with the eyes. 

no irritation in eyes washed 30 or 60 seconds 

after dose or in treated eyes that remained 

unwashed 

RH-5992 technical calssified as 

“inconsequentially irritating to the eye.” 

No mortality or clinical signs of toxicity. At 

1, 24, 48, and 72 hours, positive corneal and 

conjunctival effects were observed in 2/6 

rabbits; effects no longer evident by day 7. 

Rohm and Haas classifies Mimic®240 LV 

“MODERATELY IRRITATING” (i.e., a 

positive test that is reversible at $ 24 hours 

but #7 days. 

Hazleton and 

Quinn 1995b 

MRID 43781708 

(hazard 

evaluation and 

toxicity summary) 

Gingrich and 

Parno 1994s 

MRID 

444727706 

Appendix 3-4 
Appendix 4: Toxicity of tebufenozide to birds after oral administration. 

Animal Dose	 Response Reference 

ACUTE 

Bobwhite quail, 13­

days old, 10 per 

dose group 

Ducks, Mallard, 8­

days old, 10 per 

dose group 

LONGER-TERM 

Bobwhite quail, 29­

weeks old, five 

males and five 

females per dose 

group 

Ducks, Mallard, 25­

weeks old, 16 males 

and 16 females per 

dose group 

0, 312, 625, 2500, or 5000 ppm 

a.i. in diet for 5 consecutive days 

followed by a 3-day recovery 

period. 

Food consumption was about 

13% of body weight during the 

exposure period (Tables III and 

IV). Thus, the dietary 

concentrations correspond to 

doses of 0, 41, 81, 325, 650 

mg/kg bw/day. 

0, 312, 625, 1250, 2500 or 5000 

ppm in diet for 5 consecutive 

days followed by a 3-day 

recovery period 

0, 1470, or 2150 mg a.i./kg via 

gelatin capsules for 21 days. 

0, 100, 300, or 1000 ppm ai in 

the diet for 20 weeks 

LD50 >5000 ppm a.i.	 Fletcher. 

1990a 

MRID 

42436235 

LD50 >5000 ppm a.i. 

No mortality, no signs of toxicity, 

and no statistically significant 

difference in body weights, 

compared with controls. No 

abnormal tissue alterations were 

observed at necropsy. 

Acute LD50 >2150 mg a.i./kg bw 

No mortalities or treatment related 

adverse effects at any dose level; no 

adverse effects observed on body 

weight, food consumption, or 

reproductive endpoints. 

NOEL = 1000 ppm ai 

Fletcher 

1990b 

MRID 

42436237 

Fletcher 1987 

MRID 

42436234 

Beavers et al. 

1993a 

MRID 

42991503 

Appendix 4-1 
Appendix 4: Toxicity of tebufenozide to birds after oral administration. 

Animal Dose Response Reference 

Bobwhite quail, 18­ 0, 100, 300 or 1000 ppm ai in No treatment-related mortalities, Beavers et al. 

weeks old, 16 males the diet for 20 weeks. overt signs of toxicity, or effects on 1993b 

and 16 females per body weight or food consumption at MRID 

dose group Based on reported food any concentration. 42991501 

consumption rates of about 15% Reinert et al. 

of body weight (see special note Reproductive effects: 1993a 

below), the dietary at 300 ppm, possible slight reduction MRID 

concentrations correspond to in number of eggs laid (reflected in 42991502 

doses of 0, 15, 45, and 150 14-day old survivors as % maximum 

mg/kg/day. See special note eggs set and number of 14-day old 

below. survivors per hen per day A 

substantial drop in feed consumption 

was observed during weeks 8 and 9. 

At 1000 ppm, slight decreases in 

number of eggs laid and number of 

viable embryos. 

NOEL (for reproductive parameters) 

= 100 ppm 

SPECIAL SUPPLEMENTAL NOTES ON BEAVERS ET AL. 1993b [MRID 42991501, MRID 42991502] 

mg/kg bw doses: Average doses in units of mg/kg bw are not provided in the study. Table 2, p. 34. Average 

food consumption is estimated at 30 g per bird. There was a slight transient decrease food consumption at weeks 

10 and 11 in all dosed animals and weeks 13/14 in the two higher dose groups. The magnitude of the decrease 

was about 16% to 33%below that of controls. The average body weights of the animals was about 200 g over the 

course of the study. Thus, food consumption is taken as 15% of body weight (30 g/200 g). The methods 

specifically state that food and water were available ad libitum. “No attempt was made to quantify the amount of 

feed wasted by the birds, as the wasted feed is normally scattered and mixed with water and excreta.” (p. 16). 

Effects: See Supplemental Table 1 at the end of this appendix. 

Reinert et al. 1993a [MRID 42991502], which is a supplemental report indicates that two orders of magnitude 

difference between the NOEL for bobwhite quail (100 ppm) and mallard duck (1000 ppm) is not consistent and 

concludes that many of the endpoints in the bobwhite study are confounded by the usual variability in long-term 

studies and that the lack of dose-response in many parameters when judged against available data in avian studies 

does not support a conclusion of adverse effects at 300 ppm ai in the diet and that the NOEL probably approaches 

1000 ppm, as supported in the mallard study. 

Appendix 4-2 
Appendix 4: Toxicity of tebufenozide to birds after oral administration. 

Animal Dose Response Reference 

Bobwhite quail, 18­ 0, 150, 240, 385, or 615 ppm ai No treatment-related mortalities, Reinert 

weeks old, 15 males in diet for 20 weeks. Based on overt signs of toxicity or effects on 1995a 

and 15 females per reported food consumption rates body weight or feed consumption; no MRID 

dose group of about 8% of body weight (see apparent effects on reproductive 43781701 

special note below), the dietary endpoints. 

concentrations correspond to Reinert 

doses of 0, 12, 19.2, 30.8, 49.2 NOEL = 615 ppm (highest dose 1995b 

mg/kg/day. tested) MRID 

43781702 

LOAEC >615 ppm (Supplemental 

report) 

Reinert 

1995c 

MRID 

43781703 

Supplemental report 

of statistical 

analysis) 

SPECIAL SUPPLEMENTAL NOTES ON REINERT 1995a,b [MRID 43781701 AND MRID 43781702]: 

mg/kg bw doses: Average doses in units of mg/kg bw are not provided in the study. Table 3b, p. 24. Average 

food consumption is estimated at 16 g per bird. This is only about one-half of the food consumption in the 

Beavers et al. 1993b study - i.e., about 30 g/bird - summarized in the previous entry. The average body weights 

of the animals was about 200 g over the course of the study, similar to the body weights in the Beavers et al. 

1993b study. Thus, food consumption is taken as 8% of body weight (16 g/200 g). The food consumption 

estimates did explicitly consider measurements of food wastage - i.e., food scattered from the container and not 

consumed. Ad libitum feeding is assumed but not specified. 

Effects: See Supplemental Table 2 at the end of this appendix. 

Appendix 4-3 
Supplemental Tables for Appendix 4 

Appendix 4, Supplemental Table 1: 

Details of reproductive parameters in bobwhite quail (from Beavers et al. 1993b, 

Table 3, p. 36) 

Parameter PPM in Diet 

0 100 300 1000 

Eggs Laid 714 769 570 508 

Eggs Cracked 12 15 9 14 

Eggs Set 627 680 496 435 

Viable Embryos 595 616 451 367 

Live 3-W eek Embryos 592 609 451 367 

Hatchlings 569 564 429 348 

14-Day Old Survivors 544 516 387 322 

Eggs Laid/Hen 48 48 38 36 

Eggs Laid/Hen/Day 0.68 0.69 0.54 0.52 

14-Day Old Survivors/Hen 36 32 26 23 

Appendix 4, Supplemental Table 2: 

Details of reproductive parameters in bobwhite quail (from Reinert 1995a, pp. 24-29) 

Parameter PPM in Diet 

0 150 240 385 615 

Eggs Laid 640 632 514 671 516 

Eggs Cracked 2 2 1 0 0 

Eggs Set 576 587 476 623 483 

Viable Embryos - Day 5 Candeling 492 550 409 589 449 

Viable Embryos - Day 11 Candeling 488 545 398 578 446 

Live 18-Day Embryos 476 540 392 573 441 

Hatchlings 449 474 345 522 408 

14-Day Old Survivors 418 429 323 491 375 

Eggs Laid/Hen 42.7 42.1 36.7 44.7 34.5 

Appendix 4-4 
 

Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species 

Insects 

Honey bee, adult 

Mite, predatory 

Stethorus punctum 

Tufted apple bud 

moth larvae 

(Platynota 

idaeusalis) [target 

species] 

Tufted apple bud 

moth larvae 

(Platynota 

idaeusalis) [target 

species] 

Cydia pomonella 

codling moth 

[target species] 

Hyssopus pallidus, 

Hymenopteran 

parasitoid on 

codling moth eggs 

Exposure Effects Reference 

0, 59, 117, and 234 μg/bee; 

96 hour observation period. 

Mortality rates in exposed bees were 

about 3.4% to about 5% and were less 

Atkins 1990 

MRID 42436244 

than control mortality (5.88%) 

NOEC = 234 μg/bee 

Tests on larvae, pupae, and 

adults by 24-hour dry film 

exposures, with 

concentrations ranging from 

9-90 ppm. 

Tests on eggs placed on 

Not toxic to eggs, but survival of larva 

was reduced compared to untreated 

controls. Larval mortality likely due to 

contact with residues on leaf (not 

delayed effect of exposure during egg 

stage) 

Biddinger and 

Hull 1995 

treated leaves (92 ppm) 

Note: unclear if 

In contact assay, tebufenozide was not 

toxic to adults and did not effect pupal 

survival. Less toxic than 

concentrations are 

concentrations of solutions 

leaves were dipped in or 

diflubenzuron. 

concentration on leaf 

material. 

Dietary exposure. 7-Day LC = 1.63 ppm 50

14-Day LC50 = 1.12 ppm 

Somewhat lower LC50 values in 

sensitive laboratory strain. 

Biddinger et al. 

1998 

Dietary exposure. 0.03 or 

0.05 ppm 

No effect on larval or pupal 

development. 

Biddinger and 

Hull 1999 

Decreased fecundity in matings when 

both sexes were exposed. 

Dietary exposure. LC = 0.025 ppm 50 Brown 1996 

Dose-related decrease in number of 

viable eggs from exposed females, 

especially at concentrations > than the 

LD50. No effect if males only were 

exposed. Dose-dependent decreased in 

time to emergence of adult insect from 

pupal case. Effect more pronounced in 

females than males. 

Exposure via codling moth 

exposed to up to 40 ppm 

tebufenozide in diet [24x 

LC50] 

No adverse effects on egg or larval 

development of parasitoid at 40 ppm 

tebufenozide [24x LC50] 

Brown 1996 

Appendix 5 - 1 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species Exposure 

Ascogaster sp 

Hymenopteran 

endoparasitoid on 

codling moth eggs 

Codling moth exposed to 40 

ppm tebufenozide [24x 

LC50] 

Honey bee (Apis 

mellifera) 

24-hour and 72-hour 

exposure by direct contact, 

indirect contact (test 

substance on filter paper) 

and inhalation to 0.1% v/v 

(equivalent to 1.05 kg/ha in 

1000 L/ha) tebufenozide 

formulation Hoe 105540 SC 

(a 24% a.i. water soluble 

formulation) 

3-hour (250 :g a.i./bee) 

feeding and 24-hour feeding 

(dose range approximately 

2.4 to 800 :g a.i./bee) 

Note: for all contact and 

inhalation exposures, it is 

unclear is concentrations are 

given in terms of 

formulation or a.i. Authors 

state that 0.1% v/v is 

equivalent to twice the 

application rate 

Honey bee (Apis 

mellifera) 

tebufenozide formulation 

Hoe 105540 SC (a 24% a.i. 

water soluble formulation) 

applied at rate of 1.05 

kg/300 L applied at rate of 

0.2 kg/ha. [Appears to be 

given in terms of 

formulation, although this 

was not specifically stated] 

Effects Reference 

LC = 0.07971 ppm, 3x LC values 50 50 Brown 1996 

for moth 

Direct exposure Chan 1995 

24-hr: 2% mortality in treatment group MRID 43797702 

and 0% in controls 

72-hr: 14% mortality in treatment 

group and 12% in controls 

Indirect exposure 

24-hr: 0% mortality in treatment and 

control. 

72-hr: 10% mortality in treatment 

group, 8% in controls. 

Inhalation exposure 

24-hr: 0% mortality in treatment and 

2% mortality in control 

72-hr: 10% mortality in treatment and 

control. 

Oral exposure 

3-hr: 0% mortality in treatment and 

50 control. LD > 250 :g/bee 

24-hr: 0% mortality in highest dose 

group. 2% mortality in controls. No 

dose-dependent mortality was 

observed. LD > 800 50 :g/bee. 

No behavioral effects noted for any 

route of exposure or duration of 

exposure. 

Bee colonies tested in laboratory. Chan 1995 

MRID 43797702 

No increased in treatment-related 

mortality was observed. No effects of 

treatment on flight activities or 

behavior. No effects on brood (as 

measured by dead pupae). 

Appendix 5 - 2 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species 

Trichogramma 

pretiosum 

(parasitic wasp) 

Mexican rice borer 

(Eoreuma loftini) 

Exposure 

Exposure to T. pretiosum by 

dipping parasitized host 

eggs of Ephestia kuehniella 

in solutions of tebufenozide. 

Eggs dipped for 5 seconds 

on tebufenozide solution of 

25 g a.i./100 L. 

laboratory study. Exposure 

via leaves collected from 

sprayed field as follows: 

1996 season 

leaves collected 1 day after 

field application of low 

dose Confirm (0.14 kg 

a.i./ha) and high dose 

Confirm (0.2 kg a.i./ha). 

Insects were 1st instar larvae 

1997 season 

leaves collected 1 and 4 

days after application of 

Confirm (rate of 0.28 kg 

a.i./ha). Insects were 2nd 

and 3rd instar larvae 

Effects Reference 

Three different development stages of 

parasitized host eggs tested – egglarvae, pre-pupae, and pupae. 

Consoli et al. 

1998 

No significant increase in T. pretiosum 

mortality compared to untreated 

controls. 

Decreased development time was 

slightly significantly decreased for 

tebufenozide applied at the pupae stage 

(tebufenozide 9.68 days in control 

group and 9.35 day in tebufenozide 

group), but not when applied at the 

egg-larvae and pre-pupae stages. 

For parasite, parasitism capacity 

reduced when tebufenozide was 

applied at the egg-larvae and pre-pupae 

stages, but not when applied at the 

pupal stage, 

For the 1996 season 

Cumulative mortality as follows:

 low dose: 34.4%

 high dose: 39.4%

 untreated control: 0% 

Legaspi et al. 

1999 

For the 1997 season 

For organisms exposed to leaves 

collected 1 day after field application: 

after 9 days of exposure, mortality was 

approximately 80% (data presented 

graphically). 100% mortality after 12 

days of exposure 

For organisms exposed to leaves 

collected 4 days after field application: 

after 9 days of exposure, mortality was 

approximately 20% (data presented 

graphically). Mortality not assessed 

after 9 days. 

Appendix 5 - 3 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species Exposure 

braconid parasitoid 

Allorhogas 

pyralophagus 

exposure via leaves 

collected 1 and 4 days after 

field after applications of 

Confirm in 1996 and 1997. 

1996: low dose 0.14 kg 

a.i./ha and high dose0.2 kg 

a.i./ha 

1997: 0.28 kg a.i./ha 

Effects Reference 

Using 1997 field treatments [according Legaspi et al. 

to figure 5 legend, p 809], no mortality 1999 

was observed in for A. pyralophagus 

exposed to leaves (collected 1 day and 

4 days after field application) for 4 and 

24 hrs. 

Using 1997 field treatments [according 

to figure 6 legend, p 809], no 

difference was observed between 

control and high dose tebufenozide, but 

longevity was decreased for low dose 

tebufenozide. 

Note on Legaspi et al. 1999: From the methods section, it appears that 2 application rates of Confirm were 

tested in 1996 and one was tested in 1997. However, results for 1997 are presented for low and high dose groups. 

Beet army worm, tebufenozide (Confirm 2F) Susceptibility of field collected insects Mascarenhas et 

3rd instar in food at 22.7 % a.i. (9 strains) compared to ECOGEN al. 1998 

(Lepidoptera: (wt/wt) after exposure to laboratory strain using LC50 values 

noctuidae) diet for 120 hours 

ECOGEN LC50: 17.6 ppm 

Field organisms LC50 values range from 

39.7 to 176.3 ppm 

Appendix 5 - 4 
 

 

 

 

Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species 

predatory lacewing 

adults (Chysoperla 

carnea) 

predatory lacewing 

3rd instar lavae 

(Chysoperla 

carnea) 

Exposure 

tebufenozide (TEB), 18, 90 

and 180 ng/insect, applied 

topically [authors note that 

90 mg/insect is the 

maximum field 

recommended (MFRD) 

dose] 

Diflubenzuron (DBB) 

applied at 150 (2xMFRD) 

Topical application of 

tebufenozide (TEB, Mimic 

24% a.i.) applied at 0, 90 

and 180 ng a.i./insect and 

diflubenzuron (DFB, 25% 

a.i.) applied at doses 

ranging from 0.5-75 ng 

a.i./insect 

Authors note that for TEB, 

90 ng/insect is the 

maximum field 

recommended dose 

(MFRD) 

Effects Reference 

Tebufenozide did not fecundity and Medina et al. 

egg fertility. In contrast, diflubenzuron 2002 

reduced egg hatchability to 0% 

(compared to control 87%). 

To explore differences, compared 

cuticle penetration, distribution and 

excretion of compounds. 

Cuticle penetration: 

DFB 16% 

TEB 26% 

Excretion: 

DFB 24.8% of penetrated amount 

excreted in feces in 7 days 

TEB aprpox, 50% of penetrated 

amount excreted in feces in 7 days 

For DFB, only very small amounts of 

dose recovered in ovaries and 

deposited eggs. No TEB detected in 

ovaries or deposited eggs. 

TEB had no effect on pupation, adult Medina et al. 

emergence, fecundity or egg fertility. 2003 

DFB LD50: 2.26 ng a.i./insect. At the 

lowest dose tested (0.5 ng a.i./insect), 

no effect on fecundity or egg fertility 

compared to control 

Presented results of cuticle penetration 

and excretion studies as summarized 

above for Medina et al. 2002 

Appendix 5 - 5 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species 

Indian meal moth 

(Plodia 

interpunctella) 

spruce budworm 

(Choristoneura 

fumiferana) 

predaceous 

insidium flower 

bug (Orius 

inisidoisus), 

parasitic wasp 

(Cotesia plutella) 

Exposure Effects 

dietary exposure of 1st instar Larvae monitored for weight and 

lavae to tebufenozide (RH­ mortality until metamorphosis. 

5992) at concentrations of 

0, 0.1, 1, 5, 10, and 25 ppm Weight gain: No effect on wt gain at 

for up to 31 days concentrations up to 1.0 ppm. 

Exposure to 5 and 10 ppm results in 

decreased wt gain. Exposure to 25 

ppm results in larval weight loss. 

Mortality: At concentrations of 0.1 and 

1 ppm, no effect on mortality. 

Mortality increased compared to 

control at concentrations 5 and 10 ppm. 

100% mortality at 25 ppm. 

In cell culture (PID2 imaginal disc 

line), exposure to 0.005 :M 

tebufenozide significantly increased 

glucosamine uptake (increase by 30% 

of control level). 

not reported in Keller and RH-5992 is effective in inducing a 

Brown 1998a summary incomplete molt when fed to worms 

prior to appearance of the endogenous 

ecdysteroid peak, but when 

administered after the peak. However, 

incomplete molts are observed for 

subsequent molts, presumably due to 

the persistence of tebufenozide in cells. 

Confirm applied cotton O. insidoisus: exposure to 2- and 24­

plants at an application rate hour leaves for 24 or 48 hours did not 

of 0.125 lb a.i./acre. Insects results in an increase in mortality 

were tested on plants 2 and compared to control insects. 

24 hours after application. 

C. plutella: no significant increase in 

Insects exposed to fresh percent mortality compared to control 

foliar residues for 24 and 48 exposed to 2-hour old leaves. 

hours. 

Reference 

Oberlander et al. 

1998 

Palli et al. 1995, 

as summarized 

in Keller and 

Brown 1998a 

Pietrantionio and 

Benedict 1999 

Appendix 5 - 6 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species 

spruce budworm 

(Chorisroneura 

fumiferana) 

spruce budworm 

(Chorisroneura 

fumiferana), 6th 

instar stage 

two lacewing 

species – 

Chrysoperla 

carnea (Stephens) 

and Micromus 

tasaniae (Walker) 

Exposure 

1-100 ng/insect 

tebufenozide by ingestion 

Insects force-fed 0.1 :g a.i. 

tebufenozide (aqueous 

flowable RH-5992) 

Petri dishes sprayed with 

tebufenozide (Minic 20 

flowable liquid) at 

concentations of 0.08 to 0.8 

% a.i.) and film left to dry. 

To test for 

acetylcholinesterase activity 

(AchE), insects were 

exposed for 2 and 24 hours. 

For Glutathione-Stransferase (GST), insects 

were exposed for 10 hours. 

Effects 

In 6th instar insects, treatment induced 

lethal precocious molt. Lack of 

development of new cuticle due to lack 

of gene expression of 

dopadecarboxylase. Effect observed in 

100% of insects administered a dose of 

70 ng. 

th th For 4 and 5 instars, 100% effect for
 

lethal precocious molt was observed at
 

lower dose (20 ng/insect)
 

Topical exposure did not induce effects
 

at doses up to 10,000 ng/insect.
 

Effects observed at time points after
 

exposure:
 

6 hr – insects stop feeding.
 

12 hr – head capsule slips partially.
 

24 hr – pronounced head capsule
 

slippage and mid-dorsal split of old
 

cuticle.
 

Insect remains in this state and
 

ultimately dies of starvation and
 

dessication.
 

Microscopy if integument showed
 

hypertrophy of golgi complex and
 

alterations in the cutlicular
 

components, and organelles of
 

epidermal cells.
 

For both species, no inhibition of head
 

AchE or whole body GST.
 

Reference 

Retnakaran et al. 

1997a 

Retnakaran et al. 

1997b 

Rumph et al. 

1997a 

Appendix 5 - 7 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species 

lacewing 

Micromus tasaniae 

(Walker) (3rd 

instars) 

Codling moth 

(Cydia pomonella) 

– 3 strains 

Exposure 

Test materials applied to 

petri dishes. 

Tebufenozide 7.44 2 :g/cm (according to

authors, this is 10x the 

recommended field rate). 

For tebufenozide-exposed 

larvae, effects in offspring 

were also examined, but 

offspring were not exposed 

to any test substance. 

Diflubenzuron (DFB) 0.07 

:g/cm2 

Tebufenozide (Confirm) 

dose range 10-10,000 

ng/insect, applied topically 

Effects Reference 

Examined effects of tebufenozide and Rumph et al. 

DFB on life-table parameters (sex ratio, 1998 

longevity, sterility and fecundity) in 

adults derived from treated larvae. 

Tebufenozide: No mortality observed. 

No treatment effect for sex ratio, 

longevity or number of sterile pairs for 

either first or second generation. Total 

number of eggs in reduced by 30% in 

nd st 2 generation, but not 1 generation. 

Decreased in oviposition period for 1st 

generation (33.3 days) and 2nd 

generation (30.5 days), compared to 

control (39.8 days). Only 2nd 

generation change significant. No 

change in preoviposition period for 

either generation. 

DFB: Higher proportion of females in 

DFB (64.9% ) compared to controls 

(53.0%). Longevity reduced for 

females in DFB (34.1 days) compared 

to controls (46.1 days). No treatment 

effect for in number of sterile pairs, 

although a strong trend observed 

toward an increase in infertility. Daily 

number of eggs reduced. Increased 

preoviposition period. Significant 

decrease in oviposition period. 

In susceptible strains of diapausing 

larvae, tebufenozide breaks the 

diapausing period and induces molting 

and reduces the pre-emergent period. 

In resistant strains, treatment did not 

break the diapausing state. 

LC50 values of various strains – 

Sv: 27.4 ng/insect 

Rv: 362 ng/insect 

Rt: 1570 ng/insect 

Sauphanor et al. 

1999 

Appendix 5 - 8 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species Exposure 

Larvae of Galleria, topical application of RHSarcophaga and 5992 (dose range not 

Calliphora specified in Keller and 

Brown 1998a summary) 

Spodoptera Exposure by topical or oral 

exempta (Walker) routes. Topical application 

(beet armyworm), of 0.01 to 40,000 ng/insect. 

Spodoptera exigua Oral exposure by feeding 

(Hubner) (beet leaves or prey dipped in 

armyworm), tebufenozide solutions or 

Spodoptera tebufenozide in honey water 

littoralis (Egyptian 

armyworm), (technical grade 

Mamestra tebufenozide) 

brassicae (cabbage 

moth), Galleria 

mellonella (greater 

Wax moth) 

Effects 

Galleria: stimulation premature molt. 

ED50 = 1.75 :g/insect 

Sarcophaga and Calliphora: did not 

induce molt 

S. exempta 

LD50 (topical application):

 6.75 mg/insect for 6th instar 

LC50 (fed dipped leaves - values are 

concentration of test material leaves were 

dipped in)

 3rd instar 0.034 mg/L


 4th instar 0.095 mg/L


 5th instar 0.085 mg/L


 6th instar 0.084 mg/L
 

S. exigua 

LD50 (topical application):

 59.2 mg/insect for 5th instar 

LC50 (fed dipped leaves)

 1st instar 9.7 mg/L
 

2nd instar 10.5mg/L
 

3rd instar 8.5mg/L
 

4th instar 10.0 mg/L


 5th instar 2.5 mg/L
 

Dose-dependent decrease in fecundity 

following oral exposure to tebufenozide in 

honey water (1, 10, and 100 mg/L), 

although all deposited eggs were viable 

S. Littoralis 

LD50 (topical application):

 11.02 mg/insect for 6th instar 

M. brassicae 

LD50 (topical application):

 8.53 mg/insect for 6th instar 

G. mellonella 

LD50 (topical application):

 571 mg/insect for 6th instar 

For Lepidoptera larvae, tebufenozide 

induced lethal molt within 24 hours of 

exposure. Other effects included inhibition 

of weigh gain and feeding, extrusion of 

hindgut, and loss of hemolymph. 

Reference 

Slama 1995, as 

summarized in 

Keller and 

Brown 1998a 

Smaggje and 

Degheele 1994a 

Appendix 5 - 9 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species Exposure 

larvae of Exposure by topical or oral 

Leptinotarsa routes. Topical application 

decemlineata of 0.01 to 40,000 ng/insect. 

(Colorado potato Oral exposure by feeding 

beetle), leaves or prey dipped in 

Diabrotica tebufenozide solutions or 

virgifera virgifera tebufenozide in honey water 

(western corn 

rootworm), (technical grade 

Locusta migratoria tebufenozide) 

migratoria 

(migratory locust), 

and nymphs of 

Podisus sagitta 

(predatory stink 

bug) 

Spodoptera For LC50 determination, 

exempta (Afrian insects were fed leaves 

army worm), dipped in tebufenozide 

Spodoptera exigua (technical grade) solutions. 

(beet armyworm), 

Lepinotarda 

decemlineata 

(Colorado potato 

beetle) 

Effects 

No activity observed in any species at 

any dose or concentration tested. 

LC50 values (last instars) 

S. exempta: 0.034 mg/L 

S. exigua: 2.5 mg/L 

L. decemlineata: no mortality at 

concentrations up to 50 mg/L. At 100 

mg/L, sings of neurotoxicity (tremor 

and paralysis) were noted. 

For S. exempta and S. exigua, dosedependent decreased in larval weights. 

No affect of treatment on larval weight 

for L. decemlineata. 

Resistance of L. decemlineata and 

differences in sensitivities of S. 

exempta and S. exigua apparently not 

due to differences in pharmacokinetics. 

All three species showed similar 

pharmacokinetic parameters for 

absorption, excretion, distribution and 

metabolism of tebufenozide 

Reference 

Smagghe and 

Degheele 1994b 

Smagghe and 

Degheele 1994b 

Appendix 5 - 10 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species Exposure Effects Reference 

Podisus nymphs exposed orally to No effect in either species for any Smagghe and 

nigrispinus and P. RH-5992 via feeding on exposure. Degheele 1995, 

Maculiventris larvae of Spodoptera exigua as summarized 

(predatory soldier treated with 20 :g/larvae or No chemosterilizing effects observed in in Keller and 

bugs) in drinking water (100 adults Brown 1998a 

mg/L) or exposed topically 

to up to 100 :g/nymph. 

Adults treated orally via 

feeding on larvae of 

Spodoptera exigua treated 

with 20 :g/larvae or in 

drinking water (100 mg/L) 

Cotton leafworm tebufenozide (RH-5992 2F Repeated exposure over 5 generations Smagghe and 

(Spodoptera flowable) did not result in the development of Degheele 1997 

littoralis), tolerance to tebufenozide. 

laboratory strain For repeated exposures to 

and field strain induce tolerance, exposure For 3rd instar insects, laboratory strain 

was dietary via leaves (LC50 2.47 mg/L) was more susceptible 

dipped in 0.6 mg a.i./L than the field strain (LC50 11.31 mg/L). 

tebufenozide solution. 

For LC50 determination, 

tebufenozide applied 

uniformly to food [unclear 

if concentrations are final 

concentration in food or 

concentration of fluid 

applied to food.] 

Spodoptera exigua Dietary exposure via leaves S. exigua: Smagghe et al. 

 (beet armyworm) dipped in solution of 3 mg In control insects, major hemolymph 1995 

and Leptinotarsa a.i./L tebufenozide ecdysteroid peaks appeared ~3-4 days. 

decemlineata (technical grade) for S. After treatment with tebufenozide, 

(Colorado potato exigua and 50 mg a..i/L hemolymph ecdysteroid peaks was 

beetle) tebufenozide abolished. Treatment resulted in 

decreased weight gain. Typical 

precocious molting observed. 

L. decemlineata: 

In control insects, major hemolymph 

ecdysteroid peaks appeared ~8-9 days. 

Peak unaffected by tebufenozide 

treatment. No affect of treatment on 

larval weight gain. No precious 

molting observed. 

Appendix 5 - 11 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species 

Chrysodeixis 

chalcites (tomato 

looper), last instar 

Spodoptera exigua 

(beet armyworm) 

Spodoptera exigua 

(beet armyworm) 

and Ostrinia 

nubilalis 

(European corn 

borer) 

Exposure Effects Reference 

exposure to diet containing Symptoms of premature molting Smagghe et al. 

100 :g a.i./g diet observed within 12 hours of treatment. 1997 

Significant reduction in larval weight 

tebufenozide RH-5992 2F and feeding. 

Ultrastructural changes of the 

integument included increase in 

endoplasmic reticulum, hypertrophy of 

golgi complex, increase in nuclear 

volume, numerous oval and elongated 

mitochondria. Prothoracic gland cells 

were reduced in size, show loss of cell 

organelles, and autophagic vacuoles 

appeared. In foregut epithelium, 

prominent vacuoles formed and most 

cell oragnelles disappeared. 

Ultrastructural changes also observed 

in muscle cells, with absent 

mitochondria. 

Exposure via artificial diet Continuous exposure of all larval Smagghe et al. 

with concentrations of instars to LC25 doses for over 12 1998 

tebufenoxide varying generations revealed no loss in 

according to generation. susceptibility for up to 5 generations. 

G0-5: 0.5 mg/L 

G6-10: 1 mg a.i./L From G4 onwards, generationG11-12: 2 mg a.i./L dependent reduction in oviposition. 

For G , 65% of G 4 0 oviposition, for 

For disposition studies, all G12, 0% oviposition. 

insects were exposed to the 

same amount of test Higher tissue concentrations of 14Cmaterial (20,000 dpm) tebufenozide in hemolymph, carcass, 

consumed on leaf material. and gut in susceptible larvae compared 

to G0 larvae. All insects were exposed 

to the same amount of test material 

(20,000 dpm consumed on a leaf). 

Spodoptera exigua Spodoptera exigua (last instar): Smagghe et al. 

exposed to tebufenozide in Chitin formation in cuticle was 1999a 

diet. 50 :L of solution increased in tebufenozide treated 

containing 1 mg/L insects compared to controls. Treated 

tebufenozide (50 ng) added insects died by day 3 after exposure 

to artificial diet in culture 

dish for exposure to 1 Ostrinia nubilalis (day-1 male pupae): 

insect. Tebudenzide exposure prevented the 

completion of adult development and 

eclosion. Time to death decreased with 

Ostrinia nubilalis exposed increasing dose. Tebufenozide 

to tebufenozide (0, 10, 25, exposure induced premature chitin 

50, 200, 300, and 400 synthesis in male claspers. 

ng/insect) by injection. 

Appendix 5 - 12 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species Exposure Effects Reference 

Spodoptera exigua 

(beet armyworm), 

last instars 

Tebufenozide applied 

topically to individual 

insects. Mortality counts 

made 7 days after exposure. 

LD50 = 7.06 mmole/insect Smagghe et al. 

1999b 

Cydia pomonella 

(codling moth) 

Exposure of adults to 

surfaces treated with 

tebufenozide solution (360 

ppm*) throughout their 

lives, including mating and 

ovipositing). 

Continuous exposure to tebufenoz

treated surfaces resulted in signific

reduction in number of eggs laid 

(control, 74.5 eggs; treatment 39.6 

eggs) and number of eggs hatched 

(control, 58.4%; treatment, 6.6%). 

ideant Sun and Barrett 

1999 

Recently emerged moths 

exposed to treated surfaces 

(360 ppm*) for 24 hours, 

then mated with unexposed 

partner (oviposit on nontreated surface) 

24-hour exposure of females mate

unexposed males resulted in redu

in fecundity (control, 97.7 eggs; 

treatment 26.8 eggs) and fertility 

(control, 86.3%; treatment, 78.7%). 

No effect if exposed male was mat

with unexposed female 

d to 

ction 

ed 

tebufenozide was RH-5992, 

2F (flowable) 

* authors state that this is 

the recommended field rate 

Orius laevegatus 

(predatory bug) 

exposure to plates sprayed 

with tebufenozide at the 

manufacturers 

No effect on development of nymp

on oviposition. 

hs or van de Veire et 

al. 1996, as 

summarized in 

recommended rate Keller and 

Brown 1998a 

Gypsy moth [target 

species] 

Tebufenozide applied to 

branches of oak trees at 

rate of “237 mL per 189 L 

final solution (label 

recommends 8 oz per 50 gal 

solution per acre), with 0.25 

5 (v/v) Bond sticker”. 

Laboratory-reared gypsy moth larv

(1 , 2 , 3rs, and 4 instars studied st nd th 

separately) were placed in bags and 

tied onto tips of treated branches 1 

hour after spraying. Larvae were 

exposed for 7-21 days. Same prot

was followed for larvae applied to 

branches 1, 2, 7, 14, 21, 28 and 35 

ae 

ocol 

days 

Webb et al. 1998 

Difubenzuron (DFB) 

“Dimilin 25W at 237 mL 

after spraying. 

per 378 L final solution, 

without added sticker”. 

For the exposure 1-hour postapplication, 100% mortality observed 

for all insects after 21 days of 

exposure. Similarly, 100% mortality 

observed for all “aged” residues. 

DFB also showed very high efficacy, 

except for 69% mortality on 14-day 

residue. However, all other DFB aged 

residues resulted in 100% mortality. 

Appendix 5 - 13 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species 

Epiphyas 

postvittana 

(lightbrown apple 

moth) 

Soil Invertebrates 

Earthworm 

(Dendrobaena 

octaedra), 40 per 

dose 

Collembola 

(Folsomia 

cundida, 

F. nivalis, 

Onychiurus 

parvicornis, and 

Hypogastrura 

pannosa) 

Round worm 

larvae (Ascaris 

suum) 

Exposure 

larvae exposed to 

tebufenozide (Mimic 70W) 

in food at concentrations of 

0, 0.5, 1. 1.5, 2, 2.5, 3, 10, 

30, 100, and 200 ppm. 

Deciduous leaves at 0 

(untreated), 10X and 100 X 

EEC for 12 weeks. 55.4 

ppm and 554 ppm based on 

reported EEC of 5.5461 

mg/kg (equivalent to the 

application rate of 70 g/ha). 

1996Coniferous substrate at 

72.1 μg/g (ppm) organic 

matter for 8 to 10 weeks

RH-5992 at concentrations 

in media of 5 and 50 ng/mL 

Effects 

Dose-mortality response determined at 

each larval stage. 

1st instar: no survival to pupation at 

concentrations >1.5 ppm 

3rd instar: no survival to pupation at 

concentrations >2.5 ppm

5th instar: dose-related decrease in 

survival to pupation. In 200 ppm 

exposure group, 14.8% survival. Time 

to mortality was less than in1st and 3rd 

instars. 

Mortality increased with increasing 

exposure time. Time to mortality for 

rd th 3 and 5 instars decreased when 

insects were exposed at 40BC compared 

to 20BC. 3rd instars more susceptible at 

higher temperature than 5th instars. 

No effects on growth or reproduction 

(numbers or proportion hatching) 

 No effect on survival or reproduction. 

Treatment had a biphasic effect on 

larval growth after 24-hour, premolt 

exposure – low concentrations (5 

ng/mL) increase growth. Higher 

concentrations decreased growth (> 50 

ng/mL) 

Reference 

Whiting et al. 

1999 

Addison 1996 

Addison 1996 

Fleming 1998, as 

summarized in 

Keller and 

Brown 1998a 

Appendix 5 - 14 
Appendix 5. Toxicity of tebufenozide to terrestrial invertebrates (standard toxicity 

studies and microcosm studies). 

Species Exposure Effects Reference 

earthworm 14-day exposure to RH- 14-day LC50 > 1000 mg ai/kg Garvey 1992, as 

(Eisenia foetida) 5992 at soil concentrations cited in Keller 

of 0, 61, 140, 270, 580, and 14-day NOAEC >1000 mg ai/kg 1994 (MRID 

1000 mg a.i/kg (Although 43367001) 

not specified, assume this is 

kg soil).No effect on 

survival at any 

concentration tested. 

Appendix 5 - 15 
Appendix 6. Terrestrial field/mesocosm studies on tebufenozide 

Application Species 

Examined 

Mimic 2F, 0.03 lb	 Gypsy moth; 

a.i./acre in mixed oak	 Other 

forest, May 1994	 macrolepidoptera 

richness and 

abundance 

Effects 

Examined effect of treatment on richness 

and abundance of arthropod family and 

macrolepidoptera. Sampling conducted 

May-Aug 1994 and May-Aug 1995. 

Marginal decrease in gypsy moth 

populations (not statistically significant 

compared to control plots). 

Nontarget arthropod richness and 

abundance: except for macrolepidoptera 

families, no effect of treatment for either 

sampling year. 

Significant decrease in the 

microlepidopteran Gelechiidae (p=0.02) in 

treatment year but not following year 

Marginal (p=0.07) decrease in sap-feeding 

Tingidae in treatment year but not 

following year. 

Macrolepidoptera richness: no effect of 

treatment in either sampling year 

(compared to control). 

Macrolepidoptera abundance: decreased 

during the last 8-13 weeks of 1994, but not 

different from control in the first 1-7 

weeks of 1994 or for any sampling period 

in 1995. 

Reference 

Butler et al. 

1997 

Appendix 6 - 1 
Appendix 6. Terrestrial field/mesocosm studies on tebufenozide 

Application Species Effects 

Examined 

Mimic 2F, 0.06 lb 

a.i./acre in mixed oak 

forest, May 1994 

Examined effect of treatment on richness 

and abundance of arthropod family and 

macrolepidoptera. Sampling conducted 

May-Aug 1994 and May-Aug 1995. 

Marginal decrease in gypsy moth 

populations (not statistically significant 

compared to control plots). 

Nontarget arthropod richness and 

abundance: except for macrolepidoptera 

families, no effect of treatment for either 

sampling year. 

Significant decrease in the 

microlepidopteran Gelechiidae (p=0.02) in 

treatment year but not following year 

Marginal (p=0.07) decrease in sap-feeding 

Tingidae in treatment year but not 

following year. 

Macrolepidoptera richness: decreased 

during the first 1-7 weeks after treatment 

in 1994 and during the first 1-8 weeks of 

the 1995 sampling period (compared to 

control). 

Macrolepidoptera abundance: decreased 

for the 1994 season and for the first 1-8 

weeks of 1995 season. 

Reference 

Butler et al. 

1997 

Additional Notes on Butler et al. 1997: Some macrolepidoptera (e.g., Melanolophia canadaria) were 

relatively insensitive while others (Lophocampa caryae [Hickory Tussock moth]) were highly sensitive. 

Mimic 2F, 70 and Spruce budworm Larval survival not significantly decreased Cadogan et al. 

140 g/ha [0.06 and at one application of 70 g/ha. Significant 1997 

0.12 lb a.i./acre] reductions at two applications at 70 g/ha or 

one application at 140 g/ha. 

Phenological development and larval and 

pupil weights significantly decreased in 

treated budworms compared to untreated 

controls. 

Appendix 6 - 2 
Appendix 6. Terrestrial field/mesocosm studies on tebufenozide 

Application 

Mimic, tested on 

apple plots in 

Australia 

1994/1995 season: 8 

applications of 15 g 

a.i./100 L applied by 

air-blast sprayer at 

1720 L/ha [258 g 

a.i./ha or 0.23 

lb/acre] 

1995/1996 season: 9 

applications of 10.5 g 

a.i./L applied by airblast sprayer at 1720 

L/ha [180.6 g a.i./ha 

or 0.16 lb/acre] 

Mimic 240 LV. 0.07 

a.i. kg/ha. Two aerial 

applications spaced 4 

days apart in June 

1994. Ontario 

Canada 

Confirm 70W RH5992 wettable 

powder applied to 

sugar cane plots in 

Texas. For the 1996 

season, two 

application rates: 

0.14 kg a.i./ha and 

0.2 kg a.i./ha [0.12 

lb/acre and 0.18 

lb/acre]. For the 

1997 season, 0.28 kg 

a.i./ha [0.25 lb/acre] 

Species 

Examined 

lepidopteran pests 

and nontarget 

arthropods and 

Tennessee warbler 

nests, 6 in control plot 

and 5 in Mimic treated 

plot. Monitored 

number of eggs laid, 

percent hatch and 

growth of the 

hatchlings 

Mexican rice borer 

(Eoreuma loftini) 

Effects 

Note: no untreated control plot. All 

comparisons were made to plots treated 

with other insecticides (azinphos-methyl 

and fenoxycarb). 

All plots treated with Mimic showed 

effective control over lepidopteran pests 

(codling moth, lightbrown apple moth, and 

early seasons caterpillars) 

Populations of natural enemies ( increased 

spiders, lacewings, and the specialist 

preditor mite Stethorus spp. adults and 

larvae. 

Decreases in both the average number of 

eggs per nest (6.3 in the control area and 

5.8 in the treated area) as well as the 

percent hatch (97.4% in the control area 

and 89.7% in the treated area). Based on 

the number of eggs, the differences in 

hatching were 37/38 in control plot and 

26/29 in treated plot. Using the Fisher 

Exact test, the p-value is 0.21 – i.e., not 

statistically significant. Decrease in 

brooding time and increase in foraging 

times in Mimic treated plot were probably 

associated with decrease in prey. 

For all application rates for the 1996 and 

1997 growing seasons - 

Treatment did not decrease the damage to 

cane caused by E. Loftini in either growing 

season. No increase in cane juice yield or 

quality in either growing season. 

Reference 

Gurr et al. 1999 

Holmes 1998 

Legaspi et al. 

1999 

Appendix 6 - 3 
Appendix 6. Terrestrial field/mesocosm studies on tebufenozide 

Application 

Confirm 2F applied 

to plots of peanuts at 

rates of 0.125 and 

0.24 lb a.i./acre. 

Treatment applied on 

Aug 7, 1998. Plots 

monitored on days 2, 

5, 7, 10, 14 and 20 

after application. 

Confirm 2F applied 

to plots of peanuts at 

0.25 lb a.i./acre. 

Treatment applied on 

Aug 7, 1998. 

Greenhouse study. 

Tebufenozide (RH5992-2F) applied at 

35, 70, 140 and 280 

g a.i./ha to potted 

white spruce trees. 

[0.03, 0.06, 0.12, and 

0.24 lb/acre] 

Species
 

Examined
 

defoliating caterpillars 

and beneficial 

arthropods (not 

specified) 

potato leafhopper, 

defoliating caterpillars 

(corn earworm, beet 

armyworm, rednecked 

peanutworm, 

and beneficial 

arthropods (not 

specified) 

spruce budworm 

(Chorisroneura 

fumiferana) exposed 

to trees for 10 days 

Effects 

For defoliating caterpillars, the only 

decreased in numbers was observed for the 

high dose Confirm on day 3 (9% of 

control) after treatment. 

Only decrease in beneficial arthropods 

observed for low dose Confirm (315 of 

control) on Day 3 after treatment but not 

on subsequent days (5 to 15 DAT). 

For beet army worm, numbers were 

decreased for low (6% of control) and high 

(5% of control) application rates on day 3 

after treatment. 

Potato leafhopper numbers increased on 

day 14 after treatment (220% of control), 

but not days 7 and 20 

Number of total defoliating caterpillars 

decreased on day 3 (52% of control) and 

day 7 (14% of control) after treatment. 

Number of beet aryworms decreased on 

day 7 (0% of control) after treatment. 

Number of beneficial arthropods not 

decreased at any time point. 

Evaluated effectiveness of treatment by 

mortality and feeding rate of 4th instar 

insects (by counting number of droppings, 

i.e., frass pellets). 

After 10 days exposure, mortality was not 

increased compared to controls for any 

treatment group. However, feeding 

inhibition was apparent and similar for all 

treatment groups. 

Reference 

Mulder and 

Prescott 1999a 

Mulder and 

Prescott 1999b 

Retnakaran et al. 

1997a 

Appendix 6 - 4 
Appendix 6. Terrestrial field/mesocosm studies on tebufenozide 

Application 

Tebufenozide 

applied (RH-5992­

2F) applied at 35, 70, 

140 and 280 g a.i./ha 

[0.03, 0.06, 0.12, and 

0.24 lb/acre] to 0.1 

ha plots of white 

spruce trees in Zee 

Casault, Gaspe, 

Quebec. 

Tebufenozide applied 

to apple plots in New 

South Wales, 

Australia.. 

Treatments applied 

between Nov to Feb 

over the 1992-1993 

and 1993-1994 

growing seasons. In 

each season, 8 

applications of 

Mimic at rate of 15 g 

a.i./100 L 

(volume/acre or ha 

not indicated) using 

conventional air-blast 

sprayer. 

No untreated control 

plots. 

Species 

Examined 

spruce budworm 

(Chorisroneura 

fumiferana) 

Several species ­

codling moth, early 

fruit caterpillars (not 

specified), lightbrown 

apple moth, the 

predatory mites 

Typhlodromus pyri 

and Typhlodromus 

occidentalis, spiders 

(Stetorus spp) and 

apple dimpling bug 

nymphs 

(Campylomma 

liebknechti) 

Effects 

For plots treated with >70 g a.i./ha, 

population reduction was 100% 

For plots treated with 35 g a.i./ha, 

population reduction was 95%. 

For all tebufenozide treated plots, 

defoliation was 1-2%, compared to 13­

16% in control plots. 

Comparisons of the effects of tebufenozide 

were made to 2 other treatments: azinphosmethyl and fenozycarb. 

No differences between treatments for fruit 

damage due to codling moth or early fruit 

caterpillars in either season. 

In the 1992-1993 seasons only, 

tebufenozide more effective than 

fenoxycarb on controlling damage due to 

lightbrown apple moth. 

Tebufenozide was ineffective in 

suppressing populations of the phyoseiids 

Typhlodromus pyri and Typhlodromus 

occidentalis. Compared to azinphosmethyl treatment, numbers of spiders 

(Stetorus spp) and apple dimpling bug 

nymphs (Campylomma liebknechti), 

numbers were higher in the tebufenzidetreated plots. 

Reference 

Retnakaran et al. 

1997a 

Valentine et al. 

1996 

Appendix 6 - 5 
Appendix 6. Terrestrial field/mesocosm studies on tebufenozide 

Application 

balsam firm tree plots 

in Newfoundland 

One application 

tebufenozide 

(Mimic) applied at a 

rate of 65.1 g a.i. in 

1.86 L/ha [authors 

also refer to this dose 

as 70g a.i./ha 

equivalent to 0.06 

lb/acre] 

Two applications 

tebufenozide 

(Mimic) at rate of 

33.4-35.4 g a.i in 

1.91-2.02 L/ha to 

[authors also refer to 

this dose as 35 g 

a.i./ha equivalent to 

0.03 lb/acre] 

Species 

Examined 

eastern hemlock 

looper 

Effects 

One higher dose application: 

• 9/10 plots showed reduction of loopers. 

• 9-11 days post-treatment, 3-93% 

reduction.

 • 3 weeks post-treatment 8-100% 

reduction. 

• Pupal populations reduced 8-99% 

• Defoliation of year-old foliage 10-51% 

(control plots 35-65%) and current-year 

foliage 0-16% (control plots 15-39%). 

Two lower dose applications: 

• 9-11 days post-treatment, in general, 

>50 % reduction.

 • 3 weeks post-treatment, in general 

>60% reduction. 

• Pupal populations reduced 76-100% 

• Defoliation of year-old foliage reduced 

1-33% (control plots 35-65%) and currentyear foliage reduced 0-8% (control plots 

15-39%). 

For both treatments, plots with poor 

efficacy were associated with low foliar 

deposition, with deposits <1.5 :g/g foliage 

(deposition measured for each plot) 

associated with ineffective control. 

Reference 

Appendix 6 - 6 
Appendix 7: Toxicity of tebufenozide to fish. 

Species Exposure Response Reference 

ACUTE 

Bluegill sunfish 

(Lepomis 

macrochirus), mean 

wt = 0.32 g, mean 

length = 24 mm, 

juveniles, 10 fish/dose 

group 

Rainbow trout 

(Oncorhynchus 

mykiss), juveniles, 

mean wet wgt = 0.39 

g, mean standard 

length = 28mm, 2 

replicates of 10 per 

dose group 

nominal concentrations of 

0, 0.5, 1.0, 2.0, 5.0, 10.0, 

25.0, or 100 mg ai/L; 

mean measured 

concentrations of 0, 0.39, 

0.90, 2.2, 4.0, 5.7, 9.4, or 

18 mg ai/L (ranging from 

18-100% of nominal 

concentrations) for 96 

hours under static 

conditions 

nominal concentrations of 

0, 0.5, 1, 2, 5, 10, 25 or 

100 mg ai/L; mean 

measured concentrations 

of 0, 0.42, 0.84, 1.9, 4.7, 

7.2, 10, or 17 mg ai/L for 

96 hours under static 

conditions 

No toxicity observed at 

concentrations #0.39 mg ai/L 

96 hr LC50 = 3.0 mg ai/L 

(95% CI = 2.2 and 4.0 mg ai/L) 

NOEC = 0.39 mg ai/L 

96 hr LC50 = 5.7 mg ai/L 

(95% CI = 4.7 and 6.5 mg ai/L) 

NOEC = 1.9 mg ai/L 

no signs of toxicity at 

concentrations #1.9 mg ai/L; 

mortality data from the highest 

dose group was not used to 

calculate the LC50 values. 

Graves and Smith 

1992b 

MRID 42436239 

Graves and Smith 

1992c 

MRID 42436240 

Appendix 7-1 
Appendix 7: Toxicity of tebufenozide to fish. 

Species Exposure Response Reference 

LONGER-TERM 

Fathead minnow 

(Pimephales 

promelas), newly 

fertilized eggs (<24 

hours after 

fertilization) used to 

initiate full life cycle 

study, 4 replicates of 

25 animals per dose 

group. 

Fathead minnows 

(Pimephales 

promelas), 30 days 

post hatch 

mean measured 

concentrations of 0, 

0.048, 0.090, 0.18, 0.35, 

or 0.72 mg ai/L (ranging 

from 92-100%) of 

nominal concentrations 

(0.048, 0.095, 0.19, 0.38, 

or 0.75 mg ai/L) under 

flow-through conditions. 

Both untreated and 

vehicle (acetone) control 

groups were assayed. 

nominal concentrations: 

0, 0.063, 0.13, 0.25, 0.50, 

or 1.0 mg ai/L; mean 

measured concentrations: 

0, 0.084, 0.14, 0.22, 0.36, 

or 0.71 mg ai/L by 

continuous exposure for 

35 days. 

Both untreated and 

solvent controls were 

used. 

No effects on egg hatchability, 

parental generation growth, 

reproductive activity, or F1 

generation survival at any test 

concentration. 

Parental generation survival 

was significantly decreased at 

the two highest dose levels (0.35 

and 0.72 mg ai/L): mean 

survival = 66% at 0.35 mg ai/L 

(mortality = 22/25, 20/25, 7/25, 

and 17/25 in replicate groups 

A,B,C, and D, respectively) and 

33% at 0.72 mg ai/L (mortality 

= 9/25, 17/25, 3/25, and 4/25 in 

replicate groups A,B,C, and D, 

respectively). 

The study and the supplement 

report no adverse effects on 

organism survival at hatch, 

larval survival and larval length 

and weight at any concentration 

levels. 

The U.S. EPA has classified the 

0.71 mg/L concentration as an 

effect level based on decreased 

survival (88%) relative to 

survival in the solvent control 

(98%). 

Rhodes and Leak 1996 

MRID 44221901 

Reinert et al. 1999 

MRID 44831501 

Bettancourt 1992 

MRID 42436242 

Surprenant 1994 

MRID 43145701 

(Supplement) 

Appendix 7-2 
Appendix 8: Toxicity of tebufenozide to aquatic invertebrates and algae. 

Plant or 

Animal 

Exposure Response Reference 

Aquatic Invertebrates 

ACUTE 

Cladoceran nominal test 48-hour LC50 = 3.8 mg ai/L 

(Daphnia concentrations: 0, (95% CI = 2.9 and 5.1 mg ai/L) 1992a 

magna), 0.25, 0.50, 1.0, 2.5, NOEC = 0.82 mg ai/L 

neonates (<24­ 5.0, 10, or 100 mg 

hours old), 2 ai/L; no signs of toxicity at concentrations #0.82 

replicates of mean measured mg ai/L; values >1.8 ai/L were considered to 

10 each per concentrations: 0, be above the functional water solubility of the 

dose group 0.22, 0.50, 0.82, test substance. 

1.8, 4.7, 6.4, or 35 

mg ai/L for 48 hours 

under static 

conditions 

Northern 1.0, 10, or 100 μg No adverse effects on survival and behavior. Dionne 1998 

lobsters ai/L Confirm 2F for 

(Homarus 96 hours under 

americanus), static conditions 

juveniles, 50­

80 mm long 

Midge larvae 0, 0.05, 0.1, 0.2, 96-hour aqueous LC50 = 0.30 mg ai/L (95%CI 

(Chironomus 0.4, or 0.8 mg ai/L = 0.23-0.40 mg ai/L 

riparius), 20 for 96 hours under 

larvae (2 static conditions 96-hour NOEC = 0.12 mg ai/L 

replicates of 

10 animal Both untreated and both values based on mean measured 

each) solvent controls concentrations. 

(acetone 0.10 

mL/L). 

Graves and Smith 

MRID 42436241 

MRID 44945701 

van der Kolk 1997 

MRID 44198301 

Appendix 8-1 
Appendix 8: Toxicity of tebufenozide to aquatic invertebrates and algae. 

Plant or Exposure Response Reference 

Animal 

Aquatic Invertebrates (continued) 

LONGER-TERM 

Daphnia Continuous Mortality: at 21 days, average mean survival McNamara 1991 

magna, 10 per exposure to 16, 29, at 240 μg ai/L group= 50%, significantly less MRID 42436243 

replicate 59, 120, or 240 μg (p<0.05), than controls (96%); survival in 

vessel ai/L for 21 days lower dose groups ranged from 93-100%. 

under flow-through 

conditions. 

Additional Notes on McNamara 1991:
 

Reproduction: at 120 μg ai/L, statistically significant decrease (p#0.05) in average rate of offspring/female
 

(n=143), compared with controls (n=188); at lower concentrations, rate of offspring/females ranged from 226 to
 

239, which is statistically comparable to control.
 

Growth: at 120 μg ai/L, statistically significant decrease (p#0.05) in mean total body length (5.0 mm), compared
 

with controls (5.4 mm); at lower concentrations, mean total body length ranged from 5.3 to 5.5, which is
 

statistically comparable to controls; 


at 59 and 120 μg ai/L, statistically significant decrease (p#0.05) in mean dry weight (1.3 and 1.6 mg,
 

respectively), compared with controls (1.9 mg); at lower concentrations, mean dry weight ranged from 1.9 to 2.0,
 

which is statistically comparable to controls;
 

LOEC = 59 μg ai/L; NOEL = 29 μg ai/L
 

21-day EC50 = 250 μg ai/L (lower 95% confidence interval of 120 μg ai/L)
 

Midge larvae 0, 0.0035, 0.0053, No effect on development rate of midge at any van der Kolk 1997
 

(Chironomus 0.0079, 0.012, concentration; at $0.040 no midge emerged, MRID 44198301
 

riparius), 2- to 0.018, 0.027, 0.040, which precluded the calculation of a
 

3-days old, 4 0.060,0.090, or development rate; at 0.0053, there was a
 

replicates per 0.135 mg ai/L for statistically significant (p#0.05) decrease in
 

dose group 28 days emergence rate; NOEC = 0.0035.
 

Both untreated and 

solvent controls 

(acetone 0.10 

mL/L). 

Appendix 8-2 
Appendix 8: Toxicity of tebufenozide to aquatic invertebrates and algae. 

Plant or Exposure Response Reference 

Animal 

Aquatic Algae 

Freshwater 

green alga 

(Scenedesmus 

subspicatus) 

Freshwater 

green alga 

(Selenastrum 

capricornutum 

) replicate 50 

mL cultures (3 

per treatment 

levels) 

0.046, 0.077, 0.15, 

0.25, or 0.66 mg 

ai/L (63-89% of 

nominal 

concentration) for 

96 hours. 

Cell density: at 0.077, 0.15, 0.25, and 0.66 

mg ai/L, respective cell densities averaged 81, 

58, 52, and 37 x 10 cells/mL and were 4 

statistically reduced compared with pooled 

control cultures (114 x 10 cells/mL); at the 4 

lowest treatment level, cell density was 

Hoberg 1992a 

MRID 42629501 

Both untreated and 

statistically similar to that of controls). 

solvent controls 

(acetone 0.10 

mL/L). 

Growth rate: at 0.15, 0.25, and 0.66 mg ai/L, 

the 72-96 hr growth rates were 0.259, 0.310, 

and 0.004 days-1, respectively and were 

statistically reduced compared with the 

growth rate of pooled controls (0.594 days ) -1 

NOEC for 72-96 hr growth rate = 0.077 mg 

ai/L. 

The 96 hr EC50 = 0.21 mg ai/L (95% 

confidence limit = 0.071-0.63 mg ai/L) 

Nominal Empirically estimated EC50 >0.64 mg ai/L Hoberg. 1992b 

concentration of MRID 42436245 

0.80 mg ai/L for NOEC (based on reduced cell density) = 0.64 

120 hours ai/L Reinert. 1993b 

MRID 42822201 

Treated algal culture reduced in density by 

9.1% compared with controls 

Appendix 8-3 
Appendix K 

DDVP (Dichlorvos) 

Risk Assessment 

Figure K-1. A sprayer unit mounted on a Model A Ford truck was used for gypsy 

moth control. 

SERA TR 04-43-05-05c 

Control/Eradication Agents for the
 

Gypsy Moth ­

Human Health and Ecological Risk Assessment for
 

DDVP (Dichlorvos) 

FINAL REPORT
 

Prepared for: 

USDA, Forest Service
 

Forest Health Protection 

GSA Contract No. GS-10F-0082F
 

USDA Forest Service BPA: WO-01-3187-0150
 

Requisition No.: 43-3187-1-0269
 

Task No. 5
 

Submitted to:
 

Dave Thomas, COTR
 

Forest Health Protection Staff
 

USDA Forest Service
 

Rosslyn Plaza Building C, Room 7129C
 

1601 North Kent Street
 

Arlington, VA 22209
 

Prepared by:
 

Patrick R. Durkin and Mark H. Follansbee a
 

Syracuse Environmental Research Associates, Inc. 

5100 Highbridge St., 42C
 

Fayetteville, New York 13066-0950
 

Telephone: (315) 637-9560
 

Fax: (315) 637-0445
 

E-Mail: SERA_INC@msn.com
 

Home Page: www.sera-inc.com


a Syracuse Research Corporation, 301 Plainfield Road, Suite 350, Syracuse, New York 13212 

August 20, 2004 (Risk Assessment)
 

July 20, 2007 (Editorial Corrections)
 
TABLE OF CONTENTS
 

TABLE OF CONTENTS ....................................................... ii
 

LIST OF TABLES ............................................................. v
 

LIST OF FIGURES ............................................................ v
 

LIST OF APPENDICES ........................................................ v
 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . vi
 

COMMON UNIT CONVERSIONS AND ABBREVIATIONS . . . . . . . . . . . . . . . . . . . . . . . . viii
 

CONVERSION OF SCIENTIFIC NOTATION ..................................... ix
 

EXECUTIVE SUMMARY ..................................................... x
 

1. INTRODUCTION ......................................................... 1-1
 

2. PROGRAM DESCRIPTION ................................................ 2-1
 

2.1. OVERVIEW ..................................................... 2-1
 

2.2. CHEMICAL DESCRIPTION AND COMMERCIAL FORMULATIONS . . . . . 2-1
 

2.3. APPLICATION METHODS ......................................... 2-2
 

3. HUMAN HEALTH RISK ASSESSMENT ..................................... 3-1
 

3.1. HAZARD IDENTIFICATION ....................................... 3-1
 

3.1.1. Overview. ................................................ 3-1
 

3.1.2. Mechanism of Action........................................ 3-1
 

3.1.3. Kinetics and Metabolism. .................................... 3-2
 

3.1.4. Acute Oral Toxicity. ........................................ 3-3
 

3.1.5. Subchronic or Chronic Systemic Toxic Effects. . . . . . . . . . . . . . . . . . . . 3-5
 

3.1.6. Effects on Nervous System. .................................. 3-6
 

3.1.7. Effects on Immune System. .................................. 3-8
 

3.1.8. Effects on Endocrine System. ................................. 3-9
 

3.1.9. Reproductive and Teratogenic Effects. . . . . . . . . . . . . . . . . . . . . . . . . . . 3-9
 

3.1.10. Carcinogenicity and Mutagenicity. . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-11
 

3.1.11. Irritation and Sensitization (Effects on the Skin and Eyes). . . . . . . . . 3-12
 

3.1.12. Systemic Toxic Effects from Dermal Exposure. . . . . . . . . . . . . . . . . 3-13
 

3.1.13. Inhalation Exposure. ...................................... 3-13
 

3.1.14. Inerts and Adjuvants. ..................................... 3-14
 

3.1.15. Impurities and Metabolites.................................. 3-14
 

3.1.16. Toxicological Interactions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-14
 

ii 
 

TABLE OF CONTENTS (continued) 

3.1.17. Studies on PVC Formulations of DDVP. . . . . . . . . . . . . . . . . . . . . . . 3-15
 

3.2. EXPOSURE ASSESSMENT ....................................... 3-18
 

3.2.1. Overview. ............................................... 3-18
 

3.2.2. Workers. ................................................ 3-18
 

3.2.2.1. General Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-18
 

3.2.2.2. Inhalation Exposures ............................... 3-19
 

3.2.2.3. Dermal Exposures ................................. 3-22
 

3.2.3. General Public. ........................................... 3-22
 

3.2.3.1. General Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-22
 

3.2.3.2. Dermal Contact ................................... 3-23
 

3.2.3.3. Oral Exposure to DDVP Strip . . . . . . . . . . . . . . . . . . . . . . . . 3-23
 

3.2.3.4. Oral Exposure to Contaminated Water . . . . . . . . . . . . . . . . . 3-24
 

3.3. DOSE-RESPONSE ASSESSMENT .................................. 3-26
 

3.3.1. Overview ................................................ 3-26
 

3.3.2. Acute Exposures .......................................... 3-26
 

3.3.2.1. Acute Oral ....................................... 3-26
 

3.3.2.2. Acute Dermal ..................................... 3-27
 

3.3.2.3. Acute Inhalation ................................... 3-28
 

3.3 3. Chronic Exposures ........................................ 3-28
 

3.4. RISK CHARACTERIZATION ...................................... 3-29
 

3.4.1. Overview ................................................ 3-29
 

3.4.2. Workers ................................................. 3-29
 

3.4.3. General Public. ........................................... 3-30
 

3.4.4. Sensitive Subgroups........................................ 3-31
 

3.4.5. Connected Actions......................................... 3-32
 

3.4.6. Cumulative Effects......................................... 3-34
 

4. ECOLOGICAL RISK ASSESSMENT ......................................... 4-1
 

4.1. HAZARD IDENTIFICATION ....................................... 4-1
 

4.1.1. Overview. ................................................ 4-1
 

4.1.2. Toxicity to Terrestrial Organisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-1
 

4.1.2.1. Mammals ......................................... 4-1
 

4.1.2.2. Birds ............................................. 4-2
 

4.1.2.3. Terrestrial Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-3
 

4.1.2.4. Terrestrial Plants (Macrophytes) . . . . . . . . . . . . . . . . . . . . . . . 4-4
 

4.1.2.5. Terrestrial Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-4
 

4.1.2.6. Terrestrial Field Studies ...... . .................. 4-4
 

4.1.3. Aquatic Organisms.......................................... 4-4
 

4.1.3.1. Fish .............................................. 4-4
 

iii 
TABLE OF CONTENTS (continued) 

4.1.3.2. Amphibians ....................................... 4-5
 

4.1.3.3. Aquatic Invertebrates ................................ 4-5
 

4.1.3.4. Aquatic Plants ..................................... 4-6
 

4.1.3.5. Other Aquatic Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . 4-6
 

4.2. EXPOSURE ASSESSMENT ........................................ 4-7
 

4.2.1. Overview. ................................................ 4-7
 

4.2.2. Terrestrial Vertebrates. ...................................... 4-7
 

4.2.2.1. Oral Exposure to DDVP Strip . . . . . . . . . . . . . . . . . . . . . . . . . 4-7
 

4.2.2.2. Oral Exposure to Water Contaminated with DDVP . . . . . . . . 4-7
 

4.2.3. Terrestrial Invertebrates ..................................... 4-8
 

4.2.4. Aquatic Species ............................................ 4-8
 

4.3. DOSE-RESPONSE ASSESSMENT ................................... 4-9
 

4.3.1. Overview. ................................................ 4-9
 

4.3.2. Terrestrial Organisms........................................ 4-9
 

4.3.3. Aquatic Organisms.......................................... 4-9
 

4.3.3.1. Fish .............................................. 4-9
 

4.3.3.2. Aquatic Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-10
 

4.4. RISK CHARACTERIZATION ...................................... 4-11
 

4.4.1. Overview. ............................................... 4-11
 

4.4.2. Terrestrial Organisms....................................... 4-11
 

4.4.3. Aquatic Organisms......................................... 4-12
 

4.4.3.1. Fish ............................................. 4-12
 

4.4.3.2. Aquatic Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-12
 

5. REFERENCES ........................................................... 5-1
 

iv 
TABLE OF CONTENTS (continued)
 

LIST OF TABLES
 

Table 2-1: Selected physical and chemical properties of DDVP . . . . . . . . . . . . . . . . . . . . Tables-1 

Table 3-1: Common effects of acetylcholinesterase inhibition . . . . . . . . . . . . . . . . . . . . . Tables-2 

Table 3-2: Parameters used in DDVP air model . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-3 

Table 3-3: Summary of risk characterization for human health risk assessment . . . . . . . . Tables-4 

Table 4-1:Summary of exposure assessments and risk characterization for 

nontarget species . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-5
 

LIST OF FIGURES 

Figure 3-1: Concentration of DDVP in air after the placement of one Shell No-Pest Strip in 

an unventilated room and a poorly ventilated room . . . . . . . . . . . . . . . . Figures-1 

NOTE: Tables followed by figures are placed after Section 5, References. 

LIST OF APPENDICES 

Appendix 1: Application and Optimization of DDVP Inhalation Exposure Model....Appendix 1-1 

Appendix 2: Estimates of dermal absorption rates for DDVP.....................................Appendix 2-1 

LIST OF SUPPLEMENTS 

Supplement 1: EXCEL Worksheets for Human Health and Ecological Risk Assessments of 

DDVP. SERA EXWS 04-43-05-05b, dated August 18, 2004. 

Located at: http://www.fs.fed.us/foresthealth/pesticide/risk.shtml 

v 
AEL 

ACGIH 

AChE 

a.i. 

BCF 

bw 

CBI 

ChE 

CI 

cm 

CNS 

DAA 

DAT 

d.f. 

