PB94160678
ffl
United States
Environmental Protection EPA/600/R-94/004
Agency January 1994
Research and
Development
AGRICULTURAL INSECTICIDE RUNOFF
EFFECTS ON ESTUARINE ORGANISMS:
CORRELATING LABORATORY AND FIELD
TOXICITY TESTS, ECOPHYSIOLOGY
BIOASSAYS, AND ECOTOXICOLOGICAL
BIOMONITORING
REPRODUCED BY:
U.S. Department ol Commerce
National Technical Information Service
Springfield, Virginia 22161
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TECHNICAL REPORT DATA
(PLEASE READ INSTRUCTIONS OH THE REVERSE BEFORE COMPLETING^
1. REPORT NO.
4. TITLE AND SUBTITLE
Agricultural Insecticide Runoff Effects on Estuarine Organisms: Correlating Laboratory
and Raid Toxicity Tests, Ecophysiology Bioassays, and Ecotoxicological Biomonitoiing
3. RECIPIENT'S ACC
PB94-160678
5. REPORT DATE
January 1994
6. PERFORMING ORGANIZATION CODE
EPA/ORD
7. AUTHOR(S)
G.I. Scott1'3, M.H. Fulton1, M.C. Crosby2, P.8. Key1, J.W. Daugomah1, J.T. Waldren1,
E.D. Strozier'. C.J. Louden3, G.T. Chandler0, T.F. Bidleman3, K.L Jackson3, T.W.
Hampton4, T. Huffman3, A. Shulz3, and M. Bradford3
8. PERFORMING ORGANIZATION REPORT NO.
9. PERFORMING ORGANIZATION NAME AND ADDRESS
'U.S. National Marine Fisheries, Southeast Fisheries Science Center, Charleston
Laboratory, Charleston, SC 29422-0607; 2U.S. National Oceanic and Atmospheric
Administration, National Ocean Survey, Office of Estuarine Sanctuaries and Reserves,
1825 Connecticut Ave., N.W. Room 714, Washington, DC 20235; University of South
Carolina, School of Public Health, Columbia, SC 29208: 'Agency for Toxic Substance,
Disease Registry, Atlanta, GA
10. PROGRAM ELEMENT NO.
11. CONTRACT/GRANT NO.
12. SPONSORING AGENCY NAME AND ADDRESS
U.S ENVIRONMENTAL PROTECTION AGENCY
ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
GULF BREEZE, R.ORIDA 32561
13, TYPE OF REPORT AND PERIOD COVERED
14. SPONSORING AGENCY CODE
15 SUPPLEMENTARY NOTES
16. ABSTRACT
This study compared in situ, field and laboratory toxidty testing results (or several Insecticides (azinphosmethyt - an organophosphate; endosulfan - an
organochlorme and fenvalerate - a synthetic pyrethroid) with ecotoxicological biomonitoring results from the macropelagic, estuarine tidal creek
community in pristine habitats and in areas receiving significant Insecticide runoff from agriculture, Field studies were conducted over a four-year
period (198546) at several coastal field sites on Wadmaiaw (Leadenwah Creek) and Johns (unnamed tidal creek near Kiawah Island) Island, coastal sea
islands located just south of Charleston, South Carolina. Results indicated that laboratory and field toxicity testing and biomonitoring methodologies
should be integrated to provide holistic environmental risk assessments for pesticides. Laboratory toxicity tests provide the initial bench mark for
estimating toxic effects. In situ, field toxicity test* provide a mechanism toSvalidate initial laboratory tests and expand their design to test differences in
formulations, lifs history stages, pulsed versus continuous dose, salinity interactions, and pesticide mixtures for more realistic estimates of effects of
field exposures. Application of this method in the environmental risk assessment for three classes of pesticides (organoehtorines-endosulfan,
pyrethroids-fenvalerate, and organophosphates-axlnphosmethyl) has been demonstrated in assessing the effects of nonpoint source agricultural runoff
on sensitive estuarine tidal creek fauna in South Carolina. Over a three year period of study, the integration of this approach has provided significant
data to assist environmental regulators trying to control recurrent problems of agricultural runoff effects in Leadenwah Creek and other areas ol the
state. Future studies should be expanded to broaden our understanding of the usefulness of this integrated approach in better assessing pesticide
runoff in other aquatic ecosystems throughout the U.S.
17. KEY WORDS AND DOCUMENT ANALYSIS
A. DESCRIPTORS
18. DISTRIBUTION STATEMENT
RELEASE TO PUBLIC
EPA Form 2220-1 (Rev. 4-77) Previous Edition
B. IDENTIFIERS/OPEN ENDED TERMS
19. SECURITY CLASS (TH/S REPORT)
UNCLASSIFIED
20. SECURITY CLASS (TH/S PAGE)
UNCLASSIFIED
C. COSATI FIELD/GROUP
21. NO. OF PAGES
314
22. PRICE
is Obsolete
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PB94-160678
EPA/600/R-94/004
January 1994
Agricultural Insecticide Runoff Effects On Estuarine
Organismsr Correlating Laboratory and Field Toxicity Tests,
Ecophysiology Bioassays, and Ecotoxicological Biomonitoring
i
'by
G.I. Scott1-3, M.H. Fulton1, MJC. Crosby2, P-B. Key1,
J.W. Daugomah1, J.T. Waldren1, E.D. Strozier1, C. J. Louden3,
G.T, Chandler3,T.F. Bidleman3, K.L Jackson3, T.W. Hampton4,
T. Huffman3, A. Shulz3, and M. Bradford3
'U.S. National Marine Fisheries
Southeast Fisheries Science Center
Charleston Laboratory
Charleston, SC 29422-0607
2U.S. National Oceanic and Atmospheric Administration
National Ocean Survey
Office of Estuarine Sanctuaries and Reserves
1825 Connecticut Avenue, N.W. Room 714
Washington, DC 20235
3University of South Carolina
School of Public Health
Columbia, SC 29208
4Agency for Toxic Substance
Disease Registry
Atlanta, GA
Pijoject CR816213
U.S. Environmental Protection Agency
Environmental Research Laboratory —
Office of Research and Development
Gulf Breeze, FL 32561
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DISCLAIMER
The information in-this document has been funded wholly or in part by the United States
Environmental Protection Agency under Cooperative Agreement CR816213 to the
University of South Carolina School of Public Health, Columbia South Carolina. Mention
of trade names or commercial products does not constitute endorsement or recommendation
for use.
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CONTENTS
EXECUTIVE SUMMARY xvi
ACKNOWLEDGEMENTS , xxiv
<
INTRODUCTION 1
MATERIALS AND METHODS ' : 6
Insecticides Studied 6
Azinphosmethyl 6
Endosulfan 8
Fenvalerate 10
Study Sites 12
Field Toxicity Tests 18
Chemical Analysis of Environmental Samples 21
Seawater Samples 21
Sediment Samples 23
Oyster, Shrimp and Fish Tissue Sampjes- 24
Quality Control 24
Oyster Field Studies, 1989-90 25
Oyster Collection and Transplantation 27
Physicochemical Measurements 28
Chemical Analyses 28
Physiological Analyses 28
Field Mortality Analyses 29
Perldnsus Marinus Analyses 30
Spat Settlement 30
StatisticaJ Analyses • Oyster Studies 31
Laboratory Toxicity Tests 31
Effects of Azinphosmethyl on Brain AChE Activity in Mummichogs . 32
Laboratory Phase 32
I
BIOMARKER STUDIES 34
Field Exposure Phase ._._ 35
Assay of AChE Activity 36
Whole Body Insecticide Residue Analysis 36
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Ecotoxicological Studies of Macropelagic Organisms: 1989 - 1990 37
Block Seining 37
Ecotoxicological Sampling Statistical Procedures 40
Water Quality Parameters 41
Push Netting 41
RESULTS ! , 43
Field Toxicity Tests 43
Daily Physicochemical Parameters 43
1989, Daily Water Quality Parameters • 43
1990, Daily Water Quality Parameters 47
Rainfall Measurements 50
1989 Study Period 50
1990 Study Period 50
Measured Insecticide Concentrations in Water Samples 56
Results for the 1989 Study Period 56
Water Samples (56); Pesticide Loadings (79); Pesticide Transport
Studies (85)
Results for the 1990 Field Study 88
Hydrolab Results for the 1989 Study Period 99
Hydrolab Results for the 1989 Study Period 99
Hydrolab Results for the 1990 Study Period 113
Survival Data for Field Toxicity Test 113
1989 Field Toxicity Test 113
Quality Assurance and Quality Control for Bioassay Organisms Used in
Field Toxicity Test during the 1989 Field Study 125
1990 Field Toxicity Tests 128
Quality Assurance and Quality Control for Bioassay Organisms Used in
Field Toxicity Tests during the 1990 Field Study 138
OYSTER ECOPHYSIOLOGY STUDIES, 1989-90 146
1989 Studies ..: ' 146
, 1990 Results 166
Discussion and Conclusions: Oyster Ecophysiology Studies 1989-90 177
Laboratory Toxicity Tests ) 178
Effects on Brain AChE Activity 178
1. Laboratory Phase • ECso Determination 178
Discussion and Conclusions 182
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Relationship Between Specific Levels of Azinphosmethyl - Induced
Brain AChE Inhibition and Sublethal Effects on Respiration,
Nitrogen Excretion and O/N Ratios 182
Biomarker Studies 185
Brain AChE in Mummichogs 185
Field Exposures 185
Field Effects on Brain AChE in Mummichogs .. 190
Discussion and Conclusions • Field Exposure Tests 194
Discussion and Conclusions 195
Sublethal Effects of Azinphosmethyl on Brain .AChE-Comparison of
Field and Laboratory Effect 195
Ecotoxicological Studies 201
Block Seining 1989-90 201
Biomass 201
P. pugio Density 206
F. heteroclitus Density 210
Total Fish Density 214
Penaied Shrimp Densities 217
Blue Crab Densities 221
Discussion and Conclusions of Ecotoxicological Studies, 1989-90 .... 224
Push Netting, 1990 .' 225
Total Biomass 225
Total Density 229
P. pugio Density 233
P. pugio Biomass 237
Discussion: Comparisons of Estimated P. pugio Densities Using Block
Seining and Push Netting Methodologies 237
1989-90 Discussion and Conclusions 242
Correlating Laboratory and Field Toxicity Test Results with Field
Ecotoxicological Biomonitoring 242
LITERATURE CITED 276
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TABLES
& Table Description Page
1. Daily physicochemical water quality parameters
measured during 1989 ' 44
2. Daily physicochemical water quality parameters
measured during 1990 48
3. Rainfall summary for 1989 , 51
4. Days of significant rainfall, 1989 , 53
5. Rainfall summary for 1990 54
6. Days of significant rainfall, 1990 57
7. Measured insecticide concentrations in grab water
samples at the CTL Site, 1989 5&
8. Measured insecticide concentrations in composite
water samples at the CTL Site, 1989 64
9. Measured insecticide concentrations in grab water
samples at the TRT Site, 1989 65
10. Measured insecticide concentrations in composite
water samples at the TRT Site, 1989 70
11. Measured insecticide concentrations in grab water
samples at the KWA Site, 1989 72
12. Measured insecticide concentrations in grab water
samples at tomato field drain-age ditches discharging
at the KWA Site, 1989 80
13. Measured insecticide concentrations in grab water
samples taken at fish kills at the KWA Site, 1989 86
14. Spiked recovery efficiencies in water samples, 1989 87
15. Measured insecticide concentrations in grab and
composite water samples at the CTL Site, 1990 89
16. Measured insecticide concentrations in grab and
composite water samples at the TRT Site, 1990 92
17. Measured insecticide concentrations in grab and
composite water samples at the KWA Site, 1990 95
18. Spiked recovery efficiencies in water samples, 1990 98
19. Survival of P. pugio during 1989 field toxicity tests 77.... 114
20. Survival of Penaeus species during 1989 field toxicity tests 117
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£ Table Description ' Page #
21. Survival ol Mysidopsis bahia during 1989 field toxicity tests 119
22. Survival of-E. heteroclitus during 1989 field toxicity tests 121
23. Survival of Cyprinodon variegatus during
1989 field toxicity tests t 123
24. Quality Control/Quality Assurance bioassay results for 1989 126
25. Survival of P. pugio during 1990 field toxicity tests 129
26. Survival of Penaeus species during 1990 field toxicity tests 132
27. Survival of Mysidopsis bahia during 1990 field toxicity tests 134
28. Survival of F. heteroclitus during 1990 field toxicity tests 136
29. Survival of Cyprinodon variegatus during 1990 field toxicity tests .... 139
30, Survival of Menidia berylina during 1990 field toxicity tests 141
31. Quality Control/Quality Assurance bioassays for 1990 143
32. Oyster ecophysiology studies, 1989-90: Physicochemical parameters . 147
33. Oyster ecophysiology studies, 1989-90: Condition Index 149
34. Oyster ecophysiology studies, 1989-90:
Perldnsus marinus infection intensities 151
35. Oyster ecophysiology studies, 1989-90:
Respiration rates and Q,0 adjusted respiration rates 154
36. Oyster ecophysiology studies, 1989-9~0:
Nitrogen excretion .rates 159
37. Oyster ecophysiology studies, 1989-90:
O/N Ratios 162
38. Oyster ecophysiology studies, 1989-90:
Q]0 Adjusted O/N Ratios 164
39. Summary of 1989 rainfall observed in field bioassays
measuring brain AChE inhibition in F heteroclitus 186
40. Summary of 1990 rainfall observed in field bioassays
measuring brain AChE inhibition in F. heteroclitus 187
41. Summary of measured insecticide concentrations
detected in selected grab water samples collected
during the 1989 field study 189
42. Summary of azinphosmethyl concentrations measured
in water samples and brain AChE levels in F. heteroclitus
during rain events analyzed for 1988-90 197
43. List of species observed during ecotoxicological
sampling, 1989-90 202
VII
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£ Table Description Page #
44. Ecotoxicological sampling, block seining:
Total biofflass, 1989-90 203
45. Ecotoxicological sampling, block seining:
P. pugio densities, 1989-90 i 207
46. Ecotoxicological sampling, block seining:
F. heteroditus densities 1989-90 211
47. Ecotoxicological sampling, block seining:
Total fish densities 1989-90 .' " 216
48. Ecotoxicological sampling, block seining:
Penaied shrimp densities 1989-90 218
49. Ecotoxicological sampling, block seining:
Callinectes sapidus densities, 1989-90 222
50. Ecotoxicological sampling, push netting:
Total biomass 1990 226
51. Ecotoxicological sampling, push netting:
Total densities 1990 230
52. Ecotoxicological sampling, push netting:
P. pugio densities 1990 ,..: 234
53. Ecotoxicological sampling, push netting:
P. pugio biomass 1990 238
54. Comparison of P. pugio densities for block seining
versus push netting during 1990 241
55. Statistical comparisons of P. pugio densities estimated
from block seining and push netting during 1990 243
56. Summary of laboratory toxicity testing with P. pugio and
F. heteroclitus exposed to azinphosmethyl, endosulfan,
and fenvalerate 247
57. Summary of field toxicity testing with P. pugio and
F. heteroclitus, 1985-88 .' 253
58. Summary of field toxicity testing with P. pugio and
F. heteroditus, 1989-90 254
59. Summary of ecotoxicity mortality estimates for
P. pugio and F. heteroclitus, 1985-88 258
60; Summary of ecotoxicity mortality estimates for
P. pugio and F. heteroclitus, 1989-90 259
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#. Table Description Page #
61. Comparison of laboratory toxicity test, field
toxicity tests, and ecotoxicity predicted
mortality in F. heteroclitus, 1989 263
62. Comparison of laboratory toxicity tests, field toxicity tests,
and ecotoxicity predicted mortality in F. heteroclitus, 1990 264
63. Comparison of laboratory toxicity tests,
field toxicity tests, and ecotoxicity predicted
mortality in P. pugio, 1989 : ' 266
64. Comparison of laboratory toxicity tests,
field toxicity tests, and ecotoxicity
predicted mortality in P. pugio, 1990 269
65. Comparison of field and laboratory derived LCM
values for P. pugio and F. heteroclirus exposed
to azinphosmethyl, endosulfan, and fenvalerate 271
66. Comparison of field and laboratory derived EC^
values for brain AChE inhibition in F. heteroclitus
exposed to azinphosmethyl 274
IX
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FIGURES
£ Figure Description Page#
1. Map of study sites used in the 1989-90 field study 14
2. Sketch of net deployment during block seizing
at the TRT Site 38
3. Measured insecticide concentrations (ug/L) and salinities (ppt)
measured at the CTL Site during the 1989 field study 62
4. Measured insecticide concentrations (ug/L) and salinities (ppt)
measured at the TRT Site during the 1989 field study 69
5. Measured insecticide concentrations (ug/L) and salinities (ppt)
measured at the KWA Site during the 1989 field study 78
6. Measured insecticide concentrations (ug/L) and salinities (ppt)
observed at the CTL Site during the 1990 field study 91
7. Measured insecticide concentrations (ug/L) and salinities (ppt)
observed at the TRT Site during the 1990 field study 93
8. Measured insecticide concentrations (ug/L) and salinities (ppt)
observed at the KWA Site during the 1990 field study 97
9. Hydrolab results for the CTL Site, 24-27 May, 1989 100
10. Hydrolab results for the TRT Site, 24-27 May, 1989 101
11. Hydrolab results for the CTL Site, 28-31 May, 1989 102
12. Hydrolab results for the TRT Site, 28-31 May, 1989 103
13. Hydrolab results for the CTL Site, 31 May-3 June, 1989 104
14. Hydrolab results for the TRT Site, 31 May-3 June, 1989 105
15. Hydrolab results for the CTL Site, 3-7 June, 1989 107
16. Hydrolab results for the TRT Site, 3-7 June, 1989 108
17. Hydrolab results for the CTL Site, 7-10 June, 1989 109
18. Hydrolab results for the TRT Site, 7- 10 June, 1989 110
19. Hydrolab results for the CTL Site, 10-13 June, 1989 Ill
20. Hydrolab results for the TRT Site, 10-13 June, 1989 112
21. Survival .of P. pugio in field toxicity tests
during the 1989 field study 115
22. Survival of juvenile Penaeus species in field
toxicity tests during the 1989 field study 118
23. Survival of Mysidopsis bahia in field toxicity
tests during the 1989 field study 120
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S. Figure Description Page#
24. Survival of Fundulus heteroclitus in field
toxicity tests during the 1989 field study 122
25. Survival of juvenile Cyprinodon variegatus
in field toxicity tests during the 1989 field jtudy . . 124
26. Survival of P. pugio in field toxicity tests during
the 1990 field study 130
27. Survival of juvenile Penaeus species in field
toxicity tests during the 1990 field study ' " 133
28. Survival of Mysidopsis bahia in field
toxicity tests during the 1990 field study 135
29. Survival of Fundulus heteroclitus in field
toxicity tests during the 1990 field study 137
30. Survival of juvenile Cyprinodon variegatus in
field toxiciry tests during the 1990 field study 140
31. Survival of juvenile Menidia berylina in field
toxicity tests during the 1990 field study 142
32. Condition index in oysters (Crassostrea virginica')
deployed at the CTL and TRT Sites during the 1989
field study 150
33. Perldnsus marinus infection intensities in
oysters (Crassostrea virginica) deployed at the
CTL and TRT Sites during the 1989 field study 152
34. Q10 adjusted respiration rates (ml/0.685g/h) in oysters
(Crassostrea virginica) at the CTL and TRT Sites for
three exposure temperatures (23°, 25°, and 30°C)
during the 1989 field study 155
35. Mean Q10 standardized respiration rates (ml/0.685g/h)
measured in oystejs (Crassostrea, virginica) at the
CTL and TRT Sites during the 1989 field study 157
36. Ammonia nitrogen excretion rates (ug atoms N/g/h)
in oysters (Crassostrea virginica) deployed at
the CTL and TRT Sites during the 1989 field study 160
37. Mean O/N Ratios in oysters (Crassostrea virginica)
deployed at the CTL and TRT Sites during the
1989 field study 163
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£. Figure Description Page#
38. Mean Q10 Temperature adjusted O/N Ratios in oysters (Crassostrea virginica)
deployed atlhe CTL and TRT Sites during the 1989 field study 165
39. Condition index in oysters (Crassostrea virginica)
deployed at the CTL and KWA Sites durinf the 1990 field study 167
40. Perldnsus marinus infection intensities in oysters
(Crassostrea virginica) deployed during the
1990 field study at the CTL and KWA Sites 169
41. Q10 adjusted respiration rates (ml/0.685g/h)
in oysters (Crassostrea virginica) deployed
at the CTL and KWA Sites during the 1990 field study 170
42. Mean Q10 standardized respiration rates (ml/0.685g/h)
in oysters (Crassostrea virginica) deployed at
the CTL and KWA Sites during the 1990 field study 172
43. Ammonia nitrogen excretion rates (ugatoms N/g/hr)
in oysters deployed at the CTL and KWA Sites
during the 1990 field study 174
44. Mean O/N Ratios in oysters (Crassostrea virginica) deployed
at the CTL and KWA Sites during the "1990 field study 175
45. Mean Q10 adjusted O/N Ratios in oysters (Crassostrea virginica)
deployed at the CTL and TRT Sites during the 1990 field study 176
46. Laboratory predicted EC50 values (ug/L) based upon brain AChE
inhibition in F. heteroclitus exposed to azinphosmethyl for 24h 179
47. Brain AChE levels in F. heteroclitus exposed to
a sublethal dose of azinphosmethyl for 24h 180
48. Whole animal respiration rates (ug O2/g/h) in F. heteroclitus
exposed to a sublethal dose of azinphosmethyl for 24h 181
49. Nitrogen excretion rates (ug atoms N/g/h) in F. heteroclitus
exposed to a sublethal dose of azinphosmethyl for 24h 183
50. Mean O/N Ratios in F. heteroclitus exposed to
a sublethal dose of azinphosmethyl for 24h 184
51. Brain AChE levels in F. heteroclitus exposed at the
CTL, TRT and KWA Sites during the 1989 field study 191
52. Brain AChE levels in F. heteroclitus exposed at the
CTL, TRT, and KWA Sites during the 1990 field study 192
XII
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i£ Figure Description Page#
53. Predicted EC50 (ug/L) value based upon brain AChE inhibition
in F. heteroclitus exposed to azinphosmethyl at the CTL, TRT
and KWA Sites during field studies using maximum measured
azinphosmethyl exposure concentrations . H 198
54. Predicted EC50 (ug/L) value based upon brain AChE inhibition in
F. heteroclitus at the CTL, TRT and KWA Sites during field studies
using 24h post-peak insecticide runoff, azinphosmethyl concentrations 199
55. Predicted EC50 (ug/L) value based upon" brain AChE Inhibition
in F. heteroclitus at the CTL, TRT and KWA Sites during field
studies at the CTL, TRT, and KWA Sites using average
(peak + 24h azinphosmethyl concentrations/2)
azinphosmethyl concentrations 200
56. Total biomass (g/50m of stream) measured in block seining
at the CTL and TRT Sites during 1989-90 204
57. P. pugio densities (#/50m of stream) measured in block seining
at the CTL and TRT Sites during 1989-90 208
58. F. heteroclitus densities (#/50m of stream) measured in
block seining at the CTL and TRTjSites during 1989-90 212
59. Total fish densities (#/50m of stream) measured in block
seining at the CTL and TRT Sites during 1989-90 216
60. Penaied shrimp (Penaeus aztecus, Panaeus duorarum, and
Penaeus setiferus) densities (#/50m of stream) measured
in block seining at the CTL and TRT Sites during 1989-90 219
61. Callinectes sapidus densities (#/50m of stream) measured in block
seining at the CTL and TRT Sites during 1989-90 223
62. Total biomass (g/50m of stream) in push netting
at the CTL and TRT Sites during 1990 227
63. Total biomass (g/50m of stream) in push netting
at the CTL and KWA Sites during 1990 228
64. Total densities (#/50m of stream) in push netting
at the CTL and TRT Sites during 1990 231
65. Total densities (#/50m of stream) measured in push netting
at the CTL and KWA Sites during 1990 232
66. P. pugio densities (#/SOm of stream) measured in push netting_
at the CTL and TRT Sites during 1990 235
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# Figure Description Page#
67. P. pugio densities (#/50m of stream) measured in push netting
at the CTL-and KWA Sites during 1990 236
68. P. pugio biomass (g/50m of stream) measured in push netting
at the CTL and TRT Sites during 1990 .. f 239
69. P. pugio biomass (g/50m of stream) measured in push netting
at the CTL and KWA Sites during 1990 240
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PLATES
£_ Plate Description Page #
1 Aerial photograph of the CTL Site located on the west
branch of Leadenwah Creek * 15
2A. Aerial photograph of the TRT Site located on the
eastern branch of Leadenwah Creek 16
2B. Retention pond constructed at the TRT Site in 1988 16
3 Aerial photograph of the KWA Site located on an unnamed
title tributary of Haulover Creek 17
4 Photograph of dead F. heteroclitus at the KWA Site following
significant rainfall and resulting fish kill 81
5 Photograph of dead P. pugio at the KWA Site following
significant rainfall and resulting fish kill 82
6A. Photograph of dead Uca pugilator at the KWA Site following significant rainfall and
resulting fish kill. There was
significant mortality in fiddler crabs at this site 83
6B. Photograph of dead polychaetes at the KWA Site following significant rainfall and
resulting fish kill 7 83
7A. Photograph of dead Mugil cephalus at the KWA Site following significant
rainfall and resulting fish kill 84
7B. Photograph of shorebirds (gulls, wading shorebirds and egrets) consuming dead fish,
crustaceans and invertebrates at the KWA Site during the fish kill 84
xv
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EXECUTIVE SUMMARY
The primary objective of this study was to evaluate and compare in situ field and
laboratory toxicity tests and ecophysiology bioassays (specific - brain AChE inhibition and
nonspecific-O2 consumption, NH4 excretion, and O/N ratios) for several insecticides
(azinphosmethyl - an organophosphate; endosulfan - an organochlorine, and fenvalerate -
a synthetic pyrethroid) with several field ecotoxicological biomonitoring approaches (block
seining and push netting) for assessing the macropelagic estuarine tidal creek community.
Studies were conducted in pristine habitats (reference or control = CTL Site) and in two
agricultural areas, one highly managed [Integrated Pest Management (IPM), Best
Management Practices (BMP), and retention ponds] for control of nonpoint source pesticide
runoff (Treatment site or Exposure site 1 = TRT Site) and a second unmanaged (no IPM,
BMP, or retention ponds) site (Kiawah or Exposure site 2 = KWA Site). Field and
laboratory studies were conducted over a two year period (1989-90) on coastal sea islands
(Wadmalaw and Johns Island) located just south of Charleston, South Carolina.
Parameters measured included:
1) Laboratory ecophysiology bioassays with mummichogs exposed to azinphosmethyl
for determination of specific (enzyme AChE) and nonspecific (bioenergetic
metabolism - O2 consumption, NH4 excretion, and O/N ratios) biomarkers of
exposure effects,:
2) In situ field toxicity tests with adult mummichogs (Fundulus heteroclitus), juvenile
tidewater silversides (Menidia berylina), juvenile Penaeid shrimp
(Penaeus aztecus, Penaeus setiferus and Penaeus duorarum), adult mysid shrimp
(Mysidopsis bahia), and adult grass shrimp (Palaemonetes pugio} at the CTL, TRT,
and KWA sites during periods of fair weather and following significant runoff events;
3) In situ ecophysiology bioassays with adult oysters (Crassostrea virginica) and
mummichogs using both specific (enzyme AChE) and nonspecific (i.e., bio-energetic
metabolism) biomarkers of exposure/effects at CTL, TRT and KWA sites during
periods of fair weather and following significant runoff events;
4) Pesticide (azinphosmethyl, endosulfan, and fenvalerate) levels in surface waters
(ng/L), sediments (ug/kg) and oysters (ug/kg) at the CTL, TRT and KWA sites
along with transboundaiy measurements from tomato field at the KWA and adjacent
tidal creeks at Seabrook Island;
xvi
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5) Daily and continuous (i.e., every 15 minutes) measurements of water temperature
(°C), salinity (ppt), dissolved oxygen (mg/L), pH, and relative water depth (M) at the
CTL, TRT, and KWA sites before and after periods of significant pesticide runoff;
6) Biomonitoring (monthly or bimonthly block seining) of the macropelagic tidal
creek community at the CTL and TRT sites;
7) Additional biomonitoring (monthly push netting) of grass shrimp (Palaemonetes
pugio) populations at the CTL, TRT and KWA sites;
8) Statistical analysts of all data using both parametric and nonparametric
procedures including calculation of laboratory and field derived LC50 and EC50
values, between site comparisons, and regression analysis (linear and logistic).
Results indicated that during 1989 there were four to five days of significant (> 1.27
cm/day) rainfall during the peak of the vegetable crop growing season (May - June) which
resulted in significant runoff of azinphosmethyl, endosulfan and fenvalerate at the
unmanaged agricultural site at KWA. Discharge of this runoff resulted in significant
mortality to all caged toxicity test species at the KWA Site. Additionally, two fish kills were
observed at this site. Significant inhibition of brain AChE in surviving mummichogs was
noted along with uptake of insecticides (i.e., endosulfan) by oysters. All toxicity observed
in field toxicity tests were the result of pesticide exposure, as a physicochemical water
quality parameters (salinity, pH, and dissolved oxygen) remained at levels within the zone
of capacity adaptation or tolerance for the crustacean and fish species tested.
Transboundary movement (movement away from the original point of discharge) of
insecticide runoff some two river miles (4.5 Km) during one tidal cycle (12-14h) was noted
at the KWA Site, which resulted in additional impacts to juvenile fish in adjacent tidal
creeks.
At the TRT Site, where agricultural management practices were in place (BMP, IMP,
and retention ponds), pesticide discharges were greatly reduced. Only elevated levels of
fenvalerate (65 - 93 ng/L) were observed, which caused significant toxicity among caged
grass shrimp and penaeid shrimp. All grass shrimp and penaied shrimp toxicity resulted
from pesticide exposure, as physicochemcial water quality parameters (i.e., dissolved oxygen)
remained within normal limits of tolerance. There was no significant toxicity observed in
other deployed bioassay species. Additionally, ecotoxicological biomonitoring indicated
significant decreases in field populations (43%) of grass shrimp at the TRT Site. Oysters
were exposed to fenvalerate (15-93 ng/L) and extremely low salinity conditions which
XVll
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caused altered Q10 adjusted respiration (increased), nitrogen excretion rates (increased) and
O/N Ratios; however, these slight changes in energetic metabolism did not translate into
major changes in body component indices (condition index) and parasite infection intensity.
While these results~clearly demonstrated the potential of oyster ecophysiology measurements
to assess NFS pesticide runoff effects, it is important to note the significance of co-factors
such as low salinity exposure, that may occur concomitantly with pesticide exposure and
confound interpretation of results. Results from oyster ecophysiology studies indicate the
need for controlled laboratory studies to confirm that observed exposure response
relationships with pesticides in the field are not influenced by concomittant changes in
physiochemical water quality parameters such a low salinity.
At the CTL or reference site, only background levels of endosulfan (< 10 ng/L) were
observed. There was extremely high survival in all species deployed in field toxicity tests,
along with normal bioenergetic metabolism in oysters and brain AChE in fish.
Biomonitoring indicated very high population densities of grass shrimp, mummichogs and
other fish/shellfish species similar to population densities measured earlier (1985-88) at this
site.
Results from 1990 indicated only two days of significant (> 1.27cm/day) rainfall
occurred during the peak of the vegetable crop growing season (May - June) which resulted
in only slight runoff of azinphosmethyl (
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precision, sampling of more replicates/site may enhance the statistical power using this
method.
Ecotoxicological sampling (block seining and push netting) during 1989-90 was
confined to CTL, TRT, and KWA Sites, No sampling upstream or downstream was
conducted to evaluate spatial distribution of impacts to macropelagic fauna. The labor
intensive method of block seining while thorough and effective, is time consuming and cost
prohibitive to spatially characterize a watershed; however, the use of pushnetting provides
a time and cost effective method for assessing spatial impacts within a watershed and should
be employed in future studies. Additional studies of benthic communities, marine mammals,
sea turtles and bird populations in agriculturally influenced watersheds should be conducted
to fully understand the ecological impacts of pesticide runoff.
Comparing results between the two agricultural sites for 1989-90, clearly indicates the
importance of various management strategies (BMP, IPM, and retention ponds) at the TRT
Site in significantly reducing pesticide risk from vegetable farming to adjacent estuarine tidal
creek nursery habitats. During a relatively dry year, (i.e. 1990) these management practices
were not necessary for protecting the environment, given the small amount of rain and
resulting runoff observed. During a relatively wet year, such as 1989, it was evident that
these management strategies greatly minimized pesticide impacts at the TRT Site, especially
when compared to results for 1985-88, prior to implementation of these practices.
Further analysis of field and laboratory results from this study have indicated:
1) Significant agreement between field and laboratory toxicity tests for
azinphosmethyl, endosulfan, and fenvalerate for grass shrimp (P. pugio) and
mummichogs (F. heteroclitus);
2) The field-derived LC^, for azinphosmethyl was 0.95 ug/L (CL = 0.86 - 1.05
ug/L) versus 96h laboratory, static renewal (SR) LC^, ranging from 0.97 -1.05
ug/L (CL = 0.77 - 1.24 ug/L) in P. pugio.
3) The field-derived LCM for P. pugio exposed to endosulfan was 0.28 ug/L (CL
= not calculated) versus 96h laboratory, SR LC^ ranging from 0.25 - 1.01
ug/L (CL = 0.14 -1.43 ug/L) in adults and 0.39 ug/L (CL = 0.27 - 0.58 ug/L
in zoeae.
XIX
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4) The field-derived LCX for P. pugio exposed to fenvalerate was 0.060 ug/L (CL
= 0.050 - 0.070 ug/L) versus 96h laboratory, SR LC50 valves ranging from
0.052 - 0.060 ug/L (CL = 0.037 - 0.097 ug/L) in adults and 0.007 - 0.020 ug/L
(CL"= 0.005 - 0.031 ug/L) in zoeae.
5) The field-derived LC^ for F. heteroclitus exposed to azinphosmethyl was
>7.002 versus 96h laboratory SE LQo ranging from 28.00 - 36.95 ug/L (CL
= 20.23 - 48.24 ug/L). Also the field derived lowest observable effect
concentration (LOEC) was 7.00 ug/L versus a 96h laboratory, SE no
observable effect concentration (NOEC) of 4.95 ug/L.
6) The field derived LCjo for F. heteroclitus exposed to endosulfan was 0.12 ug/L
(CL = not calculated) versus 96h laboratory, SR LC50s ranging from 1.29 -
1.45 ug/L (CL = 1.21 - 1.59 ug/L) for adults and 0.23 ug/L (CL= 0.14 - 0.40
ug/L) for juveniles. Most field population impacts were in juvenile F.
heteroclitus, which resulted in field derived values more closely resembling
juvenile laboratory toxicity test results.
7) The field derived LC^ for F. heteroclitus exposed to fenvalerate was 0.100
ug/L (CL = 0.090 • 0.110 ug/L) versus 96h, SR laboratory LC50 values
ranging from 1.63 - 2.86 ug/L (CL = 1.080 - 4.060 ug/L) for adults and 2.67
ug/L (CL = 1.670 - 4.260 ug/L for juveniles. The wide disparity between
field and laboratory test results for F. heteroclitus exposed to fenvalerate, may
have resulted from potential low salinity (<10ppt) pesticide, mixture
(endosulfan, fenvalerate and azinphosmethyl) interactions. Laboratory toxicity
tests indicated that fenvalerate was 1.75 times more toxic to adult F.
heteroclitus at Sppt salinity than at 20ppt. Similarly, laboratory studies
conducted with other juvenile estuarine fish [Menidia menidia (LCSO = 0.31
ug/L, CL = 021 - 0.40 ug/L) and Mugii cephalus (LCK - 0.58 ug/L, CL =
0.41 - 1.00 ug/L)] reported similar LCso values for fenvalerate.
8) Significant agreement between field and laboratory derived EC^ estimates of
brain AChE inhibition in mummichogs exposed to azinphosmethyl;
9) The 24h laboratory-derived EC^ based upon brain AChE in F. heteroclitus
exposed to azinphosmethyl was 0.90 ug/L versus field derived EC50 values
ranging from 0.63 ug/L (based upon 24h post rain pesticide insecticide
xx
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concentrations) to 1.53 ug/L (based upon peak insecticide concentrations).
A field derived £€50 value of 1.13 ug/L was determined using the average
insecticide concentration measured during each rain event. The most
significant limiting factor affecting correlations between field and laboratory
validations for biomarkers such as AChE may be in the accuracy of the
characterization of field pesticide concentrations. These results suggest, based
upon differences between peak, post rain and 24h, post rain sampling, average
measured EC^ values for azinphosmethyl AChE inhibition varied by a factor
of 2.42. For field validation determinations, it is important to report
maximum, minimum and average pesticide concentrations measured in all
field studies to accurately characterize field exposure concentrations. It is
interesting to note that the 24h field derived EC^ of 1.13 ug/L, which was
derived, based upon the average post rain concentration only varied from the
laboratory derived £€50 of 0.90 ug/L by a factor of 1.25.
10) Significant sublethal (increased respiration and increased nitrogen excretion
rates) effects of pesticide (fenvalerate and azinphosmethyl/endosulfan)
runoff on adult oysters (Crassostrea virginica) ecophysiology;
11) Significant sublethal effects (reduced nitrogen excretion) from mummichogs
exposure to azinphosmethyl^
12) Significant correlation between block seining and push netting approaches
for assessing ecotoxicology effects of pesticide runoff in macropelagic fauna;
13) Ecological recovery of the macropelagic faunal community at the highly
managed agricultural site (TRT Site) following construction of retention
ponds there to reduce nonpoint source agricultural runoff into estuarine tidal
creeks; and;
14) Insecticide runoff and related lexicological impacts to estuarine organisms
were greatly reduced at the highly managed agricultural site when compared
to the unmanaged site.
Results from this study have indicated that the application of field and laboratory
testing for both lethal (acute toxicity) and sublethal (ecophysiology) pesticide effects when
coupled with ecotoxicological biomonitoring provides an integrated method for holistic
environmental risk assessment for pesticides. Laboratory toxicity and ecophysiology studies
xxi
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provide the initial bench marks for lethal and sublethal effects for each pesticide studied.
The range between lethal and sublethal endpoints determined in the laboratory indicates
the boundary for potential and/or realized field impacts. Field toxicity tests and
ecophysiological studies provide a mechanism to validate initial laboratory studies and to
expand their design (pulsed vs. continuous dose; low vs. high salinity; the interaction of
pesticide mixtures; and adult vs. other life history stages) to better interpret field results.
Ecotoxicological biomonitoring provides an independent mechanism to confirm the validity
of both laboratory and field toxicity tests and in some instances (i.e. reproductive
impairment) ecophysiology bioassays.
Application of BMP, IPM, and retention ponds as nonpoint source insecticide runoff
control techniques was very effective at the TRT Site at reducing surface water risks and
impacts to organisms in adjacent estuarine tidal creeks. Ecotoxicological biomonitoring at
the TRT Site indicated recovery of the macropelagic fauna at this site to levels comparable
at the CTL Site. The integrated field and laboratory toxicological methods employed in this
study were not only effective in quantifying insecticides impacts, but were equally as
sensitive in documenting ecological recovery as contaminant risks were reduced. These
results imply that the methods employed in this study would be effective not only in
pesticide risk characterizations, but also in risk reduction evaluations. This is extremely
important!
Highly correlated field and laboratory results for pesticide risk assessment imply a
mechanism for simultaneously quantifying toxicological risk and evaluating risk management
options. In practice, very few studies have been able to document both. The present study
has clearly indicated the utility of this approach for a variety of insecticides
(organophosphates, organochlorines, and pyrethroids), in a variety of species (fish and
shellfish), using a variety of risk management options.
Future studies should be better focused at evaluating the implications of sublethal
effects or biomaikers in terms of population level impacts. Sublethal effects studies should
distinguish between biomarkers which are indicators of specific contaminant exposure (i.e.,
AChE) and those which are nonspecific indicators of significant sublethal effects (i.e.
bioenergetic metabolism). Linking cause and effect with generalized physiological indicators.
is tenuous at best. Utilization of both specific (i.e. AChE) and nonspecific (bioenergetic
metabolism) sublethal indicators provides a holistic method of determining if significant
contaminant exposure has occurred and if that exposure is translatable into
ecophysiologically significant effects. The approach used in this study provides a template
for evaluating this question and has partially answered this question for estuarine fish and
XXll
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shellfish exposed to azinphosmethyl. Further study of this issue is needed for other
insecticides to adequately address this question.
Future laboratory studies which utilize specific and nonspecific biomarkers should be
conducted to evaluate the persistence of any sublethal effect beyond just the initial exposure
phase. Such studies would distinguish between labile effects and permanent impairment.
Obviously, any organism when exposed to a pesticide may exhibit altered physiological
responses. Once insecticide exposure is terminated, if the effect disappears rapidly, then the
observed biomarker was not ecophysiologically significant. If the effect persists, then the
potential for permanent impairment exists. The application of this approach should begin
with better designed laboratory studies so that those biomarkers of prolonged
ecophysiological impairment may be identified and distinguished from indicators of exposure
per se. Application of laboratory results to field studies in populations would be made
easier and more directly translatable to environmental risk assessment for insecticides.
XXlll
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ACKNOWLEDGEMENTS
The authors wish to thank all the many people who have provided a significant
contribution in the completion of this large research effort. Obviously, in a research effort
so large, it is impossible to thank everyone individually. However, there are some
individuals whose contributions were so noteworthy, they are listed below.
First, the authors wish to thank the U.S. Environmental Protection Agency, Gulf
Breeze Environmental Research Laboratories (GBERL), Pensacola, Florida, for support of
this research through a Cooperative Agreement. Dr. James R. Clark, of GBERL, our
original Project Officer, has been invaluable for his assistance in designing, conducting, and
analyzing research results collected during this project. Dr. Clark has provided significant
insight and input into many of the research areas under study. His contributions have been
most significant. Dr. Mike Lewis (Final Project Officer) has also been extremely helpful in
final preparation of this report. Also, Dr. Sonny Mayer (Ecological Effects Branch Chief)
has been very supportive of this research and we acknowledge his assistance throughout this
effort. Additionally, Mr. Jim Patrick, Mr. John MacCauley, Mr. Larry Goodman, Roman
S. Stanley and Mr. George L. Craven, all of GBERL, have also made significant
contributions during the 1989-90 field studies.
Secondly, we would like to thank all of Leadenwah Creek landowners who have
allowed us virtual unlimited access to their property. Mr. Lester H. Bentz has been
invaluable for his assistance in studying fish kills on the Leadenwah Creek. His contribution
to the research effort is greatly appreciated. Mr. and Mrs. William L. (Bill and Emily)
Leland, Jr. have also allowed us the use of their dock and property. Their contribution is
also greatly appreciated.
Thirdly, we would like to thank the members of the Ad Hoc Fish Kill Committee (Mrs.
Jane Settle and Dr. Tommy Mathews - SC Wildlife and Marine Resources Department; Mr.
Wayne Fanning - SC DHEC; Dr. Von McCaskill, Dr. Randy G. Griffin, and Mr. Cam Lay -
Clemson University (Cooperative Extension Service and Plant Pest Regulatory Authority);
Mr. Barrett Lawrimore (SC Tomato Growers Association) and Mr. John Wapole (Farmer -
Wadmalaw Island). The input provided by this Fish Kill Committee has been invaluable.
This group has been able to integrate farming and marine ecology interests into a unified
management plan for minimizing the effects of agriculture runoff into marine waters. This
group is to be commended for its management activities in a regulating nonpoint source
agricultural runoff.
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Fourthly, we would like to thank Mr. Andy A. Schlon of Andy's Deli for providing a
virtually unlimited supply of five gallon buckets which were used for collection of field
samples. Additionally, we would like to thank the SC Sea Grant Consortium for providing
initial research funds through their "Seed Project" Program which allowed our group to
being toxicological studies of agricultural insecticides. The authors also wish to thank Ms.
Betsy Cooke for her preparation of the many graphics presented in this report. Ms. Rachael
Suggs and Mr. Will Suggs are acknowledged for their help during summer field work. The
authors wish to thank Mrs. Meryl Reese and Jan Carson for their unending efforts in typing
such an extensive manuscript. Their contribution was significant and greatly appreciated.
Mr. Tom Siewicki of the National Marine Fisheries Service, Charleston Laboratory is
commended for the excellent job done in editing the original text of this report. Also Dr.
James R. Clark (Exxon Biomedical) and Anthony S. Pait (US NOAA) are acknowledged
for their thorough editing of the final report.
This report was submitted in fulfillment of the Cooperative Agreement CR816213
between the U.S. Environmental Protection Agency and the University of South Carolina
School of Public Health.
XXV
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INTRODUCTION
Throughout the U.S., continued public scrutiny has been placed on the use of
pesticides in the Environment (i.e., Med-fly spraying in southern California). While
the use of pesticides in agriculture and vector control has been justified due to world
shortages of food and public health concerns, respectively, their impacts on the
environment are being more widely studied, both in the U.S. and other countries of
the world. A total of nearly one billion pounds of active ingredient (PAI) pesticide
was produced in the U.S. in 1983. Agriculture was the dominant use (77%) followed
by industrial and government use (16%) and home and garden use (7%).
There are over 50 companies which produce over 960 pesticides, in the U.S.,
sold in more than 25,000 formulations. Ten percent of all U.S. pesticide production
consists of unregistered (in the U.S.) or banned products (i.e., DDT1) which are sold
in overseas market (Revelle and Revelle, 1988). For many developing third world
nations, pesticide usage is indispensable to prevent starvation and disease (Atuma,
1985). Risk assessments of pesticide use must be balanced differently in economically
developed and underdeveloped nations of the world. Risk assessments for pesticides
must be far ranging, protective of human-consumers, farm workers, and avian,
terrestrial and aquatic resources. Particular emphasis must be placed on coastal and
estuarine ecosystems, given their ecological importance and as a commercial and
recreational source of food.
Pait et al (1989) summarized agricultural pesticide usage for 28 pesticides in
estuarine drainage areas of the U.S. Nationally some 800 million PAI pesticides were
applied to agriculture in the contiguous U.S. The 28 pesticides which were evaluated,
accounted for 50% of all pesticide applications nationwide. Over 34 million PAI of
these 28 pesticides were applied in coastal areas, representing 8% of their total use in
the U.S. The greatest amount of pesticide was applied to corn (> 10,000,000
PAI/yr), followed by soybeans (>8,000,000. PAI/yr), rice (>2,000,000 PAI/yr),
peanuts (> 1,200,000 PAI/yr) and pasture/range land (> 1,000,000 PAI/yr) in coastal
habitats of the U.S. The coastal estuarine drainage areas with the highest pesticide use
'Trade names are provided for information only and do not imply endorsement by the
National Oceanic and Atmospheric Administration.
1
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were Chesapeake Bay (5,290,000 PAI/yr), Winyah Bay (3,240,000 PAI/yr),
Albemarle Sound-(2,132,000 PAI/yr), Pamlico Sound (1,963,000 PAI/yr) and Laguna
Madre (1,902,00 PAI/yr). Pesticide usage was greatest in the Gulf and Southeast
coast, followed by Northeast coast, and West coast. The dominant pesticides applied
to crops in the coastal zone of the U.S. were Alachlor (herbicide - >6,000,000
PAI/yr), Atrazine (herbicide - 5,000,000 PAI/yr), Metolachlor (herbicide -
>2,500,000 PAI/yr), 2,4D (herbicide - >2,000,000PAI/yr, andCarbaryl (insecticide
- 1,500,000 PAI/yr). While the 28 pesticides evaluated in this study represented the
majority of pesticides applied to agricultural crops in the coastal areas of the U.S.,
several important and highly toxic pesticides were not considered, including
azinphosmethyl, fenvalerate and many of the other pyrethrins. Additionally, the size
of the drainage basin, the proportion of the drainage basin cultivated in agriculture,
and the pesticide use per unit of cropland were dominant factors affecting the overall
ranking of potentially high risk areas. As a result, some site specific effects may have
been overlooked. For example, soybeans (and the pesticide applied to this crop) were
the dominant crop in most estuarine drainage areas. As a result, the pesticide applied
to soybeans would dominate summary statistics for pesticide usage. Yet throughout
several estuarine drainage areas, other crops (i.e., tomatoes) may have much higher
pesticide application rates and consequently pose a greater toxicological risk.
Estuarine habitats adjacent to these agricultural areas would be at greatest risk
from pesticide effects and impacts. More than 70% of commercial and recreational
fisheries landings are taken from estuaries (Department of Commerce, 1988).
Additionally, these estuarine and coastal habitats provide significant recreational and
aesthetic pleasures to the public. More than $7 billion of public funds are spent
annually on outdoor marine and estuarine recreation in the 22 coastal states of the
U.S. (NOAA, 1988). It is imperative that effective methodologies for pesticide risk
be developed which adequately protect these fragile estuarine habitats.
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In many of the estuarine drainage areas of che U.S., agricultural lands comprise
a substantial portion of the land use. This is particularly true in the Southeast and
Gulf coast regions of the U.S. Pait et al, (1989) reported that 8 of the top 10
estuarine drainage basins, with the highest proportion of land use as agriculture were
located in the Southeast and Gulf coast regions of the U.S., with agriculture
comprising anywhere from 36%-75% of the land surface area in given estuarine
drainage areas. It is not surprising that nonpoint source (NFS) runoff from agriculture
is thus a major concern in terms of water quality contraventions in estuarine habitat
impacts.
Recent reports (NOAA, 1988; Humenik, 1987; EPA, 1984; and EPA, 1983)
have indicated that in most regions of the U.S., NPS runoff remains one of the more
pervasive, least understood, and poorly managed sources of water pollution. NOAA
(1988) reported in a survey of 145 marine pollution experts that NPS pollution ranked
fourth in terms of severity out of 83 marine pollution problems evaluated. The
Association of State and Interstate Water Pollution Control Administrators
(ASIWPCA) evaluated the effects of NPS runoff in 49 states in the U.S. and reported
that water quality was threatened and/or degraded due to NPS runoff in 42% of the
estuarine habitats surveyed (ASIWPCA, 1985). On a national basis the major sources
of NPS runoff are agriculture, urbanization, mining, silviculture, and
construction/building activities (EPA, 1984). These findings clearly indicate the
importance of estuarine habitats and their potential vulnerability to chemical
contaminants present in NPS runoff.
Although the effects of NPS runoff on estuarine habitats have not been-fully
studied, efforts have been made to identify estuarine impacts associated with this type
of pollution. Trim and Marcus (1990) reported that in South Carolina from 1978-88,
a total of 805 fish kills occurred, of which 354 (44%) occurred in estuarine waters.
In evaluating these estuarine fish kills, it was found that 43% were from natural causes
(i.e., depleted DOi), 35% were from anthropogenic causes and 22% were from
undetermined causes. Nearly 54% of the anthropogenic induced fish kills were from
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coastal pesticide usage including weed control around resorts (21 %), agriculture (20%)
and vector control (13%). Almost all of these pesticide related fish kills were from
NFS runoff rather than spray drift or misapplication. Of additional interest was the
fact that point source discharges from the 123 permitted estuarine discharges accounted
for only 12% of all anthropogenic related fish kill (Trim and Marcus, 1990).
Moreover, a recent assessment has also indicated that 65% of all closed shellfish
harvesting waters in South Carolina are due to bacterial pollution from nonpoint
sources of pollution (SCDHEC, 1988). These results clearly indicate the significance
of NFS pollution in the State of South Carolina. In other coastal states similar
problems with NFS pollution abound.
Fish kills and shellfish closure clearly represent environmental episodes and/or
conditions where environmental management, permitting procedures and regulatory
policies have failed. These episodic events must be considered over a long time frame
and when integrated with other available data bases (i.e., ambient monitoring data) so
that management alternatives can be formulated, tested and evaluated.
Such is often the case with pesticides registration processes within the federal
statutory authority of the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA)
used to register pesticides within the U.S. Environmental hazard evaluation is a
critical and significant part of the pesticide registration process. Scon et al (1990) in
an integrated laboratory and field study of pesticide impacts to estuarine organisms
clearly found significant statistical correlations between 96 hour laboratory toxicity
tests and in situ field toxicity tests and in stream ecotoxicological biomonitoring
approaches. These three approaches - laboratory toxicity tests, in situ toxicity tests,
and ecotoxicological biomonitoring provide the lexicological cornerstones found in
pesticide hazard evaluations and environmental risk assessments. Additionally (Fulton,
1989), also reported in Scott et al (1990) found significant statistical correlations
between sublethal physiological biomarkers (brain acetylcholinesterase) in comparisons
between fish exposed in the laboratory and to NFS pesticide runoff in the field. While
these studies were significant in defining and relating the integration and
interrelationships between field toxicity tests and ecotoxicological biomonitoring with
laboratory toxicity test methodologies, additional studies are needed to better define
these associations.
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The objective of this present study was to continue and expand the approaches
used by Scott et al (1990) in pesticide hazard evaluation processes by:
1) Continued study of in situ field toxicity testing and ecotoxicological
biomonitoring of sites impacted by agricultural NFS runoff;
2) Comparing NPS runoff in situ effects at field sites with [retention ponds,
Best Management Practices (BMP) and Integrated Pest Management (IPM)]
and without (Calendar Spray - a spray application every three - five days,
no IPM/BMP) significant NFS runoff controls measures;
3) Evaluating and comparing the significance of biomarkers (brain -AChE) as
measures of both exposure and sublethal physiological effects;
4) Evaluating the utility of bioenergetic metabolism approaches (i.e. scope for
growth) in assessing NPS runoff effects in the American oyster, Crassostrea
virginica (Gmelin); and
5) Evaluating and comparing the use of more rapid assessment (i.e., push
netting) ecotoxicological sampling approaches with more traditional
biomonitoring methods (i.e., block seining) in assessing pesticide NPS
runoff effects.
These additional studies were undertaken as a Cooperative Research Agreement
between the University of South Carolina, School of Public Health and the U.S.
Environmental Protection Agency, Gulf Breeze Environmental Research Laboratory.
