Biological Opinion

National Marine Fisheries Service Endangered Species Act 

Section 7 Consultation on Environmental Protection Agency’s 

Registration of Thiobencarb

Reprinted with permission from the photographer, Chris Haile, chrishailephotography.com, Marysville, CA.

June 30, 2012
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Table of Contents

1. Background................................................................................................. 4

2. Consultation History................................................................................... 5

2.1. Species Addressed in the BE.................................................................................. 16

3. Description of the Proposed Action.......................................................... 18

3.1.1. Thiobencarb ...........................................................................................24

4. Action Area ............................................................................................... 28

5. Approach to the Assessment ..................................................................... 31

5.1. Evidence Available for the Consultation ............................................................... 34

5.2. Application of Approach in this Consultation ....................................................... 35

5.2.1. Problem Formulation.............................................................................35

5.2.2. Designated Critical Habitat...................................................................38

5.2.3. Critical Habitat Risk Hypotheses...........................................................40

5.2.4. Evaluating Exposure and Response.......................................................40

5.2.5. Analysis Plan .........................................................................................43

5.3. Other Considerations............................................................................................. 44

6. Status of Listed Resources........................................................................ 46

6.1. Species Status......................................................................................................... 47

6.2. Conservation Role of Critical Habitat for the Species .......................................... 49

6.3. Chinook Salmon..................................................................................................... 52

6.3.2. Central Valley Spring-run Chinook Salmon..........................................56

6.3.3. Sacramento River Winter-run Chinook Salmon ....................................69

6.4. California Central Valley Steelhead...................................................................... 79

7. Environmental Baseline............................................................................ 93

7.1. Natural Mortality Factors ..................................................................................... 94

7.1.1. Parasites and/or Disease .......................................................................94
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7.1.2. Predation................................................................................................95

7.1.3. Wildland Fire .........................................................................................97

7.1.4. Oceanographic Features, Climatic Variability and Climate Change ...98

7.2. Anthropogenic Mortality Factors........................................................................ 100

7.2.1. Habitat Blockages..................................................................................101

7.2.2. Water Development................................................................................102

7.2.3. Water Conveyance and Flood Control ..................................................105

7.2.4. Land Use Activities................................................................................107

7.2.5. Water Quality.........................................................................................117

7.2.6. Pesticides ...............................................................................................118

7.2.7. Baseline Water Temperature - Clean Water Act ...................................120

7.2.8. Hatchery Operations and Practices.......................................................124

7.3. Baseline Habitat Condition ................................................................................. 126

7.4. Baseline Pesticide Detections in Aquatic Environments..................................... 128

7.5. Baseline Pesticide Consultations......................................................................... 131

7.6. Rice Production and the Use of Thiobencarb in California................................ 133

7.6.1. Water Management................................................................................138

7.6.2. Drift Minimization .................................................................................141

7.6.3. Applicator Education.............................................................................142

8. Effects of the Proposed Action to Threatened and Endangered Pacific 

Salmonids.................................................................................................... 144

8.1. Exposure Analysis................................................................................................ 144

8.1.1. Threatened and Endangered Pacific Salmonids use of Aquatic Habitats 145

8.1.2. Chemical Exposure Pathways to Salmonids Habitats...........................147

8.1.3. Summary of Chemical Fate of Thiobencarb ..........................................148

8.1.4. Exposure of salmonid habitats to the stressors of the action.................152

8.1.5. Analysis of Species Presence in Rice Growing Areas ...........................159

8.2. Modeling: Estimates of exposure to thiobencarb ............................................... 171

8.2.1. EPA exposure estimates.........................................................................171
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8.2.2. Utility of EPA-derived EECs for defining exposure to Pacific salmonid habitats

..........................................................................................................................171

8.2.3. NMFS exposure estimates for floodplain habitats.................................173

8.2.4. NMFS exposure estimates for pesticide mixtures..................................175

8.3. Monitoring Data: Measured Concentrations of Parent Compounds in Surface 

Waters......................................................................................................................... 176

8.3.1. Monitoring data considerations.............................................................177

8.3.2. USGS NAWQA Data..............................................................................180

8.3.3. Targeted monitoring of thiobencarb during applications .....................196

8.3.4. Exposure to Other Action Stressors.......................................................202

8.4. Exposure Conclusions.......................................................................................... 205

8.5. Response Analysis................................................................................................ 211

8.5.1. Thiobencarb’s Mode and Mechanism of Action ....................................212

8.5.2. Temperature and toxicity .......................................................................213

8.5.3. pH and toxicity.......................................................................................213

8.5.4. Herbicide effects to salmonids and their habitats..................................214

8.5.5. Importance of plants and other photosynthetic organisms in fueling secondary 

production within salmonid habitats ...............................................................215

8.5.6. Effects of herbicides on non-target aquatic communities......................217

8.5.7. Challenges in scaling up effects and making predictions across salmonid 

habitats.............................................................................................................221

8.5.8. Toxicity of Thiobencarb (Assessment Endpoints)..................................227

8.5.9. Indirect Effects to Salmonids (Prey effects and Habitat Modifications)240

8.5.10. Degradate Toxicity...............................................................................244

8.5.11. Tank Mixtures ......................................................................................245

8.5.12. Adjuvant Toxicity .................................................................................247

8.5.13. Uncertainties and Data Gaps Identified from Review of Available Toxicity 

Information for Thiobencarb ...........................................................................249

8.5.14. Evaluation of Data Available for Response Analysis..........................249

9. Risk Characterization ............................................................................. 252
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9.1. Exposure and Response Integration .................................................................... 254

9.2. Evaluation of Risk Hypotheses ............................................................................ 257

9.2.1. Risk Hypotheses.....................................................................................262

9.2.2. Population Level Responses ..................................................................273

10. Cumulative Effects................................................................................ 277

11. Integration and Synthesis for Listed Species........................................ 281

12. Effects of the Proposed Action on Designated Critical Habitat .......... 289

12.1. Risk Characterization ........................................................................................ 294

Summary of the Effects of the Action on PCEs................................................301

12.2. Integration and Synthesis for Designated Critical Habitat............................... 303

13. Conclusion ............................................................................................ 307

14. Incidental Take Statement..................................................................... 308

14.1. Amount or Extent of Take .................................................................................. 308

14.2. Reasonable and Prudent Measures ................................................................... 310

14.3. Terms and Conditions........................................................................................ 311

14.4 Conservation Recommendations..................................................................... 318

14.5 Reinitiation Notice .......................................................................................... 318

15. References............................................................................................. 320

16. Appendix 1 Abbreviations / Acronyms.................................................. 350

17. Appendix 2: Glossary .......................................................................... 357

18. Appendix 3: Median and 95th percentile rate of thiobencarb application 

to rice in California (1980-2010)............................................................... 366

19. Appendix 4: Co-Application of Thiobencarb with Other Pesticides in 

California (1999-2010)............................................................................... 368

20. Appendix 5: Toxicity of Eleven Pesticides to Embryonic Zebrafish ... 380

21. Appendix 6: Conservation Values of Designated Critical Habitat for 

Listed Salmonids in California’s Central Valley........................................ 400
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List of Tables

Table 1. EPA’s effects determinations for thiobencarb and molinate on Pacific salmon. 17

Table 2. Thiobencarb use patterns in the action area....................................................... 27

Table 3. Salmonid life stage and habitat assessment endpoints and measures................ 36

Table 4. Essential physical and biological features of PCEs in salmonid critical habitat 

designations....................................................................................................................... 39

Table 5. Listed Species and Critical Habitat in the Action Area. .................................... 46

Table 6. Species and population annual growth rates (Good et al., 2005) ...................... 48

Table 7. Temporal occurrence of adult and juvenile CV Spring-run Chinook salmon in 

the Sacramento River. Darker shades indicate months of greatest relative abundance. 

Note: Yearling rear in their natal streams through the first summer following their birth. 

Downstream migration generally occurs the following fall and winter. Young of year 

migrate during the first spring after they hatch................................................................. 59

Table 8. Criteria for assessing the level of risk of extinction for populations of Pacific 

salmonids in the Central Valley of California. Overall risk is determined by the highest 

risk score for any category................................................................................................ 62

Table 9. Central Valley Spring-run Chinook salmon--preliminary population structure, 

historic and most recent natural production, spawner abundance, and hatchery 

contributions (Good et al. 2005, USFWS and Reclamation 2007, Williams et al. 2011). 63

Table 10. CV Spring-run Chinook salmon CALWATER HSA watersheds with 

conservation values........................................................................................................... 68

Table 11. Temporal occurrence of adult and juvenile winter-run Chinook salmon in the 

Sacramento River. Darker shades indicate months of greatest relative abundance......... 71

Table 12. Temporal occurrence of adult and juvenile CCV steelhead in the Central 

Valley. Darker shades indicate months of greatest relative abundance........................... 81

Table 13. CCV spring-run Chinook salmon CALWATER HSA watersheds with 

conservation values........................................................................................................... 87

Table 14. Generalized life-history timing for listed salmonids in the upper Sacramento 

River................................................................................................................................ 104
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Table 15. Land uses and population density in the Sacramento and San Joaquin 

watersheds (Carter and Resh 2005). ............................................................................... 107

Table 16. Area of land use categories within the range of listed salmonids in the Central 

Valley (km2). Total area for each category is given in bold. Land cover was determined 

via the National Land Cover Database 2006, developed by the Multi-Resolution Land 

Characteristics (MRLC) Consortium, comprised of nine federal agencies (USGS, EPA, 

USFS, NOAA, NASA, BLM, NPS, NRCS, and USFWS). Land cover class definitions 

are available at: http://www.mrlc.gov/nlcd_definitions.php . Rice information was 

obtained from the California Rice Commission (2010 values) and from the Cal. DPR 

pesticide use reporting data base (2010 county reports)................................................. 114

Table 17. Kilometers of river, stream, and estuary included in state 303(d) lists due to 

temperature. Data from most recent GIS layers available from CA 2010..................... 124

Table 18. Consultations with EPA's proposed FIFRA actions. Past Opinion's jeopardy 

and adverse modification determinations pertaining to California Central Valley ESA 

listed Pacific salmonids................................................................................................... 132

Table 19. California DPR thiobencarb water management requirements summary. .... 138

Table 20. RPP Monitoring Sites (CRWQCB-CVR 2005)............................................. 141

Table 21. General life histories of Pacific salmonids which utilize habitat that overlaps 

with thiobencarb spatial use patterns. ............................................................................. 146

Table 22. Environmental fate characteristics of thiobencarb1. ...................................... 149

Table 23. Degradates of thiobencarb (EPA 2009b)....................................................... 150

Table 24. Concentration of thiobencarb and metabolites reported in field dissipation 

studies ............................................................................................................................. 150

Table 25. Co-occurrence of listed Pacific salmonids with potential application of 

pesticides to use sites within the salmonids' freshwater distribution.............................. 153

Table 26. Generalized run-timing of ESA-listed Central Valley Pacific Salmonids by life 

stage. ............................................................................................................................... 155

Table 27. Temporal occurrence of adult and juvenile CV spring-run Chinook salmon in 

the Sacramento River. Darker shades indicate months of greatest relative abundance in 

each location along the river........................................................................................... 162
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Table 28. CV spring-run Chinook relative abundance in rice areas (Figure 25) during 

thiobencarb use periods April - July. .............................................................................. 163

Table 29. Temporal occurrence of adult and juvenile winter-run Chinook salmon in the 

Sacramento River. Darker shades indicate months of greatest relative abundance in each 

location along the river. .................................................................................................. 166

Table 30. Sacramento River winter-run Chinook relative abundance in rice areas (Figure 

25) during thiobencarb use periods April – July............................................................. 166

Table 31. Temporal occurrence of adult and juvenile CV steelhead in the Central Valley. 