ECx

EC25 

EC50 

ExToxNet 

F 

FH 

FIFRA 

FOIA 

FQPA 

g 

ha 

HQ 

IAA 

IARC 

IRIS 

ka

ke

kg 

Ko/c 

Ko/w 

Kp

L 

lb 

LC50 

LD50 

LOAEL 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

adverse-effect level 

American Conference of Governmental Industrial Hygienists 

acetylcholinesterase 

active ingredient 

bioconcentration factor 

body weight 

confidential business information 

pseudo-cholinesterase 

confidence interval 

centimeter 

central nervous system 

days after application 

days after treatment 

degrees of freedom 

concentration causing X% inhibition of a process 

concentration causing 25% inhibition of a process 

concentration causing 50% inhibition of a process 

Extension Toxicology Network 

female 

Forest Health 

Federal Insecticide, Fungicide and Rodenticide Act 

Freedom of Information Act 

Food Quality Protection Act 

gram 

hectare 

hazard quotient 

indole-3-acetic acid 

International Agency for Research on Cancer 

Integrated Risk Information System 

absorption coefficient 

elimination coefficient 

kilogram 

organic carbon partition coefficient 

octanol-water partition coefficient 

skin permeability coefficient 

liter 

pound 

lethal concentration, 50% kill 

lethal dose, 50% kill 

lowest-observed-adverse-effect level 

vi 
ACRONYMS, ABBREVIATIONS, AND SYMBOLS (continued) 

LOC level of concern 

m meter 

M male 

MCL mononuclear cell carcinoma 

mg milligram 

mg/kg/day milligrams of agent per kilogram of body weight per day 

mL milliliter 

mM millimole 

MRID Master Record Identification Number 

MSDS material safety data sheet 

MW molecular weight 

NCAP Northwest Coalition for Alternatives to Pesticides 

NCI National Cancer Institute 

NIOSH National Institute for Occupational Safety and Health 

NOAEL no-observed-adverse-effect level 

NOEC no-observed-effect concentration 

NOEL no-observed-effect level 

NOS not otherwise specified 

NRC National Research Council 

NTP National Toxicology Program 

OM organic matter 

OPIDN organophosphate-induced delayed neurotoxicity 

OPP Office of Pesticide Programs 

OPPTS Office of Pesticide Planning and Toxic Substances 

OSHA Occupational Safety and Health Administration 

ppm parts per million 

PVC polyvinyl chloride 

RBC red blood cell 

RED re-registration eligibility decision 

RfD reference dose 

SERA Syracuse Environmental Research Associates 

SGOT serum glutamic oxaloacetic transaminase 

SGPT serum glutamic pyruvic transaminase 

SRC Syracuse Research Corporation 

STS Slow the Spread 

UF uncertainty factor 

U.S. United States 

USDA U.S. Department of Agriculture 

U.S. EPA U.S. Environmental Protection Agency 

USGS U.S. Geological Survey 

WCR water contamination rate 

WHO World Health Organization 

ì micron 

vii 
COMMON UNIT CONVERSIONS AND ABBREVIATIONS
 

To convert ... Into ... Multiply by ... 

acres hectares (ha) 0.4047 

acres square meters (m ) 2 4,047 

atmospheres millimeters of mercury 760 

centigrade Fahrenheit 1.8 EC+32 

centimeters inches 0.3937 

cubic meters (m ) 3 liters (L) 1,000 

Fahrenheit centigrade 0.556 EF-17.8 

feet per second (ft/sec) miles/hour (mi/hr) 0.6818 

gallons (gal) liters (L) 3.785 

gallons per acre (gal/acre) liters per hectare (L/ha) 9.34 

grams (g) ounces, (oz) 0.03527 

grams (g) pounds, (oz) 0.002205 

hectares (ha) acres 2.471 

inches (in) centimeters (cm) 2.540 

kilograms (kg) ounces, (oz) 35.274 

kilograms (kg) pounds, (lb) 2.2046 

kilograms per hectare (hg/ha) pounds per acre (lb/acre) 0.892 

kilometers (km) miles (mi) 0.6214 

liters (L) cubic centimeters (cm ) 3 1,000 

liters (L) gallons (gal) 0.2642 

liters (L) ounces, fluid (oz) 33.814 

miles (mi) kilometers (km) 1.609 

miles per hour (mi/hr) cm/sec 44.70 

milligrams (mg) ounces (oz) 0.000035 

meters (m) feet 3.281 

ounces (oz) grams (g) 28.3495 

ounces per acre (oz/acre) grams per hectare (g/ha) 70.1 

ounces per acre (oz/acre) kilograms per hectare (kg/ha) 0.0701 

ounces fluid cubic centimeters (cm ) 3 29.5735 

pounds (lb) grams (g) 453.6 

pounds (lb) kilograms (kg) 0.4536 

pounds per acre (lb/acre) kilograms per hectare (kg/ha) 1.121 

pounds per acre (lb/acre) mg/square meter (mg/m ) 2 112.1 

pounds per acre (lb/acre) ìg/square centimeter (ìg/cm ) 2 11.21 

pounds per gallon (lb/gal) grams per liter (g/L) 119.8 

square centimeters (cm ) 2 square inches (in ) 2 0.155 

square centimeters (cm ) 2 square meters (m ) 2 0.0001 

square meters (m ) 2 square centimeters (cm ) 2 10,000 

yards meters 0.9144 

Note: All references to pounds and ounces refer to avoirdupois weights unless otherwise specified. 

viii 
CONVERSION OF SCIENTIFIC NOTATION
 

Scientific Decimal Verbal 

Notation Equivalent Expression 

1 @ 10-10 0.0000000001 One in ten billion 

1 @ 10-9 0.000000001 One in one billion 

1 @ 10-8 0.00000001 One in one hundred million 

1 @ 10-7 0.0000001 One in ten million 

1 @ 10-6 0.000001 One in one million 

1 @ 10-5 0.00001 One in one hundred thousand 

1 @ 10-4 0.0001 One in ten thousand 

1 @ 10-3 0.001 One in one thousand 

1 @ 10-2 0.01 One in one hundred 

1 @ 10-1 0.1 One in ten 

1 @ 100 1 One 

1 @ 101 10 Ten 

1 @ 102 100 One hundred 

1 @ 103 1,000 One thousand 

1 @ 104 10,000 Ten thousand 

1 @ 105 100,000 One hundred thousand 

1 @ 106 1,000,000 One million 

1 @ 107 10,000,000 Ten million 

1 @ 108 100,000,000 One hundred million 

1 @ 109 1,000,000,000 One billion 

1 @ 1010 10,000,000,000 Ten billion 

ix 
EXECUTIVE SUMMARY
 

OVERVIEW
 

The USDA uses DDVP in its program to manage the gypsy moth. The primary use of DDVP is 

as a component in the pheromone baited milk carton style traps that are used primarily for 

surveying and monitoring gypsy moth populations. Because of this a very limited use in USDA 

programs, the potential for exposures to humans or nontarget ecological species is extremely 

limited. Because of this limited use and limited potential for exposure, this risk assessment 

focuses on the information that has the greatest impact on potential hazard rather than a summary 

of all of the available information that is available on DDVP and this risk assessment utilizes 

several detailed reviews conducted by agencies responsible for assessing chemical risks 

PROGRAM DESCRIPTION 

In USDA programs for the control of the gypsy moth, DDVP is used only in a 1" x 4" inch 

polyvinyl chloride (PVC) strip that contains 590 mg of DDVP. These strips are used to kill 

insects that are attracted to and enter milk carton style traps baited with the gypsy moth 

pheromone. Typically milk carton traps are deployed in widely spaced grids (inter-trap distances 

ranging from 500 m to 7 km) to survey for the presence of gypsy moth populations in the STS or 

eradication areas. Only rarely are milk carton traps deployed in mass trapping grids to control 

isolated infestations. When used in mass trapping control efforts, milk carton traps are deployed 

in tightly spaced grids (inter-trap distance of 20 to 30 meters). Mass trapping is a rarely used 

eradication tactic that targets low-density infestations (<10 egg masses per acre) occupying 

relatively small areas (<100 acres) . 

HUMAN HEALTH RISK ASSESSMENT 

Hazard Identification – DDVP is an organophosphorus insecticide that works by inhibiting 

cholinesterase. DDVP has been used since the early 1960's and has been the subject of many 

toxicity studies and review articles. Information is available on a number of case reports of 

accidental and suicidal exposures as well as human monitoring data from normal use. The 

toxicity of DDVP has been adequately evaluated using laboratory animals, although not all of 

these studies are available in the open literature. 

DDVP is readily absorbed into the body of mammals via all routes of exposure, where it is 

rapidly metabolized and eliminated. In general, the systemic effects observed after oral, 

inhalation, or dermal exposure of humans or laboratory animals to DDVP result from the 

inhibition of acetylcholinesterase. Inhibition of this enzyme in mammalian systems produces a 

variety of systemic effects, including salivation, urination, lacrimation, convulsions, increased 

bronchial secretions, respiratory depression, and even death. The nature and magnitude of the 

toxic effects produced by a given exposure to DDVP by any route are directly related to the dose 

and rate at which the exposure occurs. In the case of the USDA programs for the management of 

the gypsy moth, the use of milk carton traps containing Vaportape II (slow-release of DDVP 

from PVC strips) essentially precludes rapid exposures to high doses of DDVP. 

x 
Short-term animal studies have shown that oral exposures to doses below about 0.5 mg/kg-day 

(or inhalation exposures to 1–2 mg/m3) do not result in meaningful reductions in cholinesterase 

activity. Experiments in laboratory mammals that were exposed to DDVP during pregnancy (by 

oral or inhalation route) did not show any effect on fertility or health of the offspring, even at 

levels that produced maternal toxicity. The latest evaluation of data from assays for 

carcinogenicity and genetic toxicity classify DDVP as a “suggestive” carcinogen and determined 

that a quantitative assessment of cancer risk is not applicable. The literature contains some data 

suggesting that contact dermatitis (as well as cross-sensitization to other pesticides) may occur; 

although, this appears to be an infrequent occurrence in the general population. 

Exposure Assessment – Under normal conditions, exposure to both workers and members of the 

general public should be negligible. Workers will handle DDVP strips only during the assembly 

of milk cartoon traps. If workers wear gloves and assemble the traps outdoors or in very well 

ventilated rooms, both inhalation and dermal exposures should be negligible. Inhalation 

exposure to DDVP during transport of the traps should also be negligible if the traps are not 

transported inside of the passenger compartments of vehicles. Worker exposures will also be 

limited in most programs because foil wrapping in which the strip is distributed will not be 

removed until after the trap is transported to the field. Milk carton traps will generally be placed 

about four feet above the ground (Leonard 2004) and exposure of members of the general public 

to DDVP contained in the milk carton traps should also be negligible except in the case of 

intensional tampering. 

Notwithstanding the above assertions, exposure assessments are developed for workers who do 

not use gloves in the assembly of the milk carton traps and who assemble the traps indoors and 

transport the traps in the passenger compartments of vehicles. All of these exposure scenarios 

should be considered atypical and some are extreme. The intent is to illustrate the consequences 

of mishandling or imprudent handling. During assembly, the central estimate of dermal 

exposures in workers not wearing gloves leads to an absorbed dose of about 0.0008 mg/kg with a 

range of about 0.0003 mg/kg to 0.004 mg/kg. Inhalation exposures in workers may be highly 

variable depending on the ventilation rates in an enclosed space and the number of traps that are 

handled. Based on the handling and transport of 75 traps, inhalation exposures could reach up to 

3 3 about 0.6 mg/m in an enclosed and unventilated room and up to about 1.8 mg/m in the

passenger compartment of a vehicle. These exposure assessments are based on several site and 

situation specific assumptions which are intended to reflect plausible upper bounds of exposures. 

Exposure assessments are also developed for children who might come in contact with an 

accidentally discarded or misplaced DDVP strip. Estimated dermal doses are much higher than 

those for workers: a central estimate of about 0.02 mg/kg with a range of 0.003 mg/kg to 0.1 

mg/kg. Oral exposures from a small child sucking on the pest strip are about a factor of 10 

higher than dermal exposures: a central value of about 0.2 mg/kg with a range of 0.04 mg/kg to 

0.6 mg/kg. 

xi 
Under normal circumstances, the use of DDVP in PVC strips is not likely to result in 

contamination of water or other materials that might be consumed by members of the general 

public. Nonetheless, an exposure assessment is developed for the accidental contamination of a 

small pond by a pest strip. In this scenario, dose estimates range from about 0.000003 mg/kg to 

0.00007 mg/kg with a central estimate of about 0.00001 mg/kg. 

Dose-Response Assessment – The extensive toxicology data base has been evaluated by a 

number of governmental organizations including the U.S. EPA, the Agency for Toxic Substances 

and Disease Registry (ATSDR), National Institute for Occupational Safety and Health (NIOSH), 

Occupational Safety and Health Administration, and the World Health Organization. Following 

the approach taken in most USDA risk assessments, these sources are used for selecting levels of 

acceptable exposure. Because all of the scenarios considered in this risk assessment involve only 

acute exposures, only acute exposure criteria are considered. 

For both oral and dermal exposures, the acute RfD established by the U.S. EPA, 0.0017 mg/kg, is 

used for the risk characterization. This is based on an acute oral NOAEL of 0.5 mg/kg from a 

study in rats with the application of an uncertainty factor of 300. Acute exposure criteria 

proposed by other groups are comparable to but somewhat higher than the acute RfD. Because 

some of the accidental acute exposures may substantially exceed the acute RfD, some attempt is 

made to characterize the consequences of high oral exposures. A human NOAEL of 1 mg/kg for 

AChE inhibition has been identified. While this NOAEL is not used to modify the acute RfD, it 

can be used to assess plausible consequences of exceeding the RfD. The human data on DDVP, 

although extensive, are not sufficient to identify a minimal lethal dose. For the current risk 

assessment, the lowest reported lethal dose (16 mg/kg) is used to assess the plausibility of 

observing serious adverse effects in cases of accidental over-exposure to DDVP. 

A number of inhalation criteria for DDVP are available. Since potentially significant inhalation 

3 exposures are likely only in workers, the occupational exposure criteria of 0.1 mg/m proposed

by American Conference of Governmental Industrial Hygienists is used. This value is a factor of 

10 below the occupational criteria proposed by NIOSH and OSHA. 

Risk Characterization – In most cases, exposures to both workers and members of the general 

public should be negligible. If workers take prudent steps to limit both dermal and inhalation 

exposures, the likelihood of exposures to DDVP reaching a level of concern appears to be very 

low. Similarly, members of the general public should not be exposed to substantial amounts of 

DDVP. The DDVP is contained within a PVC strip to insure that the active ingredient is slowly 

released over a long period of time. The strip, in turn, is placed within a trap and the trap is 

placed so that the that will not be accessed except in the case of intentional tampering or trap 

monitoring. 

Nonetheless, this risk assessment develops exposure scenarios for both workers and members of 

the general pubic that are intended to illustrate the potential effects of mishandling or tampering 

with DDVP strips. For workers, the greatest risks are associated with inhalation exposures from 

xii 
assembling the traps in enclosed and poorly ventilated spaces or transporting the traps in the 

passenger compartments of vehicles. These risks can be readily avoided. Dermal exposures can 

also lead to lesser but sill undesirable levels of exposure. For members of the general public, all 

of the exposure scenarios are accidental and some are extreme. The most likely of these is the 

accidental contamination of a small body of water. This scenario leads to exposures that are 

below the level of concern by a factor of about 25. If a child were to come into contact with a 

DDVP strip, however, both dermal and oral exposures could substantially exceed a level of 

concern. While such exposures should clearly be avoided, it does not seem likely that frank signs 

of toxicity would be observed. This is consistent with human experience in the use of DDVP 

resin strips. 

ECOLOGICAL RISK ASSESSMENT 

Hazard Identification – The available data suggest that invertebrates are more sensitive to 

DDVP than other organisms. For example, the oral LD50 in honey bees is 0.29 ìg/g bee, and the 

topical LD is 0.65 ìg/g bee. DDVP is also toxic to birds with an oral LD value of < 10 mg/kg 50 50 

for the most sensitive species. Short-term repeat dose studies in mammals found that oral 

exposures to doses below about 0.5 mg/kg-day or inhalation exposures to 1–2 mg/m3 generally 

do not result in adverse effects. 

Aquatic animals are also sensitive to DDVP and, as with terrestrial animals, invertebrates may be 

more sensitive than vertebrates. The lowest reported LC50 value in fish is approximately 0.2 

mg/L. Some aquatic invertebrates are much more sensitive to DDVP than fish. For daphnids, 

the most sensitive group of invertebrate species, reported EC50 values range from 0.00007 mg/L 

to 0.00028 mg/L. 

The majority of the toxicity data in ecological receptors is limited to free DDVP, rather than a 

slow-release formulation such as the Vaportape II product used in USDA programs for control of 

the gypsy moth. Hence, the toxicity values reported for indicator species will likely be 

conservative (i.e., suggest greater toxicity) as compared to Vaportape II. U.S. EPA has assessed 

the ecological effects of DDVP; however, the exposures assessed by U.S. EPA are not specific to 

formulations where DDVP is encapsulated in PVC resin. In general, aside from those organisms 

that enter the milk carton trap or those that remove the PVC strip form the trap, toxicity resulting 

from exposure of ecological receptors to DDVP in Vaportape II milk carton traps is not likely. 

Exposure Assessment – As in the human health risk assessment, exposure of terrestrial 

mammals to DDVP from the VaporTape strips used in milk carton traps is likely to be negligible 

under most circumstances. Nonetheless, it is conceivable that some mammals such as racoons or 

bears could easily access and tamper with the milk carton trap. Depending on the proportion of 

the DDVP strip that is consumed, doses (as DDVP in the PVC strip) are estimated to range from 

10.5 mg/kg (10% of strip) to 105 mg/kg (100% of strip) and the central estimate is taken as 31.6 

mg/kg (30% of strip). In addition, contamination of water with a pest strip is plausible, although 

probably rare, and is considered in a manner similar to the corresponding scenario in the human 

health risk assessment (Section 3.2.3.4). This scenario is based on the consumption of 

xiii 
contaminated water by a small mammal and the dose to the animal is estimated at about 0.00003 

mg/kg with a range of 0.000009 mg/kg to 0.00009 mg/kg. Other exposure scenarios for 

terrestrial vertebrates, while possible, seem far less plausible and are not considered 

quantitatively. No quantitative exposure assessments for terrestrial invertebrates are developed 

because the milk carton trap will attract only male gypsy moths. Nontarget insects that 

incidentally enter the trap are likely to be killed by exposure to the DDVP vapor. Exposures to 

aquatic species are based on the same water concentrations used for terrestrial species: 0.000177 

mg/L with a range of 0.000059 mg/L to 0.00059 mg/L. 

Dose-Response Assessment – Given the limited nature of the use of DDVP in programs to 

control the gypsy moth and consequent limited number of exposure assessments, the doseresponse assessment for DDVP is relatively simple. For terrestrial mammals, a value of 240 

mg/kg from a study using DDVP in a PVC formulation is used for direct exposure to the DDVPPVC strip – i.e., a raccoon tampering with a milk carton trap and consuming all or part of the 

DDVP strip. At the dose of 240 mg/kg, no mortality or frank signs of AChE inhibition were 

observed. For the contaminated water scenario, the NOAEL of 0.5 mg/kg from a study involving 

exposure to free or unformulated DDVP is used. This NOAEL is from the study that forms the 

basis for the acute RfD used in the human health risk assessment. Although DDVP is classified 

as highly toxic to fish, the estimated levels of acute exposure for fish are far below the 30-day 

NOEC of 0.03 mg/L. Thus, this value is used for all fish and no attempt is made to consider 

differences in sensitivity among fish. A somewhat different approach is taken with aquatic 

invertebrates, some of which are more sensitive to DDVP than fish by a factor of over 2500. 

Risks to sensitive species of aquatic invertebrates – i.e., daphnids and other small arthropods – 

are characterized based on the lowest reported LC50 value, 0.00007 mg/L from a 48-hour 

bioassay in Daphnia pulex. Some other groups of aquatic invertebrates, such as snails, appear to 

be much less sensitive than small arthropods. Risks to such tolerant species are based on a LC 50 

value of 21 mg/L in a freshwater snail. 

Risk Characterization – As with the human health risk assessment, it is anticipated that typical 

exposures and consequent risks to nontarget species should be negligible. As with the human 

health risk assessment, it is anticipated that typical exposures and consequent risks to most 

nontarget species should be negligible. The containment of the DDVP within a slow release 

PVC strip combined with the target specific nature of pheromone baited traps should reduce the 

risks of inadvertent effects in non-target species. Other insects and arthropods that may 

inadvertently enter the trap will probably be killed by DDVP vapor. While such inadvertent 

contact may occur, it is not likely to impact substantial numbers of nontarget insects or 

arthropods. 

Because of the limited use of DDVP, a relatively small number of exposure scenarios – all of 

which might be considered accidental or incidental – are developed. For terrestrial mammals, 

contact with the pest strip could occur by an animal directly tampering with a trap or by an 

animal consuming water that had been accidentally contaminated with a DDVP strip. Adverse 

effects would not be expected in either case. In the case of accidental contamination of a small 

xiv 
body of water with a DDVP strip, concentrations of DDVP in the water would be below the level 

of concern for fish by factors of about 50 to 500. Some aquatic invertebrates, however, might be 

affected. For the most sensitive species of aquatic invertebrates – i.e., small aquatic arthropods 

such as daphnids – exposures could substantially exceed laboratory LC50 values by factors of up 

to about 8. Exposures to tolerant aquatic invertebrates – such as snails – would be below a level 

of concern by a substantial margin – i.e., factors of about 30,000 to 300,000. 

The exposure assessments that serve as the bases for these risk characterizations are highly 

dependent on specific conditions – i.e., how much DDVP was in the strip at the time that the 

contamination occurred and the size of the body of water that was contaminated. Because the 

hydrolysis of DDVP in water is rapid, the estimates of adverse effects in some aquatic 

invertebrates would probably apply only to a very limited area near the pest strip rather than to 

the larger area of the body of water that is contaminated. 

xv 
1. INTRODUCTION
 

The USDA uses DDVP in its program to manage the gypsy moth. The primary use of DDVP is 

as a component in the pheromone baited milk carton style traps that are used primarily for 

surveying and monitoring gypsy moth populations. This document is an update to a risk 

assessment prepared in 1995 (USDA 1995a,b) and provides risk assessments for human-health 

effects and ecological effects to support an assessment of the environmental consequences of 

these uses. 

This document has four chapters, including the introduction, program description, risk 

assessment for human health effects, and risk assessment for ecological effects or effects on 

wildlife species. Each of the two risk assessment chapters has four major sections, including an 

identification of the hazards associated with DDVP, an assessment of potential exposure to the 

product, an assessment of the dose-response relationships, and a characterization of the risks 

associated with plausible levels of exposure. These are the basic steps recommended by the 

National Research Council of the National Academy of Sciences (NRC 1983) for conducting and 

organizing risk assessments. 

Although this is a technical support document and addresses some specialized technical areas, an 

effort was made to ensure that the document can be understood by individuals who do not have 

specialized training in the chemical and biological sciences. Certain technical concepts, 

methods, and terms common to all parts of the risk assessment are described in plain language in 

a separate document (SERA 2001). 

The human health and ecological risk assessments presented in this document are not, and are not 

intended to be, comprehensive summaries of all of the available information. This is particularly 

true for DDVP used in gypsy moth programs. There is an extremely large and relatively complex 

database of literature on DDVP. For example, TOXLINE, one of several commonly used 

commercial databases containing information on toxic chemicals, has over 14,000 citations on 

DDVP. DDVP, however, has a very limited use in USDA gypsy moth programs (Section 2) and 

the potential for exposures to humans (Section 3.2) or nontarget ecological species (Section 4.2) 

is extremely limited. Because of this limited use and limited potential for exposure, this risk 

assessment focuses on the information that has the greatest impact on potential hazard rather than 

a summary of all of the available information that is available on DDVP and this risk assessment 

utilizes several detailed reviews conducted by agencies responsible for assessing chemical risks 

(e.g., ATSDR 1997; U.S. EPA 1999a, 2000a,b; WHO 1988, 1989). 

This risk assessment involves numerous calculations. Many of the calculations are relatively 

simple and the very simple calculations are included in the body of the document. Some of the 

calculations, however, are complex. For the more complex calculations, worksheets are included 

as an attachment to the risk assessment. The worksheets provide the detail for the estimates cited 

in the body of the document. The worksheets for DDVP are contained in an EXCEL workbook 

1-1
 
and are included as Supplement 1 to this risk assessment and general documentation for the use 

of these worksheets is given in SERA (2004). 

The USDA will update this and other similar risk assessments on a periodic basis and welcomes 

input from the general public on the selection of studies included in the risk assessment. This 

input is helpful, however, only if recommendations for including additional studies specify why 

and/or how the new or not previously included information would be likely to alter the 

conclusions reached in the risk assessments. 

1-2
 
2. PROGRAM DESCRIPTION
 

2.1. OVERVIEW
 

DDVP is an organophosphate insecticide that acts by inhibiting acetylcholinesterase, an enzyme 

that is very important in the nervous system of all vertebrates and many invertebrates including 

all arthropods. Thus, DDVP is not specific to the gypsy moth or other insects. In USDA 

programs for the control of the gypsy moth, DDVP is used only in a 1" x 4" inch polyvinyl 

chloride (PVC) strip that contains 590 mg of DDVP. These strips are used to kill insects that are 

attracted to and enter milk carton style traps baited with the gypsy moth pheromone. Typically 

milk carton traps are deployed in widely spaced grids (inter-trap distances ranging from 500 m to 

7 km) to survey for the presence of gypsy moth populations in the STS or eradication areas. 

Only rarely are milk carton traps deployed in mass trapping grids to control isolated infestations. 

When used in mass trapping control efforts, milk carton traps are deployed in tightly spaced grids 

(inter-trap distance of 20 to 30 meters). Mass trapping is a rarely used eradication tactic that 

targets low-density infestations (<10 egg masses per acre) occupying relatively small areas (<100 

acres) . 

2.2. CHEMICAL DESCRIPTION AND COMMERCIAL FORMULATIONS 

DDVP is the common name for O,O-dimethyl O-(2,2-dichlorovinyl) phosphate: 

Other synonyms for DDVP as well as selected chemical and physical properties of DDVP are 

summarized in Table 2-1. 

DDVP is a contact and stomach organophosphate insecticide (Gallo and Lawryk 1991, IARC 

1991). As detailed further in the human health risk assessment (Section 3) and the ecological 

risk assessment (Section 4), DDVP acts by inhibiting acetylcholinesterase, an enzyme that is very 

important in the nervous system of all vertebrates (including humans) and most other animals 

including all arthropods. 

DDVP is currently undergoing reregistration (http://www.epa.gov/pesticides/op/ddvp.htm; 

Mennear 1998) and is being considered in the U.S. EPA’s cumulative risk assessment of 

organophosphates (http://www.eps.gov/pesticides). 

Various DDVP pest strips for residential or industrial use have been registered with the U.S. 

EPA and are manufactured by AMVAC Chemical Corporation, Loveland Industries, and 

Spectrum Group (http://www.cdpr.ca.gov/docs/pressrls/ddvp.htm). However, the only strip used 

by the USDA in gypsy moth programs is the Vaportape II strip provided by Hercon 

Environmental Corp, Emigsville, PA (Hercon 1993). A contract for the supply of these strips to 

2-1
 
the USDA gypsy moth program was awarded to Hercon Environmental Corp on March 23, 1999 

(www.fbodaily.com/cbd/archive/1999/03 (March)23/Mar-1999/87awdoo1.htm). 

Vaportape II is distributed in packages of 50 strips, each of which comes in a protective pouch. 

Each strip consists of a 1" x 4" inch red, multi-layered polyvinyl chloride (PVC) strip containing 

590 mg of DDVP. The average thickness of the strip is 67.5 mil with a range of 65–70 mil or 

0.0675 inches with a range of 0.065–0.07 inches (Hercon 1994). Additional details concerning 

the composition of the strips have been disclosed to U.S. EPA (Health-Chem Corporation 19??; 

Herculite Products Incorporated 19??a,b; Starner 1993). Note that the 19?? designation indicates 

that the material is not dated and that the U.S. EPA cannot determine when the information was 

submitted. This is not uncommon for submissions that occurred in the early 1970's. The details 

of the information contained in these submissions are classified as CBI (confidential business 

Information) under Section 7(d) and Section (10) of FIFRA and this information cannot be 

specifically disclosed in this risk assessment. 

The product label specifies that in addition to DDVP, each strip contains 0.75% compounds that 

are related to DDVP and 89.25% inerts (Hercon 2004). Further details are not provided on the 

label; nonetheless the impurities in commercial DDVP have been characterized (Gillett and 

others 1972a, IARC 1991). The impurities include: Dipterex (O,O-dimethyl 

2,2,2-trichloro-1-hydroxyethylphosphonate); O,O-dimethyl 2-chlorovinyl phosphate; 

O,O-dimethyl methylphosphonate; O,O,O-trimethyl phosphate; and trichloroacetaldehyde. 

These impurities are known to be or are likely to be toxic (Gillett and others 1972a, WHO 1989). 

These impurities are encompassed in the risk assessment because the dose-response assessment 

is based on studies that used commercial grade DDVP. Consequently, the results of these studies 

are directly applicable to the risk assessment for human health (Section 3) and ecological effects 

(Section 4). 

2.3. APPLICATION METHODS 

The Vaportape II strips are used as an insecticide in large capacity pheromone traps to monitor 

gypsy moth populations. DDVP is also used in a similar way in monitoring populations of the 

beet armyworm (Lopez 1998). 

In order to minimize the ecological effects and human health effects of gypsy moth infestations, 

the USDA adopted various intervention strategies that are roughly categorized as suppression, 

eradication, and Slow the Spread (STS). Suppression efforts are conducted by the USDA Forest 

Service in areas of well established gypsy moth infestations to combat or interdict periodic gypsy 

moth population outbreaks. Eradication efforts are conducted by USDA/APHIS to completely 

eliminate gypsy moth populations in areas where new populations of the gypsy moth are found. 

Slow the spread, as the name implies, is a program to reduce the expansion of gypsy moth 

populations from areas of established populations to adjacent non-infested areas (Liebhold and 

McManus 1999). The STS project is the primary user of DDVP and milk carton traps. STS has 

purchased DDVP in the following amounts: 2002 - 540 packs (540x50 strips=27,000 strips); 

2003 - 540 packs (27,000 strips); 2004 - 500 packs (25,000 strips) (Leonard 2004). 

2-2
 
As in the previous gypsy moth programs, a Vaportape II strip is contained in the milk carton trap 

together with together with a slow release dispenser containing disparlure, the gypsy moth 

pheromone. The milk carton traps containing the strips are placed in selected areas to monitor 

gypsy moth infestations. When used in eradication efforts for mass trapping, milk carton traps 

are typically used only in low density infestations – i.e., 10 egg masses per acre or less. In 

addition, because of the labor involved in mass trapping, this method is applied to relatively 

small areas – i.e., about 100 acres or less (USDA 2001, p. 1-7 to 1-8). 

As discussed in the exposure assessments for human health (Section 3.2) and ecological effects 

(Section 4.2), the nature of the exposures to humans and other nontarget species will typically be 

extremely small and it is unlikely that significant exposures will occur under normal 

circumstances. For workers, the nature of exposure to DDVP depends on program handling 

practices, which vary from state to state. In most cases, dermal and inhalation exposure will be 

minimal, provided that recommended work practices are followed. In some states, inhalation 

exposure will be minimal because strip installation takes place outdoors, at the trap placement 

site. In other states, traps may be assembled the day before placement. Even so, the workers are 

instructed to assemble the traps only in a well-ventilated area, and the traps are sealed in plastic 

bags after assembly and prior to transport. Dermal exposure is also likely to be minimal. In most 

states, workers are given plastic gloves and instructed to use them. In other states, workers are 

instructed to touch only the plastic wrapper in which the strip is shipped. 

2-3
 
3. HUMAN HEALTH RISK ASSESSMENT
 

3.1. HAZARD IDENTIFICATION 

3.1.1. Overview 

DDVP is an organophosphorus insecticide that works by inhibiting cholinesterase. DDVP has 

been used since the early 1960's and has been the subject of many toxicity studies and review 

articles. Information is available on a number of case reports of accidental and suicidal 

exposures as well as human monitoring data from normal use. The toxicity of DDVP has been 

adequately evaluated using laboratory animals, although not all of these studies are available in 

the open literature. 

DDVP is readily absorbed into the body of mammals via all routes of exposure, where it is 

rapidly metabolized and eliminated. In general, the systemic effects observed after oral, 

inhalation, or dermal exposure of humans or laboratory animals to DDVP result from the 

inhibition of acetylcholinesterase. Inhibition of this enzyme in mammalian systems produces a 

variety of systemic effects, including salivation, urination, lacrimation, convulsions, increased 

bronchial secretions, respiratory depression, and even death. The nature and magnitude of the 

toxic effects produced by a given exposure to DDVP by any route are directly related to the dose 

and rate at which the exposure occurs. In the case of the USDA programs for the control of the 

gypsy moth, the use of milk carton traps containing Vaportape II (slow-release of DDVP from 

PVC strips) essentially precludes rapid exposures to high doses of DDVP. 

Short-term animal studies have shown that oral exposures to doses below about 0.5 mg/kg-day 

(or inhalation exposures to 1–2 mg/m3) do not result in meaningful reductions in cholinesterase 

activity. Experiments in laboratory mammals that were exposed to DDVP during pregnancy (by 

oral or inhalation route) did not show any effect on fertility or health of the offspring, even at 

levels that produced maternal toxicity. The latest evaluation of data from assays for 

carcinogenicity and genetic toxicity classify DDVP as a “suggestive” carcinogen and determined 

that a quantitative assessment of cancer risk is not applicable. The literature contains some data 

suggesting that contact dermatitis (as well as cross-sensitization to other pesticides) may occur; 

although, this appears to be an infrequent occurrence in the general population. 

3.1.2. Mechanism of Action 

The mechanism of action of DDVP in target organisms and its principal toxic effects in humans 

and animals result from inhibiting neural acetylcholinesterase (AChE). DDVP shares this 

mechanism of action with other organophosphate insecticides. A number of excellent reviews on 

the mechanism of action of the organophosphate insecticides are available in various texts (Wills 

1972; Gallo and Lawryk 1991; Taylor 1996; Ecobichon 2001). The AChE enzyme is present at 

cholinergic synapses (spaces between the nerve cells) throughout the nervous systems, and it is 

responsible for hydrolyzing acetylcholine released from the pre-synaptic terminal. If this enzyme 

is inhibited, acetylcholine accumulates in the synapse, resulting in increased stimulation of the 

postsynaptic neuron and cholinergic overstimulation. The consequences of increased cholinergic 

activity in various organ systems are listed in Table 3-1. These classical symptoms of 

3-1
 
organophosphate neurotoxicity increase in severity and rapidity of onset in a dose-dependent 

manner. 

Acetylcholinesterase is also present in erythrocytes where it is known as erythrocyte or red blood 

cell acetylcholinesterase (RBC AChE). In vitro assays have found that the erythrocyte and neural 

forms of AChE are inhibited to roughly the same extent by exposure to DDVP (ATSDR 1997). 

Measurement of RBC AChE is used as a surrogate of the inhibition of neural AChE. One of the 

major diagnostic tools and measures of exposure to DDVP and other organophosphate 

insecticides is the determination of cholinesterase activity in various tissues, most often red blood 

cells and plasma ( Ecobichon 2001; Gallo and Lawryk 1991; Murphy 1980). Plasma 

cholinesterase, sometimes referred to as pseudo-cholinesterase or ChE, is produced by the liver 

and differs from AChE in structure and substrates (ATSDR 1993). Although the normal 

physiological role of plasma ChE is not known, it is also inhibited by DDVP and is often used as 

a marker for exposure. Inhibition of RBC AChE is generally regarded as a more clinically 

significant index of organophosphate exposure, compared with inhibition of plasma ChE, as 

plasma ChE is inhibited by DDVP at lower levels of exposure than required to inhibit neural or 

erythrocyte AChE (ATSDR 1997). 

3.1.3. Kinetics and Metabolism 

DDVP is a small, lipid-soluble molecule (see Table 2-1) that is readily absorbed by passive 

diffusion through the lungs, gastrointestinal tract, or skin. Little information is available on the 

pulmonary absorption rate of DDVP, but it appears to be rapidly absorbed by the inhalation as 

well as oral and dermal routes of exposure. Due to the rapid degradation of DDVP by tissue 

esterases, particularly in the liver and the serum, measuring DDVP in vivo is difficult. Laws 

(1966) reported that DDVP is absorbed primarily by hepatic portal venous system after oral 

administration and is subject to first pass metabolism by the liver. Because of the difficulty in 

measuring DDVP in vivo, the rate of absorption is typically inferred from the time to onset of 

clinical signs of AChE inhibition (see Table 3-1). Determination of the tissue distribution of 

DDVP is also difficult to study because of rapid metabolism, but the data do not suggest 

preferential distribution or sequestration in any tissue (ATSDR 1997). A compartmental model 

has been proposed by Garcia-Repetto et al. (1995) to describe the toxicokinetics of DDVP 

following oral exposure. The model was composed of two compartments: central and peripheral. 

The central compartment was blood, and the peripheral compartment encompassed adipose, 

muscle, and liver. 

3.1.3.1. Oral Absorption – Oral absorption of DDVP is rapid. Acute oral toxicity studies have 

demonstrated toxic effects from oral DDVP exposure within minutes. ATSDR (1997) noted that 

animal studies demonstrated lethality from single gavage doses of DDVP within 9 minutes for 

Swiss mice and 15–30 minutes in crossbred swine; signs of cholinergic toxicity (vomiting and 

diarrhea) were noted in greyhound dogs 7–15 minutes after receiving oral doses of DDVP in 

gelatin capsules. Based on a suicide case, Shimizu et al. (1996) have reported the tissue 

distribution of DDVP in humans following oral exposure. Tissue to blood ratios in this 

individual ranged from <1 for brain and liver to 28 for heart and 115 for the spleen. The authors 

3-2
 
reported that the high-tissue concentrations in the heart and spleen were likely due to diffusion 

from the stomach to nearby organs (postmortem, the stomach contained approximately 250 mL 

of fluid equivalent to 300 g of DDVP). Studies in swine treated with DDVP-impregnated PVC 

pellets (veterinary use as anthelminthic) show that DDVP is absorbed from the PVC resin after 

oral exposure (Jacobs 1968, Potter et al. 1973). 

3.1.3.2. Dermal Absorption – No studies have been found on the dermal absorption rate of 

DDVP in humans. As a small, lipid-soluble compound (see Section 2.2), DDVP would likely be 

rapidly absorbed through the skin. Dermal absorption in rats has been studied by Jeffcoat (1990). 

Groups of rats were dosed with 14C-DDVP at 3.6, 36, and 360 μg/rat by applying the compound 

to the shaved back. The treated area was isolated with a protective cover for a 10-hour period. 

After 10 hours, the remaining DDVP was washed from the treated surface and animals were 

sacrificed over 24- to 102-hour periods. Based on the 14C recovered from the rats, the amount 

penetrating the skin ranged from 21.9 to 30.1% with no substantial variation among dose groups. 

For this type of a study, first order dermal absorption coefficients (k) can be calculated as: 

k = -ln(1-f)/t 

where f is the fraction absorbed and t is the duration of exposure. Based on absorption fractions 

of 0.219 to 0.301, the first-order dermal absorption rates can be calculated as 0.025 hour-1 

-1 [-ln(1-0.219)/10 hours] to 0.036 hour [-ln(1-0.301)/10 hours]. These calculations are based on 

the cumulative amount of DDVP recovered from urine, feces, expired air, blood, carcass, and 

treated skin). Excluding treated skin, only 6.4 to 11.4% of the dose was actually absorbed. 

These correspond to first order dermal absorption rates of 0.0066 hour-1 [-ln(1-0.064)/10 hours] 

to 0.012 hour-1 [-ln(1-0.114)/10 hours] and these estimates are consistent with the dermal 

absorption rate selected by EPA (2000a) for occupational and residential exposures (11% in 10 

hours of exposure). 

3.1.3.3. Metabolism – As noted above, DDVP is rapidly degraded by tissue esterases, 

particularly in the liver and the serum. The products of the esterase-catalyzed degradation of 

DDVP are dimethyl phosphate and dichloroacetaldehyde. Dimethyl phosphate is excreted in the 

urine, while dichloroacetaldehyde can be reduced to dichloroethanol or dehalogenated to glyoxal, 

which enters 2-carbon metabolism. Dichloroethanol is either conjugated to glucuronic acid and 

excreted in the urine or dehalogenated and further metabolized. There is also evidence that 

DDVP can be demethylated in a glutathione-dependent reaction (WHO 1989, ATSDR 1997). 

The in vitro half-life of DDVP in human blood is about 10 minutes (Blair et al. 1975). 

3.1.4. Acute Oral Toxicity 

As described in Section 3.1.2, DDVP exposure can result in increased cholinergic activity in the 

nervous system, producing the classical symptoms of organophosphate poisoning (See Table 

3-1). The life-threatening effects of acute exposure to DDVP are usually related to its 

cholinergic effects on the respiratory system (respiratory depression, bronchospasm, increased 

bronchial secretions, pulmonary edema, and muscle weakness). DDVP is moderately to highly 

3-3
 
toxic by the oral route when administered in single doses to a variety of animal species, and 

several cases of acute DDVP poisoning in humans have reported in the literature. Some 

individuals have committed suicide by intentionally ingesting DDVP pesticide formulations (e.g., 

Shimizu et al. 1996). This study is discussed further in Section 3.3 (Dose-Response 

Assessment). In an attempted suicide, a 56-year old woman who ingested about 100 mg/kg 

DDVP survived following intensive care for 14 days (WHO 1989). Two workers who had skin 

exposure to a concentrated dichlorvos formulation, and failed to wash it off, died of poisoning. 

In addition, four patients suffering from severe poisoning from oral exposure to dichlorvos 

survived, although they later showed delayed neurotoxic effects (WHO 1989). Thus, although 

the possibility of neuropathy in humans cannot be excluded, it is likely to occur only after almost 

lethal oral doses (see also Section 3.1.6). 

Oral LD50 values for experimental mammals range from 25 to 300 mg/kg (Jones et al. 1968, 

Gaines 1969, Muller 1970, Wagner and Johnson 1970). Signs of intoxication in these studies are 

consistent with cholinergic overstimulation, typically salivation, lacrimation, urination, 

defecation, tremors, convulsions, and death from respiratory failure. 

EPA (2000a, p. 18) identified an unpublished neurotoxicity study in rats as the basis for 

establishing a risk level for acute oral exposure to unformulated DDVP – i.e., DDVP not in a 

PVC strip. In this study (Bast et al. 1997), Sprague Dawley rats (12/sex/dose) received a single 

oral dose of DDVP (97.8%) at doses of 0, 0.5, 35, or 70 mg/kg. Behavioral testing, including a 

functional observation battery and motor activity, was conducted pretest, 15 minutes after 

treatment, and on days 7 and 14 after exposure. Cholinesterase activity was not measured in any 

tissue. The acute NOAEL was 0.5 mg/kg and the LOAEL was 35 mg/kg based on neurological 

effects related to AChE inhibition. 

The containment of DDVP in a slow-release vehicle, however, such as the PVC in the Vaportape 

II strips, will reduce the likelihood of acute toxic effects. The kinetics of DDVP release from 

PVC were investigated in a study in which DDVP was incorporated into PVC at 20% (w/w) 

(Slomka and Hine 1981). The PVC was extruded, cut into pellets, and encased in a hard gelatin 

capsule. The release of DDVP from the capsules was assayed in vitro using an artificial gastric 

fluid and in vivo in swine and humans. The release rates in the three assays were comparable; 

approximately 30% was released in the first 24 hours, and the subsequent release appeared to 

follow a first order function with a release rate of approximately 0.1 day -1. 

The effect of PVC encapsulation on the toxicity of DDVP has been quantified in parallel acute 

assays in young pigs (Stanton et al. 1979) using unformulated DDVP (undiluted technical grade 

administered in gelatin capsules) and DDVP in PVC resin (administered by gavage). For the 

technical grade liquid formulation, the LD50 was 157 (113–227) mg/kg. Signs of toxicity in these 

animals were consistent with the general signs of AChE inhibition (Table 3-1) and included 

decreased general activity, vomiting, poor coordination, and twitching. In the bioassay using the 

PVC formulation, no deaths occurred at any of the administered doses – i.e., 180 mg/kg, 240 

mg/kg, 320 mg/kg or 1,000 mg/kg. Higher doses of the DDVP-PVC formulation could not be 

3-4
 
administered because these doses produced vomiting. While not specified by Stanton et al. 

(1979), vomiting at doses >1,000 mg/kg may have been due to the physical stress associated with 

such a large gavage dose. Although no animals died, vomiting was observed at all DDVP-PVC 

doses. At the lowest dose, 180 mg/kg, vomiting with no other signs of AChE inhibition were 

observed. At the next higher dose, 240 mg/kg, no adverse effects are reported. 

Stanton et al. (1979) also conducted 30-day assays using only the PVC formulation. Aside from 

alterations in cholinesterase activity, 30 consecutive days of exposure of young swine or gravid 

sows to doses as high as 25 mg/kg-day of the DDVP-PVC formulation produced no adverse 

effect on any physical or biochemical parameter measured. The authors suggest that the lack of 

serious adverse effects was related to the slow-release of DDVP from the PVC pellet (Stanton et 

al. 1979). 

In an abstract, Singh et al. (1968) evaluated free DDVP (200 or 400 mg/day) or DDVP in V-13 

pellet (800 mg/day; 9% DDVP, 91% inert [NOS]) in gravid sows. The DDVP, whether in free 

form or in the pellet, produced no adverse effects on the number of pigs born alive, number of 

pigs born dead, average birth weight, average number of pigs weaned at 35 days, or the average 

weanling weight. Minor gross signs of organophosphate poisoning (NOS) were observed only in 

the group receiving 400 mg/day free DDVP. 

3.1.5. Subchronic or Chronic Systemic Toxic Effects 

Subchronic and chronic toxicity bioassays have been conducted in several laboratory animal 

species (e.g., rats, mice, dogs, pigs, and monkeys), exploring the adverse effects of DDVP 

exposure by oral and inhalation routes of exposure. Generally, the toxic effects of DDVP 

exposure (regardless of route of administration) are due to the inhibition of AChE (Table 3-1). 

Consequently, plasma, erythrocyte, and brain cholinesterase activity are metrics of exposure and 

toxicity. Studies have demonstrated more sensitive neurological effects than cholinesterase 

inhibition; however, the toxicologic implications of these early biomarkers of exposure are 

uncertain. For example, the correlations between the relatively low level, chronic dichlorvos 

(DDVP) exposure and early electrophysiological changes (assessed by electrocorticogram, 

cortical evoked potentials, conduction velocity, and refractory periods of peripheral nerve) 

showed the electrophysiological parameters to be sensitive biomarkers of the exposure in humans 

(Desi et al. 1998). 

In a long-term dietary study, rats fed diets containing DDVP for 2 years showed no signs of 

toxicity until the dietary exposures reached 2.5 mg/kg-day or more (WHO 1989). EPA (2000a) 

identified an unpublished dietary study in dogs (MRID No. 41593101 as summarized by U.S. 

EPA 1994) as the basis for establishing a risk level for chronic oral exposure. Groups of beagle 

dogs received DDVP orally in capsules at dose levels of 0, 0.1, 1.0, and 3.0 mg/kg/day for 52 

weeks. The 0.1 mg/kg/day dose was lowered to 0.05 mg/kg/day on day 22 due to the inhibition 

of plasma ChE noted after 12 days (the magnitude of the reduction was 21.1% in males and 

25.7% in females). After week 2, plasma ChE activity was only significantly reduced in males 

(39.1–59.2%) and females (41.0–56.7%) in the mid-dose group and in males (65.1–74.3%) and 

3-5
 
females (61.1–74.2%) in the high-dose group at all other later time intervals. RBC AChE 

activity was reduced in males (23.6%) and females (50.1%) at week 6 in the low-dose group. 

The authors attributed this to a residual effect on RBC AChE of the earlier dose of 0.1 

mg/kg/day, because much less inhibition was observed in this group after week 6. After week 6, 

RBC AChE activity was only significantly decreased in males (43.0–53.9%) and females 

(38.0–51.9%) in the mid-dose group and in males (81.2–86.9%) and females 79.2–82.5%) in the 

high-dose groups at all other later time intervals. Brain AChE activity was significantly reduced 

in males (22%) in the mid-dose group and in males (47%) and females (29%) in the high-dose 

group. The NOAEL and LOAEL selected by EPA (2000a) for chronic oral risk exposure are 

0.05 and 0.1 mg/kg/day, respectively. These effect levels are based on plasma ChE and RBC 

AChE inhibition in male and female dogs as early as the first time point measure and brain AChE 

inhibition in male dogs. 

3.1.6. Effects on Nervous System 

A neurotoxicant is a chemical that disrupts the function of nerves, either by interacting with 

nerves directly or by interacting with supporting cells in the nervous system (Durkin and 

Diamond 2002). This definition of neurotoxicant distinguishes agents that act directly on the 

nervous system (direct neurotoxicants) from those agents that might produce neurologic effects 

that are secondary to other forms of toxicity (indirect neurotoxicants). As discussed in Section 

3.1.2, DDVP, like all organophosphate insecticides, is a direct-acting neurotoxicant. DDVP 

combines with and inhibits AChE. The biochemical basis for the toxic effects of DDVP is 

related to the normal function of AChE. In the cholinergic system, neural impulses are 

transmitted between nerve cells or between nerve cells and an effector cell (such as a muscle cell) 

by the acetylcholine. When the acetylcholine reaches a certain level, the receptor cell is 

stimulated. Normally, the acetylcholine is then rapidly degraded to inactive agents (acetic acid 

and choline) by AChE. When AChE activity is inhibited by organophosphate agents (such as 

DDVP), acetylcholine persists and continues to accumulate at the synapse (the space between the 

nerve cells). Initially, this accumulation causes continuous stimulation of the cholinergic system, 

which may be followed by paralysis because of nerve cell fatigue (ATSDR 1993). 

The cholinergic effects of DDVP intoxication are well documented in studies involving humans, 

wildlife, and experimental mammals (Gillett et al. 1972a,b; IARC 1979, 1991; WHO 1989). 

DDVP also inhibits other cholinesterases and many other esterases outside of the nervous system 

and induces clinical signs of intoxication that are dependent upon the dose and duration of 

exposure (Table 3-1). In addition, some studies of lifetime exposure of rats to DDVP suggest 

that oral exposures to doses $0.97 mg/kg-day result in behavioral changes (Schultz et al. 1995, 

Institäoris et al. 1997). 

RBC AChE activity follows a circadian oscillation in both mice and humans (Jian and Zhiying 

1990). Furthermore, mortality in mice associated with exposure to DDVP is inversely related to 

the oscillation in AChE activity. These investigators report that DDVP interferes with the 

normal circadian rhythm of RBC AChE in mice and humans, although this interference is 

secondary to pronounced AChE inhibition. 

3-6
 
The effect of DDVP on AChE activity in humans has been assayed by Gledhill (1997). In this 

study, DDVP was administered to 6 male volunteers as a single dose of 70 mg DDVP in a corn 

oil solution in a gelatin capsule. The body weights of 6 individuals ranged from 67 kg to 80 kg 

(Gledhill 1997) and thus the individual dose rates ranged from 0.70 to 1.04 mg/kg bw. No effect 

on AChE activity was observed and there were no signs or symptoms of cholinergic 

overstimulation. 

Normal ChE activities can be highly variable among individuals. Consequently, interpreting 

differences between cholinesterase levels in exposed groups and control groups is more difficult 

than interpreting differences between individual ChE levels before and after exposure (ATSDR 

1993). All of the human and animal studies on PVC-DDVP formulations report AChE levels 

using the method involving treated groups and control groups. For all of the human studies on 

DDVP (Cervoni et al. 1969; Pena-Chavarria et al. 1969; Hine and Slomka 1970; Slomka and 

Hine 1981), the interpretation is further complicated because ChE levels are reported as ranges of 

inhibition, rather than mean values with standard errors. 

As discussed in the general literature and illustrated in the human studies on DDVP, inhibition of 

cholinesterase in plasma and blood is not necessarily associated with clinically significant 

adverse effects (Gage 1967; Wills 1972). ATSDR (1997) noted that the nervous system can 

accept a certain amount of acetylcholinesterase inhibition without overt toxic effects. In humans 

and animals, toxic signs are generally not seen until at least 20% of this enzyme (RBC AChE 

used as a marker) has been inhibited (ATSDR 1997). In a rat study, brain AChE after a 2-year 

inhalation exposure to DDVP was inhibited more than 90% compared to control animals (Blair 

et al. 1976), yet signs of cholinergic overstimulation were not observed. ATSDR (1997) suggests 

that the best predictor of toxicity is not necessarily the actual percentage inhibition of AChE, but 

rather how rapidly this inhibition has occurred. Rapid inhibition does not afford the nervous 

system time to adapt to AChE inhibition. This adaptation appears to involve desensitization and 

down regulation of muscarinic receptors (ATSDR 1997). 

A significant characteristic of some organophosphate insecticides is that the reversibility of 

enzyme inhibition is slow (Murphy 1980). Relatively little information is available on the 

reversibility of inhibition due to DDVP. There is one case report indicating substantial inhibition 

of ChE, 36% of normal, in an individual exposed to DDVP 3 days before the assay of ChE 

activity (Bisby and Simpson 1975), and other data suggest that cholinesterase activity levels do 

not return to normal for several months (ATSDR 1997). 

Exposure to some organophosphorus compounds cause delayed neuropathy in humans (also 

known as organophosphate-induced delayed neurotoxicity or OPIDN). Clinical manifestations 

include motor dysfunction, tingling in the extremities, and in some cases paralysis. These effects 

usually appear 7–14 days after exposure, when signs of cholinergic toxicity have resolved, and 

can persist for weeks or years (ATSDR 1997). The data concerning the potential for DDVPinduced OPIDN are inconsistent and controversial. Several studies that demonstrate that DDVP 

does not induce delayed neuropathy (WHO 1989), including a recent study in adult hens 

3-7
 
50 (Abdelsalam 1999). On the other hand, very high doses of DDVP (doses in excess of the LD ) 

produced clinical neuropathy when administered to hens (Johnson 1978, 1981). These data are 

consistent with human cases of poisoning where recovery was followed by delayed neurotoxicity 

(see Section 3.1.4) (WHO 1989). Subcutaneous doses of DDVP (single dose of 200 mg/kg or 6 

mg/kg-day for 8 weeks) in rats led to motor deficit or biochemical and behavioral deficits (Sarin 

and Gill 2000, 1998, respectively). The potential for OPIDN in humans resulting from exposure 

to DDVP in PVC resin strips is unknown. 

3.1.7. Effects on Immune System 

Immunotoxicants are chemical agents that disrupt the function of the immune system. Two 

general types of effects, suppression and enhancement, may be seen and both of these are 

generally regarded as adverse. Agents that impair immune responses (immune suppression) 

enhance susceptibility to infectious diseases or cancer. Enhancement or hyperreactivity can give 

rise to allergy or hypersensitivity, in which the immune system of genetically predisposed 

individuals inappropriately responds to chemical or biological agents (e.g., plant pollen, cat 

dander, flour gluten) that pose no threat to other individuals or autoimmunity, in which the 

immune system produces antibodies to self components leading to destruction of the organ or 

tissue involved (Durkin and Diamond 2002). 

Although the literature contains some evidence that organophosphate insecticides can impair 

immunological markers (Colosio et al. 1999), no human data are available to describe a doseresponse relationship for the immunotoxic potential of DDVP. Animal studies suggest that 

exposure to DDVP may be associated with immunosuppression. Treating rabbits with oral doses 

of 0.31–2.5 mg/kg DDVP (2.5–20% of the LD ) 5 days per week for 6 weeks resulted in 50

inhibition of both humoral and cell-mediated immune response to S. typhimurium (Desi et al. 

1978, 1980). Immunosuppression (suppressed IgM response at 48 hours) was also observed in 

mice treated with a single oral dose of 120 mg/kg DDVP (Casale et al. 1983). A decrease in 

relative spleen weight was also noted in this study; however, severe signs of DDVP neurotoxicity 

were noted and the authors stated that the immunosuppression observed in this study may have 

been related to toxic chemical stress. In addition, in vitro studies on cells from embryonic renal 

tissue of carp demonstrated a dose-related decrease in lymphocyte proliferation and myeloid cell 

respiratory burst activities, both of which indicate immunosuppression; however, no effects on 

antibody production were noted in an in vivo study of carp (Dunier et al. 1991). Bryant (1985) 

has associated the precipitation of preexisting asthma to small doses (NOS) of DDVP. 

Aside from the few positive reports above, there is very little direct information on which to 

assess the immunotoxic potential of DDVP in humans. The extrapolation of the observed 

alterations in the immune system response of experimental animals to humans is uncertain, since 

the functional relevance of these deficits in humans is unknown. The immune system has a 

functional reserve and modifications in the immune response do not always correlate with a 

measurable health effect (Vial et al. 1996; Voccia et al. 1999). 

3-8
 
The systemic toxicity of DDVP has been adequately examined in numerous acute, subchronic, 

and chronic bioassays. Although many of these studies did not focus on the immune system, 

changes in the immune system (which could potentially be manifest as increased susceptibility to 

infection among DDVP-exposed animals compared to controls) were not observed in any of the 

available long-term animal studies. In a three-generation study of Wistar rats, neurologic 

endpoints were found to be more sensitive markers of exposure than immunologic endpoints in 

all three generations (Institäoris et al. 1997). 

3.1.8. Effects on Endocrine System 

In terms of functional effects that have important public health implications, some of the effects 

on endocrine function would be expressed as diminished or abnormal reproductive performance. 

This issue is addressed specifically in the following section (Section 3.1.9). As discussed in 

Durkin and Diamond (2002), mechanistic assays are generally used to assess the potential for 

direct action on the endocrine system. DDVP has not been tested for activity as an agonist or 

antagonist of the major hormone systems (e.g., estrogen, androgen, thyroid hormone), nor have 

the levels of these circulating hormones been adequately characterized following DDVP 

exposures. Alterations in the diurnal rhythm of the pituitary/adrenal axis were observed in rats 

exposed to 2 ppm (approximately 0.3 mg/kg) DDVP in drinking water. Although effects on 

plasma ChE activity were not noted, levels of plasma adrenocorticotrophic hormones and adrenal 

cholesterol ester were altered (Civen et al. 1980). In the absence of mechanistic studies of the 

endocrine system, any judgments concerning the potential effect of DDVP on endocrine function 

must be based largely on inferences from standard toxicity studies, none of which provide 

evidence for an endocrine effect. 

3.1.9. Reproductive and Teratogenic Effects 

No data are available in humans concerning the potential for DDVP-induced reproductive or 

developmental toxicity. As a small, lipid-soluble molecule, DDVP would be expected to cross 

the placental barrier and be excreted into breast milk (Desi et al. 1998). According to some 

studies, exposure to DDVP caused reproductive and teratogenic effects in laboratory animals; on 

the other hand, there are several breeding studies in which no adverse reproductive or teratogenic 

effects were observed in rabbits or swine after exposure to DDVP (ATSDR 1997). In a study in 

which female rats were given intraperitoneal injections of 15 mg/kg DDVP on day 11 of 

gestation, herniation of the umbilical cord was observed in 3 of 41 offspring from the treated 

group (Kimbrough and Gaines 1969). The effect was not observed in offspring from the control 

group (0/65) but the effect is not statistically significant using the Fisher Exact test (p=0.074) – 

i.e., the conventional criterion for statistical significance is a p-value of # 0.05. In a threegeneration study of Wistar rats, oral gavage doses of approximately 1, 1.3, or 1.9 mg/kg-day 5 

days/week for 28 weeks found no consistent toxicity (systemic, reproductive, or immunologic) 

across generations (e.g., birth body weight was statistically decreased in generation 2 and 

increased in generation 3) (Institäoris et al. 1995, 1997). 

When rabbits were treated with 6 mg/kg DDVP during the last 10 days of gestation and the brain 

tissue of the offspring was examined by electron microscopy, there was an incidence of 

3-9
 
immaturity or delay in brain development that was not apparent in the offspring of the untreated 

rabbits (Dambska et al. 1979). The method of dosing the animals is not specified in this study. 

Groups of New Zealand White rabbits (16/dose) received DDVP (97% purity in distilled water) 

orally at dose levels of 0, 0.1, 2.5, or 7.0 mg/kg/day on gestation days 7 through 19 (U.S. EPA 

2000a, p. 19). The NOAEL for maternal toxicity was 0.1 mg/kg/day and the LOAEL was 2.5 

mg/kg/day, based on decreases in maternal body weight gain during gestation days 7–19. The 

U.S. EPA (2000a) considered the decrease in weight gain to be biologically significant even 

though the effect was not statistically significant. A dose-related increase in maternal mortality 

also was noted at 2.5 and 7 mg/kg/day. Cholinergic signs were observed at 7 mg/kg/day. No 

adverse developmental effects were noted in the fetuses. Cholinesterase activity was not 

determined. 

An early study by Schwetz et al. (1979) in New Zealand White rabbits and CF-1 mice using the 

MTD dose (based on signs of cholinesterase inhibition) for both oral (gavage of 5 mg/kg-day 

DDVP in corn oil on gestation days 5–18 and 60 mg/kg-day DDVP in corn oil on gestation days 

5–16 for rabbits and mice, respectively) and inhalation (whole body exposure to atmospheres 

3 containing 4 ìg/dL (0.4 mg/L or 400 mg/m ) DDVP for 7 hours/day on gestation days 5–18 or

5–16 for rabbits and mice, respectively) routes of exposure found no teratogenic effects that 

could be attributed to DDVP. These studies suggest that DDVP is not a selective developmental 

toxin, since adverse developmental effects only occur at doses that are maternally toxic. 

At toxic doses (i.e., where signs of organophosphorus poisoning are evident), DDVP may 

produce reversible adverse effects on spermatogenesis (WHO 1989). Adverse testicular effects 

were observed in mice after chronic exposure to average daily doses of 0, 58, or 94.8 mg/kg/day 

DDVP in drinking water (MRID 41041801 as cited by U.S. EPA 1994). There was a 

dose-related decrease in the absolute and relative weight of the testes, and testicular atrophy was 

increased at 94.8 mg/kg/day. In addition, sperm abnormalities were seen in C57BL/C3H mice 

injected intraperitoneally with 10 mg/kg/day for 5 days ( Wyrobek and Bruce 1975). About 6% 

of the sperm from DDVP-treated animals was abnormal compared to 1.8% of sperm from 

untreated animals. In a reproductive toxicity study involving male CF-1 mice, groups of 16 mice 

were exposed to atmospheres containing 0, 30, or 55 mg/m3 (0, 3.3, or 6.1 ppm, respectively) for 

16 hours or to 0, 2.1, or 5.8 mg/m3 23 hours/day for 4 weeks ( Dean and Thorpe 1972). No 

differences between control and treated mice were observed in the number of early fetal deaths, 

late fetal deaths, or live fetuses found in the pregnant females. The percentage of pregnancies for 

females mated to males exposed to DDVP was also similar to the controls (73–88%, mean 

80.9%). Under these exposure conditions, DDVP does not appear to affect the fertility of male 

CF-1 mice. No gross or histological evidence of treatment-related damage to reproductive 

tissues (prostate, testes, epididymis, ovaries, or uterus) was seen in F344 rats (4 or 8 mg/kg/day) 

or B6C3F1 mice (10, 20, or 40 mg/kg/day) orally exposed to DDVP by gavage for 2 years ( NTP 

1989). 

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3.1.10. Carcinogenicity and Mutagenicity 

Adequate data regarding the carcinogenic potential of DDVP in humans by any route of exposure 

are not available. Studies of human populations exposed to DDVP (including workplace and 

residential exposures) are constrained by the lack of adequate exposure data and other limiting 

issues. As reported in a series of case studies, some evidence suggests an association between 

childhood cancer and exposures to DDVP in resin strips during childhood or during gestation 

(Reeves et al. 1981, Davis et al. 1992, 1993, Liess and Savitz 1995). These studies have been 

reviewed by U.S. EPA (2000a) which concluded:

 “[r]eviews of these studies have identified biases and confounders that could 

explain the observed associations. The Agency concludes that the biases are a 

more likely explanation for the findings of increased cancer than exposure to 

resin strips. Additional studies that correct for the control of potential biases and 

problems of exposure determination are needed before an association between 

Dichlorvos and childhood cancer can be established” (U.S. EPA, 2000a, p. 26). 

The carcinogenic potential of DDVP has been evaluated in several animal species (mice, rats, 

dogs, and swine) via the oral route and in rats via the inhalation route. The weight of evidence 

suggests that the cancer bioassays do not offer sufficient evidence to treat DDVP as a potential 

human carcinogen (U.S. EPA 2000a,b). DDVP produced positive results in mammalian 

bioassays for carcinogenicity by the oral, but not the inhalation route of exposure. A cancer 

bioassay was conducted in which male and female mice were given gavage doses of DDVP (NCI 

1977). The doses levels were 10 and 20 mg/kg for males and 20 and 40 mg/kg for females. 

There was a significant dose-related increase in squamous-cell papillomas of the forestomach in 

both sexes. In females at the high-dose level, the incidence of squamous-cell carcinomas was 

significantly greater than in the control group (p=0.004 using the Fisher Exact test). In the same 

study, male rats were given 4 mg/kg/day DDVP by gavage and female rats were given 8 

mg/kg/day. A significant (p<0.001) dose-related increase in the incidence of acinar-cell 

adenomas of the pancreas was observed in the males. The increased incidence of fibroadenomas 

and adenomas of the mammary gland was significant (p=0.028) in the females. The increased 

incidence of the pancreatic acinar cell carcinomas in male rats and squamous cell tumors in male 

mice reported by NCI (1977) has been discounted by WHO (1989) and Mennear (1994, 1998). 

The relevance of the sex-specific increase in mononuclear cell carcinoma (MCL) reported by 

NCI (1977) has also been questioned ( Manley et al. 1997, Mennear 1998, U.S. EPA 2000b). 

The issues of concern regarding the increased incidence of MCL in male rats are not dose-related 

increases in mortality or disease severity (Mennear 1998), incidence rates among DDVP-treated 

rats statistically increased as compared to matched controls but within historical control 

incidence, and similarity in histopathology between the MCL tumors and spontaneous tumors in 

control animals (Manley et al. 1997). U.S. EPA (2000b) found compelling evidence to disregard 

the MCL finding in Fisher rats, concluding that “the high background and variability in the 

incidence of this tumor, as well as its species and strain specificity, make it an invalid response 

for human risk assessment”. Two other bioassays conducted on the carcinogenicity of DDVP 

3-11
 
after oral exposure are reviewed by IARC (1991). Neither study indicated significant evidence of 

carcinogenicity (IARC 1991). 

DDVP has been tested extensively for mutagenicity, and the results of the tests are available in 

several reviews (IARC 1979, 1991, Ramel et al. 1980, Mennear 1998, U.S. EPA 2000a,b). 

Mutagenic effects as well as covalent binding to RNA and DNA have been demonstrated in 

bacterial systems. Generally, mutagenicity is decreased by the presence of liver microsomal 

preparations; however, chromosome abnormalities in peripheral lymphocytes have been reported 

in pesticide workers who use DDVP (no quantitative exposure data are available and this appears 

to be from workers using a spray formulation of DDVP) (Desi et al. 1998). EPA (2000b) 

concluded that “the results from whole animal bioassays supercede the results in vitro tests... 

[C]ompounds that are positive in mutation tests but do not cause cancer in whole animals 

should be regulated as noncarcinogens”. 

A more detailed review of the cancer and mutagenicity literature database on DDVP is beyond 

the scope of this risk assessment. Owing to the extraordinary level of effort and Special Agency 

Reviews of the issue (U.S. EPA 2000a,b), this risk assessment will defer to the EPA’s latest 

position (U.S. EPA 2000a) concerning the carcinogenic and mutagenic potential of DDVP. In 

that assessment (U.S. EPA 2000a), which included an open meeting to discuss the issues (U.S. 

EPA 2000b), it was decided that “[t]he carcinogenicity potential of Dichlorvos has been 

classified as ‘suggestive’ under the 1999 Draft Agency Cancer Guidelines and no quantitative 

assessment of cancer risk is required”. Thus, this risk assessment for DDVP does not include a 

quantitative assessment of cancer risk. 

3.1.11. Irritation and Sensitization (Effects on the Skin and Eyes) 

The available human data, supported by studies on experimental animals, suggest that exposure 

to DDVP may cause skin irritation or allergic reactions. Human data regarding the dermal 

effects of DDVP are relatively sparse. In a case report, relatively severe contact dermatitis 

developed in an adult male after a 1% solution of DDVP leaked onto his skin (Bisby and 

Simpson 1975). This effect was accompanied by signs of cholinergic toxicity, including fatigue, 

dizziness, and labored respiration. Cases of dermatitis and skin sensitization due to DDVP have 

been described in workers handling and spraying different types of pesticides and 

cross-sensitization with certain pesticides has been seen (WHO 1989). 

The data from animal testing supports the results of human case reports. In New Zealand white 

rabbits, the application of an aqueous solution of 5–20% DDVP to the skin caused relatively 

severe irritation ( Arimatsu et al. 1977). In a skin sensitization assay, 1% DDVP in olive oil 

induced no visible effects in male albino guinea pigs ( Kodama 1968). In a guinea pig assay for 

allergenicity, 35% of the tested guinea pigs had a positive response to a 0.5% solution of DDVP ( 

Fujita 1985). In a sensitization assay, Ueda et al. (1994) reported that 1% DDVP was a 

threshold irritation concentration in guinea pigs and that cross-sensitization occurred between 

DDVP and triforine. WHO (1989) reported that in Hartley guinea pigs the primary irritant 

threshold limit value for DDVP was $2%. 

3-12
 
3.1.12. Systemic Toxic Effects from Dermal Exposure 

Most of the systemic effects observed after dermal exposure of laboratory animals (including 

monkeys, rats, and chickens) to DDVP were the result of the neurotoxicity of this chemical. In 

its risk assessment for DDVP, U.S. EPA (2000a) selected studies for short-term and 

intermediate-term risk assessment that reflect the systemic toxicity resulting from dermal 

exposures to DDVP. In both of these studies, the toxicity of DDVP is secondary to inhibition of 

cholinesterase activity. Data concerning the dermal absorption kinetics of DDVP are discussed 

in Section 3.1.3.2. 

A number of fatalities have been reported from dermal exposures to concentrated formulations of 

DDVP (spilling or splashing onto skin) (WHO 1989). The data suggests that, in those cases 

where the spilled solution was immediately washed off, the victims developed symptoms of 

organophosphorus poisoning but they recovered after treatment (WHO 1989). Such exposures 

are not relevant to this risk assessment, as the encapsulation of DDVP in PVC used in Vaportape 

II precludes rapid exposure to high doses of DDVP. 

3.1.13. Inhalation Exposure 

Exposure of pesticide manufacturing plant workers to concentrations in the air of up to 0.5 

mg/m3 were without clinical effects, and no, or only insignificant, inhibition of blood ChE 

activity was noted (WHO 1989). When DDVP is used properly, air levels of 0.01–0.03 ppm are 

achieved (ATSDR 1997). This level kills most insects within 1 hour; whereas, in human 

volunteers, exposure at about 20 times this level (0.23 ppm) for 2 hours a day for 4 days had no 

harmful effects (ATSDR 1997). Consistent with the human exposure data, harmful effects have 

not been seen in laboratory animals exposed to air levels of dichlorvos below 0.5 ppm (about 4.5 

3 mg/m ) (ATSDR 1997), and exposure of laboratory animals to DDVP air concentrations between

0.2–1 mg/m3 do not affect ChE activity significantly (WHO 1989). In a 2-year study in rats, 

breathing air every day containing low-to-moderately high levels (0.006–0.6 ppm or about 0.05 

to 5 mg/m3) of DDVP had no effect on survival or general health (ATSDR 1997). Generally, the 

systemic effects observed after inhalation exposure of laboratory animals to higher levels of 

DDVP were the result of the neurotoxicity (cholinesterase inhibition) (U.S. EPA 2000a). 

Chronic inhalation exposure of laboratory animals to DDVP produced no compound-related 

pulmonary toxicity (U.S. EPA 2000a). 

EPA (1994) selected the chronic inhalation study in rats (Blair et al. 1976) as the basis for 

establishing an RfC for DDVP. Groups of 50/sex/group Carworth E Farm (CFE) rats were 

exposed (whole body exposures) for 23 hours/day, 7 days/week to DDVP vapor (>97% purity) at 

atmospheric concentrations of 0, 0.05, 0.5, and 5 mg/m3 for 2 years. The rats were observed for 

clinical signs of toxicity, hematology, and clinical chemistry. Plasma, RBC, and brain 

cholinesterase activity were determined at study termination, but not prior to the study. No 

clinical signs of toxicity were observed, and no organ weight or organ to body weight changes or 

hematological changes were associated with DDVP exposure. Body weights were decreased as 

compared to control rats in high-dose male (up to 20% vs. control) and female rats (up to 14% 

vs. control) for large portions of the study. Dose-dependent reductions in plasma, RBC, and 

3-13
 
brain cholinesterase activity were observed. This study establishes a NOAEL of 0.05 mg/m3 and 

a LOAEL of 0.5 mg/m3 based on reductions in brain cholinesterase activity (U.S. EPA 2000a). 

3.1.14. Inerts and Adjuvants 

As discussed in Section 2.2, the DDVP used in gypsy moth control programs is contained in a 

multi-layered polyvinyl chloride (PVC) strip. The manufacturer (Hercon 2004) indicates that the 

product contains 10% DDVP, 0.75 % related compounds (Section 3.1.15), and 89.25% inert 

ingredients. The only toxicity data available on this strip itself (i.e., without DDVP) is an acute 

oral toxicity study in rats (Braun and Killeen 1975). This study used a DDVP-free strip ground 

to a “grayish-green powder”. The strip was tested at the limit dose of 5,000 mg/kg bw by gavage 

with a 14-day post-dosing observation period in 5 male and 5 female rats. No adverse effects 

were noted in any of the rats based on mortality, gross observations, body weight gain, and gross 

necropsy. While this single study has its limitations, it suggests that the PVC strip alone (i.e., 

without DDVP) is unlikely to produce acute adverse effects. Given the limited nature of the 

exposure scenarios assessed herein, these data may be sufficient information for the likely 

exposure scenario (i.e., a child putting a strip in his/her mouth). Section 3.1.17 focuses on the 

toxicity studies concerning DDVP embedded in the PVC strips. 