The purpose of this present study was to better define the relationship between
conventional 96 hour Laboratory toxicity tests and in situ field effects for three major
classes of insecticides - organochlorines (endosulfan), pyrethroids (fenvalerate> and
organophosphates (azinphosmethyl). By better understanding the toxicological
interrelationships between laboratory and field toxicity tests and ecotoxicological
studies, greater insight into effective pesticide risk assessment may be gained.
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MATERIALS AND METHODS
Insecticides Studied
Azinphosmethyl
Azinphosmethyl, also known as Guthion or Gusathion, is an organophosphate
having the chemical designation: O,O-dimethyl S-[(4-oxo-l ,2,3benzotriazin-3(4H)-yl)
methyl] phosphorodithioate (Turner, 1977). Azinphosmethyl was developed by Bayer
A.G. in 1953 and is used as a nonsystemic insecticide and acaricide. The structure
of azinphosmethyl, along with selected physicochemical factors are listed below:
Azinphosmethyl
Molecular weight: 317.3
Octanol/water partition coefficient: 360 @ 20°C
Solubility in water: 29mg/L @ 25°C
Melting Point: 73-74°C (Verttorazi, 1976; Morifusca, 1977)
Persistence: 14 days (water); 12-28 days (soils)
(Shultzet. al 1972; Staiff et. al., 1975:
Gunther et. al., 1977; and Engelhart et. al., 1984).
Like most organophosphate insecticides, azinphosmethyl acts by blocking
synaptic transmission. The disruption of the nerve impulse is caused by excessive
amounts of the neurotransmitter acetylcholine (ACh) at the synapses which is
normally broken down by acerylcholinesterase (AChE). In order for azinphos-
methyl to exert its cholinergic effect it must first be metabolized by replacing the
sulfur of the thiophosphate linkage with an oxygen. This oxygen analogue then
binds to the active site of the AChE to prevent breakdown of ACh. Once
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neurotransmission in the respiratory center of the brain or the neuromuscular
junction of the respiratory apparatus has been blocked, death rapidly ensues.
Depressed AChE activity may persist for weeks and it is possible with repeated
exposure to see an additive effect (Dubois et al., 1957; Coppage and Matthews,
1974). Of the more than 16 metabolites that have been identified only the oxygen
analogue of azinphosmethyl has been shown to exert toxicity (Yaron et al., 1974).
Despite the fact that azinphosmethyl has been registered for agricultural uses
for some time, there are relatively few studies concerning its effects on nontarget
species. Loosanoff et al. (1957) examined the effects on certain forms of plankton
and found azinphosmethyl to be nontoxic to two species of freshwater algae,
Chlorella sp. and Chlamydomonas sp. at a concentration of 1 ug/L. Further, he
found that it exerted no toxic effects on oyster larvae and at a concentration of 0.05
ug/L actually' enhanced growth (Loosanoff et al., 1957). A study by Benke and
Murphy (1974) compared toxicity of azinphosmethyl and methyl parathion, another
organophosphate, in mice and fish. They found azinphosmethyl to be
approximately 400X more toxic than methyl parathion in the sunfish Lepomis
gibbosus. This difference was explained in part by the greater sensitivity of brain
and muscle acetylcholinesterase to azinphosmethyl as indicated by in vitro I50
(Concentration causing 50% AChE Inhibition) values: 4.8xlO~'°M for brain and
2.4xlO"'°M for muscle tissue with azinphosmethyl as compared to 2xlO"BM for
brain and 4xlO"8M for muscle with methyl parathion. Additionally the authors
found additive effects with repeated exposure (Benke et al., 1973; Benke and
Murphy, 1974). In a study examining the toxicity of azinphosmethyl in salmonids,
Katz (1961) found 96h LCjoS ranging from 3.2 to 4.3 ug/L for three freshwater
species (Oncorhynchus tshawytscha, Oncorhynchus kisulch, and Salmo gairdneri)
while for a marine species, Gasterosteus aeuleatus, the 96h LCM ranged from 4.8
ug/L (salinity-25 ppt) to 12.1 ug/L (salinity-5 ppt). In a study by Adelman et al.
(1976) on the toxicity of azinphosmethyl to the Fathead minnow, Pimephales
promelas, a concentration of 0.51 ug/L was found to drastically reduce fecundity.
Meyer (1965) determined 48h LC^s for four freshwater species ranging from 25
ug/L in green sunfish, Lepomis macrochirus, and Largemouth bass, Micropterus
salmonids, to 9,000 ug/L in channel catfish, Ictalurus punctatus. Other studies
examining the toxicity of azinphosmethyl in various American freshwater fish have
reported 96h LCjoS ranging from 0.4 ug/L to 4300 ug/L with most values less than
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50 ug/L (Lewallen and Brydon, 1958; Weiss, 1961; Pickering et ai., 1962; Macek
and McCallister, 1970; and Murty, 1986). A limited number of studies have
examined the toxicity of azinphosmethyl in saltwater species with toxicity values
ranging from a 48h LCJO in Brown shrimp, Penaeus aztecus, of 2.4 ug/L to a 96h
LC5Q of more than a 1000 ug/L for the Eastern oyster, Crassostrea virginica. again
with most values less than 50 ug/L (Coppage and Matthews, 1974; Miura and
Takahashi, 1976; and Mayer, 1987).
Endosulfan
Endosulfan also known as Thiodan, Thiosulfan, Cyclodan, and several
other trade names is a chlorinated cyclodiene having the chemical designation:
6,7,8,9,10,10-Hexachloro-l,5,5a,6,9,9a-hexahydro-6,9-methane-2,4,3-
benzodioxathiepin-3-oxide (Berg, 1985). Endosulfan is a nonsystemic contact
insecticide and acaricide. It was developed in Germany by Hoechst AG in 1965 and
is distributed in the U.S. by the FMC Corporation (Thomson, 1985). The
structure of endosulfan, along with selected physicochemical properties are listed
below:
Endosulfan
Molecular weight: 407.0
Solubility in hexane: 24g/L @ 20°C
Solubility in water: alpha isomer 0.32mg/L @25CC
beta isomer 0.33mg/L @ 25 °C
Melting Point: alpha isomer 109 °C beta isomer 213.3'C
(Rao and Murty, 1980; BCPC, 1983)
Persistence: 14 days (water); 60-160 days (soil) metabolite (endosulfan
cyclic sulfate) is highly persistent. (Eichelberger and
Lichtenberg. 1979; McEwen and Stephenson, 1979; Rao and
Murty, 1980).
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The exact mode and site of action of endosulfan is not completely known.
Like most insecticides, the cyclodienes are neurotoxins and are generally
considered to be central nervous system stimulants. Biochemical studies with
dieldrin, another cyclodiene, showed an alteration of amino acid ratios and an
increase in ammofiia levels in the brain (Murphy, 1980). A study by Truhaut et
al., (1974) found endosulfan to cause inhibition of hamster serum and rat hepatic
cholinesterase. In a study by Gupta, (1976), it was found that acetylcholinesterase
activity in rat brain was decreased by 23-33% following an intraperitoneal injection
of 30-60 mg/Kg of endosulfan.
There is a considerable volume of data concerning the toxicity of endosulfan
to nontarget aquatic species. A study by Mathiessen and Logan (1984) examined
the toxicity of endosulfan to tropical cichlids (Tilapia rendalii and Sarotherodon
mossambicus) following aerial spraying of endosulfan for control of the tse tse fly.
They found 75 fewer cichlids nests in treated areas and a 25 % reduction in juvenile
recruitment. In this same study the 24h LCX for Sarotherodon mossambicus was
found to be 10.4 ug/L with a concentration of 0.6 ug/L affecting breeding behavior
(manifested as a delay in spawning). In a study by Roberts (1975), mussels and
scallops were exposed to 450 ug/L technical grade endosulfan for 24 hours.
Results indicated a 50% reduction in byssal attachment. Another study by
Netrawali et al., (1986) examined the effects of endosulfan on the sexual life cycle
of Chlamydomonas reinhardtii and found a delay in the onset of meiosis in the
zygote at a concentration of 0.25 x lO^m (10.18 ug/L). One study which
compared sediment versus superficial water exposures using the shrimp, Crangon
sepiemspinosa, reported that exposure through water was the primary factor
controlling toxicity in this species with a 96h LC50 of 0.2 ug/L in water and 3.5-49
ug/Kg in sediment exposures (McLease and Metcalfe, 1980). Haider and Inbaraj
(1986) compared the toxicity of the technical material and commercial formulations
of endosulfan in adult Channa punaatus and found the 96h LCX for the
emulsifiable concentrate (3.0 ug/L) to be 1.88 times less than that for the technical
material (5.78 ug/L). Additionally, exposed animals in both formulations exhibited
definable behavioral changes such as: 1) erratic swimming; 2) convulsions;
3) increased or difficulty in respiration; 4) loss of equilibrium; 5) pale color; and
6) excessive mucous about the gill epithelium. In a study comparing the toxicity of
the two isomers of the parent compound, the alpha isomer was found to be more
toxic than the beta isomer in the fish, Channa punctatus, with 96h LCjo values of
0.16 p,g/L versus 6.6 ug/L respectively (Devi et al., 1981). Other studies
examining the toxicity of endosulfan in various species of freshwater fish reported
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96h LC50s ranging from 0:.9 to 8.1 ug/L depending on the organism tested (Macek
et al., 1969; Johnson, 1980). In general, results from the toxicity tests on
freshwater species indicate that endosulfan is generally more toxic to fish than
invertebrates. Fewer tests have been conducted with saltwater species but available
results indicated marine species are at least equally if not more sensitive to
endosulfan than freshwater organisms. Results from toxicity tests on various
saltwater species indicated 96h LC50 ranging from 0.04 ug/L for the pink shrimp,*
Perweus duorarum, to 1.31 ug/L for the grass shrimp, Palaemonetes pugio
(Schimmel et al., 1977).
Fenvalerate
Fenvalerate also known as Pydrin, Belmark, Ectrin, Sumicidin, and other
trade names is a synthetic pyrethroid with the chemical formula Cyano(3-
phenoxyphenyl)- methyl 4-chloro alpha (1-methylethyl) benzeneacetate. Fenvalerate
is used as a selective contact and stomach poison insecticide. It was developed by
Sumitomo Chemical Co. of Japan in 1974 was originally distributed in the US by
Shell, and is currently distributed in the U.S. by DuPont (Thomson, 1985). The
structure of fenvalerate, along with selected physicochemical properties are listed
below:
Fenvalerate
Molecular weight: 419.9
Octanol/water partition coefficient: 1.58 x 10* @ 20°C
Solubility in water: 0.002mg/L @ 23°C
Melting point: viscous liquid at room temperature
Persistence: reported half-life 2-7 weeks (Mulla et. al., 1978; Ware,
1980; Schimmel et. al., 1983; Caplanef. al., 1984; and
Smith and Stratten, 1986).
10
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Fenvalerate is a type II pyrethroid which acts on the central nervous system
(Bradbury et al.. 1986). Fenvalerate causes a depolarization of the nerve
membrane by effecting its permeability to Na^and K4" ions resulting in rapid and
repetitive firing of the nerve impulses, leading to disorientation and death. It is
thought that fenvalerate wedges in the open Na+ channels so they cannot close. As
a result, the membrane potential is unable to return to resting state and remains
partially depolarized. When the level of depolarization is near the threshold
voltage, repetitive firing of the nerve cell may occur. The repetitive discharge is
manifested as hyperexcitability and convulsions in the affected animal. In the
continued presence of the insecticide, the nerve cell becomes increasingly
depolarized until impulse conduction is blocked and death results (Shell, 1977).
Numerous studies have considered the effects of fenvalerate on nontarget
aquatic organisms. A study by Coats and Jeffery (1987) compared the toxicity of
the technical and emulsifiable formulations of fenvalerate on Rainbow trout, Salmo
gairdneri. in static tests. They found the emulsifiable formulation to be 3.2 times
more toxic than the technical grade material with 24h LCjoS of 21 ug/L and 76
ug/L, respectively. However, a study by Bradbury et al., (1986) comparing the
toxicities of the technical and emulsifiable formulations of fenvalerate on the
Fathead minnow, Pimephales promelas, found no significant difference in toxicity
between formulations. Behavioral changes observed during acute toxicity studies
included: 1) rapid gill movement; 2) erratic swimming; 3) altered schooling
activity; and 4) swimming at the surface (Holcombe et al., 1982; Bradbury et al.,
1986). In a study by Dyer et al., (1986) increasing water hardness was found to
enhance the toxicity of fenvalerate to the Bluegill, Lepomis macrochints. Symonik
et al. (1986) exposed bluegills Lepomis macrochirus, to the technical material and
the individual isomers (2S, aS; 2S, aR; 2R, aS; 2R, aR) of fenvalerate and found
that the 2S, aS isomer was 100 times more toxic than the next most toxic isomer
2S, aR. All the R-acid isomers were found to be essentially nontoxic.
A study by McKenney and Hamaker (1984) examined the effects of
fenvalerate on the larval development and metabolism of grass shrimp,
Palaemonetes pugio, during osmotic stress. Flow-through exposures to a nominal
concentration of 3.2 ng/L significantly reduced the number of larvae completing
metamorphosis. Further, larvae reared continuously in a sublethal concentrations of
11
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0.1 and 0.2 ng/L showed significant increased metabolic rates when subjected to
acute fluctuations in salinity as compared to controls at salinities of 10 ppt and 30
ppt. Fenvalerate has also been shown to exhibit a negative temperature coefficient
(i.e. greater toxicity widi decreasing temperature).
Other studies (Shell, 1977) examining the toxicity of fenvalerate in various
species of freshwater fish found 96h LC50s ranging from 0.64 ug/L in bluegill to
6.2 ug/L in rainbow trout. Toxicity values in marine organisms have been found to
range from a 96h LC50 value of 0.003 ug/L for larval grass shrimp, Palaemonetes
pugio, to 1600 ug/L for Amphioxus (Schimmel et al., 1983; Clark et al., 1985).
Study Sites
The study sites for field research were located south of Charleston, South
Carolina at Leadenwah Creek (Coordinates - Latitude N32°36'12" Longitude
W80°07') on Wadmalaw Island and an unnamed tidal creek (Coordinates-Latitude
N32°36'7", Longitude W80°07') on Johns Island. (Figure 5). The eastern branch
of Leadenwah Creek is surrounded by extensive agricultural fields used for
vegetable (tomatoes, snap beans, cucumbers, and squash) fanning (Plate 1). Fields
here are drained by ditches which discharge into the eastern branch of Leadenwah
Creek. The eastern branch of Leadenwah Creek has been the site of numerous fish
kills over the past 10 years and was designated the Treatment Site (TRT).
A reference or Control Site (CTL) was selected on the western branch of
Leadenwah Creek, which lies within the drainage basin of rural, single family
dwellings bordered by upland forests and saltmarsh (Plate 2). There are
approximately 40 acres of tomato fields under cultivation which drain into a pond
at this site. Both Leadenwah Creek sites are remarkably similar in terms of
hydrographic regime, salinity, dissolved oxygen, pH, temperature, and sediment
substrates.
An additional agricultural site was selected for study on an unnamed tidal
creek, near Kiawah Island (Plate 3). This site was designated the Kiawah Site
(KWA) which lies in the drainage basin of several large agricultural fields used for
12
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vegetable crop cultivation. Salinities and pH were slightly higher ai this site and
the sediments were finer grain-sized (muds versus coarse-fine sand) than at the
Leadenwah Creek study sites. Additional differences were that agricultural fields
in this area have an extensive vegetative border when compared to the TRT site
and farmers do not use Integrated Pest Management (IPM) or Best Management
Practices (BMP) nor is runoff controlled by retention ponds.
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Figure 1. Map of study sites used in 1989-90 field study. Sites include the
Reference (REF) or Control (CTL) Site on the west branch of
Leadenwah Creek; the Exposure 1 (EXP-1) or TRT Site on the eastern
branch of Leidenwah Creek; and the Exposure 2 (EXP-2) or Kiawah
Site on an unnamed tidal tributary of Haulover Creek. Arrow (t)
denotes the Cherry Point (CP) Collection Site for mummichogs
deployed in field toxicity tests.
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Plate 1 Aerial photograph of the CTL Site located on the west branch of Leadenwah
Creek. Note the presence of scattered rural single family dwellings and
deciduous forests in the area.
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Plate 2A
Aerial photograph of the TRT Site located on the eastern branch of
Leadenwah Creek. Note the presence of extensive agricultural fields at this
site.
Plate 2B Retention pond constructed at the TRT Site in 1988.
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Plate 3 Aerial photograph of the KWA Site located on an unnamed title tributary of
Haulover Creek. Note the extensive agricultural fields surrounding this site.
17
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These three study sites offer a diverse approach for, dealing with nonpoint source
runoff. Drainage ditches are prevalent in agricultural fields at all three sites. The CTL Site,
has less agricultural acreage (relative to the other sites) and nonpoint source runoff has been
controlled by channeling all runoff into a small retention pond. The TRT Site has extensive
agricultural fields (> 1.000 acres), without extensive vegetative buffer strips. During 1985-
87, agricultural runoff at the TRT Site flowed directly into ditches that extensively drained
agricultural fields in this area, which then discharged directly into the headwaters of eastern
branch of Leadenwah Creek. Following significant droughts during 1986 and 1987, an
extensive retention pond system was constructed in agricultural fields at the TRT Site.
During 1988, this retention pond system effectively drained and retained approximately 50%
of the agricultural runoff at the TRT Site. The water in the retention ponds was used for
drip fertigation (drip irrigation of sand filtered, fertilized water) in tomato Fields. During
1988, 1989 and 1990, certain portions of agricultural fields were planted with a rye grass,
vegetative strip buffer. The farmer at the TRT Site has adhered rigorously since 1987 to
suggested BMP and utilized recommended IPM strategies. The KWA Site has extensive
agricultural fields with a natural vegetative buffer strip surrounding each field. Runoff is
channeled into an extensive ditch network which discharges directly into the headwaters of a
small unnamed tidal tributary. The farmers at the KWA Site do not utilize BMP or IPM
practices such as those employed at the TRT Site.
Field Toxicity Tests
Field toxicity tests were conducted during May-June, 1989 and May-June, 1990.
During 1989, field toxicity tests were conducted from 25 May - 27 June, 1989 at the
CTL, TRT and KWA Sites. In tests conducted from 25 May - 15 June, 1989, a total of five
test species were utilized including: 1) adult grass shrimp (15-35 mm P. pugio); 2) adult
mummichogs (35-100 mm F. heteroclitus); 3) juvenile penaied shrimp (35 - > 100 mm P.
aztecus and P. setiferus); 4)adult mysid shrimp (M. bahia); and 5) juvenile sheepshead
minnow (<20 mm - C. variegatus). In tests conducted after June 15, 1989 only four species
- sheepshead minnow, grass shrimp, mummichogs, and juvenile penaied shrimp were used.
All toxicity tests with grass shrimp, mummichogs, juvenile penaied shrimp and mysid shrimp
were of 96 hour duration. Field toxicity tests with juvenile sheepshead minnow were either
14 [Group 2 (1-15 June 1989)] or 16 day [Groups 1 (24 May-9 June, 1989) and 3 (11-27
June 1989)].
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During 1990, field toxicity tests were conducted from 24 May - 23 June, 1990 at the
CTL, TRT and KWA sites. In tests conducted from 24 May - June 1990, a total of six
species were utilized including: 1) Adult grass shrimp (15-35 mm P. pugio); 2) adult
mummichogs (35-10(7 mm F. heteroclitus); 3) juvenile penaied shrimp (35-100 mm P.
aztecus and p. setiferus); 4) adult mysid shrimp (M. bahia); 5) juvenile sheepshead minnow
(<23 mm C. variegatus) and 6) juvenile tide water silversides (13-22 mm - Menidia »
berylina). Mysid shrimp tests were only conducted through 13 June and silversides through
June 17, 1990. In tests conducted after June 17, 1990 only four species - grass shrimp,
penaied shrimp, mummichogs, and sheepshead minnow were used. All field tests with grass
shrimp, mummichogs, juvenile penaied.shrimp, and mysid shrimp were of 96 hour duration.
Field toxicity tests with juvenile sheepshead minnow (6-9 days) and juvenile menidia (4-9
days) were of variable duration due to additional growth experiments (C. variegatus) and
problems in field survival (M. berylina).
All grass shrimp and penaied shrimp were collected by seine at the CTL Site on
Leadenwah Creek. Mummichogs were collected by minnow trap from an unnamed tidal
tributary of Bohicket Creek, near Cherry Point Landing. Mysid shrimp, juvenile sheepshead
minnows and silversides were taken from existing laboratory stocks at the Gulf Breeze
Environmental Research Laboratory, Pensacola^Florida.
Each test species was deployed in different types of cages during field toxicity tests as
follows: 1) grass shrimp (rectangular, plexiglass cages - 25 x 5.3 x 5.3 cm with 2 mm nytex
screen with styrofoam floats); 2) mummichogs (rectangular, plexiglass cages - 51.5 x 12.5 x
11.7 cm with nytex 2 mm screen with and without styrofoam floats); 3) penaied shrimp
(rectangular plexiglass cage - 51.5 x 12.5 x 11.7 cm with 2 mm nytex screen); 4) Mysid
shrimp (circular, nalgene plastic cages - 8.25 cm diameter x 13.97 cm height with 0.45 ^
nytex screen); 5) sheepshead minnow (circular, nalgene plastic cages - 8.25 cm diameter x
13.97 cm height with 1.000 \i nytex screen) and 6) silversides (circular, nalgene cage - 11
cm diameter x 7 cm height with 1 \i nytex screen). All caged organisms were placed in a
larger wire cage to exclude predators. During each toxicity test, a total of three replicate
cages/species, 10 organism per replicate (n=30/species/site) were deployed at each field site.
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During the 1989-90 field toxicity testing period, new animals were deployed every 96
hours with the exception of juvenile sheepshead minnows, which were deployed every seven
days.
In each field toxicity test, the following parameters were recorded daily: 1) percent
mortality and survival; 2) water temperature (YSI Model 64 oxygen meter and Taylor Min-
Max thermometers - °C); 3) Salinity (A.O. refractometer-ppt); 4) dissolved oxygen (YSI
Model 64 oxygen meter - mg 0Z/L); 5) pH (Hanna Model 0064 pH meter; 6) Rainfall
(cm/day); and 7) surface water samples (4.25 L) were collected for pesticide residue analysis.
In addition, surface sediments and adult oysters (Crassostrea virginica) were collected weekly
for pesticide residue analysis. During rain events, additional grab and composite water
samples were collected for residue analysis at the CTL, TRT and KWA Sites. Composite
water samples (250ml/20 minutes) were normally collected over a 12 hour period, using a
Sigma water sampler. Water samples were solvent (dichloromethane) extracted in the field
and refrigerated until analyzed. Oyster samples were cleaned, shucked, placed in solvent-
cleaned glass jars and frozen until analysis. Sediment samples were also placed in solvent-
cleaned glass jars and frozen until analyzed. All samples were analyzed by capillary column
gas chromatography using methods outlined by EPA (1980).
During the May - June 1989 sampling period, continuous (every fifteen minutes)
measurements of water depth (m), salinity (ppt), conductivity (mmhos), water temperature
(°C), pH, and dissolved oxygen (mg 07/L) were recorded at the two Leadenwah Creek Sites
using a Hydrolab in-stream water monitor. Similarly during May - June 1990, Hydrolabs
were deployed at the CTL, TRT, and KWA Sites. Continuous (every fifteen minutes)
measurements of water depth (m), salinity (ppt), conductivity (mmhos), water temperature
(°C), pH, and dissolved oxygen (mg 02/L) were recorded.
To ensure uniformity among all test animals used in field toxicity tests, quality
control, static 96 hour Quality Assurance (QA) toxicity tests were conducted weekly on each
test species (grass shrimp, penaied shrimp, mummichogs, and sheepshead minnow - 1989;
and grass shrimp, penaied shrimp, mummichogs, sheepshead minnow and silversides - 1990).
The Emulsifiable Concentrate (EC) of endosulfan (24% AT) was used as the reference
toxicant. Exposure concentration varied among test species (grass shrimp and penaied shrimp
- nominal 0.01, 1.00, 1.15 and 2.50 Mg/L; mummichogs - nominal 0.01, 1.15, 2,50 and 5.00
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Mg/L; and silversides and sheepshead minnow - 0.01, 1.15 and 2.50 /ig/L). Test species
were also exposed to the carrier (0.1% acetone) in seawater. During 1989, tests were
conducted at salinities_ranging from 25 to 34 ppt, water temperature of 21.2 - 28.3 °C,
dissolved oxygen levels ranging from 2.00 - 8.80 mg/L, and pH ranging from 7.20 - 8.10.
During 1990, tests were conducted at salinities ranging from 30-34 ppt, dissolved oxygen
levels ranging from 1.40 - 7.60 mg/L, water temperatures ranging from 19.8 - 26.2°C, and
pH ranging from 7.3 - 8.2 Water changes were made daily. Pesticide concentrations were
based on nominal dilutions of a measured stock. All tests were conducted at ambient light:
dark cycles (~ 14:10 L:D cycle). Quality control tests were not conducted on mysid shrimp,
since test groups were taken from existing laboratory stocks at GBERL.
Chemical Analysis of Environmental Samples
Seawater Samples
During field toxicity testing, seawater samples (4.25 L) were collected daily and at
prescribed sampling intervals, following significant rainfall (> 1.27 cm/24 hr) events, for
pesticide residue analysis at all study sites. Samples were placed on ice, transported
immediately back to the lab and processed in the following manner:
(1) Initially, all samples were thoroughly shaken by hand and 750 ml of sample was
decanted.
(2) Two 20 ml aliquots of samples were taken from this 750 ml portion and filtered
through a preweighed filter (pore size 0.70um). Each filter was placed in aluminum foil and
frozen. Later each filter was dried in a drying oven at 65 °C for 24 hours and rewelghed.
Contents on each filter represented total filterable solids residue (TFR-g/L).
(3) Three hundred ml of dichloromethane was added to the remaining 3.SOL of sample.
Each sample was then shaken thoroughly by hand and placed on a jar mill for a minimum of
two h. The solvent layer was then placed into a 2000 ml separately funnel and decanted into
a solvent-cleaned, 500 ml, Teflon-capped amber glass bottle. Each sample bottle was sealed
and stored in a refrigerator for subsequent analysis.
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(4) Each extracted water sample was placed into a glass round bottom flask and flash
evaporated at 30°C to a final volume of 5-10 ml.
(5) Florisil columns were used to remove interfering, biogenic compounds from each
sample. During the 1989-90 studies, microflorisil columns were prepared by placing 0.5 cm
(0.70 g) of florisil and 0.50 cm of sodium sulfate into a solvent cleaned pasteur pipette
(14.40 cm). The concentrated sample was then added to the microcolumn and eluted with 7
ml of a 20% (by volume) ethyl acetate in hexane solution. Each sample was then evaporated
under a stream of dry nitrogen gas (ultra high purity 99.999%) to near dryness and diluted up
to 1 ml with isooctane.
(6) All samples were analyzed by capillary column gas chromatography (CC-GC) in
accordance with procedures outlined by EPA, (1980), using a Hewlett-Packard GC (Model
5890A) with an electron capture detector and a BP-1 bonded phase silica based capillary
column (25 m). Samples were injected in the splitless mode. Helium was used as the carrier
gas at a flow rate of 1 ml/min. Peak heights areas were determined by a Hewlett-Packard
(Model 3393 A) integrator. The injector temperature was 220°C and the detector
temperature was 300°C for the analysis of each insecticide. A temperature program with a
20°C/min ramp from 90-200°C and then 10°C/min ramp from 200-290°C with a 15 min
hold at 290°C was used.
(7) Peak heights and retention times from each sample were compared with analytical
standards (fenvalerate, endosulfan I, endosulfan II and endosulfan cyclic sulfate; ethyl
parathion; methyl parathion, and azinphosmethyl) obtained from the U.S. EPA Pesticides-
Industrial Chemicals Repository, Research Triangle Park, Raleigh, NC for compound
identification and quantification. Insecticide concentrations are reported in ng/L for summary
tables and ^g/L for figures depicting measured insecticide concentrations.
Water samples from all laboratory toxicity tests- were also analyzed in a similar manner as
that described for field toxicity tests.
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Sediment Samples
During 1989 field xoxicity testing, sediment samples were collected at discrete sampling
periods (weekly) for pesticide residue analysis. Sediment samples were collected at each
field site by scraping soils from the top 20-30 mm of surface and placing sediments into a
solvent cleaned, glass jar (0.5L)> Following collection, all samples were sealed and frozen
until analyzed.
At the lab, each sample was thawed and three, 20 mg aliquots of wet sediment were each
placed in an aluminum tin and dried at 60°C for 24 hour in a dry air incubator. At the end
of 24 hour samples were reweighed to determine % water content of each sample as
described by EPA (1980). The % water content estimate for each sample was used in the
final calculation of sediment pesticide concentration/g sediment.
Next, 50 g of wet sediment from each sample was placed in solvent-cleaned freeze drying
flask. Each sample was then freeze dried at -50°C for 24 h at 0.001 mm Hg pressure.
Following freeze drying, each sample was broken into a powder with a mortar and pestle and
then placed in a solvent-cleaned cellulose thimble (Whatman-43 x 123 mm). Each sample
was then soxhlet extracted for 24 h with dichloromethane (250 ml volume) at a rate of 1
cycle volume/h. Next, samples were flash evaporated to 5-10 ml volume and then blown to
dryness under a dry stream of nitrogen and 1 ml hexane (nanograde) was added. Then 0.5
ml of Hg (triple distilled) was added to each sample. Samples were vortexed for 30 sec,
allowed to settle and vortexed again for another 30 sec. The precipitate in each sample was
allowed to settle and the clean solvent (hexane) layer was decanted.
Samples were then placed in a florisil column and eluted into three separate fractions with
15 ml of 6, 15, and 50% ethyl ether in hexane, respectively. Each of the three fractions was
evaporated under a dry stream of nitrogen (ultra high purity - 99.999%) to one ml volume
and were then diluted up to 5 ml with isooctane.
Sediment samples were then analyzed by capillary column chromatography in accordance
with EPA (1980) methodologies using the same procedure used in the analysis of water
samples. Results are reported in jtg/kg (dry weight) for each insecticide detected.
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Qvster. Shrimp and Fish Tissue Samples
Adult oysters (Crassostrea Virginia) were collected at discrete sampling periods (weekly)
for pesticide residue analysis during field toxicity tests. During 1989, all oysters were
initially collected from a reference site located at the mouth of Leadenwah Creek on 24 May,
1989. Oysters were transported back to the lab, an initial tissue sample was taken, and the
remaining animals were deployed at each site (CTL, TRT, and KWA Sites) in plastic trays
(92 x 61 x 15 cm).
During 1990, oysters were initially collected on 24 May 1990 from a reference site
located at the mouth of Leadenwah Creek. Oysters were transported back to the lab, an
initial tissue sample was taken, and the remaining animals were deployed at the CTL, TRT
and KWA Sites in plastic trays (92 x 61 15 cm) on 25 May 1990.
Trays were affixed to the creek bottom at each site in the mid-lower intertidal zone with
reinforcing rods (rebar for concrete) and subsequent samples were collected weekly
throughout the study. Each collected sample was immediately cleaned, shucked into a
solvent-cleaned, glass jar using a solvent-cleaned, oyster knife. Samples were sealed and
frozen until further analysis. In addition to oyster tissues, penaied shrimp, grass shrimp,
blue crab, mullet and mummichogs were collected during field toxicity tests, when significant
mortality occurred. These samples were placed in solvent-cleaned glass jars, transported
back to the lab on ice, and frozen until further analysis.
Quality Control
During field toxicity tests, weekly water, sediment and oyster tissue samples from the
CTL Site were spiked with endosulfan, fenvalerate, azinphosmethyl and.methyl parathion to
determine "spiked" recovery efficiencies for the various extraction and clean up methods used
with each sample procedure. Results were reported in ng/L (water) and Mg/kg (sediment and
tissue) and as % recovery efficiency (water, sediments and tissues).
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Oyster Field Studies, 1989-90
The estuarine habitat is a very complex, dynamic and sensitive ecosystem. These tidal
creeks support many-indigenous species of ecological, commercial and/or recreational
importance. They also serve as nursery grounds and refuges for many ecologically,
recreationally and commercially important species of fish and shellfish (Bearden 1982; Cain
and Dean 1976; Hampton 1987; Patterson 1986; Scott et al. 1986). One of the most
important indigenous species of commercial, recreational and overall ecological importance is
the American oyster, Crassostrea virginica (Gmelin), which may be sensitive to inputs of
toxic chemicals due to its ability to bioconcentrate and bioaccumulate pollutants. A dominant
secondary producer in the Spartina marshes of South Carolina and the southeastern United
States, as well as a prime candidate for aquaculture, is the intertidal filter-feeding bivalve, C.
virginica. C. virginica is also a very important species in the healthy ecological functioning
of estuarine creeks as it provides hard substrate habitat for many other estuarine species.
Oysters filter and ingest a mixture of inorganic particles, phytoplankton and detrital
complexes. The energy gained through digestion of algae and detrital complexes may be
directed to any of a suite of physiological processes within the oyster. Both exogenous (i.e.,
food quality and quantity, temperature, etc) and^endogenous (i.e., age, size, reproductive
state, nutritional status, disease, etc.) factors will also affect the energy partitioning of the
organism. The degree of fecundity and growth within an oyster population, then, results
from a complex balance between gametic production, somatic production, and metabolism
(respiration and excretion) for the energy absorbed from the oysters' food ration. This
relationship has been described by Winberg (1960) as:C-F = A = R + U+PorP = A
- (R + U). Where: C = food energy consumed, F = energy lost as feces, A = energy
absorbed from the food, R = energy respired, U = energy excreted, P = energy into
somatic and gametic production. Energy available for growth and reproduction has also been
referred to as "scope for growth" (Warren and Davis, 1967). Positive scope for growth
values indicate energy is available for production of gametes and somatic growth, while
negative values are indicative of stressful conditions requiring utilization- of body reserves.
Since direct measurement of growth and fecundity (production and viability of gametes) can
be difficult in bivalve species, scope for growth has become a useful index for these
physiological parameters (for a review, see Bayne et al., 1985). The use of this index for
the measurement of the physiological performance of mussels subjected to environmental
stress has been discussed by Bayne et al. (1979, 1982) and Widdows et al. (1981) a, b).
25 .
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Bivalves with reduced scope for growth due, for example, to decreased food availability,
may produce fewer eggs which have low nutrient reserves (Sastry, 1975; Bayne et al., 1978).
The result is a reduced probability of successful larval metamorphosis and setting of spat
(juvenile oysters), resulting in population instability. Lipid content in bivalve eggs and larvae
seems to be related to parental scope for growth. Bayne et al. (1978) reported eggs produced
by Mytilus edulis having a negative scope for growth were smaller and had less organic
matter per egg than those produced by mussels with a positive scope for growth. Adult
Ostrea edulis maintained on low food rations released larvae having slower rates of growth
and lower lipid content than larvae released from adults maintained on high rations (Helm et
al., 1973). The resulting extension of the larval period may well increase mortality, thus
reducing recruitment potential. Crassostrea virginicd having a positive scope for growth
need not deplete its glycogen reserves, which are used, in pan, as a source of lipid in eggs
(Gabbott, 1976; Holland, 1978).and in part as a source of energy for adults during winter
for routine maintenance and disease fighting metabolism. The oysters would then be more
likely to produce gametes with adequate nutrient reserves which, by extension, leads to viable
larvae, greater over-wintering survival of adults, and a replenishing or increase in the
population as a whole. Scope for growth, as a measure of die energy status of an oyster,
measures not only level of stress within oysters but the potential for growth and fecundity as
«
well.
Two additional indices of stress, condition index and O:N ratios, are also useful
measurements. Scope for growth may be viewed as a measure of energy status and
fundamental adaptive responses, while condition index is a measure of alterations in die
nutritive status of an organism, and O:N ratios measure alterations in the balance between
catabolic processes. Organisms under stressful conditions (i.e., parasitic infection), which
cause their metabolic requirements to increase above normal levels, tend to utilize nutrient
reserves in order to meet the elevated metabolic demands. The ratio of oxygen utilized to the
amount of ammonia excreted is an index of the relative use of protein in an organism's
metabolism, with lower ratios indicative of greater use of protein relative to carbohydrate and
lipid. As an example, Widdows et al. (1981) found that mussels transplanted along a
pollution gradient in Narragansett Bay demonstrated a decline in O:N from 75 to 30 with
increasing contamination.
During 1989 an assessment was made using a modification of these various bioenergic
metabolic approaches to evaluate the effects of nonpoint source agricultural pesticide runoff •
26
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on the American oyster at the CTL and TRT Sites. During 1990, this assessment was
repeated at the CTL and KWA Sites. The pesticides of concern during this 1989-90
assessment were those used during the growing season in the area and included: methyl
parathion, azinphosmethyl, endosulfan and fenvalerate. This assessment consisted of
measuring pesticide uptake rates and resulting lethal and sublethal, physiological effects in
adult oysters (by use of whole animal respiration, nitrogen excretion rates, O:N ratios,
condition and gonadal index) using methods described by Scott et al., 1990 and Crosby
(1988). In addition, larval settlement, size-frequency distributions (1989-90) and Perkinsus
marinus infections were also measured. This integrated approach generally follows the
concepts of Sastry and Miller (1981) who suggest the use of multi-media sampling and
linkage to biological species for definition of effects or impacts.
Oyster Collection and Transplantation
During the 1989 study adult American oysters, Crassostrea virginica (Gmelin), of legally
harvestable size (^7.5 cm in height) were collected on 24 May 1989 by hand from the mid-
intertidal zone of a well-established, healthy reef near the mouth of Leadenwah Creek. The
total number of oysters harvested were split into three groups and placed into cages for re-
laying. Plastic trays were used as cages (dimensions: 92 cm length x 61 cm width x 15 cm
height). The interior walls of the trays were lined with nylon screening (1.00 mm) to
preclude easy access to the cages by predators.
Each group consisted of three cages in which approximately 100 oysters were placed in
two of the three. One cage held only 30 oysters, each marked with an individual
identification code. All cages were anchored to the mid-intertidal zone by reinforcing rods
and supported above the mud by concrete blocks. The two cages of 100 oysters each were
used as the sample pool for physiological and chemical measurements. The cage of 30
oysters served for in situ mortality determinations.
During 1990, three groups of oysters were again deployed at the CTL, TRT and KWA
field sites. Each group consisted of three cages in which approximately 100 oysters were
placed. These three cages were used as a sample pool for physiological and chemical
analysis. All cages at each site were deployed in the mid-low intertidal zone by placing each
cage on small cement blocks (to provide support above the mud) and were then anchored in
place by reinforcing rods.
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Phvsicochemical Measurements
A Fisher minimum-maximum thermometer was affixed to one transplant cage per site to
record overall temperature extremes (in degrees Celsius °C) between each sampling visit.
Discrete salinity measurements were made at each site during each sampling period using an
American Optical refractometer and reported in parts per thousand (ppt). Water temperature
at the time of sampling was recorded from each site using a standard stick thermometer in
°C. Rainfall in cm per day was recorded daily at each site using a standard rainfall
collection gauge.
Chemical Analyses
During 1989, analyses for methyl parathion, endosulfan, azinphosmethyl and fenvalerate
were conducted on a composite sample of 15 oysters per site at exposure days 0, 6, 13, 23,
32 and 63.
Following collection, all oysters were returned to the laboratory and washed with tap
water using a spray nozzle and brush to remove all external dirt and debris. They were then
shucked into glass containers with tinfoil lined caps that had been washed with laboratory
detergent, rinsed four times with deionized water and then three times with pesticide-grade
solvents (acetone, petroleum ether). The tin foil-lined caps were prepared in the same way.
All shucked oysters were stored at -15°C until laboratory analyses began. Analyses were
conducted by gas chromatography with all analytical procedures following USEPA Methods
(1980).
Physiological Analyses
During 1989, 10 to 15 oysters per site were sampled for respiration, nitrogen excretion,
O:N ratios, condition and gonadal index measurements at exposure days 0 (17 May 1989),
28, 54 and 72 (28 July 1989). During 1990, 10 to'15 oysters per site were similarly sampled
for respiration, nitrogen excretion, O:N ratios, condition and gonadal index measurements at
exposure days 0 (15 May 1990), 30, and 60 (14 July 1990). After washing as described for
the chemical analyses, all fouling and commensal organisms were removed, oysters were
immersed in a chlorine bleach-tap water solution for 5-10 minutes, thoroughly rinsed and
28
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placed in a flow-through respirometer chamber. Following an appropriate period of
acclimation, chambers were sealed off and measurements of whole animal respiration
determined. Following respiration determinations, each oyster was placed in an acid clean
container for one hour at room temperature (20°C). Water samples were then collected,
fixed with phenol, and stored in a refrigerator for subsequent nitrogen ammonia
determinations (Solarzano, 1979). Each oyster was then sacrificed for condition index
«
following the method of Lawrence and Scott (1982). This technique measures the cavity
volume of shell by subtraction of the dry shell weight (without soft tissues) from the total
weight (shell and soft tissues). The resultant cavity volume is then used to calculate the
condition index (CI) by utilizing the following formula:
CI = [total dry body weight (g) / cavity volume ml] x 100
The gonadal tissues were removed by cutting just above the dark area of the digestive
diverticulum and then along the adductor muscle. These tissues were dried at 60°C for 48
hours, as were the remaining soft tissues for the CI analysis. The gonadal index (GI) was
then calculated by using the following formula:
GI = [dry gonad weight (g) / total dry body weight (g)] x 100
Field Mortality Analyses
During 1989 and 1990, the level of in situ mortality was monitored using 30 oysters
maintained in one cage. After harvest from the endemic area, 30 oysters per site were
cleaned by scrubbing with a bristle brush to remove mud and debris. After allowing to air
dry at room temperature for 30 minutes, a unique identifying code from Kl to K30, (KWA
Site), Tl to T30 (TRT Site), CI to C30 (CTL Site) was assigned to each oyster and painted
onto the cleaned shell using fingernail polish.
Prior to each field visit, three replicates of 10 coded oysters per replicate were formed by
random number selection from the pool of 1 to 30 at each site. Using these three replicate
groups at each station, mortality trays were then checked for dead oysters and any such
mortality or otherwise missing oysters were recorded. The identifying codes of missing or
dead oysters were not carried into subsequent replicate formulations. Thus, by the end of the
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study, there were less than 10 oysters per replicate. The level of mortality was reported as
the mean of the percent from each replicate.
Perkinsus Marinas Analyses
During 1989-90, rectum and labial palps were dissected from the oysters used in the
physiological measurements and infection incidences and intensities of Perkinsus marinus
were estimated by culturing those tissues in thioglycolate medium fortified with dextrose
(Ray, 1966). Mycostatin and chloramphenicol were added to the medium to inhibit bacterial
growth and tissue putrefication. After incubation at room temperature (up to six months),
tissues were stained with Lugol's Iodine Solution and hypnospores were counted using a
microscope at 45X magnification.
Densities were determined by averaging the counts observed for three non-overlapping
areas (4.71 mm2) in each tissue. A number code (NC) was assigned to these averages to
facilitate interpretation (Scott et al., 1983) as follows:
No. of Intensity Class Number Code
hypnospores
0 Negative 0
1-5 Very Light 1
6-10 Light 2
11-30 Lightly Moderate 3
31-300 Moderate 4
301-1000 Moderately Heavy 5
1001-3000 Heavy 6
> 3000 Very Heavy 7
Spat Settlement
Recruitment of oyster larvae to the existing reefs at each site was investigated by
assessing larval settlement.
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During 1989, spat plates, consisting of corrugated PVC pipe (1.90 cm diameter x 183 cm
height) were deployed on 3 June, 1989 in three transects, each 30 cm apart, throughout the
upper to lower imertidal zone at the CTL, TRT, and KWA Sites. Along each transect,
collectors were spaced at 183 cm intervals from the upper to lower intertidal zone. A total
of nine spat collectors were deployed at each site. Initially steel reinforcing rods were driven
into the sediment substrate and then the larger corrugated PVC pipe was placed over each
reinforcing rod. Spat collectors were collected 365 days later and analyzed for settlement of
both oysters and barnacles by enumerating the number settled (#/cm2) throughout the entire
spat collector at respective vertical elevations (0-15, 15-30, 30-45, 45-60 and 60-75 cm above
the sediment surface). In addition, oysters were measured for shell height (cm) and weight
(g). Barnacles were identified to species, divided into two distinct size classes (1-5 mm and
6-18 mm) and enumerated.
Statistical Analyses - Ovster Studies
Tests for significant differences in means between sampling sites for selected parameters
were made using both parametric and nonparametric procedures. The Wilcoxon Rank Sum
Test and the Kruskal-Wallis one way Analysis of Variance Test (SAS 1985; Wilcoxon and
Wilcox 1964) were used. These nonparametric techniques were selected because of the usual
non-normal distribution of environmental data and the poor transformation response of the
data to a normal distribution (Gertz 1978; Wright et al. 1985). Additional parametric
procedures (T-Test and ANOVA) were used when possible.
Laboratory Toxicity Tests
Earlier studies by Scott et al., (1990) have established 96 hour LC50 and 6 hour pulsed
dosed LC50 values for grass shrimp (P. pugio) and mummichogs (F. heteroclitus) exposed to
azinphosmethyl, acephate, endosulfan, fenvalerate, and various insecticide mixtures (i.e.,
azinphosmethyl-endosulfan, endosulfan-fenvalerate, azinphosmethyl-fenvalerate).
Additionally, extrinsic (salinity) and intrinsic (life stage) factors were evaluated. The results
of these studies indicated that these factors enhanced the traditional 96 hour LC50 toxicity at
20 ppt salinity by no more than a factor of 2.86. Given the extensive data base reported by
Scott et al. (1990) additional 96 hour laboratory toxicity tests were not conducted during this
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study. Emphasis in the laboratory was rather placed on sublethal effects of azinphosmethyl
on the mummichog, Fundulus heteroditus.
Effects of Azinphosmethyl on Brain AChE Activity in Mummichogs
Laboratory Phase
Azinphosmethyl was selected for laboratory study because of its known occurrence in
runoff at field sites (TRT - 1986, 1987 and KWA - 1988) and its mode of toxicity to the
mummichog (Inhibition of AChE). Laboratory experiments were conducted to determine the
effects of short-term azinphosmethyl exposure on brain AChE activity in the mummichog.
Two groups of 12 adult (> 35mm) mummichogs (F. heteroditus), collected by minnow
trap from the Cherry Point (CP) collection site were exposed to 2.4 /xg/L of EC
azinphosmethyl for 24 hours. This exposure concentration of azinphosmethyl was selected as
the result of range finding studies which suggested this concentration would produce 80%
brain AChE inhibition after 24 hours of exposure. Two additional groups of 12 fish were
exposed only to the carrier (acetone) and served as controls.
Exposure duration was 24h in 5L glass aquaria at 20 ppt salinity. Temperature in the
aquaria was ambient and ranged from 20 - 21 °C. Azinphosmethyl exposure concentrations
ranged from 0.24 to 3.90 /xg/L. Fish utilized in the bioassays ranged in length from 45 - 80
mm. A total of six fish were exposed per concentration. Total exposure volume was 4.8L.
An additional group of six fish was maintained as a control. All test concentrations and the
control contained the same carrier (acetone) concentration. Following 24h of exposure, all
animals were removed from the exposure media and sacrificed. The brains were removed,
wrapped in aluminum foil and stored at -20°C until analyzed for AChE activity as previously
described. The level of AChE inhibition produced in each of the azinphosmethyl
concentrations was then determined based on a comparison to the level in control animals.
The results of these tests were then used to calculate a 24h ECjo for azinphosmethyl-induced
AChE inhibition.
Following the 24 hour exposure whole animal respiration rates (/xgOj/g/h) were
determined for fish in one of the treatment groups and one of the control groups. In
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addition, samples were collected for ammonia determinations. Both of these determinations
were made using methods described by Scott et al., 1987.
Respiration rates (/igO2 consumed/g of tissue/h) were determined by measuring oxygen
consumption with a dissolved oxygen (DO) meter (YSI Model 58). BOD bottles were filled
with high salinity (20 ppt), filtered (0.45^m) seawater and immersed in a water bath kept
constant at 20°C. Each BOD bottle was allowed to acclimate for one hour, then initial DO
levels and time of measurement were recorded prior to the introduction of test animals into
each BOD bottle. Additional BOD bottles containing 300 ml of filtered seawater without test
animals were examined to determine the effects of aerobic activity other than that of test
animals. Final DO determinations were made after 1 hour of immersion in the water bath.
At the end of each final respiration determination, a 20 ml sample of seawater was
collected from each BOD bottle, preserved with two (2) ml of phenol (80%) and promptly
refrigerated until later analysis for ammonia. After ammonia samples were collected, fish
were removed from the BOD bottle, measured for standard length (mm), sexed, placed in
preweighed aluminum pans, and dried for at least 72 h at 90°C before final dry weight
determinations (g) were taken. Results were expressed as /xgOj/g/hr).
All water samples collected for ammonia analysis were analyzed within two weeks of
collection using the procedure described by Solarzano (1969). An Orion Scientific Auto
Analyzer (Model 140) was used for ammonia determinations of samples and standard curves.
Results were expressed as \i% of ammonia excreted per gram of dry weight per hour
(MgNH4/g/hr).
Bayne (1975) and McKenney (1982) have suggested that alterations in balance between
the catabolism of carbohydrate, protein, and lipid substrates may be useful as a measure of
stress in aquatic organisms. Alterations in the ratio between oxygen consumption and
ammonia excreted has been used to assess stress in.aquatic organisms. Oxygen/nitrogen
(O/N) ratios were calculated for all mummichogs exposed to azinphosmethyl and in controls
by dividing the oxygen consumption (/ig atoms) by the nitrogen excretion rates (in pig atoms)
for each fish.
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Fish were also sacrificed and the brains removed and stored at -20°C for subsequent
determination of brain AChE activity using a modification of the method described by Ellman
et al., 1961.
The remaining control and treatment fish were transferred to clean water and held for 8
days. Water was changed daily and fish were fed during their depuration period.
At the end of 8 days, respiration measurements were made, samples collected for
ammonia determinations and the fish sacrificed and the brains removed as previously
described.
From these determinations, mean fish respiration rate (nitrogen excretion rate), and 0/N
ratios were determined and statistically compared using ANOVA. An alpha <0.05 was the
minimum level of significance used. Both temporal (24 hour exposure versus 192 hour
depuration: azinphosmethyl and control groups) and between group (control versus
azinphosmethyl groups) comparisons were made.
BIOMARKER STUDIES
Toxicity studies with pesticides in fish and other aquatic organisms generally assess only
the acute toxicity of individual insecticides. In order to accurately predict the ecological
impact of insecticide exposure in the environment, additional information concerning the
sublethal effects of insecticides in aquatic organisms is needed. One goal of this project was
to evaluate specific sublethal toxic responses in aquatic organisms exposed to agricultural
insecticides.
Organophosphorus (OP) insecticides are believed to produce toxicity by severely
inhibiting the enzyme, acetylcholinesterase (AChE). This inhibition causes an accumulation
of acetylcholine at the post-synaptic membrane which leads to excessive-activity at the
synapses followed by a blockade of nervous impulses (O'Brien, 1967).
Earlier studies by Fulton 1989 also reported in Scott et al., (1990), have indicated the
effect of azinphosmethyl on brain AChE enzyme. The 24 hour EC50 was 0.81 /ig/L.
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Additional statistical analyses of various intrinsic (sex) and extrinsic (salinity and the presence
ot more than one insecticide) were conducted to evaluate their importance. Results indicated:
1) ANOVA analysis indicated that salinity and brain AChE were significantly correlated.
Additional, multiple mean comparisons of high and low salinity indicated there were no
differences in brain AChE levels at high (20 ppt) and low (5 ppt ) salinities.
2) ANOVA analysis also indicated there was no significant interaction between
azinphosmethyl and other insecticides (endosulfan); and
3) ANOVA analysis further indicated no significant interactions between azinphosmethyl
exposure and sex associated with brain AChE levels.
These extensive laboratory studies in fish were then compared with measured levels of
brain AChE inhibition in field exposures to azinphosmethyl. Results indicated excellent
agreement between field and laboratory results.
This present study was designed to further examine the effects of sublethal insecticide
exposure on the level of brain AChE activity under field conditions. Additional comparisons
of whole animal AChE levels in oysters were made to compare different species. The goal
.of this research was to enhance our knowledge af the utility of AChE inhibition as a
biomarker of nonpersistent pesticide exposure in the field.
Field Exposure Phase
Study sites for the field exposure tests were those previously described in the Materials
and Methods section. Field exposure tests were conducted during May - June 1989 - 90.
During 1989 and 1990, 96h field exposure tests were conducted during the vegetable growing
season at the TRT, KWA and CTL Sites.
F. heteroclitus were collected using a minnow trap at the Cherry Point (CP) collection
site (Figure 1). All animals utilized in the field exposure tests were between 45-100 mm.
Animals were deployed by placing them in rectangular plexiglass cages - 25 (L) x 5.3 (W) x
5.3 (H) cm with 2.0 mm nytex screen. Ten animals were deployed per cage. A total of 30
animals were deployed in three cages at each site. All plexiglass cages were placed in larger
wire cages to exclude predators.
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At the end of the 96 hour field exposure, the animals deployed at each test site were
removed from the field and transported back to the laboratory in large, insulated coolers. At
the laboratory, animals from each exposure site were sorted by sex. A total of 15 - 20 fish
from each site were sacrificed. All animals were sacrificed within 12 hours of their removal
from the field. The bTains from these animals were removed, wrapped in aluminum foil (5
brains per sample) and stored at -20°C until analyzed for AChE activity.
During field exposure-tests, seawater samples were collected daily and at prescribed
sampling intervals following significant rainfall. These samples were collected and analyzed
using procedures described in the Materials and Methods Section.
Assay of AChE Activity
Brain AChE activity was determined using a continuous assay procedure modified from
Ellman et at., (1961). Each brain tissue sample was homogenized on ice with a TenBroeck*
ground glass homogenizer in 50 mm Tris-HCl buffer (pH = 8.1) at 20 mg/ml. Next, 250 n\
of this homogenate were added to a test rube containing 4.75 ml of the buffer. After
incubation for 15 minutes in a shaking water bath at 30 °C, 2.9 ml of the dilute homogenate
were added to a cuvet containing 100 pi of 0.87% 5, 5-dithiobis-(2-nitrobenzoic acid), the
color reagent. Finally, 15 v\ of 75 jxM acetylthtocholine, the substrate, were then added to
the cuvet which was covered by parafilm and inverted to mix. The absorbance was then read
continuously for 1 minute at 412 nm using a Bausch and Lomb Spectronic® 1001
spectrophotometer. Enzyme velocities were linear during the assay period. A minimum of
three subsamples were assayed for each brain tissue sample. In addition, a subsample
incubated with 10.0 ^M eserine was used to account for non-enzymatic, non-AChE
hydrolysis of the substrate. The protein content of the homogenate was determined using the
SigmaR assay procedure, a modification of the original Lowry method (Lowry et al., 1951).
