PB83-263681
Toxicity, Bioconcentration, and
Metabolism of Five Herbicides in Freshwater Fish
Wisconsin Univ.-Superior
Prepared for
Environmental Research Lab.-Duluth, MN
Sep 83
U.S. DEPARTMENT OF COMMERCE
National Technical Information Service
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.EPA-600/5-83-096
September 1983
TOXICITY, BIOCONCENTRATICN, AND METABOLISM
OF FIVE HERBICIDES IN FRESHWATER FISH
by
Daniel J. Call, Larry T. Brooke
and Raymond J. Kent
Center for Lake Superior Environmental Studies
University of Wisconsin-Superior, Superior, WI 54880
U.S. EPA Grant No. R-806196-01
Project Officer
John I. Teasley
Environmental Research Laboratory-Duluth
Office of Research and Development
U.S. Environmental Protection Agency
Duluth, Minnesota 55804
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TECHNICAL REPORT DATA
(Please read Irtsmicrions on the reverse before completing)
1. REPORT NO.
EPA-600/5-83-096
3. RECIPIENT'S ACCESSION NO.
26368 1
4. TITLE AND SUBTITLE
Toxicity, Bioconcentration, and-Metabolism of Five
Herbicides in Freshwater Fish
5. REPORT DATE
September 1985
6. PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
D.J. Call, L.T. Brooke, and R.J. Kent
8. PERFORMING ORGANIZATION REPORT NO.
9. PERFORMING ORGANIZATION NAME AND ADDRESS
Center for Lake Superior Environmental Studies
University of Wisconsin-Superior
Superior, Wisconsin 54880
10. PROGRAM ELEMENT NO.
11. CONTRACT/GRANT NO.
806196, 80020010, 806864
12. SPONSORING AGENCY NAME AND ADDRESS
U.S. Environmental Protection Agency
Environmental Research Laboratory-Duluth.
6201 Congdon Boulevard
Duluth, Minnesota 55804
13. TYPE OF REPORT AND PERIOD COVERED
14. SPONSORING AGENCY CODE
EPA/600/03
IS. SUPPLEMENTARY NOTES
16. ABSTRACT lexicological studies were conducted In two areas: (1; the TOK I city, bloconcentratlon
potential and n»et»bolls» of five harbtcldas In fish; and (2) th« toKlclty and/or metabolism of priority
pollutants and related chemicals In various aquatic organisms.
The tast herbicides Included alachlor I2-chloro-2' ,6'-d lathy l-H-(m»thc«ymethy I) acatanl I ld»l , bronacll
(5-bromo-3-£ac-buty1-6-n»thyl uracl I), dlnoseb l2-(sac-butyl)-4,6-dlnltrophanol I, dluron ( 3-(3,4-dlchloro-
phenyl )-1,1-dlmethylureal, and propanll (3,4-dlchloroproplonanl I Ide) . Ac'jte toxlclty (through 192 hr) ,
sarly llfe-stsga tovtclty (58-64 Jay), and bloconcentratlon stud I as oera conducted «lth fathaad ulnnovs
(Plmephalas prome'as) In Laka Superior vater. Harblclda matabolIsm «as Investigated In rainbow trout (Salno
galrdnarl) both In vivo and In vitro.
Twenty-two chemicals from th» EPA priority pollutant list were studied for their acute and/or chronic
toxlclty to selected freshwater organisms. Thesa Included 1,2-dlchloroethane, 1,1,2-tr Icfiloroethane,
1,1,2,2-tatrachloroethane, tietrachloroethylene, 1,2-dlchlorobeniana, 1 ,3-dlch!orobeni9ne, I ,4-dichloro-
benzane, hexachlorobenzene, hexachlorobutadf ane, dl-n-butylphthal ats, pentachlorophenol, heptachlor,
chlordane, toxaphane, arsenic , chromium , lead* /mercury- , nickel , silver* ,
selantuffl , and cyanide. Freshwater species tested Included the fathead minnow, rainbow trout, bluegill
sun fish (Lepomls macrochlrus), flagflsh (Jordanella florldae), Oaphnla rnagna, scud (Gammarus
pseudal Imnaeus), itldge (Tanytarsus d tsslml I Is) and green alga (Selanastrum capr Icornutum) . Toxlclty tests
were also conducted with pentachloroethane, hexachloroethane, 1,2,4-tr IchloroCeniane, pentachIorubenjone,
methanol and dimethyl formamlde. The uptake by fish of dl-n-butylphthalate from
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NOTICE
This document has been reviewed in accordance with
U.S. Environmental Protection Agency policy and
approved for publication. Mention of trade names
or commercial products does not constitute endorse-
ment or recommendation for use.
11
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FOREWORD
Pesticides are applied to land and water In high tonnage volumes each
year. These chemicals must be environmentally safe as well as effective. The
United States Environmental Protection Agency researches and reviews pesticides
that are currently registered and those which are under registration considera-
tion for their environmental safety.
This report contains the results of toxicity, bi concentration, and
metabolism studies with five registered herbicides and fish. This information
will assist the Office of Pesticide Programs in its evaluation of these
herbicides for their possible impacts upon the environment.
Norbert Jaworski, Ph. D.
Director
Environmental Research Laboratory
Duluth, MN
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ABSTRACT
Acute and early life-stage toxicities were determined for the herbicides
alachlor, bromacil, dinoseb, diuron, and propanil with the fathead minnow
(Pimephales promelas). Uptake, bioconcentration potential, and elimination of
C-labeled herbicides were studied in the same species. In vivo metabolism
of radiolabeled herbicides was determined with rainbow trout (Salmo gairdneri)
and fathead minnows as test organisms. In vitro metabolism was studied with
rainbow trout liver homogenates.
LCj-Q values at 96 hr in Lake Superior water for 30-day old fathead
minnows were: alachlor (5.0 mg-L" ), bromacil (182 mg-L" ), dinoseb (0.7
mg-L" ), diuron (14.2 mg-L ), and propanil (8.6 mg-L ). Mortality readings
were made at intervals up to 192 hr to obtain mortality curves for each
herbicide.
The "no-effect" estimate for a 60-day post-hatch exposure of fathead
minnows to alachlor was between 0.52 and 1.10 rng-L , based upon fish weight
and length. A "no-effect" concentration range for a 60-day exposure to broma-
cil could not be determined from the data, but was less than 1.0 mg-L . The
60-day "no-effect" estimate for dinoseb was between 14.5 and 48.5 jig-L~ ,
based upon the two most sensitive parameters, fish survival and weight. The
60-day "no-effect" estimate for diuron was between 33.4 and 78.0 yg-L~ based
upon two significantly affected parameters, abnormal fry development and fish
survival through 60 days. A 54-day "no-effect" estimate for propanil was
between 0.4 and 0.6 pg-L based upon fish length and dry weight as the two
iv
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most sensitive parameters.
14
Fathead minnow mean bioconcentration factors for C parent herbicide
equivalents were: alachlor (45.8 x), bromacil (3.2 x), dinoseb (62.8 x),
diuron (150.4 x), and propanil (90-2 x). Percent parent compound at the end
of the uptake phase for alachlor, dinoseb, diuron, and propanil were 13.2,
2.3, 1.3, and 1.8%, respectively. This was not determined for bromacil.
Parent herbicide bioconcentration factors were: alachlor (.6.0 x), dinoseb
(.1.4 x}, diuron (2-0 x), and propanil 0-6 x).
14
Rainbow trout injected with C-labeled herbicides readily metabolized
all five compounds. From 75 to 98% of the injected doses were excreted with-
in 24 hrs. Metabolites of the herbicides accumulated in the bile to varying
degrees - from 1% (bromacil) to 22% (propanil) of the injected dose. GC/MS
characterization of extracts from diuron exposed trout revealed the presence
of 3,4-dichloroaniline and demethylated products. A propanil metabolite was
identified as either 3',4'-dichloro-2-hydroxypropionanilide or 3',4'-dichloro-
3-hydroxypropionanilide.
In vitro binding of herbicides to rainbow trout liver macromolecules
occurred for all five compounds. Alachlor, bromacil, dinoseb, diuron, and
propanil had protein binding values of 75.2, 3.6, 13.2, 6.7, .and 2.3 nanomoles
herbicide-gm" of protein, respectively. The binding was not dependent on NADPH,
was not altered by the addition of glutathione, and was not altered by the
substitution of heat-inactivated microsomes for active microsomes. The binding
was apparently due to the affinity of the parent herbicides for protein and
other macromolecules, and was not dependent on mixed-function oxidase activity.
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CONTENTS
Foreword iii
Abstract iv
Figures viii
Tables x
Acknowledgements xii
1. Introduction 1
2. Conclusions 4
3. Recommendations 6
4. Materials and Methods 7
Test Organisms and Chemicals 7
Toxicity Tests 10
Biconcentration Tests 12
Measurements of Herbicide Concentrations 13
Metabolite Characterization 22
Herbicide Binding 26
Herbicide Stability Studies 27
Gas Chromatographic/Mass Spectrometric Analysis .... 28
5- Results . 29
Acute Toxicity Tests 29
Early Life-Stage Toxicity Tests 31
Bioconcentration Tests 43
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Metabolite Characterization 48
Binding of Herbicides to Macromolecules ... 52
Herbicide Stability in Solution 53
Tetrachloroazobenzene and Tetrachloroazoxybenzene
Analysis 54
6. Discussion
Acute Toxicity 55
Early Life-Stage Toxicity 60
Uptake, Metabolism and Elimination 69
Herbicide Stability 72
Tetrachloroazobenzene and Tetrachloroazoxybenzene ... 74
References 77
Appendices
A. Chemical characteristics of the toxicity test water .... 85
B. Mortality curves for acute flow-through toxicity tests ... 88
C. Water concentrations and whole fish tissue concentrations
of total '4C from herbicide bioconcentration studies
with the fathead minnow CPimephales promelas) 93
D. Mass chromatograms and reconstructed ion chromatograms
for standards of 3,3',4,4'-tetrachloroazobenzene
CTCAB), 3,3',4,4'-tetrachloroazoxybenzene (TCAOB),
p_-diiodobenzene CDIB), and technical grades of
propanil and diuron 97
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FIGURES
Number Page
1 Survival curves for fry and juvenile fathead minnows
CPimephales promelas) exposed to propanil in an early
life-stage toxicity test 42
B-l LC5Q concentrations for duplicate alachlor toxicity tests
with 30-day old fathead minnows (Pimephales promelas) . . 88
3-2 LCgQ concentrations for duplicate bromacil toxicity tests
with 30-day old fathead minnows CPimephales promelas) . . 89
B-3 LCgQ concentrations for duplicate dinosefa toxicity tests
with 30-day old fathead minnows CPimephales promelas) . . 90
B-4 LCsg concentrations for duplicate diuron toxicity tests
with 30-day old fathead minnows CPimephales prgmelas) . - 91
8-5 LC5Q concentrations for duplicate propanil toxicity tests
with 30-day old fathead minnows (.Pimephales promelas) . . 92
14
C-l Log mean exposure water concentrations of C-labeled
alachlor (ng-mL"') and log mean C± S.D.) whole fish
total '4C residues (ng*9 1 during uptake and
depuration phases 93
14
C-2 Log mean exposure water concentrations of C-labeled
dinoseb (ng-mL"') and log mean (± S.D.) whole fish
tissue '^C residues (ng-g"1) during uptake and
depuration phases 94
14
C-3 Log mean exposure water concentrations of C-labeled
diuron (ng-mL"1) and log mean (± S.D.) whole fish
total 14C residues (ng-g~') during uptake and
depuration phases ... 95
14
C-4 Log mean exposure water concentrations of C-labeled
propanil (ng-mL"') and log mean (± S.D.) whole fish
total '4C residues (ng-g ) during uptake and
depuration phases 96
vm
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Number Page
D-l Mass chromatograms and reconstructed ion chromatogram
(RIC) for standards of 3,3',4,4'-tetrachloroazobenzene
(TCAB), 3,3',4,4'-tetrachloroazoxybenzene (TCAOB), and
p_-diiodofaenzene (DIB) 97
D-2 Mass chromatograms and reconstructed ion chromatogram
(RIG) of technical grade propanil monitored for
3,3',4,4'-tetrachloroazobenzene (TCAB) and 3,3',4,4'-
tetrachloroazoxybenzene CJCAOB) 98
D-3 Mass chromatograms and reconstructed ion chromatogram
(RIC) of technical grade diuron monitored for 3,3',4,4'-
tetrachloroazobenzene (TCAB) and 3,3',4,4'-
tetrachloroazoxybenzene (TCAOB) 99
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TABLES
Number Page
1 LCcA Values and 952 Confidence Intervals for Acute
Toxicity Tests with Replicates Combined 30
2 Hatchability, Abnormal Development, Survival and Growth
of Fathead Minnows (Pimephales promelas) Exposed to
Alachlor for 64 Days 32
3 Hatchability, Abnormal Development, Survival and Growth
of Fathead Minnows (.Pimephales promelas) Exposed to
Bromacil for 64 Days 35
4 Hatchability, Abnormal Development, Survival and Growth
of Fathead Minnows CPimephales promelas) Exposed to
Dinoseb for 64 Days 37
5 Hatchability, Abnormal Development, Survival and Growth
of Fathead Minnows CPimephales promelas) Exposed to
Diuron for 64 Days 39
6 Hatchabi1ity, Abnormal Development, Survival and Growth
of Fathead Minnows (.Pimephales promelas) Exposed to
Propanil for 58 Days 41
7 Bioconcentration, Metabolism and Elimination of Five
Test Herbicide in Fathead Minnows CPiniephales promelas) ... 44
8 Binding of Herbicides to Macromolecules 53
9 96 Hr Acute Toxicity of Diuron to Various Fish Species .... 58
A-l Means, Standard Deviations and Ranges of Dissolved Oxygen
Concentrations (Percent of Saturation) in Control and
Exposure Chambers for Fathead Minnows (Pimephales
promelas) Exposed to Five Herbicides in Acute and
Early Life-Stage Tests 85
A-2 Some Water Quality Characteristics During Acute Exposures
of Fathead Minnows (Pimephales promelas) to Herbicides ... 86
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Number Page
A-3 Some Water Quality Characteristics During Early Life-
Stage Exposures of Fathead Minnows CPimephales
promelas) to Herbicides . 87
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ACKNOWLEDGMENTS
We would like to thank our Project Officer, John Teasley, from the Environ-
mental Research Laboratory-Duluth (JERL-D). U.S. Environmental Protection Agency
for his cooperation and involvement in this study. We thank the following
scientific staff members from ERL-0 for their suggestions and assistance : John
Eaton, Leonard Mueller, Gilman Veith, Douglas Kuehl, Kenneth Welsh, Nan Stokes,
Frank Puglisi, James McKim, and Glenn Christensen. Glenn Endicott and the
facilities staff at ERL-D were most helpful. We gratefully recognize the
assistance of the following University of Wisconsin-Superior technical staff
members: Michael Knuth, Steven Poirier, Catherine Morarity, Cheryl Anderson,
Pamela Shubat, James Huot, Ann Lima, Patricia Schmieder, Edward Slick, and
Kenneth Johnson. Dr. M.T. Stephen Hsia of the University of Wisconsin-Madison
kindly provided analytical standards of 3,3' ,4,4'-tetrachloroazobenzene and
3,3',4,4'-tetrachloroazoxybenzene. We thank representatives from Dow, DuPont,
Monsanto, and Rohm & Haas Chemical Companies for providing samples of technical
grade and radiolabeled herbicides for this study. We gratefully acknowledge the
work of our secretary, Joyce Barnes, in the preparation of this report.
xn
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SECTION I
INTRODUCTION
Contamination of non-target organisms* including man, by pesticides is
of common occurrence. Watersheds throughout the world are annually exposed
to pesticides, some very intensively. Certain pesticides, applied terrestrially,
enter aquatic ecosystems. Additionally, selected pesticides are applied
directly to aquatic ecosystems for control of undesirable plant and animal
species. These exposures create the potential for stress at all trophic levels
in such ecosystems.
The Federal Insecticide, Fungicide and Rodenticide Act CFIFRA) of 1947
established the registration process for pesticides used in this country. The
Act was amended by the Federal Environmental Pesticide Control Act of 1972 to
more strictly control pesticide manufacture, sale and use. In 1975 the U.S.
Environmental Protection Agency established the "Rebuttable Presumption Against
Registration" (RPAR) process to review pesticide usage patterns and to deter-
mine whether specific pesticides pose substantial questions of safety, sub-
sequently necessitating denial or cancellation of registrations under this
Act. As a result of these reviews, some pesticides will undoubtedly not be
reregistered and needs for alternative pesticides will develop. The five
herbicides of this study have been produced and marketed for several years,
and are being considered as alternative pesticides.
1 Federal Register, Vol. 40, Sec. 162.11, pp. 28268, 32329, 42746.
1
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Alachlor is a pre-emergence herbicide introduced by Monsanto Company in
1967 for control of various annual grasses and certain broadleaf weeds such as
nutsedge (Agriculture Canada, 1973). It is commonly used for weed control in
corn and soybeans.
Bromacfl, manufactured by DuPont, was first used publicly in 1963 (JU.S.
EPA, 1974). Its primary uses are for control of perennial grasses, residual
weeds, and brush at industrial sites, utility rights-of-way, railroads, and
ditch or canal Banks (U.S. EPA, 1974).
Oinoseb is manufactured tn the United States by Dow Chemical Company
(Agriculture Canada, 1973). It was first described as a weed killer in 1945,
and was the most toxic of a number of dinitro substituted phenols tested for
phytotoxicity (Crafts, 1945}. Dinosefi ammonium salt is used as a post-
emergence, selective spray in flax, beans, peas, leek, potatoes, coffee, vine-
yards and orchards; and as a dessicant in potatoes and legume crops grown for
seed (Crafts, 1975). OinoseB as the parent phenol in oil is used as a general
contact spray; and the alkanolamine salt is used to kill germinating seeds and
as early postemergence and directed sprays (.Crafts, 1975).
Diuron, manufactured by E.I. DuPont de Nemours & Company, was introduced
as an experimental compound in 1951, registered for industrial weed control in
1953 and for agricultural use in 1954 (Johnson and Julin, 1974). Diuron has
been used for many years for selective control of annual grasses and broadleaf
weeds in apple, pineapple, sugarcane, and cotton (Majka and Lavy, 1977). It
has also been extensively studied and variously used for aquatic weed control,
although it has not been registered for aquatic use (Johnson and Julin, 1974).
Propanil is a postemergence amide herbicide used in the United States to
control barnyardgrass and other weeds in rice (U-S- EPA, 1974; Smith, 1965).
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Economically, it is the most important acylanilide herbicide (.Hsia and Burant,
19791-
The present study was undertaken to determine the acute and early life-
stage tCLxicities, bioconcentration potential, and metabolism of these five
herbicides in freshwater fish. Toxicity tests and bioconcentration studies
were conducted with the fathead minnow fPimephales promelas). Metabolism
studies utilized both the fathead minnow and rainbow trout CSalmo gairdneri)
as test species.
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SECTION II
CONCLUSIONS
96 Hr LC50 values (95% confidence'intervals in parentheses) for fathead
minnows exposed to alachlor, bromacil, dinoseb, diuron, and propanil were 5.0
(4.5 -5.6), 182 (177-188), 0.7 (0.6-0.7), 14.2 (13.4-15.0), and 8.6 (7.7-9.5)
rng-L" , respectively.
Based upon one or more of the parameters of dead and abnormal fry,
juvenile fish survival, and juvenile fish growth in early life-stage tests with
fathead minnows, "no-effect" concentrations were between 0.52 and 1.0 nig-L
for alacnlor, 14.5 and 48.5 pg-L" for dinoseb, 33.4 and 70-8 ug-L~ for diuron,
and 0.4 and 0.6 yg-L" for propanil. Growth was significantly reduced at the
lowest bromacil exposure of 1.0 mg-L" , but a "no effect" range was not deter-
mined.
All five herbicides had weak potentials for concentration from the water
into fish tissue. Bioconcentration factors for the parent compounds in fat-
head minnows were 6.0, <3.2, 1!4, 2.0, and 1.6 for alachlor, bromacil, dinoseb,
diuron, and propanil, respectively.
Rainbow trout injected with the radiolabeled herbicides eliminated over
75% of the radioactivity within 24 hr. The radioactivity was eliminated both
as parent compound and as various metabolites.
One metabolite present in the bile of rainbow trout exposed to propanil
was characterized by gas chromatography /mass spectrometry (GC/MS) as either
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3',4'-dichloro-2-hydroxypropionanilide or 3',4'-dichloro-3-hydroxypropion-
anilide. Mass spectral data confirmed that 3,4-dichloroaniline had been
eliminated into the water by trout injected with radiolabeled diuron.