3.1.15. Impurities and Metabolites 

The product label Hercon (2004) specifies that, in addition to DDVP (10%), each strip contains 

0.75% compounds that are related to DDVP. Further details are not provided on the label; 

nonetheless, the impurities in commercial DDVP have been characterized (Gillett et al. 1972a; 

IARC 1991). The impurities include: Dipterex (O,O-dimethyl 2,2,2-trichloro-1­

hydroxyethylphosphonate); O,O-dimethyl 2-chlorovinyl phosphate; O,O-dimethyl 

methylphosphonate; O,O,O-trimethyl phosphate; and trichloroacetaldehyde. These impurities are 

known to be or are likely to be toxic (Gillett et al. 1972a, WHO 1989). These impurities are 

encompassed in the risk assessment because the effect levels are based on studies that used 

commercial grade DDVP. Consequently, the results of these studies are directly applicable to the 

risk assessment for human health. 

3.1.16. Toxicologic Interactions 

The major toxicologic interaction of concern is concurrent exposure to other cholinesterase 

inhibitors (e.g., organophosphate or carbamate insecticides) or cholinomimetic agents (e.g., 

agents such as pilocarpine or carbachol that mimic the action of acetylcholine). In either case, 

simultaneous exposure would likely enhance the cholinergic toxicity produced by DDVP. 

Potentiation studies using DDVP in combination with 22 other organophosphate pesticides, 

however, found little or no potentiation (WHO 1989). Chemicals that react with the serine 

residue at the active site of the “A”-type esterases (e.g., diisopropylfluorophosphate [DEP]) could 

also increase the toxicity of DDVP by interfering with its metabolism (ATSDR 1997). 

In addition, experimental data suggest that repeated exposures of rats to DDVP (5 mg/kg/day by 

intraperitoneal injection for 30 consecutive days) depletes brain glutathione levels ( Julka et al. 

1992). Reduced glutathione levels may decrease the rate of detoxification of DDVP by the 

3-14
 
glutathione-dependent metabolic pathways. The toxicologic significance of depleted brain 

glutathione on DDVP metabolism is not known. In contrast with the potentiation of DDVP 

toxicity observed when rats are pretreated with diethylmaleate ( Fukami 1980), Costa and 

Murphy (1984) reported that pretreatment with 600 mg/kg acetaminophen (which is also 

detoxified by and thus reduces glutathione levels) did not have any effect on the toxicity of 

DDVP. Although no data are available, these experiments suggest that repeat exposure to DDVP 

(resulting in a depletion of glutathione levels) may increase an organism’s susceptibility to 

toxicity by another chemical if that chemical is also detoxified by glutathione-dependent 

pathways. 

3.1.17. Studies on PVC Formulations of DDVP 

In the EPA risk assessment for DDVP (U.S. EPA 2000a), EPA noted that DDVP resin strips 

(such as the Vaportape II strip used in USDA programs) “account for a very small proportion of 

total incidences [e.g., reports of poisonings], about 33 cases per year (1% of total incidences). 

Incidence reports involving exposure to resin strips usually do not involve any significant acute 

symptoms that would require medical treatment”. In a review of DDVP-impregnated PVC strips 

(Gillett 1972a,b concluded that “even when chewed or applied directly to the skin for short 

intervals, the strips do not release excessive or hazardous amounts of DDVP”. 

When DDVP was administered orally to human volunteers (single or repeated doses of a 

slow-release PVC formulation), significant inhibition of RBC ChE activity was found only at 4 

mg/kg body weight or more (Hine and Slomka 1970; Slomka and Hine 1981). Single oral doses 

(1–32 mg/kg) of DDVP in a slow-release PVC formulation was administered to 107 male 

volunteers. Measurable reductions in erythrocyte ChE activity was observed at dose levels above 

4 mg/kg, with a maximum reduction of 46% at 32 mg/kg. Plasma ChE activity was affected at 

lower doses, with 50% reduction at 1 mg/kg and about 80% reduction at 6 mg/kg or more. 

Repeated oral doses of 1–16 mg/kg bw per day were given to 38 male volunteers for up to 3 

weeks. Plasma ChE activity was depressed at all dose levels, and RBC AChE activity depression 

was dose-related and statistically significant at doses of 2 mg/kg or more. Blood cell count, 

urine, liver function, prothrombin time, and blood urea nitrogen were all normal ( Hine and 

Slomka 1968, 1970, Slomka and Hine 1981, WHO 1989). Among these individuals, the clinical 

signs of DDVP exposure were minimal (nausea, diarrhea, lassitude, restlessness, and lightheadedness). 

Data from 32 rhesus monkeys receiving orally administered DDVP in PVC resin (as an 

anthelminthic) at 0, 5, 10, 20, 40, or 80 mg/kg once daily or 0, 8, or 20 mg/kg twice daily for 10 

to 21 days support the human data (Hass et al. 1971). None of the monkeys died or exhibited 

debilitating symptoms of organophosphorus poisoning, although some cholinergic effects were 

noted (a loss of appetite and emesis [LOAEL = 20 mg/kg]; diarrhea and salivation [LOAEL = 80 

mg/kg]). A semi-quantitative assay for cholinesterase activity demonstrated inhibition. Studies 

in swine treated with DDVP-impregnated pellets (veterinary use as anthelminthic) suggest that 

DDVP is absorbed from the pellets after oral exposure (Jacobs 1968, Potter et al. 1973). Neither 

study was reported in sufficient detail to develop dose-response relationships. 

3-15
 
Two reproduction studies investigated exposure to PVC-DDVP formulations. In one of the 

studies, swine were exposed to 5 or 25 mg/kg/day DDVP during the last 30 days of gestation 

(Stanton et al. 1979). Sows and fetuses were monitored for changes in ChE. Both plasma ChE 

and RBC AChE were inhibited in sows, and brain AChE was increased in fetuses. In a separate 

experiment conducted by these investigators, there were no significant effects on reproductive 

capacity in sows treated with 25 mg/kg/day DDVP during the last 30 days of gestation. In an 

abstract concerning DDVP encapsulated in PVC, Vogin (1971) reported that no adverse effects 

on reproduction or developmental parameters were observed in dams exposed to DDVP 

concentrations that did not cause maternal toxicity (up to 12 mg/kg). Maternal toxicity was 

evident in dams treated with 34 mg/kg. This abstract also employed exposures to PVC resin and 

dioctylphthalate to assess the potential developmental toxicity of inerts. No teratogenic effect 

was reported for any exposure regimen. 

When DDVP pesticide strips were used in hospital wards, exposure of hospitalized adults and 

children, as well as healthy pregnant women and newborn babies, did not produce any significant 

effects on plasma ChE or RBC AChE activity. Exposures were estimated TWA concentrations of 

0.05, 0.152, and 0.159 mg/m3 based on 18 hours/day (Vigliani 1971). Only those subjects 

exposed 24 hours/day to concentrations above 0.1 mg/m3 or patients with liver insufficiency 

showed a moderate decrease in plasma ChE activity (Cavagna et al. 1969). Cavagna et al. (1969) 

also calculated DDVP inhalation exposure doses (based on inhalation volumes of 10 m3/day for 

adults and 1.4 m3/day for children and continuous exposures) that would be required to produce a 

significant reduction in plasma ChE activity (25–54% reduction in activity) for healthy adults and 

children (approximately 0.03 mg/kg-day) and adults and children with liver insufficiency 

(approximately 0.006 mg/kg-day). Note that these exposure doses are not anticipated to produce 

signs or symptoms of cholinesterase inhibition (Cavagna et al. 1969). No significant effects on 

plasma ChE or RBC AChE activity were observed in people exposed to the recommended rate of 

one strip per 30 m3 in their homes over a period of 6 months, even when the strips were replaced 

at shorter intervals than that normally recommended (Zavon and Kindel 1966). The maximum 

average concentration in the air of the homes was approximately 0.1 mg/m3 (WHO 1989). In 

factory workers exposed to an average of 0.7 mg/m3 for 8 months, significant inhibition of 

plasma ChE and RBC AChE activity was found (WHO 1989). 

In a study evaluating the effects of 30 minutes of dermal exposure to a DDVP pest strip on AChE 

activity, no dermal effects were noted in 21 individuals (Zavon and Kindel 1966). Zavon and 

Kindel (1966) also reported no inhibition of plasma or erythrocyte cholinesterase from the 30 

minute dermal exposure as well as 5 consecutive days of 30 minutes of continuous dermal 

exposure to DDVP resin strips. EPA (1981) provides a summary or exposure incidents involving 

DDVP in the general public. The reports involving DDVP-impregnated resin strips involved 

dermal contact which largely resulted in DDVP-induced allergic reactions or contact dermatitis 

(this is consistent with the effects of DDVP reported in dermal contact bioassays as described in 

Section 3.1.12). Flea collar dermatitis (primary contact dermatitis) has been reported in dogs and 

cats wearing DDVP-impregnated PVC flea collars (Muller 1970), and four people who handled 

dogs wearing flea collars containing 9–10% DDVP developed contact dermatitis (patch tests 

3-16
 
using 0.25–1% DDVP in these individuals were positive). The data suggest that a very small 

proportion of the general population is susceptible to dermal irritation by DDVP (WHO 1989). 

3-17
 
3.2. EXPOSURE ASSESSMENT 

3.2.1. Overview. 

Under normal conditions, exposure to both workers and members of the general public should be 

negligible. Workers will handle strips only during the assembly of milk carton traps. If workers 

wear gloves and assemble the traps outdoors or in very well ventilated rooms, both inhalation 

and dermal exposures should be negligible. Inhalation exposure to DDVP during transport of the 

traps should also be negligible if the traps are not transported inside of the passenger 

compartments of vehicles. Worker exposures will also be limited in most programs because foil 

wrapping in which the strip is distributed will not be removed until after the trap is transported to 

the field. Milk carton traps will generally be placed about four feet above the ground and 

exposure of members of the general public to DDVP contained in the milk carton traps should 

also be negligible except in the case of intensional tampering. 

Notwithstanding the above assertions, exposure assessments are developed for workers who do 

not use gloves in the assembly of the milk carton traps and who assemble the traps indoors, 

remove the protective foil strip during assembly, and transport the traps in the passenger 

compartments of vehicles. All of these exposure scenarios should be considered atypical and 

some are extreme. The intent is to illustrate the consequences of mishandling or imprudent 

handling. During assembly, the central estimate of dermal exposures in workers not wearing 

gloves leads to an absorbed dose of about 0.0008 mg/kg with a range of about 0.0003 mg/kg to 

0.004 mg/kg. Inhalation exposures in workers may be highly variable depending on the 

ventilation rates in an enclosed space and the number of traps that are handled. Based on the 

3 handling and transport of 75 traps, inhalation exposures could reach up to about 0.6 mg/m in an

3 enclosed and unventilated room and up to about 1.8 mg/m in the passenger compartment of a

vehicle. These exposure assessments are based on several site and situation specific assumptions 

which are intended to reflect plausible upper bounds of exposures. 

Exposure assessments are also developed for children who might come in contact with an 

accidentally discarded or misplaced strip. Estimated dermal doses are much higher than those for 

workers: a central estimate of about 0.02 mg/kg with a range of 0.003 mg/kg to 0.1 mg/kg. Oral 

exposures from a small child sucking on the pest strip are about a factor of 10 higher than dermal 

exposures: a central value of about 0.2 mg/kg with a range of 0.04 mg/kg to 0.6 mg/kg. 

Under normal circumstances, the use of DDVP in PVC strips is not likely to result in 

contamination of water or other materials that might be consumed by members of the general 

public. Nonetheless, an exposure assessment is developed for the accidental contamination of a 

small pond by a pest strip. In this scenario, dose estimates range from about 0.000003 mg/kg to 

0.00007 mg/kg with a central estimate of about 0.00001 mg/kg. 

3.2.2. Workers 

3.2.2.1. General Considerations – The EPA (2000a) concluded that human exposures would 

be negligible from DDVP-impregnated strips in insect traps (such as those used in USDA 

programs). Consequently, the EPA (2000a) did not quantitatively assess the exposure or 

3-18
 
potential risks posed by the use of PVC formulations of DDVP for any route of exposure. While 

this may be a reasonable approach, the current risk assessment develops quantitative exposure 

assessments for both workers and the general public that could occur in cases of poor handling 

practices. 

The milk carton traps can be assembled in two stages. The most time consuming stage is the 

carton assembly, in which two pre-cut perforated pieces of heavy waxed paper, similar to those 

used in milk cartons, are configured. In the second stage, the DDVP strip and disparlure wick are 

attached to the twist tie, and the twist tie is placed in the trap. The second stage should proceed 

much more rapidly than the first. During assembly, two routes of exposure may be significant, 

inhalation and dermal. As discussed in the program description (Section 2.2), however, both 

routes of exposure will be negligible if proper handling procedures are followed (that is, if the 

strips are installed outdoors or in a well ventilated area, if foil wrapping in which the strip is 

distributed is removed until after the trap is transported, and dermal contact with the strip is 

avoided). 

3.2.2.2. Inhalation Exposures – During normal use and assembly, either outdoors or in well 

ventilated areas, inhalation exposures to DDVP should be negligible. The material safety data 

sheet for VaporTape II (Hercon 1993) calls for local exhaust and respirators under conditions of 

continuous handling. Estimates of concentrations of DDVP in air from release of DDVP by 

VaporTape strips under different conditions of ventilation can be based on estimates of release 

rates (Hercon 1994) and a more general air model for DDVP pest strips proposed by Gillett et al. 

(1972a). 

Hercon (1994) conducted a study on the release of DDVP from Vaportape II strips. In this study, 

two samples (referred to as A and B) were weighed and assayed for DDVP at various intervals 

for up to 12 weeks after placement outdoors. The results, expressed as the proportion of DDVP 

remaining in the strip at various intervals, are detailed in Worksheet A01. As also detailed in 

Worksheet A01, the release data fit a first order model extremely well with an adjusted squared 

correlation coefficient of 0.97 and a p-value of 2×10 -23 . The estimated first-order release 

-1 -1 coefficient is 0.04 day with very narrow confidence intervals – i.e., 0.037 to 0.043 day . 

Gillett et al. (1972a) proposed the following model for estimating concentrations of DDVP in air 

from the release of DDVP from pest strips: 

(Eq. 3-1) 

3-19
 
The terms in the above equation are defined as follows: 

t time after start of release 

C t concentration of DDVP in air at time, t (days) 

M0 mass of DDVP in strip or strips at time zero (mg) 

Va volume of room or other space (m )3 

ã apparent adsorption coefficient of DDVP on to surfaces 

exp(x) the exponential function, e , where e is the constant 2.718 and x is x 

any numeric expression 

ë first-order release rate constant (days )-1 

RH relative humidity (proportion) 

A t air flow rate (m /day) 3 

k first-order vapor phase hydrolysis rate (days )-1 

The parameters used in the model are summarized in Table 3-2. The fit of the Gillett et al. 

(1972a) model to the data from Slomka (1970) using the apparent adsorption coefficient (ã) of 

37.5 is illustrated in Figure 3-1 (which is in turn taken from Worksheet A02b). Technical details 

of the application of the model and optimization of the model parameter for adsorption (ã) are 

given in Appendix 1. 

For the current risk assessment, two scenarios are considered for inhalation exposures of workers 

to DDVP: assembly of traps with strips in a garage and driving in a vehicle containing assembled 

traps with the strips. Both scenarios assume that the worker has removed the protective foil from 

the strip during assembly of the trap. These exposure scenarios are detailed in Worksheets A03a 

(garage) and A03b (vehicle). It should be noted that these exposure assessments are based on a 

number of plausible but conservative exposure assumptions – i.e., number of traps assembled or 

transported, volume of the space in which the traps are assembled or transported, and the 

ventilation rates of these spaces. The worksheets in which these exposure assessments are given 

are designed so that these parameters may be varied and applied to specific uses of the DDVP 

strips in specific USDA programs. 

A major factor in exposure will be the number of traps that are assembled. In the previous risk 

assessment (USDA 1995a), it was assumed that a workers would assemble up to 75 traps at a 

time. No more recent information has been encountered on the number of traps that might be 

assembled by a worker or workers and the value of 75 traps is maintained in the current risk 

assessment. 

For exposures in a garage involving the assembly of the milk carton traps, the dimensions of the 

3 3 3 3 garage are assumed to be 1,500 ft (10 feet @ 10 feet @ 15 feet) or 42.48 m [1 ft =0.02832 m ]. 

For the exposure assessment involving transport of the strips in a vehicle, the volume of the 

3-20
 
3 3 driving cabin is assumed to be 160 ft (8 feet @ 5 feet @ 4 feet) or 4.5 m . Again, these assumptions 

are somewhat arbitrary but are identical to the assumption used in the previous risk assessment 

(USDA 1995a). 

The other major assumptions used in these exposure scenarios involve ventilation rates and 

-1 release rates. The release rate is taken as 0.04 day from the study by Hercon (1994) discussed

above and detailed in Worksheet A01. It should be noted that the study by Hercon (1994) was 

conducted outdoors over a period of 12 weeks. Hercon (1994) does not specify the average 

temperature or range of temperatures. As discussed in Gillett et al. (1972a), the release rate of 

DDVP from PVC test strips will increase with increasing temperature, doubling from a 

temperature of 25EC to 38EC. This variability is not explicitly incorporated into the model used 

in this risk assessment (Eq. 3-1) and release rates higher than 0.04 day-1 are possible at high 

ambient temperatures. 

Ventilation rates are likely to be highly variable. In most cases, it is likely that the milk carton 

traps will be assembled outdoors and will be transported in a cargo area and not in the driving 

cabin. In such cases, inhalation exposure would likely be negligible. For the purpose of 

illustrating the consequence of assembling traps in a garage or similar structure or transporting 

assembled traps in a vehicle, three ventilation rates (number of air turnovers per day) are used for 

-1 -1 each scenario. Rates of 0 day (no ventilation) and 60 day (poor ventilation) are used in both 

scenarios. An additional rate of 300 day-1 is used in the garage scenario and an additional rate of 

-1 3000 day is used in the vehicle scenario. These rates are referred to as “Adequate” in 

Worksheets A03a and A03b. As discussed further in Section 3.4.2, this term is used because 

3 these ventilation rates lead to concentrations in air that are about 0.1 mg/m , the chronic NOAEL

from animal studies and the TLV recommended by ACGIH (2004). 

As detailed in Worksheet A03a, the garage scenario models concentrations over a 24 hour 

period. This duration period is selected under the assumption that traps might be stored for a day 

3 prior to use. The modeled concentrations reach up to about 0.5 mg/m for no ventilation and 0.3

3 3 mg/m for poor ventilation. As noted above, peak concentrations of 0.1 mg/m are obtained with 

a ventilation rate of about 300 day -1. The vehicle scenario (Worksheet A03b) covers a period of 

only 6 hours. It is likely that the duration of transport would typically be much less. Peak 

3 3 concentrations are somewhat higher – 1.8 mg/m for no ventilation and about 1.5 mg/m for poor

ventilation. It is unclear if the no ventilation or poor ventilation assumptions are reasonable for a 

vehicle. As discussed by Fedoruk and Kerger (2003), concentrations of volatile organic 

compounds in vehicles suggest that substantial air turnover rates are likely in vehicles even when 

the ventilation system is turned off and the windows are closed. Quantitative estimates of air 

turnover rates in vehicle passenger cabins, however, have not been encountered. Nonetheless, it 

-1 -1 seems that turnover rates of 0 day or 60 day will lead to overestimates of concentrations of 

DDVP in the air of passenger compartments. Adequate ventilation for a vehicle is defined as a 

-1 3 turnover rate of 3000 day , the rate required to reach a concentration in air of about 0.1 mg/m .

3-21
 
3.2.2.3. Dermal Exposures – For assessing the likelihood of systemic toxic effects from dermal 

exposures, such as handling a pest strip during assembly, some estimate of absorbed dose is 

necessary. The method for making such an assessment for DDVP test strips, however, is highly 

uncertain. 

As an individual manipulates the strip, some material will be transferred to the surface of the 

skin. Some of the chemical will be absorbed and some will volatilize. Assuming that the nature 

of the manipulation is such that a film of DDVP is maintained on the contaminated surface, 

Fick's first law may be used to estimate absorption (U.S. EPA/ORD 1992). Fick's first law 

requires an estimation of the K in cm per hour, the concentration of the chemical in a solution in p

contact with the skin, the area of the body surface that is contaminated, and the duration of 

exposure. There is no experimentally determined K for DDVP. Based on structure-activity p 

relationships proposed by the U.S. EPA/ORD (1992), K for DDVP is may be estimated at about p

0.00090 cm/hour with a 95% confidence interval of 0.00061 cm/hr to 0.0013. Details of these 

calculations are given in Appendix 2. 

In this and other similar scenarios considered in this risk assessment, the DDVP is not in 

solution; instead, the skin is in contact with neat or undiluted DDVP. Following the 

recommendations of U.S. EPA/ORD (1992), the functional concentration of DDVP on the 

surface of the skin is assumed to be the solubility of DDVP in water, 10 mg/mL (Table 2-2) – 

i.e., the concentration of DDVP in pore water of the skin will be limited by the solubility of the 

chemical in water. 

For workers wearing gloves, dermal absorption will be negligible. For workers who do not wear 

gloves, it is possible that the tips of the fingers and perhaps other surfaces on the hands would be 

contaminated. The most likely surface for contamination would be the finger tips. The precise 

area that might be contaminated, however, is difficult to estimate. The finger tip of each digit 

2 2 will be taken as 1 cm , except for the thumb that will be taken as 2 cm . Thus, the total surface 

area of the finger tips of both hands will be taken as 12 cm . 2 This value will be used to calculate 

both lower and central estimates of absorbed dose. To account for the potential contamination of 

other parts of the hand, the upper range of exposed surface area will be taken as 24 cm . 2 The 

duration of exposure is difficult to estimate. Most of the time spent in assembling the milk 

carton trap will not involve the DDVP strip. For this exposure assessment, a central estimate of 

0.5 hours of total contact time with the strip is used and the range is taken as 0.25 hours to 1 

hour. As detailed in Worksheet B01a, the assumptions used in this exposure scenario lead to 

estimates of absorbed dose of about 0.0008 mg/kg with a range of about 0.0003 mg/kg to 0.004 

mg/kg. 

3.2.3. General Public 

3.2.3.1. General Considerations – Milk carton traps contain the strip of Vaportape II attached to 

a twist tie or simply placed in the bottom of the trap. The DDVP strip can be accessed easily and 

removed. As summarized by U.S. EPA (2000a, p. 26), incidents involving contact with DDVP 

resin strips have been reported but these incidents account for only a small proportion of the total 

3-22
 
incidents involving DDVP (1% or about 33 cases per year) and the reported incidents involving 

DDVP strips typically to do not lead to overt signs of toxicity that require medical treatment. 

In the current risk assessment, two routes of exposure are considered for the general public: 

dermal contact and ingestion. Milk carton traps will generally be placed about four feet above 

the ground (Leonard 2004) and exposure of members of the general public to DDVP contained in 

the milk carton traps should also be negligible except in the case of intensional tampering. 

Although any member of the general public could tamper with a trap, incidents such as these 

seem to be more plausible for children, compared with adults. While the traps may be place out 

of the reach of young children, the potential for exposure to the DDVP strip could occur if traps 

were accidentally dislodged or misplaced. In addition, using children as the exposed group is 

conservative because dose estimates for children, in units of mg/kg body weight, will be higher 

than those for adults. 

3.2.3.2. Dermal Contact – The exposure assessment for dermal contact with a VaporTape II 

strip is detailed in Worksheet B01b. This scenario is very similar to that for dermal contact in a 

worker (Worksheet B01a). The major differences involve body weight, the dermal surface area 

that is considered, and the duration of exposure. The body weight is taken as 13.3 kg, the 

standard value for a 2-3 year old child (U.S. EPA/ORD 1996). In this scenario, it is assumed that 

a young child comes in contact with a pest strip and holds the strip against the surface of the skin 

for a period of time. Thus, the exposed skin surface area is taken as the dimensions of the strip – 

2 i.e., 1" x 4" inches or about 26 cm ). The duration of exposure must be set somewhat arbitrarily. 

It does not seem reasonable to assume that a 2-3 year old child would be unsupervised for a 

prolonged period of time. Consistent with the approach taken in the 1995 risk assessments 

(USDA 1995a), the central estimate of exposure will be taken as 1 hour with an upper range of 

4 hours. In the current risk assessment, a lower range of 15 minutes (0.25 hours) is also used and 

may be a more reasonable estimate of a plausible duration of exposure. Other assumptions and 

calculations are identical to those in the corresponding worker exposure assessment (Worksheet 

B01a, Section 3.2.2.3). As indicated in Worksheet B01b, this exposure assessment for a young 

child handling a DDVP strip leads to an estimated dose of about 0.02 mg/kg with a range of 

0.003 mg/kg to 0.1 mg/kg. 

3.2.3.3. Oral Exposure to DDVP Strip – As with dermal exposure, it is unlikely that children 

would experience any oral exposure to DDVP strips. The strips are placed within the milk carton 

traps and 2-3 year old children will generally be closely supervised. Thus, this exposure 

assessment for oral exposure, as with the above scenario for dermal exposure, should be regarded 

as accidental. 

An assessment of oral exposure might be based on incidental sucking on a pest strip. The 

amount of DDVP that a child might absorb will depend on the proportion of the strip that is in 

the mouth, the release rate of DDVP from the strip, and duration of the activity. The durations 

will be taken as the same as in the dermal exposure scenario, a central estimate of 1 hour with a 

range of 0.25 to 4 hours. The initial release rate will be taken as 0.015 hour -1. This is calculated 

3-23
 
from the study by Slomka and Hine (1981) which indicated that approximately 30% of the 

DDVP was released in the first 24 hours – i.e., k = -ln(1-f)/t = ln(1-0.3)/24 hours = 0.01486 

-1 hour .]. The proportion of the strip that might be in the mouth of the child will be taken as 0.25 

– i.e., a area of about 1 square inch. As indicated in Worksheet B02, this exposure assessment 

results in estimates of absorbed doses of about 0.2 mg/kg with a range of 0.04 mg/kg to 0.6 

mg/kg. This scenario would also involve some dermal exposure. As indicated in Section 3.4, 

any plausible dermal exposure would likely be much less than the oral exposure and would have 

no impact on the characterization of risk. 

3.2.3.4. Oral Exposure to Contaminated Water – Under normal circumstances, the use of 

DDVP in PVC strips is not likely to result in contamination of water or other materials that might 

be consumed by members of the general public. In the recent risk assessment by U.S. EPA 

(2000a), no exposure assessment for water contamination by DDVP in PVC formulations is 

presented. 

The approach taken by U.S. EPA (2000a) seems reasonable in that the slow release DDVP from 

the test strip and rapid hydrolysis of DDVP in water is likely to limit the concentration of DDVP 

in ambient water. For example, the halftimes for the hydrolysis of DDVP in water range from 

about 11.65 days at pH 5 to 0.88 days at pH 9, with a hydrolysis halftime of 5.19 days at pH 7 

(U.S. EPA 1999a, p. 3). These values correspond to hydrolysis rates – i.e., k = ln(2)/t50 – of 0.06 

-1 -1 -1 day [pH 5], 0.13 day [pH 7], and 0.78 day [pH 9]. All of these hydrolysis rates are more 

rapid than the release rate of DDVP in air from the Hercon pest strip – i.e., 0.04 day-1 as 

discussed in Section 3.2.2.2. 

For this risk assessment, the assumption will be made that a VaporTape strip accidentally 

contaminates a small pond (e.g., it is inadvertently dropped into a pond during placement of a 

trap or a trap is dislodged and falls or is blown into a pond). No data are available to directly 

estimate the amount of DDVP that might be released over the course of a single day. For this 

exposure assessment, the assumption will be made that 30% of the DDVP in a fresh strip might 

be released over the course of a single day. This is based on the study by Slomka and Hine 

(1981), discussed in Section 3.1.4, in which 30% of the DDVP was released from a pest strip into 

gastric juices over a 24 hour period. Thus, the central estimate of the amount of DDVP in water 

is taken as 177 mg [590 mg × 0.3]. The upper range of the amount of DDVP in water is taken 

simply as the amount of DDVP in a new pest strip – 590 mg. The selection of a lower is 

somewhat arbitrary and a value of 10% or 59 mg is used. Other details of this exposure 

assessment are given in Worksheet B03 and involve standard assumptions concerning the size of 

the pond and the amount of water that might be consumed. These assumptions are standard in 

risk assessments (SERA 2001). As detailed in Worksheet B02, dose estimates range from about 

0.000003 mg/kg to 0.00007 mg/kg with a central estimate of about 0.00001 mg/kg. 

As noted above, this very simple exposure scenario does not consider the degradation or 

dissipation of DDVP. As discussed further in Section 3.4, however, this exposure assessment 

leads to concentrations in water that are far below a level of concern. Thus, the overestimates of 

3-24
 
concentrations in water developed in this section have no impact on the risk characterization for 

potential effects in humans. 

3-25
 
3.3. DOSE-RESPONSE ASSESSMENT 

3.3.1. Overview 

The extensive toxicology data base has been evaluated by a number of governmental 

organizations including the U.S. EPA, the Agency for Toxic Substances and Disease Registry 

(ATSDR), National Institute for Occupational Safety and Health (NIOSH), Occupational Safety 

and Health Administration, and the World Health Organization. Following the approach taken in 

most USDA risk assessments, these sources are used for selecting levels of acceptable exposure. 

Because all of the scenarios considered in this risk assessment involve only acute exposures, only 

acute exposure criteria are considered. 

For both oral and dermal exposures, the acute RfD established by the U.S. EPA, 0.0017 mg/kg, is 

used for the risk characterization. This is based on an acute oral NOAEL of 0.5 mg/kg from a 

study in rats with the application of an uncertainty factor of 300. Acute exposure criteria 

proposed by other groups are comparable to but somewhat higher than the acute RfD. Because 

some of the accidental acute exposures may substantially exceed the acute RfD, some attempt is 

made to characterize the consequences of high oral exposures. A human NOAEL of 1 mg/kg for 

AChE inhibition has been identified. While this NOAEL is not used to modify the acute RfD, it 

can be used to assess plausible consequences of exceeding the RfD. The human data on DDVP, 

although extensive, are not sufficient to identify a minimal lethal dose. For the current risk 

assessment, the lowest reported lethal dose (16 mg/kg) is used to assess the plausibility of 

observing serious adverse effects in cases of accidental over-exposure to DDVP. 

A number of inhalation criteria for DDVP are available. Since potentially significant inhalation 

3 exposures are likely only in workers, the occupational exposure criterion of 0.1 mg/m proposed

by American Conference of Governmental Industrial Hygienists is used. This value is a factor of 

10 below the occupational criteria proposed by NIOSH and OSHA. 

3.3.2. Acute Exposures 

3.3.2.1. Acute Oral – As summarized in Section 3.1.4, the U.S. EPA (2000a) bases the acute 

oral RfD for DDVP on the study by Bast et al. (1997) in which no effects, including assays for 

alterations in behavior, were noted at 0.5 mg/kg but neurological effects related to AChE 

inhibition were noted at 35 mg/kg. In deriving the acute RfD, the U.S. EPA (2000a, p. 18) used 

an uncertainty factor of 300 and recommended an acute RfD of 0.0017 mg/kg/day [0.5 mg/kg ÷ 

300 = 0.0017 mg/kg]. ATSDR (1997) has recommended a somewhat higher acute oral minimal 

risk level (MRL) – a value that is analogous to the RfD – of 0.004 mg/kg/day. This is based on a 

14-day LOAEL of 4 mg/kg/day in which brain AChE was inhibited by 44%. The MRL was 

calculated using an uncertainty factor of 1000 (ATSDR 1997, pp. 83-84). 

As also discussed in Section 3.1.4, the study by Stanton et al. (1979) suggests that DDVP in a 

PVC formulation will be much less toxic than unformulated DDVP. The extent of the difference 

in toxicity, however, is difficult to quantify. For unformulated DDVP, the LD50 value was 157 

(113–227) mg/kg with no mortality observed at 56 mg/kg. For the DDVP-PVC formulation, no 

deaths occurred at doses of up to 1000 mg/kg, although signs of toxicity consistent with AChE 

3-26
 
inhibition were observed at doses of 320 mg/kg and 1000 mg/kg using the DDVP-PVC 

formulation. No tremors or salivation were observed at doses of 240 or 180 mg/kg of the DDVPPVC formulation. Stanton et al. (1979) do not provide comparative data the extent of AChE 

inhibition in unformulated DDVP and the DDVP-PVC formulation. 

As detailed in Section 3.2.3.3, estimates of acute oral exposure for a small child sucking on a 

pest strip are far above the acute RfD of 0.0017 mg/kg. Thus, the potential for more severe 

effects must be considered. Based on the recent study by Gledhill (1997), no changes in AChE 

activity and no signs of toxicity were seen in a group of 6 men administered DDVP in a gelatin 

capsule at an approximate dose of 1 mg/kg. This is a factor of about 600 above the acute oral 

RfD. This study is unpublished and was submitted to the U.S. EPA by a registrant. In the U.S. 

EPA (2000a) human health risk assessment, the MRID number for this study is cited but the 

results of the study are not discussed specifically. For the current risk assessment, a dose of 1 

mg/kg from the Gledhill (1997) study is used qualitatively to characterize the risks of exposures 

that are not likely to produce clinically significant effects. 

For many pesticides, exposures that would be associated with severe and possibly fatal effects 

often can be estimated from poisoning reports. Most reports of fatal exposures to DDVP, 

however, do not provide sufficient information to estimate a lethal dose in humans. An 

approximate lethal dose, however, can be estimated from the study by Shimizu et al. (1996), 

which reports a fatal exposure of a 62.5 kg woman who intentionally consumed a pesticide 

formulation containing 75% DDVP and 25% xylene. While xylene is also a toxic agent, the oral 

LD50 for xylene in rodents is in the range of 3,500 to 8,600 mg/kg (ATSDR, 1995, p. 59). This is 

much greater than the reported LD50 values for DDVP in rodents – i.e., in the range of 25 to 300 

mg/kg as summarized in Section 3.14. The amount of DDVP that the woman ingested is unclear. 

About 300 grams (300,000 mg) of DDVP were found in the stomach and Shimizu et al. (1996, p. 

65) estimate that the woman probably absorbed about 1,000 mg/kg. Taking the estimated 

absorbed dose, a lethal dose for humans can be estimated at about 16 mg/kg [1,000 mg ÷ 62.5 

kg]. This is not necessarily a minimum lethal dose – i.e., the individual might have died after 

ingesting a lesser amount of DDVP. Other reported poisoning cases involving DDVP (e.g., 

ATSDR 1997; WHO 1988) do not have sufficient information to estimate a minimum lethal dose 

for humans. 

3.3.2.2. Acute Dermal – For short-term dermal exposure, the U.S. EPA (2000a) recommends an 

oral NOAEL of 0.1 mg/kg with a margin of exposure of 300 for residential exposure and 100 for 

occupational exposure. This would correspond to an acute RfD of 0.00033 mg/kg for residential 

exposures and 0.001 mg/kg for occupational exposures. The U.S. EPA (2000a) recommends 

using this value with dermal deposition data and an assumed dermal absorption fraction of 11%. 

These values will not be used in the current risk assessment. Following the general approach 

used in other risk assessments prepared for USDA (SERA 2001), the absorbed doses estimated in 

Section 3.2.2.3 for workers and Section 3.2.3.2 for the general public will be used with the acute 

oral RfD of 0.0017 mg/kg/day. The general rationale for this approach is given in SERA (2001). 

3-27
 
For DDVP in particular, the standard approach used in USDA risk assessments is necessary 

because the incidental or accidental handling of VaporTape strips does lead to estimates of 

dermal deposition. 

3.3.2.3. Acute Inhalation – For short-term inhalation exposures, the U.S. EPA (2000a) 

recommends the same acute toxicity value used for dermal exposures. Given the extensive 

inhalation toxicity data available for DDVP, the rationale for this approach is unclear. The U.S. 

EPA (1994) has derived an inhalation RfC for DDVP of 0.0005 mg/m . 3 This is based on an 

3 3 animal NOAEL of 0.05 mg/m with a corresponding LOAEL of 0.48 mg/m from a two year

exposure study in rats. As noted below, this chronic RfD is not relevant to the current risk 

assessment because no chronic exposures are anticipated. In addition to this value recommended 

by EPA, ATSDR (1997) has recommended an acute minimum risk level (MRL) of 0.002 ppm for 

3 3 DDVP which corresponds to a concentration of about 0.018 mg/m – i.e., 1 ppm = 9.04 mg/m . 

This value is intended to be applied to exposure periods of up to 14 days. 

As detailed in Section 3.2.2.2, all exposures for workers are short-term. OSHA and NIOSH 

share responsibility for proposing exposure criteria to protect workers. OSHA provides 

regulatory enforcement (exposure standards) and NIOSH provides science based exposure 

criteria (NIOSH 2002). For DDVP, NIOSH recommends a time-weighted average exposure 

3 limit of 1 mg/m and this value has been adopted by OSHA (NIOSH 2002). Another group 

involved in recommending criteria for occupational exposure is ACGIH (2004), which 

3 recommended a lower occupational exposure limit of 0.1 mg/m (ACGIH 1991). This lower 

value appears to have been selected by ACGIH (1991) based on an unpublished report to the 

3 TLV committee that exposures to 1 mg/m over the course of a workday resulted in an inhibition

of plasma AChE of 20%-25% in a group of workers (ACGIH 1991, p. 446). The documentation 

for the TLV, however, does not suggest that any adverse health effects were observed. The lower 

3 and more protective value of 0.1 mg/m is adopted in the current risk assessment for the

protection of workers during inhalation exposures. 

3.3.3. Chronic Exposures 

The U.S. EPA (2002), ATSDR (1997), and WHO (1998) have all recommended various criteria 

for chronic exposure to DDVP by oral, dermal, and/or inhalation routes. Because none of the 

exposure scenarios in this risk assessment involve chronic or subchronic exposures, these 

recommendations are not considered in the current risk assessment. While the previous USDA 

risk assessment (USDA 1995a) considered the potential cancer risks associated with exposure to 

DDVP, this approach is not adopted in the current risk assessment. As discussed in Section 

3.1.10, the recent re-evaluation of the cancer data on DDVP (U.S. EPA 2000a,b) has concluded 

that the data available on the carcinogenicity of DDVP is not sufficient for quantitative risk 

assessment. 

3-28
 
3.4. RISK CHARACTERIZATION 

3.4.1. Overview 

The quantitative risk characterizations for workers and members of the general public are 

summarized in Table 3-3. This table is taken directly from Worksheet C02 and is included in the 

body of the risk assessment only for convenience. 

In most cases, exposures to both workers and members of the general public should be 

negligible. If workers take prudent steps to limit both dermal and inhalation exposures, the 

likelihood of exposures to DDVP reaching a level of concern appears to be very low. Similarly, 

members of the general public should not be exposed to substantial amounts of DDVP. The 

DDVP is contained within a PVC strip to insure that the active ingredient is slowly released over 

a long period of time. The strip, in turn, is placed within a trap and the trap is placed in areas that 

will not be generally accessed except in the case of intentional tampering or trap monitoring. 

Nonetheless, this risk assessment develops exposure scenarios for both workers and members of 

the general pubic that are intended to illustrate the potential effects of mishandling or tampering 

with DDVP strips. For workers, the greatest risks are associated with inhalation exposures from 

assembling the traps in enclosed and poorly ventilated spaces or transporting the traps in the 

passenger compartments of vehicles. These risks can be readily avoided. Dermal exposures can 

also lead to lesser but sill undesirable levels of exposure. For members of the general public, all 

of the exposure scenarios are accidental and some are extreme. The most likely of these is the 

accidental contamination of a small body of water. This scenario leads to exposures that are 

below the level of concern by a factor of about 25. If a child were to come into contact with a 

DDVP strip, however, both dermal and oral exposures could substantially exceed a level of 

concern. While such exposures should clearly be avoided, it does not seem likely that frank signs 

of toxicity would be observed. This is consistent with human experience in the use of DDVP 

resin strips. 

3.4.2. Workers 

The risk characterization for workers is highly dependant on how the worker handles the DDVP 

strip during assembly of the milk carton trap. If the trap is assembled outdoors and if the worker 

wears protective gloves during the assembly of the trap, both dermal and inhalation exposures as 

well as consequent risk should be negligible. Whether or not this is common practice is unclear. 

The MSDS states that gloves (vinyl, latex, or rubber) should be worn if the strip is handled for 

prolonged periods of time (Hercon 1993). The product label (Hercon 2004) indicates that hands 

should be washed thoroughly after handling the pest strip. In addition, the Gypsy Moth Program 

Manual (USDA 2001, p. E-6) recommends that workers “use the outer package or rubber gloves 

to handle the insecticide strip. Handle the insecticide strip as little as possible”. If these 

recommendations are followed, direct dermal exposure to DDVP should be negligible. 

If workers assemble traps in enclosed areas or do not use protective gloves during the assembly 

of traps or take other measures to prevent dermal exposure, it is plausible that exposures will 

exceed a level of concern. As summarized in Table 3-3, the potential for undesirable inhalation 

3-29
 
exposures is substantial – i.e., risk quotients up to 18 – if the traps are assembled or transported 

in areas with poor or no ventilation. As discussed in Section 3.2.2.2 and detailed further in 

Appendix 1, these exposure assessments are based on a large number of site and situation 

specific factors – i.e., the volume of the room or area in which the strips are assembled or 

transported, the number of strips that are involved, and the ventilation rates of the area in which 

exposure occurs. Thus, if the pest strips are assembled indoors, it would be prudent to modify 

Worksheet A03a and ensure that the local conditions would likely lead to air concentrations that 

are below the ACGIH (1991) TLV of 0.1 mg/m . 3 

It should be noted that the risk quotients associated with transport of the pest strips in the 

passenger compartment of a vehicle are substantially higher than risk quotients during assembly 

of the traps in a room. High ventilation rates – i.e., 3000 air turnovers per day or about 2 air 

turnovers per minute as detailed in Worksheet A03b – could probably be achieved in a vehicle by 

rolling down the window and this would reduce the inhalation exposure to below the level of 

concern. Nonetheless, transporting DDVP or any volatile neurotoxic agent in the passenger 

compartment of a vehicle is clearly imprudent and should be avoided. 

Dermal exposure is of lesser and only modest concern based on the exposure assessments. As 

noted in Table 3-3, the acute RfD is modestly exceeded – i.e., a hazard quotient of 3 – at the 

upper range of estimated exposures if workers do not wear gloves . This risk quotient is 

associated with a dose of about 0.005 mg/kg bw. It seems unlikely that any adverse effects 

would be experienced at this dose level, which is a factor of 200 below the human NOAEL of 1 

mg/kg [1 mg/kg ÷ 0.005 mg/kg = 200] and a factor of 3,200 below the lowest reported lethal 

dose in humans [16 mg/kg ÷ 0.005 mg/kg = 3200]. While there are uncertainties with the 

exposure assessment on which the risk quotient of 3 is based, contamination of the skin in 

workers not wearing gloves seems to be highly likely. As noted in the product label for the 

VaporTape II strip: “After prolonged storage, a small amount of liquid may form on the strip” 

(Hercon 2004). This liquid would presumably contain DDVP which would contaminate the 

surface of the exposed skin. It is also worth noting that the exposure assessment assumes that 

only the tips of the fingers are contaminated and that the duration of exposure is only 15 minutes 

to 1 hour. If the worker were to contaminate a greater area of the skin or to spend a longer period 

of time assembling the traps, the estimated doses would be greater. 

3.4.3. General Public 

The nature of risks to the general public is substantially different from those to workers. As 

detailed in the previous section, undesirable levels of exposure are plausible for workers if 

sensible measures are not taken to limit exposure. For members of the general public, essentially 

no significant exposures are plausible. The accidental contamination of a small pond with a pest 

strip (Worksheet B02) is probably the most likely exposure scenario. As indicated in Table 3-3, 

this exposure scenario leads to levels of risk that a very low – i.e., the highest hazard quotient is 

0.04, below the level of concern by a factor of 25. 

3-30
 
The probability of a child tampering with a trap is low because the traps will not generally be 

placed in areas that the general public will frequent and will be placed so that the traps are not 

easily accessible to children. Thus, the exposure scenarios involving a child either tampering 

with a trap or otherwise coming into direct contact with a DDVP strip appear to be highly 

unlikely. As illustrated in Table 3-3, dermal exposures would lead to risk quotients of up to 60. 

These exposures would be associated with doses of up to about 0.1 mg/kg (Worksheet B01b). 

This dose is below the lowest reported lethal dose in humans by a factor of about 160 [16 mg/kg 

÷ 0.1 mg/kg], below the acute human NOAEL of 1 mg/kg by a factor of 10, and below the acute 

animal NOAEL of 0.5 mg/kg by a factor 5. Thus, while this type of exposure would be 

considered unacceptable, the plausibility of observing toxic effects seems remote. 

The plausibility and consequences of oral exposures for a child tampering with a DDVP strip are 

very difficult to assess. The unpleasant taste and smell of the pest strip should help to decrease 

the amount of exposure; however, there are reported cases of child poisoning by pest strips 

containing DDVP, although none of the exposures have been fatal. Nonetheless, the oral 

exposure scenarios developed in this risk assessment lead to the highest risk quotients for DDVP, 

a central estimate of 97 with a range of 24 to 380 (Table 3-3 and Worksheet C02). These risk 

quotients are associated with doses of about 0.2 mg/kg with a range of about 0.04 mg/k to 0.6 

mg/kg. As with the dermal exposures for a small child, these exposures should be clearly 

regarded as unacceptable. Nonetheless, it is not clear that any significant adverse effects would 

be observed since the dose estimates are below the human NOAEL of 1 mg/kg and the upper 

range of exposure is below the lowest reported lethal dose by a factor of over 25 [16 mg/kg ÷ 0.6 

= 26.7]. Thus, while these exposure scenarios may be considered extreme and could warrant 

prompt medical attention as a precautionary measure, it is possible that no serious adverse effects 

would be observed. This risk characterization is consistent with the assessment of incidents 

involving exposures to DDVP resin strips – “exposure to resin strips usually do not involve any 

significant acute symptoms that would require medical treatment” (U.S. EPA 2000a, p. 26). 

3.4.4. Sensitive Subgroups 

Children are of primary concern to this risk assessment. As noted above, imprudent handling of 

a DDVP impregnated strip would most likely involve a child. In addition, very young children 

(that is, infants less than 6-months old) may be at special risk because they have incompletely 

developed AChE systems and immature livers (ATSDR 1993). Several other groups may be at 

special risk to all cholinesterase inhibiting compounds, including DDVP. A small proportion of 

the population has an atypical variant of plasma cholinesterase. This condition is known to make 

these individuals sensitive to succinylcholine and may make them more susceptible to exposure 

to DDVP and other AChE inhibitors. Other groups known to have low plasma AChE levels are 

long-distance runners, women in early stages of pregnancy, women using birth control pills, 

individuals with advanced liver disease, alcoholics, individuals with poor nutritional status, and 

individuals with skin diseases. Asthmatics may also be at special risk because DDVP may 

induce or exacerbate respiratory distress (ATSDR 1993). 

3-31
 
3.4.5. Connected Actions 

There are no data regarding the effects of exposure to DDVP combined with exposure to the 

other agents used to control the gypsy moth or the gypsy moth itself. Inhibition of AChE is the 

most sensitive effect of DDVP. This effect is not associated with exposure to the other control 

agents or exposure to the gypsy moth. Therefore, there is no plausible basis for assuming that the 

effects of exposure to DDVP and any or all of the other control agents or the gypsy moth will be 

additive. 

Exposure to other compounds that inhibit AChE are likely to lead to an additive effect with 

DDVP. The most common examples include any other organophosphate or carbamate pesticides 

(ATSDR 1993; Gallo and Lawryk 1991). Thus, if members of the general public or workers use 

other organophosphate pesticides to the extent that AChE activity is substantially inhibited, they 

could be at increased risk if exposed to significant levels of DDVP. 

No studies were located regarding toxicological interactions between Vaportape II and other 

chemicals. There are several studies regarding combined exposures to commercial grade DDVP 

and other chemicals, all of which involve animal exposure, and, in most cases, overtly neurotoxic 

doses of DDVP administered by acute injections. Of the few studies regarding oral or dermal 

exposure to DDVP, most involve acute durations of exposure and do not provide adequate 

evidence of toxicological interactions. Nevertheless, some of these studies are discussed here 

because they concern certain interactions that are generally associated with organophosphate 

insecticides as a class and because they are relevant to the issue of whether or not such 

interactions involving DDVP are plausible. 

Phenothiazine-derived drugs such as chlorpromazine have been shown to enhance the toxicity of 

acutely administered organophosphate insecticides such as parathion (Calabrese 1991). The 

mechanism for this enhancement is not known and may involve altered metabolic activation or 

deactivation of the organophosphate. The interaction between topically applied 

DDVP/Crotoxyphos insecticide and orally administered phenothiazine anthelmintic has been 

studied to a limited extent in livestock, and no obvious interactions have been observed. A series 

of case studies were reported in which young cattle were treated with topical doses of various 

organophosphate insecticides at the end of a 30-day oral treatment with phenothiazine 

anthelmintic, followed by DDVP/Crotoxyphos insecticide 1 month later. There was no evidence 

of an interaction between the phenothiazine and DDVP/Crotoxyphos insecticide (Schlinke and 

Palmer 1973). In a more controlled study, lambs were treated orally with phenothiazine 

antihelmentic (12.5 g initially and 4 days later with 6.25 g every 3 days for nine treatments) or 

topical application of an emulsifiable mixture of 2.3% DDVP and 10% Crotoxyphos (1,550 mL 

of 0.25% emulsion sprayed every 2 weeks for three applications) or both. Erythrocyte 

acetylcholinesterase inhibition and clinical signs of acetylcholinesterase inhibition occurred 

within 40 minutes after each DDVP/Crotoxyphos mixture spray; the severity of the effects was 

not affected by the concurrent phenothiazine treatment (Mohammad and St. Omer 1983, 1985). 

3-32
 
Because of their ability to inhibit acetylcholinesterase and thereby alter the metabolism and 

deactivation of acetylcholine, organophosphate insecticides are expected to interact with drugs 

that mimic the effect of acetylcholine (cholinergic drugs) or that block the effects of 

acetylcholine (anticholinergic drugs). In fact, the anticholinergic drug, atropine, is indicated for 

treatment of severe cholinergic symptoms of organophosphate insecticide toxicity. Because both 

cholinergic and anticholinergic drugs have many other uses, inadvertent interactions in which the 

organophosphate insecticide alters the effect of the drug also should be considered. Acute 

interactions of this type involving DDVP have been studied only to a limited extent in animal 

models of peripheral cholinergic control mechanisms. In one such study, the anticholinergic 

drug, atropine, was administered to dogs (0.022 mg/kg by intramuscular injection) 90 minutes 

after an acute oral dose of 60 mg/kg DDVP, and the heart rate was monitored for cholinergic 

(decreased rate) and anticholinergic (increased rate) effects. Although the DDVP dose alone had 

no effect on heart rate, it did attenuate the acceleration of the heart rate caused by atropine. The 

DDVP dose decreased plasma and erythrocyte cholinesterase by approximately 50% (Dellinger et 

al. 1987). This study suggests that interactions in which DDVP affects the actions of 

anticholinergic drugs (for example, atropine, scopolamine, belladonna alkaloids) are plausible; 

however, there is no evidence of such interactions in humans. 

Chemicals that inhibit carboxyesterases such as the non-organophosphate insecticide, 

triorthotolyl phosphate (TOTP), have been shown to enhance the toxicity of certain 

organophosphate insecticides. Inhibition of carboxyesterases may be a mechanism by which 

certain organophosphate insecticides act synergistically (Calabrese 1991). The significance of 

this interaction mechanism to DDVP toxicity has not been thoroughly investigated. In a study 

using mice, an acute intraperitoneal dose of TOTP 3 days before DDVP treatment enhanced the 

toxicity of an acute intraperitoneal dose of either malaoxon or paraoxon but did not alter the 

toxicity of an intraperitoneal dose of DDVP. Dieldrin, administered orally 4 days before 

sacrifice, increased liver carboxyesterase activity but had no effect on the toxicity of 

subsequently administered DDVP (Ehrich and Cohen 1977). This study suggests that 

carboxyesterase inhibitors may have a more significant effect on malaoxon and paraoxon toxicity 

than on DDVP toxicity. 

The interaction of DDVP with other commonly occurring chemicals in the environment has not 

been well studied. In rats, pretreatment with acetaminophen, a common analgesic, had no effect 

on the acute toxicity of DDVP (Costa and Murphy 1984). 

Toxicological interactions of DDVP have not been studied extensively or well enough to be of 

use in quantitative risk assessment. The few studies described here suggest that certain 

interactions typical of the organophosphate insecticides as a class (for example, anticholinergic 

agents) are plausible for DDVP. Nevertheless, there is no evidence that such interactions 

actually occur in humans. Furthermore, the studies regarding those kinds of interactions in 

animals have examined single exposures and have focused only on the acute anticholinesterase 

activity as the toxic endpoint (usually assessed by measurements of plasma or blood 

cholinesterase or cholinergic symptoms). There need to be more complete interaction bioassays 

3-33
 
that examine multiple dose levels and durations, and more complete assessments of toxicity if 

risks related to possible interactions are to be assessed. 

3.4.6. Cumulative Effects 

Cumulative effects associated with DDVP exposures might be associated with repeated 

exposures during a single season or repeated exposures over several seasons. For the general 

public, the only substantial exposures will occur from tampering with traps containing DDVP. 

Such incidents have not been reported despite the long use of DDVP in traps for the gypsy moth 

as well as other species. These scenarios are considered in this risk assessment as accidental 

exposures, which occur infrequently. Consequently, it does not seem reasonable to expect that 

the same person will be involved repeatedly in such unusual exposures. 

Workers, on the other hand, may be exposed repeatedly to DDVP if they are involved in the 

assembly and placement of traps over a period of several weeks. Such exposures, however, are 

encompassed by the current risk assessment. For inhalation exposures, the risk is characterized 

using the TLV (ACGIH 1991). The TLV is intended to be protective of exposures that occur 

during a typical career (for example, 8 hours/day, 5 days/week, for 20 years). 

For some organophosphates, concern about cumulative effects is diminished because studies 

have demonstrated tolerance to repeated exposures (Gallo and Lawryk 1991). This tolerance has 

not been demonstrated for exposure to DDVP. As is true for exposures involving the general 

public, concern for repeated exposures is diminished because, under normal handling conditions, 

substantial levels of exposure are not anticipated. 

3-34
 
4. ECOLOGICAL RISK ASSESSMENT
 

4.1. HAZARD IDENTIFICATION 

4.1.1. Overview 

As described in Section 3.1.2., DDVP is an organophosphate insecticide. DDVP inhibits 

acetylcholinesterase (AChE) activity, resulting in overstimulation of cholinergic neurons. 

Inhibition of this enzyme in mammalian systems produces a variety of systemic effects, including 

salivation, urination, lacrimation, convulsions, increased bronchial secretions, respiratory 

depression, and even death. DDVP is readily absorbed by the oral, dermal, and inhalation routes 

of exposure. Because the target enzyme (cholinesterase) for DDVP is common to mammals, 

fish, fowl, and insects, toxicity due to DDVP exposure can result in all of these species. By 

contrast, DDVP exhibits low toxicity to plants. 

The available data suggest that invertebrates are more sensitive to DDVP than other organisms. 

For example, the oral LD in honey bees is 0.29 ìg/g bee, and the topical LD is 0.65 ìg/g bee. 50 50 

DDVP is also toxic to birds with an oral LD50 value of < 10 mg/kg for the most sensitive species. 

Short-term repeat dose studies in mammals found that oral exposures to doses below about 0.5 

mg/kg-day or inhalation exposures to 1–2 mg/m3 generally do not result in adverse effects. 

Aquatic animals are also sensitive to DDVP and, as with terrestrial animals, invertebrates may be 

more sensitive than vertebrates. The lowest reported LC50 value in fish is approximately 0.2 

mg/L. Some aquatic invertebrates are much more sensitive to DDVP than fish. For daphnids, 

the most sensitive group of invertebrate species, reported EC50 values range from 0.00007 mg/L 

to 0.00028 mg/L. 

The majority of the toxicity data in ecological receptors is limited to free DDVP, rather than a 

slow-release formulation such as the Vaportape II product used in USDA programs for control of 

the gypsy moth. Hence, the toxicity values reported for indicator species will likely be 

conservative (i.e., suggest greater toxicity) as compared to Vaportape II. U.S. EPA has assessed 

the ecological effects of DDVP; however, the exposures assessed by U.S. EPA are not specific to 

formulations where DDVP is encapsulated in PVC resin. In general, aside from those organisms 

that enter the milk carton trap or those that remove the strip from the trap, toxicity resulting from 

exposure of ecological receptors to DDVP in Vaportape II milk carton traps is not likely. 

4.1.2. Toxicity to Terrestrial Organisms. 

4.1.2.1. Mammals – As summarized in Section 3.1, the database includes a number of toxicity 

studies in experimental mammals. The principal adverse effects of DDVP exposure are directly 

related to inhibition of cholinesterase (the mode of action for DDVP). Inhibition of this enzyme 

in mammalian systems produces a variety of systemic effects (Table 3-1). The nature and 

magnitude of the toxicity produced by a given exposure to DDVP by any route are directly 

related to the dose and rate at which the exposure occurs. In USDA programs for the control of 

the gypsy moth, the use of milk carton traps employing slow-release of DDVP from PVC strips 

essentially precludes rapid exposures to high doses of DDVP. As described in Section 3.1.4, 

4-1
 
short-term animal studies have shown that oral exposures to free DDVP below about 0.5 

mg/kg-day (or inhalation exposures to 1–2 mg/m3) do not result in meaningful reductions in 

cholinesterase activity. Experiments in laboratory mammals that were exposed to DDVP during 

pregnancy (by oral or inhalation route) did not show any effect on fertility or health of the 

offspring, even at levels that produced maternal toxicity (see Section 3.1.9). 

Dietary administration of DDVP (free and encapsulated in PVC resin pellets) has been used as a 

veterinary anthelminthic agent in a variety of species, including dogs (Batte et al. 1966; Batte et 

al. 1967), pigs (Batte et al. 1965; Bris et al. 1968; Stanton et al. 1979; Todd 1967), horses (Himes 

et al. 1967), sheep (Bris et al. 1968), cattle (Bris et al. 1968), dromedary camels (Wallach and 

Frueh 1968), and non-human primates (Wallach and Frueh 1968). In general, oral administration 

of DDVP produced no signs of organophosphate poisoning at doses that were effective at 

reducing intestinal parasites (Wallach and Frueh 1968). For example, two consecutive days of 

dosing at 2.3 in camels or 1.7 mg/kg in non-human primates, respectively, was well tolerated by 

the animals despite debilitating intestinal infection (Wallach and Fueh 1968). In cows, Lloyd and 

Matthysse (1971) reported that diets containing DDVP (in PVC pellets) at doses 1.3, 1.8, or 2.3 

mg/kg bw for 14 days produced no adverse effect on milk production (no other effects were 

reported). No DDVP was found in the milk at 1, 3, 7, 10 or 14 days. Free DDVP – i.e., not 

encapsulated in a PVC resin – produced severe inhibition of cholinesterase activity at a dose of 

4.5 mg/kg (Tracey et al 1960). 

As discussed in Section 3.1.4, the effect of PVC encapsulation on the toxicity of DDVP has been 

quantified in parallel assays (Stanton et al. 1979), in which DDVP (undiluted technical grade) 

and DDVP (impregnated in PVC) were administered to groups of young swine. For the technical 

grade liquid formulation, the LD50 was 157 (113–227) mg/kg and the NOAEL based on lethality 

was 56 mg/kg. For the PVC formulation, no deaths occurred at any doses including 1,000 

mg/kg, the highest dose tested. 

As discussed in Section 3.1.16, simultaneous exposure to DDVP and another cholinesterase 

inhibitor (e.g., organophosphate or carbamate insecticides) or a cholinomimetic agent (e.g., 

pilocarpine and carbachol) would likely enhance the cholinergic toxicity produced by DDVP. 

This is the major toxicologic interaction for DDVP. In addition, Short et al. (1971) also reported 

that DDVP exposure in combination with the muscle relaxant succinylcholine can produce 

cardiac arrythmias, apnea, and death in Shetland ponies depending on the degree of 

cholinesterase inhibition. 

4.1.2.2. Birds – The acute oral LD50 in birds ranges from 6.5–24 mg/kg (WHO 1989, Hudson et 

al. 1984, Grimes and Aber 1988). As in mammals, the signs of DDVP intoxication in birds are 

typical of organophosphorus poisoning (e.g., tremors, and convulsions) and usually appear 

shortly after dosing. At lethal doses, death occurs within 1 hour, with survivors recovering 

completely within 24 h after dosing (WHO 1989). Tucker and Crabtree (1970) found various 

internal hemorrhages at autopsy in sacrificed pheasants and mallard ducks that survived acute 

high dose exposures. 

4-2
 
The data from unpublished egg production and hatchability studies suggests that mallard ducks 

are more sensitive to DDVP than northern bobwhite quail. In mallard ducks, 20 weeks of dietary 

exposure identified a NOEC of 5 ppm and a LOAEL of 15 ppm based on number of eggs laid, 

eggshell thickness, number of viable embryos and number of live 3-week embryos (Redgrave and 

Mansell 1997). Cameron (1996) reported no effect on bobwhite quail reproduction following 

dietary exposure to DDVP at concentrations of 12 or 30 ppm for 20 weeks. At 100 ppm, 

however, statistically significant reductions in the number of eggs laid, viable embryos, live 3­

week embryos, and survivors at 14 days. The short-term dietary LD50 in birds (5 days of 

exposure followed by three days of untreated diet) ranged from 300 ppm in Japanese quail to 

5000 ppm in mallard ducks (Hill et al. 1975). Using chick and duck eggs, injections with DDVP 

at various incubation stages revealed that the LD50 values for these avian species at the midincubation stage were comparable to the rodent oral LD50 values (i.e., >50 mg/kg) (Khera and 

Lyon 1968). 

Five days of continuous exposure of canaries, Indian finches, and budgerigars to DDVP vapor at 

0.14 mg/m3 reduced cholinesterase activity, but produced no overt signs of organophosphate 

intoxication (Brown et al. 1968, as cited by WHO 1989). 

It is important to note that the LD50 values reported from these studies are derived from the active 

ingredient, DDVP, in free form. Encapsulation in PVC resin (such as Vaportape II used in milk 

carton traps) would be expected to slow the release of DDVP, thereby reducing the acute toxicity 

and increasing the LD50 values (Section 3.1.4). No published data are available concerning the 

acute toxicity of DDVP encased in PVC resin in birds. 

4.1.2.3. Terrestrial Invertebrates – In general DDVP is highly toxic to invertebrates with effect 

levels for honey bees below 1 ìg/g bee. In laboratory studies of honey bees, Atkins et al. (1973) 

found an LD50 of 0.495 ìg/bee in 48 h (topical application of dust; 26.7 °C with a relative 

humidity 65%). Beran (1979) reported an oral LD of 0.29 ìg/g body weight and a topical LD 50 50 

of 0.65 ìg/g body weight. 

A variety of other studies are available; however, they are not reported in sufficient detail to 

provide quantitative estimates of exposures. Nevertheless, these studies support the conclusion 

that invertebrates are highly susceptible to the toxic effects of DDVP. Following the exposure of 

honeycombs to DDVP vapor emanating from DDVP resin strips for 4 months, the combs 

absorbed the insecticide and were toxic to bees for approximately one month after exposure. 

Contamination of the bees appeared to be by inhalation rather than direct contact (Clinch 1970). 

Consumption of mulberry leaves sprayed with 1.56–6.25 mg/L DDVP produced 50% mortality 

in silkworm larvae after 4 hours of feeding (Aratake and Kayamura 1973). No adverse effects 

were observed on the hatchability and general condition of silkworm larvae hatched in the 

generation following feeding of mulberry leaves pre-treated with 3 mg/kg DDVP of leaf to adults 

(Yamanoi 1980). 

4-3
 
 

4.1.2.4. Terrestrial Plants (Macrophytes) – Neither the published literature nor the review 

documents include data regarding the phytotoxicity of DDVP. Given the mode of action of 

DDVP, the U.S. EPA (1999a) has determined that toxicity testing in plants is not required for 

registration. 

4.1.2.5. Terrestrial Microorganisms – WHO (1989) reported that the effect of DDVP on 

microorganisms is variable and species dependent. Certain microorganisms are able to 

metabolize DDVP, but DDVP may interfere with the endogenous oxidative metabolism of the 

organism. In certain organisms DDVP inhibits growth, while in others it has no influence or may 

stimulate growth. The above effects have been seen over a concentration range of 0.1–100 mg/L 

(Lieberman and Alexander 1981). 

As noted earlier, the LD50 values reported from these studies are derived from the active 

ingredient, DDVP, in free form. Encapsulation in PVC resin (such as Vaportape II used by the 

Forest Service in milk carton traps) would be expected to slow the release of DDVP, thereby 

reducing the acute toxicity and increasing the LD50 values. No published data are available 

concerning the acute toxicity of DDVP encased in PVC resin in terrestrial microorganisms. 

4.1.2.6. Terrestrial Field Studies – No terrestrial field studies on the effects of free DDVP or 

DDVP in PVC resin were located. Whitehead (1971) has advised caution in the use and 

handling of DDVP, where birds might be exposed because of their particular sensitivity to the 

toxic effects of organophosphate poisoning. In the case of the USDA programs involving the 

use of DDVP in traps, however, the probability of widespread contamination of soil or aquatic 

ecosystems is very low because a small amount of DDVP (590 mg) is used in the Vaportape II 

trap and because the DDVP is released slowly from the PVC resin. 

4.1.3. Aquatic Organisms. 

4.1.3.1. Fish – DDVP is classified as highly toxic to both freshwater and estuarine fish (U.S. 

EPA 1999a). In freshwater fish, reported 96-h LC50 values range from about 0.2 mg/L for lake 

trout or cutthroat trout and 12 mg/L for fathead minnows (U.S. EPA 1999a, p. 12). In estuarine 

fish, 96-h LC50 values range from 0.23–14.4 mg/L for striped mullet and mummichog, 

respectively (U.S. EPA 1999a, p. 12). Sublethal effects – i.e., brain and liver cholinesterase 

inhibition – have been reported in fish at doses of 0.25–1.25 mg/L, but cholinesterase activity 

recovered when the fish were returned to clean water (WHO 1989). The acute toxicity of DDVP 

in cutthroat trout or lake trout was not altered by variations in water hardness from 44 to 162 

mg/L or at pH 6 to 9 (Johnson and Finley 1980). 

Studies of sublethal effects in fish, most involving exposure periods of about 30 days, have 

demonstrated that exposure to #1 mg/L DDVP may produce changes in respiratory rates, serum 

and liver enzyme activity (aside from cholinesterase), lipid and carbohydrate metabolism, and 

hemoglobin and clotting time (WHO 1989). From these reports of adverse effects in fish, WHO 

(1989) derived an NOEC of 0.03 mg/L. 

4-4
 
Only unpublished studies submitted to U.S. EPA were located regarding the chronic toxicity of 

DDVP in fish. These studies are all summarized in U.S. EPA (1999a). A NOEC of 0.0052 mg/L 

was reported for rainbow trout with a corresponding LOAEL of 0.0101 mg/L for a reduction in 

larval survival. Another study found that 0.96 mg/L produced no effects on fry of sheepshead 

minnow, whereas 1.84 mg/L produced statistically significant reductions in fry survival and 

length. As discussed in Section 3.1.7., in vitro studies on cells from embryonic renal tissue of 

carp demonstrated a dose-related decrease in lymphocyte proliferation and myeloid cell 

respiratory burst activities, both of which indicate immunosuppression; however, no effects on 

antibody production were noted in an in vivo study of carp cells (Dunier et al. 1991). The authors 

concluded that the results suggest that chronic exposure to DDVP may impair the immune 

system of fish. 

4.1.3.2. Amphibians – Neither the published literature nor the review documents include data 

regarding the toxicity of DDVP to amphibians. 

4.1.3.3. Aquatic Invertebrates – In general, invertebrates tend to be more sensitive to the toxic 

effects of DDVP than fish. Whereas the lowest reported LC50 value reported in fish is 0.183 

mg/L (the value for lake trout reported by U.S. EPA 1999a, p. 12), the lowest comparable value 

reported for aquatic invertebrates is 0.00007 mg/L (the 48-hour EC50 value for Daphnia pulex 

reported by U.S. EPA 1999a, p. 13). Based on these measures, aquatic invertebrates appear to be 

more sensitive than fish by a factor of over 2500 [0.183 mg/L ÷ 0.00007 mg/L = 2614]. WHO 

(1989) reports that the acute toxicity of DDVP to aquatic insects (stone fly) and estuarine 

crustaceans (hermit crab) is also extremely high (96-hour LC50 values ranging from 0.0001–0.045 

mg/L, respectively). 

As with the data on fish, some of the more important studies are unpublished and have been 

submitted to U.S. EPA for the registration of various uses of DDVP (U.S. EPA 1999a). As 

summarized by U.S. EPA (1999a), the 48-hour EC50 values in two species of water flea range 

from 0.00007 mg/L to 0.00028 mg/L. In an unpublished 21-day study in daphnids, the NOEC 

and LOEC are 0.0000058 mg/L and 0.0000122 mg/L, respectively. 

Not all species of aquatic invertebrates, however, are this sensitive. The most remarkable 

exception to the sensitivity of aquatic invertebrates to DDVP is the freshwater snail; 

Jonnalagadda and Rao (1996) reported a 96-hour LC50 of approximately 21 mg/L in this species. 

Exposure of prawns to DDVP concentrations of 0.31 or 0.62 mg/L for 96 hours produced a 

decrease in hepatic glycogen and an increase in the blood glucose level (Omkar and Shukla 

1984). 

Forget et al. (1998) report static 96-hour LC50 values for copepods ranging from 0.00092–0.0046 

mg/L (different sensitivity depending on life stage). Treatment of eutrophic carp ponds with 

0.325 mg/L DDVP killed Cladocera (predominantly Bosmina and Daphnia species) and 

decreased cyclopods (mainly Cyclops). These reductions were offset by increased development 

of rotifers (mainly Polyarthra and Brachionus species) and phytoplankton (mainly Scenedesmus 

4-5
 
 

and Pediastrum species), so that the total plankton biomass changed only slightly (Grahl et al. 

1981). 

4.1.3.4. Aquatic Plants – The database for DDVP does not contain many reports of its toxicity 

in aquatic plants. In an unpublished report cited by U.S. EPA 1999a), EC50 values >100 ppm are 

reported for green algae, 14 ppm for algae (NOS), and 17-28 ppm for marine diatoms. Butler 

(1977) reported that 3.5 mg/L DDVP produces 50% growth inhibition of Euglena gracilis 

(algae). 

4.1.3.5. Other Aquatic Microorganisms – Neither the published literature nor the review 

documents include data regarding the toxicity of DDVP to other aquatic microorganisms. 

4-6
 
4.2. EXPOSURE ASSESSMENT 

4.2.1. Overview 

As in the human health risk assessment, exposure of terrestrial mammals to DDVP from the 

VaporTape strips used in milk carton traps is likely to be negligible under most circumstances. 

Nonetheless, it is conceivable that some mammals such as racoons or bears could easily access 

and tamper with the milk carton trap. Depending on the proportion of the DDVP strip that is 

consumed, doses (as DDVP in the PVC strip) are estimated to range from 10.5 mg/kg (10% of 

strip) to 105 mg/kg (100% of strip) and the central estimate is taken as 31.6 mg/kg (30% of 

strip). In addition, contamination of water with a pest strip is plausible, although probably rare, 

and is considered in a manner similar to the corresponding scenario in the human health risk 

assessment (Section 3.2.3.4). This scenario is based on the consumption of contaminated water 

by a small mammal and the dose to the animal is estimated at about 0.00003 mg/kg with a range 

of 0.000009 mg/kg to 0.00009 mg/kg. Other exposure scenarios for terrestrial vertebrates, while 

possible, seem far less plausible and are not considered quantitatively. No quantitative exposure 

assessments for terrestrial invertebrates are developed because the milk carton trap will attract 

only male gypsy moths because of the pheromone bait in the milk carton trap. Nontarget insects 

that incidentally enter the trap are likely to be killed by exposure to the DDVP vapor. Exposures 

to aquatic species are based on the same water concentrations used for terrestrial species: 

0.000177 mg/L with a range of 0.000059 mg/L to 0.00059 mg/L. 