Enzyme activity was calculated as nmol product formed min'1 mg protein "'.
Whole Body Insecticide Residue Analysis
Following significant runoff events, fish deployed in field bioassays were sacrificed for
whole animal analysis of pesticide levels.
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Tissue preparation, sample cleanup and insecticide quantification were performed using
che methods described by Bush et al., (1977) and Bush et at., (1978).
After thawing,_approximately-5g of fish tissue were placed in a high speed blender jar.
About 50 g of Na:SO4 and 300 ml of ethyl acetate were added, and the sample was blended
for 5 - 10 minutes. This homogenate was then filtered with suction through a 9-cm diameter
Buckner funnel fitted with a Reeves Angel glass filter paper into a 500 ml suction flask. The
filtrate was then transferred to a boiling flask and taken to dryness using a rotary evaporator
at 50°C. This extract was then made to 10 ml with ethyl acetate-toluene (75:25). Fat was
removed by gel permeation chroma tog rap hy (GPC) using an automated GPC AutoPrep®
1001.
Gas Chromatography (GC) analysis was conducted with a Tracer gas chromatograph
(Model 550) equipped with 6 ft x 0.25 inch coiled glass columns and Ni63 electron capture
detectors. Identification and quantification of insecticide residues was based on their
retention time and peak height relative to those of reference insecticide-standard solutions.
Bioconcentration factors for the insecticides in F. heteroditus exposed in field
exposures were calculated by dividing the insecticide concentration measured in the fish by
the insecticide concentration measured in stream at each field site. Detection limits were 50
^g/kg for azinphosmethyl and 10 ^g/kg for endosulfan I, endosulfan II and endosulfan
sulfate.
Ecotoxicological Studies of Macropelagic Organisms: 1989 - 1990
Block Seining
The east (TRT) and west (CTL) branches of Leadenwah Creek were sampled for
macropelagic fauna, using a block seining technique during 1989-90 (Figure 2). At each site,
three consecutive 50 m stretches of stream were permanently marked with metal stakes.
During each sampling period (monthly, February - May, 1989; bimonthly (every 14 days),
June - August, 1990; monthly, September 1989 - May 1990; bimonthly, June - August, 1990;
and monthly, September 1990 - March, 1991), a total of four seine nets (12 m x 1.5 m x 4
mm mesh) with 2 m long poles were anchored into the sediments for each 50 m interval. At
each site, lead lines for each net were pushed into the sediments and held in place by bricks.
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marsh v!- 9and :
Figure 2. Sketch of net deployments during block seining at the TRT Site.
38
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Another net was then pulled between each set of block nets and the contents of each seine
placed into a plastic bucket, preserved in a 10% buffered formalin, and scored for subsequent
taxonomic identification.
At the laboratory, each bucket was opened and the contents poured onto a large
diameter (10 mm) screen, washed and all detritus, vegetation and algae was removed from
each sample. Each sample was then placed back into the bucket and weighed wet to
determine total sample biomass (g/50m of stream).
Following biomass measurements, all large (> 10 cm) organisms (fish, shrimp and blue"
crabs) were removed from each sample, identified to- genus or species, counted (density/50 m
of stream), and the wet weight (g/species) of each species noted. The remainder of the
sample (organisms < 10 cm) was then identified to genus or species, enumerated (density/50
m of stream) and the wet weight (g/species) of each species noted. For small biomass
samples (<2000 g/samples) this procedure was followed for each sample, but for larger
biomass samples (>2000 g/sample) a subsampling procedure for small organisms (< 10 cm)
was used for each sample. Three randomly selected subsamples (500 g each) were taken
from each sample. Each subsample was sorted and identified to genus/species, enumerated
(density/species) and weighed wet (g/species). Each subsample was then multiplied by a
sample weight (g)/subsample weight (g) conversion factor to estimate the number of
organisms in each sample. Each of the three subsamples were then averaged and a mean
density (±95% CL) and biomass (±95%CL) calculated for each species. This subsampling
procedure was used primarily for grass shrimp (P. pugio), mummichogs (F. heteroditus),
juvenile spot (L. xamhurus), and bay anchovies (A. mitchilli). Following subsampling, the
remainder of each sample was poured back out onto the screen to identify any rare species
such as sheepshead minnow (Cyprinodon variegatus) which had been excluded by the
subsampling procedure.
From this procedure the following ecological parameters were calculated for each
sample:
(1) Total sample biomass (g/50 m of stream);
(2) Total grass shrimp (P. pugio) densities (#/50 m of stream);
(3) Total mummichogs (Fundulus heteroditus) densities (#/50m of stream);
(4) Total penaied shrimp (Penaeus aztecus, Penaeus seiiferus, and Penaeus
duorarum) densities (#/50m of stream);
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(5) Total blue crab (Callmectes sapidus) densities (#/50m of stream); and
(6) Total fin fish densities (#/50m of stream).
Ecotoxicological Sampling Statistical Procedures
Ecological parameters at each sampling site (CTL and TRT) were statistically compared
using:
(1) The Mann-Whitney or Wilcoxon Rank Sums Nonparametric Method for unpaired
samples (Wilcoxon, 1964).
(2) The Wilcoxon Rank Sum Nonparametric Method for paired samples (Wilcoxon,
1964). This procedure was found to be more appropriate since statistical analysis
generally indicated a slight statistical bias in our sampling method, as higher densities
and biomass were found in our most seaward net (#1 at the CTL and #3 at the TRT
Sites). Therefore the paired procedure was found to be more appropriate for our data
analysis.
In all samples, statistical comparisons were based upon a sample size of n = 6 (three
replicates at the CTL and three replicates at the TRT Site); however, in samples which were
subsampled, statistical comparisons of paired samples were possible with an n = 8 - 18. In
samples which were subsampled, additional statistical comparisons were made using the
Wilcoxon Rank Sums Tests for paired data as previously described. An alpha level of <
0.05 was chosen as a minimum significant level in comparisons of samples.
An additional method of statistical data analysis was used in which the ecological
parameter of interest at the TRT Site was subtracted from the same paired parameter at the
CTL Site. If the two sites were similar, the numerical difference between the matched
replicate pairs should approach a theoretical zero difference. Large deviations from this zero
difference may occur if the TRT Site was impacted by pesticide runoff (I.e. toxic effects,
behavioral avoidance, or both). During periods of significant pesticide runoff, increased
densities at the CTL Site may occur relative to the TRT Site, resulting in significant
deviations from the zero difference line. Statistical difference between CTL and TRT Sites
were based upon the Wilcoxon Rank Sum Nonparametric Method. An alpha level of < 0.05
was the minimum significance level used.
40
-------
Water Qualify Parameters
During each sampling event during 1989-90, temperature (°C), dissolved oxygen fmg
02/L) and salinity (ppt) were measured by a YSI oxygen meter (Model 64) and an American
Optics Salinity Refractometer. using methods described in Standard Methods (1982) for
calibration and sample determination. The pH was also measured using either an Orion
(Model 250) or a Hanna (Model 0064) pH meter.
Push Netting
While block seining has been shown to be an effective method for assessing population
level effects in the macropelagic community, it is extremely time consuming, labor intensive,
and produces large amounts of solid (animal carcass) and hazardous wastes (i.e. formalin
waste). Additionally, certain habitats may be extremely difficult to sample in this manner.
Alternative methods which are less labor intensive, time consuming and waste generating are
needed.
Results of block seining studies have clearly indicated the importance of P. pugio in
tidal creek habitats and their known sensitivity to various agricultural pesticides. Welch
(1975) in earlier studies has demonstrated the use of push netting to characterize field
populations of P. pugio in estuarine habitats of Texas, with densities varying from 20-300
animals/m2 being reported. Welch's method was modified and evaluated as a rapid census
method for grass shrimp in estuarine tidal creeks at the CTL, TRT, and KWA Sites from
March - December, 1990. At each site, three consecutive, 50m stretches of stream were
permanently marked with metal stakes and sampled monthly with a push net. (31 cm length x
30 cm width x 5 mm mesh). Two tows (by hand), one along each bank, were made in each
stream reach at or near dead low tide. Each tow was made going against the tide. The
contents of the two tows were pooled, placed in 10% formalin and stored for subsequent
taxonomic identification.
At the laboratory, each sample was opened and poured onto a large diameter screen (10
mm mesh), washed and all detritus, vegetation, and algae were removed. The remainder of
the sample was blotted dry and weighed wet to determine total biomass (g/m2). Following
biomass measurements, all crabs (primarily juvenile Callinectes sapidus), penaied shrimp,
(primarily Penaeus aztecus or Penaeus senferus), and small fish (primarily F. heteroditus]
41
-------
were identified to genus and/or species, enumerated (density/50m) and weighed wet (g/50m).
The remainder of each sample containing grass shrimp (P.pugio) was enumerated
(density/50m) and weighed wet (g/50m). From this procedure the following ecological
parameters were calculated for each sample:
1) Total sample biomass (g/50m) and total sample density for
all species (density/50 m);
2) Total grass shrimp (P.pugio) density (density/50m)and biomass (g/50m); and
3) Total non grass shrimp biomass (g/50 m) and density (density/50m).
These ecological parameters at each sampling site were statistically compared using
nonparametric procedures (Mann-Whitney, Wilcoxon Rank Sums and Kniskal-Wallis)(Zar,
1974; Armor, 1973; and Wilcoxon, 1964). An alpha level of <0.05 was chosen as a
minimum for significance levels when comparing samples between sites.
42
-------
RESULTS
I. Field Toxicity Tests
A. Daily Physicochemical Parameters
L. 1989. Daily Water Quality Parameters
Results of physicochemical water quality parameters measured daily at each site
during the 1989 field study are listed in Table 1. Mean daily seawater temperature
ranged from 24.3 - 31.0°C, averaging 27.64°C at the CTL Site. Similarly,
temperatures at the TRT Site ranged from 23.5 - 33.4°C, averaging 27.34°C. At the
KWA Site, temperatures ranged from 22.0 - 37.0°C, averaging 28.10°C. Statistical
analysis indicates that seawater temperatures at the three field sites were not
significantly different during May - June, 1989.
Mean salinities ranged from 16 - 33.2 ppt, averaging 28.59 ppt at CTL Site.
Salinities were lower at the TRT Site, ranging from 6.0 - 32.0, averaging 22.04 ppt.
Statistical analysis indicated that salinities were significantly (p £ 0.05) lower at the
TRT when compared to the CTL Site. The lower salinities at the TRT Site were the
result of freshwater inputs of agricultural runoff following major rain events. The low
salinities at the TRT Site occurred despite the fact that most of agricultural drainage
area's runoff was channeled into retention ponds.
Salinities at the KWA Site were even lower, ranging from 2-35 ppt, averaging
15.79 ppt. Statistical analysis indicated that salinities at the KWA Site were
significantly (p £ 0.05) different from both the CTL and TRT Sites. The much lower
salinities at the KWA Site demonstrates the significance that freshwater discharge from
agriculture may have on salinity. Of particular interest is the fact that salinity
comparisons between the TRT Site, an agriculture site with BMP, IPM, and retention
ponds, were significantly different from the KWA Site, an agricultural site without
BMP, IPM, and retention ponds. Although salinities at the TRT Site were significantly
lower than the CTL Site, the retention ponds there appeared to provide some degree of
protection by moderating fresh water inputs.
43
-------
TABLE 1. Summary of physicochemical water quality parameters measured at field sites during (he 1989 Held study
Pooled means with different letters (A ,B, C) were significantly (p < 0.05) different from one another.
1989
SITE
CTL
TRT
KWA
GRP#
1
DATE
^5/24/89
through
5/29/89
Water Temperature (°C)
X
26.51
26.17
28.00
SE
0.72
0.77
1.04
Range
24.5 - 29.7
24.2 - 29.7
25.0-32.5
Salinity (ppt)
X
32.08
28.23
29.83
SE
0.28
0.71
1.45
Range
31.1 - 33.1
26.1 -30.82
25.0 - 35.0
D02 (nig 02/L)
X
4.48
3.95
4.92
SE
0.74
0.49
0.70
Range
3.31 -8.12
2.63 - 6.02
1.40 - 9.30
pll
X
7.29
7.38
7.83i
SE
0.09
0.05
0.10
Range
7.11 - 7.60
7.24-7.59
7.70 - 8.30
CTL
TRT
KWA
2
5/29/89
through
6/2/89
25.90
26.30
30.30
0.67
0.57
2.08
24.3 - 27.9
25.1 -28.2
25.0 - 36.0
32.82
31.24
34.00
0.18
0.43
0.63
32.2 - 33.2
30.3 - 32.8
32.0 - 35.0
3.52
3.68
6.93
0.30
0.25
1.19
3.10 -4.41
3.18 -4.60
4.10 - 11.2
7.17
7.24
7.82
0.02
0.04
0.10
7.11 -7.21
7.10-7.29
7.50- 8.10
CTL
TRT
KWA
3
6/2/89
through
6/7/89
28.52
28.93
31.75
0.74
1.77
2.79
25.8 - 30.7
23.5-33.4
22.0 - 37.0
30.38
24.83
23.33
1.27
4.46
6.15
25.8 - 32.7
9.0 - 32.9
2.0- 35.0
4.62
5.50
9.53
0.69
0.62
1.50
2.70 - 5.20
3.70 - 7.20
4.20 - 13.5
7.20
7.22
8.00
0.05
0.08
0.11
7.00 - 7.30
7.00 - 7.50
7,50 - 8.30
CTL
TRT
KWA
4
6/7/89
through
6/11/89
25.94
24.62
24.30
0.22
0.39
0.58
25.5 - 26.5
23.5 - 25.6
24.0 - 26.0
25.60
10.88
2.40
0.46
1.38
0.40
24.1 -27.0
6.0- 13.7
2.0 - 4.0
2.94
4.06
4.96
0.39
0.56
1.14
1.80-4.20
2.60 - 5.50
3.50-9.50
7.02
7.17
7.92
0.04
0.06
0.10
6.90-7.10
7.04 - 7.40
7.60- 8.10
CTL
TRT
KWA
5
6/11/89
through
6/15/89
28.02
27.64
26.24
0.64
1.05
0.69
26.4 - 30.3
25.3-31.4
24.0 - 28.2
27.34
19.22
5.60
0.45
2.28
1.17
25.7 - 28.0
11.5 -23.7
4.0 - 10.0
3.00
3.73
6.40
053
1 29
084
1.80-4.70
0.89 - 9.90
5.00 - 9.50
7.16
7.36
774
0.09
0.11
0.13
700- 7.50
7.10- 7.70
7.30 - 8.10
.
-------
1989
SITE
CTL
TRT
KWA
GRP/?
6
DATE
6/15/89
through
6/19/89
Water Temperature (°C)
X
29.12
29.66
27.50
SE
0.67
0.94
0.82
Range
27.7- 31.0
26.5- 31.6
25.1 - 30.6
Salinity (ppt)
X
26.40
21.40
6.17
SE
0.62
3.93
1.30
Range
24.5 - 28.0
7.0-27.5
3.0- 10.0
DO2 (mg 0,/L)
X
4.28
5.94
5.17
SE
0.51
1.33
0.65
Range
2.70 - 5.50
2.10 -9.90
2.70 - 7.40
pll
X
7.70
768
7.53
SE
0.12
0.13
0.16
Range
7.40-8.10
7.40- 8.10
7.20 - 8.30
CTL
TRT
KWA
7
6/19/89
through
6/23/89
27.76
27.36
27.52
0.26
1.02
0.55
27.7 - 28.5
25.6-31.3
25.7 - 28.6
25.20
18.20
15.30
2.33
2.11
3.27
16.0-28.0
14.0 - 26.0
6.0 - 24.0
3.25
3.55
4.69
0.23
0.66
0.65
2.70 - 3.83
2.10-5.20
2.40 - 5.90
7.36
i
748
758
0.04
0.13
0.10
7.30-7.50
7.30 - 8.00
7.30 - 7.90
CTL
TRT
KWA
8
6/23/89
through
6/27/89
28.70
27.16
28.08
0.77
0.36
0.49
27.1 -30.9
25.8 - 27.9
26.4 - 29.5
27.20
19.90
5.00
1.83
2.21
1.41
i
20.0 - 30.0
15.0-27.5
2.0 - 10.0
3.52
2.94
9.88
0.61
0.18
1.38
2.26 - 5.20
2.50-3.37
5.14 - 12.65
7.32
7.28
7.70
0.04
0.07
0.18
7.20 - 7.40
7.10 -7.50
7.30-8.30
CTL
TRT
KWA
Grp 1
through
Grp 8
5/24/89
through
6/27/89
27.64A
27.34A
28. 10*
0.30
0.45
0.62
24.3 - 31.0
23.5 - 33.4
22.0 - 37.0
28.59*
22.04"
15.79C
0.65
1.38
2.21
16.0-33.2
6.0 - 32.9
2.0 - 35.0
3.84*
4.19*
6.53B
0.22
0.31
0.50
1.80-8.12
0.89 - 9.90
1.40- 13.50
7.29A
7.36A
7.76B
0.04
0.04
0.05
6.90 - 8.10
7.00 - 8.10
7.30- 8.30
-------
Mean dissolved oxygen concentrations ranged from 1.80 - 8.12, averaging 3.84 mg/L at the
CTL Site compared to levels ranging from 0.89 - 9.90, averaging 4.19 mg/L at the TRT Site.
Statistical analysis indicated the mean dissolved oxygen concentrations were not significantly
different in comparisons between the CTL and TRT Sites. Additionally dissolved oxygen levels
were at concentrations sufficient to support crustacean and fish populations observed at these
sites.
At the KWA Site, dissolved oxygen concentrations ranged from 1.40 - 13.50, averaging
6.53 mg/L. Statistical analysis indicated that mean dissolved oxygen concentrations were
significantly (p < 0.05) higher at the KWA Site when compared to the CTL and TRT Sites.
Although dissolved oxygen levels were somewhat similar at the sites, it is interesting to note that
dissolved oxygen concentrations had a much greater range, than was observed at the CTL Site
(1.80 - 8.12 mg/L). In some instances, (i.e. afternoon ebb tides) oxygen levels were
supersaturated (> 100%) suggesting possible nutrient enrichment at the KWA Site. At the TRT
Site, similar supersaturated conditions were observed, similarly suggesting nutrient enrichment.
Additional morphometric features, such as the broad, shallow habitats conducive for benthic
diatoms and other phytoplankton growth, must also be considered.
Mean daily pH ranged from 6.90 - 8.10, averaging 7.29 at CTL Site compared to values
ranging from 7.00 - 8.10, averaging 7.36 at the TRT Site. Statistical analysis indicated that pH
was not significantly different in comparisons between the two sites.
At the KWA Site, pH ranged from 7.20 - 8.30, averaging 7.76. Statistical analysis
indicated that pH at the KWA Site was significantly (p < 0.05) higher than at the TRT and CTL
Sites. Generally pH declines with reduction in salinity due to decreased carbonate/bicarbonate
buffering capacities in the water. This trend was not observed at the KWA Site, as pH increased
with generally decreased salinities. This suggests possible agricultural influences (i.e, nutrients,
fertilizer) which may have affected pH values there. Although, there were statistically
significant differences in pH between sites, these differences were not biologically significant
(i.e., pH values were within the zone of compatibility for most organisms) in terms of survival
of estuarine organisms residing there. The CO2 - Carbonate buffering system in seawater
maintains pH at a range of 7.50 - 8.50 (Valiela, 1984). EPA (19.76) reported in the Red Book
that a pH range of 6.5 - 9.0 provides adequate protection for fresh water and marine organisms.
These slight differences however, may reflect possible eutrophic conditions due to nutrient
enrichment (i.e, increase of plankton production) as evidenced by the increased dissolved oxygen
levels, observed at the KWA Site. This may result in increased plankton densities and possible
plankton species differences which may cause increased pH.
46
-------
2. 1990. Daily Water Quality Parameters
Results of physicochemical water quality parameters measured daily at each site during
the 1990 field studies are listed in Table 2. Mean daily seawater temperatures ranged from
21.3 - 32.1°C at (Tie CTL Site, averaging 26.83°C for the May - June, 1989 study period.
At the TRT Site, seawater temperatures ranged from 21.6 - 34.4°C, averaging 27.08°C
compared to temperatures ranging from 22.5 to 34.8°C, averaging 28.59°C at the KWA
Site. Statistical analysis indicated that seawater temperatures were significantly (p < 0.05)
higher at the KWA Site than at the CTL and TRT Sites; however, these differences were
more the result of the generally later daily sampling time at the KWA Site rather than actual
physical between site differences.
Mean daily salinities ranged from 24.0 - 34.0, averaging 30.4 ppt at CTL Site compared
to salinities ranging from 14.2 - 35.0, averaging 27.78 at the TRT Site. Similarly salinities
were slightly lower at the KWA Site compared to the CTL Site, ranging from 21.6 - 35.5,
averaging 30.84 ppt. Statistical analysis indicated that salinities at the TRT Site were
significantly (P < 0.05) lower than values measured at both the CTL and KWA Sites.
Mean daily dissolved oxygen levels at the CTL Site ranged from 2.10 - 7.20, averaging
4.72mg/L. Similar levels at the TRT Site ranged from 2.40 - 9.10, averaging 4.89 mg/L.
At the KWA Site, daily dissolved oxygen levels were slightly higher, ranging from 1.60 -
11.70, averaging 6.26 mg/L. Statistical analysis indicated that dissolved oxygen levels were
significantly (p < 0.05) higher at the KWA Site, when compared to the CTL and TRT
Sites. Generally dissolved oxygen levels were somewhat similar at all sites, but it is
interesting to note that at the KWA Site the range of daily dissolved oxygen levels was again
much greater than the CTL or TRT Sites during 1990, suggesting possible nutrient
enrichment. Measurements of Chlorophyll A at the KWA, TRT, and CTL Sites during 1990
made by EPA (John Macauley, USEPA, Personal Communications) directly supported this
observation, as Chlorophyll A levels were higher at the KWA Site than the TRT or CTL
Sites. Higher Chlorophyll A levels would result from high phytoplankton biomass and
resulting hyperproduction of oxygen during periods of high photic activity. While minimum
dissolved levels were low relative to EPA water quality criteria, the organisms residing in
these tidal creeks and used in toxicity tests are. well adapted to the rigors of this
environment. None of the low dissolved oxygen levels observed were detrimental to the fish
and crustaceans tested. Recent findings (Dr. L. Burdette, University of Charleston, personal
communication) report that at dissolved oxygen levels of <2mg/L, grass shrimp become
respiro-conformers rather than respiro-regulators. This suggests that these species are quite
capable of adapting to low dissolved oxygen levels.
47
-------
TAI5LE 2. Summary of physicochemical water quality parameters measured at field sites during the 1990 field study
Pooled means with different letters (A ,B, C) were significantly (p < 0.05) different from one another.
1990
SITE
CTL
TRT
KWA
GRPtt
1
DATE
5/24/90
through
5/28/90
Water Temperature (°C)
X
25.98
25.58
29.68
SE
1.32
1.32
0.58
Range
22.3 - 29.9
22.5-30.1
27.6 - 30.8
Salinity (ppt)
X
29.38
30.78
26.74
SE
0.87
0.49
0.93
Range
26.0-31.0
30.0 - 32.3
25.0 - 30.0
DOZ (mg 02/L)
X
5.98
4.39
7.68
SE
0.65
0.52
0.83
Range
4.30- 7.20
3.08- 5.69
4.60 - 9.40
pll
X
7.50
7.31 '
7.56
SE
0.11
0.07
0.04
Range
7.20-7.80
7.10- 7.49
7.40 - 7.60
CTL
TRT
KWA
2
5/28/90
through
6/1/90
24.26
23.98
25.32
1.54
1.62
1.32
21.3-29.9
21.6- 30.1
22.5 - 30.2
27.00
20.84
25.50
1.00
2.90
1.64
24.0 - 30.0
14.2 - 30.0
21.6-30.0
4.63
4.21
4.56
0.62
042
0.77
3.45 - 6.90
3.07 - 5.69
2.90 - 7.40
7.40
7.39
7.38
0.08
0.05
006
7.10-7.60
7.27 - 7.50
7.20-7.50
CTL
TRT
KWA
3
6/1/90
through
6/5/90
25.10
25.12
28.96
0.83
0.92
1.33
22.9 - 27.9
22.0 - 27.4
25.3 - 32.8
29.50
24.20
30.10
0.50
1.11
0.72
28.0-31.0
22.0 - 28.0
28.0-32.5
4.26
4.92
8.42
0.39
0.42
0.92
3.70 - 5.00
4.20 - 6.50
6.20- 11.70
7.30
7.36
7.52
0.07
0.06
0.06
7.10- 7.50
7.20- 7.50
7.40 - 7.70
CTL
TRT
KWA
4
6/5/90
through
6/9/90
27.72
29.40
32.46
0,88
0.60
0.92
25.3 - 29.6
28.0-31.0
30.0 - 34.8
30.90
29.40
33.10
0.40
0.60
0.60
30.0-31.5
28.0-31.0
32.0 - 35.0
4.80
5.22
8.88
0.43
0.47
0.45
3.70-6.10
4.22 - 6.70
7.60- 10.10
7.28
7.40
7.58
0.02
0.06
0.08
7.10-7.50
7.30 - 7.60
7.40 - 7.80
00
-------
CTL
TRT
KWA
5
6/9/90
through
6/13/90
29.46
29.46
30.30
0.37
1.14
1.59
28.7- 30.5
26.8-31.9
26 5 - 33.5
31.60
29.40
34.20
0.37
1.08
0.58
31.0-33.0
26.0 - 32.0
32.0 - 35.0
5.68
7 36
8.24
0.31
0.60
1.06
5.10-6.80
5.80- 9.10
5.50- 10.30
7.36
7.64
7.78
0.04
0.10
0.10
7.30-7 50
7.40 - 8 00
7.40 - 8.00
CTL
TRT
KWA
6
6/13/89
through
6/17/90
26.90
25.98
26.16
1.01
0.75
0.88
24.6 - 29.3
24.5-28.5
22.8 - 27.8
31.00
28.30
32.20
0.63
2.13
1.49
29.0 - 33.0
24.0 - 32.5
27.5 - 35.5
4.38
4.38
4.54
0.90
093
045
2.10-5.90
2.40-6.80
3 20-5.80
7.24
7.28
7.34
0.04
004
0.02
7.10-7.30
7.20 7.40
7.30-7.40
CTL
TRT
KWA
7
6/17/90
through
6/21/90
28.38
29.00
28.04
1.33
1.63
1.72
25.2-31.2
24.5 - 34 4
25.1 -34.6
32.4
28.10
32.50
040
068
0.77
31.0-330
26.0 - 30 0
30.0 - 34.0
3.80
3.98
3.08
027
0.49
0.50
3.20 - 4.70
2.70- 5.40
1.60 -4.60
7.20
7.44
7.28
0.05
0.09
0.05
7.10-7.40
7.20 - 7.70
7.10- 7.40
i
CTL
TRT
KWA
8
6/21/90
through
6/23/90
30.33
32.60
29.87
1.34
1.14
2.83
27.7-32.1
30.5 - 34.4
24.8 - 34.6
33.33
32.67
34.83
0.33
1.45
0.44
33.0 - 34.0
30.0 - 35.0
34.0 - 35.5
4.33
5.10
3.54
0.19
0.85
0.70
4.1 -4.70
3.50-6.40
2.30 -4.73
7.20
7.50
7.17
0.12
0.12
0.07
7.00 - 7.40
7.30- 7.70
7.10- 7.30
CTL
TRT
KWA
Grp I
through
Grp 8
5/24/90
through
6/23/90
26.83
A
27.08
A
28.59"
0.51
0.61
0.63
21.3-32.1
21.6-34.4
22.5 - 34.8
30.40
A
27.68"
30.84
A
0.41
0.89
0.73
24.0 - 34.0
14.2 - 35.0
21.6-35,5
4.72A
4.89*
626"
0.24
0.29
0.51
2.10-7.20
2.40-9.10
1.60- 11.70
7.31A
7.41B
7.47"
0.03
0.03
0.04
7.00- 780
7.10 - 8.00
7.10- 8.00
-------
Mean daily pH ranged from 7.00 - 7.80, averaging 7.31 at the CTL Site. Values at the
TRT Site were slightly higher ranging form 7.10 - 8.00, averaging 7.41. Similarly, values
at the KWA Sites were slightly higher, ranging from 7.10 -8.00, averaging 7.47. Statistical
analysis indicated that mean pH values were significantly (p < 0.05) higher at the TRT and
KWA Sites when compared to the CTL Site. The slight differences in pH between sites
were not biologically significant (i.e. would not affect survival of most estuarine organisms)
but may be indicative of agricultural influences. During 1989, pH was also significantly
higher at the KWA Site despite very low salinities (< 5 ppt). Generally pH declines with
reductions in salinity. This trend was not observed at the TRT and KWA Sites suggesting
possible agricultural influences.
B. Rainfall Measurements
1. 1989 Study Period
Cumulative rainfall totals throughout the 1989 study period (May 24 - June 27, 1989)
were 16.36 cm (± 0.25) at the CTL Site, 17.04 cm (± 0.36) at the TRT Site, and 25.40 cm
(± 0.30) at the KWA Site (Table 3). During the study period rainfall occurred on a total of
14 days at the CTL Site, 15 days at the TRT Site, and 10 days at the KWA Site (Table 3),
The largest daily (within 24 hours) rainfall amounts were 4.75 cm (± 0.05 at the CTL Site,
4.90 cm (± 0.05) at the TRT Site and 8.46-cm (± 0.08) at the KWA Site (Table 3).
Also during the study period, the number of significant (> 1.27cm/day) rainfall days was
4 days at the CTL Site, 5 days at the TRT Site, and 5 days at the KWA Site (Table 4). At
the CTL Site the greatest rainfall amounts were observed on June 5 (4.75 ± 0.05cm) and
June 6 (3.43 ± 0.08 cm). Similarly at the TRT Site, the greatest rainfall amounts occurred
on June 5 (4.90 ± 0.05 cm) and June 6 (3.43 ± 0.08 cm). At the KWA Site the greatest
rainfall amounts occurred at June 5 (7.54 ± 0.08 cm), June 6 (8.46 ± 0.08 cm) and June
24 (4.57 ± 0.00 cm).
2. 1990 Study Period
Cumulative rainfall totals throughout the 1990 Study Period (May 24 - June 23, 1990)
were 4.88 cm (± 0.08) at the CTL Site, 5.31 cm (± 0.15) at the TRT Site, and 4.32 cm
(± 0.03) at the KWA Site (Table 5). During the study period, rainfall occurred on a total
of 4 days at the CTL Site, 4 days at the TRT Site, and 5 days at the KWA Site (Table 5).
The largest daily (within a 24 hour period) rainfall amounts were 3.02 cm (± 0.03) at the
CTL Site, 2.90 cm (± 0.13) at the TRT Site, and 2.24 cm (± 0.03) at the KWA Site
(Table 5).
50
-------
Table 3. Summary of rainfall observed during the 1989 field study.
1989
SITE
CTL
TRT
KWA
Group #
1
Date
5/24/89
through
5/29/89
Cumulative Rainfall
(cm)
X (± SE)
0
0
0
Range1
(cm)
NC
NC
NC
It Days of
Rain
(Days/Grp)
0
0
0
Greatest Rainfall
Amount /Day
(cm/day)
X
0
0
0
SE
NC
NC
NC
CTL
TRT
KWA
2
5/29/89
through
6/2/89
0
0
0
NC
NC
NC
0
0
0
0
0
0
NC
NC
NC
CTL
TRT
KWA
3
6/2/89
through
6/7/89
8.26 (±0.08)
3.37 (±0.10)
16.20 (±0.15)
8.10-8.36
8.36-8.71
15.95-16.46
4
4
4
4.75
4.90
8.46
0.05
0.05
0.08
CTL
TRT
KWA
4
6/7/89
through
6/1 1/89
1.60 (±0.08)
1.80 (±0.05)
2.21 (±0.08)
1.52- 1.78
1.75- 1 91
2.13 -239
3
3
2
1.35
1.58
2.11
0.08
0.05
0.08
CTL
TRT
KWA
5
6/11/89
through
6/15/89
0
0
0
NC
NC
NC
0
0
0
0
0
0
NC
NC
NC
-------
1989
SITE
CTL
TRT
KWA
Grpff
6
Date
6/15/89
through
6/19/89
Cumulative Rainfall
(cm)
X (± SE)
5.13 (±0.05)
1.61 (±0.01)
2.21 (±0.05)
Range
(Inches)
5.05-5.21
4.03-4.13
2.10-2.29
# Days of
Rain
(Days/Grp)
4
4
2
Greatest Rainfall
Amount/Day
(cm/day)
X
3.02
2.06
1.47
SE
0.03
0.03
0.03
CTL
TRT
KWA
7
6/19/89
through
6/23/89
3.51 (±0.02)
2.67 (0.08)
0.03 (0.03)
3.47 - 3.53
2.57-2.82
0.00-0.13
4
5
1
3.02
1.22
0.05
0.03
0.03
0.05
CTL
TRT
KWA
8
6/23/89
through
6/27/89
1.02 (±0.00)
1.42 (±0.02) '
4.83 (±0.00)
1.02- 1.02
1.40- 1.46
4.83 - 4.83
2
3
2
0.76
1.14
4.57
0.00
0.00
0.00
CTL
TRT
KWA
Grp 1
through
,Grp 8
5/24/89
through
6/27/89
16. 36 (±0.25)
17.04 (±0.36)
25.40 (±0.30)
15.95-16.84
16.46-17.65
24.92-25.98
14
15
10
4.75
4.90
8.46
0.05
0.05
0.08
1 = Between 3 Rain Gauges
X = Mean
SE = Standard Error
NC = Not Calculated
GRP - Group
-------
Table 4. Dates of signifcant rainfall (> 1.27 cm/day) during the 1989 Field study
1989
SITE
CTL
Date
6/5/89
6/6/89
6/9/89
6/16/89
6/19/89
6/24/89
Rainfall Amount (cm/day)
Range*
4.70- 4.83
3.30- 3.56
1.27 - 1.52
0.89 - 0.89B
2.97 - 3.05
0.25 - 0.25s
X
4.75
3.43
1.35
0.89
3.02
0.25
(± SE)
(0.05)
(0.08)
(0.08)
(0.00)
(0.03)
(0.00)
TRT
6/5/89
6/6/89
6/9/89
6/16/89
6/19/89
6/24/89
4.83 - 4.95
3.30-3.53
1.52- 1.65
2.03 -2.10
1.21 - 1.27B
0.19 -0.326
4.90
3.43
1.57
2.03
1.22
0.25
(0.05)
(0.08)
(0.05)
(0.02)
(0.03)
(0.03)
KWA
6/5/89
6/6/89
6/9/89
6/16/89
6/19/89
6/24/89
7.37 - 7.62
8.38- 8.64
2.03 - 2.29
1.40- 1.52
0.00 - 0.00B
4.57 - 4.57
7.54
8.46
2.11
1.47
0.00
4.57
(0.08)
(0.08)
(0.08)
(0.03)
(0.00)
(0.00)
A = Range between three rain gauges
B = Rainfall < 1.27 cm/day but included for comparative purposes
X = Mean
SE = Standard Error
53
-------
Table 5. Summary of rainfall observed during the 1990 field study
1990
SITE
CTL
TRT
KWA
Grpff
1
Date
5/24/90
through
5/28/90
Cumulative Rainfall
(cm)
X (± SE)
3. 15 (±0.03)
3.02 (±0.13)
2.34 (±0.03)
Range1
(cm)
3.10-3.18
2.77 - 3.18
2.31 -2.41
tt Days of
Rain
(Days/Grp)
2
2
2
Greatest Kainfall
Amount
cm/day
X
3.02
2.90
3.02
SE
0.03
0.13 '
0.03
CTL
TRT
KWA
2
5/28/90
through
6/1/90
3.02 (±0.03)
2.90 (±0.13)
2.24 (±0.03)
3.00-3.05
2.67 - 3.05
2.21 - 2.31
1
1
1
3.02
2.90
2.24
0.03
0.13
0.03
CTL
TRT
KWA
3
6/1/90
through
6/5/90
0
0
0
NC
NC
NC
0
0
0
0
0
0
NC
NC
NC
CTL
TRT
KWA
4
6/5/90
through
6/9/90
0
0
0
NC
NC
NC
0
0
0
0
0
0
NC
NC
NC
CTL
TRT
KWA
5
6/9/90
through
6/13/90
0.25 (±0.00)
0.25 (±0.03)
<0.13 (±0.00)
0.25 - 0.25
0.25 - 0.25
<0.13-<0.13
1
1
1
0.25
0.25
<0.13
0.00
0.00
0.00
Ul
-------
UT
UT
1990
SITE
Grp#
Date
Cumulative Rainfall
(cm)
X (± SE)
Range1
(cm)
0 Days of
Rain
(Days/Grp)
Greatest Rainfall
Amount
cm/day
X
SE
CTL
TRT
KWA
6
6/13/90
through
6/17/90
1.45 (±0.05)
2.01 (±0.03)
1.78 (±0.00)
1.45 - 1.52
1.98-2.03
1.78- 1.78
1
1
1
1.45
2.01
1.78
0.05
0.03
0.00 '
CTL
TRT
KWA
7
6/17/90
through
6/21/90
0
0
0
NC
NC
NC
0
0
0
0
0
0
NC
NC
NC
CTL
TRT
KWA
8
6/21/90
through
6/23/90
0
0
<0.13
NC
NC
<0.13-
<0.13
0
0
1
0
0
<0.13
NC
NC
0.00
CTL
TRT
KWA
Grp 1
through
Grp 8
5/24/90
through
6/23/90
4.88 (±0.08)
5.31 (±0.15)
1.70 (±0.01)
4.78 - 5.00
5.00- 5.51
4.29 - 4.39
4
4
L_ 5
3.02
2.90
2.24
0.03
0.13
0.03
I = Range between 3 rain gauges
X = Mean
SE = Standard Error
NC = Not calculated
GRP = Group
-------
A total of two days of significant (> 1.27 cm/day) rainfall occurred at each site (Table
6). At the CTL Site, significant rainfall occurred on May 28 (3.02 + 0.13 cm) and June 15
(1.45 + 0.05 cm) (Table 6). At the TRT Site, significant rainfall days were also May 28
(2.90 ± 0.13 cm) and June 15 (2.01 ± 0.03 cm) (Table 6). On May 28 (2.24 ± 0.03 cm)
and June 15 (1.78 ± 0.00 cm) significant rainfall was observed at the KWA Site (Table 6).
C. Measured Insecticide Concentrations in Water Samples
1. Results for the 1989 Study Period
a. Water Samples - Results for analysis of selected seawater samples collected during
the 1989 field study are listed in Table 7 (CTL Site - grab samples); Table 8 (CTL Site -
Composite Samples); Table 9 (TRT Site - grab samples), Table 10 (TRT Site -
Composite Samples); Table 11 (KWA Site - grab samples); Table 12 (Tomato field
discharges, KWA Site -grab samples); Table 13 (Pesticide Transport - Haulover Creek
adjacent to the KWA Site) and Table 14 (spiked recovery efficiencies). Figures 3
(CTL), 4, (TRT) and 5 (KWA) depict measured insecticide levels in grab seawater
samples from each site throughout the 1989 study.
Analysis of spiked water samples indicated generally good recovery efficiencies
ranging from 78.0 - 112.6%, averaging'93.1 % (± 6.6%) for azinphosmethyl; from 60.0
-92.2%, averaging 75.9% (± 5.1%) for endosulfan I; from 61.7 - 99.7%, averaging
80.5% (± 5.8%) for endosulfan II; from 65.3 - 99.3%, averaging 82.0% (± 5%) for
endosulfan sulfate; from 51.8 -96.5%, averaging 76.9% (± 6.9%) for fenvalerate; and
from 69.1 - 99.6%, averaging 84.3% (± 4.6) for methyl parathion (Table 14). Pooled
spiked recovery efficiencies for all pesticides was 82.1% (± 2.5%). This compares
favorably with spiked recoveries for 1986 - 88, which ranged from 77.5-84.0% (Scott
et al, 1990).
At the CTL Site, only background levels of endosulfan (£10 ng/L) were observed
in most (73%) of the grab water samples analyzed, with concentrations ranging from 2 -
10 ng/L (Table 7 and Figure 3). In the 27% of samples where levels exceeded
background, endosulfan concentrations ranged from 11-14 ng/L and were observed
during ebb tides following significant (> 1.27 cm/day) rain events. The average
endosulfan concentrations for the CTL Site during 1989 was 8.0 ng/L (± 0.60 ng/L).
None of the measured concentrations at the CTL Site during 1989 exceeded the 96h .
LC50 values for any test species deployed in field toxicity tests.
56
-------
Table 6. Dates of significant rainfall (> 1.27 cm/day) during 1990 field study
1990_
SITE
CTL
DATE
5/28/90
6/15/90
Rainfall Amount (cm/day)
RANGEA
3.00- 3.05
1.40- 1.52
X
3.02
1.45
(± SE)
(0.03)
(0.05)
TRT
5/28/90
6/15/90
2.67 - 3.05
1.98 - 2.03
2.90
2.01
(0.13)
(0.03)
KWA
5/28/90
6/15/90
2.21 -2.31
1.78 - 1.78
2.24
1.78
(0.03)
(0.00)
A = Range between three rain gauges
X = Mean
SE = Standard Error
57
-------
Table 7. Summary of measured insecticide concentrations (ng/L) in water samples
collected at the CTL Site during the 1989 field study
1989
#
1
Code #
51
Date
6/5/89
Time
1330
Site
CTL
Salinity
(PPO
32
Measured Concentration (ng/L)
Insecticide
Azinphosmerhyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
< DL
4.0 (± 0.0)
< DL
< DL
2
55
6/6/89
0015
Initial Post
Rain
CTL
26
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
7.0 (± 3.4)
< DL
< DL
3
57
6/6/89
0202
2h
Post Rain
CTL
26
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
7.0 (± 3.8)
< DL
< DL
4
59
6/6/89
0445
4h
Post Rain
Dead Low
CTL
20
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
2.0 (±0.06)
< DL
< DL
5
73
6/6/89
1615
CTL
6
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
11.0 (±1.3)
< DL
< DL
6
75
6/6/89
through
6/16/90
1842
Post Rain
Viaood
CTL
5
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
8.0 (±5.1)
< DL
< DL
_
7
84
6/7/89
0530
CTL
26
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
11.0 (± 1.5)
< DL
< DL
58
-------
1989
#
8
Code #
106
Date
6/8/89
Time
0910
1 Day
Post Rain
Site
CTL
Salinity
(PPO
NM
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
< DL
14.0 (±3.0)
< DL
< DL
9
121
6/9/89
0915
Initial
Post Rain
Dead Low
CTL
25
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
9.0 (± 4.5)
< DL
< DL
10
140
6/11/89
0930
2 days
Post Rain
CTL
26
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
12.0 (±0.6)
< DL
< DL
11
144
6/12/89
0930
CTL
27
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
11.0 (±4.0)
< DL
< DL
12
148
6/13/89
0920
CTL
' 28
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
5.0 (±0.5)
< DL
< DL
13
155
6/15/89
1240
CTL
28
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
8.0 (±3.1)
< DL
< DL
14
168
6/16/89
0030
Initial
Post Rain
Dead Low
CTL
25
Azinphosmethyl
Endosulfan
Fenvalerate -
Methyl Parathion
< DL
10.0 (± 1.0)
< DL
< DL
15
174
6/16/89
1155
ll.Sh
Post Rain
V, Ebb
CTL
23
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
9.0 (± 2.2)
< DL ,
< DL
59
-------
1989
#
16
Code#
159
Date
6/16/89
Time
1400
13. 5h
Post Rain
V3 Flood
Site
CTL
Salinity
(PPt)
24
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (+SD)
< DL
10.0 (± 0.6)
< DL
< DL
«
17
180
6/17/89
0100
24h
Post Rain
Dead Low
CTL
25
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
4.0 (± 0.6)
< DL
< DL
18
189
6/17/89
1315
37h
Post Rain
Dead Low
CTL
28
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
4.0 (± 1.0)
< DL
< DL
19
200
6/18/89
1620
54h
Post Rain
% Ebb
CTL
NM
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
12.0 (± 1.0)
< DL
< DL
20
209
6/20/89
1000
CTL
28
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
9.0 (± 0.8)
< DL
< DL
21
216
6/22/89
1100
CTL
16
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
8.0 (± 4.0)
< DL
< DL
22
221
6/23/89
0830
CTL
28
Azinphosmethyl
Endosulfan
Fenvalerate"
Methyl Parathion
< DL
10.0 (± 6.0)
< DL
< DL
23
230
6/24/89
1532
Post Rain
CTL
30
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
8.0 (± 1.5)
< DL .
< DL
60
-------
1989
#
24
Code*
242
Date
6/25/89
Time
1532
Posi Rain
M Flood
Site
CTL
Salinity
(PPt)
29
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
< DL
4.0 (± 0.5)
< DL
< DL
25
250
6/26/89
0915
Dead Low
CTL
20
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
4.0 (± 1.0)
< DL
< DL
26
260
6/27/89
1015
CTL
29
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
8.0 (± 1.0)
< DL
< DL
NM = Not measure
< DL = Less than lower limits of detection
Limits of detection:
Azinphosmethyl < 5ng/L
Endosulfan < 3ng/L
Fenvalerate < 2ng/L
Methyl Parathion < Ing/L
61
-------
1909 INSECTICIDE CONCENTRATION-CTL
o
o
cc —
HI -1
i I
UJ
o 20
0 IS
0 10
:)05
0.06
0.04
0.02-
0.06-
0.04 -
0.02-
30 -
10 -
= RAINFALL ONSET
Figure 3. Measured insecticide concentrations (ug/L) and salinities (ppt) observed at the CTL Site
during the 1989 field study. Values reported are maximum daily concentrations observed
at the CTL Site on the dates sampled (•) at ebb tide. While concentrations are shown
continuously for dates sampled, it should be noted that actual insecticide concentrations may
fluctuate temporally at each site with tidal flushing. Measured insecticide levels are in ug/L,
while results for tables arc reported in ng/L. To convert ug/L to ng/L, multiply by 1000.
62
-------
Analysis of composite water samples collected at the CTL Site during periods of
significant (> 1.27 cm/day) rainfall (Table 8) indicated only detectable levels of
endosulfan (Endosulfan I and Endosulfan Sulfate) with concentrations ranging from 2 - 9
ng/L, averaging 5.5 (± 1.18 ng/L). The peak/composite ratio (peak concentration in
grab samples/average concentration in a composite sample) for endosulfan concentrations
measured during significant rain events at the CTL Site ranged from 1.11 - 3.50,
averaging 2.12 (± 0.44). These findings suggest that composite sampling at the CTL
Site would underestimate peak endosulfan concentrations by a factor of 2 (i.e. peak
concentrations would be double average concentrations measured in composite samples).
At the TRT Site (Tables 9-10, Figure 4), only detectable levels of endosulfan were
observed during periods of fair weather. Following major rain events detectable levels
of azinphosmethyl, endosulfan, and fenvalerate were observed. For endosulfan,
concentrations ranged from 2-10 ng/L, averaging 5.2 ng/L (± 1.35 ng/L) during
periods of fair weather (21% of all samples). During rain events, (79% of all samples),
endosulfan concentrations ranged from 2-20 ng/L, averaging 8.2 ng/L (± 0.99 ng/L).
Analysis of composite sampling during these same rain events indicated endosulfan
concentrations ranging from 3 - 9 ng/L, averaging 5.4 ng/L (± 0.95 ng/L).
Peak/composite ratios ranged from 1.25 - 2.80, averaging 2.32 (± 0.58 ng/L). For the
entire 1989 study period, endosulfan concentrations ranged from 2-20 ng/L, averaging
7.54 ng/L (± 0.86 ng/L). These levels were quite comparable to the average endosulfan
concentrations of 8.0 ng/L for 1989. These 1989 results also compare favorably with
TRT Site runoff sampling results for 1987 (endosulfan concentrations of 2 - 59 ng/L and
peak/composite ratios of 2.46) and 1988 (endosulfan concentrations of 2.2 ng/L and
peak/composite ratio of 2:46) (Scott et al, 1990).
During fair weather periods azinphosmethyl was not detected at the TRT Site.
Detectable levels of azinphosmethyl of 16 ng/L were observed in one sample, 4h post
rain at dead low tide on June 6, 1989.
Similarly, fenvalerate was not detected- at the TRT Site during periods of fair
weather. Following periods of significant rainfall (> 1.27 cm/day) detectable levels of
fenvalerate were observed, ranging from 11, averaging 5.23 (± 2.44).
63
-------
Table 8. Summary of measured insecticide concentratrations (ng/L) observed in composite
water samples collected at the CTL Site during the 1989 field study.
1989
#
1
2
3
4
5
6
Code#
69
92
81
131
175
190
Date - Time
6/5 - 2320
through
6/6- 1120
6/7-0000
through
6/7 - 1300
6/7 - 1600
through
6/7 - 2400
6/9 - 0300
through
6/9 - 1515
6/16 - 0030
through
6/16 - 1400
6/16 - 1400
through
6/17 - 0030
Site
CTL
CTL
CTL
CTL
CTL
CTL
Sample
Description
Composite
(F-E-F)
Composite
C/jF-F-E-'/iF)
Composite
(F-E-MiF)
Composite
(E-F-*E)
Composite
(E-F-E)
Composite
(E-F-E)
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan I
Endosulfan [I
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
X (± SD)
< DL
< DL
*< DL
2.0 (± 0.5)
< DL
< DL
< DL
5.0 (± 0.5)
< DL
4.0 (± 0.5)
< DL
< DL
< DL
2.0 (± 1.0)
< DL
3.0 (± 1.0)
< DL
< DL
< DL
2.0 (± 0.0)
< DL
2.0 (± 0.6)
< DL
< DL
< DL
2.0 (± 0.6)
< DL
2.0 (± 0.6)
< DL
< DL
< DL
4.0 (± 2.0)
< DL
5.0 (± 0.5)
< DL
< DL
E = Ebb Tide; F = Flood Tide; Samples composited every 20 minutes for each time period.
Limits of Detection: Azinphosmethyl
Endosulfan I
Endosulfan II
< DL = Less than lower limits of detection
< 5 ng/L Endosulfan Sulfate < 1 ng/L
< 1 ng/L Fenvalerate < 2 ng/L
< 1 ng/L Methyl Parathion < 1 ng/L
64
-------
Table 9. Summary of measured insecticide concentrations (ng/L) in water samples collected
at the TRT Site during the 1989 field study
1989
#
1
Code #
52
Date
6/5/89
Time
1400
Site
TRT
Salinity
(PPt)
33
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X(±SD)
< DL
2.0 (± 1.5)
< DL
< DL
2
54
6/5/89
2350
Initial
Post Rain
TRT
24
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
7.0 (± 2.8)
< DL
< DL
3
56
6/6/89
0137
2h
Post Rain
V, Ebb
TRT
26
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
6.0 (± 2.6)
< DL
< DL
4
58
6/6/89
0415
4h
Post Rain
Dead Low
TRT
5
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
16.0 (± 6.0)
13.0 (± 4.1)
93.0 (± 17.0)
< DL
5
60
6/6/89
0725
7.5h
Post Rain
'/4 Flood
TRT
9
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
9.0 (± 1.9)
50.0 (± 12.0)
< DL
6
66
6/6/89
1055
llh
Post Rain
Flood Tide
TRT
28
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
4.0 (± 1.0)
< DL
< DL
7
70
6/6/89
1320.
Initial
Post Rain
'A Ebb
TRT
21
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
7.0 (± 3.8)
< DL
< DL
8
72
6/6/89
1614
3h
Post Rain
% Ebb
TRT
14
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
16.0 (± 3.2)
< DL
< DL
65
-------
1989
n
9
Code H
74
Date
6/6/89
Time
1820
5h
Post Rain
Dead Low
Site
TRT
Salinity
(Ppt)
5
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
< DL
20.0 (± 1.7)
40.0 (± 47.0)
< DL
10
85
6/7/89
1100
21.5h
Post Rain
Dead Low
TRT
8
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
13.0 (± 0.5)
22.0 (± 0.0)
< DL
11
122
6/9/89
0830
Initial
Post Rain
Dead Low
TRT
6
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
5.0 (± 0.5)
21.0 (± 2.0)
< DL
12
139
6/11/89
0830
48h
Post Rain
TRT
12
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
5.0 (± 0.5)
< DL
< DL
13
143
6/12/89
0845
72h
Post Rain
TRT
18
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
6.0 (± 0.5)
< DL
< DL
14
147
6/13/89
0830 .