All of the herbicides exhibited time-dependent binding to macromolecules.
The binding was not dependent on NADPH, was not altered by the addition of
glutathione, and was not altered by the substitution of heat-inactivated
microsomes for active microsomes. Thus, the binding appeared to be due to the
affinity of the parent herbicides for protein and other macromolecules and was
not dependent upon mixed-function oxidase activity.
In unsterilized Lake Superior water held at room temperature and normal
laboratory lighting conditions Cfluorescent lights), propanil concentrations
declined curvilinearly, with an estimated half-life of 65 days. The principal
breakdown product in the propanil solution was identified by GC/MS as 3,4-
dichloroaniline. Dinoseb concentrations declined by 21% in 40 days. Bromacil
and diuron concentrations remained stable over 40 days. Alachlor was not tested
for stability in solution.
The technical grade propanil used in this study also contained 0.67 mg-g
of 3,3',4,4'-tetrachloroazobenzene (TCAB) as a contaminant. No 3,3',4,4'-
tetrachloroazoxybenzene (TCAOB) was detected in the propanil. Neither TCAB
nor TCAOB were detected in the technical grade diuron used.
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SECTION III
RECOMMENDATIONS
The effects of technical grade propanil on a variety of aquatic organisms
should be further investigated in the laboratory and in the field under normal
use patterns. Propanil residues in waters where the fauna may be impacted
should be regularly monitored in conjunction with studies on possible biological
effects.
The effects of trace amounts of 3,3',4,4'-tetrachloroazobenzene, a con-
taminant of technical propanil and diuron, on aquatic animals should be
determined. Acute and chronic toxicity data as well as more field residue
data are needed.
More surface water residue studies are needed in areas where diuron and
dinoseb are used to determine if further laboratory and/or field investigations
are warranted.
The relatively high concentrations of alachlor and bromacil necessary to
produce adverse effects in fish would not likely be encountered in aquatic
ecosystems. It is recommended that further studies on the direct effects of
these two herbicides on aquatic animals be preceded by studies on other chemi-
cals for which the potential hazard to aquatic life appears to be greater.
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SECTION IV
METHODS
Water Supply and Environmental Control
Water for toxicity and bioconcentratlon tests was from Lake Superior,
Several chemical parameters (dissolved oxygen, pH, hardness, acidity, and
alkalinity) were monitored throughout the tests by standard analytical
methods (American Public Health Association, 1975). Water was not filtered
or sterilized. A portion of the water was heated (>30°C) before being dis-
tributed to the test systems.
Temperature control was maintained by proportional mixing of heated and
unheated lake water in constant head reservoirs. Electronic sensors monitored
the temperatures in the reservoirs and added heated water when necessary to
maintain a desired water temperature of 25°C. Lighting for the acute and
embryo-larval toxicity tests was supplied by two 40-watt fluorescent bulbs
centered above the exposure chambers.
Test Organisms
Fathead minnow brood fish were received from stock maintained by the
Environmental Research Laboratory-Duluth, U.S. Environmental Protection
Agency. Brood fish were maintained at 25°C, and were fed twice daily a diet
of frozen adult brine shrimp.
Asbestos pipe (12.5 cm O.D.) cut in half, longitudinally, was used as the
spawning substrate. The spawning substrates (tiles) were checked daily for
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egg deposition. Eggs were removed from the tiles the same day (<24 hr) that
spawning occurred when used in early life-stage tests. For the culturing of
fathead minnows to be used in acute toxicity or bioconcentration tests, tiles
with "eyed up" eggs were placed in temperature controlled C25°C) hatching
chambers to complete incubation. Once they had hatched (approximately 96 hrs
after spawning at 25°C) fry were transferred to rearing aquaria where they
were fed fresh, newly hatched Brine shrimp nauplii 3 times daily.
Rainbow trout ranging in size from 75 to 225 g were used for in vivo
metabolism studies. Larger fish were used for in vitro metabolism studies.
Trout were received from the Environmental Research Laboratory-Duluth, where
they had been reared at 12°C.
Test Chemicals
Alachlor. Non-radioactive alachlor or Lasscrwas supplied by Monsanto
Chemical Co. as technical grade material with a purity of 92.6% (/lot no. MHF-
183). C-Labeled alachlor was uniformly ring-labeled compound with a specific
activity of 1.9 mCi-mmol" . Alachlor [2-chloro-2' ,6'-diethyl-N-Onethoxymethyl)
acetanilide] is an amide herbicide with the formula C-j.H^QClNOg. a molecular
weight of 269.9, and the following chemical structure (U.S. EPA, 1974; Agri-
culture Canada, 1973):
8
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Bromacil. Non-radioactive bromacil was supplied by DuPont Chemical Co-
14
as technical grade material with a purity of 95.0%. C-Labeled bromacil was
supplied by DuPont Chemical Co. from New England Nuclear Clot no. 896-232).
It was radiolabeled at the number 2 carbon position, and had a specific activity
of 6.25 mCi-mmol" . Bromacil C5-bromo-3-sec_-butyl-6-methyluracil) is a uracil
derivative with the formula CgH,3BrN202» a molecular weight of 261.1, and the
following chemical structure (U.S. EPA, 1974; Agriculture Canada, 1973):
Dinoseb. Non-radioactive dinoseb (Premerg^r; was supplied by Dow Chemical
Co. as technical grade material with a purity of 98.02 Clot no. AGR 133942).
14
C-Labeled dinoseb was supplied by Dow Chemical Co. from Pathfinder Labora-
tories (Jot no. 80612). It was uniformly ring-labeled with a specific activity
of 7.80 mCi-mmol . Dinoseb (.2-sec_-6uty1-4,6-dinitrophenol) is a phenolic
herbicide with the formula c-joH<|2°5N2' a molecular weight of 240.2, and the
following chemical structure (U.S. EPA, 1974; Agriculture Canada, 1973).
CH3CH2CH
Diuron. Non-radioactive diuron (Karmejrjwas supplied by DuPont Chemical
14
Co. as technical grade material with a purity of 98.6%. C-Labeled diuron
14
was supplied by DuPont Chemical Co. as carbonyl C-labeled material with a
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specific activity of 0.98 mCi.-mmol . C-Labeled diruon in the form of
uniformly ring-labeled compound was supplied by California Bionuclear Corpora-
tion Uot no. 780627AB) with a specific activity of 12.2 raCi-mmo!"1. Diuron
[3-(3,4-dichloropheny1)-l ,1 -dimethyl urea] is a phenylurea herbicide with the
formula CgH^Cl^gO, a molecular weight of 233.1, and the following chemical
structure (U.S. EPA, 1974; Agriculture Canada, 1973):
8
Propanil. Non-radioactive propanil or STAMi was supplied by Rohm and
Haas Chemical Co. as technical grade material with a purity of 85.9% Oot no.
8771). C-Labeled propanil was also supplied by Rohm and Haas from New
England Nuclear (lot no. 862-186). It was uniformly ring-labeled, and had a
specific activity of 10.42 mCi-mmol . Propanil (.3' ,4'-dich1oropropionanilide)
is an amide herbicide with the formula CgHgNOClp, a molecular weight of 218,1,
and the following chemical structure (IJ.S. EPA, 1974; Agriculture Canada, 1973)
NH(jC2H5
Toxicity Tests
Acute Toxicity Tests. Acute tests were run in 17.5 L glass aquaria
(20 cm x 35 cm x 25 cm) containing 6 L of water. Water was delivered through.
proportional diluter systems (Mount and Brungs, 1967) with five toxicant con-
centrations and a control. Each, toxicant concentration and control was run in
10
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duplicate. The diluters delivered 0.5 L to each test tank 4 to 8 times per
hour depending upon the toxicant used.
Twenty 30-day-old fathead minnows were placed into each aquarium. Obser-
vations for mortality were made several times during the first 24 hours and
twice daily for the remaining 7 days of the test. Fish were not fed during
the test. Fish were considered dead, and removed from the aquaria if
opercular movements had ceased.
Water temperatures were measured two or more times weekly. All tests were
run at a nominal water temperature of 25 C.
Toxicant stock solutions of alachlor, dinosefc, diuron, and propanil were
generated from sand columns, adaptations of a procedure described by Veith and
Comstock (.1975). A near-saturation solution of these four herbicides in Lake
Superior water was maintained by circulating the stock solution through a sand
column containing the toxicant.
The bromacil stock solution was prepared differently due to the high con-
centrations required to kill fathead minnows. Bromacil was dissolved at near-
saturation levels in a large reservoir C230 L) of Lake Superior water. This
solution was continuously pumped to a head reservoir 07 L), from which it
was dispensed to the mixing cell of the proportional diluter.
Early Life-Stage Tests. Early life-stage tests were run using the same
basic diluter and exposure system as in the acute tests. The system was
modified to permit incubation of eggs in oscillating 130 mL specimen bottles
with 200 ym mesh nylon screen bottoms. The bottles were oscillated by a rocker
arm assembly at a rate of about 6 times per minute, causing the eggs to move
a vertical distance of about 5 cm. Modifications were also made within the
aquaria to reduce the area and confine the fish. A glass and stainless steel
11
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mesh (.200 urn) barrier reduced the volume of water to 4.5 L. The bottom color
of the aquaria was changed from black to white, permitting easier observation
of fry.
Fathead minnow eggs <24 hours old were placed into the incubation jars
(50 eggs per jar; 2 jars per treatment replicate). Dead eggs were removed
daily, and upon completion of hatch the survivors were counted and 15 fry from
each incubation jar were released into the aquarium (30 per aquarium). Feeding
was begun the day after hatching and continued to the end of the test. Finely
granulated dry fish food (Tetramirr*j and newly hatched brine shrimp were fed
for the first 30 to 45 days after hatching and supplements of frozen adult
brine shrimp were added to the diet for the last 15 to 30 days of the test.
Equal volumes of food were provided to each aquarium.
Several parameters were measured in the early life-stage tests. These
included egg hatchability, occurrence of abnormal and dead fry at time of
transfer from egg cups into the aquaria, fry survival at end of exposure
period, and wet weight and length at end of exposure period. Observations
were also made on behavior and development of the fry throughout the exposures.
Bioconcentration Tests
Bioconcentration tests were run in a modified diluter system using fluid
metering pumps to deliver the test compounds CC-labeled) to each. 27-1 treat-
ment aquarium. Water (.1 L at 25°C) was delivered 4 times per hour to each
treatment and control aquarium where it was mixed with the test compound be-
fore going into the aquarium. Each test solution was a mixture of technical
grade and radio!abeled compound dissolved in methanol. Two exposure concen-
trations of herbicide were used which differed by approximately one order of
magnitude. All aquaria including the controls received 100 mg-L methanol.
12
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One hundred fathead minnows (.30 days old) were placed Into each aquarium
and exposed to the test compounds. Five fish were removed and individually
analyzed on days 1, 2, 4, 7, 10, 14, 17, etc. until an equilibrium was achieved
14
in total C between the exposure water and the fish tissue. When equilibrium
was reached, the remaining fish were transferred to 27-L aquaria receiving only
Lake Superior water (no herbicide or methanol), with a turnover rate of 4 L-
hr" . Elimination rates were determined by individually analyzing samples of
14
5 fish on days 1, 2, 4, 7, 10, 14, 17, etc. until C activity had been almost
completely eliminated.
Measurements of Herbicide Concentrations in Water and Fish
Alachlor. For the acute and embryo-larval toxicity tests, alachlor was
extracted from water with a hexane:acetone (.90:10. v/v) mixture. A 10 ml water
sample was extracted three times with 5 mL of solvent mixture. The extract
was collected in a 100 ml volumetric flask for the acute test samples and in
a 500 ml Kuderna-Danlsh evaporative concentrator for the embryo-larval test
samples. Acute test samples were diluted with hexane to appropriate volumes
for GLC analysis. Embryo-larval test samples were concentrated on a steam
bath prior to analysis.
Gas-liquid chromatographic (GLC) analysis was performed on a Tracor MT
160 instrument equipped with a Ni electron capture detector and a glass
column C6' x 1/4") packed with 3* OV-101 on 100/120 mesh Gas-Chrom Q, Nitro-
gen was the carrier and detector purge gas with flow rates of 34 and 35 mL-
min , respectively. Operating temperatures for inlet, column and detector
were 222°C, 236°C, and 263°C, respectively. The retention time of alachlor was
3.72 minutes. At an output attenuation of 8 X, a 0.60 ng injection gave a
19% full scale deflection. The linear working range was 0.20 to 0.60 ng.
13
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Standard curves were prepared daily by injecting several volumes up to
6 pL of a 0.10 ng.yL working standard. Samples were diluted or concentrated
to yield responses within the linear range of the gas chromatograph, and con-
centrations determined from a standard curve regression equation. Based upon
20 spikes in Lake Superior water ranging in concentration from 0.2 to 3.0
mg-L , the mean alachlor recovery was 92.8 ± 9.12. Water concentrations were
corrected on the basis of this recovery.
For the alachlor bioconcentration study, C-alachlor was extracted from
water with a mixture of benzene and isobutanol (50:50, v/v). A 100 ml water
sample was extracted with 25 itl benzene:isobutanol and a 5 ml aliquot of the
solvent layer was added to 15 ml of scintillation fluid. The scintillation
(R) CR)
fluid consisted of Permafluor III, (Packard Instrument Co.) Triton X-100, and
toluene (10:33:57, v/v). Samples were counted for a minimum of 5 minutes along
with standards on a Packard Tricarb liquid scintillation counter.
14
Concentration-count relationships were determined for C-alachlor by
using 4 duplicated standards. Background counts and instrument quench curve
readings were made with each series of runs on the scintillation counter. A
computer program corrected counts from standards and samples for differences in
14
instrument quench readings, and calculated yg quantities of C parent compound
equivalents in water based on a regression line fit (least squares) from parent
compound standard curves. Sample concentrations were then calculated and
expressed in yg-L of water. Due to water solubility of the solvent mixture,
the solvent volume was reduced from 25 ml to 17.87 ml (n=6). Concentrations
14
were corrected for this reduction, and recovery of C-alachlor from Lake
Superior water averaged 102.6 t 2.92, (n=22).
C-Alachlor was measured as parent compound equivalents in whole fish at
14
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various time intervals during the bioconcentration study. Whole fish were
blotted dry, weighed, and oxidized in a Packard Tri-Carb Model B306 Sample
Oxidizer. Weighed fish were placed in a combustible cone, 0.3 mL of
CombustaiaH.Packard Instrument Co.) solution was added, and the fish were
oxidized. The liberated C-labeled C02 was trapped with 6 mL of Carbosorb
(Packard Instrument Co.), then rinsed and diluted with 14 ml of Permafluor r^
(.Packard Instrument Co.). The solution was counted in a liquid scintillation
counter, and the counts quantitatively equated to parent compound based upon
14.
an alacKlor standard curve. Recovery of C-alachlor from fish averaged
36.8 + 5-8%, (n=15). Sample concentrations were then calculated and expressed
in ng-g of fish tissue.
Bromacil. For acute and embryo-larval tests, bromacil water concentra-
tions were measured on a Beckman UV-visible double-beam spectrophotometer at
a wavelength of 280 nm. Bromacil standard solutions in methanol ranging in
concentration from 0.8 to 10 mg-L were used to prepare standard curves on
each sampling day. Exposure chamber water samples were either measured
directly, as in some of the lower concentrations in the embryo-larval test, or
diluted with distilled water to fall within the absorbance range of the
standards. The absorbances of the control chambers were subtracted as correc-
tion factors in calculating exposure tank concentrations of bromacil. The
minimum detection limit for bromacil was approximately 0.5 mg-L .
14
C-Bromacil water concentrations in the bioconcentration study were
determined by methylene chloride extraction. Water samples of 100 mL were
extracted with 25 mL of methylene chloride. The solvent was evaporated to
dryness in a 60°C hot water bath under a hood. The residue was dissolved in
5.0 mL of toluene, 15.0 mL of scintillation fluid (Permafluor Iin Triton X-
15
-------
100 ,and toluene, 10:33:57, v/v) were added, and the sample counted on a
scintillation counter as described for alachlor. Recovery of C-bromacil
from Lake Superior water was 87.5 ± 2.62, (ns20).
14
C-Bromacil was measured as parent compound equivalents in whole fish at
various time intervals during the bioconcentration study. The procedure was
the same as described previously for C-alachlor. The mean recovery of C-
bromacil was 88.1 ± 4.2%, (ji=10).
Dinoseb. In the acute toxicity test where all dinoseb concentrations
exceeded 0.10 mg-L , dinoseb was extracted from water by placing a 10.0 mL
sample into a 60 ml separatory funnel. Concentrated sulfuric acid C2-4 drops)
was added to lower the pH below 2.0, followed by the addition of 0.20 ml of
saturated NaCl solution. Oinoseb was then triple-extracted with 10 ml of
benzene. After each extraction, the benzene layer was drained through a 60°
funnel plugged with glass wool (.acetone washed) and filled with 20 g of
anhydrous sodium sulfate. The separatory funnel was rinsed with 10 ml of
benzene which was also passed through the sodium sulfate. The extracts were
collected in a 25 ml volumetric flask and diluted to volume, prior to methy-
1ation.
In the embryo-larval toxicity test, where all dinoseb concentrations were
less than 0.10 mg-L , 500 ml water samples were collected. To each sample,
approximately 3 ml of concentrated sulfuric add and 10 ml of saturated NaCl
solution were added. A triple-extraction with 50, 25, and 25 mi. of benzene
was made, with the extracts collected in a 500 mL Kuderna-Danish evaporative
concentrator. The sample was concentrated on a steam bath to approximately
2 mL, and further reduced to 1 mL under a stream of nitrogen gas. The con-
centrated extract was then methylated.
16
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Diazomethane was prepared daily with each set of samples according to the
procedure in "Manual of Analytical Methods for the Analysis of Pesticide
Residues in Human and Environmental Samples," (U.S. EPA, 1977). Under a
ventilated hood, small increments of N-methyl-N'-nitro-N-nitrosoguanidine
(Aldrich Chemical Co.) crystals were added to a test tube containing 5.0 ml
of 20% NaOH and 3 ml of hexane, until the solution was saturated with diazo-
methane. The reagent was used immediately after gas bubbles had ceased
evolving.
Dinoseb sample extracts from both the acute and embryo-larval tests were
methylated by adding approximately 12 drops of the diazomethane-saturated
hexane layer to 1.0 ml of sample extract. The reaction flask was swirled and
allowed to stand for at least 15 minutes. The methylated extract was then
diluted to an appropriate volume for GLC analysis.
6LC analysis of dinoseb was performed on Tracer MT 160 and Tracer 550
instruments, both equipped with Ni electron capture detectors. The Tracer
MT 160 glass column t6' x 1/4" i.d.) was packed with 3% OV-101 on 100/120
mesh Gas-Chrom Q. Inlet, column, and detector temperatures were 230, 205,
and 260°C, respectively. Nitrogen was the carrier and detector purge gas at
flow rates of 45 and 36 nt-min , respectively. With an attenuation of 2 X,
0.05 ng of dinoseb produced a 25% full-scale deflection. The Tracor 550 glass
column was 6' x 1/4" i.d. packed with 32 OV-101 on 80/100 mesh Chromosorb W.
Inlet, column, and detector temperatures were 223, 196, and 259°C, respectively.
Nitrogen was the carrier gas with a flow rate of 65 mL-min" and the detector
purge gas at a flow rate of 49 mL-min . With an attenuation of 4 X, 0.05 ng
produced a 28% full-scale deflection.
Dinoseb standard solutions were prepared from a 1.0 mg-L" stock solution
17
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in acetone. A 1.0 yg-mL intermediate standard was prepared by dilution of
the stock with benzene. Methylation was performed on 1.0 ml of this standard,
which was then diluted with hexane to give a working standard of 0.01 ug-mi_~ •
Standard curves were prepared daily fay injecting from 1 to 5 uL, and plotting
nanogram quantity versus peak height. The detection limit for dinoseb using
a 500 mL water sample was 0.005 ug-L .
Dinoseb recovery from water was determined by spiking Lake Superior water
over a range of concentrations from 0.50 u9'L to 1.00 mg-L . The mean
recovery was 96.4 + 8.7%, (n=22).
C-Dinoseb was extracted from water in the bioconcentration study with
benzene-.isobutanol, (50:50 v/v). A 100 ml sample was acidified with 2 drops of
concentrated H^SO. and extracted with 25 ml benzene:isobutanol. A 5 ml aliquot
of the solvent layer was added to 15 mL of scintillation fluid (same as de-
scribed previously) and counted as described for alachlor. After correction
14
for solvent volume reduction, recovery of C-dinoseb from Lake Superior water
was 102.5 t 3.2%, Cn=20).
14
C-Dinoseb concentrations in whole fish, were measured as described for
14
C-alachlor, using a sample oxidizer, and were expressed as parent compound
equivalents. The mean recovery from fish tissue was 78.2 ± 9.6Z (n=20).