Darker shades indicate months of greatest relative abundance in each location. ........... 170

Table 32. CV steelhead relative abundance in rice areas (Figure 25) during thiobencarb 

use periods April - July................................................................................................... 170

Table 33. Estimated average initial thiobencarb concentrations in a floodplain habitat 

that is 2m wide and of variable depths using AgDrift 2.0.05. ........................................ 174

Table 34. AgDrift Estimates for point deposition of thiobencarb applied at the 4 lbs a.i./A 

rate................................................................................................................................... 175

Table 35. Estimated average initial pesticide concentration in a floodplain habitat that is 

2m wide and 0.1m deep using AgDrift 2.0.05................................................................ 176

Table 36. Number of NAWQA sample sites within the freshwater distribution of listed 

Pacific salmonids in the Central Valley of California, as determined through GIS analysis

......................................................................................................................................... 181

Table 37. Detections and concentrations of thiobencarb reported in the NAWQA 

database........................................................................................................................... 182

Table 38. Number of CDPR database sample sites within the distribution of listed Pacific 

salmonids in California, as determined through GIS analysis........................................ 185

Table 39. Detections and concentrations of thiobencarb from the CDPR database ....... 187

Table 40. Summary of Irrigated Lands Study data from the Central Valley Regional 

Water Quality Control Board (2004-2006)..................................................................... 189

Table 41. Annual monitoring results for thiobencarb at stations associated with rice 

cultivation (1995-2002; LRL=0.5 ug/L)......................................................................... 192
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Table 42. Annual monitoring results for thiobencarb at stations associated with rice 

cultivation (2003-2011; Detection Limit 0.5 or 0.1 ug/L).............................................. 192

Table 43. Results of County Agricultural Commissioner's monitoring of seepage 

associated with thiobencarb applications and compliance with state regulations (2006-

2010). .............................................................................................................................. 195

Table 44. Results of County Agricultural Commissioner's monitoring of compliance with 

state regulations associated with the use of thiobencarb (2006-2010). .......................... 196

Table 45. Texas field studies that evaluated offsite drift and runoff with thiobencarb in 

three tidally influenced bayous (Lauck 1979) ................................................................ 198

Table 46. Registrant submitted field studies in Arkansas and Texas that evaluated off-site 

drift and runoff with thiobencarb and thiobencarbsulfoxide (MRID 434040-03).......... 200

Table 47. Examples of thiobencarb product ingredients. .............................................. 203

Table 48. Detection and concentrations of nonionic detergent degradates in streams of 

the U.S. (Koplin et al 2002)............................................................................................ 203

Table 49. Chemical exposure data ranges in monitoring data and modeling. ................ 206

Table 50. Relative abundance of listed salmonids near rice fields during the period when 

thiobencarb is applied and water is discharged from thiobencarb treated fields (April -

July)................................................................................................................................. 209

Table 51 Thiobencarb toxicity values (μg/L) for aquatic organisms and plants reported in 

EPA salmonid BE, CRLF BE, RED, IRED, EFED science chapter, and ECOTOX. 

Abbreviations as follows: a.i. = active ingredient; NR = Not Reported; T= Technical 

grade; F = Formulated product; sw = estuarine/marine species; [ ] = 95% Confidence 

interval. ........................................................................................................................... 229

Table 52 Study designs and results with aquatic species................................................ 241

Table 53. Summary of Toxicity Data for Thiobencarb................................................... 250

Table 54. Risk Hypotheses ............................................................................................. 261

Table 55. Summary of determinations to risk hypotheses............................................. 272

Table 56. Relative abundance of listed salmonids in rice areas during thiobencarb use 

periods, April-July. ......................................................................................................... 274

Table 57. Salmonid risk hypotheses and population level effects. ................................. 276
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Table 58. Population growth rates among major rice growing counties. ...................... 278

Table 59. Key Findings of Effects of the Proposed Action Section ............................... 284

Table 60. Summary Findings for Species....................................................................... 288

Table 61. Thiobencarb’s assessment endpoint toxicity values (μg/L) for saltwater aquatic 

organisms presented in California Red-Legged Frog (CRLF) BEs................................ 300

Table 62. Risk to the primary constituent elements (PCEs) from stressors of the proposed 

action............................................................................................................................... 302

Table 63. Key Findings from the Effects of the Proposed Action Section to Critical 

Habitat............................................................................................................................. 304

Table 64. Summary findings for Designated Critical Habitats....................................... 306

Table 65. Conclusions for EPA's Reregistration of Thiobencarb ................................... 307
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List of Figures

Figure 1. Stressors of the Action...................................................................................... 21

Figure 2. The Central Valley steelhead ESU encompasses the Sacramento River winterrun Chinook, and the Central Valley spring-run Chinook ESUs. Rice is only grown 

within these boundaries, consequently, the CV steelhead ESU comprises the area of 

interest for this consultation.............................................................................................. 30

Figure 3. Conceptual framework for assessing risks of EPA’s action to ESA listed 

resources. .......................................................................................................................... 41

Figure 4. Exposure pathways for stressors of the action, and general response of Pacific 

salmonids and habitat........................................................................................................ 42

Figure 5. Central Valley Spring-run Chinook salmon distribution.................................. 57

Figure 6. Central Valley Spring-run Chinook salmon Conservation Values per Sub-Area.

........................................................................................................................................... 69

Figure 7. Sacramento River Winter-run Chinook salmon distribution............................ 73

Figure 8. Estimated yearly adult natural production and in-river adult escapement of 

winter-run from 1967 - 2007 based on RBDD ladder counts........................................... 75

Figure 9. CCV steelhead distribution............................................................................... 83

Figure 10. California Central Valley Steelhead Conservation Value per Sub-area.......... 90

Figure 11. Landuse in California (National Land Cover Database 2006). .................... 109

Figure 12. CV spring-run Chinook ESU with associated land-use patterns.................. 110

Figure 13. Sacramento River winter-run Chinook ESU with associated land use patterns.

......................................................................................................................................... 111

Figure 14. CV steelhead DPS with associated land use patterns................................... 112

Figure 15. California dams and NPDES permit sites. ................................................... 121

Figure 16. California 303(d) list: water bodies and stream segments included in the 2010 

Integrated Report. ........................................................................................................... 122

Figure 17. 2011 Sacramento Valley Rice Acres by County (CRC 2011). .................... 133

Figure 18. Central Valley rice growing overlay with listed CV spring-run Chinook ESU.

......................................................................................................................................... 134
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Figure 19. Central Valley rice growing overlay with Sacramento River winter-run 

Chinook ESU. ................................................................................................................. 135

Figure 20. Central Valley rice growing overlay with CV steelhead DPS (note change in 

scale from previous figures)............................................................................................ 136

Figure 21. Rice Pesticide Program (RPP) Monitoring Sites.......................................... 140

Figure 22. Exposure analysis......................................................................................... 145

Figure 23. Chemical structure of thiobencarb................................................................. 148

Figure 24. Number of thiobencarb applications in California by month from 1995-2010 

(source: CDPR PUR database). Points in figure represent the number of application for a 

year. Months with zero applications are not plotted. The solid horizontal line represents 

the monthly mean number of applications over all of the years. .................................... 154

Figure 25. Distribution of rice in the Sacramento Valley based on the 2010 NASS Crop 

Data Layer. North Delta, Knights Landing (KL) and Red Bluff Diversion Dam (RBDD) 

fish sampling sites and distribution of the Central Valley Spring Run Chinnok are also 

shown. ............................................................................................................................. 157

Figure 26. CV spring-run Chinook probable and known spawning areas in relation to 

2010 rice growing areas. ................................................................................................. 160

Figure 27. Sacramento River winter-run Chinook spawning areas relative to rice growing 

areas. Spawning is confined to the 44 river miles immediately downstream of Keswick 

Dam (Yoshiyama et al. 1998). In California, Rice is primarily cultivated in the 

Sacramento Valley downstream of the Red Bluff Diversion Dam (RBDD). ................. 164

Figure 28. CV steelhead probable and known spawning areas in relation to 2010 rice 

growing areas.................................................................................................................. 169

Figure 29. Acres treated with thiobencarb in California. Data from the California 

Pesticide Use Reporting Database 1980 – 2010. ............................................................ 179

Figure 30. Distribution of NAWQA monitoring sites that have sampled for the presence 

of thiobencarb relative to the range of threatened and endangered Pacific salmonids in 

California. ....................................................................................................................... 180
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Figure 31. Distribution of CDPR monitoring sites that have sampled for the presence of 

thiobencarb relative to the range of threatened and endangered Pacific salmonids in 

California. ....................................................................................................................... 184

Figure 32. The location of seven sample sites accounting for 95% of the detections in 

thiobencarb in the CDPR monitoring database............................................................... 186

Figure 33. Sacramento Valley RPP monitoring locations (source: Moran 2012 

modification of figure from CRC 2004 RPP Report)..................................................... 190

Figure 34. Response Analysis Conceptual Model .......................................................... 211

Figure 35. Part I of a conceptual model of potential effects of herbicides to aquatic 

communities. This figure focuses on potential effects of herbicides applied to riparian 

areas adjacent to salmonid habitats. Bolded arrows and text note those effects that are 

most likely to occur based on the frequency that they are reported in the literature. ..... 224

Figure 36. Part II of a conceptual model of potential effects of herbicides to aquatic 

communities. This figure focuses on potential effects of herbicides that are applied to or 

otherwise reach salmonid habitats. Bolded arrows and text note those effects that are 

most likely to occur based on the frequency that they are reported in the literature. ..... 225

Figure 37. Species sensitivity distribution of freshwater fish 96 h LC50s derived from 

EPA denoted "acceptable" studies.................................................................................. 235

Figure 38. Schematic of Risk Characterization Phase. .................................................. 252

Figure 39. Thiobencarb exposure concentrations and effect concentrations for assessment 

endpoints......................................................................................................................... 256

Figure 40. Assessment Framework for Designated Critical Habitat Assessment 

Framework for Designated Critical Habitat.................................................................... 290

2

Marine Fisheries Service (NMFS) or the U.S. Fish and Wildlife Service (USFWS)2, 

depending upon the endangered species, threatened species, or designated critical habitat 

that may be affected by the action (50 CFR §402.14(a)). Federal agencies are exempt 

from this general requirement if they have concluded, with written concurrence from the 

U.S. Fish and Wildlife Service, NMFS or both, that an action “may affect but is not likely 

to adversely affect” endangered species, threatened species or designated critical habitat 

(50 CFR §420.14(b)).

The United States (U.S.) Environmental Protection Agency (EPA) initiated consultation 

with NMFS on its proposals to authorize use, pursuant to the Federal Insecticide, 

Fungicide, and Rodenticide Act (FIFRA), 7 U.S.C. 136 et seq., of pesticide products 

containing the active ingredients (a.i.) thiobencarb and molinate from August 1, 2002. 