TRT
24
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
7.0 (± 1.0)
< DL
< DL
15
154
6/15/89
1130
TRT
19
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
10.0 (± 2.4)
< DL
< DL
to
16
167
6/16/89
0000
Initial
Post Rain
Dead Low
TRT
15
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
17
173
6/16/89
1145
I2h
Post Rain
Dead Low
TRT
7
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
5.0 (±0.5)
< DL
< DL
>
< DL
14.0 (± 1.4)
15.0 (± 6.0)
< DL
66
-------
1989
#
Code*
Date
Time
Site
Salinity
(Ppt)
Measured Concentration (ng/L)
Insecticide
X (±SD)
18
161
6/16/89
1320
13.25h
Post Rain
- '/3 Flood
TRT
7
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
6.0 (± 2.4)
< DL
< DL
19
181
6/17/89
0030
24.5h
Posi Rain
Dead Low
20
188
6/17/89
1625
40.5h
Post Rain
Dead Low
TRT
NM
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
TRT
28
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
11.0 (± 1.5)
< DL
< DL
< DL
2.0 (± 0.5)
< DL
< DL
21
201
6/18/89
1800
66h
Post Rain
TRT
28
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
11.0 (± 1.0)
10.0 (± 12.0)
< DL
22
204
6/19/89
0600
77. 5h
Post Rain
TRT
26
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
7.0 (± 3.4)
< DL
< DL
23
210
6/20/89
1430
TRT
18
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
6.0 (± 1.7)
< DL
< DL
24
219
6/23/89
0800
TRT
18
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
5.0 (± 2.6)
< DL
< DL
25
222
6/24/89
0930
Initial
Post Rain
Vb Flood
TRT
28
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
5.0 (± 1.0)
< DL
< DL
67
-------
1989
it
26
27
Code*
241
249
Date
6/25/89
6/26/89
Time
0930
Initial
Post Rain
'/i Flood
0800
22. 5h
Post Rain
Site
TRT
TRT
Salinity
(PPt)
22
15
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X(±SD)
< DL
6.0 (± 1.2)
< DL
< DL
< DL
2.0 (± 1.7)
< DL
< DL
28
258
6/27/89
0830
47h
Post Rain
TRT
17
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
1.0 (± 1.0)
< DL
< DL
< DL = Less than lower limits of detection
NM = Not Measured
Limits of Detection:
Azinphosmethyl < 5 ng/L
Endosulfan < 3 ng/L
Fenvalerate < 2 ng/L
Methyl Parathion < 1 ng/L
68
-------
1989 INSECTICIDE CONCENTRATION-TRT
0.2C
o =>
Q —
Z
005 -
_
wi -i
a
z
0.04 -
002 -
- 0.10 H
ui -/
5 Z
30 -
>-
t ~
z a 20 -
*" 10 -
1
Figure 4. Measured insecticide concentrations (ug/L) and salinities observed at the TRT Site during the 1989
field study. Values reported are maximum daily concentrations observed at the TRT Site on the dates
(•) sampled at ebb tide. While concentrations depicted are generally representative for the dates
sampled, actual pesticide concentrations may fluctuate temporally with tidal flushing. Measured
insecticide levels reported are in ug/L rather than the ng/L levels reported in tables. To convert ug/L
to ng/L, multiply by 1000.
69
-------
Table 10. Summary of measured insecticide concentrations (ng/L) observed in
composite water samples collected at the TRT Site during the 1989 field study.
1989
ft
1
2
3
4
5
6
7
Code ft
68
95
83
129
191
247
257
Date - Time
6/5 - 2~320
through
6/6- 1120
6/7-0000
through
6/7 - 1300
6/7 - 1600
through
6/8 - 0030
6/9 - 0300
through
6/9- 1515
6/16 - 1230
through
6/17-0000
6/25 - 1000
through
6/25 - 1930
6/25 - 2000
through
6/26 - 0830
Site
TRT
TRT
TRT
TRT
TRT
TRT
TRT
Sample
Description
Composite
(F-E-F)
Composite
(F-E-F)
Composite
(%E-E-F)
Composite
C/iE-E-F-'/iE)
Composite
-------
These results agree favorably with 1988 results when post rain fenvalerate levels of
< DL - 68 ng/L were observed at the TRT Site, with peak grab ratios of 2.06 (Scott et
al, 1990). For the entire study period, fenvalerate concentrations ranged from < DL -
93 ng/L, averaging 9.0 ng/L. At the TRT Site, no measured endosulfan or
azinphosmethyl concentrations exceeded the 96h LC50 values for any of the test species
deployed. Measured fenvalerate concentrations during rain events exceeded the 96H
LC50 values for mysid shrimp and P. pugio and may have exceeded the no observable
effect concentration (NOEC) for penaied shrimp. Measured fenvalerate concentrations
did not exceed 96h LC50 values for mummichogs and sheepshead minnow during rain
events monitored during 1989.
These data suggest that during 1989, 3 days of rainfall (June 6, 1989; June 9, 1989;
and June 16, 1989) occurred which resulted in significant pesticide runoff at
concentrations high enough to pose acute toxicity risk to crustaceans (M. bahia, P.
pugio, and Penaeus sp.) deployed in field toxicity tests (Figure 16). Peak/composite
sample comparisons during runoff events suggest that peak (i.e. pulsed) insecticide
concentrations were 2.32 - 5.23 times greater than time weighted average concentrations
obtained from composite water samples.
At the KWA Site (Tables 11-13; Figure 5), significant runoff of endosulfan and
azinphosmethyl following rain events on June 5, June 6, June 9, June 16, and June 24,
1989, resulted in elevated levels of both pesticides for most of the 1989 study period.
During periods of fair weather insecticide concentrations ranged from < DL - 211 ng/L,
averaging 68.7 ng/L (± 31.05 ng/L) for azinphosmethyl; from 5 - 64 ng/L, averaging
33.0 ng/L for endosulfan; and only non detectable levels of fenvalerate. At the KWA
Site, measured azinphosmethyl and fenvalerate concentrations following major rainfall
events exceeded the 96h LC50 values for Penaeus species, P. pugio, 'M. bahia, juvenile
C. variegatus and the NOEC for F. heteroclitus. Similarly, concentrations of endosulfan
exceeded the NOEC and Lowest Observable Effect Concentration (LOEC) for some
species (P. pugio, M. bahia, Penaeas species, juvenile C. variegatus and juvenile F.
heteroclitus). In fact, the average in stream concentration of azinphosmethyl (1,078
ng.L) for the entire study period (May 23 - June 24,1989) exceeded reported 96h LC50
value (970 - 1050 ng/L for P. pugio). Similarly the average in stream concentrations of
endosulfan (48.42 ng/L) and fenvalerate (4.48 ng/L) represented 21% and 64%
respectively of the 96h LC50 value for the most sensitive species (juvenile F.
heteroclitus and zoeal P. pugio, respectively).
71
-------
Table 11. Summary of measured insecticide concentrations (ng/L) in water samples
collected at the KWA Site during the 1989 field study
1989
#
1
Code#
50
Date
6/4/89
Time
1415
Site
KWA
Salinity
(PRt)
33
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
< DL
5.0(±1.0)
< DL
< DL
2
53
6/5/89
1500
Initial
Post Rain
V3 Ebb
KWA
35
Azinphosraethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
4.0 (± 1.0)
< DL
< DL
3
63
6/6/89
0345
L3h
Post Rain
% Ebb
KWA
2
Azinphosraethyl
Endosulfan
Fenvalerate
Methyl Parathion
817.0 (± 69.0)
18.0 (± 4.5)
54.0 (± 3.8)
< DL
^
4
64
6/6/89
0545
15h
Post Rain
Dead Low
KWA
4
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
480.0 (± 47.0)
16.0 (± 4.3)
< DL
< DL
5
65
6/6/89
0745
I7h
Post Rain
Vb Flood
KWA
6
Azinphosraethyl
Endosulfan
Fenvalerate
Methyl Parathion
1,730.0 (± 137.0)
69.0 (± 5.1)
< DL
< DL
6
99
6/6/89
1200
21h
Post Rain
Flood Tide
KWA
9
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
809.0 (± 51.0)
144.0 (± 7.1)
< DL
< DL
7
100
6/6/89
1400
23h
Post Rain
V, Ebb
KWA
9.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
306.0 (±28.0)
50.0 (± 4.8)
< DL .
< DL
72
-------
1989
#
8
Code #
101
Date
6/6/89
Time
1630
26. 5h
Post Rain
% Ebb
Site
KWA
Salinity
(PPt)
NM
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
508.0 (+ 44.0)
122.0 (± 8.1)
< DL
< DL
9
102
6/6/89
1830
28.5
Post Rain
Dead Low
KWA
NM
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
350.0 (+ 41.0)
46.0 (± 5.4)
39.0 (± 5.0)
< DL
10
103
6/7/89
0400
Initial
Post Rain
ViEbb
KWA
2
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
1,078.0 (± 87.0)
71.0 (± 5.6)
< DL
< DL
11
104
6/7/89
0600
2h
Post Rain
Dead Low
KWA
0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
151. 0(± 26.0)
96.0 (± 14.0)
64.0 (± 7.0)
< DL
12
105
6/7/89
0800
4h
Post Rain
'A Rood
KWA
3
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
1, 222.0 (± 83.0)
163.0 (± 11.9)
< DL
< DL
13
88
6/7/89
1100
7h
Post Rain
Vt Flood
KWA
2
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
28.0 (± 13.0)
50.0(± 1.0)
< DL
< DL
14
115
6/7/89
1200
8h
Post Rain
Rood Tide
KWA
2
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
1,155.0 (± 192.0)
122.0 (± 12.6)
< DL
< DL
73
-------
1989
#
15
Code #
109
Date
6/8/89
Time
1030
30. 5h
Post Rain
'A Flood
Site
KWA
Salinity
(PPt)
2
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
382. 0(± 55.0)
64.0 (± 5.8)
< DL
< DL
16
124
6/9/89
0800
Initial
Post Rain
Vs Flood
KWA
2
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
76.0(± 4.0)
41.0 (+ 4.5)
< DL
< DL
17
135
6/9/89
1245 .
4h
Post Rain
High Tide
KWA
NM
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
50.0 (± 10.0)
43.0 (± 1.7)
< DL
< DL
18
134
6/10/89
19
141
6/11/89
0800
24h
Post Rain
Dead Low
KWA
2
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
31.0 (± 4.0)
43.0 (± 5.4)
< DL
< DL
-
1130
KWA
4
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
368.0 (± 90.0)
54.0 (± 6.2)
< DL
< DL
20
145
6/12/89
0800
KWA
4
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
211.0(± 74.0)
50.0 (± 8.7)
31.0 (± 36.0)
< DL
21
149
6/13/89
1030
KWA
6
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
53.0 (± 62.0)
64.0 (± 8.5)
< DL
< DL
74
-------
1989
#
22
Code #
152
Date
6/14/89
Time
0830
Site
KWA
Salinity
(PPt)
4
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
89.0 (± 13.0)
32.0 (± 4.1)
< DL
< DL
23
153
6/15/89
0900
KWA
10
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
40.0 (± 6.0)
32.0 (± 3.3)
< DL
< DL
24
169
6/16/89
0130
Initial
Post Rain
Dead Low
KWA
6
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
168.0 (± 19.0)
28.0 (± 4.0)
< DL
< DL
25
160
6/16/89
1020
9h Post Rain
VS Ebb
Fish Kill
Observed
26
27
172
179
6/16/89
1250
ll.Sh
Post Rain
X Ebb
6/16/89
1515
I4h
Post Rain
>A Flood
KWA
10
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
2457.0 (± 256.0)
25.0 (± 1.2)
< DL
< DL
KWA
KWA
3
4
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
1100.0 (± 158.0)
30.0 (± 1.3)
< DL
< DL
2222.0 (± 520.0)
38.0 (± 5.1)
< DL
< DL
28
178
6/16/89
1800
17h
Post Rain
Flood Tide
KWA
6
Azinphosmethyl
Endosulfan "
Fenvalerate
Methyl Parathion
1528.0 (± 366.0)
24.0 (± 2.1)
< DL
< DL
75
-------
1989
#
29
Code #
182
Date
6/17/89
Time
0100
23h
Post Rain
Dead Low
Site
KWA
Salinity
(PPt)
7
Measured Concentration (ng/L)
Insecticide
Azinphosmechyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
578.0 (+ 122.0)
27.0 (± 2.4)
< DL
< DL
30
186
6/17/89
1100
33.5h
Post Rain
VbEbb
KWA
3
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
716.0 (± 49.0)
27.0 (± 2.4)
< DL
< DL
31
192
6/18/89
0645
53.25h
Post Rain
'A Flood
KWA
5
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
426. 0(± 53.0)
23.0 (± 1.5)
< DL
< DL
32
202
6/18/89
1830
55h
Post Rain
Dead Low
KWA
NM
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
1710.0 (± 766.0)
32.0 (± 3.3)
< DL
< DL
-
33
205
6/19/89
1100
76.5h
Post Rain
'A Flood
KWA
6
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
1351.0 (± 267.0)
27.0 (± 30.)
< DL
< DL
34
35
211
220
6/20/89
6/23/89
0930
99h
Post Rain
Dead Low
1000
KWA
KWA
17
10
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
214.0 (± 26.0)
12.0 (±2.2)
< DL
< DL
19.0 (± 4.0)
15.0 (± 1.0)
< DL
< DL
76
-------
1989
#
36
Code #
224
Date
6/24/89
Time
1040
Initial
Post Rain
Vb Flood
Fish Kill
Observed
Site
KWA
Salinity
(PPt)
4
Fish Kill
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
3111.0 (± 377.0)
64.0 (± 6.6)
< DL
< DL
37
237
6/24/89
1930
9h
Post Rain
Dead Low
KWA
2
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
1612. 0(± 150.0)
34.0(± 1.9)
< DL
< DL
38
239
6/24/89
2200
11. 5h
Post Rain
'A Flood
KWA
NM
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
2383. 0(± 352.0)
42.0 (± 2.4)
< DL
< DL
39
243
6/25/89
1100
Initial
Post Rain
% Flood
KWA
6
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
5288.0 (± 742.0)
65.0 (± 9.6)
< DL
< DL
40
248
6/25/89
2011
9h
Post Rain
'A Ebb
KWA
3
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
1956.0 (± 439.0)
50.0 (± 2.2)
< DL
< DL
41
251
6/29/89
2011
24H
Post Rain
Dead Low
KWA
2
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
1604.0 (± 353.0)
38.0 (± 2.7)
< DL
< DL
42
261
6/27/89
1115.
48h
Pose Rain
KWA
2
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
7002.0 (± 759.0)
32.0 (± 2.4)
< DL
< DL
NM = Not Measured
-------
1989 INSECTICIDE CONCENTRATION-KWA
CO
O
Q
0.2Q
0.15
0 05-
UJ
-------
At the KWA Siteduring rain events, azinphosmethyl concentrations ranged from < DL -
7,002 ng/L, averaging 1,246.31 ng/L (± 243.79 ng/L). Endosulfan concentrations ranged
from 4 - 163 ng/L, averaging 50.92 ng/L (± 5.95 ng/L). Fenvalerate concentrations ranged
from < DL - 64 ng/L, averaging 5.2 ng/L (± 2.60 ng/L).
At the KWA Site during the entire 1989 study period, azinphosmethyl concentrations
ranged from < DL - 7002 ng/L, averaging 1078.07 ng/L (+ 218.28 ng/L). Endosulfan
concentrations ranged from 4-163 ng/L, averaging 48.42 ng/L (± 5.33 ng/L). Fenvalerate
concentrations ranged from < DL - 64 ng/L, averaging 4.48 ng/L (± 2.24 ng/L).
b. Pesticide Loadings
Table 12 lists results of water samples collected from drainage ditches from tomato fields
at the KWA Site approximately 20m upstream of the KWA Site. These samples were taken
to address pesticide loading, potential at the KWA Site. During the rain event of June 16,
1989, 1.47 cm of rain at the KWA Site caused significant azinphosmethyl runoff with
concentrations of 15, 497 ng/L (± 1796 ng/L) and slight runoff of endosulfan (concentrations
of 119 ng/L). In stream water samples 20m downstream from this tomato field site contain
much lower levels of azinphosmethyl (2457 ng/L) and endosulfan. The field/stream ratio
(insecticide concentration measured exiting the tomato field/tidal stream concentration) was
6.31 for azinphosmethyl and 4.76 for endosulfan. A fish kill was observed at the KWA Site
during this runoff event.
On June 18, 1989, 44h post rain significant azinphosmethyl (7,567 ng/L) and endosulfan
(117 ng/L) concentrations were still observed in ditches exiting the tomato field. In stream
concentrations of azinphosmethyl (426 ng/L) and endosulfan (23 ng/L) at the KWA Site 20m
downstream were much lower than levels observed in the tomato field ditch. The field stream
ratio was 17.76 for azinphosmethyl and 5.09 for endosulfan.
On June 24, 1989, 4.57 cm of rainfall occurred at the KWA Site resulting in significant
runoff of azinphosmethyl (1574 ng/L) and endosulfan (100 ng/L) from tomato fields adjacent
to the site which caused a fish kill (Plates 4-7). At the KWA Site, in stream concentrations
of azinphosmethyl (3, 111 ng/L) were nearly double concentrations measured in the ditches
exiting the tomato field. This suggests that the majority of the azinphosmethyl runoff had
exited the field and entered the tidal creek. For endosulfan,
79
-------
Table 12. Summary of measured insecticide concentratrations (ng/L) observed in composite water
samples collected from a tomato field drainage ditch as it enters an estuarine tidal creek at
the KWA Site during the 1989 field study.
1989 _
#
1
2
3
Code ft
171
193
223
Date - Time
6/16/89 - 1100
6/18/89 - 0645
6/24/89- 1100
Site
KWA
KWA
KWA
Sample
Description
Tomato field
drainage ditch
as it enters
tidal creek
Tomato field
drainage ditch
as it enters
tidal creek
Tomato field
drainage ditch
as it enters
tidal creek
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
X (± SD)
15497.0 (±
1796.0)
3.0 (± 0.5)
24.0 (± 3.0)
92.0 (± 12.0)
< DL
< DL
7567.0 (± 802.0)
< DL
24.0 (± 3.0)
93.0 (± 39.0)
< DL
< DL
1574.0 (± 247.0)
< DL
19.0 (± 1.0)
81.0 (± 6.0)
< DL
< DL
Limits of Detection:
Azinphosmethyl
Endosulfan I
Endosulfan II
< 5 ng/L
< 1 ng/L
< 1 ng/L
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
< 1 ng/L
< 2 ng/L
< 1 ng/L
< DL = Less than lower limits of detection
80
-------
Plate 4 Photograph of dead F. heteroclitus at the KWA Site following significant rainfall and
resulting fish kill. Note how all dead mummichogs were juvenile - young adult size
classes.
81
-------
Plate 5 Photograph of dead P. pugio at the KWA Site following significant rainfall and
resulting fish kill. Dead grass shrimp were found along the bank of the tidal creek
throughout this area.
82
-------
Plate 6A. Photograph of dead Uca pugilator at the KWA Site following significant rainfall and
resulting fish kill. There was significant mortality in fiddler crabs at this site.
Plate 6B.
Photograph of dead polycbaetes at the KWA Site following significant rainfall and
resulting fish kill. This was the first time that dead polycbaetes were observed at a fish
kill during the 6 years of these studies.
83
-------
Plate 7A. Photograph of dead MugU cephalus at the KWA Site following significant rainfall and
resulting fish kill.
Plate 7B. Photograph of shorebirds (gulls, wading shorebirds and egrets) consuming dead fish,
crustaceans and invertebrates at the KWA Site during the fish kill.
84
-------
a less water soluble, organochlorine insecticide, the reverse was seen. Endosulfan
concentrations in the tomato field runoff (100 ng/L) were higher than in stream endosulfan
concentrations (64 ng/L) suggesting that less water soluble insecticides may runoff from the
field at slightly different rates than more water soluble pesticides such as azinphosmethyl.
Azinphosmethyl is nearly 100 times more soluble in water than endosulfan (29 ng/L at 25 °C
versus 0.32 - 0.33 ng/L at 25°C in distilled water), which explains its greater mobility and
transport. The field/stream ratio for this rain event was 0.51 for azinphosmethyl, resulting
from the rapid runoff from the field, and 1.56 for endosulfan.
c. Pesticide Transport Studies
To further evaluate pesticide transport, samples were analyzed from Haulover Creek,
a small tidal creek 2 river miles (4.5 km) northwest of the KWA Site during periods of
significant rainfall (Table 13). On June 24, 1989, 4.57 cm of rain fell at the KWA Site
resulting in significant runoff of azinphosmethyl (3111 ng/L) and endosulfan (64 ng/L) at
the KWA Site. A resulting fish kill occurred at the KWA Site at 1040 on June 24, 1989.
At Haulover Creek at 1440 on June 24, 1989 only slight concentrations of azinphosmethyl
(116 ng/L) and endosulfan (4 ng/L) were observed. Concentrations of endosulfan and
azinphosmethyl were 16 and 26.8 time respectively higher at the KWA Site. By 1726 on
June 24, 1989, azinphosmethyl concentration at the Haulover Creek Site had increased to
1631 ng/L and were identical to concentrations at the KWA Site (1612 ng/L). Similarly
endosulfan concentrations increased to 18 ng/L at the Haulover Site. A fish kill then began
to occur at the Haulover Site. By 1945, azinphosmethyl concentrations (1700 ng/L) were still
similar to levels measured at the KWA Site. Similarly endosulfan concentrations at the
Haulover Site (23 ng/L) were nearly identical to concentrations at the KWA Site (34 ng/L).
These results clearly indicate the rapid transport and mobility of azinphosmethyl and
endosulfan in agricultural runoff. In a 12 - 14h time period, significant levels of these
insecticides were discharged from a tomato field into a small tidal creek causing a fish kill.
On the initial ebb tide these insecticides wee discharged out of this small tidal tributary into
a larger portion of Haulover Creek. As the flood tide occurred, this runoff was transported
> 2.0 km upstream in Haulover Creek, causing a second fish kill. These findings clearly
indicate that in small drainage basins with very slow flushing rates, runoff of insecticides
may cause significant spatial impacts which are not restricted to the stream site adjacent to
the point of discharge into the stream.
85
-------
Table 13. Summary of measured insecticide concentrations (ng/L) at fish kills at
Haulover Creek, adjacent to the KWA Site and at a tidal canal adjacent
to the golf links at Seabrook Island.
1989
0
1
2
3
4
5
Code#
142
233
235
238
240
Date -
Time
6/11/89
1330
6/24/89
1440
6/24/89
1726
6/24/89
1945
6/24/89
2210
Site
Seabrook
Island
Golf
Course
Haulover
Creek
Haulover
Creek
Haulover
Creek
Haulover
Creek
Sample
Description
Fish Kill
at tidal
canal
adjacent to
golf course
Fish Kill
in creek
adjacent
to KWA Site
(Flood Tide)
Fish Kill
in creek
adjacent
to KWA
Site
('/* Ebb)
Fish Kill
in cteek
adjacent
to KWA Site
(Dead Low)
Fish Kill
in creek
adjacent
to KWA
Site
(14 Flood)
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
X (± SD)
< DL
< DL
< DL
< DL
< DL
< DL
116.0 (± 27.0)
< DL
< DL
4.0 (+ 1.0)
< DL
< DL
1631.0 (± 290.0)
2.0 (± 1.0)
4.0(± 1.0)
10.0(1 1.0)
< DL
< DL
1700.0 (+ 792.0)
2.0 (1 0.5)
4.0(1 0.6)
17.0 (+ 2.0)
< DL
< DL
2670.0(1 262.0)
< DL
4.0(1 0.6)
12.0(1 1.0)
< DL
< DL
Limits of Detection:
Azinphosmethyl
Endosulfan I
Endosulfan II
< 5 ng/L
< 1 ng/L
< 1 ng/L
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
< 1 ng/L
< 2 ng/L
< 1 ng/L
-------
Table 14. Spiked recovery efficiencies (% recovery) for water samples
during the 1989 field study
Pesticide
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Spike
Concentration
L_
-------
2. Results for the 1990 Field Study
Results of analysis of selected seawater samples collected daring the 1990 field study are
listed in Tables 15 (CTL Site - Grab and Composite samples), Table 16 (TRT Site - grab and
composite samples), Table 17 (KWA Site - grab and composite samples) and Table 18
(Spiked recovery efficiencies). Figure 6 (CTL Site), Figure 7 (TRT Site) and Figure 8
(KWA Site) depict measured insecticide levels in grab samples from each site during the 1990
study. Analysis of spiked water samples indicated generally good recovery efficiencies
ranging from 55.0 -80.0%, averaging 69.2% (± 4.5%) for azinphosmethyl, from 62.0 -
81.0%, averaging 70.8% (± 3.6%) for endosulfan I. From 68.0 - 85.0%, averaging 78.2%
(± 3.3%) for endosulfan II, from 79.0- 102.0%, averaging 90.0% (± 5.3%) for endosulfan
sulfate, from 66.0 - 94.0%, averaging 83.4% (± 4.7%) for fenvalerate, and 71.0 - 92.0%,
averaging 82.2% (± 3.8%) for methyl parathion (Table 18). Pooled spiked recovery
efficiencies for all pesticides was 79.0% (± 3.2%). This compares favorably with spiked
recovery efficiencies for 1989 (82.1 %) and for results from 1986-88, which ranged from 77.5
- 84.0% (Scott et al, 1990).
At the CTL Site, only background levels of endosulfan (<10 ng/L) were observed in
water samples analyzed during the 1990 field study (Table 15 and Figure 6). Endosulfan was
the only pesticide detected, with concentrations ranging from < DL - 9 ng/L, averaging 2.3
ng/L (± 1.20 ng). Detectable endosulfan concentrations were observed in only 33% of the
samples analyzed. The water samples analyzed were those associated with the two major rain
events (May 28, 1990 and June 15, 1990) during the 1990 studies. These findings clearly
indicate that at the CTL Site, pesticides were below concentrations which cause toxicity in
acute exposures to those species deployed in field toxicity tests.
At the TRT Site (Table 16 and Figure 7), detectable concentrations of endosulfan and
fenvalerate were observed. Endosulfan concentrations ranged from
-------
Table 15. Summary of measured insecticide concentrations (ng/L) observed in water
samples from the CTL Site during 1990.
Values are for grab samples unless otherwise denoted (composite samples).
1990
ff
1
Code #
302
Date
5/24/90
Time
1445
Site
CTL
Salinity
(PPt)
29.7
Measured Concentration (ng/L)
Insecticide
Azinphosmeihyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
< DL
< DL
< DL
< DL
2
344
5/28/90
through
5/29/90
2200-1030
Initial
12. 5h
Posi Rain
CTL
Composite
F-E-F
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
3
342
5/30/90
0900
CTL
24.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
-
4
382
6/7/90
1333
CTL
31.5
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
6.0 (±0.60)
< DL
< DL
5
413
6/15/90
0808
CTL
31.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
6.0 (±0.05)
< DL
< DL
6
428
6/16/90
through
6/16/90
0015-1300
(7-20h)
Post Rain
CTL
Composite
%F-*P-*E-»F
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Paradiion
< DL
< DL
< DL
< DL
7
424
6/16/90
1100
18h
Post Rain
V3 Flood
CTL
31.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
9.0 (±1.5)
< DL
< DL
89
-------
L990
#
Code #
Date
Time
Site
Salinity
(PPO
Measured Concentration (ng/L)
Insecticide
X (±SD)
8
431
6/17/90-
1030
41. 5h
Post Rain
Dead Low
CTL
31.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
9
456
6/22/90
1400
CTL
34.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
E = Ebb
F = Flood Tide
Lower Limits of Detection:
Azinphosmethyl < 5ng/L
Endosulfan < 3 ng/L
Fenvalerate < 2ng/L
Methyl Parathion < Ing/L
-------
1000 INSECTICIDE CONCCNTF1A MONS-CTL
i. 1)02 -
Figure G. Measured insecticide concentrations (ug/L) and salinities (ppt) observed at the CTL Site during the
1990 field study. Values depicted are maximum daily concentrations at the CTL Site on the dates
(•) sampled al ebb tide. While concentrations depicted arc generally representative for the dates
sampled, actual pesticide levels may fluctuate temporally with tidal flushing. Measured insecticide
concentrations depicted arc in ug/L radicr than ng/L levels'reported in tables. To convert ug/L
10 ng/L, multiply by 1000.
91
-------
Table 16. Summary of measured insecticide concentrations (ng/L) observed
in water samples form the TRT Site during the 1990 field study.
Values are for grab samples unless otherwise denoted (i.e. composite).
1990
#
1
Code#
332
Date
5/29/90
Time
0620
8h
Post Rain
Dead Low
Site
TRT
Salinity
(PPt)
6.0
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
< DL
14.0 (± 1.7)
123.0 (±5.1)
< DL
2
343
5/28/90
through
5/29/90
2200-1100
Initial 13h
Post Rain
TRT
Composite
F-E-F
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
3
419
6/16/90
0044
8h
Post Rain
% Flood
TRT
30.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
4.0 (± 0.8)
< DL
< DL
4
425
6/16/90
0945
17h
Post Rain
Dead Low
TRT
" 31.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
5
429
6/16/90
through
6/16/90
0045-1400
8-21h
Post Rain
TRT
Composite
%F-F-E-F
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
5.0 (±3.7)
< DL
< DL
6
431
6/17/90
0900
40h
Post Rain
'A Rood
TRT
26.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
E = Ebb Tide; F = Flood Tide;
Lower Limits of Detection: Azinphosmethyl < 5 ng/L
Endosulfan < 3 ng/L
Fenvalerate < 2 ng/L
Methyl Parathion < 1 ng/L
< DL = Less than lower limits of detection
92
-------
1990 INSECTICIDE CONCENTRAT1ONS-TRT
_ o
Figure 7. Measured insecticide concentrations (ug/L) and salinities (ppt) observed at the TRT Site during Ujc
1990 field study. Values depicted arc representative for the dates sampled (•). Actual pesticide
levels may fluctuate temporally with tidal flushing. Measured insecticide concentrations depicted
arc in ug/L rather than ng/L levels reported in tables. To convert ug/L (o ng/L, multiply by 1000.
93
-------
noted peak/composite ratios ranging as high as 10.91 (Scott et al, 1990) - > 11.01 (this study). Using
those extrapolations, one would have estimated fenvalerate concentrations of approximately 11 ng/L,
just above DL. Non detectable levels of fenvalerate were noted in composite samples suggesting that
only a small "slug" of fenvalerate was discharged into the environment during this rain event. During
the rain event of June 15, 1990, only background levels of endosulfan were observed.
During 1990, at The TRT Site, no measured endosulfan concentrations exceeded the 96h LC50
values for any of the species deployed. Measured fenvalerate concentrations during the rain event of
May 28, 1990, exceeded the 96h LC50 values for several crustacean species (M. bahia and P. pugio)
and the LOEC for Penaeus species. Measured fenvalerate concentrations did not exceed 96h LC50
values for mummichogs, sheepshead minnow and silversides.
At the KWA Site (Table 17 and Figure 8), onJy detectable levels of azinphosmethyl were observed.
Azinphosmethyl concentrations ranged from < DL - 62 ng/L, averaging 13.4 ng/L (+ 6.51 ng/L).
Detectable levels of azinphosmethyl were noted in 40% of the samples, mainly in those samples
associated with June 15, 1990 rain event. During this rain event, azinphosmethyl concentrations in
grab samples ranged from < DL - 62 ng/L, averaging 22.20 ng/L (± 11.39 ng/L). Analysis of
composite samples indicated an azinphosmethyl concentrations of 24 ng/L (± 3.7 ng/L). The
peak/composite ratio was 1.13. These results for the KWA Site, suggest that during 1990 only
detectable levels of azinphosmethyl were observed which were below levels considered acutely toxic
to any of the species deployed in field toxiciry studies.
The 1990 study period (May - June) was an extremely dry period compared with results for 1989.
Generally, dry weather results in decreased numbers of crop pests which reduces the amounts and types
of insecticides applied. The 1990 study provided a period of stark contrast to the 1989 study
characterized by:
1) Relatively low rainfall
2) Relatively little, if any significant insecticide runoff; and
3) Generally very high survival in species deployed in acute
toxicity tests at each site.
94
-------
Table 17. Summary of measured insecticide concentrations (ng/L) observed in water samples
from the KWA Site during the 1990 field study. Values are for grab samples unless
otherwise denoted (composite samples).
1990
#
1
Code#
321
Date
5/28/90
Time
1830
Dead Low
Site
KWA
Salinity
(PPt)
27.6
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
.X (±SD)
< DL
< DL
< DL
< DL
2
324
5/28/89
2000
Initial
Post Rain
V, Flood
KWA
12.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
3
331
5/29/90
0600
10h
Post Rain
Dead Low
KWA
30.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
4
339
5/28/90
through
5/29/90
2000-0820
Initial
12h
Post Rain
KWA
Composite
VsF-F-E-VSF
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
5
417
6/15/90
2010
Initial
Post Rain
Dead Low
KWA
34.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
62.0 (±9.2)
< DL
< DL
< DL
6
420
6/16/90
0135
5.5h
Post Rain
% Flood
KWA
NM
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
95
-------
1990
ff
1
Code#
423
Date
6/16/90
Time Site
_ 0615
10h
Post Rain
%Ebb
KWA
Salinity
(PPt)
31.0
Measured Concentration (ng/L)
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
X (±SD)
22.0 (±1.3)
< DL
< DL
< DL
8
426
6/16/90
0845
13h
Post Rain
Dead Low
KWA
28.0
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
< DL
< DL
< DL
< DL
9
427
6/16/90
1215
16h
Post Rain
% Flood
KWA
27.5
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
27.0 (±2.0)
< DL
< DL
< DL
10
430
6/16/90
through
6/16/90
0135 rt,rough 1235
5.5h- 16h
Post Rain
KWA
Composite
%F-F-E-%F
Azinphosmethyl
Endosulfan
Fenvalerate
Methyl Parathion
24.0 (±3.7)
< DL
< DL
< DL
NM = Not Measured
E = Ebb
F = Flood Tide
Lower Limits of Detection:
Azinphosmethyl < 5ng/L
Endosulfan < 3ng/L
Fenvalerate < 2ng/L
Methyl Parathion < Ing/L
< DL = Less than lower limit of detection
96
-------
1990 INSECTICIDE CONCENTRATIONS-KWA
C 08
Figure 8. Measured insecticide concentrations (ug/L) and salinities (ppt) observed at the KWA Site
during the 1990 field study. Values depicted are representative of maximum daily
concentrations at the KWA Site on the dates (•) sampled at ebb tide. While
concentrations depicted are generally representative for the dates sampled, actual pesticide
levels may fluctuate temporally with tidal flushing. Measured insecticide concentrations
depicted are in ug/L rather than the ng/L levels reported in tables and the text. To
convert ug/L to ng/L, multiply by 1000.
97
-------
Table 18. Spiked recovery efficiencies (% recovery) for water samples during
the 1990 field study.
Pesticide
Azinphosmethyl
Endosulfan I
Endosulfan II
Endosulfan Sulfate
Fenvalerate
Methyl Parathion
Spike
Concentration
(ug)
1.260
0.321
0.331
0.271
0.292
0.532
Pooled All Insecticides
Percent Recovery (%)
X
69.2
70.8
78.2
90.0
83.4
82.2
79.0
SE
4.5
3.6
3.3
5.3
4.7
3.8
3.2
Range
55.0-80.0%
62.0- 81.0%
68.0 - 85.0%
79.0 - 102.0
66.0 - 94.0%
71.0-92.0%
55.0- 102.0
98
-------
D. Hydrolab Results for the 1989 Study Period
1. Hvdrolab Results for the 1989 Study Period
During the 1989 study, hydrolabs were deployed at the CTL and TRT Sites only. Results
of hydrolab studies (May 24 - June 13, 1989) are depicted in Figures 9 -20. Figures 9 - 14
depict hydrographic conditions during fair weather periods at both sites. There were three
periods of significant (> 1.27 cm/day) rainfall which occurred during the time of hydrolab
deployment. These occurred on June 5, June 6, and June 15, 1989. The effects of each rain
event on physicochemical water quality are depicted in Figure 15 (CTL Site - June 5, 6),
Figure 10 (TRT Site - June 5, 6), Figure 16 (CTL Site - June 9) and Figure 8 (TRT Site -
June 9). Figures 9 (CTL) and 20 (TRT) depict recovery in physicochemical water quality
following these three major rain events. No hydrolab data were available for the KWA Site;
hence no results for the KWA Site are presented.
During fair weather periods (May 24 - June 4) note the normal tidal (depth, salinity) and
diurnal (water temperature, dissolved oxygen, and pH) fluctuations at the CTL (Figures 9,
11, and 13) and TRT (Figures 10, 12, and 14) Sites. During fair weather periods, note the
small range in salinity at each site (TRT - =26-32 ppt and CTL - 30-33 ppt). Note that the
highest dissolved oxygen levels were observed during periods of maximum daily water
temperatures concomitant with maximum daily pH values. Also note the supersaturated
dissolved concentrations (> 10 mg/L) at the TRT Site on June 3 - 5, 1989 (Figures 12, 14,
16). Also note the concomitant higher pH values observed at the TRT Site during the same
time period. These hydrolab results indicate the dynamic nature of the environment at both
sites.
During the first significant rain event of June 5th, a total of 4,75 and 4.90 cm of rain fell
at the CTL and TRT Sites, respectively. At the CTL Site, this resulted in significant NFS
runoff which caused significant decline in salinity from > 31 ppt to 20 ppt on the first post
rain, ebb tide and a further decline to 7 ppt on the second post rain, ebb tide (Figure 15).
Concomitant with the declines in ebb tide salinity were declines in dissolved oxygen and water
temperature. Note the slight increase in water depth on the second post rain ebb tide. The
only insecticide detected in water samples was endosulfan (2-11 ng/L) generally at or below
background levels (< 10 ng/L) (Table 7).
99
-------
UJatcr Quality Parameters
CTL- Site
5/24-5/28/89
Figure 9. Hydrolab results for the CTL Site, 24-27 May, 1989.
100
-------
lUoter Quality Parameters
TRT- Site
5/24-5/28/89
30 -
23 -
20 -
15 -
10 -
2.5 -
2.0-
1.5 '
1.0 -
0.5-
0.0
I0.0-
8.0-
6.0-
•1.0 '
2.0-
0.0
a —
s-r
7.8-
7.0-
6.6
6.2
34
32
30
28
26
2-1
Figure 10. Hydrolab results for the TRT Site, 24-27 May, 1989.
101
-------
Water Quality Parameters
CTL- Site
5/20-5/31/89
,4=
C 0.
~ a.
a ~
a £
25 -I
20 '
15 -
10 '
5 •
0
2.5-
2.0-
1.5
1.0-
0.5-
0.0
10.0 -
8.0 -
6.0 -
4.0
2.0-
0.0
7.8
7.4
7.0 '
6.6
6.2
34-
32
30-
28
26
24
00 1/1
^ o
oo b
-------
a £
UJaler Quality Parameters
TRT- Site
5/28-3/31/89
2.0 -
1.5 '
1.0-
0.5 -
0.0
10.0 "
a.o-
6.0 J
4.0
2.0 J
0.0
7.8 -
7.4 -
6.2
34
a- JJ -
E "
5=_ 30
,no
^ n
"§
Figure 12. Hydrolab results for the TRT Site, 28-31 May, 1989.
103
-------
c a
~ a
•c —
a £
» —
a
O)
6
mater Quality Parameters
CTL- Site
5/31-6/3/89
l-r
2 .0
1.5
1.0
0.5
0.0
10. G
8.0
6.C
4.0
2.0
0.0
7.8
l.'l
7.0
6.6
6.2
31
32
30
28
26
24
Figure 13. Hydrolab results for the CTL Site, 31 May-3 June, 1989.
104
-------
Water Quality Parameters
TRT- Site
5/31-6/3/89
a. £
«i —
a
01
E
s-r
^
Figure 14. Hydrolab results for the TRT Site, 31 May-3 June, 1989.
105
-------
At the TRT Site (Figure 6) note che much sceeper drop in post rain, ebb tide salinities
from 30 to < 5 ppc. Also there were concomitant declines in dissolved oxygen, pH and
water temperature during the post rain ebb tides. Additionally, water depth was 0.3 - 0.4 m
greater during p~ost rain ebb tides when lowest salinities were observed. This represented the
discharge of significant volumes of NFS agricultural runoff as evidenced by the significant
concentrations of endosulfan (6-20 ng/L), fenvalerate (< DL - 93 ng/L), and azinphosmethyl
( 1.27 cm/day) rain event occurred during the afternoon of June
6, 1989 when 3.43 cm of rain was observed at both the CTL and TRT Sites. Note the
continued decrease in salinities at both sites, during post rain ebb tides (Figures 15 - 18). At
the CTL Site, low tide salinities generally recovered from 7 ppt to 25 ppt by June 8th (~ 48h
post rain) (Figure 17). At the TRT Site, however, ebb tide salinities only recovered slightly
from < 5 ppt to 11 ppt (Figure 18). Dissolved oxygen, water temperature and pH during
this time period remained at levels lower than those found during fair weather periods. Only
slight concentrations of endosulfan (8 -14 ng/L) were observed at CTL Site during this second
rain event (Table 7). At the TRT Site slightly elevated (above background) concentrations
of endosulfan (7-20 ng/L) were observed, although significant (> 96h LC50 values for most
sensitive crustaceans) levels of fenvalerate (< DL - 40 ng/L) were again found (Table 9).
Highest pesticide concentrations were observed at the initial post rain ebb tide (i.e. first
flush).
The third significant (> 1.27 cm/day) rain event occurred on June 9, 1989 when 1.35
cm of rain fell at the CTL Site. A total of 1.57 cm of rainfall was recorded at the TRT Site.
At the CTL Site, this rainfall appeared to have only minimal effects, as post rain low rain
salinities remained at or about 25 ppt, similar to levels measured on June 8 (Figure 19).
Endosulfan concentrations remained at levels at or below background (9 ng/L; background
= < 10 ng/L) (Table 7). The data suggests only minimal runoff occurred at the CTL Site
during, this rain event. At the TRT Site, as low tide salinities fell -from 11 ppt to 6 ppt and
generally remained suppressed (< 10 ppt) at subsequent low tides until June 11, 1989 (1235)
(Figures 18 and 20). In fact, low tide salinities were very slow to recover throughout the post
rain period (June 10-13, 1989) (Figure 20). This suggests continued runoff possibly due to
regulated discharge from the retention
106
-------
mater Quality Parameters
CTL- Site
6/3-0/7/89
" *
Figure 15. Hydrolab results for the CTL Site, 3 -7 June, 1989. Note the effects of rain events on the
5 - 6 June on salinity and other water quality parameters.
107
-------
Water Quality Parameters
TRT- Site
6/3-6/7/89
C 0.
— Q.
1C -
0
2.5 -I
2.0 -
1.5 '
L.O -
0.5 -
0.0
10.0 '
8.0 -
' e.o:
4.0 -
2.0
0.0
7.
7.4-
7.0 -
6.6
6.2
30:
28:
26:
2i:
22
Figure 16. Hydrolab results for the TRT Site, ^3 - 7 June, 1989. Note the effects of rain events on
the 5 - 6 June on salinity and other water quality parameters.
108
-------
c a
a 6
ei
E
s-r
mater Quality Parameters
CTL- Site
6/7-6/10/09
30
2S
:o
LS
1C
s
0
2.5
2.0
1.5
1.0
0.5
0.0
10.0
e.o
6.0
4
o -
2.0
0.0
7.8
7.4
7.0
6.6
6.2
34
32
30
28
26
24
Figure 17. Hydrolab results for the CTL Sitei? - 10 June, 1989. Note the slightly reduced salinities
at this site following rain events on the 5 - 6 June and 9 June, 1989.
109
-------
c a.
— a.
a. E
01
E
a —
s-r
Water Quality Parameters
TRT- Site
6/7-6/10/09
30
25
20
15
1C
5
0
2.5
2.0
1. 5
1
0 -
5-
0
0.0
10.0
8.0
S.O
1.0
2.0
0.0 =
7.8
7.4
7.0
6.6
6.2
32
30
28
26
21
22
—. o>
Ol — O>
K- O
-------
a
*• *
.'§ "
^m |
is *
Water Quality Parameters
CTL- Site
6/10-6/13/89
15
1C
5
0
2.5-
2.0-
1.5
1.0
o.s-
0.0
10.0
a .0-
6.0-
4.0-
2.0-
0.0
7.8"
7.0-
6.6
6.2
34
32-
30
28-
26"
24
CFI
5 f,
Figure 19. Hydrolab results for the CTL Site, 10-13 June, 1989. Salinity at this site has recovered
back to levels comparable to pre-rain conditions.
Ill
-------
Water Quality Parameters
Tnr- Site
6/10-6/13/89
Figure 20. Hydrolab results for the TRT Site, 10-13 June, 1989. Note the continued reduced
salinities at tins site.
112
-------
ponds at this site. Also note that immediately, post rain on June 9, low tide water depths
were 0.2 m higher than previous ebb tides suggesting discharge of fresh water. Chemical
analysis of water samples indicated only background levels of endosulfan (5-7 ng/L) but
significant levels of fenvalerate ( 1.27 cm/day) event occurred with only minimal
levels of pesticides detected at each Site. As a result, hydrolab results for 1990 were not
included in this report.
E. Survival Data for Field Toxicity Test
1. 1989 Field Toxicitv Test
Results of in situ toxicity tests conducted May 25 - June 27, 1989 are listed in Tables
19-23 and depicted in Figures 21 - 25. The dates for caged animal deployment do not
always directly overlap or correspond with the dates of analytical chemical analysis, due
to the different bioassay deployment aSd water sample collection schedules. When
evaluating in situ toxicity test results (Tables 19 - 23) and pesticide analytical results
(Tables 7-17), note dates of deployment and sample collection in interpreting results.
Results of grass shrimp (Table 19 and Figure 21) in situ toxicity tests indicated that:
1) Survival at the CTL Site ranged from 90 - 100%, averaging 96.5% (± 1.46%); 2)
Survival at the TRT Site ranged from 28.5 - 96.7%, averaging 81.8% (±8.53%); and 3)
Survival at the KWA Site ranged from 0.0 - 86.7%, averaging 26.1% (± 9.27%).
Statistical analysis indicated that survival was significantly (p <, 0.05 - 0.01) reduced at the
TRT and KWA Sites compared to the CTL Site due to exposure to fenvalerate (TRT Site -
June 2 - 7, 1989) and combined fenvalerate, endosulfan. and azinphosmethyl (KWA Site -
June 6 - 27, 1989) exposures, respectively. . Additionally, survival at the KWA Site was
significantly (p £ 0.05) lower than af. the TRT Site.
113
-------
Table 19. Summary of survial in P. pugio at all sites during the 1989 field study.
Pooled means with different letters (A,B,C) were significantly (p < 0.05)
different.
Group
#
1
2
3
4
5
6
7
8
1 -8
1989
~" Date
5/25 - 29/89
5/29 - 6/2/89
6/2 - 7/89
6/6 - 1 1/89
6/11 - 15/89
6/15 - 19/89
6/19 - 23/89
6/23 - 27/89
5/25 - 6/27/89
Site
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWAA
KWAB
CTL
TRT
KWA
CTL
TRT
KWAC
KWAD
CTL
TRT
KWA
CTL
TRT
KWAE
KWAF
KWA°-
CTL
TRT
KWA
% Survival
X
100.0
93.3
72.3
96.7
93.3
86.7
93.0
28.5
20.0
90.0
63.3
00.0
00.0
100.0
96.7
36.7
roo.o
93.3
0.0
37.4
100.0
89.3
60.0
92.6
96.7
0.0
0.0
0.0
96.5A
81. 8B
26. lc
SE
0.00
3.33
11.83
3.33
3.33
8.82
3.53
13.82
20.00
5.77
3.33
0.00
0.00
0.00
3.33
3.33
0.00
6.67
0.00
8.12
0.00
6.43
15.28
7.41
3.33
0.00
0.00
0.00
1.46 Range = 90.0- 100.0
8.53 Range = 28.5 - 96.7
9.27 Range = 0.0 - 86.7
A = Group deployed
B = Group deployed
C = Group deployed
D = Group deployed
E = Group deployed
F = Group deployed
G = Group deployed
from 6/6 - 8/89.
from 6/8 - 11/89.
from 6/15 - 19/89.
from 6/18- 23/89.
from 6/23 - 24/89.
from 6/24- 26/89.
from 6/25 - 27/89.
114
-------
GRASS SHRIMP SURVIVAL(1989)-CTL
100 -\
8G -
60 -
43 -
2C -
n -
°- - -n g — — o a ° 5
100 -
80 -
60 -
40 -
20 -
0
GRASS SHRIMP SURVIVAL(1989)-TRT
GRASS SHRIMP SURVIVAL(1989)-KWA
IO -~-
to
I
Figure 21. Survival of P. pugio in field toxicity tests during the 1989 field study. Note the
significant mortality observed at the KWA Site. Also note one period of reduced
survival(6/2-7/89) at the TRT Site.
115
-------
Results of penaied shrimp (Table 20 and Figure 22) in situ toxicity cest indicated that:
1) Survival at the CTL Site ranged from 58.3 - 100%, averaging 90.9% (± 4.9%); 2)
Survival at the TRT Site ranged from 51.9 - 100%, averaging 92.3% (± 6.01%); and 3)
Survival at the_ KWA Site ranged from 0.0 - 35.8%, averaging 5.7% (± 4.47%).
Statistical analysis indicated that survival at the KWA Site was significantly (p < 0.001)
lower than at the CTL and TRT Sites. This resulted from significant runoff of
azinphosmethyl, endosulfan, and fenvalerate at this site. Significant mortality c48%) was
also observed at the TRT Site following the rain events of June 5 - 6, 1989, when high
concentrations of fenvalerate were measured in runoff at the site.
Results of mysid shrimp (Table 21 and Figure 23) in situ toxicity tests indicated that:
1) Survival at the CTL Site ranged from 0 -100%, averaging 40.0% (± 24.5%); Survival
at the TRT Site ranged from 1 - 100%, averaging 60.0% (±24.5); and 3) Survival at the
KWA Site ranging from 0 - 100%, averaging 37.2% (± 22.87). Poor survival was
observed at all sites due to: 1) the inability of mysids to survive low salinity (< 10 ppt)
exposure following significant rainfall (i.e. Group 3 - June 2 - 7, 1989); and 2) problem
with cage fouling on the tether line, (i.e. CTL Site - Groups 1 - 2). Statistical analysis
indicated no significant between site differences in survival, despite cage deployment
problems, described above.
Results of mummichogs in situ toxicity tests (Table 22 and Figure 24) indicated that:
1) Survival at the CTL Site ranged form 96.3 - 100%, averaging 98.7% (± 0.63%); 2)
Survival at the TRT Site ranging from 96.7 - 100.00%, averaging 99.6% (± 0.41%); and
3) Survival at the KWA Site ranging from 83.3 - 100.0%, averaging 97.1% (± 2.04%).
Statistical analysis generally indicated no significant between site difference in survival.
Survival at the KWA Site, Group 8 (83.3 ± 6.67%) was significantly (p < 0.05) reduced,
however, compared to the CTL and TRT Sites due to significant azinphosmethyl runoff.
Results of juvenile sheepshead minnow in situ toxicity tests (Table 23 and Figure 25)
indicated that: 1) Survival at the CTL Site ranged from 36.1 - 86.7%. averaging 56.1%
(± 15.53%); 2) Survival at the TRT Site ranged form 80 - 100%, averaging 92.1% (±
6.14%); and 3) Survival at the KWA Site ranged from 0 -44%, averaging 24.9% (±
13.04%). Statistical analysis indicated that survival was significantly
116
-------
Table 20. Summary of survival in Penaeus Species at all sites during the 1989
field study. Pooled means with the different letters (A,B) were not
significantly (p ^ 0.05) different.
Group
ft
1
2
3
4
5
6
7
8
1 - 8
1989
Date
5/25 - 29/89
5/29 - 6/2/89
6/2 - 7/89
6/7- 11/89
6/11 - 15/89
6/15 - 19/89
6/19 - 23/89
6/23 - 27/89
5/25 - 6/27/8?
Site
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
% Survival
X
93.3
100.0
10.0
100.0
100.0
0.0
89.6
51.9
0.0
100.0
86.7
0.0
100.0
100.0
"; 0.0
58. 3l
100.0
0.0
95.8
100.0
35.8
90.5
100.0
0.0
: 90.9*
92. 3A
5.7B
SE
6.67
0.00
5.77
0.00
0.00
0.00
0.37
6.06
0.00
0.00
8.82
0.00
0.0
0.0
0.0
23.95
0.00
0.00
4.17
0.00
8.70
4.76
0.00
0.00
4.90 Range = 58.3 - 100.0
6.01 Range = 51.9- 100.0
4.47 Range = 0.0 - 35.8
Mortality caused by extremely heavy siltation in cages following
heavy rains at dead low tide which eroded large quantities of
sediment into Leadenwah Creek.
117
-------
PENAIEO SHRIMP SUR VIVAL(1989)-CTL
inc -
es -
GO -
40 -
20 -
3
100 -
80 -
60 -
40 -
20 -
0
PENAIED SHRIMP SURVIVAL(1989)-TRT
100
80 -
60 -
30 -
20 -
0
PENAIEO SHRIMP SURVIVAL(1989)-KWA
Figure 22 Survival of ft/w«tf species in field toxicity tests during the 1989 field study. Note the
F.gure 22. Survival o ^^ ^ ^^ ^.^ ^ ^ ^^ ^ AlsQ ^ Qne penod of
reduced sarvival(6/2-7/89) at the TRT Site.
118
-------
Table 21. Summary of survival in Mysidopsis bahia at all sites during the 1989 field
study. Pooled means with the same letter (A) were not significantly (p >
0.05) different.
Group
#
1
2
3
4
5
1 - 5
1989
Date
5/25 - 29/89
5/29 - 6/2/89
6/2 - 7/89
6/7 - 11/89
6/11 - 15/89
5/25 -6/15/89
Site
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
% Survival
X
o.o1
100.0
100.0
o.o2
100.0
85.8
O.O3
O.O3
O.O3
100.0
0.0
0.0
100.0
100.0
1 0.0
40.0*
60.0*
37. 2A
SE
0.00
0.00
0.00
0.00
0.00
14.20
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
24.50 Range = 0-100
24.50 Range = 0-100
22.87 Range = 0 - 14.2
2 _
Control Group mortality due to cage becoming
aerially exposed resulting in dessication.
Control Group mortality die to cage becoming hung
up due to extremely high tides.
Mortality at all site's probably due to effects of
low salinity (CTL) and low salinity -pesticide
exposures (TRT and KWA).
119
-------
MYSID SHRIMP SURVIVAL(1989)-CTL
>
tr
ce
3
CJ
in
(M
Ol
(M
in
>
K
(A
100 -
80 -
60-
40 -
20-
0
MYSID SHRIMP SURVIVAL(1989)-KWA
in
CM
IA
Figure 23. Survival of Mysidopsis bahia in field toxicity tests during the 1989 field study. Note the
generally poor survival at all field sites.
120
-------
Table 22. Summary of survival in F. heteroclitus at all sites during the 1989 field study.
Pooled means with the same letter (A) were not significantly
(p > 0.05) different.
Group
ft
1
2
3
4
5
6
7
8
1 - 8
1989
Date
5/25 - 29/89
5/29 - 6/2/89
6/2 - 6/7/89
6/7- 11/89
6/11- 15/89
6/15 - 19/89
6/19-23/89
6/23 - 27/89
.