Piuron. Diuron water concentrations in the acute toxicity test were
measured with a Beckman double-beam UV-visible spectrophctometer at a wave-
length of 250 nm. Diuron standards were prepared in methanol and ranged in
concentration from 0.6 to 4.0 mg-L . Standard curves were prepared daily
with each set of analyses, and exposure chamber concentrations were determined
from the standard curve regression equation.' Chamber water samples were
diluted as necessary with distilled water to give absorbance readings within
18
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the standard curve concentration limits. The absorbance readings of control
tanks were subtracted from samples as correction factors in calculating tank
concentrations. The minimum detection limit was 0-1 mg-L .
Diuron water concentrations in the embryo-larval test were below the UV-
visible spectrophotometer sensitivity, ranging from means of 2.0 to 61.8
pg-L . For this test, diuron water concentrations were determined by
extraction with methylene chloride and analysis by high pressure liquid chroma-
tography CHPLC), based upon the procedure of Farrtngton e_t al_. 0977).
A 500 ml water sample was triple extracted with methylene chloride C50,
25, 25 ml), and the solvent collected in a 500 nl Kudema-Danish evaporative
concentrator. The methylene chloride was reduced to a volume of approximately
2 ml on a steam bath, and further evaporated to dryness under a stream of
nitrogen gas. The residue was redissolved with 5.0 ml of methanol (spectro-
scopy grade), and diluted as appropriate with methanol for HPLC analysis.
HPLC analysis was performed on a Waters Associates Model 6000A instrument
equipped with a Model 440 UY absorbance detector with a fixed wavelength of
254 run. A stainless steel column (.300 x 2 mm l.d.) was used, packed with
Porasil-18 and operated at ambient temperature. The mobile phase was 602
methanol and 40% water (deionized and degassed) with a flow rate of 2 mL-min
and operated at a pressure of approximately 3800-4000 psi. A WISP (Waters
Intelligent Sample Processor) automatic sample injection system was used, with
a standard injection volume of 20 pL for all samples and standards.
Diuron standards in methanol ranging in concentrations from 0.125 ng-yL
to 1.00 ng-uL were run daily with each set of analyses. Sample extracts .
were adjusted to fall within this range of concentrations, and sample concen-
trations were calculated from the standard curve regression equation. The
19
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retention time of diuron was 3.83 minutes. At an attenuation of IX, a 20 ng
injection produced a 37% full-scale recorder deflection. The linear range
of the detector for diuron was 2-5 to 40 ng. At concentrations of spiked
samples ranging from 5 ug-L to 50 yg-L , the mean recovery from Lake
Superior water was 88.9 ± 6.0%, (n=19).
Propanil. For the acute toxicity test, where mean propanil concentrations
ranged from 1.9 to 15.7 mg-L , 10 nt water samples were collected and triple-
extracted with hexane 00, 5, and 5 ml]. The extracts were diluted to
appropriate volumes for GLC analysis.
In the propanil embryo-larval test, where mean propanil concentrations
ranged between 0.4 and 3.8 yg-L , 100 nl water samples were collected and
triple-extracted with 25 ml portions each of hexane. The combined extracts
were collected in a Kuderna-Danish concentrator, and concentrated on a steam
bath to appropriate volumes for GLC analysis.
Two gas chromatographs were used for analysis of propanil concentrations.
A Tracer 220 instrument equipped with a Ni electron capture detector was
used for samples between 1.0 and 20.0 mg-L . The glass column ($' x 1/4"
i.d.) was packed with 3% OV-101 on 80/100 mesh Chromosorb W-HP. Argon-methane
(55:5) was the carrier and detector purge gas with a flow rate of 50 mL-min
through both the column and detector. An output attenuation of 4X was used.
Operating temperatures for inlet, column, and detector were 205°C, 193°C, and
300°C, respectively. The retention time of propanil was 3.05 minutes, and an
injection of 0.80 ng gave a 173. full scale recorder deflection. The linear
working range was 0.4 to 1.0 ng.
A Tracor 550 instrument equipped with a Ni electron capture detector
was used for samples between 0.1 and 10.0 ug-L" . The glass column (6' x
20
-------
1/4" i.d.) was packed with 3% OV-101 on 80/100 mesh Chromosorb W. Nitrogen was
the carrier and detector purge gas with flov/ rates of 60 and 30 mL-min" ,
respectively. An output attenuation of 2x was used. Operating temperatures
for inlet, column, and detector were 218°C, 196°C, and 240°C, respectively.
The retention time of propanil was 3.22 minutes, and an injection of 0.20 ng
gave a 35% full scale recorder deflection.
A stock solution of 1.0 mg-mL" propanil (952 purity) was prepared in
acetone. Standard solutions of propanil were prepared by making further
dilutions with hexane. The working standards used for GLC analyses contained
0.01 ng*yL and 0.10 ng-^L propanil. Standard curves were prepared daily
with each set of analyses by plotting peak height versus nanogram quantity of
standards CO-05, 0.10 and 0.20 ng), and sample concentrations determined from
the standard curve regression equation.
The recovery of propanil from Lake Superior water was determined by
spiking the water over a range of concentrations between 0.20 mg-L to 200
mg-L . Based upon 24 determinations, the mean recovery was 98.5 t 3.3%.
C-Propanll was extracted from water 1n the bioconcentration study with
toluene. A 500 ml water sample was extracted with 25.0 ml of toluene. A
5.0 aliquot of the solvent layer was placed into a scintillation vial, and
15.0 ml of scintillation fluid Csame composition as before) were added.
Samples were then counted and concentrations determined, as described for
alachlor. The mean recovery from Lake Superior water was 92.8 ± 1.9% Cn=32).
14
C-Propanil concentrations in whole fish were measured as described for
14
C-alachlor, using a sample oxidizer, and were expressed as parent compound
Huivalents.
4.5i (n=20).
14
equivalents. The mean recovery from fish spiked with C-propanil was 90.9 ±
21
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Metabolite Characterization
Rainbow Trout In vivo Experiments. Rainbow trout (Sal mo gairdneri) from
stock maintained at ERL-Duluth were used for initial in vivo herbicide metab-
olism experiments. Individual fish weighing 100-150 g were tranquilized with
tricaine methane sulfonate (MS 222) at a concentration of 50 mg-L , and
transferred to a small aquarium in which the water temperature was maintained
at 10°C. The aquarium was shaded on the sides and top, and each fish was
allowed to remain undisturbed for 24 hours. On the following day, the fish
were lightly anaesthetized with MS 222 C50 mg-L ), and injected with 1 wCi
14
of C-labeled herbicide. After an additional 24 hours, the trout were
anaesthetized with MS 222 (100 mg-L ), blood taken from the caudal artery,
bile removed from the gall bladder with a 1 cc syringe, and the liver excised.
The liver, bile, blood, and aquarium water were analyzed for metabolites.
Following the initial in vivo experiments, rainbow trout weighing 100-150
g were exposed for 4-5 days to unlabeled herbicide dissolved in 10-12 L of
Lake Superior water at concentrations below acutely lethal levels (e.g. pro-
panil at 1.0 mg-L ). Water temperature was maintained at 10°C, and the water
changed after the first 24 hrs to remove fecal material. The fish were not
fed during the exposure period. One day prior to sacrifice, individual fish
were injected with 1.0 pCi of C-labeled herbicide. The fish were killed,
the liver excised, and the gall bladder carefully separated from the liver.
The gall bladder, bile, and liver were analyzed for metabolites.
Metabolite Recovery from Water. Aquarium water containing radio! abeled
alachlor, bromacil, diuron, or propanil metabolites was passed through a
column (2.5 x 30 cm) of Amberlite XAD-2 resin. The column was washed with
22
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1-2 bed volumes of distilled water, then eluted with 900 mL of methanol.
Aliquots of the water prior to XAD-2 treatment and of the methanol eluate
were counted in a liquid scintillation counter to determine total radio-
activity recovered from the column. The methanol eluate was evaporated to
dryness in a stream of air prior to TLC. Dinoseh did not absorb well to the
XAD-2 resin; consequently, dinoseb aquarium water was acidified with HC1 to a
pH of 2 and passed through a column of Sephadex LH20. The column was eluted
with methanol, and the eluate evaporated to dryness.
Metabolite Recovery from Liver. Fish liver was homogenized with 20-25
volumes of cold acetone in a Waring blender, followed by centrifugation for
5 min. at 900 X g to remove insoluble solids. The acetone extract was dried
with sodium sulfate and then evaporated in a stream of air. An aliquot of
the acetone extract was counted in a scintillation counter and the solid resi-
due processed in a Packard Sample Qxidizer to determine unextracted radio-
activity.
Metabolite Recovery from Bile. Bile was diluted 5:1 with water, acidified
with HC1 to a concentration of 0.1 M, and extracted three times with 10 ml of
ethyl ether. The ether extract was evaporated to dryness in a steam of
nitrogen, and the residue redissolved in acetone prior to TLC. After TLC,
the chromatograms were examined under ultra-violet light for absorbing spots
or bands, and radioactive areas located with a Packard radiochromatogram
scanner. Areas which were both radioactive and ultra-violet absorbing were
scraped to remove the silica gel, and the silica gel eluted with methanol.
The eluted material was then analyzed by gas chromatography/mass spectrometry
(GC/MSl-
23
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Metabolite Recovery from Blood. Whole blood was centrifuged for 20 min
at 900 X g. The plasma was diluted with 3.5 volumes of acetone followed by
centrifugation at 900 X g for 5 min. to remove precipitated protein. The
extract was loaded on an XAD-2 column 0 x 25 cm), and processed as described
above for bile. Aliquots of the extract were counted by liquid scintillation
techniques, and the precipitated solids burned in a sample oxidizer for radio-
activity determinations.
The packed cells were washed twice with isotonic saline followed by 2 ml
of acetone. The suspension was centrifuged for 5 min at 900 X g, and the
extract loaded on a 1 x 25 cm XAD-2 column as described above. Radioactivity
in the extract and residual solid was determined as described for plasma.
Thin-Layer Chromatography. Chromatography was done on commercially avail-
able silica gel coated Mylarnsheets (.Eastman Co.) with a fluorescent indicator.
Samples to be analyzed by mass spectroscopy were chromatographed on glass
plates coated with silica gel containing a fluorescent indicator. Chromato-
gram development was in an Eastman "Chromagram" development apparatus.
The following TLC solvent systems were used for alachlor and its meta-
bolites - hexane: ethyl acetate (.97:3, v/v); for bromacil and its metabolites •
to1uene:ethy1 acetate (80:20, v/v); and for dinoseb plus metabolites -
toluene:n_-butanol (60:40, v/v). Propanil extracts were initially developed
with. 85% toluene and 15% methanol. The chromatogram was developed a second
time with 65% toluene and 35% methanol with the solvent allowed to migrate
only 40% of the original distance to prevent displacement of the previously
separated less polar components. The chromatogram was developed a third time
with 35% toluene and 65% methanol to 20% of the original migration distance.
By utilizing multiple development, all metabolites were forced to migrate
24
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from the origin.
Gas Chromatographic-Mass Spectrometric (6C/MS)
Characterization of Metabolites
Analyses were performed on extracts of exposure water or fish bile with
a Finnigan Model 4023 A at the Environmental Research Laboratory-Duluth (U-S.
Environmental Protection Agency). A Finnigan INCOS 2300 data system was used
on line to acquire and process mass spectral data.
A glass GC column (6' x 1/8" I.D.) was packed with 32 OV-1 (methyl
silicone) on 60/80 mesh Gas-Chrom Q. .The following GC conditions were
typically employed: helium carrier gas flow rate, 20 mL-min ; injector
temperature, 250°C; separator oven temperature, 280°C; transfer line tempera-
ture, 280°, and column oven temperature programmed from 100°C to 225°C at
4°C-min"1.
The mass spectrometer was scanned 50 to 500 at 2.05/decade in electron
impact mode (70 ev). The instrument was tuned to provide a 442/198 amu ratio
of 0.7 - 0.9 for the spectrum of Ultramark 443 (decafluorotriphenylphospine,
PCR Research Chemicals, Inc.). Calibration was accomplished with FC43
(perfluorinated trifautyl amine).
Rainbow Trout Liver in Vitro Experiments. A freshly excised rainbow
trout liver was rinsed with 1.15% KC1, and homogenized with two volumes of
1.15% KC1 in a Potter-Elvehjem homogenizer. The homogenate was centrifuged
for 20 minutes at 10,000 X g at 4°C, and the supernatant liquid used as the
enzyme source. Incubation mixtures contained Q.I M tris-acetate, pH 7.5,
0.005 M MgCl2, 0.2 mM NADPH, 0.05 yCi of UC-labeled herbicide, and 0.2 ml
of enzyme preparation in a total volume of 1.0 mL. Incubation was for 2
25
-------
hours at 25°C. After the incubation period, the mixture was acidified with
HC1 to 0.1 M and extracted three times with 10 ml ethyl ether. The ether
extract was evaporated to dryness in a stream of nitrogen, the residue
redissolved in acetone, and the extract analyzed by thin-layer chromatography
(TLC) and radiochromatogram scanning.
In Vitro Binding of Herbicides to Macromolecules. Fresh rainbow trout
livers were homogenized and centrifuged as above, with the supernate used
as the source for metabolizing enzymes. Incubation mixtures were prepared
as above, with incubation initiated by the addition of the enzyme source.
At various times, 0.1 mL aliquots were removed from the incubation
mixture and pipetted onto 2.0 x 2.5 cm pieces of blotting paper. After the
paper absorbed the aliquot of incubation mixture, it was immersed in 70%
ethanol to precipitate protein and other macromolecules. The paper was
sequentially washed in 100% ethanol containing 5 mg-L of unlabeled
herbicide, acetone, and finally acetonitrile. The paper was then air dried,
burned in a Packard Sample Oxidizer, and counted in the liquid scintillation
counter to determine radioactivity.
Fathead Minnow Metabolite Studies. Extra fish from each of the fathead
minnow bioconcentration tests (.excluding bromacil) were frozen at the end of
the uptake phase of each test. Pooled whole fish samples 0-5 - 6.0 g) were
homogenized with 6.5 ml of water in a Potter-Elvehjem homogenizer, and a 50 uL
14
aliquot of the homogenate was analyzed for total C radioactivity. The
homogenate was acidified with 0.4 mL of concentrated HC1 and extracted three
times with 10 ml portions of ethyl ether. For the herbicides studied in this
manner (alachlor, dinoseb, diuron, and propanil), ether extraction failed to
26
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remove any radioactivity from the aqueous phase. The aqueous layer was
centrifuged and the pellet oxidized in a Packard Model 306 Sample Oxidizer
and counted in a liquid scintillation counter to determine particulate-bound
14
C. The supernatant was lyophilized, and the residue dissolved in a small
volume of acetone. Acetone solutions of the water soluble fractions were
applied on silica gel thin-layer strips, and co-chromatographed with appropri-
ate standards of parent herbicides. A standard of the propanil metabolite,
3,4-dichloroaniline, was also co-chromatographed with the extract of the
propanil-exposed fish. Thin-layer strips were developed with the following
solvents Cv/v): alachlor - 99:1 toluene:methanol; dinoseb and diuron - 90:10
toluene.-methanol; and propanil - 92:8 toluene:methano1.
14
Developed thin-layer strips were first analyzed for C activity on a
radiochromatogram scanner. Radioactive bands were scraped from the silica
gel strips, and compounds eluted from the silica gel with water followed by
methanol in the case of the propani1-exposed fish extracts, and methanol alone
for extracts of fish exposed to alachlor, dinoseb, and diuron. Eluates were
added to scintillation fluid and counted to determine relative contributions
14
of parent herbicides and metabolic products toward total C in the water-
soluble fractions.
Herbicide Stability in Solution
Solutions of bromacil, diuron, dinoseb, and propanil at concentrations
approximating the 96 hr LCgQ concentrations for fathead minnows were prepared
using Lake Superior water. Bromacil, diuron, and dinoseb solutions were
held in 1 L stoppered clear-glass bottles in duplicate, and were allowed to
stand in a light 02 hr fluorescent light and 12 hr dark) and temperature
27
-------
C25°C) controlled room for 40 days. Samples were taken from each bottle at
various time intervals. Propanil stability was studied using 500 ml stock
solutions in duplicate held under the same light conditions as the others but
at 20°C and sampled over 194 days. Analytical techniques were the same as for
the acute toxicity tests.
GC/MS Analysis of Technical Grade Propanil and Diuron
Technical grade propanil and diuron samples were analyzed for the presence
and amount of 3,3',4,4'-tetrachloroazobenzene (.TCAB). and 3,3'4,4'-tetrachloro-
azoxybenzene (TCAOB) by GC/MS. A Finnigan 4000 GC/MS with multiple ion de-
tection (MID) was used, with a 15 m x 0.2 mm 1.0. SE-30 fused silica capillary
column operated in the splttless injection mode. The following GC/MS operating
conditions were employed: He carrier gas at a linear velocity of 47 cm-min ,
injector temperature of 250°C, transfer temperature of 250°C, oven temperature
programmed from 60-180°C at an initial hold of 0.5 min. and an increase of
10°C-min , and then increased at a rate of 5°C-min" between 180-250°C, with
a 10 m1n. hold at 250°C; 70 eV electron impact, 0.35 MA emission current,
Q
1750 V EM voltage, and 10 sensitivity. Samples were quantitated using p_-
diiodobenzene as an internal standard. The following ions were monitored for
0.210 sec during each 1.00 second MID scan period: TCAB and TCAOB, m/z 145
and 147; DIB, m/z 330.
28
-------
SECTION V
RESULTS
Toxicity Tests
Acute Flow-Through Toxlcity Tests. Chemical characteristics of the test
water during the acute exposures are presented in Appendix A. Herbicide con-
centrations in the exposure chambers were corrected for recoveries of the test
compounds from water.
Alachlor was tested at six duplicated control and exposure concentra-
tions (means of 1.70, .3.19,.5-04, 8.56, and 14-8 mg-L ). The mean water
temperature was 24.7°C with a range from 23.6 to 25.8°C. LC5Q values for the
replicates combined at 96 and 192 hrs were 5.0 and 3.0 mg-L , respectively
(Table 1).
Bromacil was tested at six duplicated control and exposure concentra-
tions (means of 108, 127, 146, 172, and 207 mg-L ). The mean water tempera-
ture was 25.0°C with a range from 24.1 to 25.9°C. LC5Q values for the
replicates combined at 96 and 168 hrs of exposure were 182 and 167 mg-L ,
.respectively (Table 1). :
Dinoseb was tested at six duplicated control and exposure concentrations
(means of 0-24, 0.34, 0.50, 0.69,and 1.00 mg-L ). The mean water temperature
was 25.0°C and ranged from 24.0 to 25.8°C. LC5Q values for the replicates
combined at 96 and 192 hrs of exposure were 0.7 and 0.5 mg-L , respectively
(Table 1).
29
-------
to
o
TABLE 1. LC50 VALUES AND 95% CONFIDENCE INTERVALS FOR ACUTE
TOXICITY TESTS WITH REPLICATES COMBINED
(All Values Expressed in mg-L"1)
Compound
Alachlor
Bromaci 1
Dinoscb
Diuron
Propanil
24 Hr LC5Q
9.9 (9.3-10.6)
185 (179-191)
0.8 (0.8-0.9)
23.3 (21.0-25.9)
11.5 (10.8-12.2)
48HrLC5Q
6.6 (6.3-7.0)
183 (177-189)
0.7 (0. 6-0.8)
19.9 (19.5-20.4)
10.2 (9.6-10.8)
96
5.0
182
0.7
14.2
8.6
"r LC50
(4-5-5.6)
(177-188)
(0.6-0.7)
(13.4-15.0)
(7.7-9.5)
192 Hr LC5Q
3.0 (2.6-3.5)
167 (162-174)*
0.5 (0.4-0-5)
7.7 (6.0-9.9)
3.4 (2.4-4-8)
* Values are for 168 hr LC5Q.
-------
Diuron was tested at six duplicated control and exposure concentrations
(means of 5.54, 7.94, 11.14, 16.42, and 24.20 mg-L ). The mean water tem-
perature was 24.3°C with a range from 22.3 to 25.6°C. LC5Q values for pooled
replicates of diuron at 96 and 192 hrs were 14.2 and 7.7 mg-L , respectively
(Table 1).
Propanil was tested at six duplicated control and exposure concentrations
(means of 1.9, 4.0, 8.4, 11.1, and 15.7 mg-L" 1- The mean water temperature
was 25.3°C with a range from 24.1 to 26.6°C. LC5Q values for the pooled
replicates at 96 and 192 hrs were 8.6 and 3.4 mg-L , respectively (Table 1).