4.2.2. Terrestrial Vertebrates 

4.2.2.1. Oral Exposure to DDVP Strip – For the exposure of a young child discussed in Section 

3.2.3.3, only sucking on the strip rather than ingestion of all or part of the strip is considered. 

Various species of wildlife, however, are probably capable of consuming all or part of a pest 

strip. For the current risk assessment, it will be assumed that a racoon tampers with a milk carton 

trap and consumes part or all of the strip – i.e., 590 mg of DDVP in the PVC formulation. 

Taking a body weight of about 5.6 kg for an adult racoon (the average of the values reported by 

U.S. EPA/ORD 1993, p. 2-236) and assuming that the animal consumes between 10% and 100% 

of the strip with a central value of 30%, the dose to the racoon would be about 31.6 mg/kg with a 

range of 10.5 mg/kg to 105 mg/kg (Worksheet D01). 

4.2.2.2. Oral Exposure to Water Contaminated with DDVP – Estimated concentrations of 

DDVP in water are identical to those used in the human health risk assessment (Worksheet B02) 

and involve the accidental contamination of a small pond with a DDVP-PVC strip. The only 

major differences in this scenario compared to the scenario in the human health risk assessment 

involve the weight of the animal and the amount of water consumed. There are well-established 

relationships between body weight and water consumption across a wide range of mammalian 

species (e.g., U.S. EPA/ORD 1993). These relationships are used to estimate the amount of 

water that a 20 g mammal would consume in one day (Worksheet D02). Unlike the human 

health risk assessment, estimates of the variability of water consumption are not available. Thus, 

for this acute scenario, the only factor affecting the variability of the ingested dose estimates is 

the amount of DDVP that might be released in one day. These amounts are discussed in Section 

4-7
 
3.2.3.4 and are used in Worksheet D02. As indicated in Worksheet D02, the central estimate of 

the dose is about 0.00003 mg/kg with a range of 0.000009 mg/kg to 0.00009 mg/kg. 

4.2.3. Terrestrial Invertebrates 

As in the previous risk assessment (USDA 1995b), quantitative exposure assessments for 

terrestrial invertebrates are not considered. The only terrestrial invertebrates that are likely to 

come into close contact with the DDVP strip are male gypsy moths, which will be attracted by 

the disparlure in the trap, or carnivorous wasps and hornets that may enter the trap to feed on 

dead and dying gypsy moths. Other insects and perhaps other invertebrates such as spiders might 

incidentally enter the milk carton traps. Because DDVP in a non-specific insecticide, nontarget 

invertebrates would likely be killed by exposure to the DDVP vapor within the trap. 

4.2.4. Aquatic Species 

The exposure assessment for aquatic species is based on concentrations of DDVP in water that 

are identical to the concentrations used in the human health risk assessment (Worksheet B02) and 

the exposure assessment for a small mammal drinking contaminated water (Worksheet D02). As 

indicated in these worksheets, the central estimate of the concentration of DDVP in the pond is 

0.000177 mg/L with a range of 0.000059 mg/L to 0.00059 mg/L. 

4-8
 
4.3. DOSE-RESPONSE ASSESSMENT 

4.3.1. Overview 

Given the limited nature of the use of DDVP in programs to control the gypsy moth and 

consequent limited number of exposure assessments, the dose-response assessment for DDVP is 

relatively simple. For terrestrial mammals, a value of 240 mg/kg from a study using DDVP in a 

PVC formulation is used for direct exposure to the DDVP-PVC strip – i.e., a raccoon tampering 

with a milk carton trap and consuming all or part of the DDVP strip. At the dose of 240 mg/kg, 

no mortality or frank signs of AChE inhibition were observed. For the contaminated water 

scenario, the NOAEL of 0.5 mg/kg from a study involving exposure to free or unformulated 

DDVP is used. This NOAEL is from the study that forms the basis for the acute RfD used in the 

human health risk assessment. Although DDVP is classified as highly toxic to fish, the estimated 

levels of acute exposure for fish are far below the 30-day NOEC of 0.03 mg/L. Thus, this value 

is used for all fish and no attempt is made to consider differences in sensitivity among fish. A 

somewhat different approach is taken with aquatic invertebrates, some of which are more 

sensitive to DDVP than fish by a factor of over 2500. Risks to sensitive species of aquatic 

invertebrates – i.e., daphnids and other small arthropods – are characterized based on the lowest 

reported LC50 value, 0.00007 mg/L from a 48-hour bioassay in Daphnia pulex. Some other 

groups of aquatic invertebrates, such as snails, appear to be much less sensitive than small 

arthropods. Risks to such tolerant species are based on a LC50 value of 21 mg/L in a freshwater 

snail. 

4.3.2. Toxicity to Terrestrial Organisms 

Two different types of exposure assessments are given for terrestrial vertebrates: direct 

consumption of all or part of the DDVP-PVC stip (Section 4.2.2.1) and consumption of water 

contaminated with DDVP (4.2.2.2). The former scenario involves exposure to the formulated 

DDVP and the latter exposure scenario involves exposure to unformulated or free DDVP. For 

the exposure assessment involving direct consumption of the DDVP-PVC strip, the dose of 240 

mg/kg for neurotoxicity from the study by Stanton et al. (1979) will be used to characterize risk. 

No mortality or frank signs of AChE inhibition were observed at this dose. For exposure to free 

DDVP in water, the NOAEL of 0.5 mg/kg for changes in AChE activity and other signs of 

neurotoxicity will be used to characterize risk. This is the NOAEL selected by the U.S. EPA 

(2000a) as the basis for the acute oral RfD for DDVP. As indicated in Section 4.4., these two 

NOAEL values are substantially below the corresponding exposure levels. Thus, elaboration of 

the dose-response assessment is not necessary. 

4.3.3. Aquatic Organisms 

4.3.3.1. Fish – The U.S. EPA typically uses LC50 values as benchmark doses for developing 

acute hazard quotients and the most sensitive LC50 of 0.183 mg/L was used by U.S. EPA in it’s 

ecological risk assessment for DDVP (U.S. EPA 1999a, p. 29). USDA risk assessments typically 

prefer to use NOEC (no observed effect concentrations) when such data are available. As 

discussed in Section 4.1.3.1, WHO (1989) has identified an NOEC of 0.03 mg/L from studies 

involving exposure periods of about 30 days. This NOEC will be adopted in the current risk 

assessment. While the application of a 30-day NOEC to the acute and much shorter term 

4-9
 
exposures considered in this risk assessment is likely to be over-protective, this has no impact on 

the characterization of risk because the anticipated levels of acute exposure are substantially 

below this NOEC. Also because this conservative NOEC value is below a level of concern, 

separate assessments are not made for sensitive and tolerant species of fish. This is discussed 

further in Section 4.4. 

4.3.3.2. Aquatic Invertebrates – As noted in Section 4.1.3.3, some aquatic invertebrates are 

much more sensitive to DDVP than fish. Based on the lowest reported LC50 values in fish and 

invertebrates, some aquatic invertebrates are more sensitive than fish by a factor of over 2500. 

There is, however, a very wide range of tolerances in aquatic invertebrates. The lowest reported 

LC value is 0.00007 mg/L. This is a 48-hour LC value in Daphnia pulex reported by U.S. 50 50 

EPA (1999a, p. 13). A NOEC value is not reported by U.S. EPA (1999a). Thus, the LC 50 

0.00007 mg/L is used directly in the risk characterization for sensitive aquatic invertebrates. As 

also noted in Section 4.1.3.3, the sensitivity of aquatic invertebrates to DDVP is highly variable. 

The least sensitive group of species appears to be aquatic snails, with a reported 96-hour LC50 of 

21 mg/L (Jonnalagadda and Rao 1996). This value will be used to characterize risks in tolerant 

aquatic invertebrates. 

4-10
 
4.4. RISK CHARACTERIZATION 

4.4.1. Overview 

As with the human health risk assessment, it is anticipated that typical exposures and consequent 

risks to nontarget species should be negligible. As with the human health risk assessment, it is 

anticipated that typical exposures and consequent risks to most nontarget species should be 

negligible. The containment of the DDVP within a slow release PVC strip combined with the 

target specific nature of pheromone baited traps should reduce the risks of inadvertent effects in 

non-target species. Other insects and arthropods that may inadvertently enter the trap will 

probably be killed by DDVP vapor. While such inadvertent contact may occur, it is not likely to 

impact substantial numbers of nontarget insects or arthropods. 

Because of the limited use of DDVP, a relatively small number of exposure scenarios – all of 

which might be considered accidental or incidental – are developed. For terrestrial mammals, 

contact with the pest strip could occur by an animal directly tampering with a trap or by an 

animal consuming water that had been accidentally contaminated with a DDVP strip. Adverse 

effects would not be expected in either case. In the case of accidental contamination of a small 

body of water with a DDVP strip, concentrations of DDVP in the water would be below the level 

of concern for fish by factors of about 50 to 500. Some aquatic invertebrates, however, might be 

affected. For the most sensitive species of aquatic invertebrates – i.e., small aquatic arthropods 

such as daphnids – exposures could substantially exceed laboratory LC50 values by factors of up 

to about 8. Exposures to tolerant aquatic invertebrates – such as snails – would be below a level 

of concern by a substantial margin – i.e., factors of about 30,000 to 300,000. The exposure 

assessments that serve as the bases for these risk characterizations are highly dependent on 

specific conditions – i.e., how much DDVP was in the strip at the time that the contamination 

occurred and the size of the body of water that was contaminated. 

4.4.2. Terrestrial Organisms 

There is no indication that adverse effects in terrestrial vertebrates are likely. This assessment is 

based on the exposure scenarios for a relatively small mammal – i.e., a raccoon – consuming all 

or part of a DDVP-PVC strip as well as a very small mammal consuming water that had been 

contaminated with a pest strip. 

The former scenario, direct consumption, may be plausible but is clearly extreme. The upper 

range of the exposure assessment assumes that the animal consumes the entire strip with a 

resulting dose of about 100 mg/kg (Section 4.2.2.1). The assessment of risk is based on a 

controlled laboratory study using a DDVP-PVC formulation in which no mortality was observed 

at 1,000 mg/kg and no signs of AChE inhibition were apparent at 240 mg/kg (Section 4.3.2). 

The dose of 100 mg/kg associated with upper range of the hazard quotient, 0.4, is below the the 

NOAEL by a factor 2.5. 

The scenario for the consumption of contaminated water is based the assumption that a fresh 

DDVP strip inadvertently contaminates a small pond and, at the upper range of the estimated 

dose, the further assumption that all of the DDVP in the strip leaches into the water (Section 

4-11
 
4.2.2.2 and Worksheets D02). The estimated dose is probably higher and perhaps much higher 

than what might actually occur because degradation of the DDVP in water is not considered. 

Even with these highly protective assumptions, the upper range of the risk quotient is only 

0.0002 – i.e., below the level of concern (1) by a factor of 5,000. Thus, there is no plausible 

basis for asserting that adverse effects are likely. 

No quantitative risk characterization is presented for terrestrial invertebrates. This approach is 

taken because there is no reason to anticipate that significant exposures to nontarget invertebrates 

are likely. It is possible that some insects and perhaps other arthropods could inadvertently enter 

a milk carton trap. In such a case, it is likely that the nontarget organisms would be killed by the 

DDVP vapor. While this is the intended effect in the target species, the gypsy moth, the efficacy 

of the traps is dependant on the use of another agent, disparlure, that serves as an attractant to 

male gypsy moths. As discussed in the risk assessment for disparlure, this attractant is highly 

specific to the gypsy moth and will not attract other species. Thus, the numbers of nontarget 

species that might be killed by inadvertently entering the traps is likely to be small and 

inconsequential. 

4.4.3. Aquatic Organisms 

4.4.3.1. Fish – There is no indication that fish are likely to be adversely affected by the use of 

DDVP in PVC strips. The exposure assessment for fish (Section 4.2.4) is based on the same very 

conservative exposure assessment used for mammals – i.e., the concentrations in water are likely 

to be over-estimated. The dose-response assessment is based on a 30-day NOEC for sublethal 

effects. The resulting risk quotients – i.e., 0.002 to 0.2 – are below the level of concern by 

factors of 50 to 500. 

4.4.3.2. Aquatic Invertebrates – As discussed in Section 4.3.3.2, some aquatic invertebrates are 

much more sensitive to DDVP than fish and this difference in sensitivity impacts the 

characterization of risk. Based on the same conservative exposure assessment used for both fish 

and terrestrial vertebrates, some sensitive aquatic invertebrates could be adversely affected by 

DDVP contamination of water. As in the other exposure assessments involving contaminated 

water, this exposure scenario should be regarded as accidental rather than routine. In other 

words, under normal circumstances, water contamination from DDVP strips will be negligible 

and this is consistent with the conclusions reached by U.S. EPA (1999a, p. 25). Nonetheless, 

based on the modeled concentrations in the event of the accidental deposition of a strip 

containing 590 mg of DDVP into a small pond, concentrations of DDVP in the water would 

reach or substantially exceed the LC50 value for sensitive invertebrates and substantial mortality 

in sensitive invertebrates could occur. 

The actual extent of mortality would depend on the rate at which DDVP is released from the 

strip, the degree of mixing that occurs in the water, and the rate of breakdown and dissipation of 

DDVP. These processes cannot be generically modeled but the conservative exposure 

assessment used to estimate concentrations in water suggests that adverse effects in sensitive 

aquatic invertebrates are plausible. No effects are likely in less sensitive aquatic invertebrates 

4-12
 
such as aquatic snails. As discussed in Section 3.2.3.4, the hydrolysis of DDVP in water is rapid 

and it is likely that the estimates of adverse effects in some aquatic invertebrates would apply to 

only a very limited area near the pest strip rather than to the larger area of the body of water that 

is contaminated. 

4-13
 
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5-18
 
Table 2-1: Selected physical and chemical properties of DDVP 

Synonyms and trade names	 SD 1750; Astrobot; Atgard; Canogard; Dedevap; Dichlorman; Dichlorophos; 

Dichlorvos; Divipan; Equigard; Equigel; Estrosol; Herkol; Nogos; Nuvan: Task; 

Vapona; Verdisol (Budavari 1989) 

U.S. EPA Reg. No. 8730-50 (Hercon 2004)
 

CAS number 62-73-7 (ARS/PPD 1995; Meylan and Howard 2000)
 

Molecular weight 220.98 (Budavari 1989)
 

Molecular formula C H Cl O P (ARS/PPD 1995; Budavari 1989; Meylan and Howard 2000)
 

SMILES Notation O=P(OC)(OC)OC=C(CL)CL (Meylan and Howard 2000)
 

Appearance/state, ambient Liquid (ARS/PPD 1995; Budavari 1989)
 

mg/L to ppm conversion for 1 ppm = 9.04 mg/m3

 (NOISH 2002)
 

air concentrations 1 mg/m3

 = 0.11 ppm
 

Boiling point 120EC at 14 mm Hg (ARS/PPD 1995)
 

4 7 2 4 

251.76 EC (Meylan and Howard 2000) 

Vapor pressure 1.2×10-2 mm Hg (Budavari 1989) 

1,600 mPa (ARS/PPD 1995) 

Water solubility (mg/L) 10,000 (Budavari 1989) 

8,000 (ARS/PPD 1995) 

Specific gravity 1.44 (Shell Chemical Company 1972) 

log K ow 1.40-2.29 (ARS/PPD 1995) [i.e., Kow = 101.4 = 25.1] 

0.60 (estimated) (Meylan and Howard 2000) 

1.47 (experimental) (Meylan and Howard 2000; U.S. EPA 1992) 

3 Henry’s law constant	 0.044 Pa m /mole at 20 EC (ARS/PPD 1995) 3 8.58E-007 atm-m /mole (Meylan and Howard 2000)

Koc 40.2 (Meylan and Howard 2000) 

BCF 0.4486 (Meylan and Howard 2000) 

Hydrolysis half-time (days) 0.022 to 0.347 (ARS/PPD 1995) 

Aqueous photolysis halftime 2.295 (ARS/PPD 1995) 

(days) 

Tables - 1 
Table 3-1. Common effects of acetylcholinesterase inhibition a 

System 

Parasympathetic 

Receptor 

Type 

Muscarinic 

Organ 

Eye 

Iris muscle 

Ciliary muscle 

Glands 

Lacrimal 

Action 

Contraction 

Secretion 

Manifestation 

Miosis 

Blurred vision 

Tearing 

Salivary Salivation 

Respiratory Bronchorrhea; rhinitis; 

pulmonary edema 

Gastrointestinal Nausea; vomiting; diarrhea 

Sweat Perspiration 

Sympathetic 

(sympatholytic) 

Heart 

Sinus node Slowing Bradycardia 

Atrioventricular

 (AV) node 

Smooth Muscle 

Bronchial 

Increased refractory

 period 

Contraction 

Dysrhythmia; heart block 

Bronchoconstriction 

Gastrointestinal Vomiting; 

cramps; diarrhea 

Sphincter 

Bladder 

Fundus 

Relaxation 

Contraction 

Fecal incontinence 

Urination 

Sphincter Relaxation Urinary incontinence 

Neuromuscular nicotinic Skeletal 

Heart 

Excitation 

Excitation 

Fasciculations; cramps followed 

by weakness; pupillary dilation; 

loss of reflexes; paralysis 

Tachycardia 

Central nervous Brain/Brainstem 

a Modified from ATSDR 1993 

Excitation (early) 

Depression (late) 

Headache; malaise; dizziness; 

confusion; manic or bizarre 

behavior 

Depression, then loss of 

consciousness; respiratory 

depression; respiratory 

(diaphragm) paralysis 

Tables - 2 
Table 3-2: Parameters used in DDVP air model 

Parameter Value Units Description/Comment/Reference 

ã 37.5 Unitless Apparent adsorption coefficient based on 

optimization using relative errors. See Worksheet 

A02b and Section 3.2.2.2 for discussion. 

ë 0.023 day-1 First-order release rate from Shell No-Pest Strips 

from Gillett et al. (1972a). Used to estimate ã from 

the data reported by Slomka (1970). 

0.04 day-1 First-order release rate from VaporTape II strips 

based on data from Hercon (1994). See Worksheet 

A01. 

RH 0.4 Unitless Relative humidity used by Gillett et al. (1972a) and 

used for model application in Worksheets A02a, 

A02b, A03a, and A03b. This is a sensitive 

parameter. See text for discussion. 

k 109.3 days -1 Hydrolysis rate constant from Gillett et al. (1972a) 

At/Va 0, 60, and 

625, and 

6500 

day-1 Air turnover rate – i.e., the ratio of air flow to room 

volume. Values of 0 and 60 used by Gillett et al 

(1972a) for no ventilation and very poor ventilation, 

respectively. Values of 300 and 3000 are selected as 

adequate ventilation for a garage and vehicle, 

respectively – see Section 4.4 for discussion. 

Tables - 3 
Table 3-3: Summary of Risk Characterization for Human Health Risk Assessment 1 

Hazard Quotients 

Group Scenario Central Lower Upper Toxicity Units Section 

Value 

Workers 

Inhalation During Assembly 3 0.9 5 0.1 mg/m3 3.3.2.3 

Inhalation During Transport 15 1.0 18 0.1 mg/m3 3.3.2.3 

Dermal During Assembly 0.5 0.2 3 0.0017 mg/kg 3.3.2.2 

Child 

Incidental Dermal Contact 10 1.8 60 0.0017 mg/kg 3.3.2.2
 

Oral Exposure from Strip 97 24 380 0.0017 mg/kg 3.3.2.1
 

Oral Exposure from Water 0.008 0.002 0.04 0.0017 mg/kg 3.3.2.1 

1

 All of the exposure assessments on which these hazard quotients are based should be regarded 

as atypical and most are extreme. As noted in Section 3.2, typical exposures for workers and 

members of the general public will typically be negligible. 

Tables - 4 
Table 4-1: Summary of Exposure Assessments and Risk Characterization for Non-target 

Species 

Exposure Assessments 

Estimated Exposures 

Species Scenario Central Lower Upper Units Worksheet 

Racoon Consumption 3.16E+01 1.05E+01 1.05E+02 mg/kg D01 as 

DDVP-PVC 

Small mammal Contaminated 2.59E-05 8.64E-06 8.64E-05 mg/kg D02 as free 

Water DDVP 

Aquatic Species Contaminated 0.000177 0.000059 0.00059 mg/L D02 

Water 

Risk Characterization 

Risk Quotients 1 Toxicity Value 

Species Scenario Central Lower Upper Value Units 

Racoon Consumption 0.1 0.04 0.4 240 mg/kg as 

DDVP-PVC 

Small mammal Contaminated 

Water 

0.0001 0.00002 0.0002 0.5 mg/kg as 

free DDVP 

Aquatic Species 

Fish 0.006 0.002 0.02 0.03 mg/L NOEC 

as free 

DDVP 

Sensitive Invertebrates 3 0.8 8 0.00007 mg/L LC50 

as free 

DDVP 

Tolerant Invertebrates 0.00001 0.000003 0.00003 21 mg/L LC 50

as free 

DDVP 

1

 Risk quotients are calculated as the exposure value, given in the upper section of the table divided by the 

toxicity value specified for the non-target species. This ratio is rounded to one significant digit. 

Tables - 5 
Figure 3-1: Concentration of DDVP in Air After the Placement of One Shell No-Pest Strip in 

an Unventilated Room (At/Va=0) and a Poorly Ventilated Room (At/Va=60)(data from Slomka 

1970). See text for discussion and Worksheet A02b for details. 

Figures - 1 
Appendix 1: Application and Optimization of DDVP Inhalation Exposure Model 

Gillett et al. (1972a) proposed the following model for estimating concentrations of DDVP in air 

from the release of DDVP from PVC pest strips: 

(Eq. A-1) 

The terms in the above equation are defined as follows: 

t time after start of release 

C t concentration of DDVP in air at time, t (days) 

M0 mass of DDVP in strip or strips at time zero (mg) 

Va volume of room or other space (m )3 

ã apparent adsorption coefficient of DDVP on to surfaces 

exp(x) the exponential function, e , where is the constant 2.718 and x is x 

any numeric expression 

ë first-order release rate constant (days )-1 

RH relative humidity (proportion) 

At air flow rate (m /day) 3 

k first-order hydrolysis rate (days )-1 

and the parameters used in the model are summarized in Table 3-2. 

The above equation is modified from Equation 3 in Gillett et al. (1972, p. 126). For simplicity, 

the term RH is used above rather than the term p/p used by Gillett – i.e., the ratio of the ambient 0 

to the saturated vapor concentration of water. More significantly, the equation given in the 

Gillett publication – i.e., Equation 3, p. 126 – contains two typographical errors. Both errors are 

in the numerator to the second exponential function. The Gillett publication fails to note that the 

negative of the sum, k RH + At/Va, must be used. These are essentially two first order processes 

– i.e., hydrolysis and dilution. If the negative of these values is not used, the equation models 

first-order growth rather than dissipation. Dissipation is clearly the intent of this term in the 

equation. The second more trivial error is that the k RH + At/Va term must be multiplied by t 

within the second exponential term. Otherwise, the units of the equation do not reduce to a 

concentration in air. This is analogous to the general equation for first-order absorption and firstAppendix 1-1 
 

order elimination (e.g., Goldstein et al. 1974, p. 333). The discussion of the validation of this 

equation by Gillett et al. (1972a) and the implementation of this equation in the Worksheets uses 

the corrected form of the equation given above. Using the equation given by Gillett et al. 

(1972a) does not reproduce the results illustrated in Figure 4 of Gillett et al. (1972, p. 128) or in 

Worksheets A02a and A02b.] 

Gillett et al. (1972a) applied this model to the data from Slomka (1970) in which a single Shell 

3 No-Pest Strip containing 20,000 mg of DDVP was placed rooms with a volume of 28.3 m at

25EC and a relative humidity of 40%. Two different ventilation conditions were used, no 

ventilation and poor ventilation. No ventilation is characterized simply as a room with no air 

turnover – i.e., At/Va = 0. Poor ventilation is characterized as a room in which 20 air exchanges 

occurred per day – i.e., At/Va = 20. The apparent adsorption coefficient (ã) was treated as an 

empirical parameter and optimized to the data from Slomka (1970). All other model parameters 

were taken from the literature as specified in Table 3-2. 

Gillett et al. (1972a) report an optimized value of 44.76 for the apparent adsorption coefficient 

(ã) but do not specify how this parameter was optimized. For the current risk assessment, the 

model given above was implemented in EXCEL and the data from Slomka (1970) was taken 

from Figure 4 in the publication of Gillett et al. (1972a). The apparent adsorption coefficient was 

then optimized using the EXCEL Solver function with the quasi-Newton method (with the 

tangent estimate and forward derivative options). Two sets of optimizations were conducted. 

The first was based on minimizing the standard square of error (Worksheet A02a) and the second 

was based on square of the relative error (Worksheet A02b). These optimizations yielded 

estimates of the apparent adsorption coefficient (ã) of 54.5 and 37.5, respectively, which bracket 

the estimate of 44.76 reported by Gillett et al. (1972a). As illustrated in Worksheets A02a and 

A02b, both of the optimized values fit the data from Slomka (1970) reasonably well. For the 

current risk assessment, the worker exposure estimates are based on the apparent adsorption 

coefficient (ã) 37.5, which leads to modestly higher estimates of exposure than do the higher 

estimates of the apparent adsorption coefficient. The fit of the Gillett et al. (1972a) model to the 

data from Slomka (1970) using the apparent adsorption coefficient (ã) of 37.5 is illustrated in 

Figure 3-1 (which is in turn taken from Worksheet A02b). 

Appendix 1-2 
Appendix 2: Estimates of dermal absorption rates for DDVP 

Table A2-1: Method for estimating the dermal permeability (K in cm/hr) and 95% p

confidence intervals. 

Model parameters 

Coefficient for ko/w 

Coefficient for MW 

ID 

C_KOW 

C_MW 

Value 

0.706648 

0.006151 

Model Constant C 2.72576 

Number of data points 

Degrees of Freedom (d.f.) 

Critical value of 0.025 t with 87 d.f.a 

DP 

DF 

CRIT 

90 

87 

1.96 

Standard error of the estimate SEE 45.9983 

Mean square error or model 

variance 

MDLV 0.528716 

Standard deviation of model (s) MSD 0.727129 MDLV0.5 

XNX, cross products matrix 0.0550931 -0.0000941546 -0.0103443 

-0.0000941546 0.0000005978 -0.0000222508 

-0.0103443 -0.0000222508 0.00740677 

a

Mendenhall and Scheaffer, 1973, Appendix 3, Table 4, p. A31. 

NOTE: The data for this analysis are taken from U.S. EPA (1992), Dermal Exposure 

Assessment: Principles and Applications, EPA/600/8-91/011B, Table 5-4, pp. 5-15 through 5-19. 

The U.S. EPA report does not provide sufficient information for the calculation of confidence 

intervals. The synopsis of the above analysis was conducted in STATGRAPHICS Plus for 

Windows, Version 3.1 (Manugistics, 1995) as well as Mathematica, Version 3.0.1.1 (Wolfram 

Research, 1997). Although not explicitly stated in the U.S. EPA report, 3 of the 93 data points 

are censored from the analysis because they are statistical outliers: [Hydrocortisone-21-yl]­

hemipimelate, n-nonanol, and n-propanol. The model parameters reported above are consistent 

with those reported by U.S. EPA but are carried out to a greater number of decimal places to 

reduce rounding errors when calculating the confidence intervals. See notes to Worksheet A07a 

for details of calculating maximum likelihood estimates and confidence intervals. 

Appendix 2-1 
Table A2-2 p : Calculation of dermal permeability rate (K ) in cm/hour for DDVP. 

Parameters Value Units Reference 

Molecular weight 220.98 g/mole 

Ko/w at pH 7 29.51 unitless 

10 o/w log K 1.47 

Column vector a for calculating confidence intervals (see Worksheet A07a for definitions.) 

a_1 1 

a_2 220.98 

a_3 1.47 

Calculation of a' A (X'X) A a - see Worksheet A07b for details of calculation. -1 

Term 1 0.0190806955 

Term 2 0.001157619 

Term 3 -0.006428795 

a' A (X'X) A a -1 0.0138 calculation verified in Mathematica 3.0.1.1 

10 p 10 o/w log k = 0.706648 log (k ) - 0.006151 MW - 2.72576 Worksheet A07b 

log10 of dermal permeability 

Central estimate -3.04623542 ± 0.025 t × s × a'A(X'X)

-1Aa0.5 

Lower limit -3.21365532088 - 1.9600 × 0.727129 × 0.1174734012 

Upper limit -2.87881551912 � 1.9600 × 0.727129 × 0.1174734012 

Dermal permeability 

Central estimate 0.00090 cm/hour 

Lower limit 0.00061 cm/hour 

Upper limit 0.0013 cm/hour 

Details of calculating aNXNX a 

The term a'A(X'X)

-1Aa requires matrix multiplication. While this is most easily accomplished 

using a program that does matrix arithmetic, the calculation can be done with a standard 

calculator. See details on following page. 

Appendix 2-2 
Letting 

a = {a_1, a_2, a_3} 

and

-1 (X'X) = {
 

{b_1, b_2, b_3},
 

{c_1, c_2, c_3},
 

{d_1, d_2, d_3}
 

},
 

a'A(X'X)

-1Aa is equal to 

Term 1: {a_1 ×([a_1×b_1] + [a_2×c_1] + [a_3×d_1])} + 

Term 2: {a_2 ×([a_1×b_2] + [a_2×c_2] + [a_3×d_2])} + 

Term 3: {a_3 ×([a_1×b_3] + [a_2×c_3] + [a_3×d_3])}. 

Appendix 2-3 
Appendix L 

Gypsy Moth 

Risk Assessment 

Figure L-1. Gypsy moth caterpillars cluster at the base of a banded tree (Arlington, 

Virginia, 1905). 

SERA TR 04-43-05-07c 

Control/Eradication Agents for the
 

Gypsy Moth ­

Human Health and Ecological Risk Assessment
 

for the Gypsy Moth
 

FINAL REPORT
 

Prepared for: 

USDA, Forest Service
 

Forest Health Protection 

GSA Contract No. GS-10F-0082F
 

USDA Forest Service BPA: WO-01-3187-0150
 

Requisition No.: 43-3187-1-0269
 

Task No. 5
 

Submitted to:
 

Dave Thomas, COTR
 

Forest Health Protection Staff
 

USDA Forest Service
 

Rosslyn Plaza Building C, Room 7129C
 

1601 North Kent Street
 

Arlington, VA 22209
 

Submitted by:
 

Patrick R. Durkin
 

Syracuse Environmental Research Associates, Inc. 

5100 Highbridge St., 42C
 

Fayetteville, New York 13066-0950
 

Telephone: (315) 637-9560
 

Fax: (315) 637-0445
 

E-Mail: SERA_INC@msn.com
 

Home Page: www.sera-inc.com
 

August 23, 2004 
TABLE OF CONTENTS
 

TABLE OF CONTENTS ....................................................... ii
 

LIST OF TABLES ............................................................ . v
 

LIST OF FIGURES ........................................................... ..v
 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .vi
 

COMMON UNIT CONVERSIONS AND ABBREVIATIONS . . . . . . . . . . . . . . . . . . . . . . . . .vii
 

CONVERSION OF SCIENTIFIC NOTATION .................................... viii
 

EXECUTIVE SUMMARY ................................................... . . i x 
 

1. INTRODUCTION ......................................................... 1-1
 

2. GYPSY MOTH AS PEST SPECIES .......................................... 2-1
 

2.1. OVERVIEW ..................................................... 2-1
 

2.2. INFESTATIONS .................................................. 2-1
 

2.3. LIFE-CYCLE ..................................................... 2-3
 

2.4. DISEASE AND PREDATION ....................................... 2-5
 

3. HUMAN RISK ASSESSMENT ..................................... ......... 3-1
 

3.1. HAZARD IDENTIFICATION ....................................... 3-1
 

3.1.1. Overview. ................................................ 3-1
 

3.1.2. Mechanisms of Action....................................... 3-1
 

3.1.3. Effects on Skin............................................. 3-2
 

3.1.4. Effects on Eyes and Respiratory Tract. . . . . . . . . . . . . . . . . . . . . . . . . . . 3-4
 

3.1.5. Other Potential Effects....................................... 3-4
 

3.2. EXPOSURE ASSESSMENT ........................................ 3-6
 

3.2.1. Overview. ................................................ 3-6
 

3.2.2. Exposure Metameter. ....................................... 3-6
 

3.2.3. Intensity of Exposures ....................................... 3-7
 

3.3. DOSE-RESPONSE ASSESSMENT ................................... 3-8
 

3.3.1. Overview ................................................. 3-8
 

3.3.2. Effects on Skin ............................................ 3-8
 

3.3 3. Others Effects .......................................... 3-11
 

ii 
TABLE OF CONTENTS (continued) 

3.4. RISK CHARACTERIZATION ...................................... 3-12
 

3.4.1. Overview ................................................ 3-12
 

3.4.2. Effects on Skin ........................................... 3-12
 

3.4.3. Other Endpoints .......................................... 3-13
 

3.4.4. Sensitive Subgroups........................................ 3-13
 

3.4.5. Connected Actions......................................... 3-14
 

3.4.6. Cumulative Effects......................................... 3-14
 

4. ECOLOGICAL RISK ASSESSMENT ......................................... 4-1
 

4.1. HAZARD IDENTIFICATION ....................................... 4-1
 

4.1.1. Overview. ................................................ 4-1
 

4.1.2. Terrestrial Organisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-1
 

4.1.2.1. Mammals ......................................... 4-1
 

4.1.2.2. Birds ............................................. 4-3
 

4.1.2.3. Reptiles and Amphibians . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-3
 

4.1.2.4. Terrestrial Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-4
 

4.1.2.5. Terrestrial Plants ( M acrophytes ) . ............... 4-5
 

4.1.2.6. Soil and Terrestrial Microorganisms . . . . . . . . . . . . . . . . . . . 4-9
 

4.1.3. Aquatic Organisms......................................... 4-10
 

4.1.3.1. Fish ............................................. 4-10
 

4.1.3.2. Aquatic Invertebrates ............................... 4-10
 

4.1.3.3. Aquatic Plants .................................... 4-10
 

4.1.3.4. Aquatic Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-10
 

4.2. EXPOSURE ASSESSMENT ....................................... 4-12
 

4.2.1. Overview. ............................................... 4-12
 

4.2.2. Direct Exposure. .......................................... 4-12
 

4.2.3. Indirect Exposure. ......................................... 4-13
 

4.3. DOSE-RESPONSE ASSESSMENT .................................. 4-14
 

4.3.1. Overview. ............................................... 4-14
 

4.3.2. Terrestrial Plants .......................................... 4-14
 

4.3.3. Terrestrial Invertebrates .................................... 4-16
 

4.3.4. Other Species . . . . . . . .............................. . 4-17
 

4.4. RISK CHARACTERIZATION ...................................... 4-18
 

4.4.1. Overview. ............................................... 4-18
 

4.4.2. Direct Effects. ............................................ 4-18
 

4.4.2.1. Terrestrial Plants .................................. 4-18
 

4.4.2.2. Terrestrial Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-19
 

iii 
 

TABLE OF CONTENTS (continued) 

4.4.3. Indirect Effects............................................ 4-20
 

4.4.3.1. Terrestrial Mammals .... ..................... 4-20
 

4.4.3 2. Birds ............................................ 4-20
 

4.4.3.3. Amphibians and Reptiles . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-20
 

4.4.3.4. Terrestrial Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . 4-20
 

4.4.3.5. Aquatic Organisms ................................. 4-20
 

5. REFERENCES ........................................................... 5-1
 

iv 
LIST OF TABLES
 

Table 3-1: Individuals with skin responses to the gypsy moth in two communities . . . . . Tables-1 

Table 3-2: Statistical Analyses of Epidemiology Data from Table 3-1 . . . . . . . . . . . . . . . Tables-2 

Table 3-3: Adverse human health effects for members of the general public associated with 

exposure to the gypsy moth . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tables-3 

Table 4-1: Summary of quantitative risk characterization for forest stands . . . . . . . . . . . Tables-4 

LIST OF FIGURES 

Figure 2-1:Frequency of defoliation by the gypsy moth from 1975 to 2002 . . . . . . . . . . Figures-1 

Figure 4-1: Relationship of egg mass density to number of larvae per 50 g dry weight of 

vegetation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Figures-2 

Figure 4-2: Summary of exposure-response Assessment . . . . . . . . . . . . . . . . . . . . . . . . Figures-3 

NOTE: Tables followed by figures are placed after Section 5, References. 

v 
ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

AEL adverse-effect level 

bw body weight 

B.t.k. Bacillus thuringiensis var. kurstaki 

CI confidence interval 

cm centimeter 

d.f. degrees of freedom 

EC x concentration causing X% inhibition of a process 

EC25 concentration causing 25% inhibition of a process 

EC50 concentration causing 50% inhibition of a process 

F female 

FH Forest Health 

g gram 

ha hectare 

HQ hazard quotient 

kg kilogram 

L liter 

lb pound 

LdNPV Lymantria dispar (gypsy moth) nuclear polyhedrosis virus 

LOAEL lowest-observed-adverse-effect level 

LOC level of concern 

m meter 

M male 

mg milligram 

mg/kg/day milligrams of agent per kilogram of body weight per day 

mL milliliter 

NOAEL no-observed-adverse-effect level 

NOEC no-observed-effect concentration 

NOEL no-observed-effect level 

NOS not otherwise specified 

ppb parts per billion 

ppm parts per million 

ppt parts per trillion 

RfD reference dose 

SERA Syracuse Environmental Research Associates 

UF uncertainty factor 

U.S. United States 

USDA U.S. Department of Agriculture 

U.S. EPA U.S. Environmental Protection Agency 

: micron or microvi 
COMMON UNIT CONVERSIONS AND ABBREVIATIONS
 

To convert ... Into ... Multiply by ... 

acres hectares (ha) 0.4047 

acres square meters (m ) 2 4,047 

atmospheres millimeters of mercury 760 

centigrade Fahrenheit 1.8 °C+32 

centimeters inches 0.3937 

cubic meters (m ) 3 liters (L) 1,000 

Fahrenheit centigrade 0.556 °F-17.8 

feet per second (ft/sec) miles/hour (mi/hr) 0.6818 

gallons (gal) liters (L) 3.785 

gallons per acre (gal/acre) liters per hectare (L/ha) 9.34 

grams (g) ounces, (oz) 0.03527 

grams (g) pounds, (oz) 0.002205 

hectares (ha) acres 2.471 

inches (in) centimeters (cm) 2.540 

kilograms (kg) ounces, (oz) 35.274 

kilograms (kg) pounds, (lb) 2.2046 

kilograms per hectare (hg/ha) pounds per acre (lb/acre) 0.892 

kilometers (km) miles (mi) 0.6214 

liters (L) cubic centimeters (cm ) 3 1,000 

liters (L) gallons (gal) 0.2642 

liters (L) ounces, fluid (oz) 33.814 

miles (mi) kilometers (km) 1.609 

miles per hour (mi/hr) cm/sec 44.70 

milligrams (mg) ounces (oz) 0.000035 

meters (m) feet 3.281 

ounces (oz) grams (g) 28.3495 

ounces per acre (oz/acre) grams per hectare (g/ha) 70.1 

ounces per acre (oz/acre) kilograms per hectare (kg/ha) 0.0701 

ounces fluid cubic centimeters (cm ) 3 29.5735 

pounds (lb) grams (g) 453.6 

pounds (lb) kilograms (kg) 0.4536 

pounds per acre (lb/acre) kilograms per hectare (kg/ha) 1.121 

pounds per acre (lb/acre) mg/square meter (mg/m ) 2 112.1 

pounds per acre (lb/acre) :g/square centimeter (:g/cm ) 2 11.21 

pounds per gallon (lb/gal) grams per liter (g/L) 119.8 

square centimeters (cm ) 2 square inches (in ) 2 0.155 

square centimeters (cm ) 2 square meters (m ) 2 0.0001 

square meters (m ) 2 square centimeters (cm ) 2 10,000 

yards meters 0.9144 

Note: All references to pounds and ounces refer to avoirdupois weights unless otherwise specified. 

vii 
CONVERSION OF SCIENTIFIC NOTATION
 

Scientific Decimal Verbal 

Notation Equivalent Expression 

1 @ 10-10 0.0000000001 One in ten billion 

1 @ 10-9 0.000000001 One in one billion 

1 @ 10-8 0.00000001 One in one hundred million 

1 @ 10-7 0.0000001 One in ten million 

1 @ 10-6 0.000001 One in one million 

1 @ 10-5 0.00001 One in one hundred thousand 

1 @ 10-4 0.0001 One in ten thousand 

1 @ 10-3 0.001 One in one thousand 

1 @ 10-2 0.01 One in one hundred 

1 @ 10-1 0.1 One in ten 

1 @ 100 1 One 

1 @ 101 10 Ten 

1 @ 102 100 One hundred 

1 @ 103 1,000 One thousand 

1 @ 104 10,000 Ten thousand 

1 @ 105 100,000 One hundred thousand 

1 @ 106 1,000,000 One million 

1 @ 107 10,000,000 Ten million 

1 @ 108 100,000,000 One hundred million 

1 @ 109 1,000,000,000 One billion 

1 @ 1010 10,000,000,000 Ten billion 

viii 
EXECUTIVE SUMMARY
 

OVERVIEW 

The best documented and most obvious effect of the gypsy moth will be on terrestrial vegetation, 

particularly forest stands in which sensitive species of trees predominate. In sensitive forest 

stands, gypsy moth larvae can cause substantial defoliation. In forest stands in which tree species 

that are not favored by gypsy moth larvae predominate, even relatively high exposures may not 

result in substantial defoliation. 

The gypsy moth may also have a direct impact on human health and the most likely effects will 

involve skin irritation. In heavy gypsy moth infestations, adverse skin reactions would be 

expected in substantial numbers and the effects would likely be sufficiently severe to cause some 

individuals to seek medical attention. In extreme outbreaks, the effects will be qualitatively 

similar to those of severe infestations but could affect up to about one-third of the human 

population. 

Because the gypsy moth may substantially damage some forests in severe infestations or 

outbreaks, secondary effects in some species of wildlife are plausible and include reductions in 

populations of squirrels, mice, and other mammals which may be sensitive to changes in the 

availability of acorns are likely. Substantial adverse effects on other groups of animals – i.e., 

birds, reptiles, and aquatic species – cannot be ruled out but have not been convincingly 

demonstrated. 

GYPSY MOTH AS A PEST SPECIES 

The gypsy moth is a pest species that can cause substantial damage to some forests. In the 

eastern United States, most hardwood forests are classified as susceptible to gypsy moth 

infestation and as many as 12.5 million acres have been defoliated in a single season. The gypsy 

moth is found throughout much of New England and south to Virginia and west to portions of 

Wisconsin. The potential for substantial outbreaks is often assessed based on counts of 

overwintering egg masses, which are relatively easy to measure and can be made in time to plan 

for and take preventative measures against the outbreak. 

The life cycle of the gypsy moth consists of the egg, larval, pupal, and adult stages with one 

generation produced each year. The larvae or caterpillars go through various sub-stages, referred 

to as instars. First stage larvae (first instars) hatch in early to late May and go through additional 

larval stages between May and late June. First instars spin fine silk threads near the tops of trees 

from which they suspend themselves; in the event of sufficient wind, these threads break 

allowing the caterpillars to be transported by the wind. The distances involved in wind 

dispersion may cover several miles. Between late June and mid-July, the caterpillars spin sparse 

silken cradles and form pupae. After 7–14 days of pupation, the adult moths emerge and mate, 

and the life cycle is repeated. 

ix 
The gypsy moth is susceptible to diseases caused by gypsy moth pathogens like B.t.k., the gypsy 

moth nuclear polyhedrosis virus (LdNPV), and Entomophaga maimaiga fungi. B.t.k. and 

LdNPV are also used as control agents for the gypsy moth and these agents are addressed 

individually in separate risk assessments. The gypsy moth is a prey species for some mammals, 

birds, and other insects. In general, invertebrates are the major predators of gypsy moth larvae, 

while small mammals are the major predators of pupae. Numerous insects, including the larvae 

of various flies and wasps, act as parasites or predators to the gypsy moth. 

HUMAN HEALTH RISK ASSESSMENT 

Hazard Identification – Skin irritation after contact with larvae of many species of lepidoptera is 

common and this effect is the most common and best documented response to contact with gypsy 

moth larvae. The skin reactions seem to be associated with contact with small fine hairs that 

stick out from the body of the larva. The precise mechanism or mechanisms of action for these 

irritant effects is unclear but may involve three general responses: mechanical irritation, toxic 

reaction to a compound such as histamine, and an immediate or delayed allergic reaction. Raised 

and reddened areas of skin, known as wheals, are the most characteristic skin lesions. These 

lesions, resembling the raised patches of skin often associated with mosquito bites, may be 

approximately 0.25–0.5 inches in diameter. Contact with larvae may also cause rashes rather 

than wheals. Both wheals and rashes may cause severe itching that can persist for several days 

to 2 weeks and may be sufficiently severe to cause the affected individual to seek medical 

treatment. Other effects that may be associated with exposure to gypsy moth larvae include eye 

and respiratory irritation but these effects are less well documented, compared with dermal 

effects. 

In very severe infestations, the large numbers of larvae in an area may cause stress or anxiety in 

some individuals. Also during heavy infestations, water quality may be affected by increased 

runoff and by direct contamination with frass. Nonetheless, there are no documented cases of 

changes in water quality being associated with adverse effects in humans. 

Exposure Assessment – The number of larvae per unit area or tree might be considered the most 

direct and relevant measure of human exposure because it is contact with the larvae that causes 

skin irritation, the adverse effect typically associated with the gypsy moth. The available dose 

response data, however, are based on studies in which exposure is quantified as the number of 

eggs masses per acre and thus this is the exposure measure that is used in this risk assessment. 

As long as gypsy moth populations remain sparse, the larvae usually eat only a small proportion 

of the foliage of even their most favored host species, and contacts with people are rare. In such 

cases, egg masses generally do not exceed 50 egg masses/acre. During full-scale outbreaks, 

densities of about 5000 egg masses/acre are common and densities greater than 20,000 egg 

masses/acre are occasionally recorded. 

Dose-Response Assessment – The dose-response assessment for human health effects is based 

on reports of skin irritation in two populations: one with low exposure (an average of 32 egg 

masses/acre) and the other with high exposure (an average of 3809 egg masses/acre). The lowx 
exposure group exhibited no increase in skin irritation and 32 egg masses/acre is taken as a 

NOAEL (no adverse effect level) for humans and is used as a surrogate RfD (reference dose) for 

exposure to the gypsy moth in a manner analogous to the use of RfD values for control agents. 

The high exposure group did evidence a significant increase in skin irritation and, based on a 

dose-response model developed by U.S. EPA, egg mass densities up to 128 egg masses/acre are 

not likely to cause a detectable increase in skin irritation or rashes. In addition to these 

quantitative estimates, the severity of the response is important, particularly in a comparison of 

effects caused by exposure to the gypsy moth and effects caused by exposure to the agents used 

to control the gypsy moth. Dermal responses to the gypsy moth are sufficiently severe to have 

generated numerous case reports. While precise statistics are not available, it does appear that 

the severity of the skin irritation is sufficient to cause appreciable numbers of affected 

individuals to seek medical care. While exposure to the gypsy moth is associated with irritation 

to the eyes and respiratory tract, quantitative dose-response relationships for these endpoints 

cannot be developed. 

Risk Characterization – In sparse to moderate infestations—i.e., egg mass densities of <500 egg 

masses/acre—adverse effects involving skin irritation are not likely to be detectable in 

populations of exposed humans. Nonetheless, some individuals who come into contact with 

gypsy moth larvae could develop skin irritation. In heavy gypsy moth infestations—i.e., >500 to 

5000 egg masses/acre—adverse skin reactions would be expected in substantial numbers and the 

effects would likely be sufficiently severe to cause some individuals to seek medical attention. In 

extreme outbreaks—i.e., >5,000 to 20,000 egg masses/acre— the effects will be qualitatively 

similar to those of severe infestations but could affect up to about one-third of the population. 

Heavy infestations or extreme outbreaks could cause ocular and respiratory effects in some 

people but the likelihood of observing these effects cannot be quantified. Similarly, severe 

infestations are often considered to be a nuisance and cause aesthetic damage to the environment. 

Both of these factors can lead to stress in some individuals. Young children may be a group at 

special risk from effects of gypsy moth exposure but it is not clear whether children are more 

sensitive than adults to the effects of gypsy moth exposure or whether responses in children 

appear greater because children spend more time outdoors compared with adults. 

ECOLOGICAL RISK ASSESSMENT 

Hazard Identification – The clearest primary effect of gypsy moth infestations is on terrestrial 

plants, primarily trees. Various instars of the gypsy moth larvae will feed on host trees and can 

cause extensive defoliation which can kill some of the infested trees. On a larger scale, the 

extensive defoliation and/or death of trees may result in secondary changes to vegetation, which 

will, in turn, affect other forms of vegetation as well as various animal species (primarily related 

to changes in habitat). Gypsy moth larvae appear to have definite food preferences; oak, birch, 

poplar, and apple trees seem to be their favorite food sources. While both the European and 

Asian gypsy moth cause similar types of damage (i.e., defoliation), their feeding preferences are 

somewhat different with the Asian gypsy moth preferring a wider range of vegetation. Heavy 

defoliation is much more common among the oaks than among trees that are not particularly 

favored as food by the gypsy moth. For susceptible oaks, the effects of infestations on tree 

xi 
mortality varies according to the initial condition of the stand and the number of infestations. 

Generally, gypsy moth infestations result in mortality of less than 15% of total basal area – i.e., 

mortality of trees involving 15% the total area of the tree trunks near the ground. When heavy 

defoliation is followed by massive overstory mortality, existing shrub and herb cover increase 

dramatically due to increases in available light, moisture, and nutrients. Extensive loss of the 

existing canopy will also favor the growth of tree species that are intolerant to shade and will 

shift the forest ecosystem towards earlier successional stages. 

The only other groups of organisms that are likely to be directly affected by the gypsy moth are 

some and probably very few other lepidopteran species, including the northern tiger swallowtail 

butterfly. The mechanisms for direct adverse effects on other lepidopteran species may include 

bacterial contamination of the leaves by gypsy moth larvae and a decrease in the nutritional value 

of the leaves damaged by the gypsy moth. Most studies, however, do not indicate substantial 

direct effects on other insects, including lepidoptera. In some cases, increases may be seen in 

populations of insect predators of the gypsy moth. 

There is no indication in the literature that the gypsy moth will cause direct adverse effects in 

most groups of animals. Indirect effects, associated with damage to vegetation, may be of 

substantial consequence to some species, including squirrels, mice, and other mammals that rely 

on acorns. Although some mammals consume insects, including the gypsy moth, there is no 

evidence that gypsy moth outbreaks have a substantial impact on insectivorous mammals. 

Similarly, there is little indication that birds or aquatic species will be adversely affected by the 

gypsy moth. In some species of birds, gypsy moth infestations and subsequent defoliation may 

be beneficial, especially for species of birds that favor dead wood as a habitat. 

Exposure Assessment – As in the human health risk assessment, the exposure metameter is 

dictated by the data used to formulate the dose-response assessment. Also as in the human health 

risk assessment, egg mass density is the exposure metameter for terrestrial invertebrates and 

plants because it is the measure on which the dose-response assessment is based. Egg mass 

densities spanning a range from 5 egg masses/acre to 5,000 egg masses per acre are used to 

estimate responses in terrestrial plants and invertebrates. 

Most wildlife species are not affected directly by exposure to the gypsy moth but are more likely 

to experience indirect effects like changes in habitat or other environmental conditions secondary 

to defoliation. Consequently, the exposure assessment for most wildlife species is almost 

identical to the dose-response assessment for terrestrial plants which is expressed as defoliation 

caused by gypsy moth larvae. For this exposure assessment, categories of defoliation are defined 

normal background defoliation (<30% defoliation), moderate defoliation (30-60% defoliation), 

and high or severe defoliation (>60% defoliation). 

Dose-Response Assessment – As in the human health risk assessment, the dose metameter is egg 

masses/acre. Quantitative dose-response assessments can be made for both terrestrial plants and 

sensitive species of lepidoptera. The dose-response assessments for terrestrial plants are based 

xii 
 

on a relatively simple quantitative model for the relationship of egg mass density to defoliation. 

Three broad categories (sensitive, intermediate, and tolerant) are used to characterize the 

susceptibility of forest stands to gypsy moth induced defoliation. Estimated LOAEL values 

based on 30% defoliation, which is considered the lower range of moderate defoliation, are 

approximately 125, 1000, and 7000 egg masses/acre for sensitive, intermediate, and tolerant 

forest stands, respectively. The corresponding NOAEL values, defined as 10% defoliation, are 

estimated as 12, 20, and 125 egg masses/acre for sensitive, intermediate, and tolerant forest 

stands. 

The effects of gypsy moth exposure on sensitive terrestrial invertebrates, including some species 

of lepidoptera, are less well documented and less well characterized, compared with the effects 

on terrestrial plants. Nonetheless, available studies indicate that the NOAEL for adverse effects 

in certain other species of lepidoptera are lower than the NOAEL for sensitive forest stands—– 

i.e., about 6-72 egg masses/acre for some lepidoptera. 

No quantitative dose-response assessment is presented for other groups of organisms—e.g., 

mammals, birds, and soil or aquatic organisms. The impact of gypsy moth exposure on these 

species is most likely to result in indirect effects secondary to defoliation. 

Risk Characterization – The best documented and most obvious effect of the gypsy moth will be 

on terrestrial vegetation, particularly forest stands in which sensitive species of trees 

predominate. In some respects, the risk characterization for terrestrial vegetation is essentially a 

restatement of the hazard identification. In other words, the effects of gypsy moth larvae on 

forests is extremely well documented and relatively well understood. In sensitive forest 

stands—i.e., stands in which oak, birch, and other favored species predominate —gypsy moth 

larvae can cause substantial defoliation. In forest stands in which tree species that are not 

favored by gypsy moth larvae predominate —e.g., hemlock, various types of pine, black locust 

and white ash—even relatively high exposures, as measured by egg mass density, may not result 

in substantial defoliation. The risk assessment for direct effects on forests should be at least 

qualitatively influenced by the current range of the gypsy moth, which has not yet extended to 

some forests in the southeast that may be among the most sensitive to gypsy moth exposure. 

Thus, unless measures to contain the gypsy moth are successful, the southeastern oak forests may 

suffer serious damage in future infestations. 

Some other lepidopteran species also may be directly affected by exposure to the gypsy moth. 

Most studies, however, suggest that substantial adverse effects in terrestrial insects are unlikely 

and effects in some insect species, including some other lepidoptera, may be beneficial. 

Because the gypsy moth may substantially damage some forests in severe infestations or 

outbreaks, secondary effects in other species of wildlife are plausible. Reductions in populations 

of squirrels, mice, and other mammals which may be sensitive to changes in the availability of 

acorns are likely. Substantial adverse effects on other groups of animals – i.e., birds, reptiles, and 

aquatic species – cannot be ruled out but have not been convincingly demonstrated. 

xiii 
1. INTRODUCTION
 

This report addresses the potential human health effects and ecological effects of gypsy moth 

infestations and is part of the effort to update the 1995 Final Environmental Impact Statement 

(FEIS) for the Cooperative Gypsy Moth Management Program. The effort to update the FEIS 

involves the preparation of human health risk assessments (HHRAs) and ecological risk 

assessments (ERAs) for each of the agents used to control or eradicate gypsy moth infestations: 

Bacillus thuringiensis kurstaki (B.t.k.), Gypchek, diflubenzuron, tebufenozide, DDVP and 

disparlure. This risk assessment of the gypsy moth is intended to assist the USDA in assessing 

the consequences of “no action” alternatives in the FEIS. In addition, a separate document in this 

series will compare the effects gypsy moth infestations with the effects of the agents used to 

control the infestations. 

This documents consists of an introduction, an overview of the gypsy moth as a pest species 

(Section 2), a risk assessment for human health effects (Section 3), and a risk assessment for 

ecological effects or effects on non-target wildlife species (Section 4). Each of the two risk 

assessment chapters has four major sections, including an identification of the hazards associated 

with the gypsy moth, an assessment of potential exposure to the gypsy moth, an assessment of 

the dose-response relationships, and a characterization of the risks associated with plausible 

levels of exposure. These are the basic steps recommended by the National Research Council of 

the National Academy of Sciences (NRC 1983) for conducting and organizing risk assessments. 

The risk assessment on the gypsy moth is different from the risk assessments for chemical and 

biological agents used to control gypsy moth infestations, primarily because many standard 

physical and chemical properties used to characterize control agents and estimate certain 

exposure parameters are not at issue. Moreover, estimates of human and ecological exposure to 

all control agents—chemical and biological— are based on application rates, (i.e., known 

amounts of the agent applied under reasonably well defined conditions), which are not relevant to 

the gypsy moth. As discussed in subsequent sections of this document, estimates regarding 

gypsy moth exposure are extremely variable and difficult to define. 

A tremendous body of information is available on the biology, physiology, and population 

dynamics of the gypsy moth and this information is presented in reviews, books, and monographs 

that are available in the open literature (e.g., Davidson et al. 1999, 2001; Gansner et al. 1993a; 

Gerardi and Grimm 1979; Herrick and Gansner 1988; Liebhold 1992; Nealis et al. 1999; Sharov 

et al. 1999, 2002; Wallner 1994, 1996; Williams et al. 2000). Additional information on the 

gypsy moth is available at a USDA Forest Service web site, http://na.fs.fed.us/wv/gmdigest/. 

The current risk assessment makes no attempt to summarize all of this information. Although 

some background information is presented (Section 2), the primary focus of this document is on 

the information that can be used directly to assess the human health effects (Section 3) and 

ecological effects (Section 4) of the gypsy moth in ways that correspond to and may be compared 

to the risk assessments of agents used to control or eradicate gypsy moth infestations. 

1-1
 
This is a technical support document that addresses some specialized technical areas. 

Nevertheless, an effort was made to ensure that the document can be understood by individuals 

who do not have specialized training in the chemical and biological sciences. Certain technical 

concepts, methods, and terms common to most risk assessments are described in a separate 

document (SERA 2001). In addition, general glossaries of environmental terms are widely 

available and a custom glossary designed to be used in conjunction with USDA risk assessments 

is available at www.sera-inc.com. Some of the more complicated terms that are specific to the 

gypsy moth are defined in the text of this risk assessment. 

1-2
 
 

2. GYPSY MOTH AS PEST SPECIES
 

2.1. OVERVIEW 

The gypsy moth is a pest species that can cause substantial damage to some forests. In the 

eastern United States, most hardwood forests are classified as susceptible to gypsy moth 

infestation and as many as 12.5 million acres have been defoliated in a single season. The gypsy 

moth is found throughout much of New England and south to Virginia and west to portions of 

Wisconsin. The potential for substantial outbreaks is often assessed based on counts of 

overwintering egg masses, which are relatively easy to measure and can be made in time to plan 

for and take preventative measures against the outbreak. 

The life cycle of the gypsy moth consists of the egg, larval, pupal, and adult stages with one 

generation produced each year. The larvae or caterpillars go through various sub-stages, referred 

to as instars. First stage larvae (first instars) hatch in early to late May and go through additional 

larval stages between May and late June. First instars spin fine silk threads near the tops of trees 

from which they suspend themselves; in the event of sufficient wind, these threads break 

allowing the caterpillars to be transported for long distances by the wind. Between late June and 

mid-July, the caterpillars spin sparse silken cradles and form pupae. After 7–14 days of 

pupation, the adult moths emerge and mate, and the life cycle is repeated. 

The gypsy moth is susceptible to diseases caused by gypsy moth pathogens including B.t.k., the 

gypsy moth nuclear polyhedrosis virus (LdNPV), and Entomophaga maimaiga fungi. B.t.k. and 

LdNPV are also used as control agents for the gypsy moth and these agents are addressed 

individually in separate risk assessments. The gypsy moth is a prey species for some mammals, 

birds, and other insects. In general, invertebrates are the major predators of gypsy moth larvae, 

while small mammals are the major predators of pupae. Numerous insects, including the larvae 

of various flies and wasps, act as parasites or predators to the gypsy moth. 

2.2. INFESTATIONS 

The current scientific name is for the gypsy moth is Lymantria dispar. In the older literature 

(e.g., Gerardi and Grimm 1979), the gypsy moth is referred to by its previous scientific name, 

Porthetria dispar. Over three quarters of the hardwood forests in the eastern United States are 

classified as susceptible to the gypsy moth (USDA/FS 1990). In addition, many forests in the 

south and central regions of the country, currently beyond the range of the gypsy moth, are likely 

to be very susceptible to damage by the gypsy moth (Liebhold and McManus 1999). In a major 

outbreak, the extent of damage can be substantial and as many as 12.5 million acres have been 

defoliated in a single season (Williams 1982). Damage to vegetation is caused by feeding larvae. 

During outbreaks, gypsy moth larval populations may range from about 10,000 to 250,000 larvae 

per hectare (Colbert et al. 1995; Christie et al. 1995). 

The gypsy moth was brought into the United States intentionally in 1869 as part of an experiment 

by a naturalist, Leopald Trouvelet, to develop a hardy silk-producing insect. In the course of the 

experiments, conducted in Medford, Massachusetts, some gypsy moth eggs were lost and a 

2-1
 
population of gypsy moths was established in the Medford area. The gypsy moth population 

grew to infest about a 400 square mile area around Medford by 1880, and the first major outbreak 

occurred in 1889 (Gerardi and Grimm 1979). The gypsy moth has spread throughout much of 

New England and south to Virginia and east to portions of Wisconsin. Current and plausible 

future infestations are discussed further in the exposure assessment for human health effects 

(Section 3.2) and ecological effects (Section 4.2). Figure 2-1 summarizes information regarding 

the frequency of gypsy moth defoliations over a period of 28 years—i.e., 1975 to 2002. In any 

given year, marked defoliations associated with gypsy moth infestations may be less ubiquitous 

and may be isolated in relatively small areas, which is due both to the control measures taken to 

limit gypsy moth populations as well as to the natural variability in gypsy moth populations. 

The population pattern observed after the release of the first gypsy moths in North America—i.e., 

a period of low and inconsequential population growth followed by a major outbreak—is typical 

of gypsy moth population dynamics, which are described as bimodal (i.e., existing either at 

innocuous densities or in an outbreak or very rapid growth mode) (Campbell 1981). Following 

an initial outbreak, populations generally decline and are usually maintained at low population 

densities that cause little damage. Subsequent outbreaks are usually less severe than the initial 

outbreak. As discussed further in Section 4.2, gypsy moth outbreaks are often associated with 

the presence of favored tree species (Baker and Cline 1936; Behre 1939; Behre and Reineke 

1943). In general, gypsy moth outbreaks in North America dissipate in 1or 2 years. In rare cases, 

outbreaks can recur annually over periods of up to 20 years (Bess et al. 1947; Campbell 1973). 

For at least half a century, the gypsy moth has persisted at generally innocuous densities in the 

predominantly oak forests of northeastern Connecticut and adjacent Massachusetts (Bess et al. 

1947; Brown and Sheals 1944). During such intervals, gypsy moth larvae usually eat only a 

small proportion of the foliage of even their most favored host species. When defoliation is low, 

nearly all of it occurs on favored-food trees (Campbell and Sloan 1977b). Once a large-scale 

outbreak is underway, the gypsy moth will feed on a greater variety of vegetation and over 300 

species of broadleaf and coniferous trees and shrubs may be damaged (Leonard 1981; Liebhold et 

al. 1994). 

The potential for substantial outbreaks is often assessed based on counts of overwintering egg 

masses. Such counts are relatively easy to make and can be made in time to plan for and take 

preventative measures against a potential outbreak (Buss et al. 1999). Egg mass counts, 

however, are not absolute predictors of outbreak potential and egg masses per acre can be highly 

variable. In an infested area in Maryland, egg masses ranged from about 20/ha to 14,000/ha at 16 

sites over a 4-year period (Davidson et al. 2001). In heavily infested areas, eggs masses per acre 

can range from about 5000 to 43,000 (Hajek 1997). The relationship of egg mass density to 

subsequent damage is complicated by the fact that the survival of egg masses is also highly 

variable, ranging from <1% to about 90% (Nealis et al. 1999). In areas with extremely cold 

winters, egg masses laid below the snow line tend to have higher survival rates than those laid 

above the snow line (Nealis et al. 1999; Smitley et al. 1998). Bess (1961) reports that reduced 

2-2
 
defoliation, which followed a winter of prolonged below zero temperatures, was due to 90% 

overwintering egg mortality. 

The spread of an infestation may also be influenced by available vegetation. While some studies 

show that the quality of resources available to the female gypsy moth has only a minor effect on 

population dynamics (Erelli and Elkinton 2000), other sources indicate that the consumption of 

vegetation by larvae and subsequent larval growth may differ substantially according to 

vegetation type (Foss and Rieske 2003). 

The spatial distribution of stand susceptibility is a key characteristic in the spread of outbreaks 

and subsequent defoliation (Liebhold and McManus 1991). Outbreaks are described as 

originating in small, discrete locations. These locations, referred to as foci, are usually 

characterized by stands growing on stressed sites like ridge tops, upper slopes, and deep sands, 

frequently subject to drought (Houston and Valentine 1977). These areas can support moderate 

to high populations of gypsy moth when the insect is undetectable in surrounding areas (Liebhold 

and McManus 1991). Protected resting locations that favor larval and pupal survival are known 

to support larger gypsy moth populations and lead to outbreaks (Bess et al. 1947; Campbell and 

Sloan 1977a; Houston 1975; Houston and Valentine 1977). 

Other factors that may precipitate outbreaks include predator failure and specific climatic and 

meteorological conditions. Khanislamov and Girfanova (1964) demonstrate that weather 

variation may have more drastic effects on the natural enemies of gypsy moth than on the pest 

itself. Population collapse at the end of an outbreak appears to be the result of disease (Section 

2.4), reduced fecundity, and starvation (Campbell 1981). Although dispersal of young larvae 

plays a role in gypsy moth outbreaks, it is thought to play a relatively minor role in outbreak 

initiation. Larval dispersion may be the major cause of gypsy moth distribution enlargement and 

range expansion at innocuous densities but it does not appear to cause outbreaks to spread 

(Campbell 1976). The rate at which infestations spread may vary substantially according to 

vegetation type and the methods used to control the spread. Reported rates of infestation range 

from 12 to 145 km/year ( Sharov et al. 1999; Wallner 1996). 

There are various models to predict the effects of gypsy moth infestations on a mixed hardwood 

forest (Colbert and Racin 1995; Colbert et al. 1995; Weseloh 1996a,b; Wilder et al. 1995; 

Williams et al. 1997). These models are discussed further in the dose-response assessment for 

terrestrial vegetation (Section 4.3.2). 

2.3. LIFE-CYCLE 

As with most insects, the life cycle of the gypsy moth consists of the egg, larval, pupal, and adult 

stages (Abrahamson and Klass 1982; Cram 1990; Gerardi and Grimm 1979). In the northeast, 

the adult female lays eggs in July or August. The larvae or caterpillars go through various stages, 

referred to as instars. First stage larvae (first instars) hatch in early to late May and go through 

additional larval stages (a total of five instars in males and six instars in females) between May 

and late June. This process occurs somewhat earlier in the southeast. The transition from each 

2-3
 
stage to the next involves molting, during which time the caterpillar sheds its outer skin. It is 

during the larval stages that feeding occurs. First instars spin fine silk threads near the tops of 

trees from which they suspend themselves. After the thread breaks, the larvae can be transported 

over relatively long distances by the wind. The distances that larvae might be carried by wind is 

likely to be highly variable and has not been well or generally characterized. Gerardi and Grimm 

(1979, p. 63) note that larvae have been monitored at elevations of up to 2000 feet and have been 

found at distances of up to 35 miles from the closest known infestation. Between late June and 

mid-July, the caterpillars spin sparse silken cradles and form pupae. After 7–14 days of 

pupation, the adult moths emerge and mate, and the life cycle is repeated. 

Newly hatched larvae often remain on the egg mass for several days before climbing toward 

foliage. First-instar gypsy moth larvae have two types of hairs or setae: long thin hairs that 

appear to assist the larvae in "soaring" or transport by wind and short hairs (sometimes referred 

to as "balloon hair") that contain chemicals like nicotine which may serve as a defense 

mechanism to discourage predators (Bardwell and Averill 1996; Deml and Dettner 1995; Smith 

1985). 

Moths begin to emerge about the middle of July, with males appearing several days earlier than 

females. In the south, moth emergence may occur as early as June. The European female cannot 

fly; she emits a pheromone (sex attractant) that volatilizes and is carried in the air. Male moths 

are attracted to the pheromone for distances up to 1 mile. After fertilizing and depositing eggs 

the adult moths do not eat and soon die (Johnson and Lyon 1988). Egg masses are deposited on 

tree trunks, rocks, and litter. Although the eggs overwinter, below normal temperatures can 

cause egg mortality (Bess 1961). 

This risk assessment considers both the European and Asian gypsy moths, which are considered 

to be the same species (Lymantria dispar). Since the European gypsy moth was introduced in 

North America from closely related individuals, genetic studies indicate little variation within or 

between populations. The Asian gypsy moth, on the other hand, displays considerable variability 

within populations. The variability is expressed morphologically in the variety of larval color 

forms, behaviorally in the female flight capability, and physiologically in the capacity of larvae to 

colonize aggressively a broad spectrum of hosts (USDA/FS 1992). 

While the female European gypsy moth is flightless, the Asian female is a strong flier capable of 

flights in excess of 18 miles (30 km) (USDA/FS 1992; Wallner 1996). Since the female Asian 

gypsy moth is able to lay eggs far from the pupal site following flight, this characteristic alone 

may make it necessary to modify the control methods of detection, delimitation, and control or 

eradication developed for the European gypsy moth (USDA/FS 1992). Asian gypsy moth larvae 

tend to feed more aggressively and on a broader variety of trees than their European counterpart 

(Wallner 1996). 

2-4
 
2.4. DISEASE AND PREDATION 

The primary focus of this risk assessment is effects of the gypsy moth on other species. 

Nonetheless, many organisms may adversely affect the gypsy moth, thereby reducing the risks 

posed by gypsy moth infestations. The gypsy moth is susceptible to diseases, including diseases 

caused by pathogens like B.t.k., the gypsy moth nuclear polyhedrosis virus (LdNPV), and 

Entomophaga maimaiga fungi. Bacterial pathogens in addition to B.t.k. and other Bacillus 

species that adversely affect the gypsy moth, include Serratia marcescens, Serratia liquefaciens, 

Streptococcus, and Pseudomonas spp. These microorganisms are associated with a collective 

mortality in the gypsy moth of less than or equal to 15% (Podgwaite 1981). The gypsy moth 

nucleopolyhedrosis virus (LdNPV) is a natural component of the gypsy moth environment 

(Podgwaite 1979; 1981; Podgwaite and Campbell 1970; Lindroth et al. 1999) and is considered 

the primary natural regulator of dense gypsy moth populations in North American forests (Glaser 

and Chapman 1913; Doane 1970). High density populations of gypsy moth will eventually 

collapse, for the most part due to pathogens, especially NPV (Elkinton and Liebhold 1990). 

B.t.k. and LdNPV are also control agents for the gypsy moth, and are addressed individually in 

separate risk assessments. 