5/25 - 6/27/89
Site
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
% Survival
X
LOO.O
96.7
LOO.O
96.7
100.0
96.7
96.7
100.0
100.0
100.0
100.0
100.0
96.3
300.0
100.0
100.0
100.0
100.0
100.0
100.0
96.7
100.0
100.0
. 83.3
98. 7A
99. 6A
97.1*
SE
0.00
3.33
0.00
3.33
0.00
3.33
3.33
0.00
0.00
0.00
0.00
0.00
3.70
0.00
0.00
0.00
0.00
0.00
0.00
0.00
3.33
0.00
0.00
6.67
0.63 Range = 96.3 - 100.0
0.41 Range •= 96.7 - 100.0
2.04 Range = 83.3 - 100.0
121
-------
100 -
80 -
GO -
JO -
;o -
o
MUMMICHOG SURVIVAL(1989)-CTL
60 -
-------
Table 23. Summary of survival in juvenile Cyprinodon variegatus at all sites during the
1989 field study. Pooled means with different letters (A,B,C) were
significantly (p < 0.05) different.
Group
#
1 '
2
3
1 - 3
1989
Date
5/24 - 28/90
5/28 - 6/1/89
6/11 - 27/89
5/25 - 6/27/89
Site
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
% Survival
X
96.7
36.7
100.0
93.0
96.7
93.3
36.1
100.0
0.0
56. 1A
92.1"
24. 9C
SE
3.33
20.28
0.00
3.53
3.33
3.33
7.35
7.35
0.00
15.53 Range = 36.1 - 86.7
6.14 Range = 80.0 - 100.0
13.04 Range = 0.0 - 44.0
123
-------
CYPRINODON SURVIVAL(1969)-CTL
IfiO
80
60
100 -
80 -
60 -
40 -
20 -
0
CYPRINODON SURVIVAL(1989)-TRT
100 -
80-
60 -
40 -
20 -
0
CYPRINODON SURVIVAL(1989)-KWA
Figure 25. Survival of Cyprinodon variegatus in field toxicity tests during the 1989 field study. Note
the reduced survival at the KWA Site and variable survival at the CTL Site.
124
-------
(p < 0.05) higher at TRT Site when compared to the CTL and KWA Sites. Additionally,
survival at the CTL Site was significantly (p < 0.05) higher than at the KWA Site.
Survival at the TRT Site was high despite the significant runoff of fenvalerate observed
following significant rainfall on June 5, June 6 and June 9. Reduced survival at the CTL
Site was most likely related to prolonged exposure duration (14 - 16 days), as most
mortality was incurred beyond 10 days post deployment. This may have resulted due to
starvation. Although fish at both sites were deployed for the same time using the same
cage type, food availability was dependent upon what passed through the mesh on each
container. At the TRT Site, cages were deployed in a shallower stream stretch than at the
CTL Site, which may have resulted in greater food availability. Differences in survival
between the TRT and CTL Sites may have resulted from differences in food availability
among animals deployed at each site. This generally was not the case at the KWA Site as
mortality occurred concomitant with periods of significant runoff and pesticide exposure
(azinphosmethyl and endosulfan), irrespective of exposure (i.e. food availability) duration.
Most mortalities at the KWA Site were attributed to azinphosmethyl exposure rather than
endosulfan or fenvalerate exposure. The basis for this conclusion were two fold. First
significant AChE enzyme inhibition was observed in organisms at the KWA Site indicating
significant azinphosmethyl exposure. Secondly, levels of endosulfan and fenvalerate were
generally below 96h LC^ values reported for the species tested.
2. Quality Assurance and Quality Control for Bioassay Organisms Used in Field
Toxicity Test during the 1989 Field Study
Results of quality assurance and control bioassays conducted during 1989 using
endosulfan as a reference toxicant are listed in Table 24.
Results for adult P. pugio exposed to endosulfan indicated mortalities ranging from 0 -
40%, averaging 8% (± 8%) at 0.01 ug/L; from 60 - 100%, averaging 72% (± 8%) at
1.15 ug/L; and from 100 - 100%, averaging 100% (± 0%) at 2.50 ug/L. Control
mortality ranged from 0 - 0%, averaging 0% (± 0%) in high salinity (> 25 ppt) controls
and was also 0% (N ="l) in low salinity controls (2ppt). The 96h LC50 for endosulfan
P. pugio was 0.18 ug/L (CL = 0.10 - 0.33 ug/L), which was comparable to previous
reported LC50 values (0.25 - 1.01 ug/L) for grass shrimp (Scott et al, 1990).
Results for adult F. heteroclitus exposed to endosulfan indicated mortalities ranging
from 0 - 0%, averaging 0% (± 0%) at 0.01 ug/L; from 0 - 40%, averaging 8% (± 8%)
at 1.15 ug/L; 60% at 2.50 ug/L (n = 1); and from 100 - 100%, averaging 100%
125
-------
Table 24. Summary of Quality Control/Quality Assurance Bioassays for 1989 using cndosulfan
as the reference toxicant, for (he five test periods of the study.
Nominal
Endosulfan
Concentration
0.00'
o.oo2
0.01
1.15
2.50
Species (*)
P. pugio
(A)
% Mortality4
n
0
ND
0
60
100
#2
0
0
40
60
100
#3
0
ND
0
60
100
#4
0
ND
0
100
100
0.00
0.01
1.15
2.50
5.00
F. heteroclitus
(A)
0
0
0
60
ND
0
0,,
40
ND
100
0
0
0
ND
100
0
0
0
ND
100
O.OO1
O.OO2
0.01
1.15
2.50
Penaeus sp.
(J)
20
ND
20
80
80
0
40
0
80
100
0
ND
0
100
100
0
ND
0
33
100
#5
ND
ND
0
80
ND
Pooled Results
X
0.0
00
8.0
72.0
100.0
(± SE)
(1 0.00)
(NC)
' (± 8.00)
(± 8.00)
(± 0.00)
96h LC50 - 0.1 8 ug/L
(59%CL = 0.10 -0.33 ug/L)
0
0
0
ND
100
0.0
00
8.0
60.0
100.0
(± 0.00)
(± 0.00)
(± 8.00)
(NC)
(± 0.00)
96h LC5fl = 1 .82 ug/L
(95%CL = 1.18 - 2.83 ug/L)
ND
ND
0
60
ND
™
5.0
40.0
4.0
70.6
95.0
(± 5.00)
(NC)
(± 4.0)
(± 11.33)
(± 5.00)
48h LC5fl = 0.18ug/L
-------
Nominal
Endosulfan
Concentration
Species (*)
% Mortality4
n
#2
#3
#4
0.00J
0.01
1.15
2.50
C. variegatus
U)
0
0
0
100
0
0
100
100
40
40
100
100
0
0
25
ND
#5
Pooled Results
X
(± SE)
(95%CL = 0.13 -0.75ug/L)
ND
ND
ND
ND
10.0
10.0
56.3
100.0
(± 10.0)
(± 10.0)
(±25.8)
, (± 0.0)
96h LC50 = 0.31 ug/L
(95% CL = 0.13- 0.75 ug/L)
K)
* = Lifestage: A = Adult; J = Juvenile
ND = NotJJetermined
NC = Not Calculated •
' i
1 = High Salinity Control = 20 ppt salinity
2 = Low Salinity Control = 2 ppt salinity
3 = High Control Mortality due to handling stress
4 = Exposure periods were 96 hours for all species except Penaeus sp. which was 48 hours.
-------
(± 0%) at 5.00 ug/L. Control mortality ranged from 0-0%, averaging 0% (± 0%). The
96h LC50 for endosulfan F. heteroclitus was 1.82 ug/L (CL = 1.18-2.83 ug/L), which
was comparable to previously reported 96h LC50 valves (1.29 - 1.45 ug/'L) for
mummichogs (Scott et al. 1990).
Results for juvenile penaied shrimp (Penaeus aztecus and Penaeus setiferus) exposed
to endosulfan indicated 48h mortalities ranging from 0-20%, averaging 4% (+ 4%) at 0.01
ug/L; from 33 - 100%, averaging 70.6% (± 11.3%) at 1.15 ug/L; and from 80 - 100%,
averaging 95.% (± 5%) at 2.50 ug/L. Control mortality ranged from 0 - 20% averaging
5% (± 5%) at high salinity (> 20 ppt). At low salinities (2 ppt - used to simulate KWA
Site) mortality was 40% (N = 1). The 48h LC50 for endosulfan in juvenile penaied
shrimp was 0.18 ug/L (CL = 0.09 - 0.38 ug/L).
Results for juvenile C. variegatus exposed to endosulfan indicated mortalities ranging
from 0-40%, averaging 10% (± 10%) at 0.01 ug/L; from 0 - 100%, averaging 56.3%
(± 25.8%), and from 100 -100%, averaging 100% (± 0%). Control survival ranged from
0 - 40%, averaging 10% (± 10%). All control mortality occurred during QA Test #3,
suggesting possible handling stress. The 96h LC50 for endosulfan in juvenile C. variegatus
was 0.31 ug/L(CL = 0.13 - 0.75 ug/L). r
These results generally indicated that each group of organism were comparable to
previously reported (Scott et al, 1990) acute toxicity results in terms of their response to
endosulfan exposure. Although some variation was noted between groups, much of this
may have resulted from differences in ambient temperature, given the inverse or negative
temperature coefficient for endosulfan (i.e. less toxic at higher temperatures).
3. 1990 Field Toxicity Tests
Results of in situ toxicity tests conducted May 24 - June 23, 1990 are listed in Tables
25 - 3D and depicted in Figures 26 - 31.
i
Results of grass shrimp (Table 2$ and Figure 26) in situ toxicity tests indicated that:
1) Survival at the CTL Site ranged from 93 - 100%, averaging 98.3% (± 0.92%), 2)
Survival at the TRT Site ranged from 36.7 - 100.0%, averaging 90.4% (± 7.69%); and
128
-------
Table 25. Summary of survial in P. pugio at all sites during the 1990 field study. Pooled
means with the same letters (A) were not significantly (p > 0.05) different.
Group
#
1
2
3
4
5
6
7
8
1 - 8
Date
5/24 - 28/90
5/28 - 6/1/90
6/1 - 5/90
6/5 - 9/90
6/9 - 13/90
6/13 - 17/90
6/17 - 21/90
6/21 - 23/90
5/24 - 6/23/90
Site
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
% Survival
X
96.7
36.7'
100.0
93.0
93.7
93.3
100.0
100.0
86.7
96.7
96.7
96.7
100.0
96.7
97.0
;oo.o
roo.o
96.7
100.0
100.0
96.7
100.0
96.7
100.0
98. 3 A
90.4A
95. 9*
SE
3.33
20.28
0.00
3.53
3.33
3.53
0.00
0.00
3.33
3.33
3.33
3.33
0.00
3.33
3.03
0.00
0.00
3.33
0.00
0.00
3.33
0.00
3.33
0.00
0.92 Range = 93.0 - 100.0
7.69 Range = 36.7 - 100.0
1.52 Range = 86.7 - 100.0
= Cage at TRT Site turned on side resulting in aerial exposure and desiccation.
129
-------
GRASS SHRIMP SUR VIVAL(1990J-CTL
<
5
rr
en
a"
130 -
80 -
60 -
40 -
20 -
Q O__— — -a c a a o a.
GRASS SHRIMP SURVIVAL(1990)-TRT
100 -
80 -
60 -
40 -
20 -
0
GRASS SHRIMP SURVIVAL(1990)-KWA
100 -i
80 •
60 -
40 -
20 -
Figure 26. Survival of P. pugio in field toxicity tests during the 1990 field. Note the high survival
at all sites.
130
-------
3) Survival at the KWA Site ranged from 86.7 - 95.9%, averaging 95.9% (± 1.52%),
Statistical analysis indicated no between site differences in survival during the 1990 study
period, despite significant fenvalerate runoff (concentration > 96h LC50 value for
crustaceans) at the TRT Site (Table 16 and Figure 7). It is noteworthy that while one grab
sample exceeded the 96h LC50 values, the composite sample (time- weighted average
exposure) was < DL. This suggests that oniy a small volume discharge of fenvalerate
occurred at the TRT Site, such that even though peak concentrations at ebb tide exceeded 96h
LC50 values, rapid dilution by incoming tides, reduced in-stream concentrations to below toxic
thresholds.
Results of juvenile penaied shrimp (Table 26 and Figure 27) in situ toxicity tests indicated
that: 1) Survival at the CTL Site ranged from 96.3 - 100%, averaging 99.5% (± 0.46%);
2) Survival at the TRT Site ranging from 96.7 - 100.0%, averaging 99.2% (± 0.54%); and
3) Survival at the KWA Site ranging from 96.3 - 100.0%, averaging 98.7% (± 0.65%).
Statistical analysis indicated no between site differences in survival during the 1990 field
study.
Results of mysid shrimp (Table 27 and Figure 28) in situ toxicity tests indicated that: 1)
Survival at the CTL Site ranged from 22.2 - 100.0%, averaging 80.9% (± 12.88%); 2)
Survival at the TRT Site ranged from 8.3 - 100%, averaging 76.3% (± 22.66%); 3) Survival
at the KWA Site ranged 77.8 - 100.0%, averaging 94.5% (± 5.55%). Statistical analysis
indicated no significant between site differences in survival during the 1990 Study. Low
survival was observed at the CTL (Groups 2 and 4) and TRT (Group 1) when cages were hung
up on the tether line, resulting in aerial exposure and desiccation. Significant runoff of
fenvalerate at the TRT Site on May 28 was not assessed due to the cage deployment problems
just described.
Results of mummichog (Table 28 and Figure 29) in situ toxicity tests indicated that: 1)
Survival at "the CTL Site" ranged from"93.3 - 100.0%, averaging 98.3% (± 0.91%); 2)
Survival at the TRT Site ranged from 93.3 - 100%, averaging 98.3% (± 0.89%); and 3)
Survival at the KWA Site ranged from 100 - 100%, averaging 100.0% (± 0%). Statistical
analysis indicated no significant between site differences during the 1990 field study.
131
-------
Table 26. Summary of survival in penaeus species at all sites during
the 1990 field study. Pooled means with the same letter (A) were not
significantly (p > 0.05) different.
Group
#
1
2
3
4
5
6
7
8
I -8
Date
5/24 - 28/90
5/28 - 6/1/90
6/1 - 5/90
6/5 - 9/90
6/9 - 13/90
6/13 - 17/90
6/17 - 21/90
6/21 - 23/90
5/24 - 6/23/90
Site
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA -
CTL
TRT
KWA
% Survival
X
100.0
100. 0
100.0
100.0
100.0
100.0
100.0
100.0
96.7
100.0
100.0
100.0
100.0
foo.o
100.0
100.0
100.0
100.0
100.0
96.7
96.3
96.3
96.7
96.3
99. 5 *
99.2*
98.7*
SE
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
3.33
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
3.33
3.70
3.70
3.33
3.33
0.46 Range - 96.3 - 100.0
0.54 Range = 96.7 - 100.0
0.65 Range = 96.3 - 100.0
132
-------
PENAIED SHRIMP SURVIVAL(1990)-CTL
_.
<
5
cc
3
I/)
^
'.CO -
80 -
GO -
JO -
20 -
c B a a a o c
1 D
PENAIED SHRIMP SURVIVAL(1990)-TRT
100 -
< 80 -
5 60 -
d
= 40 -
en
20-
a a a a a a— M g
PENAIED SHRIMP SURVIVAL(1990)-KWA
'.00 -
< 80-
> 60 -
cr
3 40 -
00 20-
* n
a a Q o a o Q Q
Figure 27. Survival of Penaeus species in fie\d toxicity tests during the 1990 Held study. Note the
high survival at all sites.
133
-------
Table 27. Summary of survival in Mysidopsis bahia at all sites during the 1990 field
study. Posted means with the same letter (A) were not significantly (p >
0.05) different.
Group
1
2
3
4
1 -4
Date
5/28 - 6/1/90
6/1 - 5/90
6/5 - 9/90
6/9 - 13/90
5/28 - 6/13/90
Site
CTL
TRT
KWA
CTL1
CTL2
TRT
KWA
CTL
TRT
KWA
CTL3
CTL4
TRT
KWA
CTL
TRT
KWA
% Survival
X
96.3
8.3'
100.0
66.7
100.0
100.0
100.0
100.0
96.7
77.8
22.20
100.0
100.0
100.0
80.9*
76.3A
94. 5 A
SE
3.70
8.33
0.00
33.33
0.00
0.00
0.00
0.00
3.33
11.11
22.20
0.00
0.00
0.00
12.88 Range = 22.2 - 100.0
22.66 Range = 8.3 - 100.0
5.55 Range = 77.8 - 100.0
Low survival due to cage hanging up on tether resulting in aerial exposure
and dessication.
Group 21 deployed 6/1 - 3/90 and resulting mortality occurred due to
problems with cage deployment.
= Group 22 deployed 6/3 - 5/90 and problem was corrected.
Group 43 deployed 6/9 - 13/90 and resulting mortality was due to problems
with cage deployment.
= Group 44 deployed 6/11 - 13$0 and deployment problem was corrected.
134
-------
MYSID SHRIMP SURVI VAL(1990)-CTL
100 -
80 -
60 -
40 -
20 -
0
MYSID SHRIMP SURVIVAL(1990)-TRT
100 -
80 -
60 -
40 -
20 -
0
MYSID SHRIMP SURVIVAL(1990)-KWA
Figure 28. Survival of Mysidopsis bahia in fie(d toxicity tests during the 1990 field study. Note the
generally good survival at all sites.
135
-------
Table 28. Summary of survival in F. heteroclitus at all sites during the 1990 field study.
Pooled means with the same letter (A) were not significantly (p > 0.05)
different.
Group
#
1
2
3
4
5
6
7
8
1 -8
Date
5/24 - 28/90
5/28 - 6/1/90
6/1 - 5/90
6/5 - 9/90
6/9 - 13/90
6/13 - 17/90
6/17 - 21/90
6/21 - 23/90
b
5/24 - 6/23/90
Site
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT .
KWA-
CTL
TRT
KWA
% Survival
X
100.0
93.3
100.0
96.3
96.7
100.0
100.0
100.0
100.0
93.3
100.0
100.0
UOO.O
100.0
100.0
96.7
100.0
100.0
100.0
100.0
100.0
100.0
96.7
100.0
98.3A
98.3A
100.0*
SE
0.00
6.67
0.00
3.70
3.33
0.00
0.00
0.00
0.00
6.67
0.00
0.00
0.00
0.00
0.00
3.33
0.00
0.00
0.00
0.00
0.00
0.00
3.33
0.00
0.91 Range = 93.3 - 100.0
0.89 Range = 93.3 - 100.0
0.00 Range = 100.0 - 100.0
136
-------
MUMMICHOG SURVIVAL (1990)-CTL
_J
•t
>
->
rr
D
ffl
100 -
00 -
GO -
40 -J
20 J
o -Q a- _v- — — ~° a — a a
~^r ™
,
MUMMICHOG SURVIVAL(1990)-TRT
^
^
nr
D
tfi
100 -
80 -
60 -
40 -
20 -
g g a o a a o a
MUMMICHOG SURVIVAL(1990)-KWA
too
80
60
40
20
t
Figure 29. Survival of Fundulus heteroditus in field toxicity tests during the 1990 field study. Note
the high survival at all sites.
137
-------
Results of juvenile sheepshead minnow (Table 29 and Figure 30) in situ toxiciry test
indicated that: 1) Survival at the CTL Site ranged from 60.9 - 100.0%, averaging 89.4%
(± 9.53%); 2)_Survival at the TRT Site ranged form 64.1 - 100.0%, averaging 91% (±
8.98%) and 3) Survival at the KWA Site ranged from 83.0 - 100%, averaging 95.8% ( +
4.18%). Statistical analysis indicated no significant between site differences in survival.
The reduced deployment time (7 - 8 days) used in 1990 (versus 14 - 15 day* in 1989)
appeared to greatly enhance survival. A seven - eight day deployment time appears
optimal.
Results of juvenile Menidia berylina (Table 30 and Figure 31) survival in field toxiciry
tests indicated: 1) Survival at the CTL Site ranged from 0 - 40%, averaging 17.2% (±
5.20); 2) Survival at the TRT Site ranged from 0 - 100%, averaging 66.7% (± 22.61%);
and 3) Survival at the KWA Site ranged from 0 - 65.7%, averaging 23.4% (± 12.91%)
Statistical analysis indicated no significant between site differences in survival. Juvenile
M. berylina appeared to have poor survival at all sites. Deployment in different cages and
in different exclusion cage positions (surface and bottom) appeared to have little effect on
survival (see Group 4 - CTL Site - Table 30). Other factors, such as extreme current flow
or low dissolved oxygen levels in the tidal creek, may have been stressful enough to cause
mortality.
4. Quality Assurance and Quality Control for Bioassay Organisms Used in Field
Toxicity Tests during the 1990 Field Study
Results of quality assurance and quality control bioassays conducted during 1990, using
endosulfan as a reference toxicant are listed in Table 31.
Results for adult P. pugio exposed to endosulfan indicated mortalities ranging from 0 -
20%, averaging 10% (± 5.78%) at 9.01 ug/L; from 60 - 100%, averaging 75% (±
9.58%) at 1.15 ug/L; and from 80 - 100%, averaging 90% (± 5.78%) at 2.50 ug/L.
Control mortality ranged from 0 - 20%, averaging 5% (± 5.0%). The 96h LC50 for
endosulfan in P. pugio during 1990twas 0.18 ug/L (CL = 0.08 - 0.39 ug/L) which
compared favorable with previously reported 96h LC50 values of 0.25 - 1.01 ug/L (Scott
et al, 1990) and with 1989 Quality Control bioassay results (LC50 = 0.18 ug/L with CL
= 0.10-0.33 ug/L).
138
-------
Table 29. Summary of survival in juvenile Cyprinodon vareigatus at all sites during
the 1990 field study. Pooled mans with different letters (A) were not
significantly (p > 0.05) different.
Group
1
2
3
4
1 -4
Date
5/24 - 6/2/90
6/2 - 10/90
6/10 - 17/90
6/17 - 24/90
5/24 - 6/24/90
Site
CTL
TRT1
KWA
CTL
TRT2
KWA
CTL
TRT
KWA
CTL
TRT
KWA
CTL
TRT
KWA
% Survival
X
60.9*
64.1*
83.3
100.0
100.0
100.0
100.0
100.0
100.0
'96.7
100.0
100.0
89.4A
91.0A
95. 8 A
SE
15.48
21.12
16.70
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
9.53 Range = 60.9 - 100.0
8.98 Range = 64.1 - 100.0
4.18 Range = 83.0 - 100.0
= TRT Group 1 deployed from 5/24 - 6/1/90.
= TRT Group 2 deployed from 6/f-6/10/90.
* =
Mortality was caused by extremely heavy situations in cages following heavy rains
at ebb tide which eroded large\quantities of sediment into Leadenwah Creek.
139
-------
tr
D
00
80 -
GO -
40 -
20 -
0
CYPRINODON SURVIVAL(1990)-CTL
CYPRINODON SURVIVAL(1990)-TRT
i
5
rr
3
tfl
100 -
80 -
60 -
40 -
20 -
n -
^ o a
\"
I
cr
-------
Table 30.
Summary of survial in M. berylina at all sites during the 1990 field
study. Pooled means with same letter (A) were not significantly
(p > 0.05) different. All deployments were in Menidia cage unless
other noted.
Group
#
I
2
3
4
1 - 8
— Date
5/24 - 28/90
5/28 - 6/2/90
6/2 - 10/90
6/10 - 18/90
5/24 - 6/18/90
Site
CTL
TRT
KWA
CTL
TRT
KWA
CTL1
CTL2
CTL3
TRT
KWA4
KWA5
CTL6
CTL7
CTL8
TRT
KWA9
CTL
TRT
KWA
% Survival
X
0.0
0.0
6.7
40.0
100.0
65.7
36.7
14.0
6.7
80.0
3.3
0.0
10.0
23.3
6.7
86.7
'41.1
17. 2A
66. 7 A
23. 4A
SE
0.00
0.00
6.67
30.55
0.00
8.69
12.02
4.0
6.67
5.77
3.33
0.00
0.00
23.33
3.33
13.33
24.05
5.20 Range = 0.0 - 40.0
22.60 Range = 0.0 - 100.0
12.91 Range = 0.0 - 65.7
1 = Group 31 deployed 6/2 - 4/90 at CTL Site.
2 = Group 32 deployed 6/4 - 7/90 at CTL Site.
3 = Group 33 deployed 6/7 - 10/90 at CTL Site.
4 = Group 34 deployed 6/4 - 8/90. at KWA Site.
5 = Group 33 deployed 6/7 - 8/90 at KWA Site.
6 = Group 46 deployed 6/10 - 13/90 in Cyprinodon
cage at CTL Site on the sur&ce.
7 = Group 47 deployed 6/10 - 13/90 in Menidia cage
at CTL Site on the bottom.
8 = Group 4* deployed 6/10 - 13/90 in Cyprinodon cage at the CTL Site on
the bottom.
9 = Group 49 deployed 6/10 - 13/90 at KWA Site.
141
-------
MENIDIA SURVIVAL(1990)-CTL
'00
5 fie
1 ,r,
100
80
60
40
20
0
MENIDIA SURVIVAL(1990)-TRT
MENIDIA SURVIVAL(1990)-KWA
Figure 31. Survival of Menidia berylina in field toxicity tests during the 1990 field studies. Note
the generally poor survival at all sites.
142
-------
Table 31. Summary of Quality Control/Quality Assurance Bioassays using endosulfan as the reference toxicant I'or (lie five
toxicity tests conducted over the course or the 1990 study.
Nominal
Endosulfan
Concentration
0.00
0.01
1.15
2.50
Species (*)
P. pugio
(A)
% Mortality1
#1
0
0
60
100
n
20
20
100
80
#3
0
20
80
100
0.00
1.15
2.50
5.00
F. heteroclitus
(A)
0
0
60
100
0
0
1f
20
80
0
0
25
100
0.00
0.01
1.15
2.50
Penaeus sp.
(J)
0
20
40
80
0
0
80
80
0
0
100
100
05
0
0
60
50
Pooled Results
X
5.0
10.0
75.0
90.0
(± SE)
(± 5.00)
(± 5.78)
(± 9.58)
(± 5.78)
96h LCJO = 0.1 8 ug/L
(95% CL = 0.08 -0.39ug/L)
0
0
0
80
0.0
0.0
26.3
90.0
(± 0.00)
(± 0.00)
(± 12.48)
(± 5.78)
96h LCSO = 0.31 ug/L
(95% CL = 2.55 - 3.86ug/L)
0
0
40
60
0.0
5.0
65.0
80.0
(± 0.00)
(± 0.00)
(± 15.00)
(± 8.17)
48h LC50 = 0.31 ug/L
(95% CL = 0.13 - 0.77 ug/L)
-------
Nominal
Endosulfan
Concentration
0.00
0.01
1.15
2.50
Species (*)
C. variegatus
% Mortality1
ff\
0
0
0
100
n
0
0
0
100
#3
0
0
0
60
0.00
0.01
1.15
2.50 —
M. berylina
(J)
0
20
100
100
25
0
100
100
202
20
100
100
i ,1
#5
0
0
20
20
Pooled Results
X
0.0
0.0
50
70.0
(± SK)
(± 0.00)
(± 0.00)
(± 5.00)
(± 19.15)
96h LQ0 = 1.97*g/L
(95% CL = 1.66-2.34 ug/L)
ND
ND
ND
ND
15.0
13.3
1000
100.0
(± 7.64)
(± 6.67)
(± 0.00)
(± 0.00)
96h LC50 = 0.07 ug/L
(95% CL = NC)
2
NC
= Life Stage: A = Adult; J = Juvenile
— Exposure periods were 96h for all species expect Penaeus species, which was 48 hours.
= Handling stress caused mortality
= Confidence Limits not calculatedTable 31
-------
Results for adult F. heteroditus exposed to endosulfan indicated mortalities ranging from
0 - 0%, averaging 0% (± 0%) ac 1.15 ug/L; from 0 - 60%, averaging 26.3% (± 12.48%)
at 2.50 ug/L; and 80 - 100%, averaging 90% (± 5.78) at 5.00 ug/L. Control mortality
ranged from CK 0%, averaging (± 0%). The 96h LC50 for endosulfan in F. heteroditus
during 1990 was 3.14 ug/L (CL = 2.55 - 3.86 ug/L) which compared favorably with
previously reported LC50 values of 1.29 -1.45 ug/L (CL = 1.29 - 1.59 ug/L) by^Scott et al
(1990) and with results for 1989 QA/QC results (LC50 = 1.82 ug/L with CL = 1.18- 2.83
ug/L). The slightly higher LC50 value in F. heteroditus obtained during 1990 was largely the
result of higher exposure temperature, particularly during the second and fourth tests. Given
the inverse temperature coefficient for endosulfan, acute toxicity was reduced with these
higher temperature. Results for 1989 and 1990 were not significantly different in statistical
comparisons (upper and lower 95% CL overlap).
Results for juvenile penaied shrimp (Penaeus aztecus and Penaeus setiferus) exposed to
endosulfan indicated 48h mortalities ranging form 0 -20 %, averaging 5% (± 5%) at 0.01
ug/L; from 40 - 100%, averaging 65% (± 15%) at 1.15 ug/L; and from 60 - 100%,
averaging 80% (± 8.17%). Control mortality ranged from 0 -0%, averaging 0% (± 0%).
The 48h LCM for endosulfan in juvenile penaied shrimp was 0.31 ug/L (CL = 0.13 - 0.77
ug/L) which compared favorably with 1989.QA/QC results for penaied shrimp (LC50 = 0.18
ug/L with CL = 0.09 - 038 ug/L).
Results for juvenile Cyprinodon variegatus exposed to endosulfan indicated mortalities
ranging from 0 - 0%, averaging 0% (± 0%) at 0.01 ug/L; from 0 - 20%, averaging 5% (±
5%) at 1.15 ug/L; and from 20 - 100%, averaging 70% (± 19.15%) at 2.50 ug/L. Control
mortality ranged from 0 -0%, averaging 0% (± 0%). The 96h LCjo for endosulfan in
juvenile Cyprinodon variegatus was 1.97 ug/L (CL = 1.66 - 2.34 ug/L) which was much
higher than the QA/QC results for 1989 (96h LCM = 0.31 ug/L with CL = 0.13 - 0.75
ug/L). The statistically higher LCjo obtained during 1990 were the result of the higher
exposure temperatures during 1990 (inverse temperature coefficient previously discussed in
F. heteroditus) and the fact that U.S. EPA laboratory stocks of juvenile sheepshead minnow
during 1990 were generally larger size juveniles (late stage) compared to 1989 stocks (early
stage). The larger size juveniles wiiuld be more resistant than earlier staged juveniles.
Similar results have been reported with adult and juvenile F. heteroditus exposed to
endosulfan (Scon et al, 1990).
145
-------
Results for juvenile Menidia berylina exposed to endosulfan indicated mortalities ranging
fromO - 20%, averaging 13.3% (± 6.67%) at 0.01 ug/L; from 100- 100%, averaging 100%
(± 0%)at 1.15 ug/L; and from 100- 100%, averaging 100% (+ 0%) at 2.50ug/L). Control
mortality ranged~0 - 25%, averaging 15% (± 7.64%). The high control mortality was the
result of handling stress in some instances, as this was the first time Menidia berylina had
been used in toxicity tests. The 96h LCSO for endosulfan in juvenile Menidia berylina was
0.07 ug/L (CL = not calculated).
These results generally indicated that each group of organisms used in each field deployment
were comparable. Although some variations were observed between groups much of these
differences were related to extrinsic (exposure temperature) and intrinsic (different life history'
stage) factors encountered during the bioassay, which would account for these differences.
II. OYSTER ECOPHYSIOLOGY STUDIES, 1989-90
A. 1939 Studies
Results of oyster ecophysiology studies conducted at the CTL and TRT Sites during 1989 are
listed in Tables 32-38 and Figures 32-38.
During May, 1989 at the mouth of Leadenwah Creek, where oysters were initially collected
prior to transplantation, salinities ranged from 24 - 27 ppt, seawater temperatures ranged from
21.5 - 27°C, and dissolved oxygen from 6.45 - 7.05 mg/L (Table 32). During June, 1989 at the
TRT Site salinities ranged from 20 - 25 ppt, water temperatures from 29.1 - 30.4°C, and
dissolved oxygen levels from 6.00 - 7.20 mg/L. Physicochemical parameters were quite similar
at the CTL Site, with salinities ranging from 21.0 - 28.5 ppt, seawater temperatures from 29.0-
29.7°C, and dissolved oxygen levels from 6X30 - 6.70 mg/L. During July, 1989 salinities ranged
from 21.5 - 29.0ppt, seawater temperatures from 30.3-32.4°C, and dissolved oxygen levels from
5.80 - 7.00 mg/L at the CTL Site versus salinities ranging from 16.0 - 25.0 ppt, seawater
temperatures from 27.9 - 33.6°C, and dissolved oxygen levels from 5.90 - 6.80 mg/L at the TRT
Site. The lower salinities at the TRT Site during July were the result of heavy rainfall and runoff,
which as a result, lowered salinities.
146
-------
Table 32. Summary of physicochemical water quality parameters measured in oyster
studies at the CTL, TRT and KWA Sites during 1989 - 90. Note the
lower salinities at the TRT Site during June 1989 following periods of significant
rainfall.
Date
17-18 May, 1989
12-13 June, 1989
14-15 June, 1989
10-11 July, 1989
12-13 July, 1989
16-17 May, 1990
13 June, 1990
14 June, 1990
6 July, 1990
7 July, 1990
Site
Mouth of
Leadenwah Creek
CTL
TRT
CTL
TRT
Mouth of
Leadtnwah Creek
CTL
KWA
CTL
KWA
Salinity
(Ppt)
24.0 - 27.0
21.0- 28.5
20.0 - 25.0
21.5 - 29.0
16.0 = 25.0
28.0 - 30.0
31.0-31.0
35.0 - 36.0
34.0-35.0
35.0-36.0
Dissolved
Oxygen
(mg/L)
6.74 - 7.05
6.30 - 6.70
6.00 - 7.20
5.80-7.00
5.90 - 6.80
r 6.28 - 7.70
6.40 - 7.60
6.10-6.75
5.20-5.95
5.40- 5.85
Water
Temperature
(°Q
21.5 -27.0
29.0 - 29.7
29.1 - 30.4
30.3-32.4
27.9 - 33.6
26.5- 29.4
23.0-25.7
23.2-26.9
30.2 - 30.9
26.7 - 31 0
147
-------
During 1989, significant runoff of fenvalerate was observed at the TRT Site on 6/5 - 6/89 (0.065 -
0.093 ug/L fenvalerate), 6/9/89 (0.021-0.022 ug/'L fenvalerate) and on 6/15/89 (0.015 ug/L fenvalerate)
and possible exposure to oysters may have occurred.
Results of condition index measured in oysters from both sites (Table 33; Figure 32) indicated there
were no significant differences in between site comparisons during June - July, 1989. The initial condition
index in May prior to transplantation, was 93.10(± 2.57). Mean condition indices measured during June-
July ranged from 64.60 (± 3.48) in June to 71.49 (± 4.48) in July at the TRT Site and ranged from
84.38 (±4.31) in June to 72.72 (± 3.53) in July at the CTL Site. These data indicate that immediately
following transplantation condition indices at both sites declined nearly 30%. This was probably the result
of spawning activities in oysters as noted earlier by Scott (1979) and Scott et al. (1990).
Results of Perkinsus marinus infection intensities analysis (Table 34; Figure 33) indicated low-
moderate infection intensities of this oyster parasite at both sites. During June, infection intensities ranged
from 2.08 (±0.34) at the TRT Site to 2.33 (±0.28) at the CTL Site. During July, infection intensities
increased in oysters at both sites, ranging from 3.50 (±0.20) at the TRT Site to 3.17 (±0.39) at the CTL
Site. Statistical analysis indicated that there were no significant between site differences observed. These
results generally agree with earlier studies by Scott et al, (1990).
148
-------
Table 33. Summary of condition indices measured in oysters at the CTL, TRT, and
KWA sites 1989-90. Statistical analysis indicated no significant
differences in temporal comparisons between paired sites.
Parameter
Condition Index
Date
May, 1989
June, 1989
July 1989
May, 1990
June, 1990
July, 1990 f
Site
CTL
CTL
TRT
CTL
TRT
CTL '
CTL
KWA
CTL
KWA
X
93.10
84.38
64.60
72.72
71.49
61.11
84.49
79.15
79.77
69.35
SE
2.57
4.31
3.48
3.53
4.48
3.91
8.57
9.68
6.62
4.12
149
-------
1989 CONDITION INDICES
100
ONDITION
INDEX
90 -
80 -
70 -
60 H
50
WAY
JUNE
JULY
MONTH
Figure 32. Condition index in oysters deployed at the CTL and TRT Sites during the 1989
field study. Note the similarities in condition indices for oysters from both sites.
150
-------
Table 34. Summary of Perkinsus marinus infection intensities measured at the CTL,
TRT and KWA Sites during 1989-90. Statistical analysis indicated no
significant differences in temporally paired comparisons between the CTL
and TRT Sites (1989) and the CTL and KWA Sites (1990).
Parameter
Perkinsus marinus
Infection Intensity
Date
June, 1989
July, 1989
May, 1990
June, 1990
July, 1991 _
Site
CTL
TRT
CTL
TRT
CTL
CTL
KWA
CTL
KWA
X
2.33
2.08
3.17
3.50
2.17
3.33
2.94
2.77
3.77
SE
0.28
0.34
0.39
0.20
0.27
0.33
0.29
0.40
0.51
151
-------
1909 PERKINSUS INFECTION INTENSITY
INFECTION
INTENSITY
MONTH
Figure 33. Perkinsus marinus infection intensities in oysters from the CTL and TRT Sites
during the 1989 field study. Note the similarities in infection intensities for oysters
from both sites.
152
-------
Results of in situ, whole animal respiration rate determinations are listed in Table 35.
Initially, a mean respiration rate (23°C) of 1.020 ml/0.685 g/h (±0.070) was measured in
oysters during May, prior to transplantation. Following transplantation, June respiration rates
(25°C) ranged from 1.600 ml/0.685 g/h (±0.080) at the TRT Site to 1.290 ml/0.685 g/h
(±0.100) at the CTL Site. During July, (30°C) respiration rates ranged from 2.470 ml/0.685
g/h (±0.100) at the TRT Site to 2.250 ml/0.685 g/h (±0.110) at the CTL Site. Statistical
analysis indicated there were significant (p < 0.002) between site differences in respiration
rates observed during June following periods of low salinity associated with fenvalerate runoff
at the TRT Site. An earlier study by Scott (1979) reported similar gill and mantle respiration
rates as were observed in whole animals at the CTL Site during May - June, in oysters from
Leadenwah Creek.
Respiration rates in oysters from both sites increased during May - July, primarily due to
increased ambient seawater temperatures (May-21.5 - 27.5°C to July- 27.9 -33.6°C). Oysters
are poikilothermic organisms, whose metabolic rates conform directly with ambient
temperatures. As a result, respiration rates in oysters from each site varied temporally due to
changes in exposure temperature. To compensate for this effect, respiration rates for each
sampling period were Q10 adjusted at 23°C (May), 25°C (June), and 30°C (July), so that
temporal comparisons between groups, could be made (Table 35; Figure 34). Q10 respiration
adjustments allow the physiological effects of pesticide runoff to be better elucidated by
normalizing the effects of temperature on respiration, so that pesticide effects can be discerned.
At 23°C, the initial Q1Q adjusted respiration rate for oysters during May, prior to
transplantation was 1.020 ml/6.685 g/h (±0.070). During June, Q10 adjusted respiration rates
ranged from 1.250 ml/0.685 g/h (±0.060) at the TRT Site to 0.950 ml/0.685 g/h (±0.080)
at the CTL Site.
153
-------
Table 35. Summary of Q10 Adjusted Respiration Rates (m/02/0.685g/h) in ovsters
deployed at the CTL, TRT and KWA Sites during 1989-90. Statistical analysis
indicated significant (*) differences (p < 0.05) in paired comparisons between
the CTL and TRT Sites (1989) and the CTL and KWA Sites (1990) as
indicated in the table.
Parameter
Respiration
Q1Q Adjusted
(ml/02/0.685g/h
)
Temp
(°C)
23'
25
30
23
251
30
23
25'
30
23
25
30'
23
25
30l
23'
25
30
23
25'
30
23
25'
30
23
25
30'
23
25
30'
Date
May, 1989
June, 1989
July, 1989
May, 1990 '
June, 1990
July, 1990
Site
CTL
CTL
TRT
CTL
TRT
CTL
CTL
KWA
CTL
KWA
X
1.020
1.220
1.720
0.950
1.290
2.140
1.250 *(0.005)
1.600 *(0.002)
2.460
1. 000
1.350
2.250
1.260*(0.001)
1.600 *(0.001)
2.470
1.080
1.290
1.820
1.070
1.260
1.750
0.900 *(0.01)
1.180
1.880
1.120
1.320
1.840
1.000
1.310
2.100
SE
0.070
0.090
0.120
0.080
0.100
0.170
0.060
0.080
0.130
0.050
0.060
0.110
0.050
0.060
0.100
0.060
0.070
0.100
0.020
0.030
0.040
0.040
0.050
0.080
0.080
0.090
0.130
0.040
0.050
0.080
* = Significantly different in paired comparisons between CTL and TRT (1990) and
CTL and KWA (1990) at each respective date and temperature. Values in
parentheses ( ) are P values.
1 = Actual time periods and temperatures respiration measurements were taken.
Values for other time periods and temperatures combinations are Q10 derived
values.
154
-------
Q-10 CORRECTED RESPIRATION TO 23 DEGREES (1989)
o
E
I 0
i G
1 J
1 2
1.0
0.8
06
MAY
JUNE
MONTH
JULY
Q-10 CORRECTED RESPIRATION FOR 25 DEGREES (1989)
2.4 -
2.2-
1.8-
1.6 -
1.4 -
1 2-
1 0-
08-
*
_. — —""" I
^ " t
MAY JUNE
MONTH
•
*
i
3
JULY
Q-10 CORRECTED RESPIRATION FOR 30 DEGREES (1989)
.c
Cl
I/1)
o
us
o
CM
0
i
iL.a -
2G-
24 -
o o _
C.i
20-
8 -
6-
4 -
.2-
.0-
0.8-
— •*— CTL T ; i
...-1 t
0--- TRT .--• J -J|
.---^— -^-^^T
j^ — "~"
I
MAY JUNE JULY
I MONTH
Figure 34 Q10 adjusted respiration rates (ml/0.685g/h) for oysters at three exposure
temperatures (23°, 25°, and 30° C) during the 1989 field study. Note the
increased respiration rates in TRT Site oysters at 23° and 25°C.
155
-------
During July. Q;a adjusted respiration rates ranged from 1.260 ml/0.685 g/h (± 0.050) at the
TRTSiteto 1.000 ml/0.685 g/h (± 0.050) ac che CTL Site. Statistical analysis clearly indicated
that there were significant (p < 0.001-0.005) between site differences in QIO adjusted
respiration rates during June and July at 23°C. At 23°C, respiration rates in oysters at the
TRT Site were much higher than at the CTL Site. Also note that Q,0 adjusted respiration rates
at the CTL Site were virtually unchanged from May through July.
At 25°C, the initial Q,0 adjusted respiration rate was 1.220 ml/0.685 g/h (± 0.090) during
May, prior to transplantation. During June. Q10 adjusted respiration rates at 25 °C, ranged
from 1.600 ml/0.685 g/h (± 0.080) at the TRT Site-to 1.290 ml/0.685 g/h (± 0.100) at the
CTL Site. During July, QIO adjusted respiration rates at 25°C, ranged from 1.600 ml/0,685
g/h (± 0.060) at the TRT Site to 1.35 ml/0.685 g/h (± 0.060) at the CTL Site. Statistical
analysis indicated that there were significant (p < 0.001-0.002) between site differences in Q10
adjusted respiration rates observed during June and July at 25°C. Respiration rates were higher
in TRT Site oysters while remaining relatively constant at the CTL Site.
At 30°C, the initial Q10 adjusted respiration rate for oysters collected in May, prior to
transplantation was 1.720 ml/0.685 g/h (± 0.120). During June, Q10 adjusted respiration rates
at 30°C, ranged from 2.460 ml/0.685 g/h (± O.'ISO) at the TRT Site to 2.140 ml/0.685 g/h
(± 0.170) at the CTL Site. During July, Q10 adjusted respiration rates at 30°C, ranged from
2.470 ml/0.685 g/h (± 0.100) at the TRT Site to 2.250 ml/0.685 g/h (± 0.110) at the CTL
Site. Statistical analysis indicated that there were no significant between site differences
observed in Q,0 adjusted respiration rates at 30°C during June and July.
Figure 35 depicts the mean Q10 standardized respiration rates for oysters at the CTL and
TRT Sites from May - July for the 23-30°C temperature range observed. Note the significantly
(p < 0.0001-0.0365) higher respiration rates at all temperatures (23-30°C), in oysters from
the TRT Site when compared'to CTL Site otganisms. These results suggest that metabolic
rates in oysters from the TRT Site were significantly higher at temperatures ranging from 23-
30°C. This would be at the upper limits of their zone of compatibility or capacity adaptations
for temperature exposure. These difference^ in respiration rates may, in part, be the result of
fenvalerate exposure in oysters at the TRT Site, although other factors such as low salinity
must also be considered. Low salinity (< 5ppt) and resulting reduced paniculate levels
(including phytoplankton) may also be significant factors, which may adversely affect oysters.
For example, low salinity (< 5ppt) exposure per se, would cause oysters to close their valves
156
-------
MEAN Q-10 STANDARIZED RESPIRATION (1989)
3 0
2.5-
2.0-
1.5-
1.0-
0.5 -
0.0
22
28
30
TEMP
Flgure 35. Mean Q10 standardized respiration rates (ml/0.685g/h) measured during the 1989
field study. Note the significantly increased respiration rates at all
temperatures tested (23-30°C) for TRT Site oysters.
157
-------
and utilize reverse glycolysis to maintain metabolic activity but with resulting increased
respiration rates due to the oxygen debt incurred. Respiration rates during hypoxia would vary
directly with temperature (Scott, 1979); however, respiration rates during low salinity
exposures ranging from 7-10 ppt, would not affect gill respiration rates in oysters (Scott et at.,
1985). It is extremely difficult to differentiate effects from low salinity and fenvalerate
exposure in the field, since both factors may co-occur during runoff events. Results from this
study suggest that exposure to agricultural nonpoint source insecticide runoff and resulting low
salinity conditions increased the cost of maintenance metabolism in oysters adjacent to
agricultural sites.
Results of nitrogen excretion rate measurements are listed in Table 36 and Figure 36.
During May, the initial nitrogen excretion rate was 2.30 ug atoms N/g/h (± 0.29 ug atoms
N/g/h). During June, nitrogen excretion rates increased at the TRT Site to 9.70 ug atoms
N/g/h (± 1.54 ug atoms N/g/h) compared to only 3.84 ug atoms N/g/h (± 0.60 ug atoms
N/g/h) in CTL Site oysters. During July, nitrogen excretion rates decreased at the TRT Site
to 4.44 ug atoms N/g/h (± 0.33 ug atoms N/g/h) which was comparable to levels of 3.00 ug
atoms N/g/h (± 0.74 ug atoms N/g/h) at the CTL Site. Statistical analysis indicated that
nitrogen excretion rates during June (peak of the tomato growing season and four days post
fenvalerate exposure) were significantly (p < 0.01) higher in TRT Site oysters. On 5, 6 and
9 June, significant (> 1.25 cm/day) rainfall occurred at the TRT Site, which resulted in
substantial runoff of fenvalerate (0.065-0.093 ug/L on 6/5 - 6/89 and 0.021 - 0.022 ug/L on
6/9/89) and concomitant periods of extended low salinity (< 10 ppt) exposure. Oysters exposed
to low salinity will catabolize amino acids (ninhydrin positive - glycine, alanine and taurine)
in order to osmoregulate. An earlier study by Scott et al. (1985) reported that low salinity does
not significantly affect respiration rate; however, other studies have shown that low salinity
may significantly increase nitrogen excretion rates in mussels (Widdows et al., 1981) and fish
(Scott et al., 1987). Fenvalerate exposure caused significantly increased respiration rates in
juvenile crustaceans which was enhanced by low salinity osmotic stress (Mckenny and
Hamaker, 1984). Exposure of fish to fenvalerate appeared to have no effect on nitrogen
excretion rates (Scott et al., 1990). Results from this study indicated significantly increased
nitrogen excretion rates in oysters at the\TRT Site exposed to low salinity, fenvalerate
agricultural runoff during June. By July, nitrogen excretion rates in oysters at the TRT Site
decreased to levels comparable to CTL Site oysters.
158
-------
Table 36. Summary of Nitrogen Excretion Rates (ug atoms N/g/h) in oysters deployed
at the CTL, TRT and KWA Sites during 1989 - 90. Asterisks (*) indicate
where paired comparisons were significantly (p ^ 0.05) different.
Parameter
Nitrogen Excretion
(ug atoms N/g/h)
Date
May, 1989
June, 1989
July, 1989
May, 1990
June, 1990
July, 1990
Site
CTL
CTL
TRT
CTL
TRT
CTL
CTL
KWA
CTL
KWA
X
2.30
3.84
9.70 *(P <; o.oi)
3.00
4.44
10.06
11.70
9.34
5.36
9.51
SE
0.29
0.60
1.54
0.76
0.33
1.46
1.63
1.17
0.97
2.87
* = Signifcantly different from paired control value.
Values in () were p values.
159
-------
MEAN MONTHLY EXCRETION RATES (1989)
16
01
z
14 -
12 -
10 -
MAY
JUNE
JULY
MONTH
Figure 36. Ammonia nitrogen excretion ratfcs (ug atoms N/g/h) in oysters deployed during the
1989 field study. Note the increased nitrogen excretion rate in TRT Site oysters
during June, following periods of significant fenvalerate-low salinity runoff
conditions.
160
-------
Results of O/N Ratios are listed in Table 37 and Figure 37. The initial 0/N ratio
measured in May was 65.04 ( + 9.66). During June, O/N ratios ranged from 50.61 (± 10.12)
at the TRT Site to 74.74 (± 11.38) at the CTL Site. During July, O/N ratios ranged from
84.39 (± 10.99) at the TRT Site to 151.91 (± 25.96) at the CTL Site. Statistical analysis
indicated there were significant (p < 0.009) differences in between site comparisons during
July as higher O/N ratios were measured at the CTL Site.
Results of Q10 adjusted O/N ratios are listed in Tables 38 and Figure 38. The initial Q10
adjusted O/N ratio in May was 74.96 (± 10.13). Daring June, Q10 adjusted O/N ratios ranged
from 31.43 (± 6.48) to 46.79 (± 6.83) at the CTL Site. During July, Q10 adjusted O/N ratios
ranged from 50.54 (± 5.22) at the TRT Site to 82.20 (± 13.91) at the CTL Site. Statistical
analysis indicated there were no significant differences in between site comparisons of Q10
adjusted O/N ratios during June. During July, Q10 adjusted O/N ratios were significantly (p
< 0.03) lower at the TRT Site.
O/N ratios and Q1Q adjusted O/N ratios were generally above 40 throughout the 1989 study,
suggesting that oysters were healthy with a dominance of lipid and carbohydrate metabolism,
with minor protein catabolism (NAS, 1980). Trie generally lower values measured at the TRT
Site were indicative of higher nitrogen production by oysters there. This higher nitrogen out
put by oysters at the TRT Site was in all likelihood, a metabolic adaptation (i.e.
osmoregulation) to lower salinity conditions there, resulting from agricultural runoff. To
compensate oysters catabolize protein to maintain homeosmocity; as a result increased nitrogen
excretion will occur, with resulting decreased O/N ratios. Fenvalerate exposure must also be
considered since fenvalerate may inhibit ATPase. This enzyme is important in the maintenance
of osmotic balance (i.e. Na*K*Mg** ATPase).
161
-------
Table 37. Summary of O/N Ratios measured in oysters deployed at the CTL, TRT
and KWA Sites during 1989 - 90. Asterisks (*) indicate where paired
statistical comparisons were significantly (p < 0.05) different.
Parameter
O/N Ratio
Date
May, 1989
June, 1989
July, 1989
May, 1990
June, 1990
July, 1990
Site
CTL
CTL
TRT
CTL
TRT
CTL
CTL
KWA
CTL
KWA/
X
65.04
74.72
50.61
151.91
84.39 *
-------
MEAN MONTHLY O/N RATIOS (1989)
O/N
ISO
•so -
uo -
120 -
100 -
80 -
60 -
40 -
20 -
0
MAY
JUNE
JULY
MONTH
Figure 37. Mean O/N Ratios in oysters deployed during the 1989 field study. Note how O/N
ratios were significantly low^r at the TRT Site during July.
163
-------
Table 38. Summary of Q10 Corrected O/N Ratios in oysters deployed at the CTL,
TRT and KWA Sites during 1989 - 90. Asterisks (*) indicate where
paired statistical comparisons were significantly (P < 0.05) different.
Parameter
Q/10 Corrected
O/N Ratios
Date
May, 1989
June, 1989
July, 1989
May, 1990
June, 1990
July, 1990
Site
CTL
CTL
TRT
CTL
TRT
CTL
CTL
KWA
CTL
KWA
X
74.96
46.79
31.43
82.20
50.54 *(p <; 0.03)
23.40
16.35
23.28
40.91
40.39
SE •
10.13
6.83
6.48
13.91
5.22
5.85
1.90
5.56
5.51
12.57
* = Significantly different in comparison with paired control value.
Values in ( ) were p values.
164
-------
MEAN MONTHLY Q-10 CORRECTED O/N (1989)
O/N
ISO
160 -
MO -
120 -
100 -
80 -
60 -
40 -
20
0
MAY
JUNE
JULY
MONTH
Figure 38. Mean Q10 adjusted O/N Ratios in oysters deployed during the 1989 field study.
Note how O/N ratios were significantly lower in July at the TRT Site.
165
-------
B. 1990 Results
Results of oyster ecophysiology studies conducted at the CTL and KWA Sites during L990
are listed in Tables 32-38 and Figures 39-45.