EPA authorization of pesticide uses are categorized as FIFRA sections 3 (new product 

registrations), 4 (re-registrations and special review), 18 (emergency use), or 24(c) 

[Special Local Needs (SLN)]. At that time, EPA determined that uses of pesticide 

products containing molinate and thiobencarb would have “no effect” on most listed 

salmonids and “may affect, but is not likely to adversely affect” three species (Central 

Valley spring-run Chinook, Sacramento River winter-run Chinook, and California 

Central Valley steelhead) and designated critical habitat for these ESUs. NMFS does not 

concur with EPA’s three not likely to adversely affect determinations and conducted 

formal consultation. This document represents NMFS’ biological opinion (Opinion) on 

the impacts of EPA’s authorization of pesticide products containing thiobencarb on the 

listed ESUs. This is a partial consultation intended to comply with a court order3

requiring EPA to make a determination on the effect of thiobencarb and 53 other active 

ingredients on listed Pacific Salmonids. 4 Consultation with NMFS will not be complete 

 

2 Generally, NMFS conducts consultation for marine and anadromous species, while FWS conducts 

consultations for freshwater and terrestrial species.

3 Northwest Coalition for Alternatives to Pesticides vs. National Marine Fisheries Service, Civ. NO 07-

1791 (W.D. Wa).

4 Two additional Pacific salmonids species have been listed since the court order. Although the court’s 

order did not address these two species, NMFS would have analyzed the effects of EPA’s action to them 

because they belong to the same taxon and require consideration of the same information. However, 
3

for registration of this a.i. until EPA makes effect determinations on all other species and 

designated critical habitat under NMFS jurisdiction and consults with NMFS as 

necessary.

This Opinion is prepared in accordance with section 7(a)(2) of the ESA and 

implementing regulations at 50 CFR Part 402. However, consistent with the decision in 

Gifford Pinchot Task Force v. USFWS, 378 F.3d 1059 (Ninth Cir. 2004), we did not 

apply the regulatory definition of “destruction or adverse modification of critical habitat” 

at 50 CFR §402.02. Instead, we relied on the statutory provisions of the ESA to complete 

our analysis of the effects of the action on designated critical habitat.

This Opinion is based on NMFS’ review of the package of information the EPA 

submitted with its 2003 and 2004 requests for formal consultation on the proposed 

authorizations of the above a.i.s. It also includes our review of recovery plans for listed 

Pacific salmonids, past and current research, monitoring reports from prior research, 

previous Opinions , published and unpublished scientific information on the biology and 

ecology of threatened and endangered salmonids in the action area, and other sources of 

information gathered and evaluated during the consultation on the proposed authorization 

of the a.i. thiobencarb. NMFS also reviewed pesticide labels, available monitoring data 

and other local, county, and state information, online toxicity databases, incident reports, 

data generated by pesticide registrants, and exposure models run by NMFS. NMFS also 

considered information and comments provided by EPA, by the registrants identified as 

applicants by EPA and information and comments submitted by others during EPA’s 

public comment process. 

 

since both of these species are outside of the area considered by NMFS because of limited use of 

thiobencarb (see Consultation History, infra), NMFS did not evaluate the effects of thiobencarb application 

to these two newly listed species. 
4

1. Background

On January 30, 2001, the Washington Toxics Coalition, Northwest Coalition for 

Alternatives to Pesticides, Pacific Coast Federation of Fishermen’s Associations, and 

Institute for Fisheries Resources filed a lawsuit against EPA in the U.S. District Court for 

the Western District of Washington, Civ. No. 01-132. This lawsuit alleged that EPA 

violated section 7(a)(2) of the ESA by failing to consult on the effects of its continuing 

approval of 54 pesticide a.i.s on 26 listed Pacific salmonid ESUs.

On July 2, 2002, the court ruled that EPA had violated ESA section 7(a)(2) and ordered 

EPA to initiate interagency consultation and make determinations regarding effects of all 

54 a.i.s on listed salmonids by December 2004. Washington Toxics Coalition v. EPA, 

C01-132C (W.D. Wash. 7/2/2002).

On January 22, 2004, the court enjoined application of pesticides within 20 (for ground) 

and 100 (for aerial) feet (ft) of streams supporting salmon. Washington Toxics Coalition 

v. EPA, C01-132C (W.D. Wash. 1/22/2004). The court imposed several additional 

restrictions on pesticide use in specific settings.

On November 5, 2007, the Northwest Coalition for Alternatives to Pesticides and others 

filed a legal complaint in the U.S. District Court for the Western District of Washington, 

Civ. No. 07-1791, against NMFS for its unreasonable delay in completing the section 7 

consultations for EPA’s registration of 54 pesticide a.i.s.

On July 30, 2008, NMFS and the plaintiffs entered into a settlement agreement with the 

Northwest Coalition for Alternatives to Pesticides. NMFS agreed to complete 

consultation within four years on 37 a.i.s. EPA had concluded that 17 of the 54 a.i.s at 

issue in the first litigation would not affect any listed salmonid species or any of their 

designated critical habitat, and so did not initiate consultation on those a.i.s.
5

On November 18, 2008, NMFS issued its first Opinion for three organophosphates: 

chlorpyrifos, diazinon, and malathion.

On April 20, 2009, NMFS issued its second Opinion for three carbamates: carbaryl, 

carbofuran, and methomyl. 

On August 31, 2010, NMFS issued its third Opinion. This third consultation evaluated 

12 organophosphate insecticides: azinphos methyl, bensulide, dimethoate, disulfoton, 

ethoprop, fenamiphos, methamidophos, methidathion, methyl parathion, naled, phorate, 

and phosmet. 

On June 30, 2011, NMFS issued its fourth Opinion. This fourth consultation evaluated 

four herbicides: 2,4-D, triclopyr BEE, diuron and linuron; and 2 fungicides: captan and 

chlorothalonil. 

On May 31, 2012, NMFS issued its fifth Opinion. This fifth consultation evaluated 

herbicides: oryzalin, trifluralin, and pendimethalin.

2. Consultation History

On August 1, 2002 EPA transmitted a letter to NMFS’ Office of Protected Resources 

(OPR) requesting section 7(a)(2) consultation for the registration of the two a.i.’s: 

thiobencarb and molinate providing their effects determinations on 26 ESUs of Pacific 

salmonids listed at that time (Puget Sound steelhead and Lower Columbia River coho 

were not evaluated). In the BE, and summarized in Table 1, EPA’s Office of Pesticide 

Programs (OPP) determined that the use of thiobencarb and molinate may affect, but is 

not likely to adversely affect, three salmonid species: Central Valley (CV) spring-run 

Chinook, Sacramento River winter-run Chinook ESUs, and CV steelhead DPS. EPA 

determined that these products would have no effect on the remaining salmonid ESUs 

and DPSs.
6

On June 28, 2005, NMFS listed the Lower Columbia River coho salmon ESU as 

threatened. 

On May 22, 2007, NMFS listed the Puget Sound Steelhead Distinct Population Segment 

(DPS) as threatened. 

On December 10-12, 2007, EPA, NMFS and the USFWS met and discussed approaches 

for moving forward with ESA consultations on pesticide registrations. The agencies 

agreed to develop methodologies for filling existing data gaps. In the interim, NMFS and 

USFWS agreed to develop approaches to address these gaps within their Opinions. The 

agencies identified communication and coordination mechanisms to address technical 

and policy issues and procedures for conflict resolution.

On February 11, 2008, NMFS listed the Oregon Coast coho salmon evolutionarily 

significant unit (ESU) as threatened. This ESU was considered in EPA’s Biological 

Evaluation for the a.i.s. in the five previous Opinions referenced in the Background

section above.

On August 20, 2008, NMFS met with EPA and requested EPA to identify applicants for 

this and subsequent pesticide consultations. 

On September 17, 2008, NMFS requested EPA approval of Confidential Business 

Information (CBI) clearance for certain staff members in accordance with FIFRA 

regulations and access to EPA’s incident database so NMFS staff could evaluate CBI 

materials from the applicants and incident reports for the a.i.s under consultation. EPA 

conveyed to NMFS that no access to the incident database would be authorized and the 

reports would be sent directly from EPA to NMFS.

On September 23, 2008, NMFS staff received notification of CBI clearance from EPA.
7

On September 26, 2008, NMFS sent correspondence to EPA regarding the roles of the 

federal action agency and applicants identified by such agency during formal 

consultation. NMFS also requested incident reports and label information for this and 

subsequent pesticide consultations from EPA. The specified timeline for NMFS’ receipt 

of incident reports and label information for the two a.i.s (thiobencarb and molinate) 

considered in this Opinion was December 1, 2010.

On December 7, 2010, EPA sent to NMFS via email incident monitoring data for the 

remaining chemicals, including thiobencarb.

On December 8, 2010, NMFS received notification from EPA that registration of 

molinate and another a.i. (lindane) scheduled to be addressed in a subsequent Opinion 

were canceled. EPA and NMFS agreed that consultation on label reviews for molinate 

and lindane would not be necessary. EPA provided NMFS an internal EPA memo dated 

July 14, 2008 that stated no use of molinate will be permitted after August 31, 2009.

On December 20, 2010, NMFS sent EPA (via email) a request for the label on thiobenarb 

along with a copy of the Sept. 26, 2008, letter requesting that EPA provide NMFS 

incident data and labels for the remaining a.i.s by December 1, 2010. 

On January 10, 2011, EPA called NMFS to give an update on the status of the labels for 

the last batch of a.i.s. They also indicated that they would not provide labels for molinate 

and lindane because their registration had been canceled.

On January 11, 2011, NMFS requested from EPA (via email) final cancelation orders 

documenting cancellation of all lindane and molinate products registered by EPA.

On January 11, 2011, EPA provided to NMFS final cancelation orders documenting the 

cancellation of lindane and molinate.
8

On March 1, 2011, NMFS received preliminary thiobencarb labels from EPA. EPA said 

that they would send official labels later.

On April 27, 2011, NMFS received the official thiobencarb labels from EPA.

On June 29, 2011, EPA and NMFS received a letter from Valent U.S.A. Corporation 

(Valent), which requesting they be considered an “applicant” for purposes of EPA’s ESA 

consultation with NMFS regarding the effects of thiobencarb on threatened and 

endangered Pacific salmonids and their designated critical habitat.

On July 18, 2011, NMFS sent an email to EPA seeking clarification on thiobencarb uses. 

NMFS noticed in EPA’s Registration Eligibility Decision (RED), thiobencarb was listed 

for other uses (e.g., celery, lettuce, endive). 

On July 19, 2011, EPA confirmed (via email) that rice is the only use approved for 

thiobencarb on the west coast, within the range of the listed Pacific salmonids.

On July 19, 2011, NMFS met with applicant Valent and EPA to talk about the 

thiobencarb consultation and to exchange general information to assist with the 

consultation process. Valent, EPA and NMFS each provided presentations at this 

meeting. NMFS confirmed with Valent that thiobencarb is only approved for rice on the 

west coast, and confirmed that they are the only registrant for thiobencarb. Valent 

provided NMFS thiobencarb monitoring data and suggested it reflected current usage, 

stewardship programs, and permit conditions.

During this meeting, NMFS and EPA agreed that the consultation on thiobencarb would 

be based not on all federally authorized uses, as agreed upon in the December 2007 

meeting and as occurred for NMFS’ four previous pesticide biological opinions, but on a 

subset of authorized use, that of rice growing in California. This decision was based on 

the fact that, unlike for other pesticide active ingredients, there is currently no authorized 
9

use of thiobencarb in Washington, Oregon or Idaho. EPA and NMFS agreed that EPA 

would reinitiate consultation in the unlikely event that Washington, Oregon or Idaho 

authorized use of thiobencarb or that rice growing expanded beyond the identified 

counties in California. NMFS’ evaluation will consider the impacts of application of 

thiobencarb up to the maximum labeled amount.