In addition to viral and bacterial pathogens, several fungal pathogens will infect gypsy moth 

populations, including species of Paecilomyces, Fusarium and Verticillium (Hajek 1997). Most 

fungal pathogens, however, appear to account for insignificant levels of recorded gypsy moth 

mortality (Podgwaite 1981). A major exception, however, is Entomophaga maimaiga, which 

plays an important role in gypsy moth population dynamics on other continents and which is 

widely established in North America. Nealis et al. (1999) estimate that E. maimaiga may 

account for approximately 4-14% of mortality in gypsy moth larvae. Infections with E. 

maimaiga tend to be more prevalent than naturally occurring infections from NPV in areas with 

low egg mass density (Buss et al. 1999). In low density plots, E. maimaiga increased mortality 

substantially only in 5th instar and later instars. In high density plots, earlier instars were also 

infected (Hajek 1997; Hajek et al. 2001). Models for the influence of E. maimaiga on gypsy 

moth populations have been developed by Weseloh (1998a, 1999, 2002, 2003). 

The gypsy moth is at risk of significant predation by mammals, birds, and other insects. In 

general, invertebrates are the major predators of gypsy moth larvae, while small mammals are the 

major predators of pupae (Grushecky et al. 1998). Mice and shrews are important predators of 

gypsy moth, particularly during the pupal stage (Bess et al 1947; Jones et al. 1998) or when the 

population density of the gypsy moth is low (Elkinton et al. 1996, 2002). When the population 

density of small mammals is high, small mammals may be a major source of predation on larvae 

(Cook et al. 1995). When populations of small mammals are low, the relative importance of 

predation by terrestrial invertebrates increases (Hastings et al. 2002a,b). 

Forbush and Fernald (1896) first identified birds as predators of gypsy moth larvae. Some 

species of birds even prey on egg masses (Cooper and Smith 1995). In general, however, 

mammals appear to have a greater impact on gypsy moth populations than birds (Smith and 

Lautenschlager 1981; Elkinton and Liebhold 1990). 

2-5
 
Numerous insects act as parasites or predators to the gypsy moth, including the larvae of various 

tachinid flies and braconid wasps (Hajek 1997). Extensive efforts were made to introduce 

European and Asian gypsy moth parasitoids to North America (parasitoids are insects, especially 

flies and wasps, that complete their larval development inside the body of another insect). Ten 

species have become established (Elkinton and Liebhold 1990). Gypsy moth mortality due to 

each type of parasite is specific to a given gypsy moth life stage. The venom of the ectoparasitic 

wasp Microbracon hebetor, contains a toxin that inhibits larval growth in gypsy moth (Masler 

and Kovaleva 1999). The food preference of certain wasps species—i.e., chalcids—seems to 

depend on the sex of the pupae (Fuester and Taylor 1996). Although invertebrate predation of 

gypsy moth pupae may be minor compared with vertebrate predation (Campbell and Sloan 

1977a), Smith and Lautenschlager (1981) suggest that mortality attributed to vertebrates may be 

caused by invertebrates, like ground beetles (Elkinton and Liebhold 1990). Both adult and 

immature stages of Calosoma sycophanta, a large ground beetle introduced from Europe, are 

known to feed on gypsy moth larvae and pupae (Elkinton and Liebhold 1990). In addition, 

Weseloh (1996b, 1998b) suggest that predation by ants, particularly on gypsy moth larvae that 

fall to the forest floor, could cause significant mortality to gypsy moth larvae. 

2-6
 
3. HUMAN RISK ASSESSMENT
 

3.1. HAZARD IDENTIFICATION 

3.1.1. Overview 

Skin irritation after contact with larvae of many species of lepidoptera is common and this effect 

is the most common and best documented response to contact with gypsy moth larvae. The skin 

reactions seem to be associated with contact with small fine hairs that stick out from the body of 

the larva. The precise mechanism or mechanisms of action for these irritant effects is unclear but 

may involve three general responses: mechanical irritation, toxic reaction to a compound such as 

histamine, and an immediate or delayed allergic reaction. Raised and reddened areas of skin, 

known as wheals, are the most characteristic skin lesions. These lesions, resembling the raised 

patches of skin often associated with mosquito bites, may be approximately 0.25–0.5 inches in 

diameter. Contact with larvae may also cause rashes rather than wheals. Both wheals and rashes 

may cause severe itching that can persist for several days to 2 weeks and may be sufficiently 

severe to cause the affected individual to seek medical treatment. Other effects that may be 

associated with exposure to gypsy moth larvae include eye and respiratory irritation but these 

effects are less well documented, compared with dermal effects. 

In very severe infestations, the large numbers of larvae in an area may cause stress or anxiety in 

some individuals. Also during heavy infestations, water quality may be affected by increased 

runoff and by direct contamination with frass. Nonetheless, there are no documented cases of 

changes in water quality being associated with adverse effects in humans. 

3.1.2. Mechanisms of Action 

As discussed in Section 3.1.3, dermal irritation is the most common adverse effect associated 

with human exposure to gypsy moth larvae. Dermal reactions to contact with lepidopteran larvae 

are in general relatively common (Anonymous 1984; Gilmer 1925; Goldman et al. 1960; Hellier 

and Warin 1967; Katzenellenbogen 1955; Perlman 1965; Schmidt 1982; Wirtz 1980,1984). 

Moreover, the gypsy moth is the most common insect associated with allergies—i.e., 28.7% of 

known cases (Wirtz 1980). 

The skin reactions seem to be associated with contact with the larval setae, small fine hair-like 

protrusions from the body of the larvae (Allen et al. 1991). The precise mechanism or 

mechanisms of action for these irritant effects is unclear but may involve three general responses: 

mechanical irritation, toxic reaction to a compound such as histamine, and an immediate or 

delayed allergic reaction (Burnett et al. 1989, Shama et al. 1982). Gypsy moth larvae have four 

kinds of setae, two of which are hollow and attached to glandular cells. The hollow setae are 

suspect, but not unequivocally identified as the setae associated with skin reactions in humans 

(Anderson and Furniss 1983). According to several case reports and epidemiology studies, 

dermal effects in humans are usually associated with exposure to the first instars (Anderson and 

Furniss;1983, Tuthill et al. 1984). Whole first instars and the setae of fifth instars contain 

histamine (Shama et al. 1982), a compound that causes wheals, which are characteristic of 

dermal contact with gypsy moth larvae (Sullivan 1982). 

3-1
 
The study by Beaucher and Farnham (1982) supports the association between gypsy moth 

exposure and allergic responses. In the study, closed patch tests were conducted on 8 individuals 

who had a history of skin reactions to the gypsy moth and 11 individuals, with no such history, 

who served as controls. A positive response to the patch test was observed in each of the 

individuals who had a history of skin reactions to the gypsy moth and in only one individual in 

the control group. The observed response was consistent with the reported dermal effects of 

gypsy moth exposure. In some cases, severe itching (pruritis) kept individuals awake at night. In 

general, the time from exposure to the onset of the reaction was 24–48 hours, suggesting a 

delayed hypersensitivity similar to poison ivy reactions. In another study, 10 of 17 workers at a 

laboratory conducting research on the gypsy moth reported a history of adverse skin or 

respiratory reactions. According to the results of scratch tests, 7 of the 10 workers who reported 

a history of adverse reactions were allergic to gypsy moth parts or other gypsy moth substances. 

The intensity of the response, based on a categorical classification of skin responses, was greater 

for extracts of cast larval skins and whole larvae than for egg mass hairs (Etkind et al. 1982). 

3.1.3. Effects on Skin 

Reports of dermal responses to contact with gypsy moth larvae began with the introduction of the 

moth to the United States. A late 19th century document describes a situation in which an 

individual in Medford Massachusetts "... was poisoned by them [gypsy moths]. While killing 

them upon the trees they would get upon his neck and blister and poison it" (Forbush and Fernald 

1896, p. 16). A few years later, a physician in Boston reported a number of cases of 

"...inflammation of the skin, which were undoubtedly caused by contact with some caterpillar ... 

which must be some recently introduced species" (White 1901 p. 599). Although Dr. White 

attributed these cases to the brown-tailed moth (Euproctis chrysorrhea), they are consistent with 

the reported effects of exposure to the gypsy moth which had escaped into the area near Boston 

some years before (Section 2.2). The literature contains no further mention of human health 

effects associated with the gypsy moth for almost a century. 

In the early 1980s, there was a massive gypsy moth infestation in the northeastern part of the 

United States. In 1981, outbreaks of itchy skin rashes that coincided with the heavy infestations 

were widespread and a source of public annoyance (Marshall 1981). Coincident with this 

infestation, reports describing the human health effects associated with exposure to the gypsy 

moth appeared in the medical literature. 

Wheals, raised and reddened areas of skin, are the most characteristic skin lesions associated 

with human contact with the larvae. These lesions, resembling the raised patches of skin often 

associated with mosquito bites, may be approximately 0.25–0.5 inches in diameter and are 

surrounded by an area of redness. In severe cases, the wheals may be so numerous that they 

overlap on large areas of the skin, a condition referred to as urticaria. Contact with the larvae is 

reported also to cause contact dermatitis, characterized by a rash rather than wheals (Anderson 

and Furniss 1983). Both rashes and wheals may cause severe itching, also known as pruritis. 

This effect can persist for several days to 2 weeks and may be sufficiently severe to cause the the 

3-2
 
affected individual to seek medical treatment (Aber et al. 1982; Allen et al. 1991; Shama et al. 

1982). 

During the severe infestations in the early 1980s, there were three published reports regarding the 

development of skin reactions in school children (Aber et al. 1982; Anderson and Furniss 1983; 

Tuthill et al. 1984). In the spring of 1982, a telephone survey was conducted to collect 

information from approximately 1000 people (representing more than 90% of those selected for 

study) in one highly infested community (HI, Lunenburg) and one minimally infested community 

(LO, Medway) in Massachusetts (Tuthill et al. 1984). The risk of developing a dermal response 

over a 1-week period was 10.4% in the HI community and 1.6% in the LO community. The 

responses occurred most often in individuals who had developed rashes during the previous year 

or who had direct contact with the larvae (that is, larvae crawled on them). The combination of 

these two factors resulted in an additive increase in risk. Other variables related to increased 

response included a history of hay fever and the practice of hanging clothes outdoors to dry. The 

rates at which the dermal responses developed in individuals in the HI community were inversely 

associated with age (18.8% in 0- to 12-year olds, 10.2% in 13- to 59-year olds, and 2.1% in 60­

year olds and older individuals). The average prevalence of dermal responses in both 

communities combined, 1 week before the emergence of the first instars, was 1.3% (Tuthill et al. 

1984). 

Sometime between the end of April and the third week of May, 1981, there was an increased 

incidence of rashes among students in two schools in Northeast Pennsylvania (Aber et al. 1982). 

School A had a response rate of 42.2% (135 of 320 students), and school B had a response rate of 

25.3% (76 of 300 students). The dermal responses included pruritic rash and occasional 

urticaria, usually located on exposed areas of the body. Based on the results of a survey of 

students from the same schools who were not affected by the gypsy moth, the investigators 

determined that there was a statistical association between touching larvae (p<0.01), working in a 

garden (p<0.05), or going fishing (p<0.01) and the incidence of rashes. 

Concurrent with the infestation in Pennsylvania was an infestation in Connecticut, associated 

with an outbreak of skin reactions in students at several schools within the community (Anderson 

and Furniss 1983). Urticaria was observed in 7.2% of the 2600 students attending four schools in 

Newton, Connecticut. More than 50% of the cases of urticaria occurred during the first week in 

May, coinciding with the emergence of first instars. Very few cases (approximately 10) occurred 

during the third week of May when the larvae were predominantly in the third instar stage. In 

Burlington, Connecticut, the incidence of skin reactions was approximately 5.1% (96 of 1870). 

In another school, about 7.1% (75 of 1058) of students were affected. In Bristol, Connecticut, 

there were 1348 cases of rashes in the public schools, amounting to approximately 10.7% of the 

total student enrollment (12,500). Health officials estimated that the true prevalence may have 

been 3 times higher than reported; however, details supporting this assessment were not 

provided. Nonetheless, the estimate is consistent with the occurrence of rashes in 12 of 25 

children attending a nursery school in the same community. 

3-3
 
3.1.4. Effects on Eyes and Respiratory Tract 

The ocular and respiratory effects in humans after exposure to the gypsy moth or other 

lepidopteran larvae are less well documented, compared with dermal effects. Of the 10 workers 

with a history of adverse reactions to the gypsy moth (Etkind et al. 1982), all 10 had skin 

reactions, 4 had eye irritation, and 2 had respiratory reactions. In a survey of laboratories 

conducting research on insects, 28.7% of all reported allergies were attributed to the gypsy moth. 

The most frequent reactions among affected individuals were skin irritation (61%), sneezing or 

runny nose (67%), and eye irritation (60.9%). Labored respiration was observed in 33% of the 

affected individuals (Wirtz 1980). The frequencies of these reactions are for all individuals who 

had adverse health effects after exposure to insects in general, not just the gypsy moth. In the 

early 1980s, NIOSH conducted a survey of workers in USDA/ARS research facilities who were 

involved in rearing insects for various research projects. As in the study by Wirtz (1980), the 

most common respiratory or ocular symptoms included sneezing/runny nose (73%), ocular 

irritation (68%), cough (38%), wheezing (26%), and shortness of breath (24%) (Anonymous 

1984). An update of this survey was planned for the early 1990 (Petsonk 1994) but no such 

publication was found in the literature. 

The severity of ocular or respiratory effects in humans after exposure to the gypsy moth is not 

well characterized; however, these effects appear to be reversible. Although some respiratory 

effects may involve pain, there are no data to indicate that the respiratory effects are life 

threatening or require hospitalization (Perlman 1965; Shama et al. 1982). 

3.1.5. Other Potential Effects 

The stress or anxiety associated with gypsy moth infestations is difficult to assess. This stress 

has not been associated with frank health effects. In many communities, the stress may be 

exacerbated by disputes about appropriate approaches for dealing with the pest (Williams 1982). 

Anecdotal reports suggest that some people may be extremely anxious about infestations 

(National Gypsy Moth Management Group 1991, p. 3): 

... the mere mention of insects sends some people into fits of 

scratching, but phobia was not an adequate explanation for the 

epidemic of runny noses, irritated eyes, and rashes that happened 

to coincide with the occurrence of gypsy moth caterpillars last 

spring [1990]. Every [Pennsylvania] county and state gypsy moth 

office received numerous calls and one agency was reported to 

have received over 2,700. 

Moreover, reports regarding the willingness of populations to pay for gypsy moth control (Miller 

and Lindsay 1993a,b) suggest that gypsy moth infestations are regarded as highly undesirable by 

the general public, both in terms of aesthetic damage and the potential for adverse effects on 

human health. Among 629 individuals residing in infested areas, the most frequent reasons for a 

willingness to pay for control measures against the gypsy moth were aesthetic damage (15%) and 

3-4
 
the nuisance factor (13%) (Miller and Lindsay 1993b). Concerns regarding adverse health effects 

directly related to exposure were expressed by 4% of the responders. 

In most instances, gypsy moth defoliation will have little effect on adjacent water bodies (Corbett 

and Lynch 1987; Grace 1986). During heavy infestations, water quality may be affected by 

increased runoff and by direct contamination with frass. During active defoliation, fecal 

streptococci levels in stream water were as high as 25,000/100 mL and fecal coliform densities 

exceeded 90/100 mL (Corbett 1991). Long term studies of the impact of gypsy moth defoliation 

on water quality have included studies that show stream water chemical concentrations following 

defoliation that have included increasing amounts of strong acid anions, base cations and 

hydrogen ions, as well as decreasing concentrations of acid neutralization capacity (Webb et al. 

1995). In addition dissolved nitrogen as nitrate will increase in streams following gypsy moth 

defoliation (Eshleman et al. 1998). 

There are neither studies that directly address the contamination of water with frass nor reports in 

the literature of adverse effects on human health associated with water contamination from frass. 

In gypsy moth defoliated forests, however, frass output reached 756 kg (dry weight)/ha in a 1 

month period (Grace 1986). 

Lyme disease, which is a bacterial infection induced by Borrelia burgdorferi, causes serious 

health effects in humans. In the northeastern and central United States, the primary vector is the 

black-legged tick, Ixodes scapularis (CDC 2004). The tick can infect the white-footed mouse, 

Peromyscus leucopus, and ticks from the white-footed mouse can infect deer or humans (Ostfeld 

2002; Ostfeld et al. 1996). As discussed in Section 4.1.2.1, the white-footed mouse eats acorns 

produced by oak trees. Gypsy moth infestations or outbreaks may result in decreases in acorn 

production due to damage to oak trees, which, in turn, may cause decreases in the population of 

white-footed mice due to decreases in food abundance (Elkinton et al. 1996, 2002). There is 

speculation that this decrease in the population of mice may limit the transmission of Lyme 

disease to humans due to the adverse effect on the primary vector (Jones et al. 1998; Randolph 

1998). Currently, however, there does not appear to be sufficient information to assess the 

plausibility of this supposition. 

3-5
 
 

3.2. EXPOSURE ASSESSMENT 

3.2.1. Overview. 

The number of larvae per unit area or tree might be considered the most direct and relevant 

measure of human exposure because it is contact with the larvae that causes skin irritation, the 

adverse effect typically associated with the gypsy moth. The available dose response data, 

however, are based on studies in which exposure is quantified as the number of eggs masses per 

acre and thus this is the exposure measure that is used in this risk assessment. As long as gypsy 

moth populations remain sparse, the larvae usually eat only a small proportion of the foliage of 

even their most favored host species, and contacts with people are rare. In such cases, egg 

masses generally do not exceed 50 egg masses/acre. During full-scale outbreaks, densities of 

about 5000 egg masses/acre are common and densities greater than 20,000 egg masses/acre are 

occasionally recorded. 

3.2.2. Exposure Metameter. 

Gypsy moth populations can be monitored by estimating the numbers of egg masses (typically 

expressed as egg masses per acre), the number of larvae (which can be expressed as larvae or 

larval mass per unit area or larvae per tree), or the number of adults per unit area. For adult 

moths, population surveys usually involve the use of pheromone traps with or without an 

insecticide. Surveys of larval populations may involve band trapping, direct examination, or 

correlations between frass volume and population density. Measurements of larval populations 

can be highly variable over time and among different species of trees. For example, Naidoo and 

Lechowicz (2001) conducted larval counts on different species of trees in a deciduous forest in 

Quebec. In preferred tree species (i.e., red oak), larval populations were as high as 250 larvae 

per tree. In less preferred tree species, larval populations were much lower, ranging from about 4 

larvae per tree (white ash) to 10 larvae per tree (sugar maple). In terms of larval mass, values of 

8.4 kg/ha in the month of June and 16 kg/ha in the month of July were measured during severe 

infestations (Grace 1986). 

While the number of larvae per unit area or tree might be considered the most direct and relevant 

measure of human exposure, the available dose response data (Section 3.3) are based on studies 

in which exposure is quantified as the number of eggs masses per acre (i.e., Tuthill et al. 1984; 

O'Dell 1994). As long as gypsy moth populations remain sparse, the larvae usually eat only a 

small proportion of the foliage of even their most favored host species, and contacts with people 

are rare. In such cases, egg mass densities generally do not exceed 50 egg masses/acre. 

For several years, gypsy moth populations may exist in a density range high enough (between 50 

and 500 egg masses/acre) to make the insect a minor nuisance in wooded communities and cause 

partial defoliation. Once, however, the gypsy moth population increases to a full-scale outbreak, 

the combination of insect frass and leaf fragments, loss of shade at midsummer, and the large 

number of larvae may become a major nuisance (Williams 1982). Although the duration of such 

outbreaks is unpredictable, the principal factors that influence the pest include a variety of 

pathogens, intraspecific competition for food, and inclement weather (Campbell 1981; Podgwaite 

3-6
 
1981; Miller et al. 1989). During full-scale outbreaks, densities of about 5000 egg masses/acre 

are common and densities greater than 20,000 egg masses/acre are occasionally recorded. 

Egg mass densities in infested areas tend to be lower in areas where the human population is 

dense, compared with less densely populated areas. At the forest periphery, however, egg mass 

densities can be much higher and seem to be associated with man-made objects (Campbell et al. 

1976). Within a relatively limited geographical range, egg mass densities may vary remarkably. 

For instance, in a heavily infested area with a mean egg mass density of approximately 3800 egg 

masses/acre, egg mass counts ranged from 0 egg masses/0.1 acre surveyed to 1000 egg 

masses/0.1 acre surveyed (O'Dell 1994). Similar variability in egg mass density were observed in 

larger survey areas, as well (Reardon et al. 1993). During a heavy infestation, as many as 50,000 

larvae may inhabit a single tree. At such extremely dense concentrations, the generation of frass 

may be sufficiently intense to be audible, sounding like a light rain (Beaucher and Farnham 

1982). 

3.2.3. Intensity of Exposures 

Given the localized variability in larval populations, quantitative estimates of exposure to larvae 

cannot be made. Epidemiology studies conducted in gypsy moth infested communities suggest 

that larval density as a measure of the intensity may not be meaningful. The most important 

factor in assessing exposure may be the probability of coming into contact with one or more 

larvae, rather than the number of larvae in a population. In this respect, patterns of human 

behavior, such as the amount of time spent outdoors and certain kinds of activities likely to result 

in contact with larvae may be more important than measurements of the local larval population. 

The likelihood of human exposure to the gypsy moth is likely to increase in proportion to the 

increases in the larval population in a given area; however, it is not possible to estimate more 

precise relationships of larval population density to human exposure. 

3-7
 
3.3. DOSE-RESPONSE ASSESSMENT 

3.3.1. Overview 

The dose-response assessment for human health effects is based on reports of skin irritation in 

two populations: one with low exposure (an average of 32 egg masses/acre) and the other with 

high exposure (an average of 3809 egg masses/acre) (Tuthill et al. 1984). The low-exposure 

group exhibited no increase in skin irritation and 32 egg masses/acre is taken as a NOAEL for 

humans and is used as a surrogate RfD for exposure to the gypsy moth in a manner analogous to 

the use of RfD values for control agents. The high exposure group did evidence a significant 

increase in skin irritation. Based on the observed dose-response relationship, egg mass densities 

up to 128 egg masses/acre are not likely to cause a detectable increase in skin irritation or rashes. 

In addition to these quantitative estimates, the severity of the response is important, particularly 

in a comparison of effects caused by exposure to the gypsy moth and effects caused by exposure 

to the agents used to control the gypsy moth. Dermal responses to the gypsy moth are 

sufficiently severe to have generated numerous case reports. While precise statistics are not 

available, it does appear that the severity of the skin irritation is sufficient to cause appreciable 

numbers of affected individuals to seek medical care. While exposure to the gypsy moth is 

associated with irritation to the eyes and respiratory tract, quantitative dose-response 

relationships for these endpoints cannot be developed. 

3.3.2. Effects on Skin 

Of the several available studies that demonstrate skin irritation in humans after exposure to gypsy 

moth larvae (see Section 3.1.3), the study by Tuthill et al. (1984) is the most appropriate from 

which to derive a quantitative dose-response assessment. Tuthill et al. (1984) investigated 

adverse dermal responses in many individuals after exposure to the gypsy moth in areas of high 

and low infestation. As summarized in Table 3-1, the most relevant data are taken from two 

groups, one that consisted of 557 individuals in an area of low infestation (Medway, 

Massachusetts) and one that consisted of 508 individuals in an area of high infestation 

(Lunenberg, Massachusetts). Although the survey was conducted in the summer of 1982 over 

two time periods, prior to and after larval emergence, the exposure estimates are based on egg 

mass counts taken in the fall of 1981. In the Tuthill et al. (1984) publication, the egg mass 

counts are given only as ranges: 10 to 10,000 egg masses/acre in the area of high infestation and 

0 to 70 egg masses /acre in the area of low infestation. 

One of the coauthors of the Tuthill et al. (1984) study surveyed egg masses in the two 

communities (O'Dell 1994). In the high exposure community, surveys were conducted on 27 

0.1-acre plots throughout the infested area between October 5 and 6, 1981. In the high exposure 

areas (Lunenberg), the average egg mass density was 3809 egg masses/acre. In the low exposure 

community (Medway), 20 sites were surveyed. The arithmetic average number of egg 

masses/acre was 32, but the egg masses were unevenly dispersed. No egg masses were found at 

15 of the 20 sites, and egg mass counts at the other 5 sites were 2, 2, 3, 7, and 50. These egg 

mass counts were made in the fall, before the outbreak of rashes in the following summer. The 

use of these egg mass densities as a surrogate for estimating exposure to larvae is based on the 

assumption that there is a positive correlation between the number of viable larvae in the summer 

3-8
 
and the number of egg masses in the preceding fall. Occasionally, below normal midwinter 

temperatures have resulted in high mortality among overwintering eggs (Bess 1961). Usually, 

however, fall egg mass counts are closely related to subsequent larval density, particularly among 

early instars. For the dose response assessment, the average egg mass counts are used for each 

site—i.e., 32 egg masses/acre for Medway and 3809 egg masses/acre for Lunenberg. 

In the Tuthill et al. (1984) study, response data for both the low and high exposure areas are 

presented as the number of individuals with and without signs of dermal irritation. This type of 

data is typically termed quantal or discrete and can be used to assess the statistical significance of 

differences between two groups using the Fisher Exact Test (Uitenbroek 1997). Typically, the 

Fisher Exact Test is used to determine if there are significant differences between a control group 

and an exposed group. The Fisher Exact Test yields a p-value, the probability that the observed 

difference occurred by random chance. If the p-value is very low, the differences are considered 

statistically significant. Typically, a p-value of 0.05 is used as the maximum value for asserting 

that the differences are significant. If the p-value is greater than 0.05, the differences are not 

regarded as statistically significant. 

The Tuthill et al. (1984) study does not include an actual control group—i.e., a population in an 

area where no gypsy moth were present; however, for both Lunenberg and Medway, the 

investigators provide responses before and after larval emergence. Consequently, within each 

group, the response rate prior to larval emergence can be considered a “control” response and the 

response rate after emergence can be considered a response associated with exposure to the gypsy 

moth larvae. Using the Fisher Exact Test, the Medway population demonstrates no statistically 

significant response after larval emergence. In other words, the p-value is 0.3 for the comparison 

of response rates before and after exposure and the probability that this difference could be due to 

random variation is 0.3 or 30%. Thus, the exposure estimate of 32 egg masses/acre may be 

considered a NOAEL (no-adverse effect level). For the Lunenberg population, however, the 

post-emergence response rate of 50/508 is significantly higher than the pre-emergence response 

rate of 7/508 and the p-value is 8×10 -10 . In other words, the probability that the difference is due 

to random variation is only 8 in 100 million. Thus, the exposure estimate of 3809 egg 

masses/acre may be considered a LOAEL (lowest observed adverse effect level). 

In addition to the pre- and post-emergence dermatological response rates for all individuals in the 

two areas, Tuthill et al. (1984) also provide post-emergence data on three different age groups: 0­

12 years, 13-59 years, and >59 years. Because no pre-emergence response data are provided by 

age group, no statistical analysis on “control” vs exposed groups can be conducted. Nonetheless, 

within the high exposure groups (Lunenberg), an age-response pattern is clearly apparent with 

younger individuals being much more sensitive than older individuals. As indicated in Table 3­

1, these differences are both statistically significant and substantial, spanning a nearly 10 fold 

difference in sensitivity—i.e., 2.1% in older individuals vs. nearly 20% in individuals in the 0-12 

year age range. 

3-9
 
Whether or not these different response rates in the different age groups represents a true agespecific difference in sensitivity is unclear. Young children, compared with adults, are likely to 

spend more time out of doors and may be more likely to come into contact with gypsy moth 

larvae. As indicated in Table 3-1, there is a clear association between the number of individuals 

who reported touching larvae and the number of individuals who developed a rash after touching 

larvae. Thus, it is plausible that the age-specific pattern apparent in the data from Tuthill et al. 

(1984) could be an artifact of greater contact with gypsy moth larvae by younger individuals. 

Because Tuthill et al. (1984) included only two exposure groups and no true control group, a 

more quantitative dose-response assessment is limited. The U.S. EPA (2001) developed a series 

of models for estimating benchmark doses. As defined by U.S. EPA (2001), the benchmark dose 

is the estimate of the lower range of a confidence interval for a dose or exposure associated with 

a defined response rate. For example, a benchmark dose could be calculated as the 95% lower 

limit for an exposure associated with a 10% response. 

Using the U.S. EPA (2001) benchmark dose software, benchmark doses for both 1% and 10% 

responses were calculated for all groups combined as well as for each age-group. Because of the 

limited exposure data, the simple exponential model is used: 

P = 1 - exp($ *EM) 

where $ is the potency parameter in units of proportion responding per egg mass/acre and EM is 

the number of egg masses/acre. These analyses are summarized in Table 3-1. For all groups 

combined, the pre-exposure responses (Table 3-1) were used as zero exposure or control 

responses. For the age-group specific modeling, no control group was used—i.e., the model has 

zero degrees of freedom. Thus, the p-values shown in Table 3-1 are just an indication of whether 

or not the potency parameter was significantly different from zero. 

The results of the dose response modeling are qualitatively consistent with the use of the simpler 

Fisher Exact Test. The potency parameter is greatest in the 0- to12- year-old groups. The doseresponse relationship for the >59-year-old group is not statistically significant (p=0.15), 

indicating no substantial response in individuals more than 59 years old in either Lunenberg or 

Medway. Based on the most sensitive individuals, egg mass densities of 128 egg masses/acre are 

likely to cause adverse effects in no more than 1%—i.e., a response rate unlikely to be detectable 

in an epidemiology study. Egg masses of 1336, however, are likely to cause a response rate of at 

least 10%, which would be detectable in a well-conducted epidemiology study. Again, these 

results are essentially consistent with the NOAEL and LOAEL values discussed above. 

For the current risk assessment, the NOAEL of 32 egg masses/acre is used as a surrogate RfD for 

exposure to the gypsy moth in a manner analogous to the use of RfD values for control agents. 

While an uncertainty factor is typically applied to NOAEL values to estimate an RfD, no 

uncertainty factor is used for this risk assessment. This approach seems reasonable based on the 

3-10
 
benchmark dose modeling which indicates that egg mass densities up to 128 egg masses/acre are 

not likely to cause a detectable increase in skin irritation or rashes. 

In addition to these quantitative estimates, the severity of the response is important, particularly 

in a comparison of effects caused by exposure to the gypsy moth and effects caused by exposure 

to the agents used to control the gypsy moth. Dermal responses to the gypsy moth are 

sufficiently severe to have generated numerous case reports as well as a study by NIOSH (see 

Section 3.1.3). One of the criteria for judging the severity of any response is whether or not an 

individual will seek medical attention as the result of an exposure to a particular agent. No 

precise statistics on seeking medical attention after exposure to gypsy moth larvae are available. 

Tuthill et al. (1984) have noted that: “Less than 10 per cent of the sufferers sought medical 

care.” As discussed further in the risk comparison for these agents, a response rate of 10% is 

substantially greater than rates for any of the agents used to control the gypsy moth, based on 

comparable data. 

3.3.3. Other Effects 

As discussed in Section 3.1, exposure to the gypsy moth is associated with irritation to the eyes 

and respiratory tract as well as generalized psychological distress during severe infestations. No 

data, however, are available for developing quantitative dose-response relationships for these 

endpoints. 

3-11
 
 

3.4. RISK CHARACTERIZATION 

3.4.1. Overview 

In sparse to moderate infestations—i.e., egg mass densities of <500 egg masses/acre—adverse 

effects involving skin irritation are not likely to be detectable in populations of exposed humans. 

Nonetheless, some individuals who come into contact with gypsy moth larvae could develop skin 

irritation. In heavy gypsy moth infestations—i.e., >500 to 5000 egg masses/acre—adverse skin 

reactions would be expected in substantial numbers and the effects would likely be sufficiently 

severe to cause some individuals to seek medical attention. In extreme outbreaks—i.e., >5,000 

to 20,000 egg masses/acre— the effects will be qualitatively similar to those of severe 

infestations but could affect up to about one-third of the population. Heavy infestations or 

extreme outbreaks could cause ocular and respiratory effects in some people but the likelihood 

of observing these effects cannot be quantified. Similarly, severe infestations are often 

considered to be a nuisance and cause aesthetic damage to the environment. Both of these 

factors can lead to stress in some individuals. There is speculation that severe damage to oak 

forests from gypsy moth infestations might result in a decrease in the prevalence of Lyme 

disease. This effect of gypsy moth exposure obviously would be viewed as beneficial to human 

health. Currently, however, there does not appear to be sufficient information to assess the 

plausibility of this supposition. Young children may be a group at special risk from effects of 

gypsy moth exposure but it is not clear whether children are more sensitive than adults to the 

effects of gypsy moth exposure or whether responses in children appear greater because children 

spend more time outdoors compared with adults. 

3.4.2. Effects on Skin 

The likelihood of adverse skin reactions in humans after exposure to gypsy moth larvae can be 

quantified at least in terms of egg mass density (see Section 3.3). Skin irritation also may be 

considered the most sensitive effect. That is, if exposure levels are less than levels at which a 

substantial increase in skin irritation is observed, other effects are not likely to be seen. 

The risk characterization for the general public is summarized in Table 3-3. The ranges of risk 

for the general public are based on the ranges of exposure given in column 2 of this table. As in 

the ecological risk assessment, the stratification of sparse to extreme infestations in terms of eggs 

masses/acre is somewhat arbitrary but covers a sufficiently broad range to encompass most egg 

mass densities that are likely to be encountered —i.e., from 50 to 20,000 egg masses/acre. Egg 

mass densities of 50 egg masses/acre or less are characteristic of mild infestations that occur in 

the south central region of the United States (Davidson et al. 2001). Egg mass densities of 

20,000 egg masses/acre or more are uncommon but can occur in localized areas during gypsy 

moth outbreaks (Hajek 1997). 

Three types of risk characterizations are provided in Table 3-3. The first is based on the NOAEL 

of 32 egg masses/acre. As discussed in Section 3-3, this value is used as a surrogate RfD for 

exposure to the gypsy moth. As with all hazard quotients (HQs) based on an RfD, an HQ of less 

than one indicates that no adverse effects are plausible. The second type of risk characterization 

is based on a LOAEL of 1336 egg masses/acre. As indicated in Table 3-2, this value is the 

3-12
 
estimated benchmark dose associated with a 10% response in the most sensitive subgroup 

(children < 13 years old). This value is considered a LOAEL rather than a NOAEL because a 

response rate of 10% would be detected in an epidemiology study and because the value is very 

close to the observed LOAEL of 3809 egg masses/acre in the study by Tuthill et al. (1984). The 

interpretation of the hazard quotients based on this LOAEL is different from standard hazard 

quotients based on an NOAEL or RfD—i.e.,values greater than 1 indicate that adverse effects are 

likely to be observed in the exposed population. 

In addition to the risk characterizations based on the NOAEL/RfD and LOAEL, the last column 

in Table 3-3 gives the upper range of extra risk associated with each of the exposure categories. 

These values are derived from the U.S. EPA (2001) benchmark dose software using the one-hit 

model, as discussed in Section 3.3. 

Taken together, all three numerical expressions of risk lead to a consistent qualitative risk 

characterization. In sparse to moderate infestations—i.e., egg mass densities of <500 egg 

masses/acre—adverse effects involving skin irritation are not likely to be detectable in 

populations of exposed humans. Nonetheless, some individuals who come into contact with 

gypsy moth larvae could develop skin irritation. In heavy gypsy moth infestations, defined in 

Table 3-3 as ranging from >500 to 5000 egg masses/acre, it is likely that adverse skin reactions 

will be reported and that the effects will be sufficiently severe to cause some individuals to seek 

medical attention. In extreme outbreaks—i.e., >5,000 to 20,000 egg masses/acre— the effects 

will be qualitatively similar to those of severe infestations but could affect up to about one-third 

of the population. 

3.4.3. Other Endpoints 

As discussed in the hazard identification (see section 3.2), exposure to gypsy moth larvae is 

associated with ocular and respiratory effects in humans. In addition, infestations are often 

considered to be a nuisance and cause aesthetic damage to the environment. Both of these 

factors can lead to stress in some individuals. Also during severe infestations, water quality may 

be affected. While all of these concerns may be qualitatively associated with exposure to the 

gypsy moth and while the severity of these effects are likely to increase with the increasing 

severity of gypsy moth infestations, no quantitative dose-response assessment can be made (see 

Section 3.3.3). Accordingly, no quantitative risk characterization can be provided. 

As discussed in Section 3.1.5, there is reason to speculate that severe damage to oak forests could 

result in a decrease in the prevalence of Lyme disease. This effect of gypsy moth exposure 

obviously would be viewed as beneficial to human health. Currently, however, there does not 

appear to be sufficient information to assess the plausibility of this supposition. 

3.4.4. Sensitive Subgroups 

Young children may be a group at special risk from effects of gypsy moth exposure. Although 

this is suggested in the study by Tuthill et al. (1984) as well as by studies on school children 

affected by gypsy moth infestations (Aber et al. 1982; Anderson and Furniss 1983), it is not clear 

3-13
 
whether the finding indicates that children are inherently more sensitive than adults to the effects 

of exposure or whether children have a greater incidence of response because they spend more 

time outdoors than adults and thus have great potential for exposure to gypsy moth larvae. 

3.4.5. Connected Actions 

There is no evidence to assess the consequences of connected actions involving the various 

program activities or other common activities associated with the gypsy moth. As discussed in 

the risk assessment on Gypchek, one of the agents used to control gypsy moths, Gypchek 

contains gypsy moth parts and may cause irritant effects similar to those caused by the gypsy 

moth. Consequently, it may be that the effect of simultaneous exposure to gypsy moth larvae and 

Gypchek would be additive. 

3.4.6. Cumulative Effects 

Two types of cumulative effects may be considered in assessing the consequences of exposure to 

the gypsy moth. During an infestation, repeated exposures will occur in the population for the 

duration over which exposure to the gypsy moth instars occurs. In addition, cumulative effects 

may be induced from year to year as infestations reoccur. Cumulative effects from exposure to 

the larvae during a single season are essentially encompassed by the Tuthill et al. (1984) study, 

the epidemiology study on which the risk assessment is based, because the investigators 

monitored effects in populations during the period in which early instars were present. The 

available data do not permit a definitive assessment of the cumulative effects of exposure to the 

gypsy moth over several seasons. As discussed in the hazard identification (see Section 3.1.2), 

there is evidence to suggest that an allergic reaction may be one of the mechanisms involved in 

the dermal effects associated with exposure to the gypsy moth. Thus, it is plausible that some 

individuals may become sensitized to the gypsy moth after repeated exposures over 1 or more 

seasons. 

3-14
 
4. ECOLOGICAL RISK ASSESSMENT
 

4.1. HAZARD IDENTIFICATION 

4.1.1. Overview 

The clearest primary effect of gypsy moth infestations is on terrestrial plants, primarily trees. 

Various instars of the gypsy moth larvae will feed on host trees and can cause extensive 

defoliation which can kill some of the infested trees. On a larger scale, the extensive defoliation 

and/or death of trees may result in secondary changes to vegetation, which will, in turn, affect 

other forms of vegetation as well as various animal species (primarily related to changes in 

habitat). Gypsy moth larvae appear to have definite food preferences; oak, birch, poplar and 

apple trees seem to be their favorite food sources. While both the European and Asian gypsy 

moth cause similar types of damage (i.e., defoliation), their feeding preferences are somewhat 

different with the Asian gypsy moth preferring a wider range of vegetation. Heavy defoliation is 

much more common among the oaks than among trees that are not particularly favored as food 

by the gypsy moth. For susceptible oaks, the effects of infestations on tree mortality varies 

according to the initial condition of the stand and the severity and frequency of defoliations. 

Generally, gypsy moth infestations result in mortality losses of less than 15% of total basal area. 

When heavy defoliation is followed by massive overstory mortality, existing shrub and herb 

cover increase dramatically due to increases in available light, moisture, and nutrients. 

The only other group of organisms that are likely to be directly effected by the gypsy moth are 

some and probably very few other lepidopteran species, including the northern tiger swallowtail 

butterfly. The mechanisms for direct adverse effects may include bacterial contamination of the 

leaves by gypsy moth larvae and a decrease in the nutritional value of the leaves damaged by the 

gypsy moth. Most studies, however, do not indicate substantial direct effects on other insects, 

including lepidoptera. In some cases, increases may be seen in populations of insect predators of 

the gypsy moth. 

There is no indication in the literature that the gypsy moth will cause direct adverse effects in 

most groups of animals. Indirect effects, associated with damage to vegetation, may be of 

substantial consequence to some species, including squirrels, mice, and other mammals that rely 

on acorns. Although some mammals consume insects, including the gypsy moth, there is no 

evidence that gypsy moth outbreaks have a substantial impact on insectivorous mammals. 

Similarly, there is little indication that birds or aquatic species will be adversely affected by the 

gypsy moth. In some species of birds, gypsy moth infestations and subsequent defoliation may 

be beneficial, especially for species of birds that favor dead wood as a habitat. 

4.1.2. Terrestrial Organisms 

4.1.2.1. Mammals – As discussed in Section 3.1 (Human Health Hazard Identification), direct 

exposure to gypsy moth larvae causes various irritant effects in humans —i.e., skin, eyes, and 

respiratory tract. In most species of mammalian wildlife, however, fur is likely to reduce the risk 

of direct contact between the gypsy moth and the skin of the animal, making skin irritation an 

4-1
 
unlikely result of exposure. Evidence of irritation to the eyes and or respiratory tract in 

mammalian wildlife species after direct contact with the gypsy moth is not found in the literature. 

Although the hazard identification for the direct effects of gypsy moth exposure in mammalian 

wildlife is basically negative, indirect effects may be of substantial consequence to some species, 

as discussed in Section 4.1.2.5. For instance, gypsy moth outbreaks that cause substantial 

defoliation and mortality in some tree species, particularly oaks, could adversely affect the 

production of acorns (Gottschalk 1990a; McConnell 1988), which may limit food availability for 

some forest mammals. 

To determine the effects of a gypsy moth outbreak on a population of black bears (Ursus 

americanus), Vaughan and Kasbohm (1993) monitored the behavior of 54 radio-collared black 

bears in the Shenandoah National Park after a gypsy moth outbreak that caused widespread 

defoliation, hard mast (acorn) failures, and tree mortality. The outbreak had no apparent effects 

on cub production or mortality rates of cubs or adults. Although the bears exhibited different 

habitat preferences at all seasons, they did not avoid defoliated habitat. In the fall, before the 

gypsy moth infestation, the bears ate mostly acorns. When acorns were no longer available due 

to defoliation of oak trees by the gypsy moths , the bears switched to eating fruit, which had no 

apparent impact on the nutritional quality of their diets. Seventy-one percent of bear dens were 

in tree cavities, primarily in living oaks (mean diameter at breast height = 98 cm). Gypsy mothinduced mortality of den trees was high and by the end of the study, 54% of the living oaks used 

as dens were dead. While no short-term effects were noted, Vaughan and Kasbohm (1993) 

speculated that the long-term adverse impact of defoliation on black bears may be a reduction in 

den sites, with natural replacement possibly requiring 50 years. Conversely, black bears will use 

as dens the upturned stumps of large dead trees that have been blown over. These would be 

expected to increase after severe defoliation sufficient to cause tree mortality. 

Variations in acorn and other mast production are directly related to variations in populations of 

squirrels, mice, and other small mammals (Brooks et al. 1998; Gorman and Roth 1989; Nixon et 

al. 1975). The size of the acorn crop in the fall directly affects the population density of mice 

living in oak-dominated forests the following spring (McShea and Rappole 1992; McShea and 

Schwede 1993). By damaging oak tress, gypsy moth infestations can decrease acorn production 

and a decrease in acorn production secondary to gypsy moth infestations has been shown to 

decrease the population of white-footed mice, Peromyscus leucopus (Elkinton et al. 1996, 2002). 

Although some mammals consume insects, including the gypsy moth (see Section 2.4), there is 

no evidence that gypsy moth outbreaks have a substantial impact on insectivorous mammals. 

Also, there is no evidence that the effects of gypsy moth outbreaks on other insect populations 

will directly or indirectly affect mammals that prey on insects. Sample et al. (1996) found no 

significant effects on the consumption of insects by Virginia big-eared bats in areas of high gypsy 

moth infestation. 

4-2
 
4.1.2.2. Birds – There is little indication that birds will be adversely affected by the gypsy moth. 

Based on predation by various species of birds on the gypsy moth compared to other hairless 

lepidoptera, some species of birds appear to avoid the gypsy moth as a prey species (Smith 

1985). This suggests, at least indirectly, that the setae or hairs on the gypsy moth larvae may 

have irritant properties for birds. Direct adverse effects, however, have not been noted in the 

literature. Reported increases in nesting failures of various species of birds appear to be due to 

increased predation and/or increased weather stress, both associated with defoliation (Crocoll 

1991; Thurber et al. 1994). 

In some species of birds, gypsy moth infestations and subsequent defoliation may be beneficial, 

especially for species of birds that favor dead wood (snags) as a habitat. As a result of 

defoliation by the gypsy moth, the amount of dead wood increases, particularly in the upper story 

and the dense vegetation in lower forest strata, providing habitat that is scarce in closed-canopy 

forests. These secondary effects of gypsy moth outbreaks, which can be considered beneficial to 

numerous birds, are well documented in the gypsy moth literature (Bell and Whitmore 1997a,b; 

DeGraaf 1987; DeGraaf and Holland 1978; Showalter and Whitmore 2002). Bell and Whitmore 

(1997) report that available nesting and foraging resources increased for several bird species as 

result of more snags, windfall, and shrub cover after defoliation, while there was no substantial 

impact from upper canopy defoliation on birds residing primarily in the forest canopy. Only tree 

nesting and flycatching guilds appeared to be affected adversely by the moth infestation. Cavitynesting birds also benefitted indirectly from a gypsy moth outbreak (Showalter and Whitmore 

2002). Thurber (1993) noted that bird density increased in plots in which the defoliation was of 

low to moderate impact. Species richness increased from 19 to 23 species per plot, with declines 

noted only for tree nesters and flycatchers on high impact plots (Thurber 1993). Increases in low 

shrub and ground nesters, cavity nesters, low shrub and ground foragers, bark foragers, forest 

edge species, short-distance migrants, year-round residents, and woodpeckers were widespread, 

but most pronounced on moderate impact plots. DeGraaf and Holland (1978) reported similar 

results, finding significantly fewer numbers of only 4 out of 36 bird species examined in heavily 

defoliated areas. DeGraaf (1987) notes no substantial effects on abundance of various species of 

birds in defoliated and non-defoliated stands in central Pennsylvania studied over a two year 

period. 

4.1.2.3. Reptiles and Amphibians (Predominantly Terrestrial) – There is very little information 

regarding the effects of gypsy moth infestations or outbreaks on amphibians or reptiles. In the 

short-term, gypsy moth defoliation could have a negative impact on some habitats occupied by 

reptiles and amphibians by increasing solar radiation on dead and down material, litter, and the 

other materials found above subterranean habitats; however, in the longer term, the defoliationinduced increases in dead and down material will be beneficial to reptiles and amphibians 

(Schweitzer 1988). Peterson (1990) conducted a field study in south central New York on the 

effect of gypsy moth infestations on the timber rattlesnake, Crotalus horridus and noted that 

gypsy moth-induced defoliation had an adverse effect on rattlesnakes, primarily through 

reductions in acorn production and the consequent decrease in the population of small rodents 

that the snakes eat (see Section 4.1.2.1). 

4-3
 
4.1.2.4. Terrestrial Invertebrates – Some lepidopteran species may be adversely affected by 

gypsy moth outbreaks and at least some of these effects may be direct rather than secondary. 

Redman and Scriber (2000) examined the adverse effects of the gypsy moth on the northern tiger 

swallowtail butterfly, Papilio canadensis, and demonstrated several different mechanisms 

associated with the adverse effects. Direct effects included 100% mortality in Papilio larvae 

exposed to leaves painted with gypsy moth body fluids, and 84% mortality in Papilio larvae fed 

leaves from aspen stands infested with gypsy moth larvae. Although the cause of death in the 

Papilio larvae was not clear, the investigators speculate that it was generally due to bacterial 

contamination of the leaves by gypsy moth larvae, since sterilized leaves did not cause a 

significant increase in mortality. Moreover, the damage to aspen leaves caused by gypsy moth 

larvae decreased the nutritional value of the leaves, which led to reduced growth rate and survival 

of the Papilio larvae. In addition, fields studies conducted by Redman and Scriber (2000) 

demonstrated that proximity to gypsy moths increased the rate of parasitism of the Papilio larvae. 

The potential adverse effects of gypsy moth outbreaks to lepidoptera was also investigated by 

Sample et al. (1996) in a study designed to compare lepidopteran populations in 50 acre plots in 

mixed oak, hickory, and pine forests in West Virginia. Contaminated plots were characterized as 

stands with average egg mass densities of 235-1156 egg masses/ha (95-468 egg masses/acre), 

larval abundance of about 68-111 larvae/50 g dry foliage, and defoliation rates up to 88% over a 

3-year period. Uncontaminated plots were characterized as stands with average egg mass 

densities of about 15-180 egg masses/ha (6-72 egg masses/acre), larval abundance of about 4-18 

larvae/50 g dry foliage, and no defoliation over a 2-year period with 40% defoliation in the third 

year. Decreases in abundance and richness of larvae and adults from the family Arctiidae (tiger 

moths) were apparent in plots infested with gypsy moth larvae, compared with uncontaminated 

plots. The differences were statistically significant for both abundance (p=0.038) and species 

richness (p=0.0015). No substantial differences were observed in other lepidoptera or other 

invertebrate taxa, although a significant increase was noted in braconid wasps. Sample et al. 

(1996) suggested that the increased abundance of braconids in the plots with gypsy infestation 

was likely due to increased host (i.e., gypsy moth) availability. 

The study by Work and McCullough (2000) demonstrates further that the impact of the gypsy 

moth is negative to only a small proportion of the lepidopteran community, primarily species that 

feed on oak and for which the larval development of the affected species and gypsy moth 

presumably coincide. Although the study does not address the mechanism(s) by which the gypsy 

moths adversely affect the lepidopteran community, the investigators suggest they might include 

altered host/plant quality, increases in natural enemies, or microclimate changes. All but the 

latter mechanism are demonstrated in the study by Redman and Scriber (2000) discussed above. 

No significant effects were observed on generalist woody plant feeders. Summerville and Crist 

(2002) criticize the guild classification used by Work and McCullough (2000); however, it is not 

clear what impact the use of alternate guild classifications would have on the conclusions 

reached in the study. 

4-4
 
Contrary to studies demonstrating the adverse effects of gypsy moth infestations to some 

macrolepidoptera, anecdotal reports suggest that certain lepidopteran species respond positively 

to gypsy moth infestations. Schweitzer (1988) claims that 1981 produced the highest number of 

butterfly species ever for the New Haven, Connecticut area, which for many years stood as the 

record for eastern North America, despite the record number of acres defoliated by the gypsy 

moth that same year. 

4.1.2.5. Terrestrial Plants (Macrophytes) 

4.1.2.5.1. Gross Effects on Trees – The clearest primary effect of gypsy moth 

infestations is on terrestrial macrophytes, primarily trees. Various instars of the gypsy moth 

larvae will feed on host trees and can cause extensive defoliation which can kill some of the 

infested trees. On a larger scale, the extensive defoliation and/or death of trees may result in 

secondary changes to vegetation, which will, in turn, affect other forms of vegetation with 

consequences to various animal species (primarily related to changes in habitat). 

Trees that are defoliated by even 75% or more are likely to refoliate during the same season. The 

refoliated leaves are smaller and fewer, and repeated defoliations can cause additional reductions 

in leaf size (Wargo 1981a). According to Wargo (1981b), trees that refoliate are completely out 

of phase with the season. Visually, for example, the condition of trees in a mixed composite 

stand of oaks (red, black, scarlet, and white) in eastern New England showed rapid decline in the 

year after defoliation and continued to decline slightly during the next 5 years. Following a 

single heavy defoliation, about 10 years passed before these trees returned to their predefoliation 

condition (Campbell and Sloan 1977b). 

Parker (1981) identifies five key factors that determine the effects of tree defoliation. The 

factors include, severity (how much foliage is removed); frequency (the number of successive 

years of defoliation); timing (when in the growing season the tree is defoliated); pathogens (the 

presence and number of secondary organisms); and health and vigor (the physiological condition 

of the tree when it is defoliated). Defoliation appears to have a direct impact on root 

carbohydrates (Kosola et al. 2001, 2002). Most hardwood (or deciduous) trees are able to 

tolerate at least 2 years of defoliation before root starch content (useable energy) is depleted 

(Wargo 1981a). Since most coniferous species store carbohydrate resources necessary to 

refoliate in the leaves, they are usually unable to survive a single, complete defoliation (Johnson 

and Lyon 1988). Further decline and possible death of previously defoliated eastern hardwood 

trees are due primarily to secondary organisms like the shoestring fungus, Armillaria species, and 

the twolined chestnut borer, Agrilus bilineatus (Wargo 1981b). The defoliator and borer cause 

adverse effects in the crown of the tree. The borer affects the main stem and the fungus attacks 

the roots (Wargo 1977, 1981b). Gottschalk (1994) notes that by removing weak, sickly trees 

from the forest population, fungus (Armillaria species) and tree borers (Agrilus species) play an 

important and positive role in forest health. 

Previous stand disturbance, which may allow partial colonization of root systems by Armillaria, 

increases rhizomorph abundance (Twery et al. 1990; Wargo 1989). Even in the presence of 

4-5
 
abundant rhizomorphs, however, non-defoliated and lightly defoliated trees may remain resilient 

(Twery et al. 1990). Stressed trees also provide an environment that favors the survival of the 

twolined chestnut borer (Twery 1991; Wargo 1977), which is attracted to volatile chemicals 

released by stressed oaks (Dunn et al. 1986a). The trees most susceptible to the pest are those 

with low stores of starch reserves; however, only the trees with extremely low winter root starch 

reserves are likely to die (Dunn et al. 1986b, 1987). 

Factors that contribute to interspecies differences in response to defoliation include where in the 

tree reserve energy is stored, the amount of energy required to refoliate, and how much energy is 

needed to maintain growth during refoliation (Twery 1991). Hemlock, for example, usually will 

not survive even one complete defoliation (Stephens 1988), whereas some oaks on dry sites may 

survive repeated defoliations indefinitely (Houston and Valentine 1977; Bess et al. 1947; Twery 

1991). 

Heavy defoliation is much more common among the oaks than among trees that are not 

particularly favored as food by the gypsy moth. For susceptible oaks, the effects of infestations 

on tree mortality varies according to the initial condition of the stand and the number of 

infestations. Davidson et al. (1999) found that stands with good crown condition, had mortality 

rates of only 7% and 22% after one and two infestations, respectively; however, in stands with 

poor crown quality, the corresponding mortality rates were 36% (one infestation) and 55% (two 

infestations). Heavy defoliation usually increases mortality rates even among trees that are 

generally not preferred by the gypsy moth as food sources. For instance, a single heavy 

defoliation in eastern New England resulted in 69% mortality of trees that are eaten but not 

preferred by the gypsy moth compared with about 37% mortality in oaks, a source of food that is 

preferred by the gypsy moth (Campbell and Sloan 1977b). 

Gypsy moth infestations generally result in mortality losses of less than 15% of total basal area. 

For example, in an artificially induced gypsy moth outbreak in poplars (Populus euramericana) 

that resulted in nearly complete (70-100%) defoliation of some stands as well as a 25% decrease 

in stem production, tree mortality ranged from 6 to 10% (Agrawal et al. 2002). Losses of 

15-35% are not uncommon, and losses occasionally exceed 50% (Gottschalk et al. 1987). 

Volume growth is reduced among surviving trees for approximately 3 years after a defoliation 

episode (Picolo and Terradas 1989; Twery 1991; Muzika and Liebhold 1999). The study by 

Twery (1987) indicates that, on average, stem volume growth in oaks decreased 20% in any year 

in which a tree was defoliated, compared with the previous year in which there was no 

defoliation. This effect is due in part to the reduced leaf area in the recovering trees (Wargo 

1981a). In any given stand, heavy defoliation year after year tends to be a rare event. When such 

an event does occur, however, consequent tree mortality rates may become very high. In the area 

described by Campbell and Sloan (1977b), for example, only 7% of the mixed oaks rated "good" 

died following a single heavy defoliation. After two successive heavy defoliations, however, 

mortality rates in this category increased to 27%. 

4-6
 
Between 1911 and 1921, defoliation caused by the gypsy moth was heavy along the eastern 

seaboard of New England. During this decade the oak component suffered about 60% mortality. 

About 30% of red maples and 33% of white pines also died (Campbell and Sloan 1977b). 

During the next decade, both defoliation and the responses to it were significantly less (Baker 

1941; Campbell and Sloan 1977b). Tree mortality in response to gypsy moth outbreaks appears 

to follow a general pattern in which the most severe tree mortality occurs during and after an 

initial outbreak (Gansner and Herrick 1984; Herrick and Gansner 1986, 1988; Twery and 

Gottschalk 1989; Twery 1991). Campbell and Sloan (1977b) observed that certain trees within 

any given species consistently suffered heavier defoliation than others and were more likely to 

die, suggesting that differential intraspecific mortality could account for the subsequent 

decrement in stand vulnerability. Similarly, Byington et al. (1994) noted marked difference in 

tolerance to gypsy moth damage among nine families of red oak. 

4.1.2.5.2. Differential Feeding Preferences for Trees – Gypsy moth larvae appear to 

have definite food preferences; oak, birch, poplar, willow and apple trees seem to be their 

favorite food sources. In the northeast, preferences vary among species of oak with the greatest 

preference shown for black and burr oak (Foss and Rieske 2003). Other species, like beech, 

maple, and white pine are less favored by the gypsy moth, and hemlock and pitch pine seldom 

serve as food sources. Mortality in white pine, however, can be substantial in stands where pine 

occurs in the understory. Much less damage occurs in oak/pine stands where pine shares the 

canopy with oak (Brown et al. 1988). Other species of trees such as black locust and ash are 

generally not substantially damaged by the gypsy moth (Campbell 1979). The avoidance of 

green ash (Fraxinus pennsylvanica) by gypsy moth larvae seems to be related to the presence of 

chemicals—not clearly identified—in the leaves of the trees (Markovic et al. 1996). On the other 

hand, gypsy moth larvae seem able to adapt to even unsuitable hosts such as the alkaloid rich 

foliage of locust trees (Lazarevic et al. 2002). 

In addition to general host preferences, there appear to be regional differentiations among 

preferred moth host plants. Oak is the preferred species in the east and quaking aspen is the 

preferred species in the Midwest (Redman and Scriber 2000). Hornbeam is strongly preferred by 

gypsy moth larvae in Quebec, but in New England it is only an intermediate host, while gray 

birch and quaking aspen are both preferred by gypsy moth larvae in New England but are classed 

as only as intermediate hosts in Quebec (Mauffette et al. 1983). 

Compared with the European gypsy moth, the Asian gypsy moth feeds more voraciously and 

grows faster on white oak, larch, and paper birch. In the former Soviet Union, the Asian gypsy 

moth feeds on more than an estimated 600 tree or plant species. Moreover, the Asian gypsy 

moth may not only thrive on the same tree species eaten by the European gypsy moth, but may do 

better on many species that the European gypsy moth does not favor, such as the Douglas-fir 

(USDA/FS 1992). Waller (1994) reports that the Asian gypsy moth grows better than European 

gypsy moth on 50 plant species in the United States, with the greatest differences in growth rates 

associated with coniferous species (Wallner 1994). At least with the Asian gypsy moth, drought 

may be a predisposing factor to severe damage from infestations (Koltunov and Andreeva 1999). 

4-7
 
The Asian strain of gypsy moth is a more serious problem in western forests, compared with the 

European strain (Montgomery 1993). 

4.1.2.5.3. Effects on Stand Structure and Productivity – Subdominant trees (trees 

growing largely in shaded areas) suffer much higher rates defoliation induced mortality compared 

with the taller dominants after heavy defoliation (Brown et al 1988; Campbell 1979; McGraw et 

al. 1990; Quimby 1993). Usually, heavy and repeated defoliation results in a more one-storied 

stand. Although oak growing-stock volume in trees less than 10 inch diameter at breast height 

actually decreased between 1965 and 1989, the losses were offset by gains in larger trees 

(Gansner et al. 1993). 

Hix et al. (1991) reported increases in red maple, which corresponded with decreases in oak 

trees, after defoliation in Appalachian forests. Moreover, total density increased from 

approximately 42,000 to 62,000 stems per acre, which the investigators attributed to the increase 

in light, nutrients, and moisture reaching the forest floor after defoliation. Regeneration in the 

Allegheny Mountain region is dominated by red maple, while red maple and birch dominate the 

Ridge and Valley regions. Oak reproduction is sparse and seedlings are small, compared with 

red maple and non-commercial seedlings, in the Allegheny Mountains. According to Allen and 

Bowersox (1989), only 4-16% of the stems were northern red oak or white oak. In many heavily 

defoliated stands, oak reproduction, which greatly depends on the survival of acorns and small 

oak seedlings (0-1 ft tall) is a major concern, especially given the limited information regarding 

the ability of oak to compete successfully with birch and maple trees (Twery 1991; Galford et al. 

1993; Hix et al. 1991). 

Moderately heavy defoliation usually accelerates forest succession toward more shade-tolerant 

(and less defoliation-prone) species (Campbell and Sloan 1977b; Houston 1981b). In contrast to 

widespread scarcity of oak regeneration in other infested areas, oaks often continue to dominate 

stands in frequently defoliated areas with excessively drained, sandy soils (e.g., Cape Cod, MA, 

and the New Jersey coastal plain) or rocky, shallow, ridge top soils (e.g., those common to 

Medford, MA). Other sources indicate that a small number of oaks in young stands in central 

New England may become dominant when the stands reach 50 years of age (Oliver 1978; Twery 

1991). 

Changes in forest composition may account for the frequently-cited reductions in gypsy mothinduced effects in areas such as New England. Gottschalk (1994) reports that moderate to heavy 

overstory mortality in recent years followed heavy defoliation on about 5-20% of defoliated 

Appalachian stands. Nevertheless, tree mortality rates in these stands show no indication of 

downturn after a second wave of equally heavy defoliation. 

Even in healthy stands with little defoliation, heavy crops of hard mast (primarily acorns) are 

only produced about every 3 or more years and during intervening years, mast crops may be poor 

or nearly absent (USDA/FS 1994). Defoliation can virtually eliminate oak mast production, 

especially in the short-run and result in several consecutive years of complete mast failure 

4-8
 
(Gottschalk 1990b; Liscinsky 1984). During years of moderate and heavy defoliation, short-term 

and residual adverse effects on mast production can be attributed to three sources: direct 

consumption of flowers; abortion of immature acorns due to a low carbohydrate supply; and lack 

of flower bud initiation. Available data suggest that abortion of immature acorns is the most 

significant of the adverse effects, which can result in up to 5 years of complete acorn failure 

(Gottschalk 1990a). 

As previously noted, trees that do not die during a defoliation episode may take as long as 10 

years to recover their full vigor. On the other hand, once trees recover their vigor, significant 

overstory mortality (>60% of the basal area) in intermediate and suppressed trees (i.e., not heavy 

mast producers) is required to cause significant reductions in acorn production. Acorn 

production by surviving trees may even be stimulated by this thinning (Gottschalk, 1990a). In 

the long term, loss of acorn and other nuts is partially compensated by an increase in the number 

of flowers (Gottschalk 1990b). 

4.1.2.5.4. Effects on Shrubs and Herbs – When heavy defoliation is followed by 

massive overstory mortality, existing shrub and herb cover increases dramatically due to 

increases in available light, moisture, and nutrients. Under certain conditions, heavy defoliation 

and subsequent overstory mortality can result in dominance by shrubbery and herbaceous 

vegetation for several years. Gansner (1985) describes an understory 10 years following 

defoliation as being dominated by blueberry, witch-hazel, raspberries, and several species of 

ferns, along with some tree seedlings that were heavily browsed by deer. Hix et al. (1991) also 

noted that blueberries and raspberries were often the shrub species that increased the most 

following defoliation. Among herbaceous plants, Brackley (1985) noted that gypsy mothinduced defoliation appeared to stimulate flowering in the endangered orchid, Isotria 

meleoloides, in New Hampshire. 

4.1.2.5.5. Effects on Fire Hazard – Researchers generally agree that heavy defoliation 

caused by the gypsy moth increases fire danger, although differences in fuels have not been 

measured and the increased fire hazard has not been calculated (Gottschalk 1990b). Wildfires 

are more difficult to control in areas of extensive tree mortality (Tigner 1992). Furthermore, the 

numerous standing dead snags may act as lightning rods and further increase risk of fire starts by 

lightning. Fire caused by a lightning strike on one or more of these snags could smolder for 

several days prior to detection (USDA/FS 1994). On the other hand, fires are infrequent during 

the growing season in eastern hardwood forests. Consequently, significant increases in fire 

hazard would occur in hardwood forests during the growing season only as the result of longterm increases in woody fuels due to tree mortality (Gottschalk 1990b). 

4.1.2.6. Soil and Terrestrial Microorganisms – There is little information from which to 

directly assess the effect of gypsy moth infestations on terrestrial microorganisms. Indirect 

evidence suggests that adverse effects are unlikely. Soil microbial activity is largely influenced 

by moisture and temperature. Vaughan and Kasbohm (1993) report that defoliation increases 

maximum daily temperatures. Since microbial activity increases with temperature, defoliation is 

4-9
 
likely to increase microbial activity, thereby accelerating the process of decomposition. 

Decomposing bacteria and fungi have high nutrient requirements. Increased nutrient content in 

litter fall might enhance decomposition, which might be the case during gypsy moth defoliation 

in the spring when leaf matter is consumed before nutrient reabsorption takes place (Grace 1986). 

The effects of these increased nutrient levels and mineralization might be to enhance forest 

regeneration. 

4.1.3. Aquatic Organisms. 

4.1.3.1. Fish – Little information is available regarding the effects of gypsy moth infestations on 

fish populations. Defoliation from the gypsy moth can result in an increase in the pH and 

temperature of ambient water (Downey et al. 1994; Webb et al. 1995). Downey et al. (1994) 

have suggested that trout, which as are very sensitive to changes in pH and temperature, could be 

adversely affected by such changes. As discussed by Webb et al. (1995), however, no direct data 

are available on the biological effects of such changes due to defoliation by the gypsy moth and it 

is unclear if significant biological effects on fish are likely. 

4.1.3.2. Aquatic Invertebrates – Hutchens and Benfield (2000) detected an apparent increase in 

the rate of leaf breakdown in streams due to gypsy moth defoliation, which might result in food 

deficits during spring for shredders—i.e. caddisflies, stoneflies, and some dipterans. The number 

of shredders collected by Hutchens and Benfield (2000), however, was greater in disturbed 

streams (i.e., streams in areas of gypsy moth defoliation) than in control streams. 

4.1.3.3. Aquatic Plants – Information directly related to the effects of gypsy moth infestations 

on algae in streams is available in the study by Sheath et al. (1986), which was conducted in a 

spring-fed, headwater stream in central Rhode Island from 1979 to 1982. In the first two 

summers, a dense riparian canopy reduced surface light penetration to a range of 5-18% of 

incident radiation, and the stream macroalgal community consisted of only one to four species 

covering from less than 1 to 35% of the stream bottom. In 1981, the surrounding leaf canopy 

was removed by a massive gypsy moth larval outbreak, which increased light penetration at the 

stream surface to 73% and increased the water temperature by 3.7°C. By early August, 

macroalgal cover increased to a peak of 80% of the stream bottom. A less severe gypsy moth 

defoliation in 1982 that did not have a significant impact on light, temperature or macroalgal 

cover from 1979 and 1980. In contrast, investigators on a stream in Shenandoah National Park 

observed no significant changes in periphyton abundance due to defoliation. These investigators 

speculated that many southern Appalachian streams are so low in nutrients that increased 

sunlight penetration alone is not enough to increase algal growth (USDA/FS 1994). 

4.1.3.4. Aquatic Microorganisms – Particularly in small, first-order streams, defoliation by the 

gypsy moth provides increased sunlight at the water surface that may enhance microbial activity 

secondary to an increase in temperature (Sheath et al. 1986). As discussed in Section 4.1.3.2, 

defoliation by the gypsy moth appears to increase the rate of leaf breakdown in streams, which is 

due, in part, to greater microbial conditioning in leaf packs (Hutchens and Benfield 2000). Other 

major increases occur in fecal coliform and fecal streptococci densities in streams during periods 

4-10
 
of maximum defoliation (Corbett and Lynch 1987). These increases might be associated with 

increases in nutrients from water contamination by frass and leaf fragments. 

4-11
 
4.2. EXPOSURE ASSESSMENT 

4.2.1. Overview 

As in the human health risk assessment, the exposure metameter is dictated by the data used to 

formulate the dose-response assessment. Also as in the human health risk assessment, egg mass 

density is the exposure metameter for terrestrial invertebrates and plants because it is the measure 

on which the dose-response assessment is based. Egg mass densities spanning a range from 

5 egg masses/acre to 5,000 egg masses per acre are used to estimate responses in terrestrial plants 

and invertebrates. 

Most wildlife species are not affected directly by exposure to the gypsy moth but are more likely 

to experience indirect effects like changes in habitat or other environmental conditions secondary 

to defoliation. Consequently, the exposure assessment for most wildlife species is almost 

identical to the dose-response assessment for terrestrial plants which is expressed as defoliation 

caused by gypsy moth larvae. For this exposure assessment, categories of defoliation are defined 

normal background defoliation (<30% defoliation), moderate defoliation (30-60% defoliation), 

and high or severe defoliation (>60% defoliation). 

4.2.2. Direct Exposure 

As discussed in Section 4.1, the gypsy moth has a direct impact on terrestrial vegetation and 

certain terrestrial invertebrates, and the direct effects, especially the effects on vegetation, are 

likely to cause indirect effects in other organisms (Section 4.2.3). Like the exposure assessment 

for human health (Section 3.2), the exposure assessment for terrestrial vegetation and terrestrial 

invertebrates can be based on any of several exposure measures, including numbers of egg 

masses per acre, numbers of larvae per acre or tree, or larval mass per acre or tree. The various 

exposure measures are not necessarily correlated and may relate to more than vegetation damage. 

Sample et al. (1996) assayed egg mass density, number of larvae per 50 g dry weight of 

vegetation, and defoliation in sets of stands: control stands with no significant level of gypsy 

moth infestation and no pesticide, stands with significant levels of gypsy moth infestation and no 

pesticide, stands with both gypsy moth infestation and treatment with B.t.k., and stands treated 

with B.t.k. in the absence of significant gypsy moth infestation. Each set of stands consisted of 

six replicates of 50-acre plots in which measurements were made over a 3-year period. As 

illustrated in Figure 4-1, the relationship between egg mass density and the number of larvae per 

unit vegetation is extremely weak and scattered. More recently, Naidoo and Lechowicz (2001) 

published the results of a 13-year study in which they assayed the number of gypsy moth larvae 

per tree in different tree species in a forest in Quebec. Figure 2 in the study shows substantial 

variation in the numbers of larvae in different tree species in the forest. Red oak trees were the 

most heavily infested (up to 250 larvae per tree), and white ash trees were the least infested 

(maximum of four larvae per tree). This study clearly illustrates that gypsy moth larvae feed 

preferentially on different types of vegetation, resulting in substantial variation in infestation 

among tree species. Although larval density may be the most intuitively reasonable measure of 

exposure (i.e., to the gypsy moth larvae), the poor correlation of egg mass density to larval 

density noted in the study by Sample et al. (1996) may be due partly to larval feeding preferences 

4-12
 
or dispersal. A complicating factor, discussed further in Section 4.3, is that larval counts 

themselves do not necessarily predict defoliation uniformly across different species of trees. 

As in the human health risk assessment, the exposure metameter is dictated to some extent by the 

data used to formulate the dose-response assessment. In the human health risk assessment, the 

dose-response assessment is based on egg mass density (see Section 3.3); therefore, egg mass 

density is, by definition, the exposure metameter. As discussed in Section 4.3 and summarized 

in Table 4-1, egg mass densities spanning a range from 5 egg masses/acre to 5,000 egg masses 

per acre are used to estimate responses in terrestrial plants and invertebrates. 