Acute mortality curves for fathead minnows exposed to the five test
herbicides are presented in Appendix B (Figs. B 1-5). The 9555 confidence inter-
vals for the two replicates were generally overlapping.
Early Life-Stage Flow-Through Toxicity Tests. Early life-stage tests were
conducted for all herbicides by incubating and hatching fathead minnow eggs
in the exposure water, and continuously exposing the fry and juvenile fish to
herbicide concentrations for 54 to 60 days. Herbicide concentrations were
corrected for recoveries of the test compounds from water. Chemical character-
istics of the test water during the early life-stage exposures are presented
in Appendix A.
Alachlor
Fathead minnow eggs (95 to 111 total per incubation chamber) less than
24 hrs old were exposed to alachlor at six duplicated control and exposure
concentrations (means of 0.06, 0.14, 0.26, 0.52, and 1.10 mg-L"'). The mean
water temperature was 25.3°C with a range from 23.0 to 28.6°C. Hatching
success for controls averaged 72.8%, and ranged from 68.3 to 76.4% for the five
alachlor exposure concentrations (Table 2). Hatching success was not affected
31
-------
TABLE 2. HATCHABILITY, ABNORMAL DEVELOPMENT, SURVIVAL AND GROWTH OF FATHEAD
MINNOWS (Plmephales promelas) EXPOSED TO ALACHLOR FOR 64 DAYS
Mean Alachlor Concentration ± s.d.
Parameter
Mean percent hatch3
Mean percent abnormal
and dead"
Mean number of survivors
at 60 days post-hatch0
cj Mean wet weight at
ro 60 days post-hatch (g)
Mean total length at
60 days post-hatch (mm)
0.0 t 0.0
(n>20)
72.8
2.6
24.5
0.482
37.1
0,06 t 0.02
(n-37)
73.3
4.0
26.5
0.432
36.2
0.14 t 0.04
(n»39)
68.3
3.7
28.0
0.419*
35.6*.
0.26 t 0.04
(n-40)
76.2
2.0
28.0
0.403
35.0
(mg-L'1)
0.52 t 0.08
(n=40)
76.4
0.6
29.5
,0.423
35.9
1.10 t 0.18
(n=40)
73.4
0.0
25.0
0.316**
31.9**
Live fry/total eggs after 6 days.
Abnormal (deformed) + dead fry/total fry at time of transfer from egg cups (6 days
after Initial exposure of eggs).
Based on 30 fry transferred to duplicate exposure chambers.
* Significantly different from controls (p<0.05).
**Sign1f1cantly different from controls (p<0.01).
-------
by alachlor exposure. The mean1 percent of fry that were dead or abnormal at
the time of transfer from egg incubation chambers to the aquaria proper was
2.6% for controls, and ranged from 0 to 4.0* for the five alachlor exposures.
None of the alachlor exposures resulted in a significant increase in abnormal
newly hatched fry (p>0.05). Survival of transferred fry through 60 days of
exposure averaged 24.5 individuals out of 30 for controls, and from 25.0 to
29.5 for the five alachlor exposure concentrations. None of the alachlor
exposures resulted tn a significant decrease in fry survival through 60 days
(p>0.05).
Mean wet weight for controls at 60 days was 0.482 g. and ranged from
0.316 to 0.432 g for the five alachlor exposure levels CTable 2). Differences
in amounts of light incident upon the alachlor exposure aquaria due to non-
uniform room lighting were noticed late in the exposure. Relative light
values were obtained with a photometer at the water surface of each aquarium
and ranged from 15.6 to 36.3. Multiple linear regression analysis indicated a
significant (p
-------
concentration for fathead minnows exposed to alachlor was considered to fall
between 0.52 and 1.0 mg-L .
Bromaci1
Fathead minnow eggs (92-106 total per incubation chamber) less than 24
hrs old were exposed to bromacil at six duplicated control and exposure con-
centrations (means of 1.0, 1.9, 4.4, 12-0, and 29.0 mg-L" ). The mean water
temperature was 24.7°C with a range from 22.9 to 27.2°C. Hatching success for
controls averaged 66.OX, and ranged from 58.4 to 72.3% for the five exposure
levels (Table 3). Hatching success was not significantly affected (p>0-05) by
bromacil. The percent of fry at time of transfer that were dead or abnormal
averaged 5.9 for controls and ranged from 6.9 to 11.52 for the five exposure
levels. None of the bromacil test concentrations resulted in a significant
increase (p>0.05) in abnormal (including dead) newly hatched fry. Percent
survival of transferred fry (60 total) at 60 days post-hatch averaged 88.3%
for controls and ranged from 88.3 to 98.32 for the five bromacil exposure
concentrations. Survival at 60 days was not significantly affected (p>0.05)
by bromacil at these concentrations.
Weight was significantly reduced (p<0.05) from the mean control wet
weight of 0.479 g at all five bromacil exposures (Table 3); and with the
exception of the second lowest exposure (1.9 mg-L" ), weight was significantly
reduced at the 99% confidence level by the remaining bromacil exposures.
Length was significantly reduced (p<0-05) from the mean control length of 29.8
mm at all bromacil concentrations with the exception of the second lowest
exposure 0-9 mg-L ). Length was significantly reduced at the 99% confidence
level by the three highest bromacil exposures (4.4, 12.0, and 29.0 mg-L },
but not by the two lowest exposures (1.0 and 1.9 mg-L" ).
34
-------
TABLE 3. HATCHABILITY, ABNORMAL DEVELOPMENT. SURVIVAL AND GROWTH OF FATHEAD
MINNOWS (Plmephales promelas) EXPOSED TO BROMACIL FOR 64 DAYS
Mean Bromadl Concentration ± s.d. (mg-L" )
0.0 ± 0.0 1.0 tO.4 1.9±0.2 4.4 ± 0.5 12.0 t 1.0) 29.0+2.1
Parameter (n=34) (n=36) (n=36) (n-36) (n=33) (n=36)
Mean percent hatcha
Mean percent abnormal
and dead"
Mean number of survivors
at 60 days post-hatch0
Mean wet weight at
60 days post-hatch (g)
Mean standard length at
60 days post-hatch (mm)
66.0
5.9
26.5
0.479
29.8
68.8
6.9
28.5
**
0.410
28.4*
58.4
7.4
29.5
0.420*
28.7
58.5
8.3
29.5
0.381**
**
27.9
60.4
8.4
27.5
0.374**
**
27.7
72.3
11.5
26.5
**
0.326
**
26.4
Live fry/total eggs after five days.
Abnormal (deformed) + dead fry/total fry at time of transfer from egg cups (5 days
after Initial exposure of eggs).
c
Based on 30 fry transferred to duplicate exposure chambers.
* Significantly different from controls (p<0.05).
^Significantly different from controls (p<0.01).
-------
A "no-effect" concentration for bromacil was not determined from the
test. The fish showed an adverse response to bromacil at all tested concen-
trations using the parameter of weight at 60 days post-hatch. Only one
exposure concentration 0-9 mg-L ) did not show an adverse response using the
parameter of length at 60 days post-hatch. It can only be concluded from this
test that the "no effect" concentration for fathead minnows exposed to bromacil
as eggs and for 60 days post-hatch is less than 1.0 mg-L" .
Dinoseb
Fathead minnow eggs (96 to 109 total per incubation chamber) less than
24 hrs old were exposed to dinoseb at six duplicated control and exposure
concentrations (means of 0.4, 1.7, 4.3. 14.5, and 48.5 yg-l ). The mean water
temperature was 25.3°C with a range from 23.5 to 26.8°C. Hatching success for
controls averaged 74.92, and ranged from 73.5 to 86.3% for the five exposure
levels (Table 4). Hatching success was not affected by the dinoseb concen-
trations. The mean percent of fry that were abnormal (.including dead) at time
of transfer was 3.2% for controls, and ranged from 0.6 to 6.0% for the five
levels of dinoseb exposure. None of the dinoseb exposures resulted in a
significant increase (p>0.05) in abnormal newly hatched fry. Survival of
transferred fry (30 per exposure chamber) through 60 days of exposure averaged
26.5 for controls and ranged from 2.5 to 26.0 for the five exposure levels.
The dinoseb exposure replicates at 48.5 ug-L averaged only two survivors,
which was significantly less (p<0.01) than the survival of controls.
Wet weight at 60 days averaged 0.599 g for controls, and ranged from
0.517 to 0.733 g for the five exposure levels of dinoseb. The highest
exposure (48.5 yg-L ) caused a significant decrease (p<0.01) in weight.
Increased growth occurred at all other exposures, and at 1.7 yg.L a
36
-------
TABLE 4. HATCHABILITY, ABNORMAL DEVELOPMENT, SURVIVAL AND GROWTH OF FATHEAD
MINNOWS (Pimephales promelas) EXPOSED TO DINOSEB FOR 64 DAYS
Mean Dlnoseb Concentration t s.d.
Parameter
Mean percent hatch3
Mean percent abnormal
and deadb
Mean number of survivors
at 60 days post-hatch0
" Mean wet weight at
60 days post-hatch (g)
Mean total length at
60 days post-hatch (mm)
0.0 t 0.0
(n=19)
74.9
3.2
26.5
0.599
39.7
0.4 ± 0.2
(n=38)
85.0
0.6
23.5
0.683
41.3
1.7 ± 1.2
(n=39)
86.3
3.4
24.5
0.733**
42.2*
4.3 t 0.6
(n=39)
80.0
3.5
26.0
0.647
39.5
(M9-L-1)
14.5 ±2.6
(n=39)
73.5
6.0
16.0
0.679
40.5
48.5 ± 7.8
(n=39)
79-7
1.5
**
2.5
**
0.517
35.2
a Live fry/total eggs.
Abnormal (deformed) + dead fry/total fry at time of transfer from egg cups (5 days
after Initial exposure of eggs).
c Based on 30 fry transferred to duplicate exposure chambers.
Significantly different from controls (p<0.01).
u
Significantly longer than controls (p<0.05 from two-tailed Dunnett's test).
jBM
Significantly heavier than controls (p<0.01 from two-tailed Dunnett's test).
-------
significant increase tp<0-01) in weight occurred. Mean length at 60 days was
39.7 mm for controls, and ranged from 35-2-42.2 mm for dinoseb-exposed fish.
Dinoseb exposure did not significantly reduce (p>0.05) fish length, but at
1.7v3'L a significant length increase (p<0.05) was observed.
The "no effect" concentration for fathead minnows exposed to dinoseb was
between 14.5 and 48.5 yg-L . This was based on the two most sensitive para-
meters of fry survival and wet weight at 60 days post-hatch.
Pi uron
Fathead minnow eggs (83 to 104 total per incubation chamber) less than
24 hrs old were exposed to dluron at six duplicated control and exposure
concentrations (.means of 2.6, 6.1, 14.5, 33.4, and 78.0 pg-L ). The mean
water temperature was 25.0°C with a range from 23.8 to 27.0°C. Hatching
success for controls averaged 67.9%, and ranged from 66.1 to 77.92 for the
five exposure levels (Table 51. Hatching success was not affected by diuron
exposure. The mean percent of abnormal (including dead) fry at time of
transfer was 2.2% for controls, and ranged from 0.6 to 15.0% for the five
dluron exposures. The highest concentration of 78.0 ug-L resulted in a
significant increase (p<0.01) in percent abnormal fry at hatch. Survival of
transferred fry (30 per aquarium) averaged 24 for controls, and ranged from
8 to 28 for the five dluron exposures. The highest concentration (78.0pg'L )
resulted in a significant reduction (p<0.05) in survival through 60 days of
exposure after hatching.
Wet weight at 60 days averaged 0.568 g for controls, and ranged from
0.496 g to 0.568 g for the five dluron exposures. Fry length at 60 days
averaged 32.2 mm for controls, and ranged from 29.1 to 32.4 mm for the five
exposures. Neither wet weight nor length were significantly affected (p>0.05)
38
-------
TABLE 5. HATCHABILITY, ABNORMAL DEVELOPMENT. SURVIVAL AND GROWTH OF FATHEAD
MINNOWS (Plmephales promelas) EXPOSED TO DIURON FOR 64 DAYS
-1
Mean Dluron Concentration t s.d. (ug-L )
Parameter
Mean percent hatch3
Mean percent abnormal
and deadb
Mean number of survivors
at 60 days post-hatch0
<£ Mean wet weight at
60 days post-hatch (g)
Mean total length at
60 days post-hatch (mm)
0.0 i 0.0
(n-18)
67.9
2.2
24.5
0.568
32.2
2.6 ± 0.7
(n»38)
77.9
0.6
26.5
0.568
32.3
6.1 i 1.6
(n=36)
75.0
1.3
28.0
0.563
32.4
14.5 ± 2.0
(n-37)
71.8
3.7
21.5
0.619
32.3
33.4 ± 4.8
(n-38)
67.9
8.2
22.5
0.563
31.0
78.0 ± 8.1
(n=38)
66.1
**
15.0
it
7.5
0.496
29-1
a
Live fry/total eggs after 5 days.
Abnormal (deformed) + dead fry/total fry at time of transfer from egg cups (5 days
after Initial exposure of eggs).
c
Based on 30 fry transferred to duplicate exposure chambers.
* Significantly different from controls (p<0.05).
**S1gn1f1cantly different from controls (p<0.01).
-------
The "no effect" concentration for fathead minnows exposed to diuron was
between 33.4 and 78.0 pg-L" . This estimate was based on the two parameters
that were significantly affected at 78.0 u9'L~ . These were abnormal fry at
time of transfer and survival through 60 days.
Propani1
Fathead minnow eggs (98 to 104 total per incubation chamber) less than
24 hrs old were exposed to propanil at six duplicated control and exposure
concentrations (means of 0.4, 0.6, 1.2, 2.4, and 3.8 yg-L ). The mean water
temperature was 25.3°C with a range from 23.6 to 27.2°C. Hatching success
averaged 75.9% for controls, and ranged from 56.6 to 80.52 for the five
propanil exposures (Table 6). The highest exposure (3.8 ug-L ) resulted in
a 56.62 hatch, which was significantly lower (p<0.05) than the control hatch.
Percent abnormal (plus dead) fry at the time of transfer from egg cups was
3.0% for controls, and ranged from 5.5 to 65-8% for the five propanil ex-
posures. The concentration of 3.8ug-L resulted in a significant in-
crease (p<0.01) in percent abnormal fry. The mean percent of surviving fry
at 54 days post-hatch was 93.4% for controls, and ranged from 0.0 to 72.5%
for the five propanil exposure levels. Exposures of 1.2yg-L and above
resulted In a significant decrease (p<0.01) in survival. Survivorship curves
for the five exposures are presented in Fig. 1.
Wet weight of fry at 54 days post-hatch averaged 0.590 g for controls,
and ranged from 0.448 to 0.558 g for the three lowest propanil exposures. The
higher two exposures (.2.4 and 3.8 yg-L ) had no survivors at 54 days. Wet
weight was not affected at the three lower propanil exposures. Since many of
the fish appeared swollen, the fish were oven-dried at 50°C for 24 hrs follow-
ing wet weight measurements to determine if an edematous condition existed.
40
-------
TABLE 6. HATCHABILITY, ABNORMAL DEVELOPMENT, SURVIVAL AND GROWTH OF FATHEAD
MINNOWS (Pimephales promelas) EXPOSED TO PROPANIL FOR 58 DAYS
Mean Propanll Concentration ts.d. (yg-L)
0.0 ± 0-0
Parameter (na!3)
Mean percent hatch* 75.9
Mean percent abnormal
and deadb 3.0
Mean percent fry survival
at 54 days post-hatch0 93.4
Mean wet weight at
54 days post-hatch (g) 0.590
Mean dry weight at
54 days post-hatch (g) 0.152
Mean total length at
54 days post-hatch (mm) 38.3
0.4 ± 0.3
(n-27)
80.5
5.5
72.5
0.558
0.132
36.7
0.6 t 0.4
(n=27)
70.2
10.4
50.0
0.491
0.119*
**
34.2
1.2 ± 0.7 2.4 t 0.8 3.8 t 0.4
(n=24) (n=>14) (n=5)
63.4 64.0 56.6*
9.4 13.5 65.8**
** ** **
16.6 0 0
0.448
0.113 — —
33.1** -
a Live fry/total eggs after 4 days of Incubation.
Abnormal (deformed) + dead fry/total eggs at time of transfer from egg cups (4 days
after Initial exposure of eggs).
c Based on mortality of 30 fish/chamber through day 30 post-hatch and 20 fish/chamber
between days 30-54.
* Significantly different from controls (p<0.05).
**S1gn1f1cantly different from controls (p<0.01).
-------
»»ntr»l
•P.
ro
10 20 30 40
0«VS POST-NATCH
90
60
Figure 1. Survival curves for fry and juvenile fathead minnows
(Plmephales promelasl exposed to propanll 1n an early
life-stage toxlclty test. Each curve represents a
combination of fish from duplicate exposure chambers
(n=60 at day 0).
-------
The mean dry weight of control fish was 0.152 g. Fish exposed to 0.6 pg'L
averaged 0.119 g, a significant weight reduction (p<0.05). Fish exposed to
1.2 yg-L propanil appeared still lighter (.0.113 g), but because of the low
number of surviving fish (jv=10}, this test did not reach significance at p<0.05.
Mean length at 54 days was 38.3 mm for control fry, and ranged from 33.1 to
36.7 mm for fry from the three exposures. Fry length was significantly reduced
(p<0.01) at propanil exposures of 0-6 and 1.2 yg-L" . Every parameter tested,
with the exception of wet weight, was adversely affected by propanil. The "no
effect" concentration based on the two most sensitive parameters Cdry weight
and length at-54 days post-hatch) was between 0.4 and 0.6 pg-L .
Bioconcentration Tests
A summary of bioconcentration potential, metabolism, and elimination of
the five herbicides in fathead minnows is presented in Table 7. The raw data
were adjusted for recoveries of C-labeled herbicides from both water and
fish. Mean recoveries from water and fish, respectively, were: alachlor-102.6
and 86.8%, bromacil-87.5 and 88.12, dinoseb-102.5 and 78.22, diuron-93.2 and
91.9%, and propanil-92.8 and 90.8%.
Alachlor
14
Thirty-day old fathead minnows were exposed to C-labeled alachlor at
mean water concentrations of 0.66 and 9-95 yg-L for 21 days. The mean water
temperature was 24.9°C with a range from 23.8 to 25.8°C. Uptake of alachlor
from water was rapid with an apparent equilibrium between water and tissue
achieved in one day (Appendix c, Fig. C-l). Mean bioconcentration factors of
C equivalents in whole fish from days 1-21 were 50.2 and 41.4 for lower and
43
-------
TABLE 7. BIOCONCENTRATION. METABOLISM AND ELIMINATION OF FIVE
TEST HERBICIDES IN FATHEAD MINNOWS (Plmephales promelas)
Herbicide
Alachlor
Bromacl 1
Dlnoseb
Dluron
Propanll
Mean H20
concentration
(pg-L-i)
0.66
9.95
4.25
35.11
0.62
7.22
3.15
30.40
0.34
5.09
14C Bio-
concentration
factor
50.2
41.4
2.8
3.5
61.5
64.1
157.0
143.7
69.0
111.3
Percent parent
compound at
end of
uptake
13.2
ndb
2.3
1.3
1.8
Percent
depuration
of T4C
In 24 hr
84.0
78.3
nd
nd
67.4
73.8
83.8
76.4
80.0
80.8
Percent depuration
of «4c at end
of test (days
of depuration in
parentheses)
96.4 (14)
99.8 (14)
nd
nd
94.6 (14)
97.9 (14)
98.7 (21)
99.0 (21)
96.4 (10)
95.4 (10)
Determined as portion of ether-extractable fraction that chromatographed by TLC as parent
compound from fish exposed to higher concentrations of test herbicides.
Not determined.
-------
higher exposures, respectively (Table 7).
Elimination of alachlor was rapid, with 84.0 and 78.3% of the mean
plateau tissue concentration lost during the first 24 hrs for fish exposed to
lower and higher concentrations, respectively. Depuration was observed for
14 days, at which time 96.4 and 99.8% had been eliminated at lower and higher
exposures, respectively. At the end of the uptake phase, 23.4% of total C
in fish tissue was extracted with ether. Of this soluble fraction, 56.5%
chromatographed by thin layer chromatography (TLC) as parent compound, or
13.2% of the total tissue 14C. UnextractaBle 14C comprised 76.5% of the total
14 14
C, Indicating a major amount of total C was apparently in a "bound" form.
Bromacil
14
Thirty-day old fathead minnows were exposed to C-labeled bromacil at
mean water concentrations of 4.25 and 35.11 wg-L" for 17 days. Mean water
temperature was 24.2°C and ranged from 23.0 to 25.5°C. Bromacil did not
appreciably bioconcentrate 1n the fish. Mean bioconcentration factors of
14
C equivalents In whole fish from days 1-17 were 2.8 and 3.5 for lower and
higher exposures, respectively (Table 7). The fish were not analyzed for
metabolites.