On August 25, 2011, NMFS sought to confirm that thiobencarb is only permitted for use 

on rice and that there are no thiobencarb products labeled for use in Washington or 

Oregon. This email was sent because monitoring data showed recent detections of 

thiobencarb in both states.

On September 15, 2011, EPA confirmed that there are no uses of thiobencarb in 

Washington or Oregon. EPA suggested thiobencarb detections in the monitoring data 

were either from misuse, atmospheric transport, or errors in the monitoring data.

On September 29, 2011, Valent sent NMFS (via email) a list of studies they had 

submitted to EPA. Valent indicated that they did not have information to characterize the 

potential for off-target drift of aerially applied granular products of thiobencarb. They 

stated that the drift potential for aerially-applied granular formulations is “significantly 

less” than for liquid formulations, however, they could provide no comparisons giving 

numerical values.

On September 30, 2011, NMFS requested (via email) EPA provide copies of the 15 

studies Valent referenced in their September 29 email to NMFS.

On December 7, 2011, NMFS received (via email) one of the studies (MRID 43404005) 

requested on September 30, 2011.

On December 9, 20ll, NMFS received (via email) eight (MRID 46091402, MRID 

46091401, MRID 42460401, MRID 43252001, MRID 42257801, MRID 45695101, 
10

MRID 44628601, MRID 44628602) additional studies NMFS requested on Sept. 30, 

2011. 

On December 9, 2011, NMFS sent a request (via email) to EPA for copies of the 

remaining six studies referenced in Valent’s September 29, 2011 email.

On January 5, 2012, NMFS sent an email to EPA to ask about EPA’s assessment of risks 

of thiobencarb to the California red-legged frog and Delta smelt. 

On January 6, 2012, in response to NMFS Jan. 5, 2012, email, EPA explained that in the 

course their routine data quality review for risk assessment, the half-lives were 

recalculated from the raw data provided in the studies, using a non-linear first-order 

equation. EPA used this recalculation of half-lives in the effect determination, rather 

than the face value reported in the studies.

On January 9, 2012, NMFS sent EPA follow up questions and asked for additional 

information on the method EPA used for the recalculations. NMFS also requested a copy 

of the Louisiana study (MRID 434040-04) referenced in EPA’s Jan. 6, 2012 email. EPA 

provided a response to NMFS’ questions. NMFS sent EPA an email seeking 

clarifications to responses provided by EPA. 

On January 10, 2012, NMFS received from EPA (via email) a copy of the Louisiana 

study referenced above (MRID 42003404).

On January 30, 2012, in response to an earlier inquiry by NMFS, the California Rice 

Commission (CRC) confirmed that representatives would be available to meet with 

NMFS in Sacramento on February 28, 2012 to discuss the CRC role and to introduce 

NMFS to rice growers in the region.
11

On February 7, 2012, NMFS received from EPA (via email) the remaining six studies

(MRID 42680401, MRID 43252001 [already received on Dec. 9, 2011], MRID 

43404001, MRID 43404004, MRID 42384701, MRID 43404003) requested by NMFS 

September 30, 2011 and December 9, 2011.

On February 13, 2012, NMFS requested EPA to send MRID 25179 (Lauck, JE. 1979) 

Final Report of Field Study: Ortho Bolero 8 EC-- Rice, 1979. (Unpublished study 

received Dec 11, 1979 under 239- 2450; submitted by Chevron Chemical Co., Richmond, 

Calif.; CDL:241490-F).

On February 14, 2012, NMFS requested (via email) another study from EPA (MRID 

79986) cited in EPA’s California Red-legged frog assessment.

On February 15, 2012, EPA informed NMFS (via email) that the studies requested on 

Feb 13 and Feb 14 were old studies and that it may take a week to get an electronic copy.

NMFS responded by email back to EPA that these studies were important as they 

represent some of the higher concentrations and it was critical NMFS evaluate the study 

design.

On February 21, 2012, the court in Northwest Coalition for Alternatives to Pesticides v. 

NMFS, Civ. No. 07-1791 (W.D. WA) granted NMFS’ request, agreed to by the plaintiffs, 

to extend the schedule for completion of consultation established in the July 30, 2008 

settlement agreement. The deadline for this Biological Opinion was extended until June 

30, 2012.

On February 21, 2012, NMFS notified EPA and Valent that NMFS was granted an 

extension to June 30, 2012, for the thiobencarb final Biological Opinion. 

On February 23, 2012, EPA sent NMFS an email confirming that Bolero Ultramax (a 

thiobencarb formulation) is the same as Bolero 15G. 
12

On February 27, 2012, NMFS staff working on this Opinion flew to Sacramento to gather 

information about rice growing, water monitoring, rice regulation and over-sight, and 

Sacramento River salmonids. NMFS gave a webinar presentation to California 

Department of Pesticide Regulation staff, Valent representatives, and others on the 

process used to develop NMFS Opinions on EPA FIFRA actions.

On February 28, 2012, NMFS met with a representative of the California Rice 

Commission (CRC). The CRC took NMFS staff on a tour of the rice areas including the 

Yolo Bypass, four of the five water monitoring stations where we observed how water 

samples are taken, and the Sacramento weir bypass. The CRC also arranged a meeting 

with a rice farmer who dry seeded rice and used the liquid formulation of thiobencarb.

On February 29, 2012, EPA sent to NMFS (via email) the two studies requested in 

NMFS February 13,14, and 15, 2012, email. 

On February 29, 2012, NMFS met again with the CRC who arranged to meet with a 

second rice farmer that wet seeded rice and used the granular formulation of thiobencarb. 

Later in the morning, NMFS met with California Department of Water Resources 

(CDWR) to gather information on salmon use on the Yolo Bypass floodplain. In the 

afternoon, NMFS met with California Department of Pesticide Regulation (CDPR) to 

gather information on rice pesticide use enforcement issues and rice pesticide modeling. 

NMFS also met with CDPR Endangered Species Division to learn more about 

California’s County Bulletins.

On March 1, 2012, NMFS met with U.S. Geological Survey (USGS) staff to gather 

information on water monitoring, results from recent studies related to thiobencarb use, 

and to discuss future data needs. Later in the day, NMFS met with CDWR and EPA 

Region 9 staff to discuss a new pesticide loading model developed for the Sacramento 

and San Joaquin rivers and Bay Delta.
13

On March 2, 2012, NMFS met again with USGS and California Department of Fish and 

Game (CDFG) staff to discuss salmonids’ temporal and spatial use of the Sacramento 

River. 

On March 14, 2012, Valent sent NMFS (via email) examples of state, regional, and 

county permit requirements for thiobencarb use in California. Seven documents were 

included in this email. 

Recommended Permit Conditions for Rice Pesticides (Rice Pesticide Program), 

obtained from Sutter County but written by the California Department of

Pesticide Regulation and the Central Valley Regional Water Quality Control 

Board for all rice thiobencarb use in California

Sutter County Rice Permit Conditions for Water Holding Requirements (Rice 

Water Holding Conditions) 2011

Butte County 2011 Rice Permit Conditions for Bolero/Abolish

Yolo County Rice Pesticide Drift Control Requirements

Yolo County Rice Water Seepage Management

Yolo County 2008 Rice Pesticide Permit Conditions Cover Letter

Yolo County 2008 Rice Pesticide Permit Conditions

On March 15, 2012, NMFS sent an email to Valent and EPA seeking clarification on the 

hold times for Abolish 8 EC liquid formulation of thiobencarb. NMFS also requested 

updated labels for these products. Valent responded (via email) by providing the 

California SLN label for Abolish 8 EC. Valent confirmed the water hold time.

On March 15, 2012, Valent provided NMFS (via email) relative use proportions of active 

ingredient thiobencarb used in California as Abolish 8 EC (liquid) and Bolero UltraMax 

(granular). Valent informed NMFS the ratio has been running approximately 30 percent 

thiobencarb as Abolish 8 EC and 70 percent thiobencarb as Bolero UltraMax.
14

On April 11, 2012, NMFS sent (via email) a draft portion of the Environmental Baseline

section of the Opinion to be CRC to review for accuracy. The CRC reviewed the section 

and returned it later this same day with suggested edits and additions. The CRC also 

provided additional information on rice water seepage management.

On May 9, 2012, NMFS sent (via email) a draft of the Reasonable and Prudent Measures 

and associated Appendix to EPA for review prior to a joint meeting to discuss on May 

10, 2012.

On May 10, 2012, NMFS met with EPA to discuss the draft Reasonable and Prudent 

Measures and corresponding Appendix.

On May 11, 2012, NMFS, provided a draft of this Opinion to EPA.

On May 22, 2012, NMFS met with EPA and Valent to discuss the draft Opinion. At this 

meeting, NMFS describe the process used to reach our conclusions. Valent then asked 

some questions and provided several constructive comments.

On June 1, 2012, EPA and NMFS received via email formal comments from Valent on 

the draft Opinion released on May 11, 2012. The comments reflected those made during 

the May 22, 2012 meeting.

On June 6, 2012, NMFS received via fax CBI information from EPA on Valent’s 

thiobencarb formulations.

On June 11, 2012, NMFS retrieved from EPA’s Docket (EPA-HQ-OPP-2008-0654)

comments on the draft Opinion from the California Rice Commission, and Valent 

comments previously received.
15

On June 13, 2012, NMFS received via fax additional CBI information from EPA on 

Valent’s thiobencarb formulations.

On June 14, 2012, EPA and NMFS received comments from California’s Department of 

Fish and Game on the draft Opinion released May 11, 2012.

On June 15, 2012, EPA and NMFS received comments from California’s Department of 

Pesticide Regulation on the draft Opinion released May 11, 2012.

On June 18, 2012, NMFS received via email partial comments from EPA on the draft 

Opinion released on May 11, 2012. The comments were on the draft RPMs. EPA and 

NMFS had a follow up phone call to discuss EPA’s issues with the draft RPM number 

one and related terms and condions. EPA and NMFS agreed to discuss the issue further. 

On June 21, 2012, NMFS and EPA met again to discuss again the draft RPMs. NMFS 

agreed to go through the CDPR permit requirements and incorporate into the RPM and 

the terms and conditions the specific requirements from the CDPR permit requirements 

and recommended best management practices intended to reduce thiobencarb loading to 

salmon waters.

On June 22, 2012, NMFS provided EPA via email an amended RPM based on discussion 

on June 18 and June 21.

On June 25, 2012, EPA sent an email to inform NMFS that they intended to share the

draft RPMs with CDPR.

On June 25, 2012, EPA, NMFS, and CDPR met via teleconference to discuss the RPMs. 

EPA had some clarifying questions of CDPR about the CDPR regulatory program and 

their enforcement.
16

On June 27, 2012, EPA and NMFS received a letter from Valent via email regarding 

draft RPMs.

2.1. Species Addressed in the BE

When an action agency concludes its action will not affect any listed species or critical 

habitat, section 7 consultation is not necessary (USFWS, & NMFS 1998). If an action 

agency concludes its action may affect but is not likely to adversely affect (NLAA) listed 

species, it seeks concurrence from NMFS or USFWS on the conclusion. When an action 

may adversely affect listed species or designated critical habitat, NMFS or USFWS 

conducts a formal consultation to determine whether that action is likely to jeopardize 

listed species or destroy or adversely modify critical habitat and issues a biological 

opinion describing their determinations. 