During May, 1990 at the mouth of Leadenwah Creek, where oysters were initially
collected prior to transplantation at the CTL and KWA Sites, salinities ranged from 28-30 ppt,
seawater temperatures from 26.5-29.4°C, and dissolved oxygen from 6.28-7.70 mg/L (Table
32). During June, 1989 at the CTL Site, salinities ranged from 31.0-31.0 ppt, water
temperatures from 23.0-25.7°C, and dissolved oxygen from 6.40-7.60 mg/L. At the KWA
Site, salinities (35.0-36.0ppt), water temperatures {23.2-26.9°C), and dissolved oxygen levels
(6.40-7.60 mg/L) were quite similar to the CTL Site during June, 1990. During July, 1990
salinities ranged from 35.0-36.0 ppt, water temperatures from 26.7-31.0°C, and dissolved
oxygen levels from 5.40-5.85 mg/L at the KWA Site versus salinities ranging from 34.0-35.0
ppt, water temperatures from 30.2-30.9°C, and dissolved oxygen levels from 5.20-5.95 mg/L
at the CTL Site. Generally physicochemical parameters were similar at both sites during June
and July, 1990. The small amount of rainfall during this study resulted in very little
agricultural runoff. As a result salinities remained high throughout the study.
During 1990, only slight runoff (concentrations < 96h LCjo values for most sensitive
estuarine species) of azinphosmethyl was observed at the KWA Site on 6/15/90 (0.024-0.062
Atg/L). Oysters at the KWA Site may have been potentially exposed to azinphosmethyl as a
result. No significant levels of pesticides were observed at the CTL Site during 1990.
Results of condition index measured in oysters from both sites (Table 33; Figure 39)
indicated there were no significant differences in condition index in between site comparisons
during June and July, 1990. The initial condition index in May 1990 prior to transplantation
averaged 61.11 (± 3.91). Condition indices measured in June and July ranged from 69.35
(± 4.12) - 79.15 (± 9.68) at the KWA Site and from 79.77 (± 6.62) - 84.49 (± 8.57) at
the CTL Site. These data indicated that immediately following transplantation, condition
indices at both sites increased nearly 30%'- This was the result of glycogen accumulation in
oysters at these sites prior to spawning. Earlier studies by Scon (1979), Scott et al., (1990)
and 1989 results reported in this study indicated a slightly earlier period (May) of glycogen
accumulation and spawn at Leadenwah "Creek. Results for 1990, indicated that glycogen
accumulation were delayed until June wheh resulting spawning activity occurred in July at both
sites.
.166
-------
1990 CONDITION INDICES
!00
CONDITION
INDEX
90 -
80 -
70
60
50
MAY
JUNE
JULY
MONTH
Figure 39. Condition index in oysters deployed at the CTL and KWA Sites during the 1990 field
study. Note the similarities in coition indices in oysters at both sites.
167
-------
Results of Perkinsus marinus intensity analysis (Table 34; Figure 40) indicated low -
moderate infection intensities of this oyster parasite at both sites. The initial infection intensity
was 2.17 (± 0.27) in oysters collected from Leadenwah Creek during May, prior to
transplantation. Bnnng June, infection intensities ranged from'2.94 (+ 0.29) at the KWA Site
to 3.33 (± 0.33) at the CTL Site. During July infection intensity increased, in oysters at the
TRT Site with intensities averaging 3.77 (+ 0.51). At the CTL Site, intensities decreased
slightly, averaging 2.77 (± 0.40). Statistical analyses indicated that there were no significant
between site differences observed. These results generally agreed with earlier studies by Scott
et al., (1990) and results for 1989 in this study.
Results of in situ whole animal respiration rate determinations are listed in Table 35 and
Figure 41. Initially, respiration rates averaged 1.080 ml/0.685 g/h (± 0.060 ml/0.685 g/h)
in oysters collected during May, 1989 prior to transplantation. During June respiration rates
ranged from 1.180 ml/0.685 g/h (± 0.050 ml/0.684 g/h) at the KWA Site to 1.260 ml/0.685
g/h (± 0.030 ml /0.685 g/h) at the CTL Site. During July, respiration rates ranged from
2.100 ml/0.685 g/h (± 0.080 ml/0.685 g/h) at the KWA Site to 1.840 ml/0.685 g/hr (±
0.130 ml/0.685 g/h) at the CTL Site. Statistical analysis indicated there were no significant
between site differences in oyster respiration rates during June and July. An earlier study by
Scott (1979) indicated similar gill and mantle tissue respiration rates during May-July in
oysters at Leadenwah Creek.
Respiration rates in oysters from both sites increased from May-July, primarily due to
increased ambient seawater temperatures (May - 26.5 - 29.4°C to 26.7 - 31.0°C during July).
Oysters are poikilothermic organisms, whose metabolic rates will conform directly with
ambient temperatures as a result. Respiration rates in oysters from each site varied
temporally, due to changes in exposure temperature. To compensate for this effect, the Q-lOs
for each respiration determinations from each sampling were calculated at 23°C (May), 25°C
(June) and 30°C (July) so that temporal comparisons between groups could be made (Table
35; Figure 41). At 23°C, the initial Q10 adjusted respirations for oysters during May prior
to transplantation was 1.080 ml/0.685 g/h (± 0.060 ml/0.685 g/h). During June, Q10 adjusted
respiration rates at 23°C ranged from 0.900 ml/0.685 g/h (± 0.040 ml/0.685 g/h) at the
KWA Site to 1.070 ml/0.685 g/h (± 0.020 ml/0.685 g/h) at the CTL Site. Statistical analysis
indicated significant (p <> 0.01) between site differences, as respiration rates were decreased
during June at the KWA Site. During Juljj, Q10 adjusted respiration rates at 23°C ranged from
1.000 ml/0.685 g/H (± 0.040 ml/0.685 g/h) at the KWA Site to 1.120 ml/0.685 g/h (± 0.080
ml/0.685 g/h) at the CTL Site. Statistical analysis indicated, there were no significant
between site differences observed in Q10 adjusted respiration rates during July at 23°C.
168
-------
1990 PERKINSUS INFECTION INTENSITY
INFECTION
INTENSITY
MONTH
Figure 40. Perkinsus marinus infection intensities in oysters deployed during the 1990 field study.
Note the similarities in infection intensities for oysters at the CTL and KWA Sites.
169
-------
Q-10 CORRECTED RESPIRATION FOR 23 DEGREES (1990)
a —
CTL
2.6
24 -
2.2 -
2.0-
1 8
1.6 -
1 4 -
1 2 -
1.0 -
08 •
0.6
MAY
JUNE
MONTH
JULY
Q-10 CORRECTED RESPIRATION FOR 25 DEGREES (1990)
a —
B
CTL
MAY
JUNE
MOflTH
JULY
26
24 -
2.2-
20 -
1.8 -
1 6 -
I 2 -
1 0 -
08 -
06
Q-10 CORRECTED RESPIRATION FOR 30 DEGREES (1990)
CTL
: j -
MAY
JUNE
MONTH
JULY
Figure 41. Q10 adjusted respiration rates (m|/0.685g/h) in oysters deployed during the 1990 field
study. Generally there were no significant differences in respiration rates observed
between oysters deployed at both sites for all temperatures tested (23-30°C).
170
-------
At 25°C, the initial Q10 adjusted respiration rate during May, prior to transplantation
was 1.290 ml/0.685 g/h (± 0.070 ml/0.685 g/h). During June, QIO adjusted respiration
rates at 25°C ranged from 1.180 ml/0.685 g/h (± 0.050 ml/0.685 g/h) at the KWA Site
to 1.260 ml/0.685 g/h (± 0.030 ml/0.685 g/h) at the CTL Site. During July, Q,0 adjusted
respiration rateTat 25°C ranged from 1.310 ml/0.685 g/h (± 0.050 ml/0.685 g/h) at the
KWA Site to 1.320 ml/0.685 g/h (± 0.090 ml/0.685 g/h) at the CTL Site. Statistical
analysis indicated there were no significant between site differences observed during June
and July at 25°C.
At 30°C the initial Q10 adjusted respiration rate in May for oysters prior to
transplantation, was 1.820 ml/0.685 g/h (± 0.100 ml/0.685 g/h). During June, Q10
adjusted respiration rates at 30°C ranged from 1.880 ml/0.685 g/h (± 0.080 ml/0.685 g/h)
at the KWA Site to 1.750 ml/ 0.685 g/h (± 0.040 ml/0.685 g/h) at the CTL Site. During
July, Q10 adjusted respiration rates at 30°C ranged from 2.100 ml/0.685 g/h (± 0.080
ml/0.685 g/h) at the KWA Site to 1.840 ml/0.685 g/h (± 0.130 ml/0.685 g/h) at the CTL
Sites. Statistical analysis indicated there were no significant between site differences
observed in Q10 adjusted respiration rate comparisons during June and July at 30°C.
Figure 42 depicts the mean Q10 Standardized respiration rates for oysters at the CTL
and KWA Sites from May-June, for the 23-30°C temperature ranged observed. At 23 "C,
there were no significant (p ^ 0.61) differences in Q10 standardized respiration rates for
comparisons between the KWA (X = 0.900 "ml/0.685 g/h ± 0.040 ml/0.685 g/h) and CTL
(X = 1.070 ml 0.685 g/h ± 0.020 ml/0.685 g/h) Sites. At 258C, there were no
significant differences in Q10 standardized respiration rates in comparisons of oysters at
each site (X = 1.180 ml/0.685 g/h ± 0.050 ml/0.685 g/h at the KWA Site versus X =
1.260 ml/0.685 g/h ± 0.030 ml/0.685 g/h at the CTL Site). Also at 30°C, there were no
significant differences in Q10 standardized respiration rates in comparisons of oysters at the
KWA (X = 1.880 ml/0.685 g/h I 0.080 mi;0.685 g/h ± 0.080 ml/0.685 g/h) and CTL
(X = 1.750 ml/0.685 g/h) Sites. These results suggest that metabolic rates in oysters from
both sites were similar at all test temperatures (23-30°C), which would be at upper thermal
limits of their zone of capacity adaptations for temperature exposure. The lack of
significant differences in respiration rates in oysters at the CTL and KWA Sites during
1990 was not surprising given the small amounts of rainfall, resulting similarities in the
physicochemical environmental at both sites, and the resulting low levels of insecticide
exposure (azinphosmethyl) observed during 1990. These results are in sharp contrast in
results for 1989, when marked differences in respiration rates were observed in TRT Site
Oysters following significant fenvalerate and low salinity exposure. During 1990 low
salinities (< 5 ppt) were not observed at the KWA Site due to the small amount of'
rainfall.
171
-------
MEAN Q-10 STANDARDIZED RESPIRATION (1990)
30
25-
2.0-
1.5-
1.0 -
05-
0 0
22
26
28
30
TEMP
Figure 42. Mean Q10 standardized respiration rates (ml/0.685g/h) in oysters deployed during
the 1990 field study. Note the similarities in respiration rates in oysters from each
site at all temperatures tested (23-30°C).
172
-------
Results of nitrogen excretion rates are listed in Table 36 and Figure 43. During May.
the initial nitrogen excretion rate was 10.06 ug atoms N/h/h (± 1.46 ug atoms N/g/h).
During June, nitrogen excretion rates ranged from 9.34 ug atom N/g/h (± 1.17 ug atom
N/g/h) at the KWA Site to 11.70 ug atom N/g/h (± 1.63 ug atoms N/g/h) at the CTL Site.
During July, nitrogen excretion rates ranged from 9.51 ug atoms N/g/h (± 2.87 ug atoms
N/g/h) at the KWA Site to 5.36 ug atom N/g/h (± 0.97 ug atoms N/g/h) at the CTL Site.
Statistical analysis indicated nitrogen excretion rates during June and July were not
significantly different in between site comparisons.
Results of 0/N ratios are listed in Table 37 and Figure 44. The initial 0/N ratio
measured in May was 29.16 (± 7.13). During June O/N ratios ranged from 26.86 (±
6.70) at the KWA Site to 16.78 (± 1.79) at the CTL Site. During July, O/N ratios ranged
from 65.12 (± 17.48) at the KWA Site to 63.11 (± 7.82) at the CTL Site. Statistical
analysis indicated these were no significant differences in between site comparisons during
June and July.
Results of Q,0 adjusted O/N ratios are listed in Table 38 and Figure 45. The initial
Q10 adjusted O/N ratio in May was 23.40 (± 5.85). During June, Q10 adjusted O/N ratios
ranged from 23.28 (± 5.56) at the KWA Site to 16.35 (± 1.90) at the CTL Site. During
July, Q10 adjusted 0/N ratios ranged from 40.39 (± 12.57) at the KWA Site to 40.91 (±
5.51) at the CTL Site. Statistical analysis indicated these were no significant differences
in between site comparisons of Q,0 adjusted O/N ratios.
The lower O/N ratio values measured during May and June (<30) are indicative of
healthy oysters, but signify some possible protein catabolism. During July the higher O/N
ratios (> 40) measured were indicative of a predominance of lipid and carbohydrate
metabolism rather than significant protein catabolism. As noted by the changes in
condition index, oysters were accumulating gametes for spawning from May - July. As
this maturation process occurred, O/N ratios increased as nitrogen output and protein
catabolism decreased.
173
-------
MEAN MONTHLY EXCRETION RATES (1990)
18
16 -
12 -
10
8
6
4
2
0
MAY
JUNE
JULY
MONTH
Figure 43. Ammonia nitrogen excretion rates (ug atoms N/g/h) in oysters deployed during the
1990 field study. There were no differences in nitrogen excretion rates observed
in comparison of CTL and KWA Site oysters.
174
-------
MEAN MONTHLY O/N RATIOS (1990)
O/N
160
140
120
100
so
60
40
20
0
JUNE
JULY
MONTH
Figure 44. Mean O/N Ratios in oysters deployed during the 1990 field study. Note the
similarities in O/N ratios in oysters deployed at the CTL and KWA Sites during
1990.
175
-------
MEAN MONTHLY Q-10 CORRECTED O/N (1990)
O'N
130 -
160 -
140 -
120 -
100 -
80 -
60 -
40 -
20 -
0 -
MAY
JUNE
JULY
MONTH
I
Figure 45. Mean Q10 adjusted O/N Ratios in oysters deployed during the 1990 field study.
Note the similarities in O/N ratios in oysters deployed at the CTL and KWA Sites
during 1990.
176
-------
C. Discussion and Conclusions: Oyster Ecophysiology Studies 1989-90.
Results of ecophysiology studies clearly indicated the utility and sensitivity for an
integrated battery of physiological parameters to assess agricultural NFS pesticide runoff
effects in oysters. During 1989, significant runoff of fenvalerate may have, in part, caused
significant sublethal stress as measured by Q10 adjusted respiration rates and nitrogen
excretion rates in oysters at the TRT Site. Alterations of these physiological parameters
did not cause resulting effects in oyster condition index or parasite infection intensity.
These results suggest that while significant fenvalerate exposure occurred, with measurable
physiological differences in respiration and nitrogen excretion rates resulting, no gross
changes in body component indices (i.e., condition index) occurred. This is suggestive that
while specific physiological differences were measured, effects were not severe enough to
cause large gross scale physiological effects.
Results from 1990, indicated a slightly different seasonal patterns of physiological
change, as condition indices were suggestive of a delayed period of glycogen accumulation
(June) and spawning (July) in 1990 compared to 1989 (May -glycogen accumulations and
June - spawning activity). An absence of significant pesticide runoff, other than the small
azinphosmethyl concentrations observed at the KWA Site, was observed during 1990. As
a result, physiological parameters were not significantly different in comparisons of oysters
between both sites during 1990.
While the results for 1989-90, demonstrated the usefulness of oyster ecophysiology
measurements to assess nonpoint source pesticide runoff effects, it is important to note the
significance of confounding factors such as low salinity exposure, which may co-occur with
pesticide exposure. Only by careful study and appropriate study design may the effects of
confounding factors such, as salinity be differentiated from pesticide effects per se. Thus
it is extremely important that study design incorporate an appropriate number of controls
to address physiological responses to non-contaminant environmental fluctuations, so that
contaminant effects per se may be elucidated and statistically discerned. Low salinity is
a particularly important factor, since it will occur concomitant with NFS runoff of chemical
contaminants.
177
-------
III. Laboratory Toxicity Tests
A. Effects on~Brain AChE Activity
1. Laboratory Phase - ECX Determination
The results of the 24h laboratory exposure experiments to determine the level of brain
AChE inhibition produced in mummichogs exposed to a series of azinphosmethyl
concentrations are shown in Figure 46. The predicted 24h ECM (concentration necessary
to produce a 50% reduction in AChE activity following 24h of exposure) was 0.90 ^g/L.
2. Relationship Between Specific Levels of Azinphosmethyl - Induced Brain AChE
Inhibition and Sublethal Effects on Respiration, Nitrogen Excretion and O/N
Ratios
Figure 47 shows the effects on brain AChE observed in mummichogs exposed to
azinphosmethyl at 2.4/ig/L for 24h, both initially and following eight days of depuration.
Brain AChE activity was reduced by 81% m the mummichogs immediately following 24
hours of exposure. Following eight days of depuration, brain AChE activity had recovered
to about 70% of normal but was still significantly (p < 0.05) lower than controls.
Figure 48 shows the oxygen consumption rates observed in mummichogs exposed to
azinphosmethyl at 2.4 jtg/L, both immediately following 24h of exposure and after eight
days of depuration in clean water. Oxygen consumption rates observed in control animals
at 24h and at eight days are also shown. There was no significant (p > 0.05) difference
in oxygen consumption between the 24 h azinphosmethyl exposed animals and the
corresponding control group. Neither was there a significant (p > 0,05) difference between
the eight day control group and the eight day treatment group. The only groups which had
significantly (p > 0.05) different oxygen consumption rates were the 24h control and the
eight day treatment groups. >
178
-------
AZINPHOSMETHYL CONCENTRATION
VS
% ACHE INHIBITION (LABORATORY)
100
2 80 -
o
X 60 -
LU 40 -
O
* 20H
y = 53.734 + 63.902X
ECSO = 0.90 ug/L
RA2 = 0.831
-1 .0
-0.5 0.0 0.5
AZINPHOSMETHYL CONCENTRATION
(LOG 10)
1.0
Figure 46. Laboratory predicted 24h EC 50 (ug/L) based on 50 % brain AChE inhibition in
F. heteroditus exposed to azinphosmethyl for 24h.
179
-------
IT)
<
CC
c
i
-
I—
UJ
100 -
CONTROL TREATMENT CONTROL TREATMENT
INITIAL DEPURATION
EXPOSURE GROUP
• » GROUPS WITH"5AHE LETTER NOT 5 I5NIF 1C ANTLf
DIFFERENT AT ALPHA =0.05
Figure 47. Brain AChE levels in F. hetertclitus exposed to a sublethal dose of azinphosmethyl
in the laboratory for 24h. Note the significant AChE inhibition following initial
exposure and the partial recovery some 7 days later.
180
-------
OXYGEN CONSUMPTION IN FUNDULUS
FOLLOWING SHORT-TERM
AZINPHOSMETHYL EXPOSURE
CONTROL
TREATMENT
INITIAL
CONTROL TREATMENT
DEPURATION
EXPOSURE GROUP
* GROUPS WITH SAME LETTER NOT SIGNIFICANTLY
DIFFERENT AT ALPHA -0.05
Figure 48. Respiration rates (ug atoms 02/g/h) in F. heteroditus exposed to a sublethal dose
of azinphosmethyl for 24h followed by a 168 hour depuration period. Exposure to
azinphosmethyl did not affect respiration rates in mummichogs acutely exposed,
despite the high levels of brain AChE.
181
-------
This difference did not appear to be related to insecticide exposure, but may have been due to
the effect of the experimental confinement on the eight day groups or the handling of stress
experienced by the 24h groups. One of these possibilities seems most likely, since oxygen
consumption rates tended to be lower in both the eight day treatment and eight day control
groups than in the corresponding 24h groups. Oxygen consumption rates ranged from 81.01
Hg atoms 0:/g dry weight/h in the 24h control group to 53.00 pg atoms 02/g dry weight/h in
the eight day treatment group.
Figure 49 shows the nitrogen excretion rates observed in control mummichogs and those
exposed to azinphosmethyl for 24h at 2.4 ^g/L. Nitrogen excretion rates were significantly (p
< 0.05) lower in the 24h treatment group and both the treatment and control depuration groups
than in the initial control group. Mean nitrogen excretion rates ranged from 11.38 ^g atoms
N/g dry weight/h in the 24h control group to 3.67 fig/g dry weight/h in the eight day control
group.
The O/N ratios determined for the control munrmichogs and those exposed to
azinphosmethyl at 24h and eight days are shown in Figure 50. Mean O/N ratios ranged from
7.40 in the 24h control group to 25.65 in the 24h treatment group. There was no significant
(p > 0.05) difference in the O/N ratio for an^ of the four groups.
C. Discussion and Conclusions:
Relationship Between Specific Levels of Azinphosmethyl - Induced Brain AChE
Inhibition and Sublethal Effects on Respiration, Nitrogen Excretion and O/N Ratios
The results of these experiments indicated that short-term exposure (24h) of
mummichogs to azinphosmethyl at 2.4 ^g/L resulted in - 81% inhibition of AChE
immediately following exposure. Following eight days of depuration this activity had
recovered to ~ 70% of normal but was still significantly lower than that in control
animals. No significant effect on oxygen consumption was observed in the fish exposed
to azinphosmethyl at 2.4 jtg/L for 24ty either immediately following exposure or after eight
days of depuration.
182
-------
NITROGEN EXCRETION IN FUNDULUS
FOLLOWING SHORT-TERM
AZINPHOSMETHYL EXPOSURE
CONTROL TREATMENT CONTROL TREATMENT
INITIAL DEPURATION
EXPOSURE GROUP
* GROUPS WITH SAME LETTER NOT SIGNIFICANTLY
DIFFERENT AT ALPHA -0.05
Figure 49. Nitrogen excretion rates (ug atoms N/g/h) in F. heteroclitus exposed to a sublethal dose
of azinphosmethyl for 24h, folfbwed by a 168 h depuration period. Exposure to
azinphosmethyl resulted in significantly decreased nitrogen excretion rates. This effect
was not evident in depuration phase organisms.
183
-------
OXYGEN/NITROGEN RATIOS IN FUNDULUS
FOLLOWING SHORT-TERM
AZINPHOSMETHYL EXPOSURE
CONTROL TREATMENT CONTROL TREATMENT
INITIAL
DEPURATION
EXPOSURE GROUP
* GROUPS WITH SAME LETTER NOT SIGNIFICANTLY
DIFFERENT AT ALPHA:-0.05
I
Figure 50. Mean O/N Ratios in F. heteroclitus exposed to a sublethal concentration of
azinphosmethyl for 24h followed by a 168h depuration period. Azinphosmethyl
exposure caused no significant effect on O/N ratios in mummichogs.
184
-------
Nitrogen excretion was significantly lower in the 24h treatment group and in both the
treatment and control depuration groups than in the initial control group. It is possible that
the effect observed in the 24h treatment group may have been a result of the insecticide
exposure while ffiose seen in the control and treatment depuration groups may have resulted
due to the stress of environmental confinement.
O/N ratios were not significantly different in any of the four groups although these
ratios were generally lower in the 24h treatment groups and both the control and treatment
depuration groups than in the 24h control group.
It is of interest to note that the azinphosmethyl concentration (2.4 ng/L) which
produced - 81 % inhibition following 24h of exposure in these experiments is about 0.065
times the 96h LCW for this compound in mummichogs of 36.95 ng/L reported by Fulton
and Scon (1991). This suggests, together with the fairly minor metabolic alterations
observed in these experiments concurrent with high levels of AChE inhibition, that a fairly
large reserve of brain AChE activity exists in this species at least as it relates to acute
lethality and the sublethal metabolic parameters examined in these experiments.
IV. Biomarker Studies 7
A. Brain AChE in Mummichogs
1. Field Exposures
Ninety-six hour field exposure tests with mummichogs were conducted during June of
1989 and May-June of 1990. Four field exposure tests were conducted during each of
these years at the CTL, TRT and KWA Sites.
Rainfall data for the field exposure tests conducted during 1989 and 1990 are shown
in Tables 39 and 40. There were six periods of rainfall during the 1989 field exposure at
each of the three field sites. Four (of these events resulted in total rainfall amounts
> 1.27cm/24h at the CTL and TRT Sites while five of the rain events resulted in total
amounts > 1.27cm/24h at the KWA Site. During the 1990 field exposures,
185
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Table 39. Dates of significant rainfall (> 1.27 cm/day) during the 1989 field study.
"1989
Site
CTL
Date
6/5/89
6/6/89
6/9/89
6/16/89
6/19/89
6/24/89
Rainfall Amount (cm/day)
Range
4.70-4.83
3.30- 3.56
1.27 - 1.52
0.89 - 0.899
2.97 - 3.05
0.25 - 0.25B
Average
4.75
3.43
1.35
0.89
3.02
0.25
(±SE)
(0.05)
(0.08)
(0.08)
(0.00)
(0.03)
(0.00)
TRT
6/5/89
6/6/89
6/9/89
6/16/89
6/19/89
6/24/89
4.83 -4.95
3.30 - 3.53
1.52- 1.65
2.03 -'-2.10
1.21 - 1.27B
0.19 -0.328
4.90
3.43
1.57
2.30
1.22
0.25
(0.05)
(0.08)
(0.05)
(0.02)
(0.03)
(0.03)
KWA
6/5/89
6/6/89
6/9/89
6/16/.89
6/19/89
6/24/89
7.37 -7.62
8.38 - 8.64
2.03 - 2.29
1.40 - 1.52
0.00 - 0.00
4.57 - 4.57
7.54
8.46
2.11
1.47
0.00
4.57
(0.08)
(0.08)
(0.08)
(0.03)
(0.00)
(0.00)
\
A = Range between three rain gauges
B = Rainfall < 1.27 cm/day but included for comparative purposes
X = Mean
SE = Standard error
186
-------
Table 40. Dates of significant rainfall (> 1.27 cm/day) during 1990 field study.
1990
Site
CTL
Date
5/28/90
6/15/90
Rainfall Amount (cm/day)
Range
3.00- 3.05
1.40- 1.52
Average
3.02 .
1.45
(±SE)
(0.03)
(0.05)
TRT
5/28/90
6/15/90
2.67 - 3.05
1.98-2.03
2.90
2.01
(0.13)
(0.03)
KWA
5/28/90
6/15/90
2.21 -2.31
1.78 - 1.78
2.24
1.78
(0.03)
(0.00)
A = Range between three rain gauges
B = Rainfall < 1.27 cm/day but included for comparative purposes
X = Mean
SE = Standard error
187
-------
there were only two periods of rainfall at each of the sites. Each of these rain events
resulted in total rainfall amounts > 1.27cm/24h at each of the three field sites. Results of
insecticide analysis of water samples collected during the 1989 and 1990 field exposure
tests are shown in Tables 7-13 and 15 - 17 and Figures 3 - 5 and 6-8. In general the
highest measured insecticide concentrations were associated with periods of significant
rainfall (> 1.27cm/24h) and depressed salinities in the tidal creeks.
The maximum insecticide concentrations, cumulative rainfall and minimum salinities
measured at each of the field sites during the first field exposure test (June 3-7) of 1989
are shown in Table 41. The maximum insecticide concentration measured at the CTL Sice
during this field test was 0.014 /xg/L of endosulfan while at the TRT Site measurable
concentrations of three insecticides were detected. The highest insecticide concentration
measured at the TRT Site was fenvalerate at 0.093 /ig/L. Endosulfan and azinphosmethyl
were measured at 0.020 /zg/L and 0.016 /zg/1, respectively. The highest insecticide
concentration measured at the KWA Site was azinphosmethyl at 1.730 /ig/1. Endosulfan
and fenvalerate were detected at 0.163 jzg/1 and 0.054 /xg/1, respectively.
The maximum insecticide concentrations measured at the CTL and TRT Sites during
the second field exposure test (June 11 -15) of 1989 were endosulfan at 0.012 ng/\ and
0.010 fj.g/1, respectively. Measurable concentrations of three insecticides were again
detected at the KWA Site. The highest insecticide concentration measured at this site was
azinphosmethyl at 0.368 jzg/1, respectively.
During the field test of June 15-19, 1989 endosulfan was detected at the CTL Site at
0.012 ^g/1 while endosulfan and fenvalerate were measured at the TRT Site at 0.010 /ig/1
and 0.015 jzg/1, respectively. The highest insecticide concentration measured at the KWA
Site was azinphosmethyl at 2.457 /ig/L while endosulfan was detected at 0.038 /ig/L.
During the final field exposure test of 1989, only trace amounts of endosulfan
(<0'.010 ng/1) were detected at the CTL and TRT Sites. At the KWA Site, however,
azinphosmethyl was measured at 7.002 /xg/L. Additionally, endosulfan was detected at
0.065 jig/L. ''
188
-------
Table 41. Summary of maximum measured insecticide concentrations, minimum
salinity and cumulative rainfall measured during the 1989 field study.
A.
Field Exposure (June 3-7, 1989)
Exposure
Site
CTL
TRT
KWA
TotaT Rainfall
(on)
8.18
833
16.00
Minimum Salinitv
(0/00)
5
5
0
Maximum Insecticide
Concentrations (/ig/L)
Endosulfan (0,014)
Endosulfan (0.020)
Azinphosmethyl (0.016)
Fenvalerate (0.093)
Endosulfan (0.163)
Azinphosmethyl (1.730)
Fenvalerate (0.054)
B.
Field Exposure (June 11-15, 1989)
Exposure
Site
CTL
TRT
KWA
Total
Rainfall
0
0
0
Minimum
Salinity (ppt)
26
18
4
Maximum Insecticide
Concentrations 0
-------
No rainfall occurred at either of the field sites during the first field exposure test (May
24-28) of 1990. No water samples from either of the field sites were analyzed for
insecticide residues for this time period. The maximum measured insecticide
concentrations, cumulative rainfall and minimum salinities measured at each of the field
sites during the second (May 28 and June 1, 1990) field test of 1990 are shown in Tables
15 -17 and 40. No insecticides were measured at concentrations above the detection limit
at either the CTL or KWA Sites. At the TRT Site, endosulfan and fenvalerate were
measured at 0.014 /xg/L and 0.123 Mg-'L, respectively. The maximum measured
insecticide concentrations, cumulative rainfall and minimum salinities measured at each of
the field sites during the third (June 13-17) field exposure test of 1990 are shown in Tables
15-17 and 40. Endosulfan was detected at the CTL and TRT site at 0.009 jxg/L and
0.005 /xg/L, respectively. At the KWA Site, azinphosmethyl was measured at 0.062 ng/L.
No rainfall occurred at either of the field sites during the final (June 21-23) field test of
1990. No water samples from either of the field sites were analyzed for insecticide
residues for this period.
B. Field Effects on Brain AChE in Mummichogs
Brain AChE specific activity levels in" mummichogs deployed at the field sites during
the field studies conducted in 1989 and 1990 are shown in Figures 51 and 52, respectively.
Brain AChE specific activity levels in mummichogs deployed at the field sites during
the first field deployment (June 3-7) of 1989, ranged from 321.70 nmol mgP1 min ' at the
TRT Site to 123.02 nmol mgP1 min 'l at the KWA Site. Brain AChE levels were
significantly (P ^ 0.05) depressed in the animals deployed at the KWA Site when
compared to those deployed at the other two field sites. There was no significant
(P > 0.05) difference between activity levels in the animals deployed at the other two sites.
During the second deployment (June Ll-15) of 1989 there was no significant (P>0.05)
difference between brain AChE levels in animals deployed at the TRT Site (379.05 nmol
mgP1 min •') and those deployed at the KWA Site (352.59 nmol mgP1 min 'l). Brain
AChE levels were not determined for Animals deployed at CTL Site during this test because
of high mortality in this group that occurred as a result of depressed DO in the holding
tank after the removal of these animals from
190
-------
BRAIN ACETYLCHOLINESTERASE ACTIVITY
IN MUMMICHOGS DEPLOYED DURING 1989
FIELD STUDIES
01
C/l
tf>
UJ
o
o
LU
-------
GRAIN ACETYLCHOLINESTERASE ACTIVITY
IN MUMMICHOGS DEPLOYED DURING 1990
FIELD STUDIES
100 -
<
m
<
A <*> — „ —
" C 5! ~ 5;
<^ 10 10 10 10
i_ < -j i- <
P E G p E
Bars Represent Standard Errors
Figure 52. Brain AChE levels measured in F. heteroclitus exposed in the field during the 1990
field study. Note the similarities in fish brain AChE levels for all sites during 1990.
192
-------
the field site. Brain AChE specific activity level in animals deployed at the field sites from
June 15-19 ranged from 384.53 nmol mgP'1 min1 in animals from the TRT Site to 51.44
nmol mgP'1 min ' in animals from the KWA Site. Specific activity was significantly (P <
0.05) lower in-the animals deployed at the KWA Site than in animals deployed at either
the CTL or TRT Site. There was no significant (P>0.05) difference between the levels
at the TRT (384.53 nmol rngP'1 mm'1) and CTL (366.55 nmol mgP'1 min1) Sites. During
the final (June 23-27) field deployment of 1989, AChE specific activity levels ranged from
361.87 nmol mgP'1 mm'1 at the TRT Site to 5.90 nmol mgF1 min1 at the KWA Site.
Brain AChE specific activity was significantly (P < 0.05) lower in animals deployed at
the KWA Site than in animals deployed at either of the other two sites. Once again there
was no significant difference between the AChE activity levels in animals deployed at the
CTL and TRT Sites.
During the first deployment of 1990 (May 24-28) brain AChE levels in mummichogs
deployed at the field sites ranged from 300.24 nmol mgP1 min'1 at the TRT Site to 326.24
nmol mgP"' min' at the CTL Site. There was no significant (P > 0.05) difference in the
level of brain AChE activity in animals deployed at any of the three field sites. A similar
pattern was observed during the second field deployment (May 28-June 1) of 1990. Brain
AChE activity in animals deployed at the three sites ranged from 304.59 nmol mgP1 min'
in the KWA animals to 326.24 nmol mgP'1 min"1 in animals deployed to the TRT Site.
Once again there was no significant (P > 0.05) difference in the level of brain AChE
activity among the animals deployed at any of the three sites.
During the field exposure test of June 13-17 brain AChE levels ranged from 260.90
nmol mgp1 min'1 in animals deployed at the KWA Site to 283.32 nmol rngP"1 min'1 in the
animals from the TRT Site. There was no significant (P > 0.05) difference in the level
of brain AChE activity among fish deployed at any of the three field sites. The results of
the final field deployment (June 21-23) of 1990 were quite similar to those from the earlier
deployments conducted during the year Mummichog specific activity levels for brain
AChE ranged from 305.41 nmol mgP"1 min'1 in animals deployed at the KWA Site to
311.44 nmol mgP"1 min'1 in the animajs from the CTL Site. There were no significant (P
> 0.05) differences in the levels of btain AChE activity among animals from any of the
three field sites.
193
-------
C. Discussion and Conclusions - Field Exposure Tests
Climatic conditions during the field exposure tests conducted during 1989 and 1990
provided an extremely interesting contrast, with the 1989 studies being carried out during
an extremely wet period and the 1990 studies conducted during a relatively dry period.
During the field tests of 1989 (June 3 - 27, 1989), total rainfall amounts at the t+iree field
sites ranged from 16.36 cm at the CTL Site to 25.40 cm at the KWA Site. During 1990,
total rainfall amounts (May 24 - June 23) ranged from 4.32 cm at the KWA Site to 5.31
cm at the TRT Site. These different rainfall characteristics most probably contributed to
the very different sublethal impacts observed on brain AChE observed in animals deployed
at one of the field sites during field studies conducted during these two years.
Much higher insecticide concentrations were observed in water samples collected at
the KWA Site during the field exposure tests conducted in 1989 than at either of the other
two field sites. Azinphosmethyl was the insecticide typically measured, with highest
concentrations at the KWA Site. This compound was measured at the KWA Site during
each of the field exposure test in 1989, with maximum concentrations ranging from 0.37
jig/L to 7.00 /xg/L. Endosulfan was also detected at relatively high concentrations at the
KWA Site during 1989. The maximum -measured endosulfan concentration at this site
ranged from 0.04 /xg/L to 0.16 ^g/L for the four exposure tests, Lesser amounts of
fenvalerate (0.03-0.05 ng/L) were detected at the KWA Site during 1989.
In contrast to the high levels of insecticides measured at the KWA Site, only very
small amounts of insecticides were detected in water samples from the CTL and TRT Sites
during 1989. Endosulfan was the only insecticide detected in water samples collected at
the CTL Site during 1989 and the highest concentration measured was endosulfan at 0.01
Mg/L. Relatively small amounts of endosulfan, azinphosmethyl and fenvalerate were
detected at the TRT Site-with maximum measured concentrations of 0.02 pg/L, 0.02 jtg/L
and 0.09 jxg/L, respectively.
Very high levels of AChE inhibition were observed in mummichogs deployed at the
KWA Site during three of the four fieW exposure tests conducted during 1989 and the level
of this inhibition was closely related to azinphosmethyl concentration measured in water
samples collected at this site (Table 41). The level of AChE inhibition for the four field
194
-------
exposures ranged from 0-98% while the corresponding maximum measured
azinphosmethyl concentrations ranged from 0.37 - 7.00 /ig/L.
During 1990, only very minor insecticide concentrations were measured at any of the
three field sites. The maximum measured insecticide concentration at the CTL Site during
1990 was endosulfan at 0.01 v-g/L while at the TRT Site the maximum measured
insecticide concentration was fenvalerate at 0.123 ug/L. The maximum insecticide
concentration measured at the KWA Site was azinphosmethyl at 0.062 ng/L. No
significant effects on brain AChE activity were observed in mummichogs deployed at either
of the field sites during 1990.
A comparison of the subleihal effects on brain AChE observed in mummichogs
deployed at the KWA Site during the 1989 field studies and the lack of a similar effect in
1990 appear to demonstrate the importance of nonpoim source agricultural runoff as a
transport mechanism for the movement of insecticides from agricultural fields into the
adjacent estuarine tidal creeks. Total rainfall amounts at the KWA Site during 1989 field
studies were more than five fold higher than for a similar period in 1990.
D. Discussion and Conclusions '
Sublethal Effects of Azinphosmethyl on Brain AChE-Comparison of Field
and Laboratory Effects
The results of field studies conducted in 1989 indicated that significant concentrations
of insecticides entered the tidal creek at the KWA Site on several occasions following
significant (> 1.27 cmy24h) rainfall events. Brain AChE activity was depressed in caged
mummichogs deployed at this site on three separate occasions. In each of these events,
water samples collected at the site contained residues of the OP insecticide,
azinphosmethyl. The maximum concentrations measured during each of these events
ranged from 1.73 -7.00 >zg/L. These results are quite similar to those previously reported
by Scott et al, 1990 for field studies^conducted at the same site during 1988. In those
studies they found significant inhibition of AChE in mummichogs deployed at the KWA
Site when azinphosmethyl concentrations were ^ 0.57
195
-------
There was also excellent agreement between the sublethal effects on brain AChE
observed in the field studies and the results obtained form the laboratory experiments. As
previously discussed, a 24h EC50 for brain AChE inhibition of 0.90 jig/L was determined
for azinphosmethyl based on laboratory exposures (Figure 46). This value was
subsequently compared to azinphosmethyl concentrations and effects of AChE measured
in the field studies. Table 42 shows the azinphosmethyl concentrations and the effects on
AChE activity measured during field deployments conducted in 1988 and 1989. The 1989
data were described earlier in this report and the 1988 data were reported by Scott et al,
1990. The data shown in this table were used to calculate field derived EC50's for AChE
inhibition. Three different approaches were utilized in the treatment of these data. First,
an EC50 was calculated based on the maximum azinphosmethyl concentration measured for
a particular field deployment. This approach produced an ECM of 1.53 ^ig/L (Figure 53).
Next, an EC^ of 0.63 ng/L (Figure 54) was calculated based on the 24h azinphosmethyl
concentration (the residual azinphosmethyl concentration present in water samples collected
- 24h after the sample containing the highest azinphosmethyl concentration). Finally, an
EC50 of 1.13 jxg/L (Figure 55) was determined based on the 24h average concentration (the
average of the maximum measured azinphosmethyl concentration and the azinphosmethyl
concentration remaining 24h later). This value was quite similar to the laboratory derived
EC50 of 0.90 ng/L. These results suggestnhat a simple 24h laboratory exposure is a good
predictor of the effects on brain AChE produced following exposure to azinphosmethyl
residues present in nonpoint source agricultural runoff.
196
-------
Table 42. Summary of Insecticide Related Effects on Brain AChE Activity Observed in Field Studies
Conducted in 1988 and 1989
Field
Test
Date
6/7 - 11/88
6/11 - L5/88
6/3 - 7/89
6/11 - 15/89
6/15 - 19/89
6/23 - 27/89
Maximum Measured
Azinphosmethyl
concentrations Og/L)
3.44
0.57
1.73
0.37
2.46
7.00
Azinphosmethyl
Concentration
at 24h Otg/L)
0.57
0.55
1.12
0.21
0.72
1.60
%
AChE
Inhibition
47
22
63
0
85
98
197
-------
MAXIMUM AZINPHOSMETHYL CONCENTRATION
VS
% ACHE INHIBITION
100
O 80 H
y = 37.089 + 69.477x
EC50 = 1.53 ug/L
R*2 = 0.809
-0.5
0.0
0.5
AZINPHOSMETHYL CONCENTRATION
(LOG 10)
Figure 53. Predicted EC50 (ug/L) based upon fieldjmeasured, brain AChE levels in F. heteroctitus exposed
to azinphosmethyl. The maximum field exposure concentration was used to predict the EC50
value.
198
-------
24 HOUR AZINPHOSMETHYL CONCENTRATION
VS
% ACHE INHIBITION
100
z
g
»-
m
X
z
HI
I
o
80 -
60 -
40 -
20-
y = 71.924 + 108.7SX
EC50=0.63 ug/L
RA2 = 0.781
•1 .0
•0.5 0.0
AZINPHOSMETHYL CONCENTRATION
(LOG 10)
0.5
Figure 54. Predicted EC50 (ug/L) based upon fiel<& measured, brain AChE levels in F. heteroclitus exposed
to azinphosmethyl. The azinphosmethyl concentrations measured 24h after the maximum
concentration was observed, was used to predict the EC50 value.
199
-------
MEAN AZINPHOSMETHYL CONCENTRATION
VS
% ACHE INHIBITION
100
z
2 so-
m
r
HI
z
o
y = 45.675 + 82.007X
EC50=1.13 ug/L
60 -
40 -
20 -
-0.6
•0.4
-0.2 0.0 0.2 0.4
AZINPHOSMETHYL CONCENTRATION
(LOG 10)
Figure 55. Predicted EC50 (ug/L) based upon field measured, brain AChE levels in F. heteroclitus exposed
to azinphosmethyl. The mean azinphosmethyl concentration (maximum + 24h concentrations/2)
was used to predict the EC50 value.
200
-------
V. Ecotoxicological Studies
A. Block SeRfing 1989-90
li Biomass
Results of total biomass (g/50m of stream) measurements for all macropelagic
(> 15mm) fauna are given in Tables 43 - 44 and Figure 56. Table 43 lists the species
observed in sample collected during this time period. Total biomass (Table 44) ranged
from 524.3-7,066.7 g/50 m of stream at the CTL Site compared to a range of 1,510.0-
10,666.7 g/50 m of stream at the TRT Site for the period of January, 1989 -
September, 1990. Peak biomass was observed during September, 1989 and September,
1990 at the CTL Site and during August, 1989 and July, 1990 at the TRT Site.
Total mean biomass (January, 1989 - September, 1990) was 91,947.8 g/50 m of
stream at the CTL Site versus 100,504.6 g/50 m of stream at the TRT Site. This
indicated that total biomass for this period was 8.5% higher at the TRT Site than CTL
Site. v
Much of this between site difference in biomass may be attributed to Hurricane
Hugo. Statistical analysis indicated that prior to Hurricane Hugo, biomass was higher
at the TRT Site only in paired sample comparisons on two occasions, July and August,
1989. During these sampling periods biomass was increased primarily due to the
greater abundance of P. pugio at the TRT Site (8,781-16,508.7/50 m of stream)
compared to the CTL Site (2,031.7-7,266/50 m of stream). After Hugo, biomass was
significantly different in both paired and unpaired sample comparisons during
November, 1989 through January, 1990. Statistical analysis indicated that biomass
measurements were 6,080.4 g higher during this time period at the TRT Site than at the
CTL Site. This 6,080.4 g difference observed post Hugo would account for 71% of
the 8.5% difference observed between sites for the entire study period (1/89 - 9/90).
In fact, the biomass levels observed at the TRT Site during November, 1989 - January,
1990, were the highest ever observed at either site from 1985-1991, for winter
201
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Table 43. List of pelagic species identified during ecotoxicological sampling 1989 - 1990.
MARINE SPECIES LIST
Aleais criniius
Anchoa mitchilli*
Bairdiella chrysoura*
Brevoortia tyrannus*
Callineaes sapidus*
Caranx hippo*
Cemroprisis striata*
Cynoscion nebulosis*
Cyprinodon varieagaius
Dorsoma cepedianum
Elops saurus
Eudnostomiu guta
Fundulus Heterocliius
Fundulus majalis
Gobiosoma bosci
Lagadon rhomboides*
Leiossomos xaiuhurus*
Lolliguncula brevis*
Menidia menidia
Micropogon undulaius*
Monocanthus hispidus
Mugil cephalus*
Opsanus tail
Palaemonetes pugio
Palaemontes vulgaris
Panopeus herbstii
Paraliduhys deruatus*
Paralichthys lethostigma*
Penaeus aztecus*
Penaeus duoranun
Penaeus setifenu
Poecilia Uutipinna
Pomatomus saliMrix*
Prionotus tribuius
Selene vomer
Symphunu plagiiaa ;t
Sygnathus fiacus 1
Synodus foeiens
Sphoeroides maculatus
Sphyraena barracuda*
Uca pugilator
Pompano
Bay Anchovy
Silver Perch
Atlantic Menhaden
Blue Crab
Jack Cravelle
Black Sea Bass
Spotted Sea Trout
Sheepshead Minnow
Gizzard Shad
Lady fish
Silver jenny
Mummichog
Striped Killifish
Goby
Pinfish
Spot
Squid
Atlantic Silverside
Atlantic Croaker
File Fish
Mullet
Toad Fish
Grass Shrimp
Grass Shrimp
Mud Crab
Northern Flounder
Southern Rounder
Brown Shrimp
Pink Shrimp
White Shrimp
Salifm Molly
Blue Fish
Bighead Sea Robin
Lookdown
Tongue Fish
Pipe Fish
Lizard Fish
Puffer Fish
Barracuda
Fiddler Crab
Denotes commercial and recreational species
202
-------
Table 44. Summary of total biomass measurements (grams/50 m of stream) observed in block seining at the
CTL and TRT Sites, 1989-90. Asterisks (";") indicate dates when samples were significantly (0.05 -
0.10) different.
Parameter: Total Biomass (Grams/50 m of stream)
Date
1/89
2/89
3/89
4/89
5/89
6/7/89
6/26/89
7/89
8/89
9/89
10/89
11/89
12/89
1/90
2/90
3/90
4/90
5/11/90
5/30/90
6/90
7/90
8/90
9/90
Total
CTL Site
X
1,933.3
3,766.7
4,000.0
4,666.7
4,666.7
2,766.7
6.066.7
3,566.0
5,200.0
7,066.7
4,533.3
1,220.0
524.3
639.3
4,504.3
6,066.7
2,055.7
5,333.3
2,621.3
4.616.7
4,166^7
5,600.0
6,366.7
91,947.8
SE
999.9
2,105.8
1,738.8
1,301.7
1,371.5
1,481.4
1,139,2
1,502.1
1,951.9
1,909.9
600.9
240.2
77.4
103.6
2,246.8
3,658.0
633.3
1,965.0
321.5
696.4
578.3
529.2 -
U271.9
TRT Site
X
4,166.7
1,933.3
4,266.7
3,266.7
3,866.7
3,333.3
7,900.0
7,233.3'
10,666.7'
3,833.3
4,666.7
3,466.7'"
3.486.7"
1,510.0";'
4,833.3
2,500.0
4,593.7
6,000.0
2,774.7
3,516.7'*
6,058.0
3,000.0":*
3,631.4
100,504.6
SE
617.3
384.4
696.0
7126
592.5
592.5
2,688.9
896.9
3,023.4
433.3
961.5
39.0
1,548.6
97.1
902.5
500.0
2,461.9
763.8
868.2
1,140.3
1.810.7
0.0
1,028.8
Significantly (p £
Significantly (p £
Significantly (p £
0 05) Different in Unpaired Test; N = 6
0,10) Different in Paired Test; N=6
0.075) Different in Unpaired Test; N=l
*• Hugo
203
-------
E
35
I"
CO ,_
_ o
Ij
14000
12000 -
10000 -
8000 -
6000 -
4000 -
2000 -
Total Biomass 1989-1990 (Seine)
O— Control
Treatment
I
CO
O) O) Ok O) Ol O> O
CM t^ co co to r^-
• - CM CJ CM --".CM CM
to
to
to
0>
— O 0> CM CM i-
*" CM *" *~ "~
O fO *- —
CO CM CM fvi
co v in ui to
DATE
Figure 56. Total biomass (g/50m of stream) measured in block seining, 1989-90. Note the general
similarities in biomass at the CTL and TRT Sites during this study. Asterisks (*) indicate
samples which were significantly (0.05) different in statistical comparisons between the
TRT and CTL Sites. Arrow (t) denotes Hurricane Hugo.
204
-------
(December-January) sampling periods. Generally during the winter sampling, biomass
is < 1000 g/50 m of stream at both sites due to the reduced temperatures observed.
December, J.989 was an extremely cold month with record snowfall (> 20 cm of snow)
and a period of nearly one week when maximum daily air temperatures were below
freezing. Despite this, record biomass levels were observed at the TRT Site. Dispersal
of organisms by Hurricane Hugo may have accounted for this.
During Hurricane Hugo, winds on the back side of the hurricane eye caused
extremely low tides at Leadenwah Creek (Jimmy Green, personal communication), as
the storm surge occurred well to the north of this site. Winds would have blown from
the northwest -* southeast, pushing water and possibly small organisms from the CTL
Site towards the TRT Site and other down wind portions of Leadenwah Creek.
Sampling at the CTL and TRT Sites during September was conducted 2-3 days prior
to Hugo and indicated relatively equivalent biomass at each site. One month after
Hugo, again biomass was equivalent at both sites. Two to four months after Hugo,
elevated biomass levels were observed at the TRT Site, mostly the result of significantly
higher levels of P. pugio and F. heteroclitus and increased levels of total fish. Many
of the grass shrimp measured in December, 1989 and January, 1990 would have been
small post larvae (particularly poor sw%nmers) at the time of Hugo. In all likelihood,
these grass shrimp were displaced from the CTL Site and other down wind habitats, as
the wind direction would have prevented grass shrimp from being as readily displaced
at the TRT Site.
Runoff of fenvalerate during June 1989, at the TRT Site appeared to have no
effect as biomass was not statistically different in between site comparisons during May
- June, 1989. Biomass was significantly (p<0.10) higher at the TRT Site during July -
August, 1989 primarily the result of increased levels of P. pugio and F. heteroclitus.
Similarly, runoff of fenvalerate during 28 May, 1990, appeared to have no
immediate effect as biomass measurements, two days post rain (30 May, 1990) were
not significantly different. Biomass in June was significantly (p <, 0.10) lower at the
TRT Site; however, reduced crustacean densities (species most sensitive to fenvalerate)
did not account for these differences. Similarly in August, 1990, three months post
fenvalerate runoff, biomass was significantly (p £ 0.05) lower at the TRT Site, mostly
the result of reduced mummichog densities (3651 vs 658/50 m of stream). Similarly
205
-------
total tish densities were significantly (p < 0.05) reduced at the TRT Site in August,
1990 (4639 vs 1.109/50 m of stream). Many of these fish species would have been
juvenile fish at the time of this rain event. Deployed juvenile C. variegatus had very
poor survival at both the TRT (64% survival) and CTL (60.9%) Sites during this rain
event. As a result ascribing effects due to fenvalerate to juvenile fish species and
resulting post runoff effects on biomass, seem tenuous at best.
2_. P. pugio Density
Results of P. pugio density (#/50 m of stream) measurements and statistical
comparisons are given in Table 45 and Figure 57. Mean grass shrimp densities ranged
from 23.7 (± 4.9) - 11,514.7 (± 7062.8)/50m of stream at the CTL Site compared to
a range from 125.9 (± 46.5) - 19,763.4 (± 4,228.1)/50 m of stream at the TRT Sites.
During 1989, peak grass shrimp densities at the CTL Site were observed in
February, March, June, August, September and October. At the TRT Site during 1989,
peak grass shrimp densities were observed during February, March, April, May, June,
July, August, September and December. These time periods corresponded with periods
of recruitment (i.e., usually February,-June, August, and November). During 1989,
significant runoff of fenvalerate was observed during 6/5-6/89 runoff event at the TRT
Site. A 30% reduction in P. pugio density was observed during this period at the TRT
Site. Following this time period, no additional reductions in P. pugio density were
observed during 1989.
For 1990, during the nine months reported, peak grass shrimp densities were
observed at the CTL Site during February, March, July and September. At the TRT
Site during 1990, peak grass shrimp densities were observed during January, February,
April, and July. The absence of peak densities at the CTL Site during April, 1990
resulted from significant predation by mummichogs and other fish species. The reduced
densities at the TRT Site during September, 1990 may have resulted from significant
fenvalerate runoff observed on the 28 May 1990.
206
-------
Table 45. Summary of P. pugio density measurements (number/50 m of stream) observed in
block seining at the CTL and TRT Sites, 1989-90. Asterisks (";') indicate
samples which were significantly (p < 0.05 - 0.10) different.
Date
1/89
2/89
3/89
4/89
5/89
6/7/89
6/26/89
7/89
8/89
9/89
10/89
11/89
12/89
1/90
2/90
3/90
4/90
5/11/90
5/30/90
6/90
7/90
8/90
9/90
Total
Parameter: /
CTL
X
986.7
4,759.7
4.735.3
611,3
169.0
345.7
4,304.7
2,031.7
7,266.0
8,058.3
7,414.2
1,920.4
986.7
1,178.0
11,514.7
9,135.8
23.7
236.2
' 461.3
3,895.7
6,568.7 :
1,900.3
5,965.3
84,469
J. pugio Density (
Site
SE
573.