Dinoseb
Fathead minnows (.30-days old) were exposed to C-labeled dinoseb for
24 days at mean water concentrations of 0.62 and 7.22 ug-L • The mean water
temperature was 25-l°C with a range from 23.8 to 26.1°C. An upward trend in
14
tissue C was evident on days 21 and 24 of the uptake phase of the study
(Appendix C, Fig. C-2). Consequently, at the end of 24 days of uptake, one-
half of the fish at the higher concentration were placed into a depuration
chamber and one-half were exposed for an additional 4 days with daily
45
-------
samplings. The mean exposure level for this group over 28 days was 7.76
yg-L" . After 28 days of exposure, depuration was also followed in these
fish.
Uptake of dinoseb was rapid, with an equilibrium between water and fish
14
achieved withtn 24 hrs. Mean C btoconcentration factors for days 1-24 of
exposure were 61.5 and 64.1 for lower and higher exposures, respectively
(Table 7-). The mean C bioconcentration factor for the higher dinoseb
exposure over 28 days was 56.2.
14
Elimination of C from dinoseb-exposed ffsh was rapid, with 67.4 and
73.8% of the mean plateau tissue concentration lost within 24 hrs for fish
exposed for 24 days to lower and higher concentrations, respectively. After
14 days of depuration, 94.6 and 97.92 had been eliminated for lower and
higher exposures, respectively (Table 7). Elimination of C from fish ex-
posed to the higher concentration of dinoseb for 28 days was similarly rapid,
with 80.2 and 95.9% eliminated after 24 hrs and 10 days, respectively.
At the end of dinoseb uptake, 27.9% of the total C in fish tissue was
ether-soluble. Of this soluble fraction, 8.1% chromatographed by TLC as
parent compound, or 2.3% of the total tissue C (Table 7). Unextractable
14C comprised 72.1% of the total UC.
Pi uron
14
Fathead minnows (30 days old) were exposed to C-lafaeled diuron for 24
days at mean water concentrations of 3.15 and 30.40 ug-l . The mean water
temperature was 24.8°C with a range from 23.8 to 25.8°C. Uptake was rapid,
with an equilibrium between water and fish achieved within 24 hrs (Appendix C,
Fig. C-3). Mean bioconcentration factors of C diuron equivalents for lower
and higher exposures were 157.0 and 143.7 times, respectively (Table 7).
46
-------
14
Elimination of C from diuron-exposed fish was rapid, with 83.8 and
76.4% of the mean tissue concentration at equilibrium eliminated within 24
hrs at lower and higher exposures, respectively. After 21 days in clean water,
98.7 and 99.0% of the C had been eliminated at the lower and higher ex-
posures, respectively.
At the end of diuron uptake, 15.0% of the total C was ether-extractable
of which 8.4% chromatographed as the parent compound. This constituted 1.3%
of the total tissue 14C (Table 7). Unextractafale 14C comprised 85.0% of the
total 14C.
Propam'l
Fathead minnows (30 days old) were exposed to C labeled-propanil in
two separate tests. In the first test, fish were exposed to mean water con-
centrations of 3.96 and 53.49 ug-L for a total of 21 days. However, during
this test significant fish mortality occurred (14.5 and 16.4% in lower and
higher exposures, respectively), and the test was repeated at lower concen-
trations.
14
In the second test, fish were exposed for 17 days to C-labeled propanil
at mean water concentrations of 0.34 and 5.09 ug-L~ . The mean water tempera-
ture for the test was 24.7°C with a range from 24.0 to 26.4°C. No mortalities
or behavioral changes were observed in the test organisms. Uptake was rapid,
with an equilibrium reached within 24 hrs (Appendix C, Fig. C-4). Bioconcen-
14
tration factors of total C were 69-0 and 111.3 for lower and higher exposures,
respectively.
Elimination of C was rapid. After 24 hrs in clean water, 80.0 and
80.8% of the mean tissue concentration at equilibrium had been eliminated for
lower and higher exposures, respectively. After 10 days in clean water, 96.4
47
-------
and 95.4% had been eliminated.
14
At the end of propanil uptake, 12.4% of the total C was ether-
extractable. From TLC analysis, 14.9% of the extractable fraction chromato-
graphed as parent compound, or 1.8% of the total tissue C (Table 7). un-
extractable 14C comprised 87.6% of the total 14C.
Metabolite Characterization
Alachlor. Rainbow trout were injected with 1.0 uCi CO-16 mg) of alachlor.
Most t83-94%) of the radioactivity was recoverd in tank water, of which 60%
appeared to be parent compound. There were 3 polar metabolites of alachlor
present in the water, and 2 polar metabolites in bile. The liver extract
contained one or more components close to the TLC plate origin. No alachlor
metabolites were recovered in sufficient quantity for GC/MS characterization.
Bromacil. Three trout were injected with 1.0 wCi of labeled bromacil.
Over 95% of the C label from injected bromacil appeared in the water within
24 hrs. The remainder was In the bile 0-33.1. with traces in the blood and
liver. Of the radiolabeled material in the water, 70-80% was parent compound
as determined by TLC and GC/MS. The rest consisted of two or more polar
metabolites that did not move from the origin during TLC, and were not further
characterized. Plasma and liver extracts showed essentially the same pattern
as the water. TLC of bile extracts revealed at least two polar components
which were apparently not altered by treatment with glucuronidase. These bile
metabolites were present in quantities too small to be characterized by GC/MS.
Dinoseb. Ninety percent of the herbicide injected into trout was re-
covered in the water, of which 50% appeared to be parent compound by TLC.
There were at least 2 polar metabolites and 1 or 2 metabolites less polar than
48
-------
dinoseb In the water. The metabolite profile was similar in bile except that
the parent compound was not present. No metabolites were characterized by
GC/VIS. '
Diuron. Tank water from diuron-injected trout contained over 90% of the
radioactivity after 24 hrs. The tank water contained at least five components,
including parent diuron. A polar compound (20-25% of the total amount) failed
to move from the origin. Two additional components were present with R-
values less than or Identical to parent diuron (R^ 0.4 to 0.5). Parent diuron
represented 35-40% of the total radioactivity in the tank water. The fifth
component appeared to be less polar than diuron (Rf 0.65 to 0.7), and rep-
resented 5-10% of the total radioactivity. The bile contained 5-6 radio-
labeled derivatives of diuron.
Propanil. Each of 3 trout received 1 yd (21 ugm) of propanil by intra-
peritoneal Injection. Most of the injected dose was recovered in the water
(75-108%) within 24 hr, and significant levels of radioactivity also occurred
in the bile (9-22%). The liver contained 1-1.9% and less than 1% was re-
covered from blood fractions.
Three or more radioactive components were present in the water. The
component with the greatest mobility (40-75% of the total radioactivity on
the chromatogram) migrated with the same R^ (0.45-0-55) as propanil. A
component of lower mobility (R-=0.2) comprised 20-552 of the total radio-
activity. The least mobile of the components (R^<0.02) contained about 5%
of the total activity.
An acetone extract of liver also contained 3 radioactive components with
the same R^s as the water components. In liver, however, the component with
49
-------
the same R* as propanil was in lower proportion (30-60%), and the most polar
component in higher proportion (10%). The compound with intermediate mobility
comprised 40-50% of the total in all 3 fish.
In bile, over 95% of the radioactivity was in the form of a highly polar
compound or compounds (Rf<0.02), and 5% of the total was in a component with
an Rf of 0.2. There was no trace in the bile of unchanged propanil. The
highly polar component of bile was eluted from one chromatogram, and incubated
with glucuronidase (from Helix pomatia). This treatment resulted in the
formation of a less polar compound with an R- of 0.16. With the addition of
1,4-saccharolactone (an inhibitor of glucuronidase) in the incubation mix,
the metabolite was not altered. Thus, the highly polar bile metabolite
appeared to be a glucuronide conjugate, but the aglycone was neither propanil
nor 3,4-dichloroaniline from its TLC mobility.
For two herbicides, propanil and diuron, pre-exposure of the fish to
unlabeled herbicide resulted in a different profile of metabolites, e.g. there
were no labeled metabolites of propanil less polar than the parent compound in
fish not pre-exposed to unlabeled herbicide, but there were 2-3 less polar
labeled metabolites in fish pre-exposed to unlabeled propanil. The different
profile may be due to enzyme induction or to the establishment of pools of
metabolically active unlabeled metabolites with which the labeled metabolites
equilibrate.
GC/MS Analysis of Metabolites. No herbicide metabolites were completely
characterized by mass spectral analysis; however, some information on metabo-
lites of propanil and diuron was obtained.
Propanil. Eight possible metabolites of propanil were submitted for
50
-------
analysis, of which 2 were partially characterized. The second most polar
compound on TLC was hydroxylated in the propionic acid moiety, and was,
therefore, either 3',4'-dichloro-2-hydroxypropionanilide or 3',4'-dichloro-3-
hydroxypropionanilide. Another metabolite tentatively identified as 3,4,3',
4'-tetrachloroazobenzene (TCAB) based on mass spectral data was probably not
this compound. Authentic TCAB was synthesized by a published method (Corbett
and Holt, 1963) and was much less polar on TLC than the metabolite. The
metabolite migrated with the same R* as 3,4-dichloroaniline CDCA), and because
DCA is a known photochemical precursor of TCAB, we tentatively concluded that
the metabolite was DCA, and that TCAB was formed on the TLC plate during the
interval between chromatography and mass-spectral analysis.
Diuron. GC and MS analysis of phenylurea herbicides is complicated by
the thermal instability of these compounds. Tri-substituted ureas such as
diuron pyrolyze to an aromatic phenylisocyanate and aliphatic component.
Pyrolysis of mono- and di-substituted phenylureas results in the phenyl-
isocyanate, the aliphatic component, and the corresponding aniline (Buchert
and Lokke, 1975). We have used this information to interpret the mass spectra
of diuron metabolites.
Four bile metabolites from diuron-exposed trout were submitted for
analysis. Only the most polar was present in sufficient quantity to be
detected by the mass spectrometer. The mass spectrum of the polar metabolite
resembled that of diuron, implying that no ring alteration had occurred, nor
had the N-l position been dealkylated.
Three diuron metabolites were recovered from water, and submitted for
analysis. The spectrum of the most polar metabolite suggested that demethyla-
tion of diuron had occurred. This was supported by the presence of DCA in
51
-------
the GC profile. The metabolite with intermediate polarity migrated with the
same R^ as reference diuron but its spectrum was much like the first metabolite
indicating that at least one methyl group had been lost. DCA was also present
in the GC profile of the second metabolite. The third diuron water metabolite
migrated with OCA and mass spectral data confirmed that this compound was DCA.
Bromaci1. Two possible metabolites of bromacil were recovered from
i
exposure water and analyzed. One compound, which migrated with the bromacil
standard was identified as the parent compound.
Alachlor and DinoseS. Metabolites of these compounds were analyzed but
there was insufficient material to generate mass spectra.
In Vitro Metabolism of Propam'T
In several experiments on the in vitro hepatic conversion of propanil
to metabolites there was a consistent pattern which resembled that seen in
tank water of fish injected with the herbicide. About 5% of the total radio-
activity was converted to metabolites, of which 502 was in a polar fraction
which remained at the origin during TLC, and the remainder was in a fraction
of intermediate polarity between propanil and the immobile fraction. With
multiple chromatogram development, the polar fraction was resolved into three
and possibly more components. The fraction of intermediate polarity was
resolved into two components. Occasionally, there was the suggestion of a
small band that migrated slightly faster than propanil.
Binding of Herbicides to Macromolecules
All of the herbicides exhibited time-dependent binding to macromolecules,
with alachlor binding most extensively and propanil the least (Table 3). In
52
-------
general, the binding was not dependent on NADPH, was not altered by the
addition of glutathione, and was not altered by the substitution of heat-
inactivated microsomes for active microsomes. The binding was thus apparently
not dependent on mixed-function oxidase activity and may have been due solely
to the affinity of the parent herbicides for protein and other macromolecules.
TABLE 8. BINDING OF HERBICIDES TO MACROMOLECULES
Herbicide
Alachlor
Bromactl
Dinoseb
Diuron
Propanil
Nanomoles herbicide/gin protein
75.2
3.6
13.2
6.7
2.3
Herbicide Stability in Solution
Declines in concentration with time occurred in the dinoseb and propanil
solutions. Dinoseb declined by 212 in 40 days in Lake Superior water. Pro-
panil declined curvilinear!ly during the 194 day test with an estimated half-
life of 65 days. The principal breakdown product in the propanil solution was
identified by GC/MS analysis as 3,4-dichloroaniline. Bromacil and diuron
remained stable during the 40 day test. Alachlor was not tested for its
stability in Lake Superior water.
53
-------
GC/MS Analysis of Technical Grade Propam"j_ and Dluron For Contaminants
GC/MS analysis of technical propanil used in this study revealed the
presence of 0.67 mg^g" of 3,3',4 ,4'-tetrachloroazobenzene (TCAB). No
3,3',4,4'-tetrachloroazoxybenzene (JCAOB) was present a&ove the detection
limit of 30 yg«g . No TCAB was found in technical grade dturon at a detection
limit of 10 yg-g , nor was any TCAOB present above the 30 yg«g detection
limit. GC/MS printouts of reconstructed ion chromatograms and mass chromato-
grams for standards of TCAB and TCAOB, and for samples of technical grade
propanil and diuron are presented in Appendix D (Figs. D-l to 0-3).
54
-------
SECTION VI
DISCUSSION
Acute Toxicity
Static exposures of rainbow trout and faluegills to technical grade alachlor
of 100% purity resulted in 96 hr IC™ values of 2.4 and 4.3 mg-L~ , respectively;
while exposure to a 43% liquid formulation of alachlor resulted in LC,-Q values
of 1.4 and 3.2 mg-L" , respectively (Johnson and Finley, 1980). In another
report that did not specify test conditions, 96 hr median tolerance limits
(TLm's) for rainbow trout and bluegills were 2.3 and 13.4 mg-L" , respectively,
for a 4 Ib/gal emulsifiable concentrate of alachlor (Weed Science Society of
America, 1970). The 96 hr LC5Q for fathead minnows in the present study was
5.0 mg-L"1.
One account of bromacil toxicity toward fish was found (U.S. Environmental
Protection Agency, 1975), which had originated through a personal communication
with E.I. duPont de Nemours and Co., Inc. TL values based on an 80% bromacil
formulation for bluegill sunfish were 103 and 71 mg-L" at 24 and 48 hr,
respectively. For rainbow trout TL values were 102, 75, and 38 mg-L" at 24,
48, and 72 hr, respectively; and for carp (Cyprinus carpio), 164 mg-L at both
24 and 48 hr. No information was provided on the nature of the exposures. In
the present study, LC5Q values of 185, 183, and 182 mg-L" were obtained for
fathead minnows at 24, 48, and 96 hr, respectively. Bromacil was the least
55
-------
toxic of the five herbicides studied in acute exposures.
The 24, 43, and 96 hr LC^g values for dinoseb with fathead minnows in the
present study were 0.8, 0.7, and 0.7 mg-L, respectively. Alabaster (1969)
reported 24 and 48 hr LC5Qs of 3.4 and 3.0 mg-L , respectively, for dinoseb
(90% active ingredient) and harlequin fish (Rasbora heteromorpha) under
toxicant replacement conditions in hard water (250 mg-L" as CaC03., pH 7.2,
20°C). Dinoseb-exposed mosquitofish (Gambusia affinis) in static bioassays using
dechlorinated tap water at 21°C had 24 hr LC5Q values of 0.37 and 0.96 mg-L"
in insecticide-susceptible and insecticide-resistant populations, respectively
(Fabacher and Chambers, 1974).
Lipschuetz and Cooper (1961) obtained a 24 hr LC5Q of 0.24 mg-L in a flow-
through test (pH 8.0, 21°C) with blacknose dace (Rhinichthys atratulus). The
same authors found pH to greatly affect the toxicity of dinoseb to rainbow
trout. For rainbow trout tested at 18.3°C, 24 hr LC5Q values of 0.073 (pH 6.9)
and 0.30 (pH 8.0) mg-L" were obtained. Similarly, Woodward (1976) found de-
creased pH resulted in increased toxicity of dinoseb to cutthroat-(Salmo clarki)-
and lake trout (Salvelinus namaycush). Dinoseb toxicity also increased with
increased water temperature and water hardness, but not as greatly as with
decreased pH. LC5Q values (96 hr) of 1.35, 0.13, and 0.041 mg-L" were re-
ported for cutthroat trout at pHs of 8.5, 7.5 and 6.5, respectively, at 10°C
in standard reconstituted soft water. Lake trout IC^ values were 1.40, 0.77,
and 0.32 mg-L"1 at pHs of 8.5, 7.5, and 6.5, respectively. Woodward (1976)
stated that ionization of dinoseb (a weak acid with pKa = 4.4) at higher pHs
could decrease its ability to be transported across the gill, and therefore
reduce its toxicity. This interpretation of the effect of pH on toxicity is
supported by other reports (Maren e_t a_]_., 1968; Hunn and Allen, 1974) which
56
-------
state that ionized forms of molecules penetrate gill membranes less readily
than nonionized forms.
In another study with cutthroat and lake trout in which fish were exposed
statically at 10°C, 96 hr LC5Q values for dinoseb were 0.067 and 0.044 mg-L'1,
respectively (Johnson and Finley, 1980). A temperature increase from 10° to
15°C resulted in no substantial change in toxicity to either species. A de-
crease in temperature from 10° to 5°C did not affect the toxicity td cutthroat
trout, but LCcnS were 3 times higher for lake trout. Increased pH again de-
creased the toxicity of dinoseb. Aging of test solutions caused a twofold
increase in LC5Q values after 1 week, but no change after 4 weeks.
A 96 hr dinoseb lethal threshold concentration of 0.070 mg-L was deter-
mined for juvenile Atlantic salmon CSalmo salar) in a static test regime
(Zitko et_al_., 1976). The lethal threshold was determined by calculating the
geometric mean of the lowest concentration at which median mortality was
observed, and the highest concentration at which no mortality occurred.
Diuron LC5Q values in the present study with fathead minnows were 23.3,
19.9, and 14.2 mg-L"1 at 24, 48, and 96 hr, respectively. From a review of
diuron acute toxicity tests with fish, 96 hr LC5Q values generally ranged
between 1 and 25 mg-L for most species (Johnson and Julin, 1974). However,
the 96 hr LC5Q values ranged from 0.5 mg-L for larval striped bass (.Morone
saxatilis) (Hughes, 1973) to the solubility limit (.42 mg-L"1 at 25°C) for
other species (Bond et,al_., 1960). A reiteration of the forementioned review
with 96 hr LC5Q data is presented in Table 9. The acute toxicity of diuron was
not greatly affected by pH or hardness, but was decreased in aged solutions
(Fish-Pesticide Research Laboratory, 1974; Johnson and Finley, 1980). Increased
water temperature increased diuron toxicity to bluegills in one study (Macek et
57
-------
TABLE 9. 96 HR ACUTE TOXICITY OF DIURON TO VARIOUS FISH SPECIES
en
00
Species
Coho salmon
(Oncorhynchus
kisutch)
Lake trout
(Salvellnus
namaycush)
Rainbow trout
(Sal mo
gairdnerl)
96 hr
(mg-P"1)
1-10
2.5
2.7
2.4
2.5
2.5
1.2
3.2
3.6
11.5
3.5
4.2
7.4
9.4
Temp
Formulation ( C)
Technical
(95% a.1.) 13
10
10
10
10
5
15
10
10
10
12
12
12
12
AlkallnJty Hardness
pH (mg-L ) (mg-L ) Comments Reference
7.4
6.5
7.0
7.5
8.5
7.5
7.5
7.0
7.0
7.0
7.5
7.5
7.5
7.5
35
38
38
38
38
112
112
41
41
41
35
35
35
35
40
33
33
33
33
104
104
33
33
33
40
40
40
40
Fish-Pesticide
Research
Continuous exposure Laboratory (1974)
H II II
II II II
II II II
II II II
II II II
II II II
Solution aged 1 wk "
Solution aged 3 wks "
Solution aged 3 wks "
Solution not aged "
Solution aged 1 wk "
Solution aged 3 wks "
Solution aged 4 wks "
Fathead minnow
(Pimephales 32
promelas
brown bullhead
(Ictalurus 11
nebulosus)
Channel catfish
(Ictalurus 38
punctatus)
Blueglll
(Lepomls
macrochfrus)
10.4
80% wp
Continuous exposure, ,
non-toxic to 32 mg-L"1
Continuous exposure
Wile (1968)
Walker (1965)
80% wp
Technical
(95% a.1.)