EPA’s BE considered the effects of pesticides containing thiobencarb and molinate to 26 

species of listed Pacific salmonids and their designated critical habitat (EPA undated 

Memorandum – Effects Determinations for Molinate and Thiobencarb for Pacific 

Anadromous Salmonids; August 1, 2002). Three listed species, Central Valley springrun Chinook, Sacramento River winter-run Chinook, and Central Valley steelhead were 

determined to be may affect, but not likely to adversely affect. All other ESUs and DPSs 

were determined to be no effect. Two species, Lower Columbia River coho and Puget 

Sound steelhead, were not considered in the BE. 

NMFS did not concur with any of EPA’s “NLAA” determinations for thiobencarb and 

determined that thiobencarb may adversely affect some ESUs or DPSs. Generally, once 

NMFS enters into formal consultation it considers impacts to all species and critical 

habitat.. In this Opinion, NMFS analyzed the impacts to all ESUs/DPSs of Pacific 

salmonids present in the action area, including those salmonid species identified by EPA 

as being unaffected, and including the two species of salmonids listed after EPA provided 

its BEs to NMFS. EPA’s effect determinations are summarized in Table 1. 
17

Table 1. EPA’s effects determinations for thiobencarb and molinate on Pacific salmon.

Species ESU or DPS

Herbicide

Thiobencarb Molinate

Chinook

Central Valley spring-run Chinook NLAA NLAA

Sacramento River winter-run Chinook NLAA NLAA

Steelhead

Central Valley NLAA NLAA

Puget Sound Not evaluated Not Evaluated

Coho Lower Columbia River Not evaluated Not evaluated

All other listed species ESUs and DPSs No effect No effect
18

3. Description of the Proposed Action

The proposed action encompasses EPA’s registration of the uses (as described by product 

labels) of all pesticides containing thiobencarb. The purpose of the proposed action is to 

provide tools for pest control that do not cause unreasonable adverse effects to the 

environment throughout the U.S. and its affiliated territories. Pursuant to FIFRA, before

a pesticide product may be sold or distributed in the U.S. it must be exempted or 

registered with a label identifying approved uses by EPA’s Office of Pesticide Programs 

(OPP). Once registered, a pesticide may not legally be used unless the use is consistent 

with directions on its approved label 

(http:www.epa.gov/pesticides/regulating/registering/index.htm). EPA authorization of 

pesticide uses are categorized as FIFRA sections 3 (new product registrations), 4 (reregistrations and special review), 18 (emergency use), or 24(c) Special Local Needs 

(SLN). 

On April 7, 2004, EPA published the final cancelation notice for molinate (69 FR 18368). 

Sale of end-use products containing molinate was prohibited after June 30, 2008, and all 

uses of those products were prohibited following August 31, 2009. As there are no 

remaining legal uses of molinate, there is no need for consultation and we will not 

provide an analysis of its effects to listed salmonids.

EPA’s pesticide registration process involves an examination of the ingredients of a 

pesticide, the site or crop on which it will be used, the amount, frequency and timing of 

its use, and its storage and disposal practices. Pesticide products may include a.i.s and 

other ingredients, such as adjuvants, and surfactants (described in greater detail below). 

The EPA evaluates the pesticide to ensure that it will not have unreasonable adverse 

effects on humans, the environment, and non-target species. An unreasonable adverse 

effect on the environment is defined in FIFRA as, “(1) any unreasonable risk to man or 

the environment, taking into account the economic, social, and environmental costs and 

benefits of the use of the pesticide, or (2) a human dietary risk from residues that result 
19

from a use of a pesticide in or on any food inconsistent with the standard under section 

408 of the FFDCA (21 U.S.C. §346a)” 7 U.S.C. 136(b).

 

After registering a pesticide, EPA retains discretionary involvement and control over 

such registration. EPA must periodically review the registration to ensure compliance 

with FIFRA and other federal laws (7 U.S.C. §136d). A pesticide registration can be 

canceled whenever “a pesticide or its labeling or other material does not comply with the 

provisions of FIFRA or, when used in accordance with widespread and commonly 

recognized practice, generally causes unreasonable adverse effects on the environment.” 

On December 12, 2007, EPA, NMFS, and FWS agreed that the federal action for EPA’s 

FIFRA registration actions will be defined as the “authorization for use or uses described 

in labeling of a pesticide product containing a particular pesticide ingredient.” In order to 

ensure that EPA’s action will not jeopardize listed species or destroy or adversely modify 

critical habitat, NMFS’ analysis necessarily encompasses the impacts to listed Pacific 

salmonid ESUs/DPSs of all uses authorized by EPA, regardless of whether those uses 

have historically occurred. 

Pesticide Labels. For this consultation, EPA’s proposed action encompasses all approved 

product labels containing the a.i. thiobencarb; its degradates, metabolites, and 

formulations, including other ingredients within the formulations; adjuvants; and tank 

mixtures. These activities comprise the stressors of the action (Figure 1). The BE 

assessing the impacts to listed salmon and steelhead indicates that the subject a.i. is

labeled for use on rice only (EPA 2002c). EPA provided no labels indicating any other 

use in California, Washington, Oregon or Idaho. This partial consultation differs 

substantially from other consultations NMFS has completed or will complete on pesticide 

registrations because thiobencarb is labeled for use on one crop only, and that crop is 

grown in identified counties in one area in California. It is unlikely rice production will 

expand beyond the current area because rice production needs certain environmental 

factors that are not widely found on the west coast. Three of the four states, Washington, 
20

Oregon and Idaho, currently prohibit the use of thiobencarb, and there is no known rice 

production in those states. In addition, California has imposed restrictions on 

thiobencarb use and conducts substantial monitoring to ensure that those restrictions are 

followed and effective. NMFS was therefore able to focus this consultation on the 

impacts of thiobencarb in the area where rice is grown, and evaluate the effects to three 

ESU/DPS present in that area and their designated critical habitat.5 Because there is no 

authorized use elsewhere in the four western states, none of the other of the 28 ESU/DPS 

of Pacific salmon or their critical habitat will be exposed to the effects of the action. 

NMFS will not consider these 25 ESU/DPS or any critical habitat designated for them 

further in this opinion. 

 

5 EPA does not control where rice is grown or state authorization or restrictions on thiobencarb use. 

However, in the July 19, 2011 meeting with NMFS and the applicant Valent, EPA has agreed to reinitiation 

triggers based on changes in any state’s authorization of thiobencarb for use on rice or expansion of rice 

production beyond its current location in California that would affect additional listed salmon and their 

critical habitat. Changes in the federal FIFRA labels that would authorize use in these four states on crops 

or use sites other than rice growing would also trigger reinitiation. 
21

Figure 1. Stressors of the Action

Active and Other Ingredients. Thiobencarb is the a.i. that kills or otherwise negatively 

affects targeted organisms (listed on the label). However, pesticide products that contain 

this a.i. also contain inert ingredients. Inert ingredients are ingredients that EPA defines 

as not “pesticidally” active. EPA also refers to inert ingredients as “other ingredients.”

The specific identification of the compounds that make up the inert fraction of a pesticide 

is not required on the label. However, this does not necessarily imply that inert 

ingredients are non-toxic, non-flammable, or otherwise non-reactive. EPA authorizes the 

use of chemical adjuvants to make pesticide products more efficacious. An adjuvant 

aides the operation or improves the effectiveness of a pesticide. Examples include 

wetting agents, spreaders, emulsifiers, dispersing agents, solvents, solubilizers, stickers, 

and surfactants. A surfactant is a substance that reduces surface tension of a system, 

allowing oil-based and water-based substances to mix more readily. A common group of 

Label-recommended tank mixtures

Metabolites and Degradates 

Active ingredients

Adjuvants/surfactants added to 

formulations

Registration and uses of pesticide labels 

Other ingredients in formulations
22

non-ionic surfactants is the alkylphenol polyethoxylates (APEs), which may be used in 

pesticides or pesticide tank mixes, and also are used in many common household 

products. Nonylphenol (NP), one of the APEs, has been linked to endocrine-disrupting 

effects in aquatic animals.

Formulations. Pesticide products come in a variety of solid and liquid formulations. 

Examples of formulation types include dusts, dry flowables, emulsifiable concentrates, 

granulars, solutions, soluble powders, ultra-low volume concentrates, water-soluble bags, 

powders, and baits. The formulation type can have implications for product efficacy and 

exposure to humans and other non-target organisms. 

Tank Mix. A tank mix is a combination by the user of two or more pesticide formulations 

as well as any adjuvants or surfactants added to the same tank prior to application. 

Typically, formulations are combined to reduce the number of spray operations or to 

obtain better pest control than if the individual products were applied alone. The 

compatibility section of a label may advise on tank mixes known to be incompatible or 

provide specific mixing instructions for use with compatible mixes. Labels may also 

recommend specific tank mixes. Pursuant to FIFRA, EPA has the discretion to prohibit 

tank mixtures. Applicators are permitted to include any combination of pesticides in a 

tank mix as long as each pesticide in the mixture is permitted for use on the application 

site and the label does not explicitly prohibit the mix.

Pesticide Registration. The Pesticide Registration Improvement Act (PRIA) of 2003 

became effective on March 23, 2004. The PRIA directed EPA to complete REDs for 

pesticides with food uses/tolerances by August 3, 2006, and to complete REDs for all 

remaining non-food pesticides by October 3, 2008. The goal of the reregistration 

program is to mitigate risks associated with the use of older pesticides while preserving 

their benefits. Pesticides that meet today’s scientific and regulatory standards may be 

declared “eligible” for reregistration. The eligibility for continued registration may be 

contingent on label modifications to mitigate risk and can include phase-out and 
23

cancellation of uses and pesticide products. The terms of EPA’s regulatory decisions are 

summarized in the RED (EPA 1997b).

Registrants can submit applications for the registration of new products and new uses 

following reregistration of an active ingredient. Several types of products are registered, 

including the pure (or nearly pure) active ingredient, often referred to as technical grade 

active ingredient (TGAI), technical, or technical product. This is generally used in 

manufacturing and testing, and not applied directly to crops or other use sites. Products 

that are applied to crops, either on their own or in conjunction with other products or 

surfactants in tank mixes are called end-use products (EUPs). Sometimes companies will 

also register the pesticide in a manufacturing formulation, intended for sale to another 

registrant who then includes it into a separately registered EUP. Manufacturing 

formulations are not intended for application directly to use sites. EPA may also cancel

product registrations. With exceptions6, EPA allows the use of canceled products, and 

products that do not reflect RED label mitigation requirements, until those products have 

been exhausted. Labels that reflect current EPA mitigation requirements are referred to 

as “active labels.” Products that do not reflect current label requirements are referred to 

as “existing stocks.” EPA’s action includes all authorizations for use of pesticide 

products including use of existing stocks, and active labels, of products containing 

thiobencarb for the duration of the proposed action. 

Duration of the Proposed Action. EPA’s goal for reassessing currently registered 

pesticide a.i.s is every 15 years. Given EPA’s timeframe for pesticide registration 

reviews, NMFS’ evaluation of the affects of the proposed action on listed Pacific 

salmonids is also 15 years.

Interrelated and Interdependent Activities. No interrelated and interdependent activities 

are associated with the proposed action.