4.2.3. Indirect Exposure 

As summarized in Section 4.1, most wildlife species are not affected directly by exposure to the 

gypsy and are more likely to experience the effects of indirect exposure like changes in habitat or 

other environmental conditions secondary to defoliation. Consequently, the exposure assessment 

for most wildlife species is almost identical to the dose-response assessment for terrestrial plants 

(Section 4.3) – i.e., the assessment is expressed as defoliation caused by gypsy moth larvae. 

Defoliation can be categorized various ways, all of which are largely judgmental. For example, 

Agrawal et al. (2000) define light defoliation as 20-40%, severe defoliation as 40-90%, and 

nearly complete defoliation as 75-100%. The categories used in the previous EIS as well as in 

the study by Gottschalk et al. (1998) are used in the current risk assessment: normal background 

defoliation (<30% defoliation), moderate defoliation (30-60% defoliation), and high or severe 

defoliation (>60% defoliation). 

The use of these somewhat arbitrary categories has only a minimal impact on the current risk 

assessment, which is largely qualitative since the available data do not permit quantitative 

measures of response as a function of defoliation for most wildlife species. This issue is 

discussed further in the risk characterization (Section 4.4). 

4-13
 
4.3. DOSE-RESPONSE ASSESSMENT 

4.3.1. Overview 

The quantitative dose-response assessment for the gypsy moth is illustrated in Figure 4-2. As in 

the human health risk assessment, the dose metameter is egg masses/acre. Quantitative doseresponse assessments can be made for both terrestrial plants and sensitive species of lepidoptera. 

The dose-response assessments for terrestrial plants are based on a relatively simple quantitative 

model for the relationship of egg mass density to defoliation. Three broad categories (sensitive, 

intermediate, and tolerant) are used to characterize the susceptibility of forest stands to gypsy 

moth induced defoliation. Estimated LOAEL values based on 30% defoliation, which is 

considered the lower range of moderate defoliation, are approximately 125, 1000, and 7000 egg 

masses/acre for sensitive, intermediate, and tolerant forest stands, respectively. The 

corresponding NOAEL values, defined as 10% defoliation, are estimated as 12, 20, and 125 egg 

masses/acre for sensitive, intermediate, and tolerant forest stands. 

The gypsy moth may affect some sensitive terrestrial invertebrates, including some species of 

lepidoptera. These effects, however, are less well documented and less well characterized 

compared with the effects on terrestrial plants. Nonetheless, available studies indicate that the 

NOAEL for adverse effects in certain other species of lepidoptera are lower than the NOAEL for 

sensitive forest stands—– i.e., about 6-72 egg masses/acre for some lepidoptera. 

No quantitative dose-response assessment is presented for other groups of organisms—e.g., 

mammals, birds, and soil or aquatic organisms. The impact of gypsy moth exposure on these 

species is most likely to result in indirect effects (i.e., secondary to defoliation). Qualitative 

expressions of risk for these species are presented in the Risk Characterization (Section 4.4.4). 

4.3.2. Terrestrial Plants 

In terms of assessing direct effects of the gypsy moth on terrestrial vegetation, the most common 

endpoint used to express damage is defoliation. As discussed in Section 4.1.2.5, numerous 

factors, many of which are interrelated, influence the level of damage that gypsy moth larvae may 

cause to a forest stand or region. Several models are useful for quantifying the likely levels of 

defoliation (e.g., Davidson et al. 2001; Gansner et al. 1985; Gottschalk et al. 1998; Gribko et al. 

1995; Liebhold et al. 1993; Montgomery 1990;Weseloh 1996a; Williams et al. 1991). 

Furthermore, the USDA Forest Service developed an expert system, GypsES, to facilitate the use 

of modeling in the management of gypsy moth programs (Gottschalk et al. 1996; Williams et al. 

1997; http://www.fs.fed.us/na/morgantown/fhp/gypses /gypmain.htm). 

The common exposure factor for all of these models is egg mass density. As discussed in 

Section 3.1, egg mass density is usually measured during the fall to predict damage in the 

following season. This indicator allows individuals involved in gypsy moth control programs 

enough time to plan an intervention strategy before damage occurs. The models, some of which 

are extremely complex, incorporate several factors, in addition to egg mass density, that affect 

defoliation—e.g., types of vegetation, terrain characteristics, and various geographical 

characteristics. For example, Weseloh (1996a) developed a 23-parameter model that 

4-14
 
incorporates egg mass density, latitude, longitude, elevation, drainage, factors for the history of 

defoliation over a previous 4-year period, and 12 parameters for interaction terms. The model 

was developed using data from defoliation patterns in Connecticut from 1987 to 1994. When 

used retrospectively on defoliation rates from 1974 to 1986, the model was reasonably accurate 

in predicting years of heavy defoliation (correlations of up to 0.8) but less accurate in predicting 

years with relatively little defoliation (correlations on the order of 0.2). 

For the current risk assessment, the relatively simple four-parameter model developed by 

Davidson et al. (2001) seems most useful for describing key factors that will have an impact on 

defoliation: 

b1 b2 b3 (Y DEF = (BA /BA ) × BA × ln(EM) × Y × e S S T p 

where: 

Y = number of years since start of outbreak 

DEFS = percent stand defoliation at time Y of outbreak 

2 BAS = basal area of sensitive species in stand (m /ha)

2 BAP = basal area of pine species in stand (m /ha)

2 BAT = total basal area in stand (m /ha)

EM = egg masses per hectare 

and (, b1, b2, and b3 are model parameters. Based on defoliation and egg count data collected 

over 4 years (1992 through 1995) from seven pine-oak stands and nine pine-sweetgum stands in 

Maryland, Davidson et al. (2001) estimated the following values for the model parameters: 

0.2226 for b1, -0.2684 for b2, 2.0792 for b3, and -0.5781 for (. Notably, the parameters 

associated with the ratio of sensitive tree species (BA /BA ) and egg mass density (EM) are S T

positive indicating that damage increases as these model components increase. The parameters 

associated with the basal area of pine species (BA ) as well as the exponential function for years p

(Y since the start of the outbreak (e ) are negative. In other words, an increase in the density of 

pine species (trees generally not favored by the gypsy moth) will tend to reduce defoliation and 

the outbreak will subside over time. All of these factors in the model are qualitatively consistent 

with the behavior of gypsy moth infestations (see Section 4.1.2.5). 

The model developed by Davidson et al. (2001), though relatively simple, is still 

multidimensional, which means the estimates of defoliation depend highly on site specific 

factors. Any number of defoliation estimates could be made based on varying any of the input 

variables in the model by Davidson et al. (2001). For this risk assessment, three general types of 

forest stands are considered: sensitive, intermediate, and tolerant. Each stand is assumed to have 

2 a total basal area of 15 m /ha. This is somewhat arbitrary but since the total basal area is only 

used as a normalizing factor on the basal area of sensitive species, this assumption has no impact 

2 on the model. The basal surface area for sensitive species is taken as 13 m /ha for sensitive

2 2 stands, 6 m /ha for intermediate stands, and 2 m /ha for tolerant stands. The basal area for pine is 

2 2 2 taken as 0.25 m /ha for sensitive stands, 1.5 m /ha for intermediate stands, and 3 m /ha for

tolerant stands. The percent defoliation in all stands is calculated for 1 year from the initial time 

4-15
 
of the outbreak. Egg mass density is modeled over a range from 1 egg mass/ha (approximately 

0.4 egg masses per acre) to 25,000 egg masses/ha (approximately 10,000 egg masses per acre). 

Again, despite the arbitrary nature of these ranges and assumptions, they reflect the variability of 

and responses among the stands considered in the Davidson et al. (2001) publication. 

The variability of responses in the three different stand types is illustrated in Figure 4-2. The 

curved lines indicate the percent defoliation expected in sensitive, intermediate, and tolerant 

stands over the range of egg mass densities considered. The two thick horizontal lines represent 

breakpoints for the classifications of defoliation discussed in Section 4.2.3—i.e., normal 

background defoliation (<30% defoliation), moderate defoliation (30-60% defoliation), and high 

or severe defoliation (>60% defoliation). Sensitive stands are likely to show evidence of low 

level intermediate defoliation—i.e., 30% defoliation—at an egg mass density of about 125 egg 

masses/acre. Comparable values for intermediate and tolerant stands are about 1000 egg 

masses/acre and 7000 egg masses/acre, respectively. For sensitive stands, the egg mass density 

associated with 50% defoliation is about 600 egg masses/acre and this estimate is consistent with 

field observations for sensitive stands (Montgomery 1990). 

Risks to wildlife species from most agents used to control the gypsy moth are based on NOAEL 

values (no observed adverse effect levels) and LOAEL values (lowest observed adverse effect 

levels). As discussed in Section 4.2.3, 30% defoliation is used in this risk assessment as a border 

value between background and moderate defoliation. Thus, the egg mass densities of 125, 1000, 

and 7000 egg masses/acre for sensitive, intermediate, and tolerant stands, respectively, are 

essentially LOAEL values—the lowest egg mass density that might be associated with a level of 

defoliation classified as moderate or a minimal adverse effect. 

The model by Davidson et al. (2001) as well as other models for gypsy moth defoliation are nonthreshold. In other words, any level of exposure is assumed to carry some risk. Thus, the 

selection of a functional NOAEL is somewhat arbitrary. Following the general approach used to 

estimate a benchmark dose (U.S. EPA 2001), the defoliation rate of 10% is used as a functional 

NOAEL. Based on the dose-response curves illustrated in Figure 4-2, these NOAEL values are 

egg mass densities of about 12 egg masses/acre for sensitive stands, 20 egg masses/acre for 

intermediate stands, and approximately 125 egg masses/acre for tolerant stands. 

These NOAEL and LOAEL estimates are at best crude approximations of egg mass densities 

associated with levels of defoliation that might be considered essentially benign (NOAELs) or 

capable of causing detectable damage (LOAELs) in various forest stands. The primary use of 

these values is to provide a basis for comparing risks associated with exposure to the gypsy moth 

to risks associated with exposure to agents used to control the gypsy moth. 

4.3.3. Terrestrial Invertebrates 

The impact of gypsy moth exposure on terrestrial invertebrates cannot be quantified. As 

discussed in Section 4.1.2.4, a few relatively recent studies demonstrate that exposure to gypsy 

moth larvae during an outbreak may adversely affect some other lepidopterans (Sample et al. 

4-16
 
1996; Work and McCullough 2000). Furthermore, the study by Redman and Scriber (2000) 

suggests that at least some of these effects could be related directly to gypsy moth exposure 

rather than to effects secondary to gypsy moth-induced defoliation. 

Although the data on invertebrates are limited in terms of defining a quantitative exposureresponse relationship, the study by Sample et al. (1996) defines an apparent NOAEL of 6-72 egg 

masses/acre for effects on tiger moths. The corresponding LOAEL is 95-468 egg masses/acre 

and is associated with decreases in the abundance of tiger moth larvae and adults. These values 

are illustrated in Figure 4-2 as the mean of each range rounded to one significant place—i.e, a 

value of 40 egg masses/acre for the NOAEL and 300 egg masses/acre for the LOAEL. 

4.3.4. Other Species 

As discussed in Section 4.2.3, other species may be affected by gypsy moth infestations 

secondary to defoliation. These observations, as summarized in Section 4.1 (Hazard 

Identification), are essentially qualitative—i.e., the effects were observed in the field or are based 

on plausible assumptions in cases of severe gypsy moth outbreaks and extensive defoliation. 

Thus, no quantitative dose-response assessment is proposed for species that are indirectly 

affected, and the plausible responses are discussed qualitatively in the risk characterization 

(Section 4.4.4). 

4-17
 
4.4. RISK CHARACTERIZATION 

4.4.1. Overview 

The best documented and most obvious effect of the gypsy moth will be defoliation of terrestrial 

vegetation, particularly in forest stands in which sensitive species of trees predominate. In some 

respects, the risk characterization for terrestrial vegetation is essentially a restatement of the 

hazard identification. In other words, the effects of gypsy moth larvae on forests is extremely 

well documented and relatively well understood. In sensitive forest stands—i.e., stands in which 

oak, birch, and other favored species predominate —gypsy moth larvae can cause substantial 

defoliation. In forest stands in which tree species that are not favored by gypsy moth larvae 

predominate —e.g., hemlock, various types of pine, black locust and white ash—even relatively 

high exposures, as measured by egg mass density, may not result in substantial defoliation. The 

risk assessment for direct effects on forests should be at least qualitatively influenced by the 

current range of the gypsy moth, which has not yet extended to some forests in the southeast that 

may be among the most sensitive to the gypsy moth. Thus, unless measures to contain the gypsy 

moth are successful, these southeastern forests may suffer serious damage in future infestations. 

Some other lepidopteran species also may be directly affected by exposure to the gypsy moth. 

Most studies, however, suggest that substantial adverse effects in terrestrial insects are unlikely 

and effects in some insect species, including some other lepidoptera, may be beneficial. 

Because the gypsy moth may substantially damage some forests in severe infestations or 

outbreaks, secondary effects in other species of wildlife are plausible. Reductions in populations 

of squirrels, mice, and other mammals which may be sensitive to changes in the availability of 

acorns are likely. Substantial adverse effects on other groups of animals – i.e., birds, reptiles, and 

aquatic species – cannot be ruled out but have not been convincingly demonstrated. 

4.4.2. Direct Effects 

4.4.2.1. Terrestrial Plants – A quantitative summary of the risk characterization for forest stands 

is presented in Table 4-1. Risks are expressed as hazard quotients and estimates of percent 

defoliation for three classes of forest stands: sensitive, intermediate, and tolerant. As discussed 

in Section 4.3.2, these classifications are intended to reflect, albeit crudely, differences in 

susceptibility of various forest stands to the effects of gypsy moth exposure, which is predicated 

on the feeding preferences of gypsy moth larvae. The specific types of trees favored and not 

favored by the gypsy moth are discussed in Section 4.1.2.5.2. The NOAEL values and 

quantitative estimates of defoliation are based on the dose-response model proposed by Davidson 

et al. (2001). Although the dose-response model is relatively simple, it does appear to reflect the 

variables that have the most significant impact on the sensitivity of various forest stands to 

defoliation by gypsy moth larvae. The four categories of infestations used in the dose-response 

assessment are based on egg mass densities of 5, 50, 500, and 5000 egg masses/acre. These 

categories generally encompass the range of egg mass densities reported in the literature for 

infestations ranging in degree from negligible to severe and are similar to the categories used in 

the human health risk assessment (Table 3-3). The hazard quotients presented in Table 4-1 will 

be used primarily in the comparative risk assessment, which is a separate document that provides 

4-18
 
a quantitative comparison of the risks for each of the various agents used to control the gypsy 

moth as well as the risks posed by the gypsy moth itself. 

In some respects, the risk characterization for terrestrial vegetation is essentially a restatement of 

the hazard identification. In other words, as discussed in Section 4.1.2.5, the effects of gypsy 

moth larvae on forests is extremely well documented and relatively well understood. In sensitive 

forest stands—i.e., stands in which oak, birch, and other favored species predominate —gypsy 

moth larvae can cause substantial defoliation. In forest stands in which tree species that are not 

favored by gypsy moth larvae predominate —e.g., hemlock, various types of pine, black locust 

and white ash—even relatively high exposures, as measured by egg mass density, may not result 

in substantial defoliation. 

The quantitative risk characterization focuses on defoliation; however there are likely to be other 

effects of gypsy moth exposure. In some cases, extensive defoliation will result in tree mortality. 

In general, however, tree mortality is likely to be relatively low (on the order of 10-20%), 

although higher rates of mortality (up to about 50%) are possible when sensitive species are 

subject to multiple years of defoliation (see Section 4.1.2.5.1). Defoliation and tree mortality can 

lead to secondary effects of exposure on animals (see Section 4.4.3). Likewise, non-target 

vegetation may be subject to secondary effects of exposure, such as increases in understory 

growth (see Section 4.1.2.5.4). The extent to which the secondary effects are considered 

beneficial or detrimental depends to some extent on forest management objectives which are 

beyond the scope of this risk assessment. 

The risk assessment for direct effects on forests should be at least qualitatively influenced by the 

range of the gypsy moth. That range has not yet extended to some forests that may be among the 

most sensitive to gypsy moth exposure. Many forests in the southeast and midwest are populated 

with a high proportion of tree species that are favored by the gypsy moth. Unless measures to 

contain the gypsy moth are successful, these regions may suffer serious damage from future 

infestations by the gypsy moth (Liebhold and McManus 1999). 

4.4.2.2. Terrestrial Invertebrates – There is plausible concern regarding direct effects of the 

gypsy moth on some lepidopteran species. Nonetheless, few studies support this concern relative 

to the large number of studies regarding effects on terrestrial plants. As summarized in 4.3.3, the 

study by Sample et al. (1996) suggests a NOAEL of about 40 egg masses/acre for Arctiidae 

larvae and adults in terms of abundance and species richness. The direct effect of the gypsy 

moth, however, involved only lepidoptera from a single family, Arctiidae, which includes the 

tiger moths. No effects were seen in other lepidopteran or non-lepidopteran species; nonetheless, 

the effects observed on Arctiidae adults and larvae were statistically significant. Based on the 

approximate NOAEL of 40 egg masses/acre, this family of lepidoptera would still be less 

sensitive to gypsy moth larvae than most forest stands. 

It is difficult to assess the extent to which other lepidopteran or non-lepidopteran groups might 

be affected by exposure to gypsy moth larvae. Redman and Scriber (2000) report several 

4-19
 
mechanisms associated with the adverse effect of gypsy moth larvae on the northern tiger 

swallowtail butterfly, Papilio canadensis. There are, however, no other field studies that 

suggest the plausibility of substantial adverse effects in most insect species during gypsy moth 

infestations. In addition, gypsy moth induced defoliation may be beneficial to some butterfly 

species (Schweitzer 1988) or will have no effect on most other insect species (Sample et al. 

1996). 

4.4.3. Indirect Effects 

4.4.3.1. Terrestrial Mammals – There is only limited information regarding the potential effects 

of gypsy moth infestations on mammalian wildlife. Adverse effects on reproduction or 

nutritional status were not observed in black bears after exposure to the gypsy moth during an 

outbreak that caused substantial mortality in oak trees and decreased acorn production (Vaughan 

and Kasbohm 1993). As noted in Section 4.1.2.1, bears adjusted for the decrease in acorn 

production by switching to alternate foods. It is not clear, however, that all mammals would 

adapt to a severe decrease in hard mast production. Consequently, there is plausible concern 

about the potential for adverse effects (reductions in populations) in squirrels, mice, and other 

mammals which may be sensitive to changes in the availability of acorns. 

4.4.3.2. Birds – The effects of gypsy moth exposure on birds appear to be highly variable, with 

reports of both detrimental effects as well as beneficial effects (see Section 4.1.2.2). While the 

available data are not amenable to quantitative analyses, it would appear that these mixed effects 

are likely to be most pronounced during severe infestations. 

4.4.3.3. Amphibians and Reptiles – There are no studies that clearly indicate adverse effects in 

either amphibians or reptiles after exposure to the gypsy moth. As discussed in Section 4.1.2.3, 

substantial defoliation of mast producing trees might adversely affect snakes that consume small 

mammals (e.g., squirrels and mice), the populations of which might decrease due to mast failure 

(see Section 4.4.3.1). 

4.4.3.4. Terrestrial Microorganisms – There is no information regarding the effects of the gypsy 

moth or gypsy moth defoliation on terrestrial microorganisms. Intuitively, it seems reasonable 

that soil microbial activity would increase in response to defoliation as a result of the subsequent 

increase in ground temperature and nutrient load that would result from increases in litter and 

frass production (see Section 4.1.2.6). These effects, although not reported in the literature for 

soil microbial activity, are reported for aquatic microorganisms (see Section 4.1.3.4). 

4.4.3.5. Aquatic Organisms – There is very little information to indicate that gypsy moth 

infestations cause substantial adverse effects on aquatic organisms (see Section 4.1.3). Stream 

microorganisms are likely to be affected directly by gypsy moth infestations due to the potential 

increase of microbial activity in forest streams. The increased activity is mostly like due to the 

increased nutrient loading of streams which results from defoliation and larval frass. Although 

Hutchens and Benfield (2000) suggest that the increase in activity might cause a food deficit for 

aquatic insects that shred decomposing leaves, the investigators found that the population of such 

4-20
 
organisms (i.e., caddisflies, stoneflies, and some dipterans) were higher in streams in areas of 

gypsy moth defoliation compared with streams in the control areas. Because of increased light 

and water temperature secondary to defoliation, algal and aquatic macrophyte growth is likely to 

be increased (see Section 4.1.3.3), which might increase productivity in some streams but 

adversely affect water quality and habitat characteristics in other streams. 

4-21
 
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5-17
 
Table 3-1. Individuals with skin responses to the gypsy moth in two communities (data from 

Tuthill et al. 1984 except as noted) 

Factor Medway Lunenburg 

Average egg masses/acre 32 b 3809 b 

All groups combined #Responding/#Exposed 

Total with rash during week before infestation 6/557 7/508 

Total with rash during first 7 days after larvae emerge 9/557 50/508 

P-value for pre- vs post-emergence difference b 0.30 8×10-10 

Subgroups 

Age 0-12 years 2/84 c 13/69 d 

(2.3%) (19%) 

Age 13-59 years 7/407 c 35/342 d 

(1.7%) (10%) 

Age > 59 years 0/66 c 2/97 d 

(0%) (2.1%) 

Larval contact 

Touched larvae 8.3% 31.4% 

Rash where individuals were touched or crawled on by 29.0% 82.0% 

larvae 

a O'Dell 1994


b Based on Fisher Exact Test. See text for discussion.
 

c

 No statistically significant difference among age groups.
 

d

 Response in 0-12 years significantly greater than 13-59 year group (p=0.039) and >59 year
 

group (p=0.000245). Response in 13-59 year group significantly greater than >59 year group
 

(p=0.0048). All comparisons based on Fisher Exact Test. See text for discussion.
 

Tables - 1 
Table 3-2: Statistical Analyses of Epidemiology Data from Table 3-1. 

Age Group Back- Potency p-value Lower 95% Confidence Interval on 

ground (proportion Egg masses/acre 

responding 

per egg 

mass/acre) 1% 10% 

0-12 years 0.022 4.89e-005 0.00041 128 1336 

13-59 years 0.016 2.40e-005 <0.0001 304 3185 

>59 years 0.00 5.52e-006 0.15 697 >11,427 

All Groups 0.013 2.37e-005 <0.0001 327 3432 

Combined 

a

 Dose associated with a given extra risk – i.e., 1% or 10%. 

Tables - 2 
Table 3-3. Adverse human health effects for members of the general public associated with 

exposure to the gypsy moth 

NOAEL: 32 egg masses/acre 

Level of Infestation Exposure 

(egg masses/acre) 

Hazard Quotient a Upper Limit on 

Extra Risk b 

Sparse 50 1.6 1.4% 

Moderate >50-500 >1.6 - 16 >1.4% - 2.5% 

Heavy >500-5000 >16 - 156 >2.5% - 12% 

Extreme >5,000 - 20,000 >156 - 625 >12% - 38% 

LOAEL: 1336 egg masses/acre 

Level of Infestation Exposure 

(egg masses/acre) 

Hazard Quotient a Upper Limit on 

Extra Risk b 

Sparse 50 0.04 Same as above 

Moderate >50-500 >0.04 - 0.4 

Heavy >500-5000 >0.4 - 4 

Extreme >5,000 - 20,000 >4 - 15 

a

 Calculated as the exposure in egg masses/acre divided by the NOAEL or LOAEL. 

b

 Based on the dose-response model summarized in Table 3-2 using data on all groups 

combined. 

Tables - 3 
Table 4-1: Summary of quantitative risk characterization for forest stands. 

Forest Stands 

Sensitive Stands Intermediate Stands Tolerant Stands 

NOAEL 12 20 125 

Egg masses/acre Hazard Quotients a 

5 0.4 0.25 0.04 

50 4 2.5 0.4 

500 40 25 4 

5,000 400 250 40 

Percent Defoliation b 

5 5.3% 2.8% 1.8% 

50 21.0% 11.0% 6.9% 

500 46.0% 24.0% 14.0% 

5000 83.0% 43.0% 29.0% 

a

 Egg mass density divided by NOAEL
 

b

 Based on dose-response model of Davidson et al. (2001) detailed in Section 4.3.2.
 

Tables - 4 
Figure 2-1: Frequency of defoliation by the gypsy moth from 1975 to 2002 (Source: 

http://www.fs.fed.us/ne/morgantown/4557/gmoth/defoliation/freq75_02.jpg) 

Figures - 1
 
Figure 4-1: Relationship of egg mass density to number of larvae per 50 g 

dry weight of vegetation (Data from Sample et al. 1996). 

Figures - 2
 
Figure 4-2: Summary of Exposure-Response Assessment (see text for details) 

Figures - 3
 
Appendix M 

Risk Comparison 

Figure M-1. Ropes were used to climb trees, to treat them for gypsy moths in the 

1930s. 

SERA TR 04-43-05-08e 

Control/Eradication Agents for the Gypsy Moth ­

Risk Comparison 

– Final Report 

Prepared for:
 

USDA, Forest Service
 

Forest Health Protection
 

GSA Contract No. GS-10F-0082F
 

USDA Forest Service BPA: WO-01-3187-0150
 

USDA Purchase Order No.: 43-3187-1-0269
 

Task No. 5
 

Submitted to:
 

Hank Appleton, COTR
 

Forest Health Protection Staff
 

USDA Forest Service
 

Rosslyn Plaza Building C, Room 7129C
 

1601 North Kent Street
 

Arlington, VA 22209
 

Prepared by Patrick Durkin
 

Submitted by:
 

Syracuse Environmental Research Associates, Inc.
 

5100 Highbridge St., 42C
 

Fayetteville, New York 13066-0950
 

Telephone: (315) 637-9560
 

Fax: (315) 637-0445
 

E-Mail: SERA_INC@msn.com
 

Home Page: www.sera-inc.com
 

October 31, 2004 (Original Document)
 

July 20, 2007 and September 10, 2007 (Editorial Corrections)
 
TABLE OF CONTENTS
 

TABLE OF CONTENTS ....................................................... ii
 

LIST OF TABLES ............................................................ iii
 

LIST OF FIGURES ........................................................... iii
 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iv
 

COMMON UNIT CONVERSIONS AND ABBREVIATIONS . . . . . . . . . . . . . . . . . . . . . . . . . vi
 

CONVERSION OF SCIENTIFIC NOTATION .................................... vii
 

EXECUTIVE SUMMARY .................................................... viii
 

1. INTRODUCTION ......................................................... 1-1
 

2. COMPARATIVE PROGRAM DESCRIPTION ................................. 2-1
 

2.1. OVERVIEW ...................................................... 2-1
 

2.2. CONTROL AGENTS .............................................. 2-1
 

2.3. APPLICATION RATES ............................................. 2-2
 

2.4. USE STATISTICS ................................................. 2-3
 

3. HUMAN HEALTH RISK ASSESSMENT ..................................... 3-1
 

3.1. HAZARD IDENTIFICATION ....................................... 3-1
 

3.2. EXPOSURE ASSESSMENT ........................................ 3-6
 

3.3. DOSE-RESPONSE ASSESSMENT .................................. 3-11
 

3.4. RISK CHARACTERIZATION ...................................... 3-18
 

4. ECOLOGICAL RISK ASSESSMENT ......................................... 4-1
 

4.1. HAZARD IDENTIFICATION ....................................... 4-1
 

4.2. EXPOSURE ASSESSMENT ........................................ 4-8
 

4.3. DOSE-RESPONSE ASSESSMENT .................................. 4-12
 

4.4. RISK CHARACTERIZATION ...................................... 4-18
 

5. REFERENCES ........................................................... 5-1
 

ii 
LIST OF TABLES
 

Table 2-1: Total use of control agents by numbers of acres treated between 1995 and 2003........Tables 1
 

Table 3-1: Comparative hazard identification for potential effects in humans..............................Tables 2
 

Table 3-2: Comparative exposure assessment for human health effects.......................................Tables 3
 

Table 3-3: Comparative dose-response assessment for human health effects...............................Tables 4
 

Table 3-4: Comparative risk characterization for human health effects........................................Tables 5
 

Table 4-1: Comparative hazard identification for potential effects in non-target species.............Tables 6
 

Table 4-2: Comparative exposure assessment for ecological effects............................................Tables 7
 

Table 4-3: Comparative dose-response assessment for ecological effects....................................Tables 8
 

Table 4-4: Comparative risk characterization for terrestrial species..............................................Tables 9
 

Table 4-5: Comparative risk characterization for aquatic species................................................Tables 10
 

LIST OF FIGURES
 

Figure 3-1: Risk comparison for potential human health effects.................................................Figures 1
 

Figure 4-1: Risk comparison for potential effects in terrestrial species........................................Figures 2
 

Figure 4-2: Risk comparison for potential effects in aquatic species............................................Figures 3
 

NOTE: Tables and Figures are placed after Section 5, References. 

iii 
ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

a.i. active ingredient 

Ach acetylcholine 

AChE acetylcholinesterase 

AEL adverse-effect level 

APHIS Animal and Plant Health Inspection Service 

ARS Agricultural Research Station 

BCF bioconcentration factor 

B.t.k. Bacillus thuringiensis var. kurstaki 

BIU Billions of international units 

bw body weight 

4-CA 4-chloroaniline 

ChE pseudo-cholinesterase 

CNS central nervous system 

DFB diflubenzuron 

EC x concentration causing X% inhibition of a process 

EIS environmental impact statement 

FH Forest Health 

FS Forest Service 

FTU forestry toxic units 

HQ hazard quotient 

IRIS Integrated Risk Information System 

IU international units 

kg kilogram 

L liter 

LdNPV gypsy moth (Lymantria dispar) nucleopolyhedrosis virus 

lb pound 

LC50 lethal concentration, 50% mortality 

LD50 lethal dose, 50% mortality 

LD95 lethal dose, 95% mortality 

LOAEL lowest-observed-adverse-effect level 

m meter 

M male 

mg milligram 

mg/kg/day milligrams of agent per kilogram of body weight per day 

mL milliliter 

MSDS material safety data sheet 

NCI National Cancer Institute 

NOAEL no-observed-adverse-effect level 

NOEC no-observed-effect concentration 

NOEL no-observed-effect level 

iv 
ACRONYMS, ABBREVIATIONS, AND SYMBOLS (continued) 

NRC National Research Council 

OPP Office of Pesticide Programs 

ORD Office of Research and Development 

OTS Office of Toxic Substances 

PIB polyhedral inclusion body 

ppm parts per million 

PVC polyvinyl chloride 

RfD reference dose 

RQ risk quotients 

UF uncertainty factor 

U.S. United States 

U.S. EPA U.S. Environmental Protection Agency 

USDA United States Department of Agriculture 

v 
COMMON UNIT CONVERSIONS AND ABBREVIATIONS
 

To convert ... Into ... Multiply by ... 

acres hectares (ha) 0.4047 

acres square meters (m ) 2 4,047 

atmospheres millimeters of mercury 760 

centigrade Fahrenheit 1.8 EC + 32 

centimeters inches 0.3937 

cubic meters (m ) 3 liters (L) 1,000 

Fahrenheit centigrade 5/9 (EF-32) 

feet per second (ft/sec) miles/hour (mi/hr) 0.6818 

gallons (gal) liters (L) 3.785 

gallons per acre (gal/acre) liters per hectare (L/ha) 9.34 

grams (g) ounces, (oz) 0.03527 

grams (g) pounds, (oz) 0.002205 

hectares (ha) acres 2.471 

inches (in) centimeters (cm) 2.540 

kilograms (kg) ounces, (oz) 35.274 

kilograms (kg) pounds, (lb) 2.2046 

kilograms per hectare (hg/ha) pounds per acre (lb/acre) 0.892 

kilometers (km) miles (mi) 0.6214 

liters (L) cubic centimeters (cm ) 3 1,000 

liters (L) gallons (gal) 0.2642 

liters (L) ounces, fluid (oz) 33.814 

miles (mi) kilometers (km) 1.609 

miles per hour (mi/hr) cm/sec 44.70 

milligrams (mg) ounces (oz) 0.000035 

meters (m) feet 3.281 

ounces (oz) grams (g) 28.3495 

ounces per acre (oz/acre) grams per hectare (g/ha) 70.1 

ounces per acre (oz/acre) kilograms per hectare (kg/ha) 0.0701 

ounces fluid cubic centimeters (cm ) 3 29.5735 

pounds (lb) grams (g) 453.6 

pounds (lb) kilograms (kg) 0.4536 

pounds per acre (lb/acre) kilograms per hectare (kg/ha) 1.121 

2 pounds per acre (lb/acre) mg/square meter (mg/m ) 112.1 2 pounds per acre (lb/acre) ug/square centimeter (ug/cm ) 11.21 

pounds per gallon (lb/gal) grams per liter (g/L) 119.8 2 2 square centimeters (cm ) square inches (in ) 0.155 2 2 square centimeters (cm ) square meters (m ) 0.0001 2 2 square meters (m ) square centimeters (cm ) 10,000 

yards meters 0.9144 

Note: All references to pounds and ounces refer to avoirdupois weights unless otherwise 

specified. 

vi 
CONVERSION OF SCIENTIFIC NOTATION
 

Scientific Decimal Verbal 

Notation Equivalent Expression 

1 x 10 -10 0.0000000001 One in ten billion 

1 x 10 -9 0.000000001 One in one billion 

1 x 10 -8 0.00000001 One in one hundred million 

1 x 10 -7 0.0000001 One in ten million 

1 x 10 -6 0.000001 One in one million 

1 x 10 -5 0.00001 One in one hundred thousand 

1 x 10 -4 0.0001 One in ten thousand 

1 x 10 -3 0.001 One in one thousand 

1 x 10 -2 0.01 One in one hundred 

1 x 10 -1 0.1 One in ten 

1 x 10 0 1 One 

1 x 10 1 10 Ten 

1 x 10 2 100 One hundred 

1 x 10 3 1,000 One thousand 

1 x 10 4 10,000 Ten thousand 

1 x 10 5 100,000 One hundred thousand 

1 x 10 6 1,000,000 One million 

1 x 10 7 10,000,000 Ten million 

1 x 10 8 100,000,000 One hundred million 

1 x 10 9 1,000,000,000 One billion 

1 x 10 10 10,000,000,000 Ten billion 

vii 
EXECUTIVE SUMMARY
 

OVERVIEW
 

The current document provides a comparison of the risks posed by the gypsy moth itself to the 

risks posed by the different control agents as well as a comparison of risks among the various 

control agents. The agents used in control programs include Bacillus thuringiensis var. kurstaki 

(B.t.k.), the gypsy moth nucleopolyhedrosis virus (LdNPV), diflubenzuron, tebufenozide, DDVP, 

and disparlure. 

The gypsy moth itself poses the clearest risks in both the human health and ecological risk 

assessments. If the gypsy moth is not controlled, population outbreaks will occur and humans 

will be exposed to large numbers of gypsy moth larvae. If this occurs, a substantial number of 

individuals will experience skin irritation that is sufficiently severe to warrant medical attention. 

No more serious effects are likely. Ecologically, the gypsy moth will clearly damage some 

terrestrial vegetation and may directly affect some other species of moths. Because of the 

obvious importance of vegetation to the existence and habitat of most animals, defoliation by the 

gypsy moth will have numerous secondary effects. 

Most of the control agents also pose risks and raise concerns, the nature and certainty of which 

are highly variable. In applications used to control the gypsy moth, B.t.k. is associated with 

irritant effects in humans; however, the severity of these effects appears to be less than those 

associated with exposure to the gypsy moth itself. The potential for B.t.k. to cause more serious 

human health effects is considered but appears to be remote. B.t.k. may also cause adverse 

effects in nontarget Lepidoptera. Concern for this effect is heightened because some of the 

Lepidoptera that may be adversely affected include at least on endangered species. 

Diflubenzuron does not appear to present any substantial risk to human health, and this 

assessment encompasses 4-chloroaniline, a potential carcinogen that is formed in the degradation 

of diflubenzuron. Diflubenzuron, however, is a rather nonspecific insecticide and is likely to 

impact both terrestrial and aquatic arthropods. Tebufenozide is a somewhat more specific 

insecticide but is used at higher application rates that may lead to high exposures in some 

terrestrial mammals. The likelihood of observing adverse effects, however, is unclear. 

Tebufenozide may also impact some nontarget moths and butterflies but should not adversely 

affect any aquatic species. Although DDVP is a broad spectrum insecticide and can be highly 

toxic to humans, adverse human health and ecological effects are not expected under normal 

conditions of use. If DDVP is improperly handled, exposures could substantially exceed prudent 

levels. For disparlure, exposure estimates for aquatic invertebrates approach a level of concern. 

More significantly, there is substantial uncertainty in the risk characterization of disparlure 

because of the limited acute toxicity data, the lack of chronic toxicity data, and the high 

likelihood that many species will be exposed to this compound. 

Unlike all of the other agents considered in this risk assessment, there is no basis for asserting 

that the use of LdNPV to control or eradicate gypsy moth populations is likely to cause any 

adverse effects in any species other than the gypsy moth. 

viii 
 

PROGRAM DESCRIPTION 

The USDA control programs for the gypsy moth are intended to limit damage to forests that can 

be substantially impacted by gypsy moth outbreaks. Two biological agents that are pathogenic to 

the gypsy moth are used in broadcast applications: B.t.k. and LdNPV. In addition, three chemical 

agents are used in broadcast applications: diflubenzuron, tebufenozide, and disparlure. 

Diflubenzuron and tebufenozide are both insecticides, and, as discussed in subsequent sections of 

the this document, are quite similar with respect to their toxicological properties. The major 

difference between the two is that application rates for tebufenozide are higher than those for 

diflubenzuron and this is a controlling factor in the comparative risk assessment for these two 

agents. Disparlure is a gypsy moth pheromone that is used in broadcast applications to disrupt 

mating and in population monitoring programs to attract the male gypsy moth to sampling traps. 

In the past, disparlure was used in a slow-release flake formulation. DDVP is not used in 

broadcast applications and is used only as a PVC formulated product in milk carton traps used in 

mass trapping operations. 

The USDA adopted various intervention strategies roughly categorized as suppression, 

eradication, and slow-the-spread. Suppression programs have relied predominantly on B.t.k. and 

diflubenzuron. Slow-the-Spread programs rely predominantly on the use of disparlure flakes and 

secondarily on B.t.k. applications. Eradication efforts rely predominantly on B.t.k. NPV is used 

in all three strategies but is used on very few acres relative to B.t.k., diflubenzuron, and disparlure 

flakes. 

HUMAN HEALTH RISK ASSESSMENT 

Hazard Identification – The gypsy moth, B.t.k., and LdNPV are similar not only because they 

are biological agents but also because the primary effect associated with each agent is irritation. 

The gypsy moth causes more pronounced and severe irritation relative to either B.t.k. or LdNPV. 

Of the chemical agents used in gypsy moth control programs, diflubenzuron and tebufenozide are 

similar to each other in that both cause adverse effects on blood. The risk assessment of 

diflubenzuron is somewhat more involved than that of tebufenozide because diflubenzuron is 

degraded to 4-chloroaniline, a compound that is classified as a carcinogen. DDVP and 

disparlure, the other two chemicals used in gypsy moth control programs, have toxicologic 

profiles that are very different from each other as well as diflubenzuron or tebufenozide. DDVP 

is a a well-characterized neurotoxin which was studied extensively in mammals. Disparlure, an 

insect attractant, was not tested extensively for toxicological effects in mammals. 

Exposure Assessment – The exposure assessments of the biological agents differ substantially 

from those of the chemical agents in terms of how the exposures are expressed. Because of the 

available exposure and toxicity data, different measures of exposure are used for each of the 

biological agents – i.e., the gypsy moth, B.t.k., and LdNPV. For the chemical agents, all 

exposure assessments are based on the amount or concentration of the chemical to which an 

individual or population might be exposed via ingestion, dermal contact, or inhalation. 

Differences among the chemical agents are dictated largely by differences in how the chemicals 

are used and, to a lesser extent, on the available toxicity data. 

ix 
A very different set of exposure assessments is conducted for each of the biological agents. Both 

B.t.k. and LdNPV may also be applied in broadcast applications and the routes of plausible 

exposure are the same as those for the chemicals applied in broadcast applications – i.e., oral, 

dermal, and inhalation. For B.t.k., however, the most directly relevant data used to characterize 

risk are based on actual applications of B.t.k. formulations where exposure is best characterized 

as an application rate. For the assessment of the potential for serious adverse effects, exposures 

are measured in colony forming units (cfu). LdNPV differs from all of the other agents in that no 

clear hazard potential can be identified; consequently, the most meaningful measure of exposure 

is, in some respects, moot. Those exposures that are quantified in the human health risk 

assessment for LdNPV are based on the mass of the formulation, Gypchek. Exposures to the 

gypsy moth itself are based on an indirect measure of exposure, egg masses/acre, because this is 

the expression of exposure that is used in the dose response assessment. 

Differences in the exposure assessments among the chemicals used in USDA programs primarily 

reflect differences in how the chemicals are applied, what routes of exposure are most 

substantial, and the nature of the toxicity data. Diflubenzuron, tebufenozide, and disparlure may 

be applied in aerial broadcast applications and multiple routes of exposure (oral, dermal, and 

inhalation) are plausible. No chronic exposures for disparlure are conducted, however, because 

no chronic toxicity data are available on this chemical. DDVP, on the other hand, is used only in 

milk carton traps and exposures will be minimal under normal conditions, although much higher 

exposures are possible if the traps are not assembled properly or if individuals tamper with the 

traps. 

Dose-Response Assessment – Dose-response assessments are typically based on an RfD, an 

estimate of a dose or exposure that is not likely to induce substantial adverse effects in humans. 

The RfD, in turn, is typically based on a NOAEL (no observed adverse effect level) divided by an 

uncertainty factor. Risk is then characterized as a hazard quotient (HQ), which is the estimated 

level of exposure divided by the RfD. If the HQ is below unity—i.e., the exposure is less than 

the RfD —there is no credible risk. If the HQ is above unity, risk is characterized based on doseresponse or dose-severity relationships. The quality of the dose-response assessment depends on 

the quality of the individual studies, the relevance of the studies to potential human exposures, 

and the strength of the dose-response relationship. 

As in the exposure assessments, the dose-response assessments for the biological agents differ 

substantially from each other as well as from those of the chemical agents. The dose-response 

assessment for the gypsy moth itself is based on only a single study; however, the study involves 

two human populations and demonstrates a clear dose-response relationship. Thus, confidence in 

the dose-response assessment is high. Two endpoints are considered for B.t.k., irritant effects 

and more serious toxic effects. While the irritant effects are well documented, there is no 

apparent dose-response relationship and confidence in the dose-response assessment is classified 

as medium. The dose-response assessment for more serious effects is based on a single study on 

mice involving intratracheal exposures. While a clear dose-response relationship is apparent, 

confidence in the dose-response assessment is low because intratracheal exposures have marginal 

x 
(if any) relevance to human exposures, the response was not independently replicated, and the 

observed response might be an artifact. For LdNPV, no endpoint of concern can be identified. 

Although the individual studies conducted on LdNPV are somewhat dated, the weight of 

evidence for LdNPV as well as other similar viruses clearly indicates that no systemic effects in 

humans are anticipated. Thus, confidence in the dose-response assessment for LdNPV is 

classified as high. 

Following standard practices in USDA risk assessments, risk assessment values available from 

U.S. EPA are adopted directly unless there is a compelling basis for doing otherwise. This 

approach is taken because the U.S. EPA will typically devote substantial resources and expertise 

to the development of risk assessment values and it is not feasible to duplicate this effort in risk 

assessments prepared for the USDA. In addition, the U.S. EPA has the legislative mandate to 

develop risk values for pesticides and it is sensible for the USDA to administratively defer to 

U.S. EPA in this area. When risk values are not available from U.S. EPA, the methods used by 

U.S. EPA are employed to derive surrogate values. Except for disparlure, chronic RfD values are 

available from U.S. EPA and these values are used directly. For 4-chloroaniline, the U.S. EPA 

also derived a cancer potency factor as well as a chronic RfD and these values are used directly in 

the risk assessment. For DDVP, the U.S. EPA derived an acute RfD, and this value is also 

adopted in the current risk assessment. A complication with DDVP, however, is that this agent is 

contained within a PVC strip, which substantially impacts the bioavailability of DDVP. In order 

to consider this detail quantitatively, a single and somewhat marginal study on the toxicity of 

DDVP in a PVC strip is used, and confidence in this dose-response assessment is, in turn, 

marginal. Unlike all of the other chemicals considered in this comparative risk assessment, very 

little toxicity data are available on disparlure. The U.S. EPA did not derive an RfD for this 

chemical, and the toxicity data available on disparlure are insufficient to derive a surrogate RfD. 

Thus, confidence in the dose-response assessment for disparlure is marginal. 

Risk Characterization – Of the agents considered in this risk assessment, the gypsy moth and 

DDVP are clearly agents of marked concern, although the nature of the concerns is different. If 

the gypsy moth is not controlled, population outbreaks will occur and humans will be exposed to 

large numbers of gypsy moth larvae. If this occurs, a substantial number of individuals will 

experience irritant effects that are sufficiently severe to cause these individuals to seek medical 

attention. No more serious effects are likely. For DDVP, the potential for risk is clear but the 

likelihood of observing risk seems to be remote. Under normal conditions and proper handling, 

levels of exposure to DDVP will be negligible and risk will be inconsequential. Workers who 

mishandle a DDVP-PVC strip and/or members of the general public who handle a DDVP-PVC 

strip may be exposed to levels of DDVP that are far above acceptable levels. While such 

exposures are clearly to be avoided, they are not likely to cause frank signs of toxicity. This 

conclusion is consistent with human experience in the use of DDVP resin strips. 

Diflubenzuron and tebufenozide are agents of marginal concern. Under most foreseeable 

conditions of exposure—i.e., exposure scenarios that might be characterized as 

typical—exposure levels will be far below levels of concern. At the upper ranges of plausible 

xi 
exposure – levels that might be characterized as extreme— the hazard quotients for 

diflubenzuron approach a level of concern (HQs between 0.1 and 0.5 for both diflubenzuron and 

its 4-chloroaniline metabolite). For tebufenozide, the highest hazard quotient is 1.5, which is 

characterized as undesirable; however, exposure is not likely to cause overt signs of toxicity. 

The somewhat higher hazard quotients for tebufenozide, compared with those of diflubenzuron, 

are due solely to higher application rates. 

Among the agents of minimal concern, B.t.k. is somewhat problematic. Based on the risk for 

serious adverse effects, there is clearly no cause for concern (the highest HQ is 0.04). As 

detailed in the dose-response assessment, this lack of concern is reenforced by a very aggressive 

and protective interpretation of the available toxicity data. Nonetheless, there is some residual 

concern with irritant effects. These effects are quite plausible in accidental cases of gross overexposure – e.g., splashing a formulation into the eye. These kinds of concern are minimal and 

are common to almost all chemical or biological agents. The more troubling concern involves 

studies of workers and non-workers who report irritant effects, primarily throat irritation. 

Whether or not these effects should be attributed to the B.t.k. exposure is unclear. 

The risk characterization for LdNPV and disparlure is unequivocal. Based on the available 

information, there is no basis for asserting that any serious adverse effects are plausible. Again, 

various accidental exposures, including splashing the agent into the eyes, could cause transient 

irritant effects. 

ECOLOGICAL RISK ASSESSMENT 

Hazard Identification – Unlike the human health risk assessment, in which the potential effects 

of the biological agents were similar, the ecological effects profile of each of the biological 

agents considered in this risk assessment are quite distinct. The gypsy moth will primarily affect 

sensitive trees, and these effects may be substantial. Because of the obvious importance of 

vegetation to the existence and habitat quality of most animals, a large number of secondary 

effects may be produced in many other groups of organisms. There is little indication, however, 

that the gypsy moth will have marked direct effects on groups of organisms other than sensitive 

plants. LdNPV, on the other hand, is not likely to have any effect on any species other than the 

gypsy moth. B.t.k. is toxic to nontarget Lepidoptera as well as the gypsy moth and some other 

lepidopteran species. There is very little indication that direct effects on other groups of 

organisms are plausible. Thus, the potential effects of all of the biological agents are considered 

relatively specific, with LdNPV showing the greatest degree of specificity (only the gypsy moth), 

followed by the gypsy moth itself (several types of plants) and B.t.k. (several types of 

Lepidoptera). 

The chemical agents also differ in specificity: disparlure is most specific, tebufenozide is 

relatively specific to Lepidoptera, diflubenzuron is less specific and may affect many arthropods, 

and DDVP is a nonspecific biocide toxic to many groups of animals, especially arthropods and 

vertebrates. As a pheromone, disparlure is almost as specific as LdNPV. It will attract the gypsy 

moth and two other closely related species, the nun moth (Lymantria monacha) and the pink 

xii 
gypsy moth (Lymantria fumida). As with the gypsy moth, both of these Lymantria species are 

forest pests, and adverse effects on these species are not a substantial concern for this risk 

assessment. In addition, the pink gypsy moth is native to Japan and is not found in the United 

States. A major qualification with the assessment of the specificity of disparlure is that, as in the 

human health risk assessment, the information on the toxicity of disparlure to nontarget species is 

very limited. At least in Daphnia magna, a commonly used test species in aquatic toxicity 

studies, the toxicity of disparlure is relatively high. Both diflubenzuron and tebufenozide are 

clearly toxic to mammals and at least some arthropods. In mammals, both chemicals will cause 

adverse effects in blood (methemoglobinemia), as detailed in the human health risk assessment. 

In both terrestrial and aquatic arthropods, both chemicals will interfere with growth and 

development. Because of differences in the mechanism of action of diflubenzuron and 

tebufenozide, tebufenozide appears to be somewhat more selective. Effects in birds have been 

clearly demonstrated for tebufenozide but not for diflubenzuron. While somewhat speculative, it 

seems plausible to assert that both diflubenzuron and tebufenozide are likely to affect the blood 

of birds in a way similar to that seen in mammals. In terms of the mechanism of action, DDVP is 

a general neurotoxin. In all animals that have nervous systems that involve acetylcholinesterase 

(AChE) and use acetylcholine (ACh) as a neurotransmitter (a substance necessary to make the 

nerves work properly), DDVP will be toxic, and sufficiently high exposures to DDVP will be 

lethal. The definition of sufficiently high, however, is critical and variable. Although DDVP is 

not selective mechanistically, differences in sensitivity among species are substantial. For 

instance, insects are much more sensitive than mammals or other higher organisms to DDVP 

exposure. This difference in sensitivity is what characterizes DDVP as an effective insecticide 

that can be used safely. 

Exposure Assessment – Diflubenzuron, tebufenozide, LdNPV, and disparlure may be applied in 

broadcast applications, which means that the potential for exposure is high and, in many cases, 

unavoidable. Disparlure, in addition to being used in broadcast applications, is used in traps as 

an attractant. Under those conditions of use, exposure to disparlure will be variable and 

primarily incidental. Exposures to the gypsy moth itself also vary, depending on the state of the 

gypsy moth population—i.e., from low level infestation to outbreak conditions. 

Some differences between the human health exposure assessment and the ecological exposure 

assessment, however, are notable. Table 4-2 does not give a measure of exposure for each agent. 

This is because the measure of exposure will vary both among agents and among the target 

groups for each agent. For example, exposures to the gypsy moth are measured as egg 

masses/acre in the human health risk assessment and this is the same measure of exposure used 

for terrestrial vegetation. As in the human health risk assessment, egg masses/acre are used as 

the measure of exposure because this is the primary determinant in the dose-response assessment 

for plants. For all other species, however, effects from the gypsy moth are likely to be secondary 

rather than primary. Thus, the exposure assessment for these indirectly affected species is based 

on defoliation – i.e., the result of the dose-response assessment for terrestrial vegetation is used 

as the exposure assessment for most other groups of organisms. 

xiii 
Other differences in the exposure assessments for nontarget species are mostly superficial. For 

each of the chemical agents, the mass of the chemical is typically used as the measure of 

exposure. Depending on the group, the measure of exposure may be expressed as dose (mg 

agent/kg bw for most terrestrial species), concentration (mg agent/L of water for aquatic species), 

or simply as application rate (lb agent/acre). This last measure is used primarily when field 

studies are the basis for the dose-response assessment. 

As in the human health risk assessment, different measures of exposure are used for each of the 

biological agents. For B.t.k., most of the exposures are characterized simply as an application 

rate in units of BIU/acre. However, colony forming units are used for some of the mammalian 

exposure scenarios. Also as in the human health risk assessment, no clear hazard potential is 

identified for LdNPV. The very few exposure scenarios that are quantified in the ecological risk 

assessment for LdNPV are based on the mass of the formulation, Gypchek. 

The level of detail used in the exposure assessments for the different chemicals reflects both 

differences in use patterns and the nature of the available toxicity data. Full sets of exposure 

assessments in several groups of animals are developed for diflubenzuron and tebufenozide. As 

in the human health risk assessment, the exposure assessment for diflubenzuron is elaborated by 

the consideration of 4-chloroaniline and the exposure assessment for tebufenozide is elaborated 

by the consideration of multiple applications. 

Disparlure, which also may be applied in aerial broadcast applications, has a much more 

restricted set of exposure scenarios on far fewer groups of organisms. This difference is due 

completely to the sparse toxicity data available on this compound. In other words, while a very 

elaborate set of exposure scenarios could be prepared, these scenarios would serve little purpose 

because they could not be combined with a dose-response assessment to characterize risk. The 

exposure assessment for DDVP is also restricted but this restriction is due to the very limited 

exposures that are plausible because DDVP is used only in milk carton traps and exposures for 

nontarget species will be minimal under normal conditions. 

Dose-Response Assessment – In general, confidence in any dose-response relationship is 

enhanced if a clear dose-response relationship can be demonstrated and both effect and no-effect 

exposures have been identified. In the case of LdNPV, however, there is simply no indication 

that LdNPV or the Gypchek formulation will cause toxicity in any nontarget species at any dose 

level. All of the risk values for LdNPV are based on no-effect concentrations or doses. While 

additional studies could be conducted at higher doses and while these studies would enhance 

confidence in the risk assessment, the NOAEL and NOEC values that have been identified are far 

above any plausible exposures. Thus, while based on limited data in terms of dose-effect 

characterization, the dose-response assessment for LdNPV is adequate for risk characterization. 

For most of the other agents, the dose-response assessments are reasonably good for the species 

of greatest concern. Dose-response assessments for DDVP are derived only for mammals, fish, 

and aquatic invertebrates. This limited approach is taken with DDVP because of the limited use 

xiv 
of DDVP in programs to control the gypsy moth. The DDVP is contained in a PVC strip that is 

placed in a milk carton trap that includes disparlure as an attractant for the gypsy moth. This type 

of use limits potential exposure to most nontarget species. A formal dose-response assessment is 

not conducted for terrestrial invertebrates. This is not due to any lack of data. The toxicity of 

DDVP to insects and many other invertebrates is very well characterized. DDVP is such a potent 

insecticide that no formal dose-response assessment is needed. Insects and many other species 

that enter the trap are likely to be killed by exposure to DDVP. 

Disparlure is the other agent for which a full set of dose-response assessments is not developed. 

As discussed in the hazard identification, this is due to the very limited data that are available on 

the toxicity of disparlure to nontarget species. 

Relatively full dose-response assessments on groups of greatest concern are given for the gypsy 

moth, B.t.k., diflubenzuron and its 4-chloroaniline metabolite, and tebufenozide. For the gypsy 

moth, the effect of primary concern is damage to vegetation. While data are available on both 

lethality in trees as well as defoliation, defoliation is used as the sublethal effect of primary 

concern. A dose-response assessment is also given for nontarget lepidopterans. While effect and 

no-effect levels can be identified, the significance of this effect is questionable. In terms of direct 

effects, terrestrial vegetation is the primary target of concern. 

The primary nontarget group of concern for B.t.k. involves Lepidoptera. A relatively rich set of 

studies is available in which the sensitivities of nontarget Lepidoptera as well as some other 

insects can be quantified reasonably well based on studies involving exposures that encompass 

the application rates used to control the gypsy moth. Sensitive nontarget lepidoptera include 

larvae of the endangered Karner blue butterfly as well as several other types of moths. 

Similar types of information are available on diflubenzuron and tebufenozide, and dose-response 

assessments can be made for the species of primary concern. For both chemicals, this includes 

nontarget Lepidoptera and aquatic invertebrates. Other terrestrial arthropods are also considered 

for diflubenzuron. In addition, because of the standard tests required by U.S. EPA for the 

registration of most pesticides, adequate toxicity data are available on mammals, birds, and fish. 

The toxicity data base for diflubenzuron is somewhat more extensive and sensitivities in 

nontarget organisms are somewhat better defined in both laboratory and field studies than is the 

case with tebufenozide. 

Risk Characterization – Ecological risk assessments involve, at least implicitly, considerations 

of thousands of different species and relationships among these species and their habitats. 

Invariably, however, data are available on only a small subset of these species and field studies 

provide only limited insight into the complex interrelationships and secondary effects among 

species. Thus, as in the human health risk assessments, ecological risk assessments cannot offer 

a guarantee of safety. They can and do offer a means to identify whether or not there is a basis 

for asserting that adverse effects are plausible and what the nature of these effects might be. 

xv 
Within these limitations, only LdNPV clearly qualifies as an agent of minimal concern. While 

there are limitations in the available studies on LdNPV, there is simply no basis for asserting that 

LdNPV will adversely affect any species except the gypsy moth. 

Agents of marked concern included the gypsy moth, B.t.k., and diflubenzuron. The types of 

concern with each of these agents, however, are quite different. For both the gypsy moth and 

B.t.k., the concerns are narrow. The gypsy moth will clearly damage some terrestrial vegetation. 

B.t.k. is likely to affect sensitive Lepidoptera. Concern with the use of diflubenzuron is broader 

and includes effects on both terrestrial and aquatic invertebrates. 

The designation of the gypsy moth as an agent of marked concern is obvious. The effects of 

gypsy moth larvae on forests are extremely well documented and well understood. In sensitive 

forest stands—i.e., stands in which oak, birch, and other favored species predominate —gypsy 

moth larvae can cause substantial defoliation and tree mortality. While some other lepidopteran 

species also may be directly affected by exposure to the gypsy moth, most of the other effects 

caused by the gypsy moth will be secondary. Reductions in populations of squirrels, mice, and 

other mammals which may be sensitive to changes in the availability of acorns are likely and 

have been well documented. Substantial secondary adverse effects on other groups of animals – 

i.e., birds, reptiles, and aquatic species – cannot be ruled out but have not been convincingly or 

consistently demonstrated. 

Diflubenzuron is also clearly an agent of marked concern. Exposures to diflubenzuron at 

application rates used in gypsy moth control programs will adversely affect both terrestrial and 

aquatic invertebrates that rely on chitin for their exoskeleton. This effect is demonstrated in 

controlled toxicity studies as well as multiple field studies. 

B.t.k. is considered an agent of marked concern because recent studies convincingly demonstrate 

that adverse effects in nontarget Lepidoptera will occur in the applications of B.t.k. used to 

control the gypsy moth. Concern is heightened because some of the Lepidoptera that may be 

adversely affected include at least one endangered species. 

Tebufenozide, DDVP, and disparlure are all classified as agents of marginal concern. For 

tebufenozide, the numeric expressions of risk may be less relevant than a more qualitative 

assessment. The highest risk is associated with the consumption of contaminated vegetation by a 

large mammal after two applications at the highest labeled application rate. It is not clear, 

however, that any frank signs of toxicity would be seen. Risks to nontarget Lepidoptera may be 

of greater concern, but the available data are insufficient to quantify potential risk. Risks to other 

invertebrates, both terrestrial and aquatic, appear to be insubstantial. DDVP is of marginal 

concern in that highly localized effects may be expected: nontarget insects entering a milk carton 

trap or some aquatic invertebrates affected by the accidental contamination of a small body of 

water with a pest strip. In both cases, the effects would be relatively minor, in terms of the 

number of organisms affected. Marginal concern for disparlure is associated with the relatively 

high toxicity of this agent to Daphnia. The very limited information on the toxicity of disparlure 

xvi 
argues for a persistent level of vigilance for this agent that may be applied to large areas in 

broadcast applications. 

xvii
 
1. INTRODUCTION
 

The USDA is preparing an update to the 1995 Environmental Impact Statement (EIS) for the 

Cooperative Gypsy Moth Management Program (USDA 1995). As part of this effort, updated 

risk assessments were developed on each of the chemical and biological control agents that are 

used in the USDA programs:—i.e., Bacillus thuringiensis var. kurstaki (B.t.k.) (SERA 2004a), 

the gypsy moth nucleopolyhedrosis virus (LdNPV) (SERA 2004b), diflubenzuron (SERA 

2004c), tebufenozide (SERA 2004f), DDVP (SERA 2004e), and disparlure as an active 

ingredient in materials used to attract the gypsy moth (SERA 2004d). In addition, a separate risk 

assessment was prepared on the gypsy moth (Lymantria dispar) itself. 

The current document not only compares the risks posed by the gypsy moth itself with the risks 

posed by the different control agents, but also compares the risks associated with the various 

control agents. The risk comparison is structured like the individual risk assessments and 

includes comparisons of uses (Section 2), potential human health effects (Section 3), and 

potential ecological effects (Section 4). As in the individual risk assessments, each of the 

comparative risk assessment sections (Sections 3 and 4) has four major subsections, including an 

identification of the hazards associated with the agents, an assessment of potential exposure, an 

assessment of the dose-response relationships, and a characterization of the risks associated with 

each agent. 

Each of the individual risk assessments cited above are complex, detailed, and often very large 

documents. The risk comparison does not attempt to summarize this information again in detail. 

Instead, it focuses on discussing the nature and quality of the data that support each step of the 

risk assessments and the uncertainties and limitations in the conclusions that are reached. Thus, 

with few exceptions, individual studies are not discussed or referenced in the current document. 

The exceptions primarily involve relatively recent studies that substantially impact the 

assessment of risk. Most of these studies involve B.t.k. (Herms et al. 1995; Hernandez et al. 

1999, 2000; Peacock et al. 1998; Petrie et al. 2003). 

1-1
 
2. PROGRAM DESCRIPTION
 

2.1. Overview 

The USDA control programs for the gypsy moth are intended to limit damage to forests that can 

be substantially impacted by gypsy moth outbreaks. Two biological agents that are pathogenic to 

the gypsy moth are used in broadcast applications: B.t.k. and LdNPV. In addition, three chemical 

agents are used in broadcast applications: diflubenzuron, tebufenozide, and disparlure. 

Diflubenzuron and tebufenozide are both insecticides and, as discussed in subsequent sections of 

the this document, have similar toxicologicalal properties. The major difference between the two 

is that application rates for tebufenozide are higher than those for diflubenzuron, which is a 

controlling factor in the comparative risk assessment for these two agents. Disparlure is a gypsy 

moth pheromone used in broadcast applications to disrupt mating and in population monitoring 

programs to attract the male gypsy moth to sampling traps. In the past, disparlure was used in a 

slow-release flake formulation. DDVP is not used in broadcast applications and is used only as a 

PVC formulated product in milk carton traps used in mass trapping operations. 

The USDA adopted various intervention strategies roughly categorized as suppression, 

eradication, and slow-the-spread. Suppression programs have relied predominantly on B.t.k. and 

diflubenzuron. Slow-the-Spread programs rely predominantly on the use of disparlure flakes and 

secondarily on B.t.k. applications. Eradication efforts rely predominantly on B.t.k. NPV is used 

in all three strategies but is used on very few acres relative to B.t.k., diflubenzuron, and disparlure 

flakes. 

2.2. Control Agents 

Gypsy moth is a pest species that can cause substantial damage to some forests. In the eastern 

United States, most hardwood forests are classified as susceptible to gypsy moth infestation and 

as many as 12.5 million acres have been defoliated in a single season. The gypsy moth is found 

throughout much of New England and south to Virginia and west to portions of Wisconsin. 

In past years, USDA employed chemical and biological agents in gypsy moth control programs. 

The biological control agents consist of B.t.k. and LdNPV. Both of these biological agents are 

pathogenic to the gypsy moth. The chemicals that may be used in the control of the gypsy moth 

include diflubenzuron, tebufenozide, and disparlure. Diflubenzuron and tebufenozide are used as 

direct insecticidal control agents, similar to the uses of B.t.k. and LdNPV. All of these agents are 

used in broadcast aerial or ground applications. 

DDVP and disparlure are used in mass trapping. Disparlure attracts the male gypsy moth to a 

large milk carton trap and the DDVP kills insects that enter the trap. While DDVP functions as 

an insecticide in the trap, it is not considered a control agent for the gypsy moth because mass 

trapping is used only in population surveys. Disparlure, in a flake formulation, is also used in 

broadcast aerial applications. While the disparlure does not cause any direct toxic effects to the 

gypsy moth, the mass application of disparlure will impair the ability of the male gypsy moth to 

2-1
 
find female gypsy moths and thus will limit the ability of gypsy moth populations to propagate. 

Thus, disparlure is used as a control agent. 

All of the agents used in gypsy moth control programs are applied in various types of 

formulations—i.e., the active ingredient combined with various other chemicals or materials. To 

the extent possible, these materials are discussed in each of the individual risk assessments. 

Specific information on inerts, however, is classified as CBI (confidential business Information) 

under Section 7(d) and Section (10) of FIFRA, and this information cannot be specifically 

disclosed in a risk assessment. In terms of a comparative risk assessment, however, the most 

important distinctions involve the formulations of B.t.k. and LdNPV in complex mixtures and the 

use of DDVP in polyvinyl chloride (PVC) strips. 

B.t.k. and LdNPV are both applied as very complex mixtures that are not fully or clearly defined. 

B.t.k. is cultured or grown in a medium containing water and nutrients, including sugars, 

starches, proteins, and amino acids. The nutrients, which are, themselves, chemically complex 

consist of variable biological materials, including animal foodstuffs, various flours, yeasts, and 

molasses. Similarly, LdNPV is prepared by isolating the virus from infected gypsy moth larvae. 

The active material consists of the virus, gypsy moth parts, and residual materials used to isolate 

and purify the virus. Complex mixtures can pose substantial difficulties in a risk assessment; 

however, the data on B.t.k. and LdNPV involve adequate studies on the toxicity of the complex 

mixtures. This is particularly true for B.t.k. in which much of the information on risk is based on 

applications of commercial formulations in the field. 

DDVP is used only in a PVC strip. Each strip contains 590 mg of DDVP and 89.25% inerts, 

which consist primarily of the PVC in the strip and plasticizers. The limited use of DDVP and 

its containment in the PVC strip have a major impact on the risk posed by DDVP, relative to the 

other compounds used in gypsy moth control programs. This impact is discussed at some length 

in the DDVP risk assessment and in subsequent sections of this document. 

2.3. Application Rates 

Application rates for the different control agents differ substantially both in magnitude and, for 

the biological agents, in the manner in which the application rate is expressed. 

For B.t.k., application rates are expressed in billions of international units (BIU), which is a 

measure of the activity or potency of the formulation rather than an expression of mass. The 

range of application rates used in USDA programs is 20-40 BIU/acre. For LdNPV, the 

recommended application rate is 0.43 oz Gypchek/acre for suppression and 1.08 oz Gypchek/acre 

for eradication. The application rate of 0.43 oz/acre corresponds to about 4x1011 PIB (polyhedral 

inclusion bodies)/acre and the application rate of 1.08 oz/acre corresponds to about 1x1012 

PIB/acre. 

Broadcast application rates are expressed in units of lb a.i./acre. For diflubenzuron, the range of 

labeled application rates is 0.0078-0.0624 lbs a.i./acre. For tebufenozide, higher labeled 

2-2
 
application rates are permitted: 0.03-0.12 lbs/acre. Multiple applications of tebufenozide are also 

permitted, and the maximum annual application rate is 0.24 lb a.i./acre. The application rates for 

tebufenozide may vary among USDA programs—i.e., suppression, eradication, and slow-thespread. For the tebufenozide risk assessment, a range of application rates—i.e., 0.015- 0.12 lb 

a.i./acre—are considered. All exposure assessments are conducted at the maximum application 

rate of 0.12 lb/acre, assuming that two applications are made with three days between 

applications. This worse-case scenario involves the use of two applications that reach the 

maximum annual application rate of 0.24 lb/acre and the shortest interval between applications. 

As noted in Section 3.4, the higher application rates for tebufenozide, compared with application 

rates for diflubenzuron, are the determining factor in the risk comparison. The application rate 

for disparlure is about 0.064 lb a.i./acre, near the maximum application rate allowed for 

diflubenzuron. Disparlure, however, is always applied in a slow-release formulation, either 

flakes or microspheres. DDVP is not applied in broadcast applications. Accordingly, the 

application rate is not a meaningful measure of exposure for this agent. 

2.4. Use Statistics 

In order to minimize the ecological effects and human health effects of gypsy moth infestations, 

the USDA adopted various intervention strategies roughly categorized as suppression, 

eradication, and slow-the-spread (Liebhold and McManus 1999). Suppression efforts are 

conducted by the USDA Forest Service in areas of well established gypsy moth infestations to 

combat or interdict periodic gypsy moth population outbreaks. Eradication efforts are conducted 

by USDA/APHIS to eliminate gypsy moth populations in areas where new populations of the 

gypsy moth are found. Slow-the-Spread, as the name implies, is a program to reduce the 

expansion of gypsy moth populations from areas of established populations to adjacent noninfested areas. 

The use of the various control agents in USDA programs is summarized in Table 2-1. This table 

gives the total number of acres treated with each of the control agents between 1995 and 2003. 

Suppression programs rely predominantly on B.t.k. and diflubenzuron. Slow-the-Spread 

programs rely predominantly on the use of disparlure flakes and secondarily on B.t.k. 

applications. Eradication efforts rely predominantly on B.t.k. NPV is used in all three strategies 

but is used on very few acres relative to B.t.k., diflubenzuron, and disparlure flakes. As 

discussed in the risk assessment on NPV, the production of Gypchek is very expensive and the 

application of this agent is currently limited to areas that are considered environmentally 

sensitive. As noted above, tebufenozide is not used in gypsy moth programs but may be used in 

the future. Given the similarities between tebufenozide and diflubenzuron, the use of 

tebufenozide is likely to be similar to that of diflubenzuron—i.e., primarily in suppression 

programs. 

2-3
 
3. HUMAN HEALTH RISK ASSESSMENT
 

3.1. HAZARD IDENTIFICATION 

3.1.1. Overview 

An overview of the comparative hazard identification for the gypsy moth and the agents used in 

USDA programs to control the gypsy moth is given in Table 3-1. The gypsy moth, B.t.k., and 

LdNPV are similar not only because they are biological agents but also because the primary 

effect associated with each agent is irritation. The gypsy moth causes more pronounced and 

severe irritation relative to either B.t.k. or LdNPV. Of the chemical agents used in gypsy moth 

control programs, diflubenzuron and tebufenozide are similar to each other in that both cause 

adverse effects on blood. The risk assessment of diflubenzuron is somewhat more involved than 

that of tebufenozide because diflubenzuron is degraded to 4-chloroaniline, a compound that is 

classified as a carcinogen. DDVP and disparlure, the other two chemicals used in gypsy moth 

control programs, have toxicological profiles that are very different from each other as well as 

from diflubenzuron or tebufenozide. DDVP is a well-characterized neurotoxin and the toxicity 

of DDVP in mammals has been studied extensively. Disparlure is an insect attractant that has 

not been extensively tested for toxicological effects in mammals. 

3.1.2. Biological Agents 

The biological agents—i.e., B.t.k., LdNPV, and the gypsy moth itself—present similar 

toxicological profiles. All three agents are irritants and cause similar irritant effects. The most 

likely effect from exposure to the gypsy moth is skin irritation. Gypsy moth larvae, as well as the 

larvae of many species of Lepidoptera, cause skin irritation in humans. The skin reactions seem 

to be associated with contact with small fine hairs that stick out from the body of the larva. 

Other effects associated with exposure to gypsy moth larvae include eye and respiratory 

irritation; however, these effects are not as well documented as the dermal effects. 