17
12
7.5
7.5
250
35
40
Continuous exposure
Continuous exposure,
finger! 1ng fish
(0.5-1.5g)
Dayfield (1971)
Fish-Pesticide
Research
Laboratory (1974)
(Continued)
-------
TABLE 9(continued)
en
«£>
Species
Blueglll
(Lepomls
macrochlrus)
Small mouth
96 hr
/LC5p-K
(mg.C ')
8.5
25
9.1
8.9
7.6
5.9
2.1
10.0
8.3
8.0
10.2
10.4
7.0
19
bass (Mlcropterus
dolomieui)
Largemouth
42
Temp Alkalinity Hardnes.s
Formulation (°C) pH (mg-L'1) (mg-L"')
80% wp
»
Technical 7
(95% a. 1.)
13
18
24
" 30
12
12
12
12
12
" 12
80% wp
- 20
-
7.1
7.1
7,1
7.1
7.1
8.0
8.0
8.0
6.5
8.5
9.5
_
7.4-7.7
-
-
35
35
35
35
35
235
235
235
35
35
35
_
41-71
•
.-
40
40
40
40
40
44
170
300
40
40
40
_
—
bass (Mlcropterus
salmoides)
Striped
bass (Horone
saxatllis)
0.5
6.0
80% wp 21
21
-
-
-
-
-
-
Comments
Continuous exposure,
flngerllng fish;
2"-3" size
4"-5" size
Continuous exposure
ii n
u ii
u u
n u
n n
ii n
ii ii
u ii
u u
n n
Test conducted with
fry and tap water
Continuous exposure
Continuous exposure.
larval fish
Continuous exposure,
Reference
Walker (1965)
"
Macek et al . ,
(19697 ~
ii
"
"
u
F1sh Pesticide
Research
Laboratory (1974)
".
11
11
"
"
U1le (1968)
Bond et al . ,
(19607"
Hughes (1973)
"
flngerllng fish
-------
al_., 1969), but had little effect upon trout or bluegills in another study
(Johnson and Finley, 1980).
Propanil LC5Qs with fathead minnows in the present study were 11.5, 10.2,
and 8.6 mg-L" at 24, 48, and 96 hr, respectively. In the mosquitofish, static
propanil exposures produced LC5Q values of 11.21, 8.45, and 7.62 mg-L at 24,
48, and 72 hr, respectively; and in the green sunfish CLeponris cyanellusl, LC
values of 9.10, 8.26, and 5.85 mg-L at 24, 48, and 72 hr, respectively
50
(Davey et al_., 1976). They used well water of unspecified hardness or pH, and
a propanil formulation of 3 lb-gal , effective concentration. A static
exposure with a 3 lb-gal liquid formulation of propanil in aged tap water of
unspecified character yielded mosquitofish TL values of 11.30, 11.00, 10.17,
and 9.46 mg-L at 24, 48, 72, and 96 hr, respectively CChaiyarach e_t al_., 1975),
In a brief account, propanil at 20 mg-L in continuously flowing dechlorinated
tap water produced 16.6% mortality at 24 hr with young channel catfish
Clctalurus punctatus) averaging 6.8 g in weight (.Chambers and Fabacher, 1974).
A 96 hr LCgQ value of 3.8 mg-L was obtained for finger!ing channel catfish in
a static test (McCorkle et.a]_., 1977).
The order of acute toxicity to fathead minnows of the five herbicides
studied after 24 to 192 hr of exposure was: dinoseb>a1achlor>propanil>diuron>
bromacil. LC5Q concentrations ranged from 0.4 to 185 mg-L for these compounds
over this exposure period. Such concentrations would not likely be encountered
in streams or lakes from field runoff.
Early Life-Stage Toxicity
Chronic (long-term) laboratory exposures often include concentrations of
toxicants which might be encountered in the field. These exposures may also
extend over a complete life cycle or at least over -the most sensitive life-
60
-------
stages. Such tests may help predict potentially harmful ecological effects
(Mawdesley-Thomas, 1971).
Several studies with selected toxicants and fish species have demonstrated
that early life-stages are the most sensitive (Benoit, 1975; Eaton, 1974). Egg
and fry C30-60 day-old) bioassays provided good estimates of chronic toxicity
with bluegills (Lepomis macrochirus) exposed to cadmium when compared to com-
plete life-cycle exposures (Eaton, 1974). In a review of 56 life-cycle toxicity
tests with 34 chemicals and four fish species, the embryo-larval and early
juvenile life-stages were the most, or among the most, sensitive (McKim, 1977).
In 822 of the cases, the maximum acceptable toxicant concentration (MATC) esti-
mated by embryo-larval or early juvenile exposures was identical to the MATC
established by the longer, more involved, and more costly partial or complete
life-cycle toxicity tests. Additional tests with several fish species and
selected metals (copper and cadmium) substantiated the observation that embryo-
larval tests reliably estimate complete life-cycle chronic tests (Eaton et al.,
1978; McKim et al_., 1978).
In the present study, herbicide exposures extended from the fertilized egg
stage through 54-60 days post-hatch. For alachlor, the early exposure parameters
(% hatch, % abnormal and dead fry at time of transfer from egg cups) were not
significantly affected. However, length and weight of juvenile fish at 60 days
post-hatch were affected, with a highly significant reduction (p<0.01) in
weight and length at the highest exposure of 1.10 mg-L" (Table 2). Fish weight
and length parameters were also reduced (p<0-05) at the concentration of 0.14
mg-L . However, since there were no significant effects at the next two higher
exposures of 0.26 and 0.52 mg-L , the "no-effect" estimate was considered to
be between 0.52 and 1.10 mg-L . Concentrations in this range would not likely
61
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be encountered from field run-off into streams and lakes.
Length and weight at 60 days post-hatch were the parameters significantly
affected by bromacil (Table 3). Weight was reduced (p<0.05) at all exposures
ranging between 1.0 and 29.0 mg-L , while length was significantly reduced
(p<0.05) at all exposures except 1.9 mg-L . The "no effect" concentration was
below the lowest test concentration of 1.0 mg-L . However, in view of the low
order of the acute toxicity of bromacil to fish and the "clean" nature of the
diluter tubes leading to the exposure chambers (i.e., lack of algal growth),
it was felt that the reductions in length and weight of bromacil-exposed fish
may have been due partly to its algacidal properties in the exposure system.
Young fathead minnows feed upon algae under natural conditions (Eddy and Under-
hill, 1976). Thus, although all exposure levels as well as the controls were
fed equal volumes of brine shrimp, the controls may have had the additional
advantage of an imported algal community as a dietary supplement.
Dinoseb did not affect egg hatch nor the incidence of abnormalities or
mortalities among newly hatch fathead minnow fry (Table 4). Dinoseb at a con-
centration of 10.0 yg-L caused a 24% decrease in the survival of lake trout
(Salvelinus namaycush) alevins through the yolk absorption stage (Woodward,
1976); and reduced fry survival over the period from yolk absorption to 60
days post-hatch to 22% compared to 86% for controls. In the present study, fat-
head minnow fry survival at 60 days post-hatch was not significantly affected by
dinoseb exposures up to 14.5 yg-L ; however, the 14.5 yg-L~ exposure decreased
the number of survivors to a mean of 16.0 out of 30 individuals, compared to
26.5 for controls. At 48.5 yg-L of dinoseb, the number of survivors was de-
creased to a mean of 2.5 individuals out of 30, which was significant (p<0.01).
Length of fathead minnows through 60 days post-hatch was not significantly
62
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reduced by any of the dinoseb exposures; however, weight was significantly
reduced tP<0-01) at the highest exposure of 48.5 vg-L. Weight and length of
lake trout fry at 60 days post-hatch were significantly reduced (p<0.05) by
dinoseb exposures of 0.5 ug-L and greater (Woodward, 1976). The rate of yolk
absorption of lake trout fry exposed to dinoseb also decreased at the same
exposure levels.
The "no-effect" concentration of dinoseb for lake trout was less than the
lowest exposure level of 0.5 ug'L . based on the parameters of length and
weight (.Woodward, 1976). The "no-effect" concentration of dinoseb for fathead
minnows in our study was between 14.5 and 48.5 pg-L , based upon the parameters
of fry survival and fry weight through 60 days post-hatch. These "no-effect"
levels are lower than the time-independent IC™ values (TILC 1 Of 54 and 102
yg-L for lake and cutthroat trout, respectively (Johnson and Finley, 1980).
values are mathematically derived toxicant concentrations at which 50%
of the test animals would be expected to survive indefinitely upon exposure to
toxicants under flow-through conditions for up to 30 days. However, TILC5Q
estimates would not encompass exposure to the early life-stages of the embryo
and newly hatched fry. A cumulative toxicity index (96 hr LC50/TILC50) of 1.5
was reported for both cutthroat and lake trout, indicating little cumulative
toxic action of dinoseb (Johnson and Finley, 1980).
Dinoseb has a water solubility of approximately 50 mg-L (.Agriculture
Canada, 1973, Weed Society of America, 1970). While no information was found on
watershed concentrations of dinoseb, several studies have been conducted on the
herbicide atrazine, which has a water solubility of 33 mg-L" (Wauchope, 1978),
slightly less than that of dinoseb. A comparison of the two herbicides may be
useful, sfnce dinoseb was degraded quite slowly in water in our study and in other
63
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toxicity studies (Woodward, 1978; Johnson and Finley, 1980), although atrazlne
is more persistent in soil (Klingman and Ashton, 1975). Atrazine residues were
commonly detected in Iowa surface waters (Richard e_t al_., 1975), with concen-
trations as high, as 42 ug-L occurring in June. The overall June mean from
three different surface waters was 4.9 ug-L" . Atrazine residues in a South-
western Ontario watershed also reached their highest concentrations in June,
with a mean monthly concentration of 3.6 wg-L (Roberts e_t al_., 1979).
The highest concentrations of atrazine residues in surface waters were re-
lated to rainfall events, the severity of such events, and the proximity of
these events to herbicide application (Richard et al_., 1975; Roberts et al.,
19791. In the case of dinoseb, with a slightly greater water solubility than
atrazine, it may be possible for surface water residue loads to develop similar
to those for atrazine, dependent upon variables such as application timing and
rate, nature of the watershed and portions treated, rainfall timing, and rain-
fall severity.
If, in certain areas, the surface water dinoseb concentrations did,
indeed, approximate the residue concentrations of atrazine (i.e. 1-5 ug-L ,
with occasional heavier loadings), dinoseb concentrations might periodically
reach levels above 14.5 ug-L and adversely affect those aquatic organisms
with sensitivities to dinoseb similar to or greater than that of fathead
minnows. Such water concentrations of dinoseb would certainly appear to have
the potential to cause adverse effects upon more sensitive aquatic forms such
as the salmonids, where the MATC was below 0.5 yg-L" for lake trout (Woodward,
1976). Studies are needed on surface water concentrations of dinoseb,
especially from selected areas where it is extensively used.
Diuron affected two parameters in the fathead minnow embryo-larval test
64
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at the highest exposure concentration of 78.0 ug-L" . There was an increased
incidence (p<0.01) in percent abnormal and dead fry at the time of transfer
from egg cups to exposure chambers, and a decrease (pO-05) in number of sur-
vivors (mean of 7.5) after 60 days of exposure. Based on these parameters, the
"no-effect" concentration was between 33.4 and 78.0 ug-L . Flow-through tests
produced TILC5Q values of 140 and 500 ug-L for rainbow and cutthroat trout,
respectively (Johnson and Flnley, 1980). As with dinoseb, TILC5Q estimates were
higher than our "no-effect" range for fathead minnows. Cumulative toxicity
indices were 12.3 and 3.7 for rainbow and cutthroat trout, respectively, in-
dicating a moderate degree of cumulative toxic action in rainbow trout (Johnson
and Finley, 1980}.
The growth of bluegill finger!ings in ponds treated with one application of
diuron at concentrations ranging from 0.5 to 3.0 mg-L was less than that of
controls over a 6-month period (McCraren e_t al_., 1969). This may have been
partly due to oxygen stress, as low oxygen levels persisted for several days
following diuron application with the resultant destruction of vascular plants.
Diuron has a water solubility of 42 mg-L (Wauchope, 1978; Weed Sci- Soc. of
Amer., 1970). Diuron, like atrazine, is an herbicide which provides full season
weed control (Klingman and Ashton, 1975). However, it is more persistent in
soil than atrazine, with a half-life of 7.0 months at 15°C as compared to 6-0
months for atrazine (Freed and Haque, 1973). In one study (Khan et_al_., 1976),
phytotoxic diuron residues were detected in the soil 3 years after the last of
a 7 year series of applications. Annual rates of diuron breakdown range from
30% to approximately 80% (.Leonard et_a]_., 1974; Dawson et al_., 1968; Khan et. .
al., 1976). Dissipation is mainly dependent upon microbial activity and not
on chemical degradation (Majka and Lavy, 1977; Hill et_ aJL, 1955; McComvick and
65
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Hiltbold, 1966; Murray ejtal_., 1969; Sheets, 1964). Diuron has relatively low
downward mobility in soil (Majka and Lavy, 1977; Khan e_t al_., 1976), allowing
for the compound to remain near the soil surface.
Diuron regularly appeared in streams and in sediments of streams and
estuaries in Hawaii, where such streams received runoff from pineapple and sugar-
cane fields (Green et_al_., 19771. Dturon concentrations in estuarine waters
ranged from 0.1 to 1 ng-L" . In another field plot study, less than 0.12% of
the annual application of diuron to cotton was lost as runoff (.Willis et al.,
19751. Their data suggested that with proper use on agricultural land in the
lower Mississippi River Valley, diuron posed little threat to adjacent aquatic
areas.
Oiuron residues were detectable in water for 21 days in ponds treated at
rates of 0.5 and 3.0 mg-L , and in plants and mud for 95 and 122 days,
respectively (McCraren et_aK, 19691. The highest water residue concentration
of 35 yg-L was reported on day 21 in a pond treated with 3.0 mg«L diuron.
More information is needed on water concentrations of diuron from agricultural
areas where it is extensively used to determine its potential for adversely
affecting aquatic life at low exposure concentrations.
Propanil adversely affected all parameters studied, with the exception
of juvenile fish wet weight at 54 days, at mean concentrations of 0.6 ug'L
and above in the present investigation. Dry weight of the survivors was
J'L'
-1
significantly reduced at a concentration of 0.6 yg-L . The "no-effect"
estimate for propanil was between 0.4 and 0.6
The observation that wet weight was not significantly affected by propanil
exposure while dry weight was at 0.6 ug'L , in combination with observations
of grossly swollen fry in propanil-exposed groups suggested that this herbicide
66
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caused a water imbalance in the fish. Physiological studies to determine if
and how propanil affects excretion or osmotic regulation would be of interest.
In some of the swollen fry reddish patches were visible through the skin
in the lateral abdominal areas, which appeared to have been due to localized
hemorrhaging or abnormal condition of the blood vessels. In a 60-day pond
exposure to propanil at a single application rate of 7 kg-ha" , propanil caused
a hemolytic effect in immature ides or roaches starting on day 10 tPopova,
19731- Livers of propanil-exposed ides on day 10 had overfilled blood vessels,
many with heomolyzed erythrocytes; and had focal blood discharges and dis-
aggregation of liver cells.
In the field study by Popova 0973), immature ides were placed into control
and exposure ponds one week after propanil application. Fish growth was con-
siderably less in the treated pond after 60 days, as live weight gains of 63
and 142 were noted in the control and treatment ponds, respectively. Fish loss
was 20%. Morphophysiological characteristics of propanil-exposed fish were
changed, and did not return to normal even though fish were maintained for a
long time in water free of propanil and its metabolites.
The solubility of propanil in water has been reported as 200 mg-L
CGordon et a]_., 1964) and 500 mg-L CWeed Society of America, 1970; Bailey and
White, 1965). Propanil has been found to dissipate rapidly from rice paddy
water. When propanil was applied to a rice field at the normal application
rate of 3.4 kg-ha" in an impounded system, propanil concentrations of
approximately 680 and 60 ug-L were measured at the time of flooding and 24 hr
later, respectively; while in a flowing system, propanil concentrations were
approximately 400 and 50 ug-L at 0 and 24 hrs, respectively (Deuel et al.,
1977). Thus, propanil concentrations in the water were reduced by approximately
67
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90% in one day. The dissipation of propanil in the flood water corresponded to
the appearance of 3,4-dichloroaniline (Deuel e_t al_., 1977}. The principal break-
down product in Lake Superior water from our study was also identified as
3,4-dichloroaniline.
In another rice field study at a single propanil application of 1 kg-ha
(Popova, 1973), the propanil concentration in water was 200 yg-L after 1 day
and approximately 50 yg-L for the subsequent 5 days. At 10-15 days, only
metabolites were observed in the water. On days 20 and 60 no herbicide was
observed in water, bottom sediments, or flora.
This study has shown that low levels of propanil adversely affect fish
hatchability, the normal development and survival of newly hatched fry, and the
growth and survival of juvenile fish.. Even though propanil has been shown to
dissipate quite rapidly under field conditions (Deuel etal_., 1977; Popova,
1973}, residues in the water during the first week after application may be
sufficiently high to adversely affect fish populations if rice paddy flood
water is inhabited by fish or if it passes over fish spawning and nursery
grounds.
Several other fish chronic or early life-stage toxicity studies with
herbicides have been reported. "No-effect" ranges for fathead minnows exposed
to trifluralin and acrolein were 1.95 - 5.1 yg-L and 11.4 - 41.7 yg-L ,
respectively (Macek et a]_., 1976}. Picloram affected lake trout CSalvelinus
namaycush) fry growth at the lowest concentration tested (35 yg-L ), thus
placing the "no-effect" range below this value (Woodward, 1976). Macek et al.
(1976) determined a "no-effect" range for atrazine-exposed bluegills (Lepomis-
macrochirus) of 90 - 500 yg'L based on equilibrium loss, and for atrazine-
exposed fathead minnows of 210 - 87Q yg-L based on fry mortality. A "no-
68
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effect" range of 300 - 1500 yg-L was obtained in a study with 2,4-D
(butoxyethanol ester} and fathead minnows (Mount and Stephan, 1967).
All five herbicides studied apparently inhibit plant photosynthesis
either by affecting the electron transport process within the chloroplasts or
by acting as inhibitory uncouplers of photophosphorylation in both chloroplasts
and mitochondria (Moreland, 1980). In addition to these primary mechanisms of
action within the chloroplasts and mitochondria, some of the herbicides studied
are also known to produce secondary effects in plants (Moreland, 1980). Bro-
macil.dinoseb, and diuron interfere with cell membrane permeability and integrity.
Bromacil and diuron also interfere with cell membrane lipid synthesis. Dinoseb
inhibits DMA, RNA or protein synthesis.
Growth reduction of herbicide-exposed fish may have resulted from some of
these same mechanisms of action having occurred within the cells of the fish.
Moreland (.1980) stated that plant and animal mitochondria are usually affected
similarly by inhibitors and uncouplers, but with different degrees of sensitiv-
ities.
Uptake. Metabolism and Elimination
The five herbicides of the present study did not appreciably accumulate
14
in fish tissue. Quantities of total C were present in fish tissue at con-
centrations slightly greater than water concentrations for alachlor (45.8 x),
dinoseb (j52.8 x), diuron 050.4 x) and propanil (90.2 x); while for bromacil
14
the total C concentrations were similar in both fish tissue and water (total
C bioconcentration factor of 3.2 x). The percentages of total C present
in fish tissue determined to be parent herbicide at the end of the uptake phase
were 13.2, 2.3, 1.3, and 1.82 for alachlor, dinoseb, diuron, and propanil,
69
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respectively. This was not determined for bromacil due to the low level of
14
total C activity. From the contributions of the parent herbicides toward
14
total C activity in fish tissue, parent herbicide bioconcentration factors
were: alachlor 05-0 x), bromacil (<3.2 x), dinoseb (J-4 x), diuron C2.0 x), and
propanil 0-6 x).
Most herbicides that have Been studied do not have a strong potential for
tissue accumulation in fish. Negligible bioconcentration factors were obtained
for 2,4-D dimethyl ami ne salt CSikka et al_-» 1977), atrazine tMacek et al.,
1976)., and simazine (Rodgers, 1970). However, trifluralin bioconcentrated in
fish approximately 1,000 times the water exposure level CMacek et.a]_., 1976).
By contrast, many of the organochlorine pesticides or their metabolites (p,p'-
DDT, p,p'-DDE, chlordane, hexachlorobenzene, mirex, dieldrin, toxaphene,
heptachlor, heptachlor epoxide, methoxychlor, and endrin) were found to
accumulate in fish from 4,500 to 51,000 times their concentrations in water
CVeith et al_., 1979; Hamelink et al.., 1971; Jarvinen and Tyo, 1978).