 

6

 For example, EPA placed an end date on authorized use of molinate.
24

Registration Information of Pesticide a.i.s under Consultation. As discussed above, the 

proposed action encompasses EPA’s registration of the uses (as described by product 

labels) of all pesticides containing thiobencarb. EPA provided copies of all active 

product labels for this active ingredient. The following descriptions represent 

information acquired from review of these labels as well as information conveyed in the 

EPA BEs, REDs, and other documents. 

3.1.1. Thiobencarb

Thiobencarb is a thiocarbamate. Thiobencarb is the common name for S-((4-

chlorophenyl)methyl)diethylcarbamothioate (CAS 28249-77-6). Thiobencarb is a 

systemic, preemergence herbicide that acts by inhibiting shoots of emerging seedlings. It 

is used to control grasses, sedge and broadleaf weeds in food crops such as rice 

(nationwide rice represents 95% of use), lettuce, celery, and endive. Thiobencarb was 

first registered for use on rice in 1982. In 1991, thiobencarb was issued regional 

tolerances for use on celery, endives, and lettuce in the State of Florida; however, 

thiobencarb is not authorized for use on these crops in Washington, Oregon, Idaho or 

California. Currently there are eight products containing thiobencarb registered for use 

under Section 3 of FIFRA. They consist of one technical (manufacturing use) product 

containing 97% a.i., emulsifiable concentrate end-use products containing 84% a.i., and 

granular end-use products containing 10 to 15% a.i. RiceBeaux®, which contains a 

mixture of thiobencarb and propanil is not registered for use in California (EPA Reg. No. 

71085-30). Other trade names for thiobencarb products currently used in the U.S. are 

Bolero and Abolish. There is one Special Local Needs (SLN) product registered for use 

in California (EPA Reg. No. 59639-79, EPA SLN CA-930003) under Section 24(c) of 

FIFRA (EPA 1997a). This SLN product is aerial spray applied for water seeded rice, and 

is to be applied to non-flooded fields only. While thiobencarb is registered for use 

nationwide, California is the only state within the range of listed Pacific salmon that has 

approved thiobencarb for use (rice only). As noted above, Oregon, Idaho, and 

Washington have not approved the use of thiobencarb. 
25

3.1.1.1. Usage Information

Reported annual use of thiobencarb in California declined from over one million lbs in 

2000 to approximately 300 thousand lbs per year during the most recent surveys, 2006-

2009 (cite most recent report CDPR 2009). Use of thiobencarb products within the state 

is reflective of the distribution of rice, the only registered use site for thiobencarb. 

Roughly 95 percent of the state’s rice acreage is in the Sacramento Valley, with most of 

the remaining rice grown in northern and central San Joaquin Valley (CDPR 2009 

report). Although rice acreage and thiobencarb use has remained relatively stable in 

California in recent years, differences have been observed among the rice producing 

counties in both the amount of thiobencarb used, and the selection in thiobencarb 

products (CDPR 2010a). As reported by California Department of Pesticide Regulation, 

from 1980 to 2010, thiobencarb has been typically applied at or near the maximum 

application rate of 4 lbs/A (Appendix 3). 

3.1.1.2. Use Sites Authorized

Agricultural. Use of thiobencarb within the distribution of listed Pacific salmonids is 

limited to rice fields. Among the four states where listed salmon and steelhead occur, 

California is the only state that has approved the use of thiobencarb. According to the 

registrants, there are no foreseeable plans to extend the use of thiobencarb in the Pacific 

Northwest States (July 19, 2011 meeting between EPA, NMFS, and the applicant).

Developed. Thiobencarb is not registered for residential, commercial, or industrial uses. 

Forestry. Thiobencarb is not registered for forestry uses.

Aquatic. Thiobencarb is not registered for aquatic uses, other than rice production in 

ponded fields.
26

Other. Thiobencarb is not registered for other uses such as rights-of-way.

Registered Formulation Types. Thiobencarb enduse products are typically formulated as 

liquid or granular. None of the products registered for use in California have more than 

one a.i. in the formulation.

3.1.1.3. Methods and Rates of Application.

Methods. Thiobencarb may be applied via ground and aerial applications. Application 

methods include broadcast spray, granular applicator, high pressure sprayer, and dilute 

high volume spray. The timing of application of recommended application is either late 

preemergence (e.g., 5 to 9 days after planting of the rice), or early post emergence of 

weeds (e.g., rice leaf stage 2-3). Several products recommend use in combination with 

propanil for the control of specific weeds.

Application Rates. Active labels within the action area allow a maximum single and 

seasonal application rate of up to 4 lbs thiobencarb/A to rice (Table 2). Current labels 

allow for a single application, or they do not specify a limit on the number of 

applications. 
27

Table 2. Thiobencarb use patterns in the action area.

Use(s) Use Site Land Use 

category

Max. 

Single 

App. Rate

(lbs 

a.i./A)

Number 

of App. 

per Year

Annual 

App. 

Rate

(lbs 

a.i./A)

App. 

Interval 

(days)

App. 

Method

Label

Number

Rice

Crop Agriculture 4 NS 4 NS Aerial 

and 

ground 

broadcast

59639-791

Rice

Crop Agriculture 3.5 NS 4 NS Aerial 

and 

ground 

broadcast

59639-112

Notes:

1 EPA label for Abolish 8 EC formulation specifies not to release flood water within 14 days of application. Additionally, 

this product can be applied to non-flooded fields only.

2 EPA label for Bolero Ultramax granular flake formulation specifies not to release flood water within 30 days of 

application. 

NS = not specified 

3.1.1.4. Metabolites and Degradates. 

Thiobencarb has two major degradates (defined by EPA as those representing 10% or 

more of the applied radiation of the parent test substance) identified. These are 4-

chlorobenzoic acid (56% of applied radiation at 30 days) and 4-chlorobenzaldehyde 

(29.4% at 14 days), both in a sensitized aquatic photodegradation study. Both of these 

degradates are expected to be soluble and mobile in water. Also, because of their simple 

molecular structure, both should be subject to further degradation by metabolism (EPA 

2009b).
28

4. Action Area

The action area is defined as all areas to be affected directly or indirectly by the federal 

action and not merely the immediate area involved in the action (50 CFR §402.02). 

Given EPA’s nationwide authorization of these pesticides, the action area would 

encompass the entire U.S. and its territories. This geographic area would include all 

listed species and designated critical habitat under NMFS jurisdiction. 

In this instance, as a result of the 2002 order in Washington Toxics Coalition v. EPA, 

EPA initiated consultation on its authorization of 37 pesticide a.i.s and their effects on 

listed Pacific salmonids under NMFS’ jurisdiction and associated designated critical 

habitat in the states of California, Idaho, Oregon, and Washington. However, in this 

consultation with EPA, NMFS is determining the effect of EPA’s proposed 

reauthorization of thiobencarb. Thiobencarb is only authorized for use on rice. In this 

case, rice is only grown in California’s Central Valley (CV) and nowhere else where 

Pacific salmon are listed. Within the CV there are three listed Pacific anadromous

salmonids: the CV steelhead, Sacramento River winter-run Chinook salmon, and CV 

spring-run Chinook salmon. Consequently, for this consultation, the action area consists 

of the entire range of these three listed salmonids and their designated critical habitat in 

California’s Central Valley (Figure 2).7 The action area encompasses the following 

aquatic components of the central valley: all freshwater, estuarine, marsh, swamps, and 

nearshore marine surface waters. 

NMFS’ analysis focuses on the effects of EPA’s action on listed Pacific salmonids in the 

California Central Valley. It includes the effects of thiobencarb on CV steelhead, 

Sacramento River winter-run Chinook and CV spring-run Chinook salmon and their 

designated critical habitats. 

 

7 As noted above, EPA will reinitiate consultation if Washington, Oregon or Idaho authorize use of 

thiobencarb or if the rice growing in California expands beyond its current area. If this occurs, the action 

area for a reinitiated consultation would most likely be larger.
29

EPA’s consultation with NMFS remains incomplete until it analyzes the effects of its 

authorization of pesticide product labels with this compound for all remaining threatened 

and endangered species and all designated critical habitat under NMFS’ jurisdiction. 

EPA must ensure its action does not jeopardize the continued existence or result in the 

destruction or adverse modification of critical habitat for other listed species and 
30

Figure 2. The Central Valley steelhead ESU encompasses the Sacramento River winter-run 

Chinook, and the Central Valley spring-run Chinook ESUs. Rice is only grown within these 

boundaries, consequently, the CV steelhead ESU comprises the area of interest for this 

consultation.
31

designated critical habitat under NMFS’ jurisdiction throughout the U.S. and its 

territories. 

5. Approach to the Assessment

NMFS uses a series of steps to assess the effects of federal actions on endangered and 

threatened species and designated critical habitat. The first step of our analysis identifies 

those physical, chemical, or biotic aspects of proposed actions that are likely to have 

individual, interactive, or cumulative direct and indirect effects on the environment (we 

use the term “potential stressors” for these aspects of an action). As part of this step, we 

identify the spatial extent of any potential stressors and recognize that the spatial extent 

of those stressors may change with time. The spatial extent of these stressors is the 

“action area” for a consultation.

The second step of our analyses identifies the listed resources (endangered and threatened 

species and designated critical habitat) that are likely to occur in the same space and at 

the same time as these potential stressors. If we conclude that such co-occurrence is 

likely, we then try to estimate the nature of co-occurrence (these represent our Exposure 

Analyses). In the exposure analysis, we try to identify life stages and life histories of the 

individuals that are likely to be exposed to an action’s effects and the populations or 

subpopulations those individuals represent. Spatial analyses are used to overlay each 

species range with land types or crop types that pesticides are used on including 

agriculture, urban/residential, forested, and right of ways, to evaluate co-occurrence of 

pesticides and salmonids.

Once we identify which listed resources are likely to be exposed to potential stressors 

associated with an action and the nature of that exposure, in the third step of our analysis 

we examine the scientific and commercial data available to determine whether and how 

those listed resources are likely to respond given their exposure (these represent our 

Response Analyses). We integrate the exposure and response analyses within the Risk 
32

Characterization section to assess the risk to listed individuals and their habitat from the 

stressors of the action. 

In the Risk Characterization Section, we also determine whether population level effects 

are anticipated (these analyses are conducted within the risk characterization phase). 

NMFS’ analysis is ultimately a qualitative assessment that draws on a variety of 

quantitative and qualitative tools and measures to address risk to listed resources.

In the final steps of our analyses, we establish the risks posed to listed species and to 

designated critical habitat. This part of the analysis is found within the Integration and 

Synthesis section. 

Our jeopardy determinations for listed species must be based on an action’s effects on the 

continued existence of threatened or endangered species as those “species” have been 

listed, which can include true biological species, subspecies, or distinct population 

segments of vertebrate species. Because the continued existence of listed species 

depends on the fate of the populations that comprise them, the viability (that is, the 

probability of extinction or probability of persistence) of listed species depends on the 

viability of the populations that comprise the species. Similarly, the continued existence 

of populations are determined by the fate of the individuals that comprise them; 

populations grow or decline as the individuals that comprise the population live, die, 

grow, mature, migrate, and reproduce (or fail to do so).