LdNPV also causes irritant effects. It is likely that the irritant effects are due at least in part to 

the presence of body parts of gypsy moth larvae in LdNPV preparations. Based on the available 

animal data, there is clear evidence that Gypchek, the commercial formulation of LdNPV, can 

cause eye irritation. There is little indication, however, that Gypchek is likely to cause dermal or 

respiratory irritation, which may have something to do with the processing of the gypsy moth 

parts during the preparation of Gypchek. 

The irritant effects of B.t.k. are probably due to the formulation of the bacteria rather than the 

bacteria itself. As noted in Section 2, commercial preparations of B.t.k. are very complex 

mixtures of the bacteria, fermentation byproducts, and adjuvants. B.t.k. formulations, however, 

are not strong irritants to either the eyes or the skin, except in the cases of accidental and gross 

contamination of the eyes. Instead, the most consistent effect appears to be irritation of the 

respiratory tract, particularly the throat. 

The irritant effects of the gypsy moth appear to be notably more severe than those of B.t.k. The 

wheals and rashes that result from exposure to the gypsy can cause severe itching which may 

3-1
 
persist from several days to two weeks. Moreover, these effects can be severe enough to cause 

the affected individual to seek medical treatment. The relatively consistent set of epidemiology 

studies following B.t.k. applications note a very different outcome. Despite many reports of 

irritant effects among exposed individuals, there is not a corresponding increase in the incidence 

of individuals seeking medical care. Thus, unlike the case in severe gypsy moth infestations, the 

severity of the irritant effects does not appear to be severe enough for individuals to seek medical 

care. 

There is very little indication that these biological agents will be associated with other more 

serious effects. LdNPV and Gypchek formulations of LdNPV were tested in relatively standard 

toxicity studies as well as in pathogenicity studies with no indication of serious effects even at 

very high doses. The gypsy moth has not been formally tested in human or animal studies; on the 

other hand, this species has infested North America for more than 100 years and no cases of 

frank adverse effects associated with gypsy moth exposure are to be found in the available 

literature. Hence, there appears to be no risk of serious adverse effects from exposure to LdNPV, 

Gypchek, or the gypsy moth itself. 

The potential for B.t.k. to produce serious adverse effects is somewhat more complicated than the 

assessment of LdNPV and the gypsy moth. As discussed in the B.t.k. risk assessment, severe 

adverse effects associated with exposure to B.t.k. are not reported in any of several epidemiology 

studies or standard animal toxicity studies on B.t.k. or formulations of B.t.k. A recent study by 

Hernandez et al. (2000), however, reports mortality in mice after intranasal instillations of B.t.k. 

Intranasal instillations of bacteria are analogous to inhalation exposures in that the bacteria are 

inhaled and transported to the lungs during the course of the study. This route of exposure is 

used to screen qualitatively for potential toxic effects, particularly for biological agents, and is 

not commonly used in a quantitative risk assessment because of uncertainties in extrapolating 

from intranasal doses to inhalation exposures that may occur in humans. In the B.t.k. risk 

assessment, some very conservative assumptions are made in the application of the Hernandez et 

al. (2000) study to provide an estimate of risk. As with LdNPV and the gypsy moth, this analysis 

(considered further in Sections 3.3 and 3.4) suggests that the risk of adverse effects is likely to be 

very low under foreseeable conditions of exposure. 

The Hernandez et al. (2000) study also reports that pre-treatment of mice with an influenza virus 

substantially increased mortality in mice exposed to various doses of B.t.k., again by intranasal 

instillation. This effect raises concern about the susceptibility of individuals who have influenza 

or other viral respiratory infections to severe adverse responses to B.t.k. exposure. The viral 

enhancement of bacterial infections is not uncommon and the enhancement of B.t.k. toxicity by a 

viral infection is, in some respects, not surprising. The relevance of this observation to public 

health cannot be assessed well at this time. No such effects are reported in the epidemiology 

studies conducted to date. It is, however, not clear that the epidemiology studies would detect 

such an effect or that such an effect is plausible under the anticipated exposure levels (typical or 

extreme) used in programs to control the gypsy moth. The viral enhancement of B.t.k. toxicity is 

likely to be an area of further study in the coming years. 

3-2
 
3.1.3. Chemical Agents 

In terms of potential human health effects, diflubenzuron and tebufenozide are similar to one 

another in that both cause adverse effects on blood. DDVP and disparlure, the other two 

chemicals used in gypsy moth control programs, have toxicological profiles that are very 

different from one another as well as from diflubenzuron or tebufenozide. The toxicity of 

DDVP, which is a well-characterized neurotoxin, has been studied extensively in mammals. 

Disparlure is an insect attractant that has not been extensively tested for toxicologicalal effects in 

mammals. 

3.1.3.1. Diflubenzuron and Tebufenozide – For both diflubenzuron and tebufenozide, the most 

sensitive effect in mammals involves damage to hemoglobin, a component in red blood cells that 

is responsible for transporting oxygen throughout the body. If this function is impaired, either 

because of damage to hemoglobin or lack of oxygen in the air, serious adverse effects (i.e., 

equivalent to suffocation) can occur. Both diflubenzuron and tebufenozide cause the formation 

of methemoglobin, a form of hemoglobin that is not able to transport oxygen. Both chemicals 

causes other effects on the blood; however, methemoglobinemia is the most sensitive 

effect—that is, the effect that occurs at the lowest dose. 

While effects on the blood are well documented, there is less of an indication that diflubenzuron 

or tebufenozide will cause other specific forms of toxicity. Neither diflubenzuron nor 

tebufenozide appears to be carcinogenic, mutagenic, neurotoxic or immunotoxic. Furthermore, 

these chemicals do not appear to cause birth defects or affect endocrine function in laboratory 

mammals. Diflubenzuron does not appear to cause reproductive effects. Tebufenozide, on the 

other hand, is associated with adverse reproductive effects in experimental mammals. These 

reproductive effects, however, occur at doses higher than those associated with 

methemoglobinemia. Neither diflubenzuron nor tebufenozide have a high order of acute oral 

toxicity. Diflubenzuron is relatively nontoxic by oral administration, with reported single-dose 

LD50 values ranging from greater than 4640 to greater than10,000 mg/kg. Similarly, single oral 

gavage doses of tebufenozide at 2000 mg/kg caused no observable signs of toxicity in mice or 

rats. 

Diflubenzuron is degraded to 4-chloroaniline in the environment. While most chemicals are 

metabolized in some way, the formation of 4-chloroaniline from diflubenzuron must be and is 

explicitly considered in the risk assessment because 4-chloroaniline is classified as a carcinogen. 

This is the only identified carcinogen associated with any of the agents used to control the gypsy 

moth. 

3.1.3.2. DDVP – DDVP is an organophosphorus insecticide that works by inhibiting 

cholinesterase. Inhibition of this enzyme in mammalian systems produces a variety of systemic 

effects, including salivation, urination, lacrimation, convulsions, increased bronchial secretions, 

respiratory depression, and even death. The nature and magnitude of the toxic effects produced 

by a given exposure to DDVP by any route are directly related to the dose and rate at which the 

exposure occurs. 

3-3
 
In the case of the USDA programs for the management of the gypsy moth, the use of milk carton 

traps containing Vaportape II (slow-release of DDVP from PVC strips) precludes rapid 

exposures to high doses of DDVP. The decrease in toxicity of DDVP in a PVC formulation has 

been studied directly. For the technical grade liquid DDVP, the acute oral LD50 in young pigs is 

about 160 mg/kg and signs of toxicity in these animals were consistent with the general signs of 

acetylcholinesterase (AChE) inhibition. In a similar bioassay using a PVC formulation, no 

deaths occurred at doses up to 1000 mg/kg. This key study on the comparative toxicity of DDVP 

and DDVP-PVC formulations is discussed further in the dose-response assessment (Section 3.3). 

DDVP is a very well studied compound and threshold doses for cholinesterase inhibition are well 

characterized. Short-term animal studies using technical grade DDVP indicate that oral 

exposures to doses below about 0.5 mg/kg-day (or inhalation exposures to 1-2 mg/m3) do not 

result in meaningful reductions in cholinesterase activity. Experiments in laboratory mammals 

exposed to DDVP during pregnancy (by oral or inhalation route) did not show any effect on 

fertility or health of the offspring, even at levels that produced maternal toxicity. The latest 

evaluation of data from assays for carcinogenicity and genetic toxicity classify DDVP as a 

“suggestive” carcinogen and determined that a quantitative assessment of cancer risk is not 

applicable. The literature contains some data suggesting that contact dermatitis (as well as crosssensitization to other pesticides) may occur; although, this appears to be an infrequent occurrence 

in the general population. 

3.1.3.3. Disparlure – In the registration of most pesticides, the U.S. EPA requires a relatively 

standard set of toxicity data covering multiple routes and durations of exposure as well as a 

number of specific endpoints of concern (e.g., carcinogenicity, reproductive toxicity, 

neurotoxicity, etc.). These requirements have been applied to diflubenzuron, tebufenozide, and 

DDVP but not to disparlure. Because of the apparently low toxicity of most pheromones to 

mammals and because of the low concentrations that are expected in the environment, U.S. EPA 

requires less rigorous testing of insect pheromones than is required of insecticides (U.S. EPA 

2004). 

The prudence of these assumptions may be argued but this issue is beyond the scope of the 

current risk assessment except to note that the application rate for disparlure is somewhat higher 

than the application rate for diflubenzuron—i.e., up to 0.0624 lbs a.i./acre for diflubenzuron and 

about 0.064 lb a.i./acre for disparlure (see Section 2). Nonetheless, as noted in Section 2, 

disparlure is always applied in a slow-release formulation (either flakes or microspheres) and the 

limited available monitoring data (Section 3.2), do support the assumption that exposures to 

disparlure are likely to be very low. 

In terms of the hazard identification, the result of the U.S. EPA position and the more general 

lack of concern with the toxicity of insect pheromones is that the toxicity of disparlure to 

mammals has not been studied extensively. Except for some standard acute toxicity studies in 

laboratory mammals, few data are available regarding the biological activity of disparlure in 

mammals. Results of acute exposure studies for oral, dermal, ocular, and inhalation exposure to 

3-4
 
disparlure show no indication of adverse effects. The acute toxicity of disparlure in mammals is 

very low. The LD50 of a single dose administered to rats by gavage exceeds 34,600 mg/kg. No 

studies investigating the effects of chronic exposure of mammals to disparlure or studies 

investigating the effects of disparlure on the nervous system, immune system, reproductive 

system, or endocrine system were identified. The carcinogenic potential of disparlure has not 

been assessed. In a single study on mutagenicity, there was no indication that disparlure is 

mutagenic. 

A case report of an accidental exposure indicates that disparlure may persist in humans for years. 

This case report concerns an individual involved in the early testing of disparlure who came into 

contact with the chemical. For more than 10 years after exposure to disparlure, the individual 

tended to attract male gypsy moths. This nuisance effect is the only well documented result of 

exposures to disparlure that might occur in USDA programs. 

3-5
 
3.2. EXPOSURE ASSESSMENT 

3.2.1. Overview 

A summary of the exposure assessments for each of the agents covered in the risk assessment is 

given in Table 3-2. The exposure assessments of the biological agents differ substantially from 

those of the chemical agents in terms of how the exposures are expressed. Different measures of 

exposure are used for each of the biological agents—i.e., the gypsy moth, B.t.k., and LdNPV. 

For the chemical agents, all exposure assessments are based on the amount or concentration of 

the chemical to which an individual or population might be exposed via ingestion, dermal 

contact, or inhalation. Differences among the chemical agents are dictated largely by differences 

in how the chemicals are used and, to a lesser extent, on the available toxicity data. 

A very different set of exposure assessments is conducted for each of the biological agents. Both 

B.t.k. and LdNPV may also be applied in broadcast applications and the routes of plausible 

exposure are the same as those for the chemicals applied in broadcast applications—i.e., oral, 

dermal, and inhalation. For B.t.k., however, the most directly relevant data used to characterize 

risk are based on actual applications of B.t.k. formulations where exposure is best characterized 

as an application rate. For the assessment of the potential for serious adverse effects, exposures 

are measured in colony forming units (cfu). LdNPV differs from all of the other agents in that no 

clear hazard potential can be identified. Thus, the most meaningful measure of exposure is in 

some respects moot. Those exposures that are quantified in the human health risk assessment for 

LdNPV are based on the mass of the formulation, Gypchek. Exposures to the gypsy moth itself 

are based on an indirect measure of exposure, egg masses/acre, because this is the expression of 

exposure that is used in the dose-response assessment. 

Differences in the exposure assessments among the chemicals used in USDA programs primarily 

reflect differences in how the chemicals are applied, what routes of exposure are most 

substantial, and the nature of the available toxicity data. Diflubenzuron, tebufenozide, and 

disparlure may be applied in aerial broadcast applications that lead to multiple routes of exposure 

(oral, dermal, and inhalation). No chronic exposures for disparlure are conducted, however, 

because no chronic toxicity data are available on this chemical. DDVP, on the other hand, is 

used only in milk carton traps and exposures will be minimal under normal conditions, although 

much higher exposures are possible if the traps are not assembled properly or if individuals 

tamper with the traps. 

3.2.2. Biological Agents 

The exposure assessments for the biological agents —i.e., the gypsy moth, B.t.k., and LdNPV 

differ substantially from each other, and these differences are largely dictated by the nature of the 

toxicity data available on each agent and the resulting dose-response assessments (Section 3.3.2). 

3.2.2.1. Gypsy Moth – For the gypsy moth, the most direct and relevant measure of human 

exposure is probably the number of larvae per unit area or tree because it is contact with the 

larvae that causes skin irritation, the adverse effect typically associated with the gypsy moth. 

Nonetheless, the available dose-response data are based on studies in which exposure is 

3-6
 
quantified as the number of eggs masses/acre. Accordingly, egg masses/acre is the exposure 

measure used in this risk assessment. As long as gypsy moth populations remain sparse, the 

larvae usually eat only a small proportion of the foliage of even their most favored host species, 

and contacts with people are rare. In such cases, egg masses generally do not exceed 50 egg 

masses/acre. During full-scale outbreaks, densities of about 5000 egg masses/acre are common 

and densities greater than 20,000 egg masses/acre are occasionally recorded. In such outbreaks, 

the numbers of gypsy moth larvae can reach up to 50,000 larvae per tree and exposure to the 

larvae will be essentially unavoidable for individuals near infested trees. 

3.2.2.2. B.t.k. – The exposure assessment for B.t.k. is unusual in two respects. First, the most 

directly relevant data used to characterize risk are based on actual applications of B.t.k. 

formulations where exposure is best characterized as an application rate. As summarized in 

Section 3.3.2, epidemiology studies are available that report responses in populations after 

applications of B.t.k. in the range of those used in USDA programs to control the gypsy 

moth—i.e., 20-40 BIU/acre. Thus, these studies are used directly in the risk characterization and 

explicit exposure assessments and dose-response assessments are not needed. 

Second, the apparent lack of a specific mechanism of toxicity for B.t.k. makes selecting the most 

appropriate measure of exposure somewhat arbitrary. The potency of B.t.k. is often expressed as 

BIU or FTU and exposures or application rates are expressed in units of BIU or FTU per acre. 

Although these units may be meaningful expressions of exposure for the gypsy moth, they are not 

necessarily or even likely to be a meaningful measures of human exposure. Exposure data are 

available, however, on colony forming units or cfu. When B.t.k. formulations are applied, either 

by aerial spray or ground spray, one or more viable spores contained in droplets or particulates is 

suspended in the air and deposited on sprayed surfaces. These droplets may be collected, either 

by air sampling or direct deposition, onto various types of filters. The filters are then cultured in 

a nutrient medium under conditions conducive to bacterial growth. As the bacteria grow, visible 

masses of bacteria, referred to as colonies, appear on the media. The significance of cfu as a 

measure of human exposure is limited because there is little indication that B.t.k. is a human 

pathogen. Consequently, the number of viable spores, albeit an important measure of exposure 

for the gypsy moth, does not appear to be toxicologicalally significant to humans. In this 

respect, cfu, like BIU, are of limited significance. 

Nonetheless, at least for short-term exposures, cfu can be used as a practical measure of relative 

exposure to a B.t.k. formulation. Based on cfu, ground workers may be exposed to much higher 

concentrations of B.t.k. than other groups—i.e., 200,000-15,800,000 cfu/m . 3 Much lower 

3 exposures, 400-11,000 cfu/m , have been measured in workers involved in aerial applications. 

During spray operations, members of the general public may be exposed to concentrations in the 

ranging from about 200-4000 cfu/m . 3 

3.2.2.3. LdNPV – Given the failure to identify any hazard associated with LdNPV or the 

Gypchek formulation, there is little need to conduct a detailed exposure assessment for Gypchek. 

Gypchek contains gypsy moth parts, and these constituents, like the gypsy moth larvae 

3-7
 
themselves, cause irritant effects in humans. The use of Gypchek, however, will not substantially 

increase the overall adverse effects of gypsy moth exposure in infested areas. On the contrary, 

the use of Gypchek will decrease the potential for human exposure to gypsy moth larvae by 

reducing larval populations. Based on simple physical processes associated with the application 

of any pesticide, it is possible to construct any number of exposure scenarios for Gypchek. The 

risk assessment for LdNPV focuses on one extreme exposure scenario involving the accidental 

spray of a home garden. While Gypchek is not intentionally applied to such vegetation, the 

inadvertent spray scenario is plausible. Based on this accidental exposure scenario, the estimated 

dose to an individual is 0.034 mg Gypchek/kg bw, with an upper range of 0.66 mg Gypchek/kg 

bw. 

3.2.3. Chemical Agents 

3.2.3.1. Diflubenzuron and Tebufenozide – Diflubenzuron and tebufenozide are applied in 

broadcast applications. The available data regarding the toxicity and environmental fate of these 

chemicals support a standard set of exposure scenarios involving worker exposure (both routine 

and accidental) and exposures of the general public to direct spray, dermal contact with 

contaminated vegetation, as well as the acute and longer-term consumption of contaminated food 

and water. For both of these chemicals, all exposure assessments are conducted at the maximum 

application rates. For diflubenzuron, all exposure assessments are conducted at the maximum 

single application rate for diflubenzuron of 0.0625 lb/acre, which is also the maximum amount 

that can be applied in a single season. The exposure assessments of tebufenozide are somewhat 

more elaborate because both single and multiple applications must be modeled—i.e., one or two 

applications at 0.12 lb/acre. While diflubenzuron is modeled at only the single maximum 

application rate, the exposure assessment for diflubenzuron is made elaborate by the quantitative 

consideration of the formation of 4-chloroaniline as an environmental metabolite. As noted in 

Section 3.1, the quantitative consideration of this metabolite is necessary because 4-chloroaniline 

is classified as a carcinogen and cancer risk is considered quantitatively in the risk 

characterization. 

The exposure patterns for both diflubenzuron and tebufenozide are similar. For workers, three 

types of application methods are modeled: directed ground spray, broadcast ground spray, and 

aerial spray. In these general applications, the maximum exposures to workers are similar: 0.009 

mg/kg/day for diflubenzuron and 0.02 mg/kg/day for tebufenozide. The differences in worker 

exposure levels merely reflect the differences in application rates for the two chemicals. 

Accidental dermal exposures for workers can be much higher: 0.4 mg/kg/day for diflubenzuron 

and 4 mg/kg/day for tebufenozide. These differences in exposure levels reflect the differences in 

the concentrations of the two chemicals used in field solutions as well as the differences in the 

estimated dermal absorption rates. 

For members of the general public, the exposure profiles for diflubenzuron and tebufenozide are 

also similar. The maximum acute exposure levels for both chemicals are associated with 

contaminated water in an accidental spill scenario: doses of 1.5 mg/kg bw for diflubenzuron and 

1.2 mg/kg bw for tebufenozide. Longer-term exposure to both agents, which involves the 

3-8
 
 

consumption of contaminated fruit rather than water, will result in much lower levels of 

exposure: 0.002 mg/kg/day for diflubenzuron and 0.03 mg/kg/day for tebufenozide. Like 

workers, members of the general public can be at risk of dermal exposure to diflubenzuron or 

tebufenozide, and dermal exposure concentrations can be estimated quantitatively. Estimates of 

dermal exposure, however, are lower than estimates of oral exposure: a maximum of 0.05 mg/kg 

bw for diflubenzuron and about 0.4 mg/kg bw for tebufenozide. 

Exposure assessments for 4-chloroaniline as an environmental metabolite of diflubenzuron are 

made only for members of the general public. Workers are not considered at risk because 

significant amounts of 4-chloroaniline are not likely to form during the application of 

diflubenzuron. For the general public, estimates of exposure to 4-chloroaniline from 

contaminated vegetation are likely to be about a factor of 50 below the corresponding estimates 

of exposure to diflubenzuron. The lower estimate of exposure to 4-chloroaniline is due to its 

expected rapid dissipation from diflubenzuron deposited on vegetation. In water, however, 

estimated concentrations of 4-chloroaniline are likely to be equal to or greater than anticipated 

water concentrations of diflubenzuron under certain circumstances. Finally, peak exposures to 

4-chloroaniline differ from peak exposures to diflubenzuron in the environment, usually 

occurring at different times (later after the application of diflubenzuron) and under different 

conditions of precipitation. These differences are due to the relatively slow rate in the formation 

of 4-chloroaniline from diflubenzuron in soil. 

3.2.3.2. Disparlure – Disparlure is like diflubenzuron and tebufenozide in that all three can be 

applied by aerial broadcast and multiple routes of acute and longer-term exposure are possible. 

The exposure assessment for disparlure, however, is much less elaborate than those for 

diflubenzuron and tebufenozide because of the very limited toxicity data base on disparlure. As 

discussed in Section 3.3 (Dose-Response Assessment), the U.S. EPA did not derive RfD values 

for acute or chronic exposure and the available toxicity data do not support the derivation of 

surrogate values. Thus, in the absence of toxicity data, an elaborate exposure assessment would 

not be useful in evaluating risk. 

For disparlure, dermal exposure is most likely to be the predominant route for occupational 

exposure and is a possible route of exposure for the general public. A case report involving the 

accidental exposure of a worker to disparlure indicates that the only notable effect in the worker 

was the persistent attraction of gypsy moths. Since the available acute systemic toxicity of 

disparlure in mammals appears to be very low, the absence of dermal absorption data does not 

add significant uncertainty to this risk assessment. While dermal exposure of workers is 

expected to be non-toxic, dermal exposure is likely to cause the persistent attraction of gypsy 

moths. 

Both workers and the public may be exposed to disparlure by inhalation, and the magnitude of 

the exposure can be estimated from available monitoring studies. At application rates more than 

15 times the normal application rate (i.e., about 200 g a.i./acre compared with 29.1 g/acre), peak 

air concentrations ranged from 0.022 to 0.030 μg/m . 3 Adjusted to the normal application rate, 

3-9
 
3 these values correspond to about 0.003-0.004 μg/m , which is far below the air concentration of

3 5.0 mg/L —equivalent to 5000 μg/L or 5,000,000 μg/m —that did not cause mortality or signs of

toxicity in experimental animals. 

3.2.3.3. DDVP – Unlike the other chemicals used in gypsy moth control programs, DDVP is not 

applied in broadcast applications. DDVP is used only in a PVC strip that is placed in milk carton 

traps. Consequently, exposures of both workers and members of the general public should be 

negligible under normal conditions—i.e., the workers use proper procedures during assembly of 

the traps and members of the general public do not tamper with the traps. The risk assessment 

for DDVP does develop exposure scenarios for both workers and members of the general public 

to encompass improper handling of the DDVP strips by workers or tampering with the traps by 

members of the general public. These exposures, however, should be considered atypical, and 

some are extreme. 

During assembly, the central estimate of dermal exposures in workers not wearing gloves leads 

to an absorbed dose of about 0.0008 mg/kg with a range of about 0.0003 mg/kg to 0.004 mg/kg. 

Inhalation exposures in workers may be highly variable depending on the ventilation rates in an 

enclosed space and the number of traps that are handled. Based on the handling and transport of 

3 75 traps, inhalation exposures could be as high as 0.6 mg/m in an enclosed and unventilated

3 room and as high as 1.8 mg/m in the passenger compartment of a vehicle. These exposure 

assessments are based on several site-specific and situation-specific assumptions intended to 

reflect plausible upper bounds of exposure. 

Exposure assessments are also developed for children who might come in contact with an 

accidentally discarded or misplaced DDVP strip. Estimated dermal doses are much higher than 

those for workers: a central estimate of about 0.02 mg/kg with a range of 0.003-0.1 mg/kg. Oral 

exposures from a small child sucking on the pest strip are about a factor of 10 higher than dermal 

exposures: a central value of about 0.2 mg/kg with a range of 0.04-0.6 mg/kg. 

Under normal circumstances, the use of DDVP in PVC strips is not likely to result in 

contamination of water or other materials that might be consumed by members of the general 

public. Nonetheless, an exposure assessment is developed for the accidental contamination of a 

small pond by a pest strip. In this scenario, dose estimates range from about 0.000003 to 0.00007 

mg/kg with a central estimate of about 0.00001 mg/kg. 

3-10
 
3.3. DOSE-RESPONSE ASSESSMENT 

3.3.1. Overview 

A summary of the dose-response assessments for each of the agents covered in the risk 

assessment is given in Table 3-3. Dose-response assessments are typically based on an RfD, an 

estimate of a dose or exposure that is not likely to induce substantial adverse effects in humans. 

The RfD, in turn, is typically based on a NOAEL (no-observed-adverse-effect level) divided by 

an uncertainty factor. Risk is then characterized as a hazard quotient (HQ) which is the estimated 

level of exposure divided by the RfD. If the HQ is below unity—i.e., the exposure is less than 

the RfD —there is no credible risk. If the HQ is above unity, risk is characterized based on doseresponse or dose-severity relationships. The quality of the dose-response assessment depends on 

the quality of the individual studies, the relevance of the studies to potential human exposures, 

and the strength of the dose-response relationship. 

As in the exposure assessments (see Section 3.2), the dose-response assessments for the 

biological agents differ substantially from one another as well as from those of the chemical 

agents. The dose-response assessment for the gypsy moth itself is based on only one study; 

however, the study involves two human populations, and a demonstrates a clear dose-response 

relationship. Thus, confidence in the dose-response assessment is high. Two endpoints are 

considered for B.t.k., irritant effects and more serious toxic effects. While the irritant effects are 

well documented, there is no apparent dose-response relationship and confidence in the doseresponse is classified as medium. The dose-response assessment for more serious effects is 

based on a single study in mice that involves intranasal exposures. Although the study 

demonstrates a clear dose-response relationship, confidence in the dose-response assessment is 

low because intranasal exposures have marginal (if any) relevance to human exposure, the 

response was not independently replicated, and the observed response may be an artifact. For 

LdNPV, no endpoint of concern can be identified. Although the individual studies conducted on 

LdNPV are all somewhat dated, the weight of evidence for LdNPV and similar viruses clearly 

indicates the unlikelihood of systemic effects in humans after exposure to LdNPV. Thus, 

confidence in the dose-response assessment for LdNPV is classified as high. 

Following standard practices in USDA risk assessments, risk assessment values available from 

U.S. EPA are adopted directly unless there is a compelling basis for doing otherwise. This 

approach is taken because the U.S. EPA will typically devote substantial resources and expertise 

to the development of risk assessment values and it is not feasible to duplicate this effort in risk 

assessments prepared for the USDA. In addition, the U.S. EPA has the legislative mandate to 

develop risk values for pesticides and it is sensible for the USDA to administratively defer to 

U.S. EPA in this area. When risk values are not available from U.S. EPA, the methods used by 

U.S. EPA are employed to derive surrogate values. Except for disparlure, chronic RfD values are 

available from U.S. EPA and these values are used directly. For 4-chloroaniline, the U.S. EPA 

derived a cancer potency factor as well as a chronic RfD, and these values are used directly in the 

risk assessment. For DDVP, the U.S. EPA derived an acute RfD, and this value is also adopted 

in the current risk assessment. A complication with DDVP, however, is that this agent is 

contained within a PVC strip, which substantially impacts its bioavailability. In order to consider 

3-11
 
this matter quantitatively, a single and somewhat marginal study regarding the toxicity of DDVP 

in a PVC strip is used, and confidence in the dose-response assessment is, in turn, marginal. 

Unlike all of the other chemicals considered in this comparative risk assessment, very little 

toxicity data are available on disparlure. The U.S. EPA did not derive an RfD for this chemical, 

and the available toxicity data are insufficient to derive a surrogate RfD. Thus, confidence in the 

dose-response assessment for disparlure is marginal. 

3.3.2. Biological Agents 

3.3.2.1. Gypsy Moth – The dose-response assessment for human health effects is based on 

reports of skin irritation in two populations: one with low exposure (an average of 32 egg 

masses/acre) and the other with high exposure (an average of 3809 egg masses/acre). The lowexposure group exhibited no increase in skin irritation. Accordingly, 32 egg masses/acre is taken 

as a NOAEL (no-observed-adverse-effect level) for humans and is used as a surrogate RfD for 

exposure to the gypsy moth in a manner analogous to the use of RfD values for control agents. 

The high exposure group had a significant increase in skin irritation, and, based on a doseresponse model developed by U.S. EPA, egg mass densities up to128 egg masses/acre are not 

likely to cause a detectable increase in skin irritation or rashes. 

While the dose-response relationship is based on only two exposure levels, the strength of the 

dose-response relationship is strong. Typically, an association is judged to be statistically 

significant if the p-value (the probability that the association occurred by chance) is 0.05 or less. 

For the study on which these dose-response relationships are based, p-values are on the order of 

0.0004 or less for most groups. The only exception involves individuals over the age of 59 years. 

In this group, it is unclear if the lack of a significant response is related to a lesser sensitivity to 

the gypsy moth or less exposure—i.e., less time spent outdoors. 

In addition to these quantitative estimates of response, the severity of the response is important, 

particularly in a comparison of effects caused by exposure to the gypsy moth and effects caused 

by exposure to the agents used to control the gypsy moth. Dermal responses to the gypsy moth 

are sufficiently severe to have generated numerous case reports. While precise statistics are not 

available, it does appear that the severity of the skin irritation is sufficient to cause appreciable 

numbers of affected individuals to seek medical care. While exposure to the gypsy moth is 

associated with irritation to the eyes and respiratory tract, quantitative dose-response 

relationships for these endpoints cannot be developed. 

3.3.2.2. B.t.k. – Two types of dose-response assessments are presented for B.t.k., one for 

irritant effects and the other systemic toxicity. There is relatively high confidence that 

formulations of B.t.k. will cause various types of irritant effects in humans and experimental 

animals; however, confidence in the quantitative assessment of these effects is limited by a very 

weak dose-dependency in the incidence of the response. The quantitative assessment for 

systemic toxic effects is extremely tenuous because it is based on a very conservative 

interpretation of a single study using a route of exposure (intratracheal instillation) that typically 

is not used in quantitative risk assessments. 

3-12
 
The estimate for irritant effects is actually a set of observations rather than a formal doseresponse assessment. Several epidemiology studies were conducted after B.t.k. applications at 

rates within the range of those used in USDA programs to control the gypsy moth—i.e., 20-40 

BIU/acre. Two key epidemiology studies, one involving workers (Cook 1994) and the other 

involving members of the general public (Petrie et al. 2003), suggest that irritant effects, 

particularly throat irritation, may be reported in groups of humans during or after applications of 

B.t.k. In the worker study, the data demonstrate a statistically significant increase in the 

incidence of irritant effects in workers. The significantly increased effects include generalized 

dermal irritation (dry or itchy skin and chapped lips), irritation to the throat, and respiratory 

irritation (cough or tightness). Furthermore, the overall incidence of all symptoms combined was 

increased significantly among the workers, compared with the controls . In the study involving 

the general public, several types of irritant effects are reported; however, the only effect that is 

clearly statistically significant involves throat irritation (p=0.002). 

Confidence in accepting whether these reports are biologically significant, however, is reduced 

by the apparent lack of a strong dose-response relationship. The workers were exposed to up to 

3 about 16 million cfu/m and the reported incidence of throat irritation is about 24%. In the study 

involving members of the general public, no measures of exposure are given. Based on 

monitoring data from similar applications, however, it is likely that members of the general 

public may have been exposed to air concentrations ranging from approximately 100 to 4000 

3 cfu/m during or shortly after aerial applications of B.t.k. This range is a factor of 3950 to 

158,000 less than exposures in the worker study. The apparent incidence of throat irritation in 

the study on members of the general public, however, is about 19%. Thus, while these much 

lower exposures lead to a somewhat lesser response, the dose-response relationship appears 

weak. Nonetheless, these studies are taken together to characterize risk semi-quantitatively, as 

discussed further in Section 3.4. 

There is essentially no information indicating that oral, dermal, or inhalation exposure to B.t.k. or 

B.t.k. formulations will cause serious adverse health effects. Extremely severe inhalation 

exposures that coat the test species with commercial formulations of B.t.k. are associated with 

decreased activity, discolored lungs, and other effects but not mortality. Although the animal 

data are consistent with data regarding human exposure B.t.k., the animal studies are all based on 

single concentrations and cannot be used in a meaningful dose-response assessment. 

Few studies (David 1990; Hernandez et al. 1999,2000) report mortality after exposure to B.t.k., 

and these studies, while related to inhalation toxicity, involve atypical routes of exposure. One 

such study (David 1990) was conducted on a B.t.k. Dipel formulation after intratracheal 

instillations. Intratracheal instillations of bacteria are analogous to inhalation exposures in that 

the bacteria is essentially inserted into the lungs. Toxic responses including death were observed 

in treated animals, and the time-to-clearance (estimated from linear regression) was prolonged. 

Hernandez et al. (1999, 2000) assayed the toxicity of B.t.k. after intranasal instillations in mice. 

This method of dosing is also analogous to inhalation exposures in that the material is deposited 

in nasal passages and the B.t.k. is gradually transported to the lungs by inhalation. Doses of 10 ,2 

3-13
 
4 6 8 10 , and 10 cfu/mouse caused only local inflamation. A dose of 10 cfu/mouse resulted in 80% 

lethality. 

In terms of the human health risk assessment, the data from Hernandez et al. (1999, 2000) are not 

directly useful. Furthermore, the route of exposure (intranasal instillation) makes any use of 

these data somewhat tenuous. Concern with the use of this atypical route of exposure in a doseresponse assessment is exacerbated because the Hernandez et al. (2000) study does not specify 

whether or not the instillations were adjusted to a constant volume. If the installations were not 

adjusted to a constant volume, it is possible that the observed dose-response relationship could be 

due to differences in volumetric bronchial obstruction or a combination of bronchial obstruction 

and B.t.k. 

Notwithstanding these reservations, the Hernandez et al. (1999, 2000) studies provide the best 

dose-response data available in experimental mammals. Based on a consideration of the 

Hernandez et al. (2000) study and the estimates of equivalent human exposures, it seems 

3 plausible that cumulative exposures up to 1.4x1010 cfu/m x hour will not cause adverse effects in

humans. This estimate is supported by the worker study (Cook 1994) from which an apparent 

3 NOAEL of 3x108

 cfu/m x hours for adverse health effects in humans can be calculated, and this

value is used quantitatively to characterize the potential for serious adverse effects in humans. 

3.3.2.3. LdNPV – The dose-response assessment for LdNPV and its formulation as Gypchek is 

extremely simple, compared with the other biological and chemical agents, except disparlure, 

used to control the gypsy moth. Due to the lack of systemic toxic effects associated with any 

plausible route of exposure (i.e., oral, dermal, or inhalation), the U.S. EPA did not derive an 

acute or chronic RfD for Gypchek. Although this approach is reasonable, the risk assessment for 

LdNPV, which is used in the EIS, derives a surrogate acute RfD of 26 mg/kg bw. The surrogate 

RfD, which is based on an experimental acute NOAEL of 2600 mg/kg bw in rats and an 

uncertainty factor of 100, provides a quantitative basis for comparison between the extremely 

low risks associated with the application of Gypchek and the risks posed by the other gypsy moth 

control agents. Confidence in this value is limited because no adverse effect levels were 

identified—i.e., the true NOAEL for Gypchek may be higher than 2600 mg/kg. This uncertainty 

in the LdNPV risk assessment is relatively minor, given that even extreme exposures are far 

below any level of concern (Section 3.4). 

Technical grade Gypchek is an eye irritant. While not quantitatively considered in the risk 

assessment, the distinction between the irritant properties of technical grade Gypchek and the 

lack of eye irritation associated with Gypchek formulations applied in the field is emphasized in 

order to highlight areas in which prudent handling practices are likely to be most important. 

3.3.3. Chemical Agents 

3.3.3.1. Diflubenzuron and Tebufenozide – As discussed in the hazard identification and the 

exposure assessment, diflubenzuron and tebufenozide are similar to one another in terms of their 

toxicological profiles. Both chemicals were tested in a similar and relatively standard set of 

3-14
 
toxicity studies required by the U.S. EPA for the registration of pesticides. Their most sensitive 

endpoint, hematological effects (including methemoglobin formation and several other endpoints 

characteristic of hemolytic anemia) was observed in all mammalian species tested. 

Quantitatively, the similarities between diflubenzuron and tebufenozide are further expanded and 

even more striking in the dose-response assessment. The U.S. EPA derived RfDs for both 

compounds and the values are virtually identical: 0.02 mg/kg/day for diflubenzuron and 0.018 

mg/kg/day for tebufenozide. Even this minor difference is an artifact of rounding. The U.S. 

EPA agency-wide workgroup, which derived the RfD for diflubenzuron, typically rounds all 

RfDs to one significant place. The U.S. EPA Office of Pesticides, which derived the RfD for 

tebufenozide, often reports RfDs to two significant places. If the agency-wide criteria had been 

applied to tebufenozide, the two RfDs would be identical—i.e., 0.02 mg/kg/day. Since the 

molecular weights of diflubenzuron (310 g/mole) and tebufenozide (352 g/mole) are so similar, 

the RfDs would be identical even when expressed in moles—i.e., 7x10-5 mMoles/kg/day for 

diflubenzuron and 5x10-5 mMoles/kg/day for tebufenozide. 

The RfDs for both chemicals are based on dietary studies in rats, and the respective NOAELs are 

quite similar: 2 mg/kg/day for diflubenzuron and 1.5-2.4 mg/kg/day for tebufenozide. Again, 

these minor differences are an artifact of the way in which the dietary concentrations (i.e., mg 

agent/kg diet) used in the studies were converted to dose estimates expressed as mg/kg bw/day 

based on food consumption. Both RfDs are also based on an uncertainty factor of 100, a factor 

of 10 for interspecies differences—i.e., extrapolation of animal data to humans—and a factor of 

10 for intraspecies variability—i.e., individuals who might be most sensitive to the chemical. 

For both chemicals, the U.S. EPA determined that an additional uncertainty factor of 10 for the 

protection of infants and children, a factor that must be considered under the Food Quality 

Protection Act (FQPA), is not required. Finally, confidence in both RfDs is high, which is stated 

explicitly in the Agency wide RfD for diflubenzuron and is implicit in the discussion of the 

chronic RfD for tebufenozide derived by the U.S. EPA Office of Pesticides—i.e., no data gaps 

are identified. 

The acute dose-response assessments on diflubenzuron and tebufenozide prepared by U.S. EPA 

are similar in that the U.S. EPA elected not to derive an acute RfD for either compound. This 

approach is taken because the agency concluded that no endpoint for acute dietary exposure 

could be identified for either chemical. U.S. EPA identifies an acute NOAEL of 10,000 mg/kg 

bw for diflubenzuron and an acute oral NOAEL of 2000 mg/kg bw for tebufenozide. For the 

USDA risk assessments on gypsy moth control agents, surrogate acute RfDs are derived for both 

chemicals according to the methods typically employed by the U.S. EPA, because many areas of 

greatest concern involve potential acute effects after accidental or incidental exposures. 

For diflubenzuron, a surrogate acute RfD of 100 mg/kg could be derived using the NOAEL of 

10,000 mg/kg identified by U.S. EPA. A more conservative approach is taken, however, using 

the NOAEL of 1118 mg/kg from an acute study (single dose) in which Dimilin 4L, a formulation 

containing petroleum oil, was used. The resulting surrogate acute RfD is 11 mg/kg. A similar 

3-15
 
approach is taken for tebufenozide. Rather than using an acute NOAEL of 2000 mg/kg, a 

NOAEL of 1000 mg/kg/day in pregnant rats and rabbits, identified by U.S. EPA, is used to 

derive a surrogate acute RfD of 10 mg/kg/day. Like the chronic RfDs, the acute RfDs are nearly 

identical. 

The dose-response assessment for diflubenzuron is somewhat more complicated than that for 

tebufenozide because of the need to consider 4-chloroaniline quantitatively. As noted in the 

hazard identification (see Section 3.1), 4-chloroaniline is an environmental metabolite of 

diflubenzuron and 4-chloroaniline has been classified as a potential human carcinogen. The U.S. 

EPA derived a chronic RfD for 4-chloroaniline of 0.004 mg/kg/day, and this value is used to 

characterize risks from 4-chloroaniline for longer-term exposures. This RfD is based on a 

chronic oral LOAEL of 12.5 mg/kg/day using an uncertainty factor of 3000, three factors of 10 

for interspecies extrapolation, sensitive subgroups, and the use of a LOAEL with an additional 

factor of 3 due to the lack of reproductive toxicity data. As with diflubenzuron, the U.S. EPA 

has not derived an acute RfD for 4-chloroaniline. For 4-chloroaniline, a conservative approach is 

taken in which a surrogate acute RfD of 0.03 mg/kg is based on a subchronic (90-day) NOAEL 

of 8 mg/kg/day. Consistent with the approach taken by U.S. EPA for the chronic RfD, an 

uncertainty factor of 300 is used. For cancer risk, the U.S. EPA proposes a human cancer 

potency factor for 4-chloroaniline of 0.0638 (mg/kg/day) -1. This potency factor is used to 

calculate a dose of 1.6x10-5 mg/kg/day that could be associated with a plausible upper limit of 

cancer risk of 1 in 1 million. 

3.3.3.2. Disparlure – As noted in the hazard identification (see Section 3.1.3.3), the U.S. EPA 

does not require extensive testing of insect pheromones, including disparlure. This approach is 

taken because insect pheromones are generally regarded as nontoxic to mammals and because 

these pheromones are commonly employed in very low environmental concentrations. While the 

merits of this approach may be argued, the result is that there is little information regarding the 

toxicity of disparlure, and no RfD values, acute or chronic, have been or can be derived. 

The only information that can be used to assess the consequences of exposure to disparlure are 

LD or LC values: an oral LD value greater than 34,600 mg/kg; a dermal LD value greater 50 50 50 50 

than 2025 mg/kg, and an inhalation LC50 value greater than 5 mg/L @ 1 hour. Notably, each of the 

values is expressed as “greater than”. In other words, less than half of the organisms died at the 

specified exposure. In the case of disparlure, these values are actually NOEC values for 

mortality in that none of the animals died during any of the exposures. 

3.3.3.3. DDVP – Like diflubenzuron and tebufenozide, and perhaps to an even greater extent, 

DDVP has an extensive toxicology data base that has been evaluated by numerous government 

organizations, including U.S. EPA, the Agency for Toxic Substances and Disease Registry 

(ATSDR), National Institute for Occupational Safety and Health (NIOSH), Occupational Safety 

and Health Administration, and the World Health Organization. As noted above, these sources 

are used when possible for selecting levels of acceptable exposure. Because all of the scenarios 

3-16
 
considered in this risk assessment involve only acute exposures, only acute exposure criteria are 

considered. 

The acute RfD established by the U.S. EPA for oral and dermal exposure to DDVP, 0.0017 

mg/kg, is used for the risk characterization. The RfD is based an acute oral NOAEL of 0.5 

mg/kg from a rat study, and the application of an uncertainty factor of 300. Acute exposure 

criteria proposed by other groups are comparable to but somewhat higher than the acute RfD. 

Because some of the accidental acute exposures may substantially exceed the acute RfD, some 

attempt is made to characterize the consequences of high oral exposures. A human NOAEL of 1 

mg/kg for AChE inhibition has been identified. While this NOAEL is not used to modify the 

acute RfD, it can be used to assess plausible consequences of exceeding the RfD. The human 

data on DDVP, although extensive, are not sufficient to identify a minimal lethal dose. For the 

current risk assessment, the lowest reported lethal dose (16 mg/kg) is used to assess the 

plausibility of observing serious adverse effects in cases of accidental overexposure to DDVP. 

A number of inhalation criteria are available for DDVP. Since potentially significant inhalation 

3 exposures are likely only in workers, the occupational exposure criterion of 0.1 mg/m proposed

by American Conference of Governmental Industrial Hygienists is used. This value is a factor of 

10 below the occupational criteria proposed by NIOSH and OSHA. 

A major factor and a major complication in the dose-response assessment of DDVP involves the 

formulation of DDVP in a PVC strip. Some of the accidental exposures considered in this risk 

assessment involve a small child gaining access to a DDVP-PVC strip and being subject to both 

oral and dermal exposure. While there is little doubt that the PVC strip will slow the rate of 

exposure and reduce the risk, this is extremely difficult to quantify. Despite the availability of 

numerous studies regarding the toxicity of DDVP itself, the number of studies regarding the 

toxicity of DDVP-PVC strips is relatively small. By far the most relevant study is that conducted 

by Stanton et al. (1979), which clearly indicates that DDVP in a PVC formulation will be much 

less toxic than unformulated DDVP. The extent of the difference in toxicity can only be semiquantitatively characterized. For unformulated DDVP, the LD50 value was 157 (113–227) mg/kg 

with no mortality observed at 56 mg/kg. For the DDVP-PVC formulation, no deaths occurred at 

doses of up to 1000 mg/kg, although signs of toxicity consistent with AChE inhibition were 

observed at doses of 320 and 1000 mg/kg. Neither tremors nor salivation were observed at doses 

of 240 or 180 mg/kg. Stanton et al. (1979) do not provide comparative data on the extent of 

AChE inhibition in unformulated DDVP and the DDVP-PVC formulation. 

3-17
 
3.4. RISK CHARACTERIZATION 

3.4.1. Overview 

Risk characterization is the process of comparing the exposure assessment with the doseresponse assessment to express the level of concern regarding a specific exposure scenario or set 

of scenarios. For systemic toxic effects, risk characterizations are presented as hazard quotients 

(HQs). A hazard quotient is the ratio of a projected level of exposure divided by some index of 

an acceptable exposure, such as an RfD. If the HQ is substantially less than one – i.e., the level 

of exposure is less than the level of acceptable exposure—there is no apparent cause for concern. 

If the hazard quotient is greater than unity, there is cause for concern. 

Because the hazard quotient does not describe dose-response or dose-severity relationships, a 

comparison of the magnitudes of the hazard quotients among different agents may not be a 

reliable index of relative risk and other types of information need to be considered. Hazard 

quotients that are close to a level of concern—i.e., between about 0.1 and 10—may be more 

difficult to interpret because of uncertainties in both the exposure estimates as well as the doseresponse relationships. While the range from 0.1 to 10 is somewhat arbitrary in terms of 

classifying the nature of concern, this is similar to the approach recently adopted by ATSDR 

(2004) in which concern for interactions of chemicals is triggered when individual hazard 

quotients exceed a value of 0.1. 

In order to reflect these gradations of concern in the general interpretation of hazard quotients, 

the comparative risk characterization is not organized by biological and chemical agents (as in 

the previous sections) but is organized by the nature of the hazard quotients: agents of marked 

concern (HQ>10), agents of marginal concern (HQ>0.1 but <10), and agents of minimal concern 

(HQ<0.1). The word minimal is emphasized because of the inherent limitation in all risk 

assessments. Risk assessments can never prove absolute safety—i.e., it is impossible to prove 

the negative, that something does not exist, in this case risk. Risk assessments, however, can be 

and are used to determine whether or not there is a basis for asserting that risk is plausible. 

An overview of the comparative risk characterization is summarized in Table 3-4 and illustrated 

in Figure 3-1. Of the agents considered in this risk assessment, the gypsy moth and DDVP are 

clearly agents of marked concern, although the nature of the concerns is different. If the gypsy 

moth is not controlled, population outbreaks will occur and humans will be exposed to large 

numbers of gypsy moth larvae. If this occurs, a substantial number of individuals will experience 

irritant effects that are sufficiently severe to cause these individuals to seek medical attention. 

No more serious effects are likely. For DDVP, the potential for risk is clear but the likelihood of 

observing risk seems to be remote. Under normal conditions and proper handling, exposures to 

DDVP will be negligible and risk will be inconsequential. Workers who mishandle a DDVPPVC strip or members of the general public who handle a DDVP-PVC may be exposed to levels 

of DDVP that are far above levels that would be considered acceptable. While such exposures 

clearly should be avoided, it seems unlikely that they would result in frank signs of toxicity. This 

conclusion is consistent with human experience in the use of DDVP resin strips. 

3-18
 
Diflubenzuron and tebufenozide are agents of marginal concern. Under most foreseeable 

conditions of exposure—i.e., exposure scenarios that might be characterized as typical—levels of 

exposure will be far below levels of concern. At the upper ranges of plausible exposure— levels 

that might be characterized as extreme—the hazard quotients for diflubenzuron approach a level 

of concern (HQs between 0.1 and 0.5 for both diflubenzuron and its 4-chloroaniline metabolite). 

For tebufenozide, the highest hazard quotient is 1.5, indicating that, although unlikely to cause 

overt signs of toxicity, the exposure would be characterized as undesirable. The somewhat 

higher hazard quotients for tebufenozide are attributed solely to the higher application rates for 

this compound, compared with diflubenzuron. 

Among the agents of minimal concern, B.t.k. is somewhat problematic. Based on the risk for 

serious adverse effects, there is clearly no cause for concern (the highest HQ is 0.04). As 

discussed in the dose-response assessment, this lack of concern is reenforced by a very 

aggressive and protective interpretation of the available toxicity data. Nonetheless, there is some 

residual concern with irritant effects. These effects are quite plausible in accidental cases of 

gross overexposure—e.g., splashing a formulation into the eye. These kinds of concern are 

minimal and are common to almost all chemical or biological agents. The more troubling 

concern involves studies of workers and non-workers who report irritant effects, primarily throat 

irritation. Whether or not these effects should be attributed to the B.t.k. exposure is unclear. 

The risk characterization for LdNPV and disparlure is unequivocal. Based on the available 

information, there is no plausible basis for concern that exposure will cause serious adverse 

effects. Again, various accidental exposures, such as splashing the agent into the eyes, might 

cause transient irritant effects. 

3.4.2. Agents of Marked Concern 

3.4.2.1. Gypsy Moth – Although the quantitative dose-response assessment is based on only one 

study, the study demonstrates a clear dose-response relationship and is supported by less 

quantitative reports of irritant effects associated with exposure to the gypsy moth as well as other 

lepidopteran larvae. In sparse to moderate infestations—i.e., egg mass densities of more than 

500 egg masses/acre—adverse effects involving skin irritation are not likely to be detectable in 

populations of exposed humans. Nonetheless, some individuals who have contact with gypsy 

moth larvae might develop skin irritation. In heavy gypsy moth infestations—i.e., from more 

than 500 to 5000 egg masses/acre—the occurrence of adverse skin reactions is expected to be 

hight, and the effects are likely be severe enough to cause some individuals to seek medical 

attention. In extreme outbreaks—i.e., greater than 5000 egg masses/acre— the effects will be 

qualitatively similar to those of severe infestations but may affect up to one-third of the 

population. Heavy infestations or extreme outbreaks may cause ocular and respiratory effects in 

some people; nonetheless, there is no way to quantify the likelihood of observing these effects. 

Similarly, severe infestations are often considered to be a nuisance and cause aesthetic damage to 

the environment. Both of these factors can lead to stress in some individuals. Young children 

may be a group at special risk from effects of gypsy moth exposure; however, it is not clear 

whether children are more sensitive than adults to the effects of gypsy moth exposure or whether 

3-19
 
responses in children appear greater because children spend more time outdoors compared with 

adults. 

3.4.2.2. DDVP – In most cases, exposures to both workers and members of the general public 

should be negligible. If workers take prudent steps to limit both dermal and inhalation 

exposures, the likelihood of exposures to DDVP reaching a level of concern appears to be very 

low. Similarly, members of the general public should not be exposed to substantial amounts of 

DDVP. The DDVP is contained within a PVC strip to ensure that the active ingredient is slowly 

released over a long period of time. The strip, in turn, is placed within a trap and the trap is 

placed so that it will not be accessed except in the case of intentional tampering or trap 

monitoring. 

Nonetheless, the risk assessment for DDVP develops exposure scenarios for both workers and 

members of the general pubic, which are intended to illustrate the potential effects of 

mishandling or tampering with DDVP strips. For workers, the greatest risks are associated with 

inhalation exposures from assembling the traps in enclosed and poorly ventilated spaces or 

transporting the traps in the passenger compartments of vehicles. These risks can be readily 

avoided. Dermal exposures can also lead to lesser but sill undesirable levels of exposure. For 

members of the general public, all of the exposure scenarios are accidental and some are extreme. 

The most likely of these is the accidental contamination of a small body of water. This scenario 

leads to exposures that are below the level of concern by a factor of about 25. If a child were to 

come into contact with a DDVP strip, however, both dermal and oral exposures could 

substantially exceed a level of concern. Although such exposures clearly should be avoided, it 

seems unlikely that they would result in frank signs of toxicity. This conclusion is consistent 

with human experience in the use of DDVP resin strips. 

3.4.3. Agents of Marginal Concern 

3.4.3.1. Diflubenzuron – The risk characterization for potential human health effects associated 

with the use of diflubenzuron is relatively unambiguous: none of the hazard quotients reach a 

level of concern at the highest application rate that could be used in USDA programs. In that 

many of the exposure assessments involve very conservative assumptions—that is, assumptions 

that tend to overestimate exposure—and because the dose-response assessment is based on 

similarly protective assumptions, there is no plausible basis for concluding that this use of 

diflubenzuron poses a hazard to human health. 

Notwithstanding the above assertion, it is worth noting that the greatest relative risk concerns the 

contamination of water with 4-chloroaniline rather than exposure to diflubenzuron itself. The 

highest hazard quotient for diflubenzuron is 0.1, a factor of 10 below a level of concern. Since 

this hazard quotient is based on toxicity, an endpoint that is considered to have a population 

threshold, it is reasonable to assert that the risk associated with exposure to diflubenzuron is 

essentially zero. 

3-20
 
Such is not the case with 4-chloroaniline, which is classified as a probable human carcinogen and 

is an environmental metabolite of diflubenzuron. For 4-chloroaniline, the highest hazard 

quotient is 0.4, below the level of concern by a factor of only 2.5. The scenario of greatest 

concern involves cancer risk from drinking contaminated water. This risk would be most 

plausible in areas with sandy soil and annual rainfall rates ranging from about 50 to 250 inches. 

The central estimate of the hazard quotient for the consumption of water contaminated with 

4-chloroaniline and based on a cancer risk of 1 in 1million is 0.09, which is 10 times lower than 

the level of concern. 

3.4.3.2. Tebufenozide – The similarities between tebufenozide and diflubenzuron have been 

emphasized throughout this comparative risk assessment. As noted in the dose-response 

assessment, the toxicities of these two compounds are virtually identical. While both 

diflubenzuron and tebufenozide are classified as agents of marginal concern—i.e., risk quotients 

between 0.1 and 10 —tebufenozide does exceed the level of concern, whereas diflubenzuron 

does not. This difference is due to the higher application rates that may be used with 

tebufenozide. These higher application rates for tebufenozide increase the levels of exposure, 

which results in somewhat higher hazard quotients for tebufenozide, compared with 

diflubenzuron. 

Nonetheless, as with diflubenzuron, there is no clear indication that adverse effects are likely to 

result from exposure to tebufenozide. At the maximum application rate considered in this risk 

assessment, two applications at 0.12 lb/acre spaced three days apart, there is little indication that 

adverse effects on human health are likely and only one scenario exceeds a risk quotient of 1. 

Based on central estimates of exposure— those that might be considered typical and 

expected—hazard quotients including workers and members of the general public range from 

0.00003 to 0.03, below a level of concern by factors ranging from approximately 30 to 33,000. 

At the upper range of plausible exposures, the hazard quotient for ground spray workers reaches 

a level of concern—i.e., a hazard quotient of 1. For members of the general public, the upper 

range of exposure leads to a hazard quotient of 1.5 for the longer-term consumption of 

contaminated vegetation following two applications at 0.12 lb/acre. Because of the linear 

relationship between exposure and application rate, two applications at 0.08 lb/acre would reach 

but not exceed a level of concern. With a single application at the maximum rate of 0.12 lb/acre, 

the hazard index is 0.8, below the level of concern. While the longer-term consumption of 

contaminated vegetation is probably not a likely scenario, it is a standard exposure scenario used 

in USDA risk assessments to consider the longer-term consumption of food items, like berries, 

that might be sprayed during the broadcast application of a pesticide. This risk assessment 

suggests that two applications at 0.08 lb/acre or more should be avoided in areas where members 

of the general public might consume contaminated fruits or other contaminated vegetation. 

3.4.4. Agents of Minimal Concern 

3.4.3.1. B.t.k. – The risk characterization regarding exposure to B.t.k. and its formulations is 

generally consistent with that of the previous USDA risk assessment as well as more recent risk 

assessments conducted by the U.S. EPA and the World Health Organization: B.t.k. and its 

3-21
 
formulations are likely to cause irritation to the skin, eyes, and respiratory tract; however, serious 

adverse health effects are implausible. Whether irritation is caused by B.t.k. in typical field 

applications used to control the gypsy moth is uncertain. While epidemiology studies involving 

self-reporting of symptoms do suggest that reports of irritant effects are to be expected, the 

biological plausibility of these effects is called into question because of an insubstantial dosedependency for the irritant effect. 

B.t.k. applications to control or eradicate the gypsy moth are not expected to cause serious 

adverse health effects in humans. At the extreme upper range of exposure in ground workers, 

exposure levels are estimated to be below the functional human NOAEL for serious effects by a 

factor of 25. For members of the general public, exposure levels are estimated to be below the 

functional human NOAEL by factors ranging from about 28,000 to 4,000,000 [4 million]. This 

assessment is based on reasonably good monitoring data, conservative exposure assumptions, 

and an aggressive and protective use of the available toxicity data. Based on these data, it is not 

likely that overt signs of toxicity will be observed in any group— ground workers, aerial 

workers, or members of the general public—exposed to B.t.k. as the result of gypsy moth control 

and eradication programs conducted by the USDA. 

There is no documented evidence of a subgroup of individuals who are more sensitive than most 

members of the general public to B.t.k. formulations. According to a recent epidemiology study, 

asthmatics are not likely to be affected adversely by aerial applications of B.t.k. The literature on 

B.t.k. includes one anecdotal claim of a severe allergy to a carbohydrate in a B.t.k. formulation; 

however, neither the claim nor observations of similar effects are substantiated in the available 

published epidemiology studies. On the other hand, B.t.k. formulations are complex mixtures, 

and the possibility that individuals may be allergic to some of the components in the formulations 

is acknowledged by a state health service. 

As noted in Section 3.1, pre-treatment with an influenza virus substantially increased mortality in 

mice exposed to various doses of B.t.k. Although the relevance of this observation to public 

health cannot be assessed well at this time, the viral enhancement of B.t.k. toxicity is likely to be 

an area of further study in the coming years. 

3.4.3.2. LdNPV (Gypchek) – There is no plausible basis for concern that either workers or 

members of the general public are at risk of adverse effects from the use of Gypchek to control 

the gypsy moth. This statement follows from the failure to identify any hazard associated with 

exposures to Gypchek or LdNPV and is essentially identical to the risk characterization given by 

the U.S. EPA. 

As discussed in both the exposure and dose-response assessments, the current risk assessment 

extends the U.S. EPA risk assessment by proposing a surrogate acute RfD and presenting a very 

conservative exposure assessment based on the accidental spray of a home garden. This 

approach is taken simply to facilitate the comparison of risks (or lack of risk) associated with 

Gypchek to the risks associated with other agents used to control the gypsy moth. Based on a 

3-22
 
relatively standard dose-response assessment and very conservative exposure assumptions, 

plausible exposures to Gypchek are below a level of concern by factors ranging from about 50 to 

more than 750. While more typical exposures—i.e., incidental exposure to Gypchek in water or 

air—are not provided, they will be substantially less than the range of accidental exposure 

scenarios used to quantify risk. 

3.4.3.3. Disparlure – Although there are studies regarding the acute toxicity of disparlure in 

laboratory animals, the lack of subchronic and chronic toxicity data precludes a quantitative 

characterization of risk. The available data regarding the acute toxicity of disparlure indicate that 

the potential hazard from exposure to the compound is low. 

The reliance on acute toxicity data introduces uncertainties into the risk assessment of disparlure 

that are quite different from the other better studied agents, and these uncertainties cannot be 

quantified. Other uncertainties in this analysis are associated with the exposure assessment and 

involve environmental transport and dermal absorption. These uncertainties are relatively minor 

compared with the lack of subchronic or chronic toxicity data. Thus, while there is no reason to 

believe that longer-term exposure to disparlure will produce adverse effects, this assumption can 

not be substantiated due to the lack of chronic toxicity data. The significance of this uncertainty 

is at least partially offset by the very low exposures that are plausible given the limited use of 

disparlure. For example, as noted in the dose-response assessment, inhalation exposures of mice 

3 to 5 mg/L (5,000,000 μg/m ) for 1 hour caused no mortality or signs of toxicity. As noted in the 

exposure assessment, likely concentrations of disparlure in air after applications comparable to 

3 those used in programs to control the gypsy moth are likely to be on the order of 0.004 μg/m , a

factor of 1,250,000,000 (1.25 billion) below the apparent NOEC for acute toxicity. This 

relationship is consistent with the general assumption made by the U.S. EPA that exposures to 

insect pheromones will be far below levels of concern (U.S. EPA 2004). 

3-23
 
4. ECOLOGICAL RISK ASSESSMENT
 

4.1. HAZARD IDENTIFICATION 

4.1.1. Overview 

An overview of the comparative hazard identification is given in Table 4-1. Unlike the human 

health risk assessment, in which the potential effects of the biological agents are similar, each of 

the ecological effects profiles of the biological agents considered in this risk assessment is quite 

distinct. The principal effect of the gypsy moth is damage to sensitive trees, which can be 

substantial. Because of the obvious importance of vegetation to the existence and habitat of most 

animals, defoliation by the gypsy moth will have numerous secondary effects in many other 

groups of organisms. There is, however, no indication that the gypsy moth will have direct 

effects on groups of organisms other than sensitive plants. LdNPV, on the other hand, is unlikely 

to have effects on species other than the gypsy moth. B.t.k. is toxic to nontarget Lepidoptera as 

well as the gypsy moth and some other lepidopteran species, but is unlikely to have direct effects 

on other groups of organisms. Thus, the potential effects of all of the biological agents are 

considered relatively specific, with LdNPV showing the greatest degree of specificity (only the 

gypsy moth), followed by the gypsy moth itself (several types of plants) and B.t.k. (several types 

of Lepidoptera). 

The chemical agents also differ in specificity: disparlure is most specific, tebufenozide is 

relatively specific to Lepidoptera, diflubenzuron is less specific and may affect many arthropods, 

and DDVP is a nonspecific biocide toxic to most groups of animals. As a pheromone, disparlure 

is almost as specific as LdNPV. It will attract the gypsy moth and two other closely related 

species, the nun moth (Lymantria monacha) and the pink gypsy moth (Lymantria fumida). Like 

the gypsy moth, both of these Lymantria species are forest pests, and adverse effects on these 

species are not a substantial concern for this risk assessment. In addition, the pink gypsy moth is 

native to Japan and is not found in the United States. A major qualification regarding the 

specificity of disparlure is the limited amount of information available on nontarget species. The 

data that are available indicate that the relative toxicity of disparlure to Daphnia magna, a 

commonly used test species in aquatic toxicity studies, is high. Diflubenzuron and tebufenozide 

are clearly toxic to mammals and at least some arthropods. In mammals, exposure to either 

chemicals causes adverse effects in blood (methemoglobinemia), as discussed in the human 

health risk assessment. In terrestrial and aquatic arthropods, exposure to either chemical 

interferes with growth and development. Because of differences in the mechanism of action of 

diflubenzuron and tebufenozide, the toxicity of tebufenozide appears to be somewhat more 

selective. For instance, effects in birds have been clearly demonstrated for tebufenozide but not 

for diflubenzuron. Nonetheless, it is plausible to speculate that both diflubenzuron and 

tebufenozide are likely to cause adverse hematological effects in birds, similar to those observed 

in mammals exposed to these chemicals. In terms of its mechanism of action, DDVP is a general 

neurotoxin. In all animals that have nervous systems that involve acetylcholinesterase (AChE) 

and use acetylcholine (ACh) as a neurotransmitter (a substance necessary to make the nerves 

work properly), DDVP will be toxic and sufficiently high exposures to DDVP will be lethal. The 

definition of sufficiently high, however, is critical and variable. Although DDVP is not selective 

4-1
 
mechanistically, differences in sensitivity among species are substantial. Insects are much more 

sensitive than mammals or other higher organisms to DDVP. This difference in sensitivity is 

what characterizes DDVP as an effective insecticide that can be used safely. 

4.1.2. Biological Agents 

4.1.2.1. Gypsy Moth – The clearest primary effect of gypsy moth infestations is on terrestrial 

plants, primarily trees. Various instars of the gypsy moth larvae will feed on host trees and can 

cause extensive defoliation which can kill some of the infested trees. On a larger scale, the 

extensive defoliation and/or death of trees may result in secondary changes to vegetation, which 

will, in turn, affect other forms of vegetation as well as various animal species (primarily related 

to changes in habitat). Gypsy moth larvae appear to have definite food preferences; oak, birch, 

poplar, and apple trees seem to be their favorite food sources. While both the European and 

Asian gypsy moth cause similar types of damage (i.e., defoliation), their feeding preferences are 

somewhat different, with the Asian gypsy moth preferring a wider range of vegetation. Heavy 

defoliation is much more common among the oaks than among trees that are not particularly 

favored as food by the gypsy moth. For susceptible oaks, the effects of infestations on tree 

mortality varies according to the initial condition of the stand and the number of infestations. 

Generally, gypsy moth infestations result in mortality of less than 15% of total basal area—i.e., 

mortality of trees involving 15% the total area of the tree trunks near the ground. When heavy 

defoliation is followed by massive overstory mortality, existing shrub and herb cover increase 

dramatically due to increases in available light, moisture, and nutrients. Extensive loss of the 

existing canopy will also favor the growth of tree species that are intolerant to shade and will 

shift the forest ecosystem towards earlier successional stages. 

The only other groups of organisms likely to be affected directly by the gypsy moth are some and 

probably very few other lepidopteran species, including the northern tiger swallowtail butterfly. 

The mechanisms for direct adverse effects on other lepidopteran species may include bacterial 

contamination of the leaves by gypsy moth larvae and a decrease in the nutritional value of the 

leaves damaged by the gypsy moth. Most studies, however, do not indicate substantial direct 

effects on other insects, including Lepidoptera. In some cases, increases may be seen in 

populations of insect predators of the gypsy moth. 

There is no evidence in the literature of direct adverse effects of the gypsy moth on most groups 

of animals. Indirect effects, associated with damage to vegetation, may be of substantial 

consequence to some species, including squirrels, mice, and other mammals that rely on acorns. 

Although some mammals consume insects, including the gypsy moth, there is no evidence that 

gypsy moth outbreaks have a substantial impact on insectivorous mammals. Similarly, birds and 

aquatic species are not likely to be affected directly or adversely by the gypsy moth. In some 

species of birds, gypsy moth infestations and subsequent defoliation may be beneficial, especially 

for those species favoring dead wood as a habitat. 

4.1.2.2. B.t.k. – The hazard identification for mammals is closely related to the hazard 

identification for the human health risk assessment in that both are based, in part, on numerous 

4-2
 
standard toxicity studies in experimental mammals. Although B.t.k. may persist in mammals for 

several weeks after exposure, there is little indication that oral or dermal exposure leads to 

serious adverse effects. Most inhalation studies do not suggest a potential for adverse effects 

even at B.t.k. concentrations much greater than those likely to be encountered in the environment. 

The lack of a positive hazard identification is supported by field studies which demonstrate a lack 

of adverse effects in populations of mammals after applications of B.t.k. 

Toxicity studies in birds are limited to standard acute exposures required by U.S. EPA for 

product registration. The studies all involve either single-dose gavage administration or five 

daily dose gavage administrations, and none of the studies reports signs of toxicity or 

pathogenicity at single oral doses up to 3333 mg formulation/kg bw or at multiple oral doses up 

to 2857 mg formulation/kg bw. Due to the lack of toxicity of B.t.k. formulations as well as other 

B.t. strains, the U.S. EPA did not require chronic or reproductive toxicity studies in birds. This 

apparent lack of toxicity is supported by numerous field studies in birds. In one field study, a 

transient decrease in abundance was noted in one species, the spotted towhee (Pipilo maculatus). 

This observation is inconsistent with other field studies on B.t.k., and, according to the 

investigators, may be an artifact of the study design. 

The mechanism of action of B.t.k. in Lepidoptera is relatively well characterized. B.t.k. 

vegetative cells produce spores and crystals. After the insect consumes the crystals, toxins are 

formed that attach to the lining of the mid-gut of the insect and rupture the cell walls. The B.t.k. 

spores germinating in the intestinal tract enter the body cavity through the perforations made by 

the crystal toxins and replicate causing septicemia and eventually death. While various strains of 

B.t. are often characterized as selective pesticides, B.t.k. is toxic to several species of target and 

nontarget Lepidoptera. Sensitive nontarget Lepidoptera include larvae of the Karner blue 

butterfly, two species of swallowtail butterflies, a promethea moth, the cinnabar moth, and 

various species of Nymphalidae, Lasiocampidae, and Saturniidae. 

While some nontarget lepidopteran species appear to be as sensitive as target species to B.t.k., 

most studies indicate that effects in other terrestrial insects are likely to be of minor significance. 

There is relatively little information regarding the toxicity of B.t.k. or B.t.k. formulations to 

terrestrial invertebrates other than insects. Some oil-based B.t.k. formulations may be toxic to 

some soil invertebrates; however, the toxicity is attributable to the oil in the formulation and not 

to B.t.k. There is no indication that B.t.k. adversely affects terrestrial plants or soil 

microorganisms. 

The U.S. EPA classifies B.t.k. as virtually non-toxic to fish, and this assessment is consistent 

with the bulk of experimental studies reporting few adverse effects in fish exposed to B.t.k. 

concentrations that exceed environmental concentrations associated with the use of B.t.k. in 

USDA programs. Although there are no data regarding the toxicity of B.t.k. or its formulations to 

amphibians, other strains of B.t. appear to have low toxicity to amphibians. The effects of B.t.k. 

on aquatic invertebrates is examined in standard laboratory studies and in numerous field studies. 

At concentrations high enough to cause decreases in dissolved oxygen or increased biological 

4-3
 
oxygen demand, B.t.k. may be lethal to certain aquatic invertebrates, like Daphnia magna. Most 

aquatic invertebrates, however, seem relatively tolerant to B.t.k. This assessment is supported by 

several field studies that have failed to note remarkable effects in most species after exposures 

that substantially exceed expected environmental concentrations. As with effects on terrestrial 

plants, the toxicity of B.t.k. to aquatic plants has not been tested. 

The U.S. EPA’s Office of Pesticides (U.S. EPA/OPP 1998) has raised concerns that some 

batches of B.t. may contain heat labile exotoxins that are toxic to Daphnia. The production of 

these toxins is an atypical event thought to be associated with abnormal or poorly controlled 

production processes. The U.S. EPA requires manufacturers to submit a daphnid study on each 

new manufacturing process to demonstrate that heat labile exotoxin levels are controlled. 