Alachlor appeared to be readily eliminated as parent compound and also
quite readily metabolized by rainbow trout, although the metabolites were not
recovered in sufficient quantity for GC/MS characterization. In a model eco-
system study CYu et_al_., 1975), alachlor was degraded to numerous unidentified
polar metabolites.
Bromacil was rapidly eliminated from rainbow trout, mainly as parent
compound. Two or more polar metabolites were produced, but were present in
insufficient quantity to be characterized by GC/MS. Gardiner et aj_. (1969)
found the primary urinary metabolite of bromacil in the rat to be a conjugate-
of 5-bromo-3-sec-butyl-6-hydroxymethyl uracil, and that bromacil was also
hydroxylated at the 2 and 3 positions of the sec-butyl side chain. Some
70
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debrominated metabolites were also found.
Dinoseb was readily eliminated as parent compound and also readily metab-
olized by rainbow trout. At least two polar metabolites and one or two metab-
olites less polar than dinoseb were present in the water of dinoseb-injected
trout. A similar metabolite profile was observed in bile, except that no
parent compound was present. However, no metabolites were characterized by
GC/MS. Rats and mice readily metabolized dinoseb to a variety of products
(Bandal and Casida, 1972). Glucuronic acid and sulfate were conjugated to the
phenolic hydroxyl, and metabolites with either of the side chain methyl groups
oxidized to carboxyls were found. The rat was able to reduce one of the nitro
groups and to acetylate the resulting amino group. The mouse was unable to
reduce either of the nitro groups. More than a dozen minor metabolites were
present but were not characterized in that study CBandal and Casida, 1972).
Rabbits also produced a propionic acid metabolite and a glucuronide containing
a reduced nitro group (.Ernst, 1967).
Diuron was readily eliminated both as parent compound and as metabolites
14
in rainbow trout injected with C-labeled diuron. A metabolite present in
the bile in sufficient quantity to be partially characterized by GC/MS had a
mass spectrum resembling that of diuron, implying that neither ring alteration
nor dealkylation in the N-l position had occurred. Metabolites recovered from
the water included 3,4-dichloroaniline and two different but unidentified
demethylated products.
In man, diuron ingestion produced urinary metabolites identified as 1-
C3,4-dichlorophenyl)-3-methyl urea, l-C3,4-d1chlorophenyl) urea, and 3,4-
dichloroaniline (Geldmacher et. al_., 1971). Bohme and Ernst (1965) found N-
C3,4-dichlorophenyl)-urea and N-C2-hydroxy-4,5-dich1orophenyl) urea to be the
71
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predominant urinary metabolites in rats fed diuron. Dogs fed diuron on a
chronic basis produced demethylated urinary metabolites of N-(3,4-dichloro-
phenyl)-urea and N-(3,4-dichlorophenyl)-N'-methyl urea, as well as some 3,4-
dichloroaniline and 3,4-dichlorophenol CHodge etaj.., 1967).
Propanil was readily metabolized by rainbow trout in the present study,
forming at least ten metabolites. It was also readily eliminated as parent
compound. One metabolite recovered from trout bile was identified by GC/MS as
either 3',4'-d1chloro-2-hydroxypropionanilide or 3',4'-d1chloro-3-hydroxy-
propionanilide. In mammals, propanil has been reported to be hydro!yzed to
3,4-dichloroaniline by an acylamidase enzyme (Williams and Jacobson, 1966}.
Other mammalian studies tend to confirm this (Singleton and Murphy, 1973; Chow
and Murphy, 1975). However, 1n the present study, 3,4-dichloroaniline forma-
tion from propanil could not be consistently demonstrated.
Herbicide Stability in Lake Superior Water
Dinoseb concentrations in Lake Superior water declined by 212 in 40 days
1n the present study. In a study of dlnoseb toxicity to cutthroat trout
(Woodward, 1976), a solution of the chemical in test water that was aged for
4 weeks was slightly less toxic than an unaged solution. The 96 hr LC5Q of
the aged dlnoseb solution was 87 yg-L" , as compared to 71 ug-L for the
unaged solution. This would represent a decreased toxicity of about 22% in
the aged solution. However, in both the present study and in the study by
Woodward 0976). microbial metabolism may not have been as important as it
would be in actual field situations. Lake Superior is a cold oligotrophic
body of water, and Woodward used standard reconstituted deionized water.
Dinoseb is readily photolysable in aqueous solution in the laboratory
72
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when irradiated by ultraviolet light with an absorption maximum at 375 nm wave-
length, which includes that part of the solar ultraviolet that reaches the
earth (.Crosby and Li, 1969). Thus, photolysis may be important in the dissipa-
tion of dinoseb in the field.
Propanil had an estimated half-life of 65 days in the present study, with
the principal breakdown product identified as 3,4-dichloroaniline. The
dissipation of propanil in Lake Superior water was much slower than its dissipa-
tion 1n actual field studies (Deuel e_t al_., 1977; Popova, 1973), possibly due
to differences in the richness of the microbial communities.
Pure chemical hydrolysis is considered a minor route of propanil degrada-
tion as sterile buffered solutions were stable for more than 4 months (E1-D1b
and Aly, 1976a). Propanil is readily degraded by the microbial community in
soil (Kearney et_a]_., 1970), and 3,4-dichloroaniline has been identified as a
microbial transformation product in soil and water (El-Oib and Aly, 1976b;
Oeuel e_t al_., 1977; Sharabi and Bordeleau, 1969). Although propanil dissipated
rapidly in field studies (Deuel elaL, 1977; Popova, 1973), El-D1b and Aly
(1976b) found that mixed microflora from river water and sewage did not utilize
propanil over a 4 month period and that active degradation occurred only after
the addition of a heavy inoculum of Bacillus cereus.
Bromacil and diuron remained stable in Lake Superior water in the present
study. Bromacil declined in concentration by 6% over 40 days, while diuron
concentrations changed by less than 1% over a 36 day period. Alachlor was not
studied for Its stability in water.
Under intense ultraviolet radiation in the laboratory, a 1 mg-L solution
of bromadl was completely decomposed in 10 minutes (Kearney e_t aj_., 1969).
Under conditions of natural summer sunlight for 4 months and laboratory irradia-
73
-------
tion which, simulated natural sunlight conditions, bromacil was photo decomposed
very slowly (Jtoilanen and Crosby, 1974), indicating that photo decomposition
probably makes only a minor contribution to the environmental disappearance of
bromacil, unless in the presence of natural photosensitizers. Indeed, the
addition of certain photosensitizers to bromactl solutions has since been shown
to result In photodecomposition by sunlight (Acher and Saltzman, 1980).
Diuron concentrations remained stable for more than 4 months 1n a labora-
tory study of hydrolysis rate conducted at 20°C and at pH values of 5, 7, and
9 CEI-Dib and Aly, 1976a). They stated that in the pU range of most natural
surface waters (pH 6 to 9), the phenyl amide herbicides (.including diuron) would
be expected to persist for long periods, and that pure chemical hydrolysis in
the aquatic environment would be a minor route of degradation. Field and labora-
tory studies Indicate that diuron 1s moderately persistent in the aquatic
environment.
Alachlor was rapidly degraded 1n a model ecosystem study CYu et al.,
1975). Of the total radioactivity present 1n water after 33 days, only 1.8X
was parent alachlor with the remainder consisting of at least 8 degradation
products. However, information from field studies 1s needed to determine if
alachlor 1s also rapidly degraded under actual field conditions.
Tetrachloroazobenzene (.TCAB) as Contaminant of Propanil
TCAB has been identified as a contaminant of commercial propanil at con-
centrations up to 2.9 mg-g tBunce e_t al_. , 1979), and in diuron at concen-
trations up to 28 yg-g"1 (.Sundstrom et al... 1978; Hill et il. , 1981). TCAOB
has been reported as present in diuron formulations at concentrations of 1-2
yg-g CSundstrom et. a]_. , 1978). TCAB was present in the technical propanil
used at a concentration of 0.67 mg-g . No TCAOB was detected. Neither
74
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contaminant was detected in the diuron used here.
Both TCAB and TCAOB are formed during the production of 3,4-dichloroaniline
or during its further conversion to herbicides (Poland et al_., 1976). TCAB
residues have been detected in soils treated with propanil (Poland et.al_.,
19761. These residues may have arisen either from the commercial formulation
itself (Bunce e_t al_., 1979} or as microbial transformation products (JJartha
et.al_., 1968; Chisaka and Kearney, 1970; Bourdeleau et.al_., 1972; Lay and
Ilnicki, 1974}. TCAOB has also been found in soil as a microbial degradation
product of propanil (Still and Herrett, 1976; Kaufman e_t al_., 1972).
TCAB is approximately Isosteric to 2,3,7,8-tetrachlorodibenzo-p_-dioxin
and 2,3,7,8-tetrachlorodiBenzofuran, two extremely potent toxins and teratogens
(Hsia and Kreamer, 1979a; Hsia e_t aJL, 1977}. All three compounds were potent
inducers of hepatic aryl hydrocarbon hydroxylase (AHH) activity (Poland et al.»
1976}. TCAB was found to be weakly mutagenic by the Salmonella microsome test
and to be toxic to mammalian cells in vitro (hsia e_t a]_., 1977). TCAB induced
unscheduled DMA synthesis in rat hepatocytes in vitro (Hsia and Kreamer,
1979b), and was considered a potential carcinogen. Hsia and Kreamer (1979a)
found that the potency of TCAB and TCAOB as microsomal enzyme inducers in in_
vivo exposures with rats surpassed that of mirex and kepone, which were studied
by Kaminsky et al_. 0978} under comparable conditions. TCAOB was a more potent
inducer of cytochrome P-448 than TCAB (Hsia and Kreamer, 1979a). TCAB and
TCAOB exposures in rats have resulted in liver hypertrophy, histopathological
alterations of hepatocytes, and induction of mitogenesls (Schrankel et al.,
1980).
The significance to aquatic organisms of the presence of trace amounts
of TCAB and TCAOB in water is not known. Deuel e£ al_., (1977) found only
75
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trace quantities of TCAB approximating the detection limit of 0.2 ug-L in
rice paddy water 24 hrs after flooding in 1973, and none at longer time inter-
vals after flooding. No TCAB was detected in water samples collected up to
24 hrs after flooding in 1974 or 1975. Teratogenic effects were observed in
fathead minnow fry of the present study at 3.8 ug-L of propanil, but whether
these effects were due to propanil itself, to the TCAB contaminant, or to a
combination of these compounds is not known. TCAB would have been present at
an extremely low concentration at a propanil exposure of 3.8 pg-L" ,
approximating a concentration in the water of 2-5 ng-L . However, single
injections of 10 ng TCAB fnto chicken eggs resulted in severe edema in the
embryos (Schrankel ejt al., 1982).
76
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REFERENCES
Acher, A.J. and S. Saltzman. 1980. Dye-sensitized photooxidation of bromacil
in water. J. Environ. Qua!. 9:190-194.
Agriculture Canada. 1973. Guide to the Chemicals Used in Crop Protection
(publ. 1093, 6th ed.}. Information Canada, Ottawa.
Alabaster, J.S. 1969. Survival of fish tn 164 herbicides, insecticides,
fungicides, wetting agents, and miscellaneous substances. Int. Pest. Control
11:29-35.
American Public Health Association. 1975. Standard Methods for the Examination
of Water and Wastewater, 14th ed. American Public Health Association, Washing-
ton, D.C.
Bailey, G.W. and J.L. White. 1965. Herbicides: a compilation of their physical,
chemical, and biological properties. Residue Rev. 10:97-122.
Bandal, S.K. and J.E. Casfda. 1972. Metabolism and photoalteration of 2-sec-
butyl-4,6-dinitrophenol (DNBP herbicide) and its isopropyl carbonate derivative
CDinobuton acaricide}. J. Agrfc. Food Chem. 20:1235-1245.
Bartha, R., H.A.B. Linke, and D. Pramer. 1968. Pesticide transformations:
production of chloroazobenzenes from chloroanilines. Science 161:582-583.
Benoit, D.A. 1975. Chronic effects of copper on survival, growth, and
reproduction of the bluegill (.Lepomis macrochirus). Trans. Amer. Fish. Soc.
104:353-358.
Bohme, L. and W. Ernst. 1965. Metabolism of urea herbicides in the rat.
II. Oiuron and afalon. Food Cosmet. Toxicol. 3:797-802.
Bond, C.E., R.H. Lewis, and J.L. Fryer. 1960. Toxicity of various herbicidal
materials to fishes. Transactions of the 1959 Seminar on Biological Problems
in Water Pollution. The Robert A. Taft Sanitary Engineering Center, Cincinnati,
Ohio. Technical Report. No. W60-3.
Bordeleau, L.M., J.D, Rosen, and R. Bartha. 1972. Herbicide-derived
chloroazobenzene residues: pathway of formation. J. Agr. Food Chem. 20:573-
578.
77
-------
Buchert, A. and H. Lokke. 1975. Gas chromatographic-mass spectrophotometric
identification of phenylurea herbicides after N-methylation. J. Chromatogr,
115:682-686.
Bunce, N.J., C.T. Corke, R.L. Merrick, and H.H. Bright. 1979. 3,3',4,4'-
Tetrachloroazobenzene as a contaminant in commercial propanil. Chemosphere
5:283-284.
Chaiyarach, S., V. Ratananun, and R.C. Harrel. 1975. Acute toxicity of the
insecticides toxaphene and carbaryl and the herbicides propanil and molinate
to four species of aquatic organisms. Bull. Environ. Contarn. Toxicol. 14:281-
284,
Chambers, H. and D.L. Fabacher. 1974. Acute toxicity of pesticides to channel
catfish fingerlings. Miss. Farm Res. 37:1.
Chisaka, H. and P.C. Kearney. 1970. Metabolism of propanil in soils. J. Agr.
Food Chem. 18:854-858.
Chow, A.Y.K. and S.D. Murphy. 1975. Propanil (3f4-dichloropropionanilide)-
induced methemoglobin formation in relation to its metabolism in vitro.
Toxicol. Appl. Pharmacol. 33:14-20.
Corbett, J.F. and P.P. Holt. 1963. Dehalogenation during the reduction of
halogennitroarenes with lithium aluminum hydride. J. Chem. Soc. 1963: 2385-
2387.
Crafts, A.S. 1945. A new herbicide, 2,4-dinitro,secondary butyl phenol.
Science 101:417-418.
Crafts, A.S. 1975. Modern Weed Control. Univ. of Cal. Press, Berkely.
Crosby, D.G. and M-Y Li. 1969. Herbicide photodecomposition. pp. 321-363
In: Degradation of Herbicides (P.C. Kearney and D.D. Kaufman, eds.). Marcel
Dekker, New York.
Davey, R.B., M.Y. Meisch, and F.L. Carter. 1976. Toxicity of five rice-
field pesticides to the mosquitofish, Gambusia affinis, and green sunfish,
Lepomis cyanellus, under laboratory and fteld conditions in Arkansas.
Environ. Entomol. 5:1053-1056.
Dawson, J.H., V.F. Bruns, and H.J. Clore. 1968. Residual monuron, diuron,
and simazine in a vineyard soil. Weed Sci. 16:63-65.
Dayfield, R.H. 1971. The toxicity of diuron to selected aquatic organisms.
Master's Thesis, Dept. of Biol. Sciences, Marshall Univ.
Deuel, I.E., Jr., K.W. Brown, F.C. Turner, D.G. Westfall, and J.D. Price.
1977. Persistence of propanil, DCA, and TCAB in soil and water under flooded
rice culture. J. Environ. Qual. 6:127-132.
78
-------
Eaton, J.G. 1974. Chronic cadmium toxicity to the bluegill (J-epomis
macrochirus Rafinesque). Trans. Amer. Fish Soc. 103:729-735.
Eaton, O.G., J.M. McKim, and G.W. Holcomfae. 1978. Metal toxicity to embryos
and larvae of seven freshwater fish species—I. Cadmium. Bull. Environ. Contam.
Toxicol. 19:95-103.
Eddy, S. and J.C. Underhill. 1976. Northern Fishes with Special Reference to
the Upper Mississippi Valley, 3rd ed. Univ. of Minn. Press, Minneapolis.
El-Oib, M.A. and O.A. Aly. 1976a. Persistence of some phenylamide pesticides
in the aquatic environment-I. Hydrolysis. Water Res. 10:1047-1050.
El-Uib, M.A. and O.A. Aly. 1976b. Persistence of some phenylamide pesticides
in the aquatic environment. III. Biological degradation. Water Res. 10:1055-
1059.
Emst, W. 1967. Der stoffwechsel von pesticiden in saugetieren. Residue Rev.
18:132-157.
Fabacher, D.L. and H. Chambers. 1974. Resistance of herbicides in insecticide
resistant mosquitofish, Gambusia affints. Environ. Letters 1:15-20.
Farrington, D.S., R.G. Hopkins, and J.A.H. Ruztcka. 1977. Determination of
residues of substituted herbicides in grain, soil and river water by use of
liquid chromatography. Analyst 102:377-381.
Fish-Pesticide Research Laboratory. 1974. Unpublished toxicity data. Fish-
Pesticide Research Laboratory, U.S. Fisft and Wildlife Service, Columbia,
Missouri.
Freed, V.H. and R. Haque. 1973. Adsorption, movement and distribution of
pesticides in soils, pp. 441-459 In: Pesticide Formulations (VI. Van
Valkenburg, ed,). Marcel Dekker, New York.
Gardiner, J.A., R.W. Reiser, and H. Sherman. 1969. Identification of the
metabolites of bromacil in rat urine. J. Agric. Food Chem. 17:967-973.
Geldmacher, V., M. Mallinckrodt, and F. Shussler. 1971. Toxicity of diuron
1 -(3,4-dichlorophenyl)-3,3-dimethyl urea and its metabolism in man. Arch.
Toxicol. 27:187-192.
Gordon, C.F., A.L. Wolfe, and L.D. Haines. 1964. Stam. Ch. 24 In: Analytical
Methods for Pesticides, Plant Growth Regulators and Food Additives. Vol. IV.
Herbicides CG. Zweig.ed.)- Academic Press, New York.
Green, R.E., K.P. Goswanrf, M. Mukhtar, and H.Y. Young. 1977. Herbicides from
cropped watersheds in stream and estuarine sediments in Hawaii. J. Environ.
Qua!. 6:145-154.
79
-------
Hamelink, J.L., R.C. Waybrant, and R.C. Ball. 1971. A proposal: exchange
equilibria control the degree chlorinated hydrocarbons are biologically
magnified in lentic environments. Trans. Amer. Fish Soc. 100:207-214.
Hill, G.D., J-W. McGahen, H.M. Baker, D.W. Finnerty, and C.W. Bingeman. 1955.
The fate of substituted urea herbicides in agricultural soils. Agron. 0.
47:93-104.
Hill, R.H., Jr., Z.R. Rollen, R-D. Kimbrough, D.P. Groce, and L.L. Needham.
1981. Tetrachloroazobenzene in 3,4-dichloroaniline and its herbicidal
derivatives: propanil, dluron, linuron, and neburon. Arch. Environ. Health
36:11-14.
Hodge, H.C., W.L. Downs* B.S. Fanner, D.W. Smith, E.A. Maynard, J.W. Clayton,
Jr. and R.C. Rhodes. 1967. Oral toxicity and metabolism of diuron (N-3,4-
dich1orophenyl)-N',N'-dimethyl urea in rats and dogs. Food Cosmet. Toxicol.
5:513-531.
Hsia, M.T.S., F.V.Z. Bairstow, L.C.T. Sfiih, J.G. Pounds, and J.R. Allen. 1977.
3,4,3',4'-tetrachloroazobenzene: a potential environmental threat. Res.
Commun. Chem. Pathol. Pharmacol. 17:225-231.
Hsia, M.T.S. and C.F. Burant. 1979. Preparation and spectral analysis of
3,3' ,4,4'-tetrachloroazobenzene and the corresponding azoxy and hydrazo
analogs. J. Assoc. Off. Anal. Chem. 62:746-750.
Hsia, M.T.S. and B.L. Kreamer. 1979a. Induction of hepatic microsomal
cytochrome. P-448 by 3,3' ,4,4'-tetrachloroazobenzene and the corresponding
azoxy and hydrazo analogs. Res. Commun. Chem. Pathol. Pharmacol. 25:319-331.
Hsia, M.T.S. and B.L. Kreamer. 1979b. Induction of unscheduled DNA synthesis
in suspensions of rat hepatocytes by an environmental toxicant, 3,3',4,4'-
tetrachloroazobenzene. Cancer Letters 6:207-212.
Hughes, J-S- 1973. Acute toxicity of thirty chemicals to striped bass
(Morone saxatilis). Presented at the Western Association of State Game and
Fish Commissioners, Salt Lake City, Utah.