3,686.3
4,671.9
553.1
63.7
119.9
1,285.6
890.0
2,554.0
2,303.7
5,026.9
755.6
199.6 ?
276.5
7,062.8
9.073.5
4.9
217.8
270.8
1,717.2
2,966.7
881.6
3,026.3
\
#/50 m of stream)
TRT Sit
X
1,672.7
3.554.0
9,288.3
5.630.3";-
5,045. 3":'
3,526. Q"'
9,970.7
8,781.3'"
16,508.7
17,045.3
11,099.8*
1,857.9
15,469.5'
5,510.3'V
19,763.4
3,309.1
7, 426.2' '='
125.9
829.2
1,272.1
6,004.0
1,905.9
2,591.1
158,187
e
SE
1199.2
1,682.0
3,508.5
1,515.6
482.3
1,473.7
5,085.8
609.7
6,899.5
5,131.1
5,244.5
272.5
10,874.7
876.8
4,228.1
1,554.6
6,162.0
46.5
161.2
242.6
2,750.9
119.6
496.3
- Hugo
Significantly (p <, 0.05) Different
Significantly (p <, 0.10) Different
in Unpaired Test; N=6
in Paired Test; N=6
207
-------
P. pugio Density 1989-1990 (Seine)
E
a
O) «
a. _
E
o
in
20000 -
1 0000 -
•
0
_ ;
Jt\
d d C
CO 00 0
T <0 0
CM ^- C
- Control
Treatment
*
' * *
}
N
n 0* d o c
o oo ee eo e
NJ r- n co c
o co a
a r— a
i-<
; J
f ^
n e
9 a
3 U
I ,
*
*
\
L v
i
i—
• T
n e
9 a
1 U
1
J
t
I
i i
> i
' i
^
n O) c
9 co a
5 r>-
*
j
J
1
t i
\ t
\ 1 .
\ t
* ' /
* ' /
1 I /
I y
j— a
not
0 O> C
- o c
*J -- e
«
t
'i
i
N
*
i
, <
i
3 C
n c
n c
;\
3 C
It C
J C
*
\ /s/^vl
ll -fl^"*
D O O O O O 0
D Ot Ot O) 01 01 O)
NJ v~ ^3 o — »- »-
- DATE
Figure 57 P pugio densities (#/50m of stream) measured in block seining, 1989-90. Note toe general
' similarities in grass shrimp densities during 1989-90 at both sites. Asterisks (*) indicate
samples which were significantly (p < 0.05) different in statistical comparisons between
the CTL and TRT Sites.
208
-------
Although no toxicity was observed in caged adult P. pugio, larval P. pugio are four
to five times more sensitive to low salinity, fenvalerate exposure than adults. Measured
fenvalerate levels were more than 15 times greater than the 96h LC50 values for larval
P. pugio exposed to fenvalerate at low salinities.
A total of 84,469 grass shrimp/50 m of stream were collected at the CTL Site
versus 158,187 grass shrimp/50 m of stream at the TRT Site, during the 21 months of
this study. Previous studies (Scott et al, 1990; Hampton, 1987) conducted during 1986-
87, reported annual grass shrimp densities ranging from 55,293 - 114,000/50 m of
stream at the CTL Site, compared to densities ranging only from 26,200-54,000/50 m
of stream at the TRT Site. During 1986-87, there were significant impacts to P. pugio
at the TRT Site observed in both caged toxicity tests and in biomonitoring studies (Scott
et al., 1990).
During 1989, annual P. pugio densities were 44,764 at the CTL Site compared
to 109,448 at the TRT Site. In comparing these results, generally P. pugio densities
at the CTL Site have remained fairly constant. (Mean densities varied by less than a
factor 2.5 at the CTL Site versus 4.3 at the TRT Site). At the TRT Site, during
episodes of significant agricultural pesticide runoff during 1986-87, P. pugio densities
declined dramatically. During 1989, despite one period of significant fenvalerate runoff
at the TRT Site, P. pugio densities dramatically increased, approaching peak annual
densities observed at the CTL Site. These data may be suggestive that P. pugio
populations are extremely resilient, being able to flourish despite some low
concentrations of fenvalerate runoff occurring at the site. During 1986-87, fenvalerate
concentrations were much higher (> 100 - 890 ngVL) than were measured during 1989
(< 100 ng/L). Another factor may be that fenvalerate concentrations measured during
1989 were residues of Asana rather than pydrin. Fenvalerate results from 1986-87,
were measured as pydrin residues, rather than Asana.
An additional factor which may explain the higher P. pugio densities observed
included the much lower predator pressures at the TRT Site, as evidenced by low
populations of mummichogs and other fishes when compared to the CTL Site. This
would allow P. pugio to flourish at the TRT Site due to the reduced predator pressures.
209
-------
During the 9 months (January - September, 1990) sampled during 1990, total
mean P. pugio densities were 39,705/50 m of stream at the CTL Site compared to a
mean density of 48,739/50 m of stream at the TRT Site. During 1989, total grass
shrimp densities at the CTL Site were only 40% of densities measured at the TRT Site.
During 1990, grass shrimp densities at both sites were quite similar, with levels at the
CTL Site approaching 81% of the TRT Site population densities. The much higher
densities of mummichogs and other fish at the TRT Site during 1990, resulted in greater
predatory pressures in P. pugio populations. As result P. pugio densities at the TRT
Site were greatly reduced in 1990 as fish populations began to re-establish higher
population abundance.
3^ F. heteroditus Density
Results of F. heteroditus density measurements (#/50 m of stream) are listed in
Table 46 and depicted in Figure 58. Mean mummichog densities ranged from 18-
4391/50 m of stream at the CTL Site compared to levels ranging from 12 - 2,419/50
m of stream at the TRT Site. During 1989, peak mummichog densities at the CTL Site
were observed during March, April, May, July, August, September and October with
densities > 1000/50 m of stream. A the TRT Site during 1989, peak mummichog
abundance was measured during July, August and September. These time periods
generally correlated with entry of young of the year from, late spawning during 1988
entering size classes measurable by our seining technique (March - May) and
recruitment of first spawning (March) of 1989 young of the year entering measurable
size cohorts. During 1990, peak abundance at the CTL Site was observed during June -
September. At the TRT Site, peak densities were measured during June - July, 1990.
Total mean mummichog densities for the 21 months sampled during 1989-90,
were 31,651.6/50 m.of stream at the CTL Site versus 15,520.7/50 m of stream at the
TRT Site. Previous studies during 1986-88 (Scott et al., 1990; Hampton, 1987)
reported annual mummichog densities ranging from 17,224 - 24,100/50 m of stream at
the CTL Site, compared to levels ranging from 5,600 - 5,802.3/50 m of stream at the
TRT Site. During 1986-87, significant runoff of azinphosmethyl, endosulfan and
fenvalerate at the TRT Site resulted in depressed mummichog densities there, only 23-
33% of CTL Site populations.
210
-------
Table 46. Summary of F. heteroclitus density measurements (number/50 m of stream)
observed in block seining at the CTL and TRT Sites, 1989-90. Asterisks
(";") indicated samples which were significantly (p ^ 0.05 - 0.179) different.
Parameter: F. heteroclitus Density (#/50 m of stream)
Date
1/89
2/89
3/89
4/89
5/89
6/7/89
6/26/89
7/89
8/89
9/89
10/89
11/89
12/89
1/90
2/90
3/90
4/90
5/11/90
5/30/90
6/90
7/90
8/90
9/90
Total
CTL Site j
X
163.7
861.7
1,426.7
1,601.3
1,215.0
1,490.3
264.0
1,012.0
2,111.0
3,705.3
2,094.2
497.9
130.7
18.0
570.1
733.5
611.1
396.7
923.6
1,336.5
4,391.3
3.650.7
2,446.3
31,651.6
SE
82.6
311.7
785.6
1,352.0
363.0
1,437.0
64.4
501.5
1,378.8
L 1,353.8
1,044.1
111.2
70.9
14.5
168.9
309.1
216.8
226.8
453.0
832.7
-.1.305.9
198.9
840.2
H
TRT Site
X
12.0
77.3";*
142. 3";*
364.3"'
163.3"'
282.3
155.3
2,328.0*
2,419.0
1030.0*
995.8*
742.1
319.7
119.3*
116.0";*
202.4*
475.6
474.7
352.5'
1,947.1
1,817.3*
657.8"A;'A
' 326.6":*
15,520.7
SE
5.1
30.8
63.7
102.2
111.4
74.4
34.2
1,005.4
1,420.7
522.4
533.8
210.1
114.7
49.8
78.5
66.7
88.3
81.7
157.7
700.1
799.7
52.1
213.2
= Significantly (p
= Significantly (p
= Significantly (p
= Significantly (p
0.05) Different in Unpaired Test; N=6
0.10) Different in Paired Test; N=6
0.08) Different in Unpaired Test; N=5
0.179) Different in Paired Test; N=5
- Hugo
211
-------
F. heteroclitus Density 1989-1990 (Seine)
6000
~ 5000 -
3 ~
= "
o
l_ -_
3 o
a
"-S
CVJ -- CM CM
«- CM n *r
C*5 CO CO ^- CD
CM «~'- e\J CM •-
-> CO- — -~ --
in co r- aa
— — CM -> CM
o *- ^ »-
CMCM — OCO — >-CO
»-»-T-e)CMCMCMCN
v vo in co
DATE
Figure 58. F. heteroclitus densities (#/50m of stream) measured in block seining, 1989-90. Note the
higher densities at the CTL Site during most of 1989-90. Asterisks (*) indicated samples
which were significantly (p < 0.05) different in statistical comparisons of the TRT and
CTL Sites.
212
-------
During 1989, total mean mummichog densities were 16,753.8 at the CTL Site
compared to 9150.7 at the TRT Site. Mummichog densities at the TRT Site during
1989 were only 55% of CTL Site populations. Statistical analysis indicated that
populations~densities were significantly (p < 0.05-0.10) higher at the CTL Site during
February - May, September and October, 1989. During July, 1989 mummichog
densities at the TRT Site were significantly (p < 0.10) higher than CTL Sue levels.
The higher TRT Site populations observed during July were probably the result of
recruitment of young-of-the-year fish whose densities flourished due to the higher grass
shrimp densities measured at the TRT Site. Significant runoff of fenvalerate at the TRT
Site during June, 1989 caused no toxicity in caged mummichogs deployed at that Site.
Measured fenvalerate concentrations were below levels toxic to adult muminichogs.
Similarly, biomonitoring results for 1989 appeared to support these results, as no
toxicity directly attributable to runoff events was observed during June, 1989 sampling
results. The increased population densities observed during 1989 at the CTL Site are
likely the residual effects of previous fish kills at the TRT Site (June, 1985; May, 1986
and August, 1988).
During 1990 (January - September), total mean mummichog densities ranged from
15,077.8 at the CTL Site compared 1*6,370 at the TRT Site. Mummichog densities
at the TRT Site during 1990 were only 42% of CTL Site populations. Statistical
analysis indicated that CTL Site densities were significantly (p £ 0.05 - 0.10) higher
than TRT Site densities during February, March, May, July, August and September,
1990. During January, 1990, mummichog densities at the TRT Site were significantly
(p < 0.10) higher than CTL Site levels. Significant fenvalerate runoff during late May,
1990 at the TRT Site caused no toxicity in caged mummichogs as levels were below
concentrations acutely toxic to adult mummichogs. Although mummichog populations
were significantly lower at the TRT Site, two days post runoff, it is not likely that these
differences were directly attributable to fenvalerate runoff. Rather these differences
were probably related to earlier pesticide runoff effects at the TRT Site (1985-88).
Results from 1988-89, indicated that despite significant reductions in pesticide
runoff at the TRT Site for 1989-9d, mummichog densities remained significantly lower
than at the CTL Site. These data are suggestive that mummichog populations at the
TRT Site were slower to recover than other species (i.e. P. pugio).
213
-------
4_. Total Fish Density
Results of total fish density measurements (#/50 m of stream) are listed in Table
47 and depicted in Figure 59. Mean total fish densities ranged from 26.3 -4,723.9/50
mg of stream at the CTL Site compared to levels ranging from 108.3 - 3,787.0/50 m
of stream at the TRT Site. During 1989, peak total fish densities (> 1500/50 m of
stream) were observed at the CTL Site during March - early June, August - October,
and in December, 1989. At the TRT Site, peak total fish densities were observed in
early June, July and August, 1989. These peak periods of total fish density generally
coincided with periods of time when juvenile mummichog young-of-the-year entered
size cohorts measurable by our sampling methods. Mummichogs were the dominant
fish species observed accounting for 72.7% of total fish density at CTL Site and 65.1 %
of total fish abundance at the TRT Site during 1989.
During 1990, peak abundances at the CTL Site were observed during March,
June, July, August and September, 1990. At the TRT Site during 1990, peak densities
were observed during June and July, 1990. These periods of peak densities observed
during 1990, generally coincided with peaks of mummichogs density, especially
intervals when juvenile mummichog yo&ng-of-the-year entered size cohorts measurable
by our sampling methods. In 1990, mummichogs were the dominant fish species
observed accounting for 70.2% of total fish density at the CTL Site and 57.1 % of total
fish abundance at the TRT Site.
Total mean total fish density for the 21 months sampled during 1989-90, was
44,274.5/50 m of stream at the CTL Site and 25,213.3/50 m of stream at the TRT Site.
Previous studies during 1986-88 (Scott et a/., 1990; Hampton, 1987), reported annual
total fish densities ranging from 24,314.3 - 39,060/50 m of stream at the CTL Site and
from 12,255.7 - 17,505/50 ra of stream at the TRT Site. During 1986-87, significant
runoff of endosulfari, azinphosmethyl, and fenvalerate at the TRT Site resulted in
several fish kills which reduced annual total fish populations 10 densities only 44.8 -
50.4% of CTL Site total fish densities.
\
214
-------
Table 47. Summary of Total Fish Density measurements (number/50 m of stream)
observed" in block seining at the CTL and TRT Sites, 1989-90. Asterisks
(";") indicate when samples were significantly (p < 0.05 - 0.179) different.
Parameter: Total Fish Density (#/50 m of stream)
Date
1/89
2/89
3/89
4/89
5/89
6/7/89
6/26/89
7/89
8/89
9/89
10/89
11/89
12/89
1/90
2/90
3/90
4/90
5/11/90
5/30/90
6/90
7/90
8/90
9/90
Total
Density
CTL Site
--x
538.5
1.207.3
2,107.7
2,422.3
1,989.3
1,913.7
1,132.3
1,295.7
2,716.0
4,284.0
2,400.3
611.3
165.7
26.3
955.7
2,632.3
1,045.3
848.6
1,149.7
1,741.1
4,723.9
4,639.2
3,728.5
44,274.5
SE
291.3
421.6
775.7
584.0
384.9
1.618.8
349.8
622.8
1,618.3
1,668.5
1,027.9
105.5
67.2
13.2
327,0
1,208.3
348.7
338.7
380.4
780.4
1,094.7 ^
381.4
1,277.3
TRT Site
X
108.3
431.7'
726.3'
700. 0"'
449.7";<
1,562.0
605.0
2,446.7'
3,787.0
1,109.0*
1,087.0"
682.1
' 360.2
191.7'"
711.0
978.1
813.7
822.1
898.3
2,164.7
2,242.7'
1,109.4"A;'A
1,226.6
25,213.3
SE
52.1
' 170.7
296.9
206.5
227.7
325.5
108.1
1,067.3
1,199.0
498.5
535.6
138.7
118.5
71.0
255.9
384.8
187.6
440.7
304.6
782.6
974.9
•6.9
490.4
Significantly (p <. 0.05) Different in Unpaired Test; N=6
Significantly (p < 0.10) Different in Paired Test; N=6
Significantly (p < 0.08) Different in Unpaired Test; N=5
Significantly (p <. 0.179) Different in Paired Test; N = 5 .
Hugo
215
-------
Total Fish Density 1989-1990 (Seine)
E
to
o
£
e
in
C\l T- cvj pj CM
OJ
-.
-------
During 1989. total fish densities at the CTL Site were 22,783/50 m of stream
compared to 14,055/50 m of stream at the TRT Site. Total fish densities at the TRT
Site during 1989 were only 57% of the CTL Site populations. Statistical analysis
indicated trrat population densities were significantly (p < 0.05 - 0.10) higher at the
CTL Site during February - May, September and October, 1989. During July, 1989
total fish densities at the TRT Site were significantly (p < 0.10) higher than CTL Site
levels. This same pattern was observed in mummichogs during 1989.
During 1990 (January - September), total fish densities ranged from 21,490.6/50
m of stream at the CTL Site compared to 11,158.3/50 m of stream at the TRT Site.
Total fish densities at the TRT Site were only 52% of CTL Site populations. Statistical
analysis indicated that CTL Site densities were significantly (p < 0.05 - 0.10) higher
than TRT Site total fish densities during January, July and August, 1990.
Significant fenvalerate runoff during June, 1989 and May, 1990 appeared to have
little effect, as total fish density comparisons were not significantly different in both
intra- and inter-site comparisons
5. Penaied Shrimp Densities v
Results of penaied shrimp [Penaeus aztecus (brown), P. duorurum (pink), and P.
setiferus (white)] density (#/50 m of stream) measurements and statistical comparisons
at each site are listed in Table 48 and depicted in Figure 60. It was noted that juvenile
brown shrimp first migrated into both branches of Leadenwah Creek during May, 1989
and generally remained at each site until November (TRT Site) or December (CTL
Site). Juvenile white and pink shrimp first appeared in mid-June, 1989 and also
remained until November, 1989. In other months of the year, penaied shrimp were not
detected. A similar pattern of penaied shrimp migration was observed in 1990.
Mean penaied shrimp densities during 1989 ranged from 0.7 - 5,580.7/50 m of
stream at the CTL Site and from 82.6 - 8,499.7/50 m of stream at the TRT Site.
During 1990 (January - September), penaied shrimp densities ranged from
217
-------
Table 48. Summary of Penaeus species density measurements (number/50 m of stream)
observed in block seining at the CTL and TRT Sites, 1989-90. Asterisks
(":") indicate when samples were significantly (p < 0.05 - 0.08) different.
Parameter: Penaeus Species Density (#/50 m of stream)
Date
1/89
2/89
3/89
4/89
5/89
6/7/89
6/26/89
7/89
8/89
9/89
10/89
11/89
12/89
1/90
2/90
3/90
4/90
5/11/90
5/30/90
6/90
7/90
8/90
9/90
Total
CTL Site
X
0.0
0.0
0.0
0.0
330.0
285.3
5,580.7
590.0
800.3
1,290.7
61.7
0.7
1.0
0.0
0.0
0.0
0.0
50.5
48.6
117.1
145.3
21.0
26.7
10,403.6
SE
0.0
0.0
0.0
0.0
169.9
128.1
426.6
405.5
264.7
320.4
11.3
0.7
1.0
0.0
0.0
0.0
0.0
18.1
16.1
16.4
: 84.6
5.3
13.4
TRT Site
X
0.0
0.0
0.0
0,0
217.3
213.0
8,499.7
2,280.3
2,357.7
227.7"
138.7
82.6"
0.0
0.0
0.0
0.0
0.0
96.9
102.8
361.7
: 1,465.0
148. 7"A
43.3
15,235.4
SE
0.0
0.0
0.0
0.0
110.8
94.7
3,388.2
1,659.3
518.9
137.2
65.7
56.6
0.0
0.0
0.0
0.0
0.0
50.7
22.2
232.7
677.1
18.4
8.4
Significantly (p ^
Significantly (p £
0.05) different
0.08) different
in Unpaired Test; N=6.
in Unpaired Test; N=5.
-Hugo
218
-------
Penaied Density 1989-1990 (Seine)
a
«3
2 ""
UJ
0.
E
a
in
Control
H— Treaiment
0j 0) 0i 0i 0) o) o) o) o) o> d
DATE
Figure 60. Penaied shrimp (Penaeus aztecus, Penaeus duorarum, and Penaeus setiferus) densities
(#/50m of stream) measured in block seining during 1989-90. Asterisks (*) indicated
samples which were significantly (p < 0.05) different in statistical comparisons between
the CTL and TRT Sites.
219
-------
21.0 - 117.1/50 m of stream at the CTL Site and from 43.3 - 1,465/50 m of stream at
the TRT Site.
Peak shrimp densities during 1989 were observed during late June and September
at the CTL Site and during late June, July and August .at the TRT Site. Statistical
analysis indicated penaied shiimp densities were significantly (p < 0.05) higher at the
CTL Site compared to the TRT Site, during September, 1989. During November, 1989
penaied shrimp densities at the TRT Site were significantly (p < 0.05) higher than at
the CTL Site. Between site density differences observed during 1989 were noted and
related in part to significant fenvalerate runoff observed at the TRT Site during June,
1989. Differences in recreational fishing pressure between the CTL (higher) and TRT
(lower) sites probably account in part for some of the observed between site differences.
During 1990, peak penaied shrimp densities were observed during June and July
at both the CTL and TRT Sites. Statistical analysis indicated that penaied shrimp
densities were significantly (p < 0.08) higher at the TRT Site during August, 1990,
when compared to the CTL Site. Between site density differences observed during 1990
were not related to significant fenvalerate runoff observed at the TRT Site during May,
1990. Differences in recreational fishing pressures between the two sites may in part
account for most observed between site differences.
A total of 10,403.6 penaied shrimp/50 m of stream at the CTL Site versus
15,235.4/50 m of stream at the TRT Site were observed during the 21 months of this
study. Annual penaied shrimp densities ranged from 1463.2 (1990) to 8,940.4 (1989)
at the CTL Site and from 2,218.4 (1990) to 13,017 (1989) at the TRT Site. Generally
penaied shrimp densities at the CTL Site were reduced 31.3-33.4% compared to the
TRT Site. These results compared favorably with earlier results (Scott et al., 1990;
Hampton, 1987) which reported penaied shrimp densities during 1986-88 ranging from
2,000 - 11,605/50 m of stream at the CTL and TRT Sites. During 1986-88, penaied
shrimp densities at the CTL Site were reduced by 43-83% compared to the TRT Site,
primarily due to higher recreational fishing pressure at the CTL Site.
220
-------
(L Blue Crab Densities
Results of blue crab (Callinectes sapidus) density (#/50 m of stream)
measurements and statistical comparisons at each site are listed in Table 49 and depicted
in Figure 61. Generally highest blue crab densities were observed during the spring -
early summer months and lowest densities were observed the late fall - early winter
months. This pattern was observed at both sites during 1989 and 1990.
During the study period (1/89 - 9/90) mean blue crab densities ranged from 0.7 -
37.0/50 m of stream at the CTL Site and from 0.0 - 31.3/50 m of stream at the TRT
Site. During 1989, peak blue crab densities at the CTL Site were observed during
April, May, June and July, 1989. At the TRT Site, peak densities were observed
during May, July and August, 1989. Statistical analysis indicated significantly (p <
0.05 - 0.10) higher densities at the CTL Site during March, April, June, September and
December, 1989. Blue crab densities were significantly (p < 0.10) higher at the TRT
Site during August and October, 1989.
During 1990, peak blue crab densities were observed during February, March,
April, May and June at the CTL Site-and during May, June and July at the TRT Site.
Statistical analysis indicated significantly (p ^ 0.05 - 0.10) higher blue crab densities
at the CTL Site during February, April and May, 1990. At the TRT Site, blue crab
densities were significantly (p ^ 0.05 - 0.10) higher during July - August, 1990.
Total mean blue crab densities of 275.5/50 m of stream at the CTL Site and
188.5/50 m of stream at the TRT Site were measured during the 21 months of this
study (1/89 - 9/90). Annual blue crab densities ranged from 118 (1990) - 157.5
(1989)/50 m of stream at the CTL Site and from 87.6 (1989) - 100.9 (1990)/ 50 m of
stream at the TRT Site. Generally, annual blue crab densities were reduced by 14.5 -
44.4% at the TRT Site compared to'the CTL Site during 1989-90. These results agree
favorably with earlier results (Scott et al., 1990; Hampton, 1987) which reported that
from 1986-88 blue crab densities, ranged from 111 - 138.6/50 m of stream at the CTL
Site and from 55 - 61.6/50 m ojfistrearn at the TRT Site. From 1986-88, blue crab
densities were 50-55.4% lower at the TRT Site compared to
221
-------
Table 49. Summary of Callinectes sapidus Density measurements (number/50 m of stream)
observed in block seining at the CTL and TRT Sites, 1989-90. Asterisks (" " M)
indicated when samples were significantly (p < 0.05 - 0.10) different.
Parameter: Callinectes sapidus Density (#/50 m of stream)
Date
1/89
2/89
3/89
4/89
5/89
6/7/89
6/26/89
7/89
8/89
9/89
10/89
11/89
12/89
1/90
2/90
3/90
• 4/90
5/11/90
5/30/90
6/90
7/90
8/90
9/90
Total
CTL Site
^X
3.3
9.7
8.3
25.0
37.0
14.7
19.7
13.3
5.7
7.7
2.7
0.7
9.7
2.3
22.0
11.0
17.0
25.7
12.7
9.3
6.3
6.7
5.0
275.5
SE
1.5
6.9
3.8
6.1
22.5
0.9
2.2
8.3
3.3
0.7
2.2
0.7
3.3 T
1.2
9.9
5.1
4.0
9.2
3.3
1.9
0.3 -.
1.5
0.6
i"
TRT Site
X
1.3
0.0
1.7'"
5.7'"
12.7
5.3-"
5.3*"
11.0
31.3'
4.0";>
5.3'
2.3
1.7
3.0
0.3"'
4.7
9.0'
26.7
8.7'
13.0
20.3'"
8.5'A
6.7
188.5
SE
1.3
0.0
0.9
2.2
8.6
2.3
3.2
4.6
17.0
0.6
2.9
0.3
1.2
1.7
0.3
3.3
2.1
0.7
3.8
5.1
2.3
0.5
3.3
Hugo
" = Significantly (p < 0.05) Different in Unpaired Test; N=6
" = Significantly (p < 0.10) Different in Paired Test; N=6
"A = Significantly (p < 0.079) Different in Unpaired Test; N = 5
222
-------
C. sapidus Density 1989-1990 (Seine)
E
ra
2
3 —
2 M
Q.
ra .fe-
at o
o
%
O1O1OOOOOOOOOO
CM — OJ CM CM -- CM CM
— O Ol CM CM •- O
CM ~ CM ~ ~ ~ "
CO CM CM CM CM
ui to r^ CD a>
DATE
Fieure 61 Callinectes sapidus densities (#/50m of stream) measured in block seining during 1989-90.
' Blue crab^densities were quite similar at the CTL and TRT Sites during 1989-90. Asterisks
(*) indicated samples which were significantly (p < 0.05) different in statistical between
site comparison of the TRT and CTL Sites.
223
-------
[he CTL Site despite heavy recreational fishing pressures at the CTL Site. From 1986-
90, blue crab densities at the TRT Site have steadily recovered from impacts resulting
from significant runoff of azinphqsmethyl. endosulfan, and fenvalerate. By 1990,
annual blue crab densities only varied by 14.7% between the CTL and TRT Sites.
~L Discussion and Conclusions of Ecotoxicological Studies, 1989-90
During 1989-90, significant runoff of fenvalerate was observed during 5-7 June,
1989 (< 100 ng/L) and 28-29 May, 1990 (122 ng/L) at the TRT Site. Measured
fenvalerate concentrations were below levels considered acutely toxic to adult fish and
blue crabs but were considered potentially toxic to adult and juvenile grass shrimp,
juvenile penaied shrimp and possibly juvenile fish.
In situ toxicity tests conducted at the TRT Site during 1989 indicated:
1) P. pugio survival was 28.5% (± 13.8%) at the TRT Site;
2) Penaeus aztecus survival was 51.9% (+ 6.06%) at the TRT Site; and
3) No mortality (high survival) was observed in caged adult mummichogs and
juvenile sheepshead minnow.
These results clearly indicated acute toxicity in caged grass shrimp and penaied
shrimp following this rain event. Analysis of ecotoxicological data for P. pugio during
May - June, 1989 indicated between site mortality of 34% compared to within site
estimated mortality rates of 69%. These results agree favorably with the 72% mortality
observed in field toxicity tests. Laboratory toxicity tests for fenvalerate predicted
similar levels of toxiciry in P. pugio (Scott et al, 1990). Additionally if one examines
P. pugio recruitment 90 days post rainfall (September, 1989) and extrapolates the
number of grass shrimp recruited; into the population during the 90 day time period at
both sites relative to the standing stock of adults at the time of the rain event, a
mortality estimate of 75.2% is obtained. This estimate assumes equal growth and
predation rates at both sites. ',
Analysis of 1989 penaied shrimp ecotoxicological data indicated that no significant
toxicity in penaied shrimp was observed at the TRT Site, using both within and between
site comparisons. These results may suggest in part, that juvenile penaied shrimp may
224
-------
actively avoid pesticide runoff following periods of heavy rainfall unlike grass shrimp
which appeared to be adversely affected.
During 1990, results of in situ toxicity tests indicated:
1) High survival in P. pugio at both sites;
2) High survival in P. aztecus at both sites: and
3) High survival in adult mummichogs and juvenile sheepshead minnows at
both sites.
Analysis of 1990 ecotoxicological data indicated that grass shrimp and penaied
shrimp populations at the TRT Site were unaffected by fenvalerate runoff. While grab
samples collected at dead low tide indicated potentially toxic levels of fenvalerate (122
ng/L) composite samples collected during this same time period, indicated nondetectable
fenvalerate levels (< 3 ng/L). These findings suggest that although significant
fenvalerate runoff occurred, it may have only been a small volume of runoff which was
diluted quickly with the incoming flood tide. As a result, no mortality was observed
in field populations and caged bioassay organisms. The retention ponds at the TRT Site
may have served to reduce overall runoff volume; hence preventing field mortality and
large transboundary pesticide movement.
B. Push Netting, 1990
l± Total Biomass
Results of total biomass (g/50 m of stream) estimates from push netting are listed
in Table 50 and depicted graphically in Figures 62 (CTL vs TRT Sites) - 63 (CTL vs
KWA Sites). Mean monthly biomass ranged from 2.4 - 80.7 g/50 m of
225
-------
Table 50. Summary of Total Biomass (grams/50 m of stream) measured in push
net sampling during the 1990 field study. Asterisks (") indicate
when samples were significantly different from the CTL Site.
Parameter: Total Biomass (g/50 m of stream)
Date
3/26/90
4/20/90
6/4/90
6/20/90
7/20/90
8/22/90
10/13/90
11/20/90
12/13/90
Total
Biomass
(3/26-I2/I3/90)
CTL
X
11.9
27.1
70.6
80.7
63.9
42.7
31.7
11.2
2.4
342.2
(SE)
9.2
18.0
57.4
47.0
30.3
17.4
2,9
5.5
1.3
TRT
X
13.2
90.1
72.3
170.2
56.5
43.8
"ro.2'
3.2'
9.9
469.4
(SE)
3.5
39.7
41.1
96.6
23.7
10.1
7.3
1.3
7.1
KVVA
X
12.0
41.2
27.7
51.6
50.8
27.3
63.7'
24.4
2.6
301.3
(SE)
7.8
21.0
3.4
20.5
15.2
11.2
8.5
16.7
2.2
* = Significantly (p < 0.05) different from ControlsTable 50
226
-------
£
a
-------
Total Biomass 1990 (Push Net)
300
DATE
Figure 63. Total biomass (g/50m of stream) measured in push netting at the CTL and KWA Sites
during 1990. Generally total biomasses were quite similar at both sites during much of
1990. Asterisk (*) indicated samples which were significantly (p < 0.05) different in
statistical comparisons between the GTL and KWA Sites.
228
-------
Total mean biomass for March - December, 1990 was 342.2 g/50 m of stream at the
CTL Site, 464g/50 m of stream at the TRT Site, and 301.3 g/50 m of stream at the
KWA Site ."Statistical analysis indicated:
1) Significantly (p < 0.05) higher biomass at the CTL Site compared to the
TRT Site during October and November, 1990; and
2) Significantly (p < 0.05) higher biomass at the KWA Site compared to both
the CTL and TRT Sites during. October, 1990.
Peak biomass was observed during June-July at the CTL Site, April - June at the
TRT Site, and June, July and October at the KWA Site. These periods of peak biomass
at each site- generally coincided with peaks in P, pugio biomass which accounted for
78.4 - 83.2% of total biomass. Other species, including Penaeus aztecus, Penaeus
setiferus, Penaeus duorarum, F. heteroclitus, Mugil Cephalus, Poecilia latipinna, C.
variegcuus, M. menidia, A. mitchilli, and Callinectes sapidus, accounted for the other
16.8 - 21.6% of the total biomass.
2. Total Density
Results of total density (#/50 m of stream) estimates from push netting are listed
in Table 51 and depicted in Figures 64 (CTL vs. TRT Sites) and 65 (CTL vs KWA
Sites). Mean monthly total density ranged from 22.7 - 531.7/50 m of stream at the
CTL Site, 15 - 724.7/50 m of stream at the TRT Site, and 21.7 - 468.0/50 m of stream
at the KWA Site. Total mean densities for March - December, 1990 were 1,947.7/50
m of stream, at the CTL Site, 2,003.1/50 m of stream at the TRT Site, and 1743/50 m
of stream at the KWA Site. Statistical analysis indicated:
1) Significantly (p ^ 0.05) higher faunal densities at CTL Site compared to
the TRT Site during October and November, 1990; and
>
2) Significantly (p <£ 0.05) higher faunal densities at the KWA Site compared
to both the CTL and TRT Sites during October, 1990.
229
-------
Table 51. Summary of Total Faunal Density (number/50 m of stream) measured in push net
sampling during the 1990 field study. Asterisks (") indicate when samples were
significantly (p < 0.05) different from the CTL Site.
Parameter: Total Faunal Density (#/SO m of stream)
Date
3/26/90
4/20/90
6/4/90
6/20/90
7/20/90
8/22/90
10/13/90
11/20/90
12/13/90
Total Density
(3/26-12/13/90)
CTL
X
57.3
22.7
470.7
531.7
371.3
215.7
192.0
63.0
23.3
1,947.7
(SE)
47.4
16.9
390.8
289.7
167.6
113.2
25.3
31.0
14.0
TRT
X
44.0
194.7
355.7
724.7
316.0
239.0
3i.r
15.0'
82.7
2,003.1
(SE)
18.2
94.5
165.6
412.9
139.2
48.0
23.1
5.5
61.4
KWA
X
30.3
123.7
154.7
253.3
152.0
199.7
468.0'
144.3
21.7
1,743.0
(SE)
15.7
83.6
18.2
82.2
76.5
94.9
50.4
101.7
16.2
= Significantly (p ^ 0.05) different from Controls.
230
-------
Total Density 1990 (Push Net)
1200
DATE
Figure 64. Total densities (if/50m of Stream) measured in push netting at the CTL and TRT
Sites during 1990. Generally, total densities were quite similar at both sites
during much of 1990. Asterisks (*) indicated samples which were significantly
(p < 0.05) different in between site statistical comparisons of the CTL and TRT
Sites.
231
-------
Total Density 1990 (Push Net)
-------
Peak total faunal densities were observed during June and July at the TRT and
CTL Sites and during June and October at the KWA Site. These periods of peak faunal
densities generally coincided with peaks in P. pugio density which accounted for 83.2 -
96.4% of=all total faunal densities. At CTL and TRT Sites, P. pugio accounted for
95.2 and 96.4%, respectively of the total faunal density. At the KWA Site, P. pugio
accounted for only 83.2% of total faunal densities. Species other than .P. pugio
accounted for 3.6, 4.8 and 16.8% of the total faunal densities, respectively at the TRT.
CTL and KWA Sites.
3.. P. pugio Density
Results of P. pugio abundance or density (#/50 m of stream) estimates from push
netting are listed in Table 52 and depicted in Figures 66 (CTL vs TRT Sites) and 67
(CTL vs KWA Sites). Mean monthly P. pugio densities ranged from 15 - 520.7/50 m
of stream at the CTL Site, 14 - 705.7/50 m of stream at the TRT Site, and 19 -
462.3/50 m of stream at the KWA Site. Total mean densities for March -December,
1990 were 1,854.4/50 m of stream at the CTL Site, 1931.1/50 m of stream at the TRT
Site, and 1449.6/50 m of stream at the KWA Site. Statistical analysis indicated:
v
1) Significantly (p ^ 0.05) higher P. pugio densities at the CTL Site
compared to the TRT Site during October and November, 1990; and
2) Significantly (p < 0.05) higher P. pugio densities at the KWA Site
compared to the TRT Site during October, 1990.
Peak P. pugio densities were observed during June - July, 1990 at the CTL and
TRT Sites while peak densities at the KWA Site were observed during June and
October, 1990.
Earlier studies (Welch, 1975) reported P. pugio densities ranging from 20-300/m2
using push netting in small tidal preeks in the Gulf Coast. Converting measured P.
pugio densities/50 m of stream iJtto densities/m? (using a conversion factor of 30 x
measured density/50 m of stream) results in estimated P. pugio densities ranging from
0.5 - 15/m2 at the CTL Site, 0.5 - 23.6/M2 at the TRT Site,
233
-------
Table 52. Summary of P. pugio Biomass (grams/50 m of stream) measured in push net
sampling during the 1990 field study. Asterisks (") indicate when samples
were significantly (p < 0.05) different from the CTL Site.
Parameter: P. pugio Biomass (grams/50 m of stream)
Date
3/26/90
4/20/90
6/4/90
6/20/90
7/20/90
8/22/90
10/13/90
1 1/20/90
12/13/90
Total P pugio Biomass
(3/26-12/13/90)
CTL
X
11.0
3.4
61.8
76.5
52.9
35.1
30.5
11.2
2.2.
284.6
(SE)
9.7
3.3
53.5
46.4
32.6
18.7
3.5
5.5
1.4
TRT
X
8.8
40.4
56.7
154.1
50.2
39.4
*- 7.3'
1.9'
9.1
367.9
(SE)
4.1
16.9
29.9
96.8
22.2
8.8
4.4
0.8
6.6
KWA
X
6.2
16.9
17.9
34.7
18.1
24.7
59. T
24.2
2.1
238.6
(SE)
2.9
12.8
4.2
9.8
8.9
12.1
5.1
16.8
1.7
= Significantly (p ^ 0.05) different from Controls.
234
-------
P. pugio Density 1990 (Push Net)
1200
DATE
Figure 66. P. pugio densities (#/50m of stream) measured in push netting at the CTL
and TRT Sites during 19901 Generally, grass shrimp densities were quite
similar at both sites during much of 1990. Asterisks (*) indicated samples
which were significantly (p < 0.05) different statistical comparisons.
235
-------
P. pugio Density 1990 (Push Net)
DATE
Figure 67. P. pugio density (#/50m of stream) measured in push netting at the CTL
and KWA Sites during 199CL Generally, grass shrimp densities were quite
similar at both sites during much of 1990. Asterisks (*) indicated samples
which were significantly (p < 0.05) different in between site statistical
comparisons.
236
-------
and 0.7 - 15.4/m2 at the KWA Site. On an annual basis, P. pugio densities would be
82.3/m2 at the CTL Site, 85.1/m2 at the TRT Site, and 64.2/m2 at the KWA Site. Results
from this study in mesotidal, (1 - < 3m tidal range) creeks are in agreement with chose
reported forfnicrotidal (< 1m) creeks in terms of P. pugio densities.
4^ P. pugio Biomass
Results of P. pugio biomass (g/50 m of stream) estimates from push netting are listed
in Table 53 and depicted in Figures 68 (CTL vs TRT Sites) and 69 (CTL vs KWA Sites).
Mean monthly P. pugio biomass ranged from 2.2 - 76.5g/50 m of stream at the CTL Site,
1.9 - 154.1 g/50 m of stream at the TRT Site, 2.1 - 59.1 g/50 m of stream at the KWA
Site. Total mean P. pugio biomass for March - December, 1990 was 284.6 g/50 m of
stream at the CTL Site, 367.9 g/50 m of stream at the TRT Site, and 238.6 g/50 m of
stream at the KWA Site. Statistical analysis indicated:
1) Significantly (p <• 0.05) higher P. pugio biomass at the CTL Site compared
to the TRT Site during October and November, 1990; and
2) Significantly (p < 0.05) higher P. pugio biomass at the KWA Site compared
to the CTL and TRT Sites during October, 1990.
Peak P. pugio biomass was observed during June - July, 1990 at the both the CTL
and TRT Sites while peak P. pugio biomass at the KWA Site was observed during June
and November, 1990.
5. Discussion; Comparisons of Estimated P. pugio Densities Using Block
Seining and Push Netting Methodologies.
Tables 54 - 55 list results comparing estimated P. pugio densities using block seining
and push netting methodologies at the CTL and TRT Sites from March - December, 1990.
At the CTL Site, P. pugio densities^ranged from 900 - 31,242/50 m of stream using push
netting compared to densities ranging from 24 - 9,136/50 m of stream using block seining
(Table 54). At the TRT Site, P. pugio densities ranged 1,842 - 42,342/50 m of stream
using push netting compared to densities ranging from 829 - 7,426/50 m of stream using
block seining (Table 54). These results suggest that
237
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Table S3. Summary of P. pugio Density (number/50 m of stream) measured in push net
sampling during the 1990 field study. Asterisks (*,*A) indicated when samples were
significantly (p < 0.05) different.
Parameter: P. pugio Density (if/50 m of stream)
Date
3/26/90
4/20/90
6/4/90
6/20/90
7/20/90
8/22/90
10/13/90
11/20/90
12/13/90
Total Density
(3/26-12/13/90)
CTL
X
55.0
15.0
456.7
520.7
328.7
204.3
188.0
63.0
23.0
1,854.4
(SE)
48.5
14.0
392.7
293.8
178.6
116.6
26.3
31.0
14.2
TRT
X
39.0
187.3
344.7
705.7
300.0
229.0
, 30.7*
14.0'
80.7
1,931.1
(SE)
16.9
93.5
159.4
409.2
133.3
45.7
22.5
5.5
60.4
KWA
X
19.0
83.0
140.3
245.7
141.0
194.0
462. 5 "A
144.0
20.3
1,449.6
(SE)
6.1
70.6
24.3
80.5
76.3
95,8
46.7
102.0
14.9
* = Significantly (p £ 0.05) different from the CTL Site
*A = Significantly (p <. 0.05) different from the TRT Site but not CTL Site.
238
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P. pugio Biomass 1990 (Push Net)
300
DATE
Figure 68. P. pugio biomass (g/50m of ^tream) measured in push netting at the CTL and
TRT Sites during 1990. Generally, grass shrimp biomass was quite similar at
both sites during much of 1990. Asterisks (*) indicated samples which were
significantly (p < 0,05) different in statistical between site comparisons.
239
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P. pugio Biomass 1990 (Push Net)
Figure 69. P. pugio biomass (g/50m of stream) measured in push netting at the CTL and
KWA Sites during 1990. Generally, grass shrimp biomass was quite similar at
both sites during much of 1990. Asterisks (*) indicated the one sample that
was significantly (p < 0.05) different in between site comparisons.
240
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Table 54. Summary of P. pugio Density (#/50 m of stream) measured in push net sampling
during the 1990 field study.
Parameter: P. pugio Density (#/50 m of stream)
Date
3/90
4/90
6/4/90
6/30/90
7/90
8/90
9-10/90
CTL
Push Net1
X (SE)
3,300(± 2,910)
900 (± 840)
27,402 (± 23,562)
31,242 (± 17,628)
19,722 (± 10,716)
12,258 (± 6,996)
1 1,280 (± 1,578)
Seine
X (SE)
9,136(± 9,074)
24 (±5)
461 (± 271)
3,896(± 1717)
6,569 (± 2.967)
1,900 (± 882) "
5,965 (± 3,026)
TRT
Push Net1
X (SE)
2,340(± 1,014)
1 1,238 (± 5,610)
20,682 (± 9,564)
42,342 (± 24,552)
18,000 (± 7,998)
13,740 (± 2,742)
1,842(± 1350)
Seine
X (SE)
3,309(± 1,555)
7.426 (± 6,162)
829 (± 161)
1,272(± 243)
6.004 (± 2,751)
1,906(± 120)
2,591 (± 496)
1 = Conversion Factor of 60 used based on gear size differences between push net and seine net gear.
241
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push netting P. pugio densities do not directly correspond numerically to absolute
densities obtained using block seining; however, relative abundance comparisons are
possible.
In statistical comparisons of P. pugio density differences between the CTL and
TRT Sites using block seining and push netting, 86% of the conclusions reached in
statistical tests were the same using both methods (Table 55). In 7% of the
comparisons, block seining was more sensitive in detecting statistically significant
differences than push netting (p < 0.05 seine vs. p < 0.10 push netting). In 14% of
the comparisons, push netting detected significant differences when block seining did
not. When this occurred mean P. pugio densities were numerically higher by 56.5%
using block seining at the CTL Site, but were not statistically different due to skewness
among replicates (i.e., most of the grass shrimp were congregated in one stream stretch
rather than being randomly distributed). An error rate of 7-14% may be expected using
the push net method. When one considers the added cost, time and effort required for
block seining, push netting provides a reasonable sampling alternative, particularly in
situations requiring rapid and large scale sampling such as oil/hazardous substance spills
and for large regional-scale sampling efforts (i.e., E-MAP).
1989-90 Discussion and Conclusions
A. Correlating Laboratory and Field Toxicitv Test Results with Field
Ecotoxicological Biomonitoring
Earlier studies by Scott et al. (1990) have reported that the integration of field
laboratory toxicity testing with ecotoxicological and ecophysiological biomonitoring
provides a holistic method of environmental risk assessment for pesticides. Similarly,
Swartz et al. (1986) and others (Chapman et al. 1983, 1984; Olla et al. 1984) have
defined a Triad of toxicity tests, utilizing a combination of field assessments and
laboratory toxicity testing to accucately define sediment toxicity and develop sediment
quality criteria. The approach used in these methods involves: 1) Standard laboratory
toxicity testing to define initial toxicity benchmarks and a battery of nonconventional
toxicity tests to define an array of toxicant-ecological (abiotic-physicochemical and
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Table 55. Summary of statistical results for P. pugio densities using seining and
push netting during 1990. Note the excellent agreement between the two
sampling methods.
Date
3/90
4/90
6/4/90
6/30/90
7/90
8/90
9-10/90
Seine
c=r
T>C (p < 0.05 unpaired)
T>C (p < 0.10 paired)
C=T
C = T
C = T
C=T
C=T
(x = 5965 vs 2591)
Push Net
C=T
T>C (unpaired p < 0.12)
T>C (paired p < 0.10)
C = T
C = T
C=T
C=T
C>T (p < 0.05 unpaired)
OT(p <, 0.10 paired)
1 = All C = T are for paired and unpaired tests
C = Control Site; T = Treatment Site
I. Number of Sampling Date Comparisons = 14 (7 paired; 7 unpaired)
II. Number of Times of Agreement = 12 = 86%
III. Number of Time Seining was more sensitive than Push Net
(False Negative) = 1=7%
IV. Number of Times Push Netting was more sensitive than seining
(False Positive) = 2 = 14%
243
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biotic-species, sex, lifestage differences) interactions; 2) A battery of field toxicity tests
to define potential field effects; and 3) Biomonitoring to address field population effects
and confirm field and laboratory toxicity test results. The approach used in this present
study utilized a similar approach but added a series of ecophysiological studies to assess
sublethal effects in field populations using both specific and nonspecific physiological
parameters. The goals of these approaches are to develop protocols for establishing
laboratory toxic tests which accurately predict field effects and to establish a paradigm
for field validations in assessing both acute toxicity and acute/chronic sublethal effects.
Acute, laboratory toxicity tests provide the initial benchmark for the environment
risk assessment process in determining pesticide safety. Most laboratory toxicity tests
are designed to expose an organism to a number of sequential concentrations, over a
defined (usually 96h) continuous exposure period. The results of such tests provide
information on the no effect concentration, the lowest concentration causing 100%
mortality, and the LCjo concentration. Additionally, these laboratory tests may provide
identification of toxic threshold concentrations. Extrapolation of environmentally safe
concentrations for a compound is possible if a number of different animal species (fish
and invertebrates) are tested.
Acute toxicity testing in previous studies (Chandler and Scott, 1990; Scott et at.,
1990: Chandler, 1989; Fulton, 1989; Williams, 1989; Moore, 1988; Baughman, 1986;
and Trim, 1986) have focused on the acute toxicity of azinphosmethyl, acephate,
endosulfan and fenvalerate on the grass shrimp (P. pugio) and the mummichog (F.
heteroclitus). Toxicity tests in these earlier studies were designed to:
1) Differentiate between the toxicity of EC and TG pesticides formulations
(this is important since most conventional laboratory tests are conducted
with TG material whereas field exposures are to various formulations).
2) Differentiate between different life history stage.sensitivities. (This is
important in both the selection of the most sensitive test species in field
toxicity tests and in the prediction of impacts in field populations.)
3) Differentiate between acute toxicity in continuous and intermittent, pulsed
exposures (field toxicity testing in semidiurnal mesotidal estuaries has
indicated that a 6h pulsed dose exposure is representative of most field
244
-------
exposures. Clark et al. (1987) have used 12h pulsed exposures in
simulating fenthion toxicity in diurnal, microtidal environments).
4) Determine the inceractive effects of low salinity conditions, which
accompany pesticide exposure during runoff events (field toxicity testing has
identified concomitant low salinity conditions generally accompany pesticide
exposure during runoff events in small estuarine tidal creeks).
5) Determine the joint or additive toxicity potential of pesticide mixtures
present in nonpoint source agricultural runoff. (Field toxicity testing has
identified the presence of endosulfan/fenvalerate and
azinphosmethyl/fenvalerate mixtures.)
6) Evaluate differences in toxicity between 6h pulsed dose and 96h continuous,
dose exposures, by evaluating the entire dose-response curve. (Hazard
Analysis)
7) Evaluate differences in pesticide toxicicy at high and low salinities with
continuous (96) and pulsed-(6h) does exposures, by evaluating the entire
dose-response curve. (Hazard Analysis)
8) Evaluate the effects of intrinsic factors (body length-size) on pesticide
toxicity.
9) Determine the sublethal (respiration, nitrogen excretion, and O/N ratios)
effects of fenvalerate exposure at high and low salinities.
10) Determine specific enzyme (AChE) responses to organophosphorus
insecticide (azinphosmethyl) exposure.
11) Determine the lethal (acute toxicity) and sublethal (egg production and %
hatch=fecundity) effects of sediment-bound fenvalerate and endosulfan in
benthic invertebrates (copepods) and polychaete larvae.
245
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Results from ecotoxicity tests in these earlier studies (Table 56) indicated:
1) Azinphosmethyl, endosulfan and fenvalerate were supertoxic (LC50 value <
10 ng/L) in tests with P. pugio.
2) Endosulfan and fenvalerate were supertoxic in tests with F. heteroclitus while
azinphosmethyl was extremely toxic (LC50 value > 10 and < 100^g/L). The
SP formulation of acephate was rated relatively harmless (LC50 value > 1 x
106 ulL) to F. heteroclitus.
3) There were no significant differences in LCM values between TG and EC
formulations of azinphosmethyl, endosulfan, and fenvalerate in tests with adult
P. pugio and F. heteroclitus.
4) Adults and zoeal P. pugio were more sensitive to endosulfan exposure than
post larval forms. Similarly juvenile F. heteroclitus were more sensitive to
endosulfan exposure than adults.
5) Adults and zoeal P. pugio were more sensitive to fenvalerate exposure than
post larvae. Adult and juvenile F. heteroclitus were equally sensitive to
fenvalerate.
6) The 6h pulsed dose LCM values for azinphosmethyl, endosulfan and fenvalerate
in P. pugio ranged from 4.31-6.24 times the 96h LCW value for these
insecticides. The 6h pulsed dose LCX values in F. heteroclitus exposed to
endosulfan and fenvalerate ranged from 5.02-6.90 times the 96h LC^ value.
The 6h Maximum Tolerated Pulse Does (MTPD) values for these three
insecticides ranged from 0.48-1.12 times the 96h LCW value in P. pugio and
from 2.18-3.45 times the 96h LCjo value in F. heteroclitus. These findings
indicate field toxicity would occur at concentrations ranging from 4-6 times the
96h LCjo values for these three insecticides.
246
-------
Table 56. Summary of 96h static renewal and 6h pulsed dose acute toxicity tests exposing adult P. pugio to azinphosmeih)I.
endosutfan and fcnvalerate; zoeal P. pugio to fenvalerate, and adult F. heteroditus to azinphosmethyl, endosulfan.
fcnvalerate and acephate at high (20 ppt) and low (5-10 ppt) salinities. Note that the most any intrinsic or extrinsic
factor affected acute toxcity (when compared to 96h LC50 at 20 ppt at 20°C) was <. a factor of 2.86. (From Scott
ct al., 1990).
Insecticide
EC Azinphosmethyl
EC Endosulfan
EC Fcnvalerate
«
SP Acephite
Test
Organisms'
P. pugio (A)
F. helerociiiia (A)
P. pugio (A)
F. hetcroclilus (A)
P. pugio (A)
P. pugio (Z)
F. hatrociitut (A) „
F fiaeroeliaa (A)
Type of
Toxicily Test
96h. SR
6h. PD
96h. SR
%h. SR
6h, PD
96h, SR
96h. SR
6h, PD
96h, SR
96k. SR
6h, PD
i
96h. SR *
Salinity
20
5
20
5
20
5
20
20
5
20
5
*20
5
20
5
20
10
20
5
20
5
20
5
9«h LCM
(95% CL) in ^g/L
1.05 (0.91-1.21)
0.97(0.77-1.24)
6.68 (5.83-7.66)
8.14(723-9.131
36.95(28.30-48.24)
28.00(20.23-38.76)
1.01 (0.72-1.43)
4.35(3.09-6.14)
3.81 (3.01-4.83)
1.45(1.32-1.59)
1.29(1.21-1.37)
0.052 (0.043-0.063)
0.060 (0.037-0.097)
0.314 (0.260-0.380)
0.235(0.106-0.522)
0.020(0.013-0.031)
0.007 (0.005-0.009)
2.86 (2.02-4.06)
1.63(1.08-2.47)
14.35(11.15-18.48)
8.55 (5.88-12.45)
26,79xtO>
(21. 61-33.21 « tO1)
35.36x10'
(29.35X10M2.58X101)
Toxicitv Ratio Valued
96h LC50 at 20 DPI = i 08
96h LCJO at 5 ppt
6h PDLCJOat 20 DDI =082
6h PDLC50 al 5 ppt
96h LC50 ai 20 DDI =1.32
96h LC50 at 5 ppt
96h LC50 at 20 ppt =163
96h LC50 11 5 ppt
6h PDLC50 at 20 pot =1.1*
6n PDLC50 at 5 ppt ]
96h LC50 at 20 opt -1.12
96h 1X50 at 5 ppt
96h LC50 at 20 DDI =0.87
96h LC50 at 5 ppt
6h PDLC50 at 20 ppi =1.34
6h PDLC50 ai 5 ppt
96h LC50 at 20 ODt =2.86
96h LC50 at 5 ppt
96h LC50 at 20 not =1.75
96h 1X50 at 5 ppi
6h PDLCSO at 20 out =1 68
6h PDLC50 at 5 ppt
96h LC50 at 20 ow =0.75
96hLC50at5 ppi
EC = Emulsifiable Concentrate
SP = Soluble Powder
1 = Test Organism: f. pugio (A)» Adult (15-25 mm); f. pugio (Z)-Zoul sage lirvte 1-2 days old: and F. kmroctitui (A)-Adulo (35-70 m)
2 = Type of Toxicity Tests; 96h. SR-96 hour static renewal and 6h. PD»Si* hour, pulled dose.