4.1.2.2. LdNPV – Similar to the hazard identification for the human health risk assessment, the 

hazard identification for nontarget wildlife species fails to identify any adverse effects of 

concern—i.e., there is no indication that LdNPV or the Gypchek formulation of LdNPV has the 

potential to cause adverse effects in any nontarget species. The mammalian toxicity data base for 

LdNPV is reasonably complete and indicates that LdNPV is not pathogenic or otherwise toxic to 

mammals. One specific study conducted on wildlife mammals that may consume contaminated 

gypsy moth larvae indicates no adverse effects in mice, shrews, and opossums. Relative to the 

large number available studies in mammals, few studies are available in birds but the results of 

these studies are nearly identical to those in mammals indicating that exposures to LdNPV at 

levels substantially higher than those likely to occur in the environment will not be associated 

with adverse effects. Based on bioassays of LdNPV on numerous nontarget insect species and 

supported by the generally high species specificity of related baculoviruses, the hazard 

identification for LdNPV in nontarget insects is strikingly similar to that in birds and mammals. 

There is no indication LdNPV will cause adverse effects in nontarget insects at any level of 

exposure. Relatively few studies regarding the toxicity of LdNPV have been conducted in fish or 

aquatic invertebrates; nevertheless, these studies are consistent with studies in terrestrial species, 

indicating a lack of toxicity to fish and aquatic invertebrates. No data are available on the effects 

of LdNPV on amphibians, aquatic or terrestrial plants, or other microorganisms. While this lack 

of information does, by definition, add uncertainty to this risk assessment, there is no basis for 

asserting that effects on these or other organisms are plausible. 

4.1.3. Chemical Agents 

4.1.3.1. Diflubenzuron and Tebufenozide – The toxicity of diflubenzuron and tebufenozide is 

well characterized in most groups of animals, including mammals, birds, terrestrial invertebrates, 

fish, and aquatic invertebrates. In general, both of these compounds are much more toxic to 

some invertebrates, specifically arthropods, than to vertebrates or other groups of invertebrates. 

This differential toxicity of these two compounds involves fundamentally different and well 

understood mechanisms of action, with tebufenozide being somewhat more selective than 

diflubenzuron. Toxicity of diflubenzuron to sensitive invertebrate species is based on the 

inhibition of chitin synthesis. Chitin is a polymer (repeating series of connected chemical 

4-4
 
subunits) of a glucose-based molecule and is a major component of the exoskeleton, outer body 

shell, of all arthropods. The inhibition of the formation of chitin disrupts the normal growth and 

development of insects and other arthropods. Both terrestrial and aquatic arthropods are affected, 

but there seems to be some substantial differences in sensitivity. The toxicity of tebufenozide to 

sensitive invertebrates is based on the mimicking of 20-hydroxyecdysone, an invertebrate 

hormone that controls molting. The effectiveness of tebufenozide in mimicking 20­

hydroxyecdysone activity, however, appears to vary markedly among orders and species of 

invertebrates. In general, moths are sensitive to tebufenozide but other insects are much less 

sensitive. 

The most sensitive effects in wildlife mammalian species and possibly other vertebrates exposed 

to diflubenzuron or tebufenozide are likely to be the same as those in experimental mammals 

(i.e., effects on the blood). The major difference between the hazard identification for 

diflubenzuron and tebufenozide concerns potential reproductive effects. As noted in the 

comparative human health risk assessment, tebufenozide may cause reproductive effects in 

mammals, while this effect has not been noted for diflubenzuron. Similarly, the reproductive 

effects of tebufenozide but not diflubenzuron are of concern for birds, although the data are 

somewhat inconsistent. The available studies on tebufenozide include a reproduction study 

investigating effects in mallard ducks and two reproduction studies investigating effects in 

bobwhite quail. In one of the quail studies, dietary concentrations of 300 and 1000 ppm caused 

reproductive effects. The effects were not observed in that study at 100 ppm; moreover, the 

effects were not observed in the more recent quail study or in the study on mallard ducks. A field 

study regarding the effects of tebufenozide on reproductive performance in birds noted trends 

that were not statistically significant but, nonetheless, suggestive of adverse reproductive effects 

in a warbler species. Thus, consistent with the interpretation by the U.S. EPA, reproductive 

effects in both mammals and birds are considered endpoints of concern for tebufenozide. For 

diflubenzuron, there is only one relatively old report of reproductive effects in birds and the 

effects reported have not been noted in other studies. Thus, also consistent with the approach 

taken by U.S. EPA, reproductive effects are not identified as an endpoint of concern for 

diflubenzuron. 

Terrestrial invertebrates appear to be much more sensitive to diflubenzuron and tebufenozide 

than are vertebrates, and tebufenozide appears to affect a narrower group of invertebrates than 

does diflubenzuron. The terrestrial species most sensitive to diflubenzuron are arthropods, a 

large group of invertebrates, including insects, crustaceans, spiders, mites, and centipedes. In 

terrestrial organisms, the species most sensitive to diflubenzuron include lepidopteran and beetle 

larvae, grasshoppers, and other herbivorous insects. More tolerant species include bees, flies, 

parasitic wasps, adult beetles, and sucking insects. Tebufenozide is toxic to a much narrower 

range of terrestrial insects. In general, moths are sensitive to tebufenozide but other insects are 

much less sensitive. 

Both diflubenzuron and tebufenozide are also more toxic to aquatic invertebrates than they are to 

fish. U.S. EPA has classified diflubenzuron as practically non-toxic to fish, with LC50 values that 

4-5
 
range from 25 to 500 mg/L. Tebufenozide is somewhat more toxic to fish, with LC50 values that 

range from 2.2 to 6.5 mg/L for fish, categorized as moderately toxic using the U.S. EPA 

classification system. Invertebrates are affected at much lower concentrations and the relative 

potency of the two compounds is reversed, with diflubenzuron being substantially more toxic 

than tebufenozide to aquatic invertebrates. The NOEC values in invertebrates for diflubenzuron 

are as low as 0.3 μg/L in acute studies and 0.04 μg/L in longer-term studies. Tebufenozide is 

substantially less toxic to invertebrates, with NOEC values as low as 120 μg/L in acute studies 

and 3.5 μg/L in longer-term studies. 

4.1.3.2. DDVP – Although DDVP is a general neurotoxin, the available data suggest that 

invertebrates are more sensitive than other organisms to DDVP. For example, the oral LD50 in 

honey bees is 0.29 mg/kg bee, and the topical LD50 is 0.65 mg/kg bee. Although DDVP is also 

toxic to birds, the oral LD50 value is about 10 mg/kg for the most sensitive species. Short-term 

repeat dose studies in mammals found that oral exposures to doses below about 0.5 mg/kg-day or 

inhalation exposures to 1–2 mg/m3 generally do not result in adverse effects. Thus, no effects are 

apparent in experimental mammals at doses that are clearly lethal to bees. 

Aquatic animals are also sensitive to DDVP. As with terrestrial animals, invertebrates appear to 

be more sensitive than vertebrates. The lowest reported LC50 value in fish is approximately 0.2 

mg/L. Some aquatic invertebrates are much more sensitive than fish to DDVP. For daphnids, 

the most sensitive group of invertebrate species, reported EC50 values range from 0.00007 to 

0.00028 mg/L. 

Most of the toxicity data on ecological receptors is limited to free DDVP, rather than a slowrelease formulation like the Vaportape II product used in USDA programs to control the gypsy 

moth. Hence, the toxicity values reported for indicator species are likely to be conservative (i.e., 

suggest greater toxicity), as compared with Vaportape II. Although U.S. EPA assessed the 

ecological effects of DDVP, the exposures assessed are not specific to formulations in which 

DDVP is encapsulated in PVC resin. In general, aside from those organisms that enter the milk 

carton trap or those that remove the PVC strip form the trap, toxicity resulting from exposure of 

ecological receptors to DDVP in Vaportape II milk carton traps is not likely. 

4.1.3.3. Disparlure – There is very little information regarding the toxicity of disparlure to 

nontarget wildlife species. As noted in the human health risk assessment, the U.S. EPA does not 

require extensive testing of insect pheromones. Thus, the only studies available are acute studies 

in bobwhite quail, mallard ducks, rainbow trout, bluegill sunfish, Daphnia magna, and Eastern 

oysters. No chronic exposure studies were identified. 

Results of acute gavage and dietary exposure studies in mallard ducks and bobwhite quail show 

that disparlure has very low toxicity in these species, with no mortalities observed following 

exposure to up to 2510 mg/kg in bobwhite quail. Limited data are available regarding the 

toxicity disparlure to aquatic animals. Relative to mammals and birds, Daphnia appear to the 

4-6
 
most sensitive species tested, with an LC50 value of 0.098 mg/L. In rainbow trout, 20% mortality 

was noted at a concentration of 100 mg/L. 

4-7
 
4.2. EXPOSURE ASSESSMENT 

4.2.1. Overview 

Table 4-2 summarizes the exposure assessments on nontarget species for each of the agents 

covered in the risk assessment. Table 4-2 is similar to the corresponding table for the human 

health risk assessment (see Table 3-2) because the applications and uses for each control agent 

are identical. Since diflubenzuron, tebufenozide, LdNPV, and disparlure can be applied in 

broadcast applications, exposure potential is high and in many cases unavoidable, as is true for 

the human health risk assessment. When disparlure is used as an attractant in traps, exposures 

will be variable and primarily incidental. Exposures to the gypsy moth itself are also variable 

and depend on the extent of the gypsy moth population, which can range from low level 

infestation to outbreak conditions. 

There are, however, notable differences between the human health exposure assessment and the 

ecological exposure assessment. Unlike Table 3-2, Table 4-2 does not provide measures of 

exposure for each agent, because the measures of exposure for ecological effects vary not only 

among the control agents but also among the target groups for each agent. For example, 

exposures to the gypsy moth are measured as egg masses/acre in the human health risk 

assessment, which is the same measure of exposure used for terrestrial vegetation, because it is 

the primary determinant in the dose-response assessment for plants. For all other species, 

however, effects from the gypsy moth are most likely to be secondary, which means the exposure 

assessment for these indirectly affected species is based on defoliation—i.e., the result of the 

dose-response assessment for terrestrial vegetation is used as the exposure assessment for most 

other groups of organisms. 

Other differences in the exposure assessments for nontarget species are mostly superficial. For 

each of the chemical agents, the mass of the chemical is typically used as the measure of 

exposure. Depending on the group, the measure of exposure may be expressed as dose (mg 

agent/kg bw for most terrestrial species), concentration (mg agent/L of water for aquatic species), 

or simply as application rate (lb agent/acre). This last measure is used primarily when field 

studies are the basis for the dose-response assessment. 

As in the human health risk assessment, different measures of exposure are used for each of the 

biological agents. For B.t.k., most of the exposures are characterized simply as an application 

rate in units of BIU/acre. Nevertheless, colony forming units are used for some of the 

mammalian exposure scenarios. Also as in the human health risk assessment, no clear hazard 

potential is identified for LdNPV. The very few exposure scenarios that are quantified in the 

ecological risk assessment for LdNPV are based on the mass of the formulation, Gypchek. 

The level of detail used in the exposure assessments for the different chemicals reflects 

differences in the use patterns and the nature of the available toxicity data. Full sets of exposure 

assessments in several groups of animals are developed for diflubenzuron and tebufenozide. As 

in the human health risk assessment, the exposure assessment for diflubenzuron is elaborated by 

4-8
 
the consideration of 4-chloroaniline and the exposure assessment for tebufenozide is elaborated 

by the consideration of multiple applications. 

Disparlure, which may also be applied in aerial broadcast applications, has a much more 

restricted set of exposure scenarios on far fewer groups of organisms. This difference is due 

completely to the sparse toxicity data available on this compound. In other words, while a very 

elaborate set of exposure scenarios could be prepared, these scenarios would serve little purpose 

because they could not be combined with a dose-response assessment to characterize risk. The 

exposure assessment for DDVP is also restricted due to the limited number of plausible 

exposures, given that DDVP is used only in milk carton traps and minimal exposures for 

nontarget species are anticipated under ordinary conditions. 

4.2.2. Biological Agents 

4.2.2.1. Gypsy Moth – As in the human health risk assessment, the exposure metameter is 

dictated by the data used to formulate the dose-response assessment—i.e., egg mass density is the 

exposure metameter for terrestrial invertebrates and plants because it is the measure on which the 

dose-response assessment is based. Egg mass densities ranging from 5 to 5000 egg masses/acre 

are used to estimate responses in terrestrial plants and invertebrates. 

Most wildlife species are not affected directly by exposure to the gypsy moth but are more likely 

to experience indirect effects due to changes in habitat or other environmental conditions 

secondary to defoliation. Consequently, the exposure assessment for most wildlife species is 

almost identical to the dose-response assessment for terrestrial plants which is expressed as 

defoliation caused by gypsy moth larvae. For this exposure assessment, categories of defoliation 

are defined as normal background defoliation (<30% defoliation), moderate defoliation (30-60% 

defoliation), and high or severe defoliation (>60% defoliation). 

4.2.2.2. B.t.k. – Based on the hazard identification, exposure assessments are presented for three 

groups: small mammals, terrestrial insects, and aquatic species. While a number of different 

exposure scenarios could be developed for terrestrial mammals, the only positive hazard 

identification for B.t.k. involves inhalation exposures. As in the human health risk assessment, 

inhalation exposures ranging from 100 to 5000 cfu/m3

 are used to assess potential risks of serious 

adverse effects in terrestrial vertebrates. These concentrations are applied to a 20 g mouse and 

correspond to inhaled doses of 0.00336-0.168 cfu/mouse. While there is no basis for asserting 

that any oral and/or dermal exposures are likely to cause adverse effects in terrestrial vertebrates, 

an extremely conservative exposure assessment is developed for combined oral (water and 

vegetation) and dermal (direct spray) exposures that yields an estimated maximum dose of about 

184 mg/kg body weight. For terrestrial insects, the toxicity values used to assess the 

consequences of observing effects is given in units of BIU/ha. Consequently, the exposure 

assessment for this group is simply the range of application rates used in USDA programs —i.e., 

approximately 49-99 BIU/ha. For aquatic organisms, toxicity data are expressed in several 

different units such as mg formulation/L, IU/L, and cfu/L. Based on application rates used in 

USDA programs and conservative assumptions concerning the depth of water over which B.t.k. 

4-9
 
might be sprayed, concentrations in water would be expected to be at or below 0.24 mg 

formulation/L. As discussed in the hazard identification, there is no basis for asserting that 

adverse effects in birds, plants, soil microorganisms, or soil invertebrates other than insects are of 

plausible concern. Consequently, explicit exposure assessments are not conducted for those 

groups. 

4.2.2.3. LdNPV – Numerous wildlife species might be exposed to Gypchek or LdNPV as a result 

of ground and aerial applications of the Gypchek formulation. The need for any formal risk 

assessment is questionable, however, because neither Gypchek nor LdNPV appear to cause 

systemic adverse effects. Nonetheless, to provide some basis for comparing the potential risks of 

Gypchek with other agents used to control the gypsy moth, two extreme exposure assessments 

are developed: one for a terrestrial herbivore consuming contaminated vegetation and the other 

for aquatic organisms in a small pond directly sprayed with Gypchek at the highest application 

rate. For the terrestrial herbivore, the dose estimates range from 1.1 to 3.2 mg Gypchek /kg bw. 

For aquatic organisms, concentrations are expressed in units of PIB/L because this unit is used in 

the corresponding toxicity studies. For a small pond directly sprayed with Gypchek at the highest 

5 application rate, the estimated initial concentration is 2.5x10 PIB/L. Several less extreme 

exposure assessments could be developed but they would not alter the risk assessment given that 

the extreme exposure assessments are substantially below any level of concern. 

4.2.3. Chemical Agents 

4.2.3.1. Diflubenzuron and Tebufenozide – As in the human health risk assessment, the 

exposure assessments for diflubenzuron and tebufenozide are similar. The same set of exposure 

scenarios are used with the same set of potential target species. The difference in their 

application rates dominates the quantitative difference in projected exposure to these two 

chemicals: a single application rate of 0.0625 lb/acre for diflubenzuron and one or two 

applications at 0.12 lb/acre for tebufenozide. As a result of the higher application rate for 

tebufenozide, all exposures are higher for tebufenozide than for diflubenzuron. Also as in the 

human health risk assessment, the exposure assessments for diflubenzuron are elaborated to 

include 4-chloroaniline as an environmental metabolite of diflubenzuron. 

Not withstanding the quantitative differences in the application rates, the patterns of exposure for 

terrestrial species for diflubenzuron and tebufenozide are similar except for the maximum acute 

exposure. For diflubenzuron, this exposure is associated with direct spray of a small mammal 

and could reach 10 mg/kg. For tebufenozide, the maximum acute exposure is associated with a 

fish-eating bird and could be as high as 85 mg/kg. For other acute and longer-term exposures, 

the consumption of contaminated vegetation results in higher levels of exposure to both 

compounds than does the consumption of contaminated water. Estimates of longer-term daily 

doses for a small mammal consuming contaminated vegetation at the application site range up to 

0.005 mg/kg for diflubenzuron and 0.08 mg/kg/day for tebufenozide. The consumption of 

contaminated water by a small mammal results in estimated doses of up to 0.00001 mg/kg/day 

for diflubenzuron and 0.0002 mg/kg/day for tebufenozide. Exposures of terrestrial organisms to 

4-chloroaniline as a degradation product of diflubenzuron tend to be much lower than the doses 

4-10
 
for diflubenzuron. The highest acute exposure to 4-chloroaniline is about 0.2 mg/kg, the 

approximate dose for the consumption of contaminated water by a small mammal and the 

consumption of contaminated fish by a predatory bird. The highest longer term exposure to 

4-chloroaniline is 0.0002 mg/kg/day, the dose associated with the consumption of contaminated 

vegetation by a large bird. 

As dicussed in Section 4.3, the toxicity data on terrestrial invertebrates are much more extensive 

for diflubenzuron than tebufenozide, which is directly related to differences in the numbers of 

field studies available on diflubenzuron (many), compared with tebufenozide (very few). The 

difference reflects the long-time, extensive use of diflubenzuron, compared with tebufenozide, 

which is a more recently introduced insecticide. For both chemicals, exposure of terrestrial 

invertebrates is generally expressed as an application rate from a field study, and no formal 

exposure assessment is given. 

Exposures of aquatic organisms to diflubenzuron or tebufenozide are based essentially on the 

same information used to assess the exposures of terrestrial species from contaminated water. At 

the maximum application rates, the upper range of the expected peak concentration in surface 

water is estimated at 16 μg/L for diflubenzuron and 40 μg/L for tebufenozide. 

4.2.3.2. Disparlure – Given the apparent low acute toxicity of disparlure and the lack of any 

chronic toxicity data, an exposure assessment for terrestrial species would not add to the 

assessment of risk. Acute exposure studies in Daphnia and rainbow trout show that aquatic 

animals appear more sensitive than terrestrial animals to disparlure. Therefore, an exposure 

assessment for aquatic species is made based on aerial spray of a pond at an application rate of 

29.1 g a.i./acre, with an estimated concentration in pond water of 0.0072 mg a.i./L. 

4.2.3.3. DDVP – As in the human health risk assessment, exposure of terrestrial mammals to 

DDVP from the VaporTape strips used in milk carton traps is likely to be negligible under most 

circumstances. Nonetheless, it is conceivable that some mammals such as racoons or bears could 

easily access and tamper with the milk carton trap. Depending on the proportion of the DDVP 

strip that is consumed, doses (as DDVP in the PVC strip) are estimated to range from 10.5 mg/kg 

(10% of strip) to 105 mg/kg (100% of strip) and the central estimate is taken as 31.6 mg/kg (30% 

of strip). In addition, contamination of water with a pest strip is plausible, although probably 

rare, and is considered in a manner similar to the corresponding scenario in the human health risk 

assessment (see Section 3.2.3.4). This scenario is based on the consumption of contaminated 

water by a small mammal, and the dose to the animal is estimated at about 0.00003 mg/kg with a 

range from 0.000009 to 0.00009 mg/kg. Other exposure scenarios for terrestrial vertebrates, 

while possible, seem far less plausible and are not considered quantitatively. No quantitative 

exposure assessments for terrestrial invertebrates are developed because the milk carton trap will 

attract only male gypsy moths. Nontarget insects that incidentally enter the trap are likely to be 

killed by exposure to the DDVP vapor. Exposures to aquatic species are based on the same 

water concentrations used for terrestrial species: 0.000177 mg/L with a range from 0.000059 to 

0.00059 mg/L. 

4-11
 
4.3. DOSE-RESPONSE ASSESSMENT 

4.3.1. Overview 

An overview of the dose-response assessment for groups of nontarget species is presented in 

Table 4-3. The information in this table categorizes the data descriptively rather than in terms of 

data quality. The categories reflect whether the data are sufficient to quantify risk or 

quantitatively characterize differences in sensitivity among several species in the designated 

group (!), whether the dose-response assessment is based on both an effect and no-effect level 

(�), whether the dose-response assessment is based only on a no-effect level (G), or whether the 

assessment is based only on an effect level ("). These categories are reasonable measures of data 

quality for all of the agents covered in this risk assessment except LdNPV. 

All of the risk values for LdNPV are based on no- effect concentrations or doses. In general, 

confidence in any dose-response relationship is enhanced if a clear dose-response relationship 

can be demonstrated and both effect and no-effect exposures have been identified. In the case of 

LdNPV, however, there is simply no indication that LdNPV or the Gypchek formulation will 

cause toxicity in any nontarget species at any dose level. While additional studies could be 

conducted at higher doses and while these studies would enhance confidence in the risk 

assessment, the NOAEL and NOEC values that have been identified are far above any plausible 

exposures. Thus, while based on limited data in terms of the dose-effect characterization, the 

dose-response assessment for LdNPV is adequate for risk characterization. 

For most of the other agents, the dose-response assessments are reasonably good for the species 

of greatest concern. As noted in Table 4-3, dose-response assessments for DDVP are derived 

only for mammals, fish, and aquatic invertebrates. As discussed in the exposure assessment, this 

limited approach is taken with DDVP because of the limited use of DDVP in programs to control 

the gypsy moth. The DDVP is contained in a PVC strip that is placed in a milk carton trap that 

includes disparlure as an attractant for the gypsy moth. This type of use limits potential exposure 

for most nontarget species. A formal dose-response assessment is not conducted for terrestrial 

invertebrates. This is not due to any lack of data. The toxicity of DDVP to insects and many 

other invertebrates is very well characterized. DDVP is such a potent insecticide that no formal 

dose-response assessment is needed. Insects and many other species that enter the trap are likely 

to be killed by exposure to DDVP. 

Disparlure is the other agent for which a full set of dose-response assessments are not conducted. 

As discussed in the hazard identification, this is due to the limited amount of data regarding the 

toxicity of disparlure to nontarget species. 

Relatively full dose-response assessments on groups of greatest concern are given for the gypsy 

moth, B.t.k., diflubenzuron and its 4-chloroaniline metabolite, and tebufenozide. For the gypsy 

moth, the effect of primary concern is damage to vegetation. While data are available on both 

lethality in trees as well as defoliation, defoliation is used as the sublethal effect of primary 

concern. A dose-response assessment is also given for nontarget lepidopterans. While effect and 

4-12
 
no-effect levels can be identified, the significance of this effect is questionable. In terms of direct 

effects, terrestrial vegetation is the primary target of concern. 

Lepidoptera are the primary nontarget group of concern for B.t.k. exposure. A relatively rich set 

of studies is available regarding the sensitivities of nontarget Lepidoptera and some other insects. 

The sensitivities of the nontarget insects can be quantified reasonably well from exposures that 

encompass the application rates used in USDA programs to control the gypsy moth. Sensitive 

nontarget Lepidoptera include larvae of the endangered Karner blue butterfly as well as several 

other types of moths. 

Similar types of information are available on diflubenzuron and tebufenozide, and dose-response 

assessments can be made for the species of primary concern. For both chemicals, this includes 

nontarget Lepidoptera and aquatic invertebrates. Other terrestrial arthropods are also considered 

for diflubenzuron. In addition, because of the standard tests required by U.S. EPA for the 

registration of most pesticides, adequate toxicity data are available on mammals, birds, and fish. 

The toxicity data base for diflubenzuron is somewhat more extensive and sensitivities in 

nontarget organisms are somewhat better defined in both laboratory and field studies than is the 

case with tebufenozide. 

4.3.2. Biological Agents 

4.3.2.1. Gypsy Moth – As in the human health risk assessment for the gypsy moth, the dose 

measure for the gypsy moth is egg masses/acre. Quantitative dose-response assessments can be 

made for both terrestrial plants and sensitive species of Lepidoptera. The dose-response 

assessments for terrestrial plants are based on a relatively simple quantitative model for the 

relationship of egg mass density and vegetation type to defoliation. Three broad categories of 

vegetation (sensitive, intermediate, and tolerant) are used to characterize the susceptibility of 

forest stands to gypsy moth induced defoliation. Estimated LOAEL values based on 30% 

defoliation, which is considered the lower range of moderate defoliation, are approximately 125 

egg masses/acre for sensitive stands, 1000 egg masses/acre for intermediate stands, and 7000 egg 

masses/acre for tolerant stands. The corresponding NOAEL values, defined as 10% defoliation, 

are estimated as 12, 20, and 125 egg masses/acre for sensitive, intermediate, and tolerant forest 

stands. 

The effects of gypsy moth exposure on sensitive terrestrial invertebrates, including some species 

of Lepidoptera, are less well documented and less well characterized, compared with the effects 

on terrestrial plants. Nonetheless, available studies indicate that the NOAEL for adverse effects 

in certain other species of Lepidoptera are lower than the NOAEL for sensitive forest stands—– 

i.e., about 6-72 egg masses/acre for some Lepidoptera. No quantitative dose-response 

assessment is presented for other groups of organisms—e.g., mammals, birds, and soil or aquatic 

organisms. The impact of gypsy moth exposure on these species is most likely to result in 

indirect effects secondary to defoliation. This is discussed further in the risk characterization. 

4-13
 
4.3.2.2. B.t.k. – As summarized in Table 4-3, exposure assessments are presented for four 

groups: mammals, terrestrial insects, fish, and invertebrates. While a number of different 

exposure scenarios could be developed for terrestrial mammals, the only positive hazard 

identification for B.t.k. involves inhalation exposures. As in the human health risk assessment, 

inhalation exposures of 100-5000 cfu/m3

 are used to assess potential risks of serious adverse 

effects in terrestrial vertebrates. These concentrations are applied to a 20 g mouse and 

correspond to inhaled doses of 0.00336-0.168 cfu/mouse. While there is no basis for asserting 

that any oral and/or dermal exposures are likely to cause adverse effects in terrestrial vertebrates, 

an extremely conservative exposure assessment is developed for combined oral (water and 

vegetation) and dermal (direct spray) exposures that yields an estimated maximum dose of 

approximately 184 mg/kg body weight. 

For terrestrial insects, the toxicity values used to assess the consequences of observing effects is 

given in units of BIU/ha over a range of applications similar to those used in gypsy moth control 

programs. The magnitude of response to B.t.k. in sensitive nontarget species appears similar to 

that of the gypsy moth. Tolerant species appear to be about 30-fold less sensitive than the gypsy 

moth to B.t.k.. The designations of sensitive and tolerant species are not intended to imply 

absolute ranges on tolerance among all possible insects. Instead, the dose-response assessments 

for this group simply indicate that some nontarget species, such as the Karner blue butterfly and 

cinnabar moth, appear to be as sensitive to B.t.k. as target species such as the gypsy moth and 

cabbage looper. The range of sensitivities among various insect species appears to follow a 

continuum, and it is possible that some species may be more or less sensitive to B.t.k. than those 

insects on which toxicity data are available. 

For aquatic organisms, toxicity data are expressed in several different units such as mg 

formulation/L, IU/L, and cfu/L. Based on application rates used in USDA programs and 

conservative assumptions concerning the depth of water over which B.t.k. might be sprayed, 

concentrations in water would be expected to be at or below 0.24 mg formulation/L. Toxicity 

values for fish are 1.4 mg formulation/L (an LOEC for sensitive species) and 1000 mg 

formulation/L (an NOEC for tolerant species). For aquatic invertebrates, the NOEC values for 

sensitive and tolerant species are 0.45 and 36 mg/L, respectively. 

4.3.2.3. LdNPV – Because no hazards can be identified for any species, a quantitative doseresponse assessment is not required. Consequently, no dose-response assessments were proposed 

by U.S. EPA and none were used in the previous gypsy moth risk assessment for Gypchek. In 

order to provide a quantitative comparison of the risks of using Gypchek relative to the other 

agents, dose-response assessments are made for both terrestrial mammals and aquatic species. 

For terrestrial mammals, the NOAEL of 2600 mg/kg bw is used. This is the same NOAEL that 

serves as the basis for the surrogate acute RfD for LdNPV in the human health risk assessment 

for this agent. For aquatic species, only NOEC values are available, and the highest NOEC of 

8x10 PIB/L is used to characterize risk.

4-14
 

9 
4.3.3. Chemical Agents 

4.3.3.1. Diflubenzuron and Tebufenozide – As summarized in Table 4-3, the dose-response 

assessments for diflubenzuron and tebufenozide are far more complete, in terms of the number of 

groups encompassed, than are the corresponding assessments for other agents considered in this 

risk assessment. This difference reflects both the nature of the available data and an assessment 

of the need to characterize risk quantitatively. Despite their specific modes of action in target 

species, diflubenzuron and tebufenozide induce toxicological responses in many different groups 

of animals. Furthermore, both chemicals are used in broadcast aerial applications, making 

exposure to many different groups of organisms likely. 

Both diflubenzuron and tebufenozide are relatively non-toxic to mammals and birds. As noted in 

the human health risk assessment, the acute and chronic toxicities of these two chemicals in 

mammals appear to be virtually identical in terms of NOAELs. This is also true for birds. The 

toxicity values used in the ecological risk assessment for mammals are identical to those used in 

the human health risk assessments: an acute NOAEL of 1118 mg/kg and a chronic NOAEL of 2 

mg/kg/day for diflubenzuron and an acute NOAEL of 1000 mg/kg and a chronic NOAEL of 1.8 

mg/kg/day for tebufenozide. The differences between the values for the chemicals are clearly 

insubstantial. For birds, the acute NOAEL for diflubenzuron is taken as 2500 mg/kg and the 

longer-term NOAEL is taken as 110 mg/kg/day. For tebufenozide, the values are again very 

similar: an acute NOAEL of 2150 mg/kg and a longer-term NOAEL of 15 mg/kg/day. For both 

chemicals, the longer-term NOAEL is taken from standard assays on reproduction. 

In terms of potential effects on terrestrial invertebrates, the data set for diflubenzuron is much 

richer than the data set for tebufenozide. Many laboratory toxicity studies and field studies have 

been conducted on diflubenzuron. Field studies are used in the dose-response assessment of 

diflubenzuron because the standard toxicity studies are extremely diverse and many are not 

directly applicable to a risk assessment. Despite the difficulty and uncertainty in interpreting 

some of the field studies, the relatively large number of field studies on diflubenzuron appears to 

present a reasonably coherent pattern that is at least qualitatively consistent with the available 

toxicity data and probably a more realistic basis on which to assess risk to nontarget species. The 

most sensitive species appear to be grasshoppers which may be adversely affected at an 

application rate of about 0.02 lb/acre [22 g/ha]. Somewhat high application rates—in the range 

of 0.027-0.031 lb/acre [30 to 35 g/ha]—will adversely affect macrolepidoptera and some 

beneficial parasitic wasps. At the maximum application rate considered in this risk assessment— 

0.062 lb/acre [70 g/ha]—some additional herbivorous insects are likely to be affected. No 

adverse effects in several other groups of insects are expected at this or much higher application 

rates. Honeybees are among the most tolerant species and are not likely to be adversely affected 

at application rates of up to 0.35 lb/acre [400 g/ha]. Invertebrates that do not synthesize chitin 

are also relatively tolerant to diflubenzuron. 

Although there are fewer and generally less detailed field studies on tebufenozide, compared with 

diflubenzuron, it appears to be less toxic to nontarget species (e.g., lacewing). In general, the 

field studies indicate that tolerant insect species are not affected by tebufenozide at application 

4-15
 
rates up to 0.24 lb/acre. The true NOEC may be higher – i.e., an LOEC has not been identified 

for tolerant species of terrestrial insects. Conversely, application rates as low as 0.03 lb/acre 

have been shown to have adverse effects on sensitive nontarget insects, primarily Lepidoptera. A 

NOEC for sensitive species was not identified. 

For both diflubenzuron and tebufenozide, the toxicity values for aquatic species follow a pattern 

similar to that for terrestrial species: arthropods appear to be much more sensitive than fish or 

non-arthropod invertebrates. Both compounds are about equally toxic to fish with virtually 

identical chronic NOEC values: 0.05 mg/L for diflubenzuron and 0.048 mg/L for tebufenozide. 

There are major and substantial differences regarding the toxicity of diflubenzuron and 

tebufenozide to aquatic invertebrates. Diflubenzuron is much more toxic. In acute toxicity 

studies, the NOEC for the most sensitive species is 0.0003 mg/L diflubenzuron, which is 400 

times less than the corresponding NOEC of 0.12 mg/L for tebufenozide. Chronic toxicity studies 

indicate a similar pattern. The NOEC for the most sensitive species is 0.00004 mg/L for 

diflubenzuron and 0.0035 mg/L for tebufenozide. The difference is a factor of about 90 [0.0035 

mg/L / 0.00004 mg/L]. Even though the number of available NOEC values is greater for 

diflubenzuron (seven acute and seven chronic), compared with tebufenozide (three acute and two 

chronic), and variability can be expected to increase as the number of species tested increases, it 

is unlikely that the apparent differences in toxicity are artifacts of sample size. For example, 

based on acute and chronic NOEC values in Daphnia, which are available for both compounds, 

diflubenzuron is more toxic than tebufenozide by a factor of about 2700 in acute studies and a 

factor of 725 in chronic studies. The toxicity to aquatic invertebrates is one of the few areas in 

which diflubenzuron and tebufenozide differ remarkably, and this difference has an impact on the 

risk characterization (Section 4.4). 

4.3.3.2. Disparlure – The limited amount of toxicity data on disparlure precludes making a 

standard dose-response assessment for terrestrial species. Disparlure is identical or similar to 

pheromones produced by other species of moths and is able to attract male nun moths. Since, 

however, there are no quantitative data available regarding the efficacy of disparlure in nontarget 

moths, a dose-response assessment for this effect in a nontarget species of moths cannot be made. 

For aquatic species, NOEC values and limited data on effect levels are available from acute 

exposure studies in rainbow trout and Daphnia. No LC50 values are available in fish. The doseresponse assessment is limited to NOEC values of 10 mg/L in trout and 300 mg/L in bluegills. 

The only information on toxic effects in fish consists of a report of 20% mortality in trout after 

acute exposure to disparlure at 100 mg/L. Thus, disparlure does not appear to be highly toxic to 

fish. Daphnia magna are much more sensitive with a 48-hour LC50 of 0.098 mg/L and an NOEC 

for mortality of 0.017 mg/L. Based on the LC50 value, disparlure is classified as highly toxic to 

aquatic invertebrates. 

4.3.3.3. DDVP – Given the limited nature of the use of DDVP in programs to control the gypsy 

moth and consequent limited number of exposure assessments, the dose-response assessment for 

DDVP is relatively simple. For terrestrial mammals, a value of 240 mg/kg from a study using 

4-16
 
DDVP in a PVC formulation is used for direct exposure to the DDVP-PVC strip—i.e., a raccoon 

tampering with a milk carton trap and consuming all or part of the DDVP strip. At the dose of 

240 mg/kg, no mortality or frank signs of AChE inhibition were observed. For the contaminated 

water scenario, the NOAEL of 0.5 mg/kg from a study involving exposure to free or 

unformulated DDVP is used. This NOAEL is from the study that forms the basis for the acute 

RfD used in the human health risk assessment. Although DDVP is classified as highly toxic to 

fish, the estimated levels of acute exposure for fish are far below the 30-day NOEC of 0.03 mg/L. 

Thus, this value is used for all fish and no attempt is made to consider differences in sensitivity 

among fish. A somewhat different approach is taken with aquatic invertebrates, some of which 

are more sensitive than fish to DDVP by a factor of more than 2500. Risks to sensitive species 

of aquatic invertebrates —i.e., daphnids and other small arthropods—are characterized based on 

the lowest reported LC50 value, 0.00007 mg/L from a 48-hour bioassay in Daphnia pulex. Some 

other groups of aquatic invertebrates, such as snails, appear to be much less sensitive than small 

arthropods. Risks to such tolerant species are based on a LC50 value of 21 mg/L in a freshwater 

snail. 

4-17
 
4.4. RISK CHARACTERIZATION 

4.4.1. Overview 

The comparative risk characterization for the ecological risk assessment is expressed similarly to 

that in the human health risk assessment. Numerically, the risk characterizations are given as 

hazard quotients (HQs), the level of exposure divided by some measure of effect, typically an 

NOAEL or NOEC. As in the human health risk assessment, the comparative risk 

characterization for ecological effects typically categorizes concern with the agents as marked 

(HQ>10), marginal (HQs between about 0.1 and 10), and minimal (HQ<0.1). One exception is 

made for B.t.k., which is classified as an agent of marked concern although the highest HQ is 9.4. 

An overview of the comparative risk characterization is summarized in Table 4-4 for terrestrial 

species and Table 4-5 for aquatic species. The risk characterizations are illustrated in Figure 4-1 

(terrestrial) and Figure 4-2 (aquatic). As in the human health risk assessment, the HQs for each 

agent are presented as a range. The upper end of the range is typically the highest hazard 

quotient associated with a plausible exposure scenario. The lower end of the range is not 

necessarily the lowest HQ calculated in each of the risk assessments. For some agents, the lower 

range is taken from sets of exposure scenarios that provide similar HQs for exposures that may 

be regarded as typical. For these agents, the lowest HQs reported in the individual risk 

assessments are close to zero. In some cases, the numerical expressions of risk do not adequately 

convey the potential for hazard. These cases are noted in Figures 4-1 and 4-2 with comments. 

Ecological risk assessments involve, at least implicitly, considerations of thousands of different 

species and the relationships among these species and their habitats. Invariably, however, data 

are available on only a small subset of these species and field studies provide only limited insight 

into the complex interrelationships and secondary effects among species. Thus, as in the human 

health risk assessments, ecological risk assessments cannot offer a guarantee of safety. They can 

and do offer a means to identify whether or not there is a basis for asserting that adverse effects 

are plausible and what the nature of these effects might be. 

Within these limitations, only LdNPV clearly qualifies as an agent of minimal concern. While 

there are limitations in the available studies on LdNPV, there is simply no basis for asserting that 

LdNPV will adversely affect any species except the gypsy moth. 

Agents of marked concern include the gypsy moth, B.t.k., and diflubenzuron. The types of 

concern with each of these agents, however, are quite different. For both the gypsy moth and 

B.t.k., the concerns are narrow. The gypsy moth clearly will damage some terrestrial vegetation. 

B.t.k. is likely to affect sensitive Lepidoptera. Concern with the use of diflubenzuron is broader 

and includes effects on both terrestrial and aquatic invertebrates. 

The designation of the gypsy moth as an agent of marked concern is obvious. The effects of 

gypsy moth larvae on forests are extremely well documented and well understood. In sensitive 

forest stands—i.e., stands in which oak, birch, and other favored species predominate —gypsy 

moth larvae can cause substantial defoliation and tree mortality. While some other lepidopteran 

4-18
 
species also may be directly affected by exposure to the gypsy moth, most of the other effects 

caused by the gypsy moth will be secondary. Reductions in populations of squirrels, mice, and 

other mammals which may be sensitive to changes in the availability of acorns are likely and 

have been well documented. Substantial secondary adverse effects on other groups of animals— 

i.e., birds, reptiles, and aquatic species—cannot be ruled out but have not been convincingly or 

consistently demonstrated. 

Diflubenzuron is also clearly an agent of marked concern. Exposures to diflubenzuron at 

application rates used in gypsy moth control programs will adversely affect both terrestrial and 

aquatic invertebrates that rely on chitin for their exoskeleton. This has been demonstrated in 

controlled toxicity studies as well as multiple field studies. 

The designation of B.t.k. as an agent of marked concern is somewhat judgmental. As noted in 

Table 4-4, the highest hazard quotient is 9.4. Based on this HQ and the classification scheme 

used generally, B.t.k. would be classified as an agent of marginal concern. However, recent 

studies convincingly demonstrate that adverse effects in nontarget Lepidoptera will occur in the 

applications of B.t.k. used to control the gypsy moth. Concern is heightened because some of the 

Lepidoptera that may be adversely affected include at least one endangered species. 

Tebufenozide, DDVP, and disparlure are all classified as agents of marginal concern. For 

tebufenozide, the numerical expressions of risk may be less relevant than a more qualitative 

assessment. The highest HQ is 4 and is associated with the consumption of contaminated 

vegetation by a large mammal after two applications of the compound at the highest labeled 

application rate. While this exposure would be considered undesirable, it is not clear that any 

frank signs of toxicity would be seen. Risks to nontarget Lepidoptera may be of greater concern 

but the available data are insufficient to quantify potential risk. Risks to other invertebrates, both 

terrestrial and aquatic, appear to be insubstantial. DDVP is of marginal concern in that highly 

localized effects may be expected: nontarget insects entering a milk carton trap or some aquatic 

invertebrates affected by the accidental contamination of a small body of water with a pest strip. 

In both cases, the effects would be relatively minor, in terms of the number of organisms 

affected. Marginal concern for disparlure is associated with the relatively high toxicity of this 

agent to Daphnia and is reenforced by the very scant data on the toxicity of an agent that may be 

applied to large areas in broadcast applications. 

4.4.2. Agents of Marked Concern 

4.4.2.1. Gypsy Moth – The best documented and most obvious effect of the gypsy moth will be 

on terrestrial vegetation, particularly forest stands in which sensitive species of trees 

predominate. In some respects, the risk characterization for terrestrial vegetation is essentially a 

restatement of the hazard identification. In other words, the effects of gypsy moth larvae on 

forests is extremely well documented and relatively well understood. In sensitive forest 

stands—i.e., stands in which oak, birch, and other favored species predominate —gypsy moth 

larvae can cause substantial defoliation. In forest stands in which tree species that are not 

favored by gypsy moth larvae predominate —e.g., hemlock, various types of pine, black locust 

4-19
 
and white ash—even relatively high exposures, as measured by egg mass density, may not result 

in substantial defoliation. The risk assessment for direct effects on forests should be at least 

qualitatively influenced by the current range of the gypsy moth, which has not yet extended to 

some forests in the southeast that may be among the most sensitive to gypsy moth exposure. 

Thus, unless measures to contain the gypsy moth are successful, the southeastern oak forests may 

suffer serious damage in future infestations. 

Some other lepidopteran species also may be directly affected by exposure to the gypsy moth. 

Most studies, however, suggest that substantial adverse effects in terrestrial insects are unlikely 

and effects in some insect species, including some other Lepidoptera, may be beneficial. 

Because the gypsy moth may substantially damage some forests in severe infestations or 

outbreaks, secondary effects in other species of wildlife are plausible. Reductions in populations 

of squirrels, mice, and other mammals which may be sensitive to changes in the availability of 

acorns are likely. Substantial adverse effects on other groups of animals—i.e., birds, reptiles, and 

aquatic species—cannot be ruled out but have not been convincingly demonstrated. 

4.4.2.2. B.t.k. – Terrestrial insects are the only organisms likely to be adversely affected by 

exposure to B.t.k. or its formulations. Separate dose-response curves can be generated for both 

sensitive and tolerant terrestrial insects. At the application rates used to control gypsy moth 

populations, mortality rates among sensitive terrestrial insects are likely to range from 

approximately 80% to 94% or more. All sensitive terrestrial insects are Lepidoptera and include 

some species of butterfly, like the endangered Karner blue and some swallowtail butterflies and 

promethea moths. 

The effects in sensitive species have been convincingly demonstrated in the study by Herms et 

al. (1997). In this study, the toxicity of Foray 48B was assayed in larvae of both the gypsy moth 

and the Karner blue butterfly, an endangered species of butterfly indigenous to the northern 

United States (Minnesota to New Hampshire). Bioassays in both species involved applications 

of Foray 48B to vegetation (wild lupine leaves for the Karner blue and white oak leaves for the 

gypsy moth) at treatment levels equivalent to either 30-37 BIU/ha per ha (low dose) or 90 BIU/ha 

(high dose). A negative control consisted of untreated vegetation. The insect larvae (either 1st or 

nd nd 2 instar for the Karner blue and 2 instar for the gypsy moth) were placed on the vegetation 7 to 

8 hours after treatment and allowed to feed for 7 days. Survival rates for Karner blue larvae 

were: 100% for controls, 27% at the 30-37 BIU/ha treatment rate, and 14% at the 90 BIU/ha 

treatment rate. Survival rates for gypsy moth larvae were: 80% for controls; 33% for low-dose 

treatment, and 5% for high-dose treatment. Based on a statistical analyses of these data, the 

gypsy moth and Karner blue appear to be equally sensitive to B.t.k. This study is supplemented 

by the series of bioassays conducted by Peacock et al. (1998) which suggest that various other 

lepidopteran species may be as sensitive as the gypsy moth to B.t.k.. 

For some Lepidoptera, sensitivity to B.t.k. is highly dependent on developmental stage. This is 

particularly evident for the cinnabar moth, where late instar larvae are very sensitive to B.t.k. and 

4-20
 
early instar larvae are very tolerant to B.t.k. Given the mode of action of B.t.k.—i.e., it must be 

ingested to be highly toxic to the organism— effects on even the most sensitive species will 

occur only if exposure coincides with a sensitive larval stage of development. In tolerant species, 

including non-lepidopteran insects and certain larval stages of some Lepidoptera, the anticipated 

mortality rates are much lower (on the order of less than 1% to about 4%). 

The risk characterization for terrestrial mammals is unambiguous: under foreseeable conditions 

of exposure, adverse effects are unlikely to be observed. Similarly, based on a very conservative 

exposure assessment for aquatic species, effects in fish and aquatic invertebrates appear to be 

unlikely. As discussed in the hazard identification, effects in birds, plants, soil microorganisms, 

or soil invertebrates other than insects are not of plausible concern. Thus, quantitative risk 

characterizations for these groups are not conducted. For oil-based formulations of B.t.k. (or any 

other pesticide), effects in some soil invertebrates are plausible. 

4.4.2.3. Diflubenzuron – While the data base supporting the risk assessment of diflubenzuron is 

large and somewhat complex, the risk characterization is relatively simple and unequivocal. 

Diflubenzuron is an effective and general insecticide. Application rates used to control the gypsy 

moth are likely to have effects on some nontarget terrestrial insects. Species at greatest risk 

include grasshoppers, various macrolepidoptera (including the gypsy moth), other herbivorous 

insects, and some beneficial predators to the gypsy moth. These species are at risk because of the 

mode of action of diflubenzuron (i.e., inhibition of chitin) and the behavior of the sensitive 

insects (the consumption of contaminated vegetation or predation on the gypsy moth). Some 

aquatic invertebrates may also be at risk but the risks appear to be less than risks to terrestrial 

insects. The risk characterization for aquatic invertebrates is highly dependant on site-specific 

conditions. If diflubenzuron is applied when drift or direct deposition in water is not controlled 

well or in areas where soil losses from runoff and sediment to water are likely to occur, certain 

aquatic invertebrates are at risk of acute adverse effects, and exposure could cause longer-term 

effects on more sensitive species. 

Direct effects of diflubenzuron on other groups of organisms—that is, mammals, birds, 

amphibians, fish, terrestrial and aquatic plants, microorganisms, and non-arthropod 

invertebrates—do not appear to be plausible. Nontarget species that consume the gypsy moth or 

other invertebrates adversely affected by diflubenzuron may be at risk of secondary effects (for 

example, a change in the availability of insect prey). There is no indication that 4-chloroaniline 

formed from the degradation of diflubenzuron will have an adverse effect on any terrestrial or 

aquatic species. 

4.4.3. Agents of Marginal Concern 

4.4.3.1. Tebufenozide – The use of tebufenozide to control the gypsy moth may result in 

adverse effects in nontarget Lepidoptera but these effects have not been well characterized or 

clearly demonstrated. There is little indication that other species will be impacted under normal 

conditions of use even at the highest application rate. Tebufenozide is an insecticide that is 

effective in controlling populations of lepidopteran pests. No data, however, are available on 

4-21
 
toxicity to nontarget Lepidoptera. For the risk assessment of this compound, the assumption is 

made that nontarget Lepidoptera may be as sensitive as target Lepidoptera to tebufenozide. 

Thus, adverse effects in nontarget Lepidoptera would be expected after applications that are 

effective for the control of lepidopteran pest species. 

There is no indication that short-term exposures to tebufenozide will cause direct adverse effects 

in any terrestrial vertebrates or non-lepidopteran invertebrates even at the upper range of 

plausible exposures or as a result of accidental exposures. Similarly, direct adverse effects from 

longer-term exposures in birds and mammals appear to be unlikely under most conditions. 

Effects on birds due to a decrease in available prey—i.e., terrestrial invertebrates—are considered 

plausible. In extreme cases, exposure levels in some large mammals might exceed the NOEC, 

but would remain below levels associated with frank signs of toxicity. This point is reflected in 

the HQ of 4 for a large mammal consuming contaminated vegetation after two applications of 

tebufenozide at the highest labeled rate. Under normal conditions of use, tebufenozide is not 

likely to cause adverse effects in aquatic species; however, in the case of a large accidental spill 

into a relatively small body of water, adverse effects might be expected in aquatic vertebrates, 

invertebrates, and plants. 

4.4.3.2. DDVP – As in the human health risk assessment of DDVP, typical exposures and 

consequent risks to nontarget species should be negligible. The containment of the DDVP within 

a slow-release PVC strip combined with the target specific nature of pheromone baited traps 

should reduce the risks of inadvertent effects on nontarget species. Other insects and arthropods 

that may inadvertently enter the trap will probably be killed by DDVP vapor. While such 

inadvertent contact may occur, it is not likely to have a substantial impact on the number of 

nontarget insects or arthropods. 

Because of the limited use of DDVP, a relatively small number of exposure scenarios —all of 

which might be considered accidental or incidental—are developed. For terrestrial mammals, 

contact with the pest strip could occur by an animal directly tampering with a trap or by an 

animal consuming water accidentally contaminated with a DDVP strip. Adverse effects would 

not be expected in either case. In the case of accidental contamination of a small body of water 

with a DDVP strip, concentrations of DDVP in the water would be below the level of concern for 

fish by factors ranging from about 50 to 500. Some aquatic invertebrates, however, might be 

affected. For the most sensitive species of aquatic invertebrates —i.e., small aquatic arthropods 

like daphnids—exposure levels could substantially exceed laboratory LC50 values by factors of 

up to about 8. Exposures to tolerant aquatic invertebrates—like snails—would be below a level 

of concern by a substantial margin—i.e., factors ranging from about 30,000 to 300,000. 

The exposure assessments that serve as the bases for these risk characterizations are highly 

dependent on specific conditions—i.e., how much DDVP was in the strip at the time that the 

contamination occurred and the size of the body of water that was contaminated. Because the 

hydrolysis of DDVP in water is rapid, the estimates of adverse effects in some aquatic 

4-22
 
invertebrates would probably apply only to a very limited area near the pest strip rather than to 

the larger area of the body of water that is contaminated. 

4.4.3.3. Disparlure – There is little data available on terrestrial and aquatic animals to allow for 

a quantitative characterization of risk. The lack of chronic toxicity data in any species adds 

uncertainty to any risk characterization. Thus, for both terrestrial and aquatic species, the 

potential for the development of toxicity from long-term exposure to disparlure cannot be ruled 

out. Concern with the lack of toxicity data on disparlure is exacerbated by the fact that this 

compound may be applied to large areas in broadcast applications. 

Nonetheless, based on the available data, clear hazards to nontarget species have not been 

identified. Disparlure may disrupt mating in some moths other than the gypsy moth. The two 

species that are known to be affected, however, are both forest pests like the gypsy moth and only 

one of these other species is native to North America. For aquatic species, hazard quotients for 

both rainbow trout and Daphnia are below one, although the hazard quotient of 0.4 for Daphnia 

approaches one. Thus, while 0.4 is below the level of concern of one, there is uncertainty in the 

risk characterization because of the limited acute toxicity data, the lack of chronic toxicity data, 

and the high likelihood that many species will be exposed to this compound. 

4.4.4. Agent of Minimal Concern: Gypchek 

Unlike all of the other agents considered in this risk assessment, there is no basis for asserting 

that the use of Gypchek to control or eradicate gypsy moth populations is likely to cause any 

adverse effects in any species other than the gypsy moth. While no pesticide is tested in all 

species under all exposure conditions, the data base on LdNPV and related viruses is reasonably 

complete and LdNPV has been tested adequately for pathogenicity in a relatively large number of 

species, particularly terrestrial invertebrates. LdNPV appears to be pathogenic and toxic to the 

gypsy moth and only to the gypsy moth. 

For Gypchek, quantitative expressions of risk are in some respects more difficult because clear 

LOEC values are not defined—i.e., if an agent is not shown to cause an effect, the threshold for 

effects is not a meaningful concept. Nonetheless, general but very conservative exposure 

assessments demonstrate that plausible upper ranges of exposures are clearly below any level of 

concern by a factor of 1000 for terrestrial species and 30,000 for aquatic species. 

4-23
 
5. REFERENCES
 

ATSDR (Agency for Toxic Substances and Disease Registry). 2004. Guidance Manual for the 

Assessment of Joint Toxic Action of Chemical Mixtures. Available from U.S. Department of 

Health and Human Services, Public Health Service, ATSDR, Division of Toxicology. 

Cook GJ. 1994. Bacillus thuringiensis kurstaki exposure in ground-spray workers. Major Paper 

Submitted in Partial Fulfillment of MHSc Degree, (Community Medicine) in the Department of 

Health Care and Epidemiology, Faculty of Medicine, The University of British Columbia. 

David R. 1990. Acute pulmonary toxicity/pathogenicity study of Dipel technical material. 

(Bacillus thuringiensis var. kurstaki) in rats: Lab Project No. G-7239. 001. Unpublished study 

prepared by Microbiological Associates Inc. 66 p. MRID 41443402. 

Herms CP; Mccullough DG; Baue LS; Haack RA; Miller DL; Dubois NR. 1997. Susceptibility 

of the endangered Karner blue butterfly (Lepidoptera: lycaenidae) to Bacillus thuringiensis var. 

kurstaki used for gypsy moth suppression in Michigan. Great Lakes Entomol. 30:125-141 

Hernandez E; Ramisse F; Cruel T; Vagueresse R; Cavallo JD. 1999. Bacillus thuringiensis 

serotype H34 isolated from human and insecticidal strains serotypes 3a3b and H14 can lead to 

death of immunocompetent mice after pulmonary infection. FEMS Immunol Med Microbiol. 

24:43-47. 

Hernandez E; Ramisse F; Gros P; Cavallo JD. 2000. Super-infection by Bacillus thuringiensis 

H34 or 3a3b can lead to death in mice infected with influenza A virus. FEMS Immunology Med 

Microbiol. 29:177-181. 

Liebhold AM; McManus M. 1999. The evolving use of insecticides in gypsy moth 

management. J For. 97(3): 20-23. 

Peacock JW; Schweitzer DF; Carter JL; Dubois NR. 1998. Laboratory assessment of the effects 

of Bacillus thuringiensis on native lepidoptera. Environ Entomol. 27: 450-457. 

Petrie K; Thomas M; Broadbent E. 2003. Symptom complaints following aerial spraying with 

biological insecticide Foray 48B. New Zealand Medical Journal. 116: 1170-1177. 

SERA 2004a. Control/Eradication Agents for the Gypsy Moth – Human Health and Ecological 

Risk Assessment for Bacillus thuringiensis var. kurstaki (B.t.k.) FINAL REPORT. SERA TR 

04-43-05-02c dated June 8, 2004. 

SERA 2004b. Control/Eradication Agents for the Gypsy Moth – Human Health and Ecological 

Risk Assessment for Gypchek – a Nuclear Polyhedrosis Virus (NPV) FINAL REPORT. SERA 

TR 04-43-05-02b dated June 16, 2004. 

5-1
 
SERA 2004c. Control/Eradication Agents for the Gypsy Moth – Human Health and Ecological 

Risk Assessment for Diflubenzuron (Dimilin) FINAL REPORT. SERA TR 04-43-05-03b dated 

July 30, 2004. 

SERA 2004d. Control/Eradication Agents for the Gypsy Moth – Human Health and Ecological 

Risk Assessment for Disparlure (a.i.) FINAL REPORT. SERA TR 04-43-05-04b dated August 

27, 2004. 

SERA 2004e. Control/Eradication Agents for the Gypsy Moth – Human Health and Ecological 

Risk Assessment for DDVP FINAL REPORT. SERA TR 05-43-05-5b dated August 20, 2004. 

SERA 2004f. Control/Eradication Agents for the Gypsy Moth – Human Health and Ecological 

Risk Assessment for Tebufenozide (Mimic) FINAL REPORT. SERA TR 04-43-05-06c dated 

August 8, 2004. 

SERA 2004g. Control/Eradication Agents for the Gypsy Moth – Human Health and Ecological 

Risk Assessment for the Gypsy Moth FINAL REPORT. SERA TR 04-43-05-07c dated August 

23, 2004. 

Stanton HC; Albert JR; Mersman HJ. 1979. Studies on the pharmacology and safety of 

dichlorvos in pigs and pregnant sows. Am J Vet Res. 40:315-320. 

USDA (U.S. Department of Agriculture). 1995. Gypsy Moth Management in the United States: 

A Cooperative Approach. Final Environmental Impact Statement. Appendix F (Human Health 

Risk Assessment) and Appendix G (Ecological Risk Assessment). 

U.S. EPA/OPP (U.S. Environmental Protection Agency/Office of Pesticide Programs). 1998. 

Reregistration Eligibility Decision (RED): Bacillus thuringiensis. Available at: 

http://cfpub.epa.gov/oppref/rereg/status.cfm?show=rereg 

U.S. EPA/OPP (U.S. Environmental Protection Agency/Office of Pesticide Programs). 2004. 

Lepidopteran Pheromones Fact Sheet. Prepared by the Office of Pesticide Programs. Available 

at: http://www.epa.gov/pesticides/biopesticides/ingredients/factsheets/ 

factsheet_lep_pheromones.htm. 

5-2
 
Tables 

Table 2-1: Total use of control agents by numbers of acres treated between 1995 and 2003 

Table 3-1: Comparative hazard identification for potential effects in humans 

Table 3-2: Comparative exposure assessment for human health effects 

Table 3-3: Comparative dose-response assessment for human health effects 

Table 3-4: Comparative risk characterization for human health effects 

Table 4-1: Comparative hazard identification for potential effects in nontarget species 

Table 4-2: Comparative exposure assessment for ecological effects 

Table 4-3: Comparative dose-response assessment for ecological effects 

Table 4-4: Comparative risk characterization for terrestrial species 

Table 4-5: Comparative risk characterization for aquatic species 
Table 2-1: Total use of control agents by numbers of acres treated between 1995 and 2003 

Suppression Eradication Slow-the-Spread Grand Total 

(Total acres) (Total acres) (Total acres) (Acres) 

B.t.k. 1,484,486 1,057,201 367,722 2,909,409 

NPV 36,518 7,376 9,140 53,034 

Diflubenzuron 657,671 6 6,883 664,560 

Tebufenozide 0 0 0 0 

Disparlure flakes 0 60,090 1,567,199 1,627,289 

Mass Traps * 0 1,912 0 1,912 

*

 Mass traps contain DDVP in a PVC strip and disparlure as an attractant. 

Tables - 1 
Table 3-1: Comparative hazard identification for potential effects in humans 

Gypsy 

Moth 

Agents used in Gypsy Moth Program 

Endpoint B.t.k. LdNPV DFB Tebufen 

-ozide 

DDVP 

Lethality " G " G G !a 

Disparlure 

"

Sub-lethal effects 

Irritation ! ! � G " � G 

Blood ! ! "b 

Carcinogenicity � c " G 

Neurotoxicity " " G " ! 

Immunotoxicity " " " " G 

Reproduction " � �

Endocrine Effects " G 

Pathogenicity G d " N/A N/A N/A N/A 

a

 Risks are mitigated by formulation in PVC.
 

b

 Excluding inhibition of plasma and RBC AChE.
 

c

 An environmental metabolite, 4-chloroanaline, poses a carcinogenic risk.
 

d B.t.k. itself does not appear to be pathogenic. Possible enhancement of influenza virus.
 

Key: ! Effect/risk demonstrated in humans 

� Effect is plausible 

G Marginal evidence for potential effect 

" No plausible basis for risk 

Blank No data are available 

Tables - 2 
Table 3-2: Comparative exposure assessment for human health effects 

Agent Measure of 

Exposure 

Plausibility 

of Exposure 

Comments 

Gypsy moth Eggs masses 

per acre 

Variable Exposure potential is high during outbreaks and decreases 

as intensity of infestation decreases. 

B.t.k. Application 

rate and 

cfu/m3

 x hour 

High During broadcast applications, exposure potential is high 

and can be reasonably well characterized. 

LdNPV mass of 

formulation 

High During broadcast applications, exposure potential is high 

and can be reasonably well characterized. 

Diflubenzuron mass of 

chemical 

High Can persist on vegetation and water contamination is 

plausible. 

Tebufenozide mass of 

chemical 

High Can persist on vegetation and water contamination is 

plausible. 

DDVP mass of 

chemical 

Very low Except in cases of intentional or incidental tampering with 

a trap, exposures will be very low. 

Disparlure mass of 

chemical 

Variable Very little compound is used in traps and exposures are 

likely to be very low. 

Tables - 3 
Table 3-3: Comparative dose-response assessment for human health effects 

Agent	 Toxicity Endpoint Quality Comment 

Value 

Gypsy Moth Acute Irritation ! Based on human data with a clear dose-response 

relationship. 

B.t.k. Acute Irritation # Based on human data but no dose-response 

relationship is apparent. 

Toxicity G Based on a single study in mice using a 

marginally relevant route of exposure. 

LdNPV Acute None ! High confidence because no endpoint of concern 

can be identified. 

Diflubenzuron Acute Blood # No EPA acute RfD. Conservative approach 

based on petroleum formulation. 

Chronic Blood ! Agency-wide EPA RfD adopted by OPP. 

4-Chloroaniline* Acute Blood # No EPA acute RfD. Conservative approach 

based on 90-day study. 

Chronic Blood G EPA chronic RfD. Confidence classified as low 

by EPA. 

Cancer Potency Cancer # EPA cancer potency factor 

Tebufenozide Acute Repro # No EPA acute RfD. Based on reproduction 

studies in two species 

Chronic Blood # EPA/OPP chronic RFD. 

DDVP Acute Neuro !/G 

For DDVP itself, value is based on an EPA 

acute RfD. For DDVP in PVC strip, the value is 

based on marginal data. 

Disparlure Acute N/A G No acute RfD can be derived. 

Key for quality of Toxicity Values: ! High 

# Medium 

G Low 

*

 An environmental metabolite of diflubenzuron. 

Tables - 4 
Table 3-4: Comparative risk characterization for human health effects a 

Hazard
 

Quotient (HQ) b


Agent Comments 

Lower Upper 

Gypsy Moth 1.6 625	 Irritant effects (dermal, ocular, and/or respiratory) are well 

documented. Lower range is based on sparse infestations, where 

effects might be seen in about 1% of the population. Upper range is 

based on major outbreaks where responses might be seen in about 

40% of the population. 

B.t.k.	 0 0.04 HQs are for serious adverse effects, which are highly unlikely to 

occur. Irritant effects could be reported in about 20% of exposed 

individuals – both workers and members of the general public. 

LdNPV 0 0.02	 No risks are plausible. Upper range of HQ is calculated from a 

free-standing NOAEL. 

Diflubenzuron 

Workers 0.05 0.5	 The upper range is associated with the upper range of plausible 

exposures in ground spray applications. Under typical conditions, the 

HQ will be about 0.05. 

Public 0.09 0.1	 This narrow range of HQs reflects the higher HQ for any longer term 

exposure (0.09) and the highest HQ for acute exposures (0.1). Most 

other HQs are below 0.01. 

4-Chloroaniline as an environmental metabolite of diflubenzuron 

Toxicity 0.02 0.4	 Lower value is based on acute consumption of contaminated water 

(peak concentration) by child. Upper range based on acute 

consumption of contaminated fish by subsistence populations after 

accidental spill. Other HQs are insubstantial. 

Cancer 0.09 0.4	 HQs based on cancer risk of 1 in 1 million. Both lower and upper are 

based on consumption of contaminated water (central and upper 

ranges). Other scenarios lead to much lower risks. 

Tebufenozide 0.03 1.5	 Lower range is based on the central estimate of contaminated fruit 

(longer-term) after 2 applications. Highest HQ is for the upper range 

of longer-term consumption of contaminated fruit following 2 

applications at the highest application rate. Other HQs are much less 

than 0.03. 

DDVP 0 380	 Lower range of risk is essentially zero because exposures are unlikely. 

Upper range is based on oral exposure from a child tampering with 

the strip. Likelihood of clinically significant effects seems remote. 

Disparlure 0 0	 No potential risk can be identified. 

a

 See Figure 3-1 for illustration.


b Hazard quotients less than 0.01 are given as zero. For B.t.k., the lower range of the HQ is 0.000036. For NPV
 

and disparlure, risks are essentially zero. For DDVP, exposure is unlikely and the risk is also essentially zero
 

except for accidental exposures.
 

Tables - 5 
Table 4-1: Comparative hazard identification for potential effects in nontarget species 

Agents used in Gypsy Moth Program 

Gypsy 

Endpoint Moth B.t.k. LdNPV DFB Tebufen DDVP Dispar­

-ozide lure 

Terrestrial species 

Mammals � " " ! ! ! " 

Birds � " " G � ! " 

Nontarget Lepidoptera � ! " ! ! ! �* 

Other arthropods G " " ! ! ! " 

Other invertebrates G G " " " ! " 

Plants ! " " " " " 

Microorganisms � " G 

Aquatic species 

Fish G " " G G ! " 

Invertebrates G � " ! ! ! �

Plants G " " G �

Microorganisms � " G

 Effects in other pest Lepidoptera pest species only. * 

Key: ! Direct effects demonstrated in species of concern 

�

G 

Effects are plausible 

Marginal evidence for effect 

" No plausible basis for risk 

Blank No data are available 

Tables - 6 
Table 4-2: Comparative exposure assessment for ecological effects 

Agent Plausibility Primary Comments 

of Exposure Route 

Gypsy moth Variable N/A Exposure potential is high during outbreaks and 

decreases as intensity of infestation decreases. 

B.t.k. High Oral During broadcast applications, exposure potential is 

high. 

LdNPV High Oral During broadcast applications, exposure potential is 

high and can be reasonably well characterized. 

Diflubenzuron High Oral Can persist on vegetation and water contamination 

is plausible. 

Tebufenozide High Oral Can persist on vegetation and water contamination 

is plausible. 

DDVP Very low Inhalation 

/Oral 

Except in cases of insects entering the trap or other 

animals tampering with trap, exposures will be very 

low. 

Disparlure Variable Variable Very little compound is used in traps and exposures 

are likely to be very low. 

Tables - 7 
Table 4-3: Comparative dose-response assessment for potential effects in nontarget species 

Agents used in Gypsy Moth Program 

Gypsy 

Endpoint Moth B.t.k. Ld- DFB 4-CA Tebufen DDVP DisparNPV -ozide lure 

Terrestrial species 

Mammals �/G a G � � � �

Birds G � � �

Nontarget Lepidoptera � ! !	 " 

Other arthropods !	 G 

Other invertebrates G	 " 

Plants ! 

Microorganisms �

Aquatic species 

Fish G/"b G "/�c G � � �

Invertebrates G G !	 G � " �

Plants G G G 

Microorganisms G " a

 NOEC value only for oral exposure. NOEC and LOEC for inhalation. 

b NOEC value only for tolerant species. LOEC only for sensitive species. 

c

 Effect level only for acute exposures. 

Descriptive Key: !	 Effect and no-effect levels clearly identified. Response or differences in 

sensitivities among species can be quantified. 

� Effect and no-effect levels identified. 

G Based on no-effect level only. 

" Based on effect level only. 

Blank No quantitative dose-response assessment is made. 

Tables - 8 
Table 4-4: Comparative risk characterization for terrestrial species a 

Hazard 

Agent Quotient (HQ) b 

Comments 

Lower Upper 

Gypsy Moth 0.25 400 All HQs based on defoliation. Lower HQ based on low infestation 

(5 egg masses/acre) in intermediate stands. Upper HQ based on 

damage to sensitive stands in an outbreak (up to 83% defoliation). 

Effects secondary to defoliation will occur in some animal 

populations. 

B.t.k. 0.36 9.4 All HQs based on lethality to terrestrial invertebrates using 10% as a 

benchmark. A maximum mortality of 3.6% for tolerant invertebrates 

and 94% for sensitive invertebrates 

LdNPV 0 0 No toxicity to terrestrial species is likely. The upper range of the 

HQ is 0.001 and is based on the consumption of contaminated 

vegetation and an acute free-standing NOAEL in mammals. 

Diflubenzuron 0.18 32 All HQs based on responses in terrestrial invertebrates. The lower 

range is based on tolerant species and the upper range on sensitive 

species. 

4-Chloroaniline 0 0.02 The upper range based on the consumption of fish by a predatory 

bird after an accidental spill (acute scenario). 

Tebufenozide 0 4 The upper range is based on the consumption of contaminated 

vegetation by a large mammal after 2 applications at the maximum 

application rate. While not quantified, effects on some nontarget 

Lepidoptera are possible. 

DDVP 0 0 Typically, exposures will be minimal. Insects entering the traps are 

likely to be killed. 

Disparlure 0 0 No potential hazard can be identified except possible mating 

disruption in other pest Lepidoptera.

a See Figure 4-1 for illustration. Note that the magnitude of the HQ among different agents is not a measure of 

relative risk or severity of effects. See text for discussion.

b Hazard quotients less than 0.01 are given as zero. For tebufenozide, the lower range of the HQ is 0.0002. For 

4-chloroaniline the lower range of the HQ is 0.00002. For NPV and disparlure, lower range of the HQs are 

essentially zero. For DDVP, exposure is unlikely and the risk is also essentially zero except for accidental 

exposures. 

Tables - 9 
Table 4-5: Comparative risk characterization for aquatic species a 

Hazard 

Agent Quotient (HQ) b 

Comments 

Lower Upper 

Gypsy Moth 0 0 No basis for asserting that adverse effects will be observed. 

B.t.k. 0 0.5 All HQs based on aquatic invertebrates. Lower range is 0.007 for 

tolerant species. The upper range is based on sensitive species 

LdNPV 0 0 No basis for asserting that adverse effects will be observed. The 

upper range is 0.00003 and is based on a free-standing NOEC. 

Diflubenzuron 0 5 Upper range is based on acute effects in sensitive aquatic 

invertebrates (Daphnia) after peak exposures. 

4-Chloroaniline 0 0.2 Upper range is based on acute exposures to aquatic invertebrates and 

aquatic plants. 

Tebufenozide 0 0.4 Upper range is based on longer-term toxicity in sensitive aquatic 

invertebrates. 

DDVP 0 0 8 No risks are plausible in normal use. The HQ for aquatic 

invertebrates could reach up to 8 in accidental exposures. 

Disparlure 0 0.4 Upper range based on acute exposures to sensitive aquatic 

invertebrates (Daphnia).

a See Figure 4-1 for illustration. Note that the magnitude of the HQ among different agents is not a measure of
 

relative risk or severity of effects. See text for discussion.
 

b

 Hazard quotients less than 0.01 are given as zero.
 

Tables - 10 
Figures 

Figure 3-1: Risk comparison for potential human health effects 

Figure 4-1: Risk comparison for potential effects in terrestrial species 

Figure 4-2: Risk comparison for potential effects in aquatic species 
Figure 3-1: Risk comparison for potential human health effects. 

Figures - 1
 
Figure 4-1: Risk comparison for potential effects in terrestrial species. 

Figures - 2
 
Figure 4-2: Risk comparison for potential effects in aquatic species. 

Figures - 3
 
 

 

 

 

 

 

 

 

 

 

 

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