The structure of our risk analyses reflects the relationships between listed species, the 

populations that comprise each species, and the individuals that comprise each 

population. Our risk analyses begin by identifying the probable risks actions pose to 

listed individuals that are likely to be exposed to an action’s effects. Our analyses then 

integrates those individual-level effects to identify consequences to the populations those 

individuals represent. Our analyses conclude by determining the consequences of those 

population-level risks to the species those populations comprise. 
33

We evaluate risks to listed individuals by determining whether an individual’s “fitness” 

defined as changes in an individual’s growth, survival, annual reproductive success, or 

lifetime reproductive success is reduced. In particular, we examine the scientific and 

commercial data available to determine if an individual’s probable response to an action’s 

effect on the environment (which we identify in our Response Analyses) is likely to have 

consequences for the individual’s fitness.

Reductions in abundance, reproduction rates, or growth rates (or increased variance in 

one or more of these rates) of the populations those individuals represent is a necessary

condition for reductions in a population’s viability, which is itself a necessary condition 

for reductions in a species’ viability. On the other hand, when listed plants or animals 

exposed to an action’s effects are not expected to experience reductions in fitness, we

would not expect that action to have adverse consequences on the viability of the 

population those individuals represent or the species those populations comprise 

((Anderson et al. 2006), (Mills and Beatty 1979), (Stearns 1982)). If we conclude that 

individuals of listed species are not likely to experience reductions in their fitness, we 

would conclude our assessment because an action that is not likely to affect the fitness of 

individuals is not likely to jeopardize the continued existence of listed species.

If, however, we conclude that listed plants or animals are likely to experience reductions 

in their fitness, our assessment determines if those fitness reductions are likely to be 

sufficient to reduce the viability of the populations those individuals represent (measured 

using changes in the populations’ abundance, reproduction, spatial structure and 

connectivity, growth rates, or variance in these measures to make inferences about the 

population’s extinction risks). In this step of our analyses, we use the population’s base 

condition (established in the Status of Listed Resources and Environmental Baseline 

sections of this Opinion) as our point of reference. Finally, our assessment determines if 

changes in population viability are likely to be sufficient to reduce the viability of the 

species those populations comprise.
34

We conduct a separate analysis on species’ designated critical habitat. The analysis 

focuses on reductions in the quality, quantity, or availability of primary constituent 

elements (PCEs) from exposure to the stressors of the action. Since chemicals are the 

stressors of the action for this Opinion, PCEs potentially affected are freshwater 

spawning sites, freshwater rearing sites, freshwater migration corridors, estuarine areas, 

and nearshore marine areas. The PCE attributes of prey availability, water quality, and 

natural cover are the primary assessment endpoints addressed when evaluating the effects 

of an herbicide registration on designated critical habitat. Information evaluated for 

effects to prey include prey survival, prey growth, prey drift, prey reproduction, 

abundance of prey, health of invertebrate aquatic communities, and recovery of aquatic 

communities following pesticide exposure. Information evaluated for degradation of 

water quality include measured and anticipated exposure concentrations leading to toxic 

responses within aquatic organisms (including salmonids, their prey, and primary 

producers that support riparian areas and inwater natural cover) as well as instances of 

water bodies not meeting local, state, or federal water quality standards and criteria. 

5.1. Evidence Available for the Consultation

We search, compile and use a variety of resources to conduct our analyses including:

EPA’s BEs, REDs, IREDS, other documents developed by EPA

Peer-reviewed literature 

Gray literature 

Books

Available pesticide labels

Any correspondence (with EPA or others)

Available monitoring data and other local, county, and state information

Pesticide registrant generated data

Online toxicity databases (PAN, EXTOXNET, ECOTOX, USGS, NPIC)

Pesticide exposure models run by NMFS and EPA

Information and data provided by the registrants identified as applicants

Comments on the draft Opinion from EPA, applicants, and others

Incident reports
35

Collectively, this information provides the basis for our determination as to whether and 

to what degree listed resources under our jurisdiction are likely to be exposed to EPA’s 

action and whether and to what degree the EPA can ensure that its authorization of 

pesticides is not likely to jeopardize the continued existence of threatened and 

endangered species or is not likely to result in the destruction or adverse modification of 

designated critical habitat.

5.2. Application of Approach in this Consultation

For this consultation, we adapt our general approach to incorporate elements of EPA’s 

ecological risk assessment (ERA) framework (EPA 1998). 

Figure 3 shows the overall framework used in this Opinion. This risk assessment 

framework organizes the available information in three phases: problem formulation, 

analysis of exposure and response, and risk characterization (EPA 1998). We adapted the 

EPA framework to address ESA-specific considerations. The NMFS framework follows 

a process for organizing, evaluating, and synthesizing the available information on listed 

resources and the stressors of the action. We separately evaluate the risk to listed species 

and the risk to designated critical habitat from the stressors of the action (See Effects of 

the Proposed Action to Threatened and Endangered Pacific Salmonids and Effects of the 

Proposed Action to Designated Critical habitat). Below, we briefly describe the problem 

formulation phase used to evaluate risk of thiobencarb products.

5.2.1. Problem Formulation 

Problem formulation includes conceptual models based on our initial evaluation of the 

relationships between stressors of the action (pesticides and other identified chemical 

stressors) and receptors (listed species and habitat). Unlike OPP’s pesticide ERAs8, 

which begin with the use, fate, and toxicity properties of thiobencarb, and evaluate risk 

based on a small number of standard toxicity test organisms exposed to thiobencarb, 

NMFS begins with the species’ range and life history to determine relevant assessment 

endpoints, identifies if those endpoints are likely to be affected by the stressors of the 

 

8 Which may be referred to as ERAs, BEs (Biological Evaluations) or pesticide risk assessments in various 

locations throughout this document. 
36

action, and seeks data with which to evaluate those effects. In brief, we employ a 

species-centric approach, rather than a chemical-centric approach. Assessment endpoints 

and measures may vary by life stage and are presented in Table 3. Many of the relevant 

assessment endpoints and measures are not ones typically considered or used in EPA’s 

registration of pesticide active ingredients. 

Table 3. Salmonid life stage and habitat assessment endpoints and measures.

Salmonid Life 

Stage

Assessment Endpoint Assessment Measure

Individual fitness Measures of changes in individual 

fitness

Egg

Development Size, hatching success, morphological 

deformities

Survival Viability (percent survival)

Alevin (yolk-sac fry)

Respiration Gas exchange, respiration rate

Swimming: predator 

avoidance and/or site fidelity

Swimming speed, orientation, burst 

speed, predator avoidance assays

Yolk-sac utilization, growth

rate, size at first feeding

Rate of yolk absorption, growth

weight and length

Development Morphology, histology

Survival LC50, (dose-response slope), percent 

dead at a given concentration

Fry, juvenile, smolt

First exogenous feeding (fry)–

post yolk-sac absorption Time to first feeding, starvation

Survival LC50, (dose-response slope). Percent 

dead at a given concentration 

Growth Stomach contents, weight, length, 

starvation, prey capture rates

Feeding Stomach contents, weight, length, 

starvation, prey capture rates

Swimming: predator 

avoidance behavior, migration, 

use of shelter

Swimming speed, orientation, burst 

swimming speed, predator avoidance 

assays, swimming rate, downstream 

migration rate, fish monitoring, 

bioassays
37

Salmonid Life 

Stage

Assessment Endpoint Assessment Measure

Individual fitness Measures of changes in individual 

fitness

Olfaction: kin recognition, 

predator avoidance, imprinting, 

feeding

Electro-olfactogram (EOG) 

measurements, 

behavioral assays

Smoltification Na/K ATPase activity, sea water 

challenge tests

Returning adult

Development Length, weight, malformations

Survival

LC50, (dose-response slope). Percent 

dead at a given concentration

Feeding Prey consumption rates, stomach 

contents, length and weight

Swimming: predator 

avoidance, migration, 

spawning, feeding

Behavioral assays, numbers of adult 

returns, numbers of eggs fertilized or 

redds, stomach contents

Sexual development Histological assessment of 

ovaries/testis, measurements of intersex

Olfaction: predator avoidance, 

homing, spawning

Electro-olfactogram (EOG) 

measurements, 

behavioral assays

Habitat 

In-stream:

Aquatic primary producers,

salmonid prey 

abundance, dissolved 

oxygen and pH,

natural cover for 

salmonids

Growth inhibition bioassays (EC25 or 

EC50), prey survival (EC50); field 

measured community metrics 

direct measurement

Riparian zone:

Riparian zone vegetation, 

natural cover for salmonids, 

sedimentation, temperature

Growth inhibition (EC25 or EC50),

salmonid monitoring (field)

direct measurements

These assessment endpoints consider effects on all life stages of the salmon (direct 

effects), as well as effects on plants and prey items (indirect effects). Based on the 

assessment endpoints, we evaluate the following risk hypotheses for the species.
38

5.2.1.1. Species Risk Hypotheses

1. Exposure to thiobencarb and other chemical stressors of the action via drift or runoff is 

sufficient to:

a. kill salmonids from direct exposure;

b. reduce salmonid survival through impacts to growth;

c. reduce reproduction;

d. impair swimming;

e. impair respiration

f. reduce salmonids growth through impacts on the availability and quantity of 

prey.

2. Exposure to the thiobencarb via drift or runoff is sufficient to:

a. reduce numbers of aquatic primary producers, thereby affecting salmonid prey 

communities, salmonids and salmonids instream cover;

b. reduce riparian vegetation to such an extent that stream temperatures are 

elevated, erosion increases, and reductions in natural cover results through 

reduced inputs of woody debris and vegetation.

3. Exposure to degradates of thiobencarb will cause adverse effects to salmonids and their 

habitat.

4. Exposure to adjuvants, tank mixes and other chemicals within pesticide products 

containing thiobencarb will cause adverse effects to salmonids and their habitats. 

5. Exposure to other pesticides present in the action area will act in combination with 

thiobencarb to increase effects to salmonids and their habitats.

6. Exposure to elevated temperatures will enhance the toxicity of thiobencarb.

5.2.2. Designated Critical Habitat

When designated critical habitat for the species is identified, primary constituent 

elements (PCEs) of that habitat are also identified in Table 4. To determine potential 

effects to designated critical habitat, we evaluate the effects of the action by first looking 

at whether PCEs of critical habitat are potentially affected by the stressors of the action. 
39

Effects to PCEs include changes to the functional condition of salmonid habitat caused 

by the action in the action area. Properly functioning salmonid PCEs are important to the 

conservation of the ESU/DPS. NMFS convened Critical Habitat Analytical Review 

Teams (CHARTs) that have ranked the conservation value of particular watersheds 

within the designated critical habitat of most Pacific salmonids as high, medium , or low

based on their review of the PCEs. The stressors of the action for this Opinion are 

chemicals introduced into the environment by application of pesticide products 

containing thiobencarb. Key PCEs potentially affected are freshwater spawning sites, 

freshwater rearing sites, freshwater migration corridors, and estuarine areas. Based on 

the action we do not anticipate nearshore marine areas or offshore marine areas to be 

exposed because thiobencarbs use is limited to the Central Valley of California. 

Table 4. Essential physical and biological features of PCEs in salmonid critical habitat 

designations.