Hunn, J.B. and J.L. Allen. 1974. Movement of drugs across the gills of
fishes. Ann. Rev. Pharmacol. 14:47-55-
Jarvinen, A.W. and R.M. Tyo. 1978. Toxicity to fathead minnows of endrin
in food and water. Arch. Environ. Contam. Toxciol. 7:409-421.
Johnson, W.W. and M.T. Finley. 1980. Handbook of acute toxicity of chemicals
to fish and invertebrates. U.S. Dept. of the Interior, Fish and Wildlife
Service Resource Publ. 137. Superintent of Documents, U.S. Gov't. Printing .
Office, Washington, D.C.
80
-------
Johnson, W.W. and A.M. Julin. 1974. A review of the literature on the use of
diuron in fisheries. National Technical Information Service, U.S. Dept. of
Commerce CPB-235 446).
Kaminsky, L.S., L.J. Piper, D.N. McMartin, and M.J. Fasco. 1978. Induction of
hepatic microsomal cytochrome P-450 fay mirex and kepone. Tox. Appl. Pharmacol.
43:327-338.
Kaufman, D.D., J.R. Plinmer, J. Iwan, and U.I. Klingebiel. 1972. 3,4',4,4'-
Tetrachloroazobenzene from 3,4-dichloroaniline tn microbial culture. J. Agr.
Food Chem. 20:916-919.
Kearney, P.C., R.J. Smith, Jr., J.R. Plimmer, and F-S. Guardia. 1970. Propanil
and TCAB residues in rice soils. Weed Sci. 18:464-465.
Kearney, P.C., E.A. Wool son, J.R. Plimmer, and A.R. Isensee. 1969. Decontamina-
tion of pesticides in soils. Residue Rev. 29:137-149.
Khan, S.U., P.B. Marriage, and W.J. Saidak. 1976. Persistence and movement of
diuron and 3,4-dichloroaniline in an orchard soil. Weed Sci. 24:583-586.
Klingman, G.C. and F.M. Ashton. 1975. Weed Science: Principles and Practice.
Wiley-Interscience, New York.
Lay, M.M. and R.D. Ilnicki. 1974. Peroxfdase activity and propanil degradation
in soil. Weed Res. 14:111-113.
Leonard, O.A., W.B. McHenry, and L.A. Ltder. 1974. Herbicide residues in soil
of the vine row 21 months following 9 successive annual applications. Proc.
Ann. Calif. Weed Conf. 26:115-122.
Lipschuetz, M. and A.L. Cooper. 1961. Toxicity of 2-secondary-fautyl-4,6-
dinitrophenol to blacknose dace and rainbow trout. N.Y. Fish and Game J. 8:
110-121.
Macek, K.J., C. Hutchinson, and 0-B. Cope. 1969. The effects of temperature
on the susceptibility of bluegills and rainbow trout to selected pesticides.
Bull. Environ. Contam. Toxicol. 4:174-183.
Macek, K.J., M.A. Lindberg, S. Sauter, K.S. Buxton, and P.A. Costa. 1976.
Toxicity of four pesticides to water fleas and fathead minnows. U.S. Environ.
Prot. Agency Ecol. Res. Series Publ. No. EPA-600/3-76-099- Environ. Res. Lab.-
Duluth, U.S. Environ. Prot. Agency, Duluth, MN.
Majka, J.T. and T.L. Lavy. 1977. Adsorption, mobility, and degradation of
cyanazine and diuron in soils. Weed Sci. 25:401-406.
Maren, T.H., R. Embry, and L.E. Broder. 1968. The excretion of drugs across
the gill of the dogfish, Squalus acanthi as. Comp. Biochem. Physiol. 26:853-
864.
81
-------
Mawdesley-Thomas, I.E. 1971. Toxic Chemicals—the risk to fish. New
Scientist Jan. 14 issue:74-75.
McCorkle, P.M., J.E. Chambers, and J.D. Yarforough. 1977. Acute toxicities of
selected herbicides to fingerling channel catfish, Ictalurus punctatus. Bull.
Environ. Contam. Toxicol. 18:267-270.
McCormick, L.L. and A.E. Hiltbold. 1966. Microbial decomposition of atrazine
and diuron in soil. Weed Sci. 14:77-81.
McCraren, J.P., O.B. Cope, and L. Eller. 1969. Some chronic effects of
diuron on bluegills. Weed Sci. 17:497-504.
McKim, J.M. 1977. Evaluation of tests with early life stages of fish for
predicting long-term toxicity. J. Fish. Res. Board Can. 34:1148-1154.
McKim, J.M., J.G. Eaton, and G.W. Hoicombe. 1978. Metal toxicity to embryos
and larvae of eight species of freshwater fish—II. Copper. Bull. Environ.
Contam. Toxicol. 19:608-616.
Moilanen, K.W. and D.G. Crosby. 1974. The photodecomposition of bromacil.
Arch. Environ. Contam. Toxicol. 2:3-8.
Moreland, D.E. 1980. Mechanisms of action of herbicides. Ann. Rev. Plant
Physio!. 31:597-638.
Mount, D.I. and C.E. Stephan. 1967. A method for establishing acceptable
toxicant limits for ftsh—malathion and the butoxyethanol ester of 2,4-D.
Trans. Amer. Fish. Soc. 96:185-193.
Mount, O.I. and W.A. Brungs. 1967. A simplified dosing apparatus for fish
toxicology studies. Water Res. 1:21-29.
Murray, D.S., L.R. Rieck, and J.W. Lynd. 1969. Mtcrobial degradation of
five substituted urea herbicides. Weed Sci. 17:52-55.
Poland, A., E. Gover, A.S. Kende, M. DeCamp, and C.M. Giandomenico. 1976.
3,4,3',4'-Tetrach1oroazoxyfaenzene and azobenzene: potent inducers of aryl
hydrocarbon hydroxylase. Science 194:627-630.
Popova, G. 1973. Change in the morphophysiological indices of some fish
caused by Stam F-34. Eksp. Vod. Toksikol. 4:38-49 (Russian Translation).
Richard, J.J., G.A. Junk, M.J. Avery, N.L. Nehring, J.S. Fritz, and H.J.
Svec. 1975. Analysis of various Iowa waters for selected pesticies:
atrazine, DDE, and dieldrin-1974. Pestic. Monit- J. 9:119-123.
Roberts, G.C., G.J. Sirons, R. Frank, and H.E. Collins. 1979. Triazine
residues in a watershed in southwestern Ontario (1973-1975). J. Great Lakes
Res. 5:246-255.
32
-------
Rodgers, C.A. 1970- Uptake and elimination of simazine by green sunfish
(Lepomis cyanellus R). Weed Science 18:134-136.
Schrankel, K.R., M.T.S. Hsia, and J.G. Pounds. 1980. Hepatocellular
pathotoxicology of 3,3',4,4'-tetrachloroazofaenzene in the rodent. I. In vivo
studies. Res. Comm. Chem. Pathol. Pharmacol. 28:527-540.
Schrankel, K.R., B.L. Kreatner, and M.T.S. Hsia. 1982. Emfaryotoxicity of
3,3',4,4'-tetrachloroazobenzene and 3,3',4,4'-tetrachloroazoxybenzene in the
chick embryo. Arch. Environ. Contain. Toxicol. 11:195-202.
Sharabi, N.E. and L.M. Bordeleau. 1969. Biochemical decomposition of the
herbicide karsil and related compounds. Appl. Microbiol. 18:369-375.
Sheets, R.S. 1964. Review of disappearance of substituted urea herbicides
from soil. J. Agrfc. Food Chem. 12:30-33.
Sikka, H.C., H.T. Appleton, and E.O. Gangstad. 1977. Uptake and metabolism
of dimethy1amine salt of 2,4-dichlorophenoxyacettc acid by fish. J. Agr.
Food Chem. 25:1030-1033.
Singleton, S.D. and S.D. Murphy. 1973. Propanil C3,4-dichloropropionan1lide)-
induced methemoglobin formation 1n mice in relation to acylamidase activity.
Toxicol. Appl. Pharmacol. 25:20-29.
Smith, R.J. 1965. Propanil and mixtures with propanil for weed control 1n
rice. Weeds 13:236-238.
Still, G.G. and R.A. Herrett. 1976. Methylcarbamates, carbamates, and
acyIan11 ides. pp. 609-664 In: Herbicides: Chemistry, Degradation, and Mode
of Action, Vol. 2 (P-C. Kearney and O.D. Kaufman, eds.}. Marcel Dekker, New
York.
Sundstrom, G., B. Jansson, and L. Renfaerg. 1978. Determination of the toxic
impurities 3,3',4,4'-tetrachloroazobenzene and 3,3',4,4'-tetrachloroazoxy-
benzene in commercial diuron, linuron, and 3,4-dichloroaniline samples.
Chemosphere 12:973-979.
U.S. Environmental Protection Agency. 1974. Herbicide Report: Chemistry
and Analysis, Environmental Effects, Agricultural and Other Applied Uses.
Publ. no. EPA-SAB-74-001. Science Advisory Board, U.S. Environ. Prot. Agency,
Washington, D.C.
U.S. Environmental Protection Agency. 1975. Initial Scientific and Mini-
economic Review of Bromacil. Publ. no. EPA-540/1-75-006. Office of Pesticide
Programs, Criteria and Evaluation Div., U.S. Environ. Prot. Agency, Washington,
D.C.
33
-------
U.S. Environmental Protection Agency. 1977. Manual of Analytical Methods
for the Analysis of Pesticide Residues in Human and Environmental Samples.
J.F. Thompson, ed., U.S. Environ. Prot. Agency Health Effects Research
Laboratory, Environmental Toxicology Division, Research Triangle Park, N.C.
Veith, G.D. and V.M. Comstock. 1975. Apparatus for continuously saturating
water with hydrophobic organic compounds. J. Fish. Res. Bd. Can. 32:1849-1851.
Veith, G.D., D.L. Defoe, and B.V. Bergstedt. 1979. Measuring and estimating
the bioconcentration factor of chemicals in fish. J. Fish. Res. Board Can.
36:1040-1048.
Walker, C.R. 1965. Diuron, fenuron, monuron, neburon, and TCA mixtures as
aquatic herbicides in fish habitats. Weeds 13:297-301.
Wauchope, R.D. 1978. The pesticide content of surface water draining from
agricultural fields - a review. J. Environ. Qua!. 7:459-472.
Weed Science Society of America. 1970. Herbicide handbook of the Weed
Science Society of America. The W.F. Humphrey Press, Inc., Geneva, New York.
368 pp.
Wile, I. 1968. Aquatic plant and algae control with diuron. Research Rep.,
Biol. Branch, Ontario Water Resources Comm., Toronto, Ontario, Can.
Williams, C.H. and K.H. Jacobson. 1966. An acylamidase in mammalian liver
hydrolyzing the herbicide 3,4-dichloropropionanilide. Toxicol. Appl.
Pharmacol. 9:495-500.
Willis, G.H., R.L. Rogers, and L.M. Southwick. 1975. Losses of diuron,
linuron, fenac, and trifluralin in surface drainage water. J. Environ.
Qua!. 4:399-402.
Woodward, D.F. 1976. Toxicity of the herbicides dinoseb and picloram to
cutthroat CSalmo clarki) and lake trout CSalvelinus namaycush). J. Fish. Res.
Board Can. 33:1671-1676.
Yu, C-C., G.M. Booth, D.J. Hansen, and J.R. Larsen. 1975. Fate of alachlor
and propachlor in a model ecosystem. J. Agric. Food Chem. 23:877-879.
Zitko, V., D.W. McLeese, W.G. Carson, and H.E. Welch. 1976. Toxicity of
alkyldinitrophenols to some aquatic organisms. Bull. Environ. Contam. Toxicol.
16:508-575.
84
-------
CO
in
APPENDIX A
CHEMICAL CHARACTERISTICS OF THE TOXICITY TEST WATER.
TABLE A-l. MEANS, STANDARD DEVIATIONS AND RANGES OF DISSOLVED OXYGEN
CONCENTRATIONS (PERCENT OF SATURATION) IN CONTROL AND EXPOSURE CHAMBERS
FOR FATHEAD MINNOWS (Plmephales promelas) EXPOSED TO FIVE HERBICIDES
IN ACUTE AND
EARLY LIFE-STAGE
TESTS
Acute Tests
Herbicide
Alachlor
Bromaci 1
Dinoseb
Dluron
Propanl 1
Control
Exposures
Control
Exposures
Control
Exposures
Control
Exposures
Control
Exposures
N
3
5
3
6
3
6
2
2
3
5
Mean
98.6
98.0
96.3
96.1
96.1
95.8
90.1
92.0
91.6
90.0
± S.D.
± 0.4
± 0.6
± 0.2
t 0.4
t 3.2
t 2.3
± 2.1
t 3.5
t 0.5
± 1.4
Range
97.3-98
98.3-99
96.1-96
95.5-96
92.8-99
92.0-98
88.6-91
89.6-94
91.2-92
88.1-91
.5
.3
.5
.6
.2
.9
.6
.5
.2
.3
N
11
2
5
11
5
6
5
6
7
2
Early
Mean i
79.3 ±
76.4 ±
88.8 t
83.7 i
90.3 ±
87.3 t
90.7 ±
91.9 t
79.8 ±
73.7 t
Life-Stage
S.D.
14.9
10.6
5.7
11.7
6.3
3.0
1.6
1.1
9.6
28.9
44
68
81
52
82
82
88
89
66
53
Tests
Range
.2-95.7
.8-83.3
.9-95.4
.2-92.5
.1-99.5
.2-91.4
.3-92.6
.9-92.9
.9-92.1
.2-94.1
-------
TABLE A-2. SOME WATER QUALITY CHARACTERISTICS DURING ACUTE EXPOSURES OF
FATHEAD MINNOWS (Plmephales promelas) TO HERBICIDES.
VALUES ARE MEANS ± S.D.
Herbicide
Alachlor
Uroiaacl 1
Dlnoseb
Oluron
Propanl 1
N
4
2
3
3
3
PH
Nd*
7.4 ± 0.0
7.6 t 0.3
7.5 ± 0.1
7.5 ± 0.0
Total
alkalinity
(mg/1 as CaC03)
40.9 t 0.3
42.1 t 0.3
45.6 t 0.4
42.1 t 2.0
42.1 ± 0.1
Total
hardness
(mg/1 as CaC03)
43.7 ± 1.0
49.4 t 0.3
49.0 t 0.9
46.4 t 2.2
47.2 t 1.4
Total
acidity
(mg/1 as CaC03)
2.0 t 0.4
2.0 ± 0.0
2.0 t 0.0
2.0 t 0.0
2.0 t 0.6
No determinations.
-------
00
TABLE A-3. SOME WATER QUALITY CHARACTERISTICS DURING EARLY LIFE-STAGE
EXPOSURES OF FATHEAD MINNOWS (Plmephales promelas) TO HERBICIDES.
VALUES ARE MEANS t S.D.
Herbicide
Alachlor
Bromacl 1
Dlnoseb
Dluron
Propanll
N
7-8
3
7
5
5-6
PH
7.4 i 0.1
7.4 ± 0.1
7.5 ± 0.1
7.5 t 0.1
7.2 ±0.4
Total
alkalinity
(mg/1 as CaC03)
40.8 t 3.9
42.0 t 0.5
45.9 t 0.4
46.9 ± 2.9
41.7 ±1.6
Total
hardness
(mg/1 as CaC03)
47.1 ± 2.0
50.6 ± 0.7
49.7 ± 1.0
48.4 ± 4.3
49.1 ± 2.0
Total
acidity
(mg/1 as CaC03)
2.0 ± 0.2
1.9 ± 0.2
1.9 ± 0.0
1.6 ± 0.4
2.0 ± 0.1
-------
APPENDIX B
MORTALITY CURVES FOR ACUTE FLOW-THROUGH TOXICITY TESTS
co
00
12.0
£ 9.6
Z
O
H
2 »•••
I-
Z
111
S 4."
O
o
o
o° a.4
TIME COAY8)
Figure B-1. LC50 concentrations for duplicate alachlor toxlclty
tests with 30-day old fathead minnows (Plmephales
promelas).
-------
• 190-
175
00
to
O
z
o
o
o
10
180
14ft
4 6
TIME (DAYS)
Figure B-2. LC5Q concentrations for duplicate bromacll
toxiclty tests with 30-day old fathead minnows
(Plmephales promelas).
-------
1.6
6 1.2
Z
O
Ul
o o.e
z
o
o
S 0.3
U
a
TIME IOAY8I
Figure B-3. LC5Q concentrations for duplicate dlnoseb
' toxTdty tests with 30-day old fathead minnows
I (Plmephales promelas).
-------
£
X
o
28.0
20.0
18.0
ui 10.0
O
z
O
O
O 8.0
to
u
TIME IOAV8I
Figure B-4. LC5o concentrations for duplicate dluron toxiclty
i tests with 30-day old fathead minnows (Plmephales
promelas).
-------
UJ
ro
20.0
16.0
O 12.0
CC
I-
ui 8.0
U
O
U
g 4-0
TIMEldiysl
Figure B-5. LC^n concentrations for duplicate propanll
toxlclty tests with 30-day old fathead minnows
(Plmephales promelas).
-------
APPENDIX C
WATER CONCENTRATIONS AND WHOLE FISH TISSUE CONCENTRATIONS
OF TOTAL 14C FROM HERBICIDE BI CONCENTRATION STUDIES
WITH THE FATHEAD MINNOW (Plmephales promelas)
4.0
UJ
-a. oi
14 21
DAYS
as
Otpurallan-
Figure
C-l.
14
concentrations of C-labeled
Log mean exposure water c
alachlor (ng-nL"1) and lo
total 14c residues (ng»g~T) during uptake and
depuration phases.
og mean (t S.D.) whole fish
~T)
-------
10
4.01
- a.o
2.0
C FUk n*l)4u*» .Hlfhtr Cip»»ur*
14 II IS
DAY*
* . 'Depuration
14,
Figure C-2. Log mean exposure water concentrations of C-labeled
dlnoseb (ng-rnL"') and log mean (t S.D.) whole fish
tissue '4C residues (ng-g"1) during uptake and
depuration phases.
-------
UD
in
• .0
- 4.0
f
m
O £ » 0
a.o
W S
§o.
O X
I4C FUh IU» !<•>••-Higher C«p»»ur*
-I 1
I4C Fl*h ••
Ei*»iur*
'
o - 0.
-2.0
HO C*»(.-HI|h»r
27
OAY8
.Upl«k«
3«
D*pur«llon-
14
Figure Cj-3. Log mean exposure water concentration of C-labeled
dturon (ng-mL"1) and log mean (t S.O.) whole fish
' total 14C residues (ng-g~1) during uptake and
I depuration phases.
!i
-------
4.0l
to
_
*""*""*"""* "
Figure C-4. Log mean exposure water concentrations of C-labeled
propanll (ng-mL"') and loo mean (+ S.D.) whole fish
total 14C residues (ng-g"T) during uptake and
depuration phases.
-------
APPENDIX D
MASS CHROMATOGRAMS AND RECONSTRUCTED ION CHROMATOGRAMS
FOR STANDARDS OF 3,3',4,4'-TETRACHLOROAZOBENZENE
(TCAB), 3,3.'4,4'-TETRACHLOROAZOXYBENZENE (TCAOB),
£-DIIODOBENZENE (DIB), AND TECHNICAL GRADES OF
PROPANIL AND DIURON
to
MS
XM
III.7-
n,
it
turn «M » tm
torn.
.'13
MIMM.
i,
Figure D-jl
Mass chromatograms and reconstructed Ion chromatogram
(RIC) for standards of 3,3',4,4'-tetrachloroazobenzene
(TCAB), 3.3',4,4'-tetrachloroazoxybenzene (TCAOB). and
p_-d11odobenzene (DIB).
-------
CO
Nt.»i
l«
nt.
1IC
I L.
n,
«tMW.
2W1IM.
."IS
in.
Figure 0-2. Mass chromatograms and reconstructed Ion chromatogram
(RIC) of technical grade propanll monitored for
a.a'.M'-tetrachloroazobenzene (TCAB) and 3,3' ,4,4'-
tetrachloroazoxybenzene (TCAOB).
-------
IO
(£>
M.2-1
MS.
Ul IIC > IUSS
H'3I>
i l
> t
torn MM BM
tiM UKli II •. 4.* IMb * t. l.« MSli V ». J
».*-
MJ
m.t-i
Ml.Jn
IIC
sim
•MM«.
n,
Figure D-3. Mass chromatograms and reconstructed ton chromatogram
(RIC) of technical grade diuron monitored for S.aM.
tetrachloroazobenzene (TCAB) and 3.3',4,4'-
tetrachloroazoxybenzene (TCAOB).
------- |