3 = Toxicity Ratio Value: 1) 96h LC50 at 20 ppt Ratio « the potency or enhancement of toikity by low salinity
96h LC50 at 5 ppc (5 or 10 ppi). insecticide exposun during 96h loxicity teso. .
2) 6h PDLCSO at 20 ppt Ratio =• The potency or enhancememof toxicity by low (5 ppt) salinity,
6h PDLCSO u 5 ppt insecticide exposure dunng 6h. pulsed doei toxicity rests
247
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7) Low salinity did not significantly affect the toxicity of azinphosmethyl or
endosulfan in continuous 96h exposures of adult P. pugio and F. heteroclitus.
In continuous exposures to fenvalerate, low salinity slightly enhanced the
toxicity of fenvalerate to adult F. heteroclitus and significantly enhanced the
toxicity of fenvalerate to zoeal P. pugio but not adult P. pugio.
8) Low salinity did not significantly enhance the acute toxicity of azinphosmethyl,
endosulfan and fenvalerate to adult P. pugio in 6h pulsed does exposures.
9) Endosulfan/fenvalerate insecticide mixtures were slightly less than additively
toxic in exposures of adult and zoeal P. pugio and adult F. heteroclitus. In
exposures of post larval P. pugio, this insecticide mixture was additively toxic.
In toto, these findings indicate that endosulfan/fenvalerate mixtures were slightly
less than additively toxic.
10) The fenvalerate/azinphosmethyl insecticide mixture was slightly less than
additively toxic to adult P. pugio.
11) The azinphosmethyl/endosulfan'mixture was slightly more than additively toxic
to F. heteroclitus at 20 ppt salinity and approached simple additive toxicity at
5 ppt salinity.
12) The acephate/fenvalerate mixture was simply additively toxic to F. heteroclitus
at 20 ppt salinity and less than additively toxic at S ppt salinity.
It is interesting to note that no intrinsic (life stage) or extrinsic (salinity, exposure
duration) factors enhanced toxicity ^ a factor of 2.86 of the 96h UCX value at 20 ppt
salinity and 20°C.
Results of earlier laboratory toxicity tests with benthic copepods exposed to sediment
bound fenvalerate (22.5-90 jug/kg) .indicated significant (p £ 0.05-0.01) reductions in
both the incidence of egg production and number of eggs produced/female for several
species (Microarthridion Morale and Paronychocamptus wilsoni) at fenvalerate
concentrations as low as 22.5 ^g/kg (Chandler, 1990). Similarly, Chandler and Scott
(1991) reported that exposure of benthic copepods and larval polychaetes to sediment-
248
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bound endosulfan resulted in significant (p < 0.05 - 0.01) effects on survival (Nannopus
palustris) and reductions in larval settlement, feeding and growth (Streblospio benedicti)
at endosulfan_concentrations ranging from 50-200 Mg/kg. Results from these laboratory
toxicity tests using field collected sediments clearly indicated that agricultural pesticide
runoff may result in significant impacts to benthic copepods and polychaetes by adversely
affecting survival, growth, settlement, feeding and reproduction.
The results of these earlier studies summarized by Scott et al., (1990) clearly
demonstrated the utility and practicality of using a battery of toxicity tests to assess acute
and chronic, lethal and sublethal effects .in aquatic and benthic marine species.
Laboratory toxicity tests were designed to not only establish acute toxicity baselines but
also to better define environmental risk assessment models by evaluating intrinsic and
extrinsic factors affecting toxicity.
Laboratory toxicity experiments conducted in this present study were designed to:
1) Refine the £€» for azinphosmethyl in mummichogs based upon brain AChE
inhibition; and
2) Evaluate the effects of azinphosmethyl exposure and resulting brain AChE inhibition
on general physiological performance in the mummichog.
Earlier studies (Scott et al., 1990; Fulton, 1989) had reported a 24 ECjo based upon
% AChE inhibition of 0.81 jig/L for mummichogs. Additional toxicity testing has
further refined this 24h ECX estimate to 0.90 pg/L. By conducting experiments using
a larger number of exposure concentrations, a more precise 24h EC*, for azinphosmethyl
AChE inhibition in mummichogs was achieved. The greater precision obtained with
these laboratory experiments will enable more accurate field and laboratory comparisons
to be made.
Results of laboratory toxicity tesjjs exposing mummichogs to a subacute, 24h dose of
azinphosmethyl and then assessing'resulting general physiological responses indicated
that:
249
-------
1) 24h exposure of mummichogs to 2.4 ^g/L azinphosmethyl resulted in 81 % brain
AChE inhibition;
2) AftePS days of depuration in clean seawater brain AChE had recovered to only
70% of controls (p < 0.05);
3) Whole animal oxygen consumption was unaffected by azinphosmethyl exposure;
4) Whole animal nitrogen excretion rates were significantly (p < 0.05) lower in
azinphosmethyl exposed mummichogs after 24h exposure; however, after 8 days
of depuration nitrogen excretion rates were not significantly different in
comparisons between treatment and control groups;
5) O/N rates were not significantly different in comparisons between control and
treatment group fish either immediately following 24h azinphosmethyl exposure
or after 8 d of depuration.
Results of these experiments clearly indicated that exposure of muminichogs to 2.4
/ig/L azinphosmethyl (6.5% of the 96 LC^) resulted in significant (p ^ 0.05) inhibition
of brain AChE (81 %). While significant brain AChE inhibition was noted, no significant
effect on respiration was observed, although nitrogen excretion was inhibited at 24h.
After 8d of depuration, however, nitrogen excretion rates were comparable to those in
control animals. Previous studies by Scott et ai., (1987) and Trim (1987) have reported
a similar phenomena in mummichogs exposed to a subacute dose of endosulfan.
Following 96h of exposure, nitrogen excretion was significantly reduced in endosulfan
exposed fish. After 7 days of depuration, nitrogen excretion rates returned to levels
comparable to control fish in a manner similar to what was observed in this study. O/N
ratios were not significantly different in comparisons between treatment and control fish,
suggesting that the rigors of azinphosmethyl exposure and resulting brain AChE
inhibition did not cause major alterations and shifts in substrate utilization (i.e., from
carbohydrate to protein or lipid to protein) and resulting whole animal physiology. The
decreased nitrogen excretion rate observed in azinphosmethyl exposed fish may either
reflect a shift in substrate utilization away from normal proportions of protein, or may
signify inhibition of normal nitrogen excretion at the gill. Previous studies with
endosulfan exposed mummichogs (Scott et a/., 1987; Trim, 1987) reported reduced
250
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nitrogen excretion may have resulted in a build up of blood ammonia levels which may
ultimately be a factor in the acute toxicity of endosulfan. Whether the decreased nitrogen
levels observed in this study signify a similar inhibition of nitrogen excretion and
resulting buildup of blood ammonia concentrations is unclear. Given the fact that the
head-kidney of the mummichog is located in the gill, it is possible that azinphosmethyl
may potentially affect the function of Na*K*Mg* pump which serves the dual function
of ion and osmotic regulation in the mummichog. Additionally, given the high levels of
brain AChE inhibition and only minor metabolic (i.e., nitrogen excretion) alterations
observed at azinphosmethyl concentrations of only a fraction (6.5%) of the 96h LQ0, is
suggestive that a fairly large reserve of brain AChE activity exists, as it relates to acute
lethality.
Results from these laboratory toxicity tests and bioassays with azinphosmethyl are
extremely important in better defining risk assessments for aquatic organisms exposed
to this pesticide. When reduced brain AChE levels are found in azinphosmethyl exposed
fish in the field, are they indicative of only azinphosmethyl exposure, or are they
indicative of both azinphosmethyl exposures and significant physiological effects? In a
more general sense, the real underlying hypothesis is: Does AChE inhibition measure
organophosphorus pesticide exposure, effects, or both? Clearly results from this srudy
indicate that for azinphosmethyl, AChE inhibition is an indicator of exposure. The
relatively minor metabolic effects observed associated with AChE inhibition in this study,
suggest additional work is needed to better explain the relationship between AChE
inhibition and other metabolic bioenergetic perturbations. Additionally, given the short
term persistence of many organophosphate insecticides in the environment and the
apparent greater persistence of AChE inhibition following exposure, suggests that AChE
inhibition may provide a reliable and moderately persistent biomarker of
organophosphorus insecticide exposure.
Field toxicity tests provide the second tier in the environmental risk assessment
process. Field toxicity tests are designed to expose an organism- to the compound being
studied at different geographical s^es. Each site will typically have a wide range of
physicochemical environmental exposure conditions. Pesticide exposure regimes at each
site will be intermittent and discontinuous. Additionally, physicochemical water quality
factors such as salinity, temperature, pH, and dissolved oxygen which are held constant
in the lab, may vary significantly during field exposure and thus may potentially enhance
251
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pesticide toxicity. Another factor to consider in assessing agricultural discharges is that
runoff may contain pesticide mixtures rather than individual compounds per se. Results
of field toxicity tests generally indicate: 1) field mortality rates at different toxicant
concentrations; 2) provide evidence for the potential interaction of physicochemical water
quality parameters: 3) evidence of the water solubility, bioavailability, persistence, and
degradation potential for various toxicants; and 4) some prediction of relative toxicity in
field populations.
Results of earlier field toxicity studies (Scott e: ai, 1990) conducted during 1985-88
at the same field sites used in the present study indicated (See Table 57):
1) A total of 10 major dates of rainfall (> 1.25 cm/day) were observed which
resulted in significant runoff of azinphosmethyl (0.0005-3.920 Mg/L), endosulfan
(0.003-0.998 Atg/U, and fenvalerate (0.011-0.890
2) A total of three fish kills were observed at two field sites (TRT and KWA
Sites).
3) During fish kills both dead P. pugio and F. heteroclitus were observed among
endemic fauna.
4) In P. pugio, field mortality was observed in seven out of 10 rain events with
mortality rates ranging form 0-100%.
5) In F. heteroclitus, field mortality was observed in four out of 10 rain events
with mortality rates ranging from 0-100%.
During 1989-90, a jotal of eight days of significant (> 1.25 cm/day) rainfall was
observed, which resulted in significant runoff of azinphosmethyl (< DL - 7.002 ng/L),
endosulfan (< DL - 0. 163 ug/L), and fenvalerate (< DL - 0. 123 /ig/L) (Table 58). A
total of three fish kills were observed, two at the KWA Site and one at the adjacent
Haulover Creek Site. During these fish kills both dead P. pugio and F. heteroclitus were
observed as well as other fish species (M. cephalus, crustaceans (Uca and penaied
Shrimp) and other invertebrates (polychaetes and other annelids).
252
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Table 57.
Summary of field toxicity tests results from 1985-88 for P. pugio and
F. heteroclitus following dates of significant (> 1.27 cm/day) rainfall.
Rainfall
Date
6/8/85
6/27/85
5/14-15/86
6/9/86
6/4/87
6/19/87
6/23/87
6/24/87
6/25/87
6/9-10/88
Insecticide
Endosulfan
Fenvalerate
Endosulfan
Fenvalerate
Azinphosmethyl
Azinphosmethyl
Fenvalerate
Endosulfan
Fenvalerate
Endosulfan
Azinphosmethyl
Endosulfan
Fenvalerate
Endosulfan
Fenvalerate
Endosulfan
Fenvalerate
Azinphosmethyl .
Endosulfan
Insecticide
Concentration1
0.003
0.107
0.249
0.079
3.920
0.560
0.032
0.012
0.031
0.004r
0.005A-0.024
0.005A-0.012
0.011A-0.013
0.024A-0.058
0.11(^-0.890
0.018
0.070
3.440
0.998
% Mortality
P. pugio
52
100.0
F. heteroclitus
NM
NM
(Fish Kill = 189 dead fish and
crustaceans/50 m of stream)
NM 1 NM
(Fish Kill = Dead F. heteroclitus
observed in endemic fauna)
90
0
0
53
93-100
22-50
50
65-100
0
0
0
0
0
0
(Fish Kill » 150 dead fish and
crustaceaos/SOm of stream)
NM = Not Measured i
1 = Peak concentrations measured in grab samples unless otherwise noted
A = Composite Sample
253
-------
Table 58. Summary of field toxicity test results in P. pugio and F. heteroditus measured during periods of
significant (> 1.27 cm/day) rainfall for the 1989-90 field study.
Rainfall
Date
6/5 - 6/89
6/9/89
6/16/89
6/19/89
6/24/89
5/28/90
6/15/90
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Insecticide
Concentration1
(ug/L)
1.730
0.163
0.064
% Mortality
P. pugio
80.0 (± 20.0)
F. heteroditus
10 (± 0.0)
Fenvalerate
Azinphosmethyl
Endosulfan
0.065A - 0.093
0.368
0.054
71.5 (± 13.8)
0.0 (± 0.0)
0.0 (±0.0)
0.0 (± 0.0)
Fenvalerate
Azinphosmethyl
Endosulfan
0.022A - 0.021
36.7(±3.3)
2.457
0.038
Fenvalerate
Azinphosmethyl
Endosulfan
0.015
1.351
0.027
100.0(±0.0)-
62.6(±8.12)
0.0 (± 0.0)
0.0 (± 0.0)
(Fish Kill= Dead F. heteroditus, M.
cephalus, penaied shrimp, P. pugio
6.6 (± 6.6)
40.0 (1 15.28)
0.0(10.0)
3.3 (± 3.3)
Fenvalerate
Azinphosmethyl
Endosulfan
<0.003
7.002
0.065
Fenvalerate
Fenvalerate
Azinphosmethyl
Endosulfan
•C0.003
<0.003A- 0.123
0.024A - 0.062
0.005* - 0.004
10.7 (± 6.43)
100.0 (± 0.0)
0.0(± 0.0)
16.6 (± 6.7)
Fish Kill = Dead F. heteroditus. P.
pugio, Uca, and Polychaetes)
3.3 (±3.3
3.3 (±3.3)
3.3 (±3.3)
0.0 (± 0.0)
0.0 (± 0.0)
3.3 (± 3.3)
0.0 (± 0.0)
0.0 (± 0.0)
Peak concentrations in grab samples unless otherwise denoted.
Composite Samples
254
-------
In F. heteroclitus, significant (> 5%) field mortality was observed in one out of eight
rain events with mortality ranging from 0 - 16.6% ( + 6.7%). In P. pugio, significant (>
5%) mortality was observed in six out of eight rain events with mortality ranging from 0-
100%. All observed mortality was attributed to insecticide exposure as all
physicochemical parameters such as dissolved oxygen and salinity, were within known
tolerance ranges for the fish and crustacean species studied.
During 1989, six periods of significant rainfall were observed. Significant mortality
occurred at the KWA Site and TRT Site during six and two of the six rain events,
respectively. During the first, two rain events (6/5-6/89 - combined due to two days of
consecutive rain) deployed organisms were exposed to insecticide runoff. Significant
concentrations of azinphosmethyl (1.73 ^ig/L) and endosulfan (0.163 Mg/L) at the KWA
Site and fenvalerate (0.093 /*g/L) at the TRT Site caused 80 and 71.5% mortality,
respectively, in P. pugio deployed during field toxiciry tests. No mortality was observed
in F. heteroclitus at all sites during these two initial rain events. All mortality in this rain
event was attributed to insecticide exposure, as all physicochemical parameters, such as
dissolved oxygen and salinity, were within known tolerance ranges for the fish and
crustacean species studied.
During the third 1989 rain event (6/9/89), significant concentrations of azinphosmethyl
(0.368 ng/L) and endosulfan (0.054 jig/L) were observed at the KWA Site and significant
levels of fenvalerate (0.022 Mg/U at the TRT Site. In P. pugio, significant mortality
(36.7%) was observed at the TRT Site but not at the KWA Site. No mortality was
observed in F. heteroclitus at all sites during the third rain event. All P. pugio mortality
was attributed to insecticide runoff, as all physicochemical parameters such as dissolved
oxygen and salinity, were within known tolerance ranges for this crustacean.
During the fourth 1989 rain event (6/16/898), significant concentrations of
azinphosmethyl (2.457 jig/L) and endosulfan (0.038 ng/L) were observed at the KWA
Site, along with a fish kill involving dead F. heteroclitus, M. cephalus, penaied shrimp,
and P. pugio. At the TRY Site, significant runoff of fenvalerate (0.015 ng/L) was
observed. In P. pugio, significant mortality was observed at the KWA Site which ranged
from 62.6 - 100%. No P. pugio mortality occurred at the TRT Site. In F. heteroclitus,
no mortality was observed during the fourth rain event. All crustacean mortality was
attributed to insecticide exposure as all physicochemical parameters such as dissolved
oxygen and salinity, were within known tolerance ranges for these species.
255
-------
During the fifth rain event of 1989 (6/19/89), significant runoff of azinphosmethyl (1.35
Atg/L) and endosulfan (0.027 jug/L) was observed at the KWA Site. Significant mortality
(40%) was observed in P. pugio deployed at the KWA Site. At the TRT Site slight P. pugio
mortality (10.7%) w_as observed, despite the absence of detectable pesticide levels. No
mortality attributable to pesticides was observed in field deployed F. heteroclitus
(0-3.3%). All P. pugio mortality was attributed to insecticide exposure, as all
physicochemical parameters, such as dissolved oxygen and salinity, were within known
tolerance ranges for this organism.
During the sixth and final rain event of 1989 (6/24/89), significant runoff of azinphosmethyl
(7.002 pg/L) and endosulfan (0.065 pg/L) was observed at the KWA Site, which resulted in
significant mortality in field deployed P. pugio (100%) and F. heteroclitus (16.6%).
Immediately following this rain event, a fish kill was observed at the KWA Site, and later at a
tidal creek (Haulover) adjacent to the KWA Site. No mortality attributable to insecticide
exposure was observed at the TRT Sites in either P. pugio or F. heteroclitus. All mortality
was attributed to insecticide exposure, as all physicochemical parameters, such as dissolved
oxygen and salinity, were within known tolerance ranges for these species.
During 1990. only two periods of significant (> 1.25 cm/day) rainfall were observed (5/28
and 6/15/90). No significant mortality was observed among field deployed P. pugio and F.
heteroclitus, despite the presence of a significant concentration of fenvalerate (< DL - 0.123
Aig/L) at the TRT Site (5/28/90) and azinphosmethyl (0.024 - 0.062 /ig/L) at the KWA Site
(6/15/90). The absence of toxicity in P. pugio exposed to potentially toxic levels of fenvalerate
(0.123 ng/L) was puzzling. Analysis of composite water samples for the initial = 12h post
rain period indicated that fenvalerate levels were < DL. The high fenvalerate levels (0.123
Mg/L) were observed at dead low tide. These data suggest that only a small amount of low
salinity runoff water containing high fenvalerate concentrations was discharged at the TRT Site
during this runoff event. The retention pond at the TRT Site, by retaining a large portion of
tomato field runoff, may haves reduced the overall runoff volume sufficient that, the incoming
flood tide was able to rapidly dilute fenvalerate concentrations to levels < DL. The potential
runoff loading capacity (volume of runoff/volume of stream) may decrease by 86% from ebb
tide to flood tide, due to the simple increase in stream volume associated with normal mesotidal
tidal exchange. These data are suggestive that the retention ponds at the TRT Site may in part
help enhance the assimilative capacity of a watershed, by reducing runoff volumes and in turn
resulting pesticide concentrations.
256
-------
Biomonitoring or ecotoxicological studies of endemic field populations provide a third tier in
[he environmental risk assessment of pesticides. The approach used in most biomoniioring
studies is quantitative, replicated ecological (usually pelagic or benthic) sampling. The results
of such studies provide estimates of field population changes in response to toxicant exposure.
Biomonitoring provides an independent mechanism to confirm che validity of toxicity cest
results. In addition, direct linkage between biomonitoring and laboratory/fie Id toxicity tests is
needed and should include the use of species in toxicity tests which are endemic to the habitat
being studied. The two organisms used in laboratory toxicity tests in this present study, P.
pugio and F. heteroclitus, are the most dominant crustacean and fish species, respectively in the
Leadenwah tidal creek drainage basin, accounting for over 80% of the annual total abundance.
Results of earlier ecotoxicological biomonitoring studies (Scott et ai, 1990; Hampton, 1987;
and Patterson, 1986) conducted from 1985-88, at the TRT and CTL Sites (See Table 59)
indicated:
1) In the absence of significant pesticide runoff, traditional ecological comparisons such as
species richness, evenness, diversity, index of similarity, total abundance, and total biomass
were virtually identical, from January - May of each year, prior to peak periods of
agricultural runoff (late May - June);
2) From 1985-88, ecotoxicological biomonitoring indicated significant mortality in eight out of
the 10 rain events, with mortality rates ranging from 0-99% in P. pugio and from 0-95%
in F. heteroctitus; and
3) Following periods of significant pesticide runoff during 1985-88, significant (p < 0.05)
reductions in total biomass, total abundance, and densities of P. pugio, F. heteroclitus,
Penaeus species, Callinectes sapidus, and total fish were observed at the TRT Site.
During 1989-90 (Table 6), a total of eight days of significant (> 1.25 cm/day) rainfall were
observed, which resulted in significant runoff of azinphosmethyl (< DL - 7.002 ^g/L),
endosulfan (< DL - 0.163 ftg/L), and fenvalerate (< DL - 0.123 ng/L). A total of three fish
kills were observed, two at the\KWA Site and one at the adjacent Haulover Creek Site. During
these rain events, significant (> 5%) mortality was observed in three (including fish kills) out
of the eight rain events in F. heteroclitus, with mortality ranging from 0 - 24.4% based upon
block seining. F. heteroclitus results for 19,90, indicated significant mortality in push netting
during both rain events, with mortality estimates ranging from 21.2 - 45.4%. In P. pugio,
block seining results for 1989-90, indicated significant (> 5%) mortality in five out of the eight
rain events (including fish kills), with mortality estimates ranging from 0-43.1%. P. pugio
push netting results for 1990, only indicated significant mortality in one out of the two rain
events, with mortality estimates ranging from 0 - 24.6%.
257
-------
Table 59. Summary of ecotoxicological biomonitoring estimates of field mortality in P. pugio
and F. heteroditus following dates of significant (> 1.27 cm/day) rainfall.
Rainfall
Date
6/8/85
Insecticide
Endosulfan
Fenvalerate
Insecticide
Concentration1 0*g/L)
0.003
0.107
% Predicted Mortalitv1
P.
pugio
72
F.
heteroditus
75
6/27/85
Endosulfan
Fenvalerate
0.249
0.079
98
80
(Fish Kill)
5/14-15/86
Azinphosmethyl
3.920
61
69
(Fish Kill)
6/9/86
Azinphosmethyl
Fenvalerate
0.560
0.032
99.9
83
6/4/87
Endosulfan
Fenvalerate
0.012
0.031
0
0
f
6/19/87
Endosulfan
0.004
0
0
6/23/87
Azinphosmethyl
Endosulfan
Fenvalerate
0.005A-0.024
0.005A-0.012
0.011A-0.013
82B-94
69" -95
6/24/87
Endosulfan
Fenvalerate
0.024A-0.058
0.1 10^0.890
82B-94
69B-95
06/25/87
Endosulfan
Fenvalerate
"0.018
0.070
82B-94
698-95
•
6/9-10/88
Azinphosmethyl
Endosulfan
\ 3.440
0.998
NM
NM
(Fish Kill)
1 = Peak Concentrations measured in grab samples unless otherwise noted.
2 = Between site mortality estimates unless otherwise noted.
A = Composite Sample
B = Within site mortality estimate.
NM = Not Measured
258
-------
Table 60. Summary of ecotoxicological estimates of mortality in P. pugio and F. heteroclitus
observed during significant (> 1.27 cm/day) rainfall events monitored during the
1989-90 field studv.
Rainfall
Date
6/5 - 6/89
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Insecticide
Concentration1
(us/L)
1.730
0.163
0.064
% Mortality*
P. pugio
NM
F. heteroclitus
NM
Fenvalerare
0.065A - 0.093
43.1
0
6/9/89
Azinphosmethyl
Endosulfan
0.368
0.054
NM
NM
Fenvalerate
0.022* -0.021
22.7
0
6/16/89
Azinphosmethyl
Endosulfan
2.457
0.038
NM
NM
Fish Kill
Fenvalerate
0.015
0
0.0 (± 0.0)
-
6/19/89
Azinphosmethyl
Endosulfan
1.351
0.027
NM
3.3 (± 3.3)
Fenvalerate
< 0.003
0
0.0 (± 0.0)
6/24/89
Azinphosmethyl
Endosulfan
7.002
0.065
NM
16. 6 (± 6.7)
Fish Kill
Fenvalerate
<0.003
0
0
-
5/28/90.
6/15/90
Fenvalerate
Azinphosmethyl
Endosulfan
<0.003A- 0.123
0.024f: - 0.062
0.005* - 0.004
tf - 24. 63
0
0.0 (± 0.0)
21. 2J -24.42
O1 - 45. 43
0.0 (± 0.0)
= Between site mortality estimate unless otherwise noted.
= Composite Sample
= Peak concentrations in grab samples unless otherwise denotes
= Block Seine Estimate
= Push Net Estimate
259
-------
During the six days of significant rainfall during 1989, the only significant mortality
observed at the TRT site was in the crustacean, P. pugio exposed to fenvalerate at
concentrations ranging from 0.065 (composite) - 0.093 (peak grab) ng/L during the rain
events of 6/5-6/89 and at concentrations ranging from 0.021 (peak grab) - 0.022
(composite) jig/L during the rain event of 6/9/89. P. pugio predicted mortalities from
block seining were 43.1% and 22.7% respectively, for these two rain events.
Unfortunately, during 1989, biomonitoring was not conducted at the KWA Site.
During the two days of significant rainfall during 1990, significant runoff of
fenvalerate at concentrations ranging from 70%) of a defined area, than does push netting. As a
result, when predicting P.pugio mortality from push netting, it is imperative to view any
predicted mortality estimate within the context of statistical analysis for relative P. pugio
abundance. Unless statistical analysis indicates that the relative P. pugio abundances are
significantly different, then predicted mortality estimates are unreliable and may instead
reflect simple population variability.
260
-------
Comparisons between observed mortality in field toxicity tests and ecotoxicological
biomonitoring studies with mortality estimates predicted for laboratory toxicity tests are
difficult. Earlier studies by Scott et al, (1990) have indicated that each method of environmental
risk assessment (acute^aboratory toxicity test, in situ toxicity tests, and ecotoxicological
biomonitoring) was useful in the assessment of the acute toxicity for three classes of pesticides
(organochlorine, organophosphates, and pyrethoids) commonly used in agriculture. Generally
most methods were similar in their prediction of the acute toxicity of azinphosmethyl,
endosulfan, and fenvalerate on P. pugio and F. heterodiius. Linear regression analysis for all
pesticides (azinphosmethyl, endosulfan, and fenvalerate) and all species (P. pugio and F.
heteroditus) were significantly (p < 0.01 - 0.05) correlated (R - 0.47 - 0.64) in comparisons
between field and laboratory toxicity tests (R = 0.63, p < 0.03), field toxicity tests and
ecotoxicity estimates (R = 0.64, p < 0.01), and laboratory toxicity tests versus ecotoxicity
estimates (R = 0.47, p < 0.05). Regression analysis for individual species, rather than
combined species, gave much higher correlations in comparisons between the various toxicity
assessment methods. For example, regression analysis for all pesticides and P. pugio were
significantly (P < 0.01 - 0.001) correlated (R = 0.78 - 0.95). In further comparing these
regression results, Scott et al. (1990) reported that:
1) The % mortality results comparisons forall pesticides and all species indicated that the
mean differences in mortality estimates ranged from 19-50% for the various methods
(lab, field, and ecotoxicity).
2) In P. pugio. the mean difference in % mortality estimates by the various methods (lab,
field, and ecotoxicity) was much smaller, ranging from 7-18%.
3) The major source of error in making accurate predictions from laboratory toxicity tests
is in the precision of field dose determination (peak versus composite insecticide
concentrations).
4) The major sources of error in making accurate predictions from field toxicity tests
include:
\
a) Limited size class estimates of mortality (generally only adults are utilized);
b) Confounding physicochemical factors such as low salinity, total filterable residue,
dissolved organic carbon, and temperature; and
261
-------
c) Presence and potential interaction of other insecticides.
5) The major sources of error in making accurate predictions from ecotoxicological
biomonitoring include:
a) Inter, vs Intrasite comparisons
b) Cumulative (multiple rain events) in field populations versus single rain event/dose
response effects in field and laboratory.
6) Comparisons between field derived and laboratory toxicity tests LC50 values indicated
generally excellent agreement between field results and 96h, static renewal laboratory
test results. Laboratory 6h pulsed dose tests results greatly underestimated field
mortality effects.
Tables 61-62 (F. heteroditus - 1989 and 1990, respectively) and 63-64 (P. pugio - 1989 and
1990, respectively) list the predicted mortality in raummicnogs and grass shrimp from various
laboratory toxicity tests at measured field concentration of azinphosmethyl, endosulfan, and
fenvalerate detected during runoff events for 1989-90. Also listed are the mortality rates in
caged P. pugio and F. heteroditus observed in fieM toxicity tests and estimated mortality rates
for P. pugio and F. heteroditus from ecotoxicity sampling (block seining - 1989-90; push
netting - 1990).
In F. heteroditus, laboratory toxicity testing predicted no significant (< 5% = Control)
mortality in all 1989-90, rain events but one. During the rain event of 24 June, 1989, (Table
61), significant runoff of azinphosmethyl (7.002 ng/L) was observed at the KWA Site.
Laboratory toxicity tests with F. heteroditus predicted 10% mortality which compares with
observed field mortality of 16.6% (± 6.7%). The laboratory-derived NOEC for
azinphosmethyl of 4.95 Mg/L, whjch was exceeded in the field with resulting mortality (16.6 ±
6.7%), again suggests significant agreement between the laboratory and field.
The only measured field effects in F. heteroditus were observed during 1990 rain events
(Table 62). During the first rain event significant fenvalerate concentrations (0.123 /*g/L) were
observed. Ecotoxicity measurements (block seining and push netting) predicted mortality
ranging from 21.2 - 24.4% mortality. No significant mortality was observed in field toxicity
testing nor was any significant mortality predicted in laboratory toxicity tests. Laboratory and
262
-------
Table 61. Summary of field and laboratory toxicity tests results and ecotoxicological estimates of
mortality in F. heteroclitus during the 1989 field study.
Rainfall
Date
6/5 - 6/89
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Insecticide
Concentration'
(ug/L)
1,173
0.163
0.064
% Mortality in F. heteroclitus
Lab
0
0
0- 3%
Field
0.00 (± 0.0)
Ecotox2
NM
Fenvalerate
0.065A-
0.093
0 - 3%
0.00- (± 0.00)
0
6/9/89
Azinphosmethyl
Endosulfan
0.368
0.054
0
0
0.00 (± 0.00)
NM
Fenvalerate
0.22A-0.021
0
0.00 (± 0.00)
0
6/16/89
Azinphosmethyl
Endosulfan
2.457
0.038 ~-
0
0
L0.00(± 0.00)
NM
Fenvalerate
0.015
0
0.00 (± 0.00)
0
6/19/89
Azinphosmethyl
Endosulfan
1.351
0.027
0
0
3. 3 (±3.3)
NM
Fenvalerate
< 0.003
0
0.00 (± 0.00)
0
6/24/89
Azinphosmethyl
Endosulfan
7.002
0.065
10
0
16.6 (± 6.7)
NM
Fenvalerate
< 0.003
0
0.00 (±0.00)
0
I _
2 —
= Composite Samples
Peak concentrations in grab samples unless otherwise denoted
Block Seine Estimate
263
-------
Table 62. Summary of field and laboratory toxicity tests results and ecotoxicological estimates of
mortality in F. heteroclitus during the 1990 field study.
Rainfall
Date
5/28/90
Insecticide
Fenvalerate
Insecticide
Concentration1
(ug/L)
<0.003A-0.123
% Mortality in F. heteroclitus
*
Lab
0- 3%
Field
3.3 (± 3.3)
Ecotox2
24.43 - 21. 22
6/15/90
Azinphosmethyl
Endosulfan
0.024A - 0.062
0.005A - 0.004
0 -0
0
0.00 (± 0.00)
O2 -45.4%3
= Composite Samples
Peak concentrations in grab samples unless otherwise denoted
Block Seine Estimate
Push Net Estimate for non P. pugio Totaf Faunal Density
264
-------
field toxiciry tests are based upon results for adult F. heteroclitus where as ecotoxiciry estimates include
adult and juvenile F. heteroclitus. Juvenile F. heteroclitus may be more sensitive to fenvalerate exposure.
although earlier tests exposing larval mummichog (1-2 day old larvae) to fenvalerate did not indicate
increased sensitivity (i.e., adult and larval LC50s were comparable). These laboratory tests with larval F.
heteroclitus were at 20 ppt salinities and not at the low salinities (5-10 ppt) observed in the field during this
rain event. At lower salinities fenvalerate may be more toxic to larval mummichogs. Another factor to
consider is that laboratory toxicity tests with fenvalerate were considered with pydrin where as all field
exposures during 1989-90 for fenvalerate were Asana. Additional toxicity testing with larval F. heteroclitus
is needed to better resolve these differences.
During the second 1990 rain event (15 June, 1990), only ecotoxicity estimates predicted from push
netting were suggestive of potential toxic effects in F. heteroclitus. As earlier discussed, push netting does
not provide accurate censusing of mummichogs and other fish species and as a result spurious conclusions
may be reached. During this second rain event, this appeared to be the case as only push netting indicated
potential F. heteroclitus mortality. Other assessment methods (lab and field toxicity testing and block
seining) indicated no significant toxiciry occurred.
In P. pugio, laboratory toxicity tests predicted significant (> 5%) mortality in seven out of the eight
rain events observed during 1989-90. Field toxicity tests with P. pugio measured significant mortality in
six out of the eight rain events. Ecotoxicity sampling (block seining - 1989-90; push netting - 1990)
predicted significant mortality in seven out of the eight rain events observed during 1989-90.
During the first two rain events (5-6 June, 1989) significant runoff of azinphosmethyl (1.73 ^g/L),
endosulfan (0.163 ^g/L) and fenvalerate (0.064 ^g/L) was observed at the KWA Site (Table 63). Also,
significant runoff of fenvalerate (0.09?.jig/L) was observed at the TRT Site. Laboratory toxicity tests
predicted mortalities of 51% (azinphosmethyl), 6-15% (endosulfan), 53-80% (fenvalerate), and combined
pesticide mortality of 100% (assuming simple additive toxicity) in P. pugio at the KWA Site. At the TRT
Site, P. pugio mortality from laboratory toxicity tesls was predicted at 85-95% (fenvalerate). Observed
mortalities in field toxicity tests were 80% (± 20%) at the KWA Site and 72% (± 13.8%) at the TRT
Site. No ecotoxicity sampling was conducted at the KWA Site during 1989, but block seining at the.TRT
Site predicted 43.1% mortality in P. pugio. These results between field, lab and ecotoxicity
265
-------
Table 63. Summary of field and laboratory toxicity tests results and ecotoxicological estimates of
mortality in P.pugio during the 1989 field study.
Rainfall
Date
6/5 - 6/89
6/9/89
6/16/89
6/19/89
6/24/89
_
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Fenvalerate
Azinphosmethyl
Endosulfan
Fenvalerate
Azinphosmethyl
Endosulfan
Fenvalerate
Azinphosmethyl
Endosulfan
Fenvalerate
Azinphosmethyl
Endosulfan
Fenvalerate
Insecticide
Concentration1
(uf/Lt
1.173
0.163
0.064
0.065A-0.093
0.368
0.054
0.022*-0.021
2.457
0.038
0.015
1.351
0.027
< 0.003
7.002
0.065 .
< 0.003
Lab
51
06- 15
53 -80
85-95
0
0-6
0-6
96
0-6
0-6
51
0-6
0
100
0-6
0
% Mortality in P. pugio
Field
80 (± 20)
72 (± 13.8)
0.00(± 0.00)
36.7 (± 3.3)
lOO.OCKtO.OO)-
62.6(±8.12)
6.6 (± 6.6)
40 (± 15.3)
10.7 (± 6.4)
100.0 (± 0.0)
3.3 (± 0.00)
Ecotox3
NM
43.1
NM
22.7'
NM
22.7Z
NM
22. 73
NM
22.7
Peak concentrations in grab samples unless otherwise noted.
Mortality Estimate in P. pugio for period 6/9 - 27/89.
Ecotox Estimates Derived from Block Seining Data
Composite Samples
266
-------
estimates of mortality closely agree. Additionally, these results were very similar to
results obtained during the rain event of 9 June 1986 at the TRT Site. When similar
concentrations of azinphosmethyl (0.58 fig/L) and fenvalerate (0.032 ng/L) caused
significant mortally in field toxicity tests (90%) and field populations of P. pugio (90%).
During the third rain event (9 June, 1989), significant runoff of azinphosmethyl (0.368
Mg/L) and endosulfan (0.054 jxg/L) was observed at the KWA Site. Also, significant
runoff of fenvalerate (0.022 fig/L) was observed at the TRT Site. Laboratory toxicity
tests predicted P. pugio mortalities ranging from 0%(azinphosmethyl) to 0-6%
(endosulfan) at the KWA Site which were very similar to 0% observed in field toxicity
tests. Laboratory toxicity tests with P. pugio also predicted mortalities ranging from 0-6%
(fenvalerate) compared to observed mortalities of 36.7% in field toxicity tests and 22.7%
in ecotoxicity sampling. This lack of close agreement between field and laboratory
toxicity tests may have resulted in part from poor characterization of peak fenvalerate
concentrations during this rain event. The fenvalerate concentration measured in the
composite sample exceeded the peak grab concentration, suggesting that the grab sampling
schedule employed may have missed the actual peak fenvalerate concentration. As
reported by Scott et al. (1990), the primary factor affecting correlation between the
laboratory and the field is accurate characterization of field pesticide concentrations to use
in laboratory toxicity dose response curves. 'Another factor may be that laboratory toxicity
tests with fenvalerate utilized Pydrin where as field exposures were from Asana.
During the fourth rain event (16 June, 1989), significant runoff of azinphosmethyl
(2.457 jig/L) and endosulfan (0.038 pg/L) was observed at the KWA Site and significant
runoff of fenvalerate (0.015 ng/L) was observed at the TRT Site. Laboratory toxicity
tests with P. pugio predicted mortalities ranging from 96% (azinphosmethyl), 0-6%
(endosulfan) and combined mortality of 96-100% (azinphosmethyl and endosulfan) at the
KWA Site. Field toxicity tests with P. pugio measured mortalities ranging from 62.6% (2
days post rain deployment^ to 100% (initial post rain deployment) at the KWA Site, which
were very similar to predictions from laboratory results. At the TRT Site, observed P.
pugio mortalities ranged form 6.6% (field toxicity tests) to 22.7% (ecotoxicity sampling)
compared to laboratory toxicity tests which predicted mortalities of 0-6%. Excluding the
ecotoxicity predicted mortality of 22.7%, (which was an integrated mortality prediction for
the time period of 6/7-6/27/89 which encompassed four rain events), there was close
agreement between field and laboratory results at the TRT Site during the fourth rain
event.
267
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During the fifth rain event of 19 June, 1989, there was significant runoff of
azinphosmethyl (1.351 /ig/L) and endosulfan (0.027 /ig/L) at the KWA Site. Laboracory
toxicity tests with P. pugio predicted mortalities ranging from 51%( azinphosmethyl), 0-
6% (endosulfan), jind 51-57% (combined azinphosmethyl and endosulfan) compared to
field toxicity tests with P. pugio which measured 40% (± 1.53%) mortality. This
suggests excellent agreement between field and laboratory toxicity testing for this rain
event. At the TRT Site, no detectable levels of pesticides were observed. As a result,
laboratory toxicity tests predicted P. pugio mortality was 0% compared to field toxicity
test mortalities of 10.7% (± 6.4%), which was just above maximum field control
mortality (5%). This is in close agreement between field and laboratory results.
During the sixth rain event of 21 June, 1989, there was significant runoff of
azinphosmethyl (7.002 /ig/L) and endosulfan (0.065 /ig/L) at the KWA Site. Laboratory
toxicity tests with P. pugio predicted mortalities ranging from 100% (azinphosmethyl) to
0-6% (endosulfan), and combined mortalities of 100% (azinphosmethyl and endosulfan)
compared to field toxicity tests results which reported 100% (± 0.0%). This again
suggests close agreement between field and laboratory toxicity tests results at the KWA
Site. At the TRT Site, no detectable pesticide concentrations were observed. As a result,
laboratory toxicity tests predicted P. pugio mortality was 0% compared to 3.3% (± 0.0%)
field toxicity tests.
During the seventh rain event (28 May, 1990-Table 64) significant runoff of fenvalerate
(< DL - 0.123 ng/L) was observed at the TRT Site. Laboratory toxicity tests predicted
P. pugio mortalities ranging from 0% (composite sampling) to 100% (peak grab sample).
Observed field mortality was 3.3% (± 3.3%) in field toxicity tests and ranged from 0%
(block seine) - 24.6% (push netting) in ecotoxicity sampling estimates. Results from this
rain event graphically illustrate the problem in accurately characterizing the field exposure
concentration to use in laboratory toxicity test dose response models. Using the composite
concentration, significant correlation exists between field and laboratory results. Using
peak grab concentration, field laboratory results do not agree. Which measured field
concentration (peak or composite) is most
\
268
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Table 64. Summary of field and laboratory toxicity tests results and ecotoxicological
estimates of mortality in P. pugio during the 1990 field study.
Rainfall
Date
5/28/90
Insecticide
Fenvalerate
Insecticide
Concentration1
(ug/L)
< 0.003M). 123
% Mortality in P. pugio
Lab
100
Field
3.3 (± 3.3)
Ecotox2
O2 - 24 63
6/15/90
Azinphosmethyl
Endosulfan
0.024A - 0.062
0.005* -0.004
0-0
0
3.3 (± 3.3)
0
= Composite Samples
Peak concentrations in grab samples unless otherwise denoted
Block Seine Estimate
Push Net Estimate
269
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appropriate to use in laboratory toxicity tests predictions? Results from this study suggest
that both composite and grab sampling are needed to accurately characterize field dose to
obtain correlative results between the field and the laboratory. Geographical (acres of
agricultural fieldiTetc.) and hydrological (stream volume, stream flow) factors and
agricultural management practices (vegetative filter strips, retention ponds) may all
ultimately affect and define the field pesticide exposure regime. During this rain event,
perhaps only a small first flush "slug" of fenvalerate was discharged from retention ponds
following the initial rain event. Very little additional pesticide discharge must have
occurred after this initial runoff was observed, as evidenced by the nondetectable
fenvalerate concentrations observed in composite samples. Results from this rain event
are also suggestive that when "slugs" of pesticides are rapidly diluted with no resulting
field toxicity , that the assimilative capacity of the stream has been maintained. The
construction and operation of retention ponds along with IPM and BMP at the TRT Site
may have contributed, in part to the return of the assimilative capacity in this water shed.
The use of grab and composite sampling in conjunction with hydrolab measurements is an
integrated procedure for not only accurately determining the pesticide field exposure
regime, but also to evaluate the assimilative capacity of a watershed to predict its
vulnerability to nonpoint source pesticide runoff. In the future the use of peak
grab/composite sample ratios may provide seme measure of a water shed's vulnerability to
NFS agricultural runoff when evaluated with other hydrological and toxicological
information.
During the eighth and final rain event of 15 June, 1990 (Table 64), significant
concentrations of azinphosmethyl (0.062 /ig/L) were observed the KWA Site. Laboratory
toxicity tests with P. pugio, predicted 0% mortality which was highly correlated with field
toxicity tests (0%) and ecotoxicity sampling results (0% in block seine and push netting).
Results from all field toxicity tests and ecotoxicity sampling for rain events studied
form 1985-90 are listed in'Table 65 as field derived LCjo values. These results are then
compared with laboratory derived LCX values for a variety of toxicity tests [96h static
renewal (SR) and 6h pulsed dose (PD) at high (20 ppt) and low (5 ppt) salinities]. These
results indicated generally excellent agreement between field and laboratory derived LC^
values for P. pugio and F. heteroditus exposed to azinphosmethyl, endosulfan and
fenvalerate.
270
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Table 65. Comparison of field and laboratory derived LC50 values in P. pugio and
F. heteroclitus exposed to azinphosmethyl, endosulfan and fenvalerate.
Note the similarities between lab and field derived 96h LC50 values.
Insecticide
Azinphosmethyl
Endosulfan
Fenvalerate
Test Organism
P. pugio
F. heteroclitus
P. pugio
F. heteroclitus
P. pugio
F. heteroclitus
Field Derived*
96h LCso
(± 95% CL) in ng/L
0.95
(0.86 - 1.05)
> 7.002
(NC)
LOEC = 7.00
0.28
(NC)
> 0.998
(NC)
0.06
(0.05- 0.07)
0.23
(0.19-0.28)
Laboratory Derived
LCM Value
in ng/L
0.97 - 1.05 SR
6.68 - 8.14PD
28.00 - 36.95 SR
NOEC =4.95
0.25 - 1.01 SR
0.27 - 0.58 SR, Z
3.81 -4.35 PD
1.29- 1.45 SR
0.14-0.40SR, J
0.052 - 0.060 SR
0.013 - 0.031 SR, Z
0.235 - 0.314 PD
1.63-2.86
1.67-4.26SR, J
A = Field Derived LCjo values were based upon a compilation of ecotoxicology and
field toxicity test results.
NC = Confidence Limits Not Calculated.
SR = 96h Static Renewal Toxicity Tests at low (5 ppt) and high (20 ppt) salinities.
PD = 6h Pulsed Dose Toxicity Testsfet low (5 ppt) and high (20 ppt) salinities.
Z = Zoeae, l-2d; other values are for adults unless otherwise noted.
J = Juvenile; other values are for adults unless otherwise noted.
271
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In P. pugio field derived LC50 values were:
1) 0.95 ng.'L for azinphosmethyl (CL = 0.86 - 1.05 jzg/L) versus 96h laboratory SR
LC50 values ranging from 0.97 - 1.05 ^g/L (CL = 0.77-1.24
2) 0.28 jxg/L for endosulfan versus 96h laboratory SR LC;o values ranging from 0.25 -
1.01 fjig/L (CL = 0.14 - 1.43 Atg/L) in adults and 0.39 jig/L (CL = 0.27 - 0.58 /ig/L)
in zoeae; and
3) 0.06 fjLg/L (CL = 0.05 - 0.07 Mg/L) for fenvalerate versus 96h laboratory SR LC50
values ranging from 0.052 - 0.060 jig/L (CL = 0.037 -0.097 jig/L) in adults and
0.007 - 0.020 0g/L (CL = 0.005 - 0.031 p.g/L) in zoeae.
These results indicated significant agreement between field results and 96h SR laboratory
toxicity tests. The 6h pulsed dose laboratory toxicity tests LCW values for all three pesticides
were not as highly correlated with field results, as they greatly underestimated field toxicity in
P. pugio.
In F. heteroclitus, field derived LC50 values^vere:
1) > 7.002 jig/L for azinphosmethyl versus 96h SR LCjo values ranging from 28.00 -
36.95 jig/L (CL = 20.23 - 48.24 /zg/L). Also the field derived LOEC was 7.00
versus a 96h SR NOEC of 4.95
2) 0.12 Atg/L for endosulfan versus 96h SR LC*, values ranging from 1.29 - 1.45
(CL = 1.21 - 1.59 jig/L) for adults and 0.23 ng/L (CL = 0.14 - 0.40 pg/L) for
juveniles; and
3) 0.10 pig/L (CL = 0.09"- 0.11 ng/L) for fenvalerate versus 96h SR LCjo values
ranging from 1.63 - 2.86 /ig/L (CL = 1.08 - 4.06 jig/L) for adults and 2.67
(CL = 1.67 - 4.26 jig/L) for juveniles.
k
These results generally indicated good agreement between field and laboratory results.
Generally field derived LCW values were lower than laboratory derived values. This was
because field derived LC^ values used both field toxicity test results and ecotoxicity sampling
272
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estimates, which included both adult and juvenile F. heteroclitus. Reported laboratory results,
were for adults, except where indicated (i.e., endosulfan). For endosulfan, the juvenile F.
heteroclitus LC50 value (0.23 ^g/L) was almost 5 times lower than for adults (1.29 - 1.45 ng/L).
As these results indicate, field derived LC50 values for F. heteroclitus, which include both
juvenile and adults, may be lower than laboratory derived LC50 values for adults only.
Additional toxicity testing with juvenile F. heteroclitus at high and low salinities and in pesticide
mixture combination are needed to better define these field and laboratory comparisons.
Ecophysiological studies, using both specific (i.e., biomarkers - AChE inhibition) and
nonspecific (i.e., general physiology - 02 respiration, nitrogen excretion, and O/N ratios)
measures of sublethal effects, provides a fourth tier of ecological assessment which may be used
to assess acute and/or chronic sublethal physiological stress responses in aquatic organisms
exposed to pesticides and other toxic chemicals.
In this present study, during 1989-90, significant runoff of azinphosmethyl (1.73 - 7.00
/ig/L) at the KWA Site resulted in inhibition of brain AChE in F. heteroclitus. Earlier studies
by Fulton (1989) and also reported in Scott et at., (1990) found significant brain AChE
inhibition in mummichogs following exposure to azinphosmethyl at the KWA Site during 1988.
Laboratory toxicity tests exposing F. heteroclitus to azinphosmethyl reported:
1) Reduced whole animal nitrogen excretion rates following 24h sublethal azinphosmethyl
exposures; and
2) 24h EC50 (based upon % brain AChE inhibition) of 0.90 /ig/L.
A comparison between field and laboratory derived EC^s for F. heteroclitus exposed to
azinphosmethyl is listed in Table 66. Note the similarities between the field derived ECM (0.63
- 1.53 fig/L) and laboratory derived ECjo (0.90 fig/L). These findings clearly demonstrate field
validation of laboratory-derived EC^ value in P. heteroclitus exposed to azinphosmethyl. As
was previously mentioned in the acute toxicity field and laboratory comparisons section the
greatest single factor affecting field and laboratory comparisons, is the accuracy of the field
pesticide exposure concentration. This was rtyilected in the range of field derived ECso estimates
obtained (0.63 - 1.53, average = 1.13 jxg/L). Never the less, field derived EC^ values using
brain AChE were very close to the ECM reported for laboratory toxicity tests with
azinphosmethyl.
273
-------
Table 66. Comparison of field and laboratory derived EC^ (% Brain AChE
Inhibition in ug/L) in F. heteroclitus. Note the similarity between lab and
field derived ECM values.
Pesticide
Azinphosmethyl
Species
F. heteroclitus
Laboratory Derived EC^1
in ug/L (95% CL in ug/L)
0.90
(NC)
Field Derived EC^1
in ug/L (Range ;)
1.13
(0.83 - 1.53)
1 = Based upon Brain AChE Inhibition
2 = Range based upon % inhibition using maximum insecticide concemation and
24h concentration.
NC = Not calculated.
274
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Additionally AChE may be an excellent biomarker for AChE inhibiting pesticides such as
organophosphate and carbamate insecticides. Many of these chemicals are difficult to
monitor, due to their short half-lives in the environment.
Brain AChE may thus be useful as a surrogate biomarker of exposure for many
AChE inhibiting pesticides with short half lives. This is particularly true given the
persistence of AChE inhibition as observed during laboratory studies, which showed only
partial AChE recovery following 7 days of depuration. Further studies with other AChE
inhibiting pesticides and other fish species are needed to fully understand the usefulness of
AChE as a field biomarker of exposure and sublethal physiological stress.
Nonspecific biomarkers such as bioenergetic metabolism were used to evaluate
physiological alterations in basal metabolism in the lab (mummichogs - azinphosmethyl
exposure) and field (oysters). In both field and laboratory studies, the nonspecific
biomarkers (respiration rate, nitrogen excretion, and O/N ratio) assessed were useful in
identifying effects in the species tested; however, interpretation of these results and
identifying cause and effect relationships with pesticide exposure is an extremely difficult
task using these tools. The concomittant exposure of pesticide and low salinity conditions
particularly may confound interpretation of field results. Extensive laboratory studies to
quantify and identify the importance of confounding factors, such as salinity, on bioenergetic
metabolism are essential to better understanding and identifying pesticide and low salinity
effects.
One limitation of this study, was that spatial, watershed-wide impacts (upstream and
down stream) were not assessed. Of particular interest is the assessment of pesticide runoff
impacts on the fauna of larger stream reaches, marine mammals, sea turtles, wading and
shore birds, and benthic communties. Funding restraints obviously limited the scope of this
project to address basin-wide impacts, although the limited studies conducted clearly
indicated that pesticide transport up to two river miles away was observed along with
significant lexicological impact at the KWA Site during 1989. Despite these limitations,
results from these studies clearly indicate the need to protect estuarine ecosystems in the
most vulnerable, upper reaches of small estuarine tidal creeks, which are the true nursery
ground for many marine fish and invertebrate species. If water quality and environmental
integrity is preserved in these nursery ground areas, water quality and environmental
integrity will be maintained in larger tidal stream reaches, further downstream, assuring a
safe and healthy ecosystem for most estuarine species.
275
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