Site Essential Physical and Biological features Species Life Stage and Functional 

Developmental Response

Freshwater Spawning Water quality, water quantity, and substrate Spawning, incubation larval 

development

Freshwater rearing Water quantity and floodplain connectivity Juvenile growth and mobility

Water quality and forage Juvenile growth and development

Natural coveraJuvenile mobility and survival

Freshwater migration Free of obstructions, water quality and 

quantity, and natural covera

Juvenile and adult mobility and survival

forage Juvenile growth and development

Estuarine areas Free of obstruction, water quality and 

quantity, and salinity

Juvenile and adult physiological 

transitions between salt and freshwater

Natural covera and forageb and water 

quantity

Growth and maturation

Nearshore Marine areas Free of obstruction, water quality and 

quantity, natural covera and forageb

Growth and maturation, survival

Offshore marine areas Water quality and forageb Growth and maturation

a Natural cover examples include shade, large wood, riparian systems, log jams, beaver dams, aquatic 

vegetation, large rocks and boulders, side channels, and undercut banks.

b Forage includes aquatic and terrestrial invertebrates and fish and shellfish species that support growth and 

maturation of salmonids.
40

Based on the PCEs and life stage potentially affected Table 4, we developed risk 

hypotheses for critical habitat. Properly functioning salmonid PCEs are important to the 

conservation of the ESU/DPS. The stressors of the action for this Opinion are chemicals

introduced into the environment by application of pesticide products. 

5.2.3. Critical Habitat Risk Hypotheses

1. Exposure to the stressors of the action is sufficient to degrade water quality,

natural cover, and/or reduce prey availability in freshwater rearing and migrating 

areas;

2. Exposure to the stressors of the action is sufficient to degrade riparian areas 

adjacent to rearing and migration corridors;

3. Exposure to the stressors of the action is sufficient to degrade water quality, 

natural cover, and/or reduce prey availability in estuarine areas;

5.2.4. Evaluating Exposure and Response

As part of the problem formulation phase, we consider the toxic mode and mechanism of 

action of thiobencarb to provide insight into potential consequences following exposure. 

Identification of the mode and mechanism of action allows us to identify other chemicals 

that might co-occur and affect the response (i.e., identify potential toxic mixtures in the 

environment). We consider authorized thiobencarb use sites to determine spatial overlap 

between its use on rice and the species and its designated critical habitat. We consider 

fate properties of thiobencarb to determine its persistence in aquatic systems. 

Conceptual diagrams are shown in 

Figure 3 and Figure 4.
41

Figure 3. Conceptual framework for assessing risks of EPA’s action to ESA listed 

resources.

Effects on individuals Effects on habitat

Effects on populations

Effects on species

(ESU or DPS)

Effects on water quality 

and prey availability within 

primary constituent 

elements

Effects on the function of 

critical habitat to support 

the intended conservation 

role for the species

Risk 

Characterization

Analyzed within the 

context of the 

Environmental Baseline 

(including multiple 

stressors such as 

temperature and 

environmental mixtures of 

pesticides); the Status of 

the Species; and 

Cumulative Effects

Can EPA ensure its action 

is not likely to jeopardize 

the continued existence of 

the species?

Can EPA ensure its action 

is not likely to adversely 

modify or destroy the 

designated critical habitat?

Stressors of the 

Action 

Co-occurrence of pesticide 

products and geographic 

range of ESA-listed species

Distribution of 

individuals

Distribution of 

habitat

Exposure Profile

Effects of pesticide products 

on ESA-listed species and 

their habitat 

Individual 

responses

Habitat 

responses

Response Profile

Exposure 

Analysis

Response 

Analysis Analyses based 

on the best 

scientific and 

commercial data 

available on 

pesticide 

products use, 

transport, fate, 

toxicity, and 

species ecology
42

exposure to 

terrestrial 

invertebrates 

(salmon prey)

exposure to 

aquatic 

invertebrates

(salmon prey)

exposure to 

individual 

salmon life 

stages

Environmental 

Matrices

Exposure of

Receptors

Responses

Interactions with 

water quality stressors 

in environmental 

baseline:

- other pesticides

-temperature

Life stage 

responses

egg alevin fry/juvenile/ 

smolt

adult

Stressors of 

the Action

other chemicals 

in formulated 

products

thiobencarb and associated degradates 

and metabolites

tank 

mixtures 

adjuvants

+

terrestrial

environment

water column sediment/ pore 

water

aquatic biota

effects to 

habitat

effects to 

individuals

exposure to 

riparian zone 

vegetation 

exposure to 

aquatic primary 

producers 

Figure 4. Exposure pathways for stressors of the action, and general response of Pacific salmonids and habitat.
43

5.2.5. Analysis Plan

5.2.5.1. Status of the Species

In this section, we present information regarding each of the ESUs and DPSs considered in this 

Opinion. We discuss life history, population abundance and trends and overall viability of the 

species. This provides part of the context in which we evaluate the effect of the proposed action.

5.2.5.2. Environmental Baseline

In this section we discuss all stressors affecting salmon populations including natural predators, 

events and disease; and anthropogenic effects such as pollution and habitat modification. This 

also provides part of the context in which we evaluate the effect of the proposed action.

5.2.5.3. Effects of the Proposed Action to Threatened and Endangered Pacific Salmonids

In the Exposure section we discuss life histories of the various species which may make them 

more or less likely to be exposed to stressors of the actions. Then we evaluate measured and 

estimated environmental concentrations of the stressors from various sources. In this section we 

also evaluate spatial and temporal co-occurrences of rice growing and salmon habitat. The 

Response section details information for the assessment endpoints identified in the problem 

formulation. In the Risk Characterization sections for listed species and designated critical 

habitat, we integrate the exposure and response information and evaluate the risk hypotheses. 

Risk Characterization may also include population-level analyses to determine if effects on 

individual fitness are sufficiently large to affect population parameters.

5.2.5.4. Integration and Synthesis

We begin Integration and Synthesis with a summary of risk as described/identified in the Risk 

Characterization associated with each of the a.i.s. In separate sections for listed species and 

critical habitat, we combine these risk conclusions regarding the effects of the proposed action 

with information in the Status of the Species, Environmental Baseline and Cumulative Effects

(those effects of future State or private activities, not involving Federal activities, that are
44

reasonably certain to occur within the action area), to determine potential effects on populations 

and species. 

5.2.5.5. Conclusion

Based on the potential effects for each species, we determine if the proposed action is likely to 

jeopardize the species or cause destruction or adversely modify of destroy designated critical 

habitat.

5.3. Other Considerations

In this Opinion, we evaluated lines of evidence constructed as species-specific risk hypotheses to 

ensure relevant endpoints were addressed. The analysis weighs each line of evidence by 

evaluating the best commercial and scientific data available that pertain to a given risk 

hypothesis. Overall, the analysis is a qualitative approach that uses some quantitative tools to 

provide examples of potential risks to listed salmonids and their habitat. Multiple methods and 

tools currently exist for addressing contaminant-induced risk to the environment. Hazard-based 

assessments, probabilistic risk assessment techniques, combinations of the two, and deterministic 

approaches such as screening level assessments have been applied to questions of risk related to 

human health and the environment. 

We also evaluated a recent assessment tool to quantify spatial and temporal loadings of 

pesticides in the Central Valley (Hoogeweg et al. 2011). We found that much of the information 

on species, pesticide fate, monitoring data, and rice culture was similar to the information we 

used in this assessment. 

In recent pesticide risk assessments, probabilistic techniques have been used to evaluate the 

probability of exceeding a “toxic” threshold for aquatic organisms by combining pesticide 

monitoring data with species sensitivity distributions (Giddings 2009, Geisy et al. 1999). There 

is utility in information generated by probabilistic approaches if supported by robust data. 

NMFS considered the use of probabilistic risk assessment techniques for addressing risk of 

thiobencarb use at population and species (ESU and DPS) scales for the stressors of the action. 
45

However, we encountered significant limitations in available data that suggested the information 

was not sufficient to define exposure and/or response probabilities necessary to determine the 

probability of risk to thiobencarb. Probabilistic techniques were not used in the Opinion due to 

issues with data collection, paucity of data, non-normal distributions of data, and quality 

assurance and quality control. For example, it was not deemed appropriate to pair the salmonid 

prey responses with exposure probabilities based on monitoring results given the limitations of 

that data set discussed in the Effects of the Proposed Action. 

To evaluate population consequences associated with potential lethality from thiobencarb 

exposure in salmon, NMFS selected the lowest reported salmonid LC50 from the available 

information to ensure risk was not underestimated. When we consider the data limitations 

coupled with the inherent complexity of EPA’s proposed action in California’s Central Valley, 

we find that probabilistic assessments at population and species scales introduce an 

unquantifiable amount of uncertainty that undermines confidence in derived risk estimates. 

These same studies do not factor the status of the species, baseline conditions of the environment 

or anticipated cumulative effects into their assessment. At this time, the best available data do 

not support such an analysis and conclusions from such an analysis would be highly speculative.
46

6. Status of Listed Resources

The purpose of this section is to characterize the condition of the three salmonid species9under 

consultation relative to their likelihood of viability and to describe the conservation role and 

function of their respective critical habitats. NMFS has determined that the following species 

and critical habitat designations may occur in the action area for EPA’s registration of 

thiobencarb containing products (Table 5). More detailed information on the status of these 

species and critical habitat are found in a number of published documents including recent 

recovery plans, status reviews, stock assessment reports, and technical memorandums. Many are 

available on the Internet at http://www.nmfs.noaa.gov/pr/species/. 

Table 5. Listed Species and Critical Habitat in the Action Area.

Common Name (Evolutionarily Significant Unit, 

or Distinct Population Segment) Scientific Name Status

Chinook (Sacramento River winter-run) Oncorhynchus tshawytscha Endangered

Chinook (California Central Valley spring-run) Threatened

Steelhead (California Central Valley) Oncorhynchus mykiss Threatened

The following narratives summarize the biology and ecology of threatened and endangered 

Pacific salmonids that are relevant to EPA’s proposed action. This includes a description of the 

timing and duration of each life stage such as adult river entry, spawning, egg incubation, 

freshwater rearing, smolt outmigration, and ocean migration. These summaries provide a 

foundation for NMFS’ evaluation of the effects of the proposed action on listed salmonids. We 

also highlight information related to the viability of salmonid populations and the primary 

constituent elements (PCEs) of designated critical habitat. 

 

9 We use the word “species” as it has been defined in section 3 of the ESA, which include “species, 

subspecies, and any distinct population segment of any species of vertebrate fish or wildlife which 

interbreeds when mature (16 U.S. C 1533).” Pacific salmon other than steelhead that have been listed as 

endangered or threatened were listed as “evolutionarily significant units (ESU), which NMFS uses to 

identify distinct population segments of Pacific salmon. Any ESU or DPS is a “species” for the purposes 

of the ESA.
47

6.1. Species Status 

The status of an ESU or DPS is determined by the degree that it (1) maintains sufficient genetic 

and phenotypic diversity to ensure continued fitness in the face of environmental change, (2) 

maintains spatial distribution of populations so that not all populations would be affected by a 

catastrophic event, and (3) maintains sufficient connectivity among populations within the ESU 

or DPS to maintain long-term demographic and evolutionary processes (ICTRT 2007, McElhany 

et al. 2000, Spence et al. 2008). We describe the current condition of the spatial structure and 

major life histories within the ESUs or DPSs. In order to maintain a spatial distribution and 

diversity that support a viable ESU or DPS, a species must maintain multiple viable populations 

that are sustainable in the long-term in the face of environmental variability. 

Before assessing population viability, we first identify the historic and current populations that 

constitute a species. How NMFS defines a population and its function are found in McElhany et 

al. (2000) and in Bjorkstedt et al.(2005), NMFS’ Pacific salmon Technical Recovery Teams 

(TRTs) have identified historic populations within ESUs/DPSs. These historical populations 

have been categorized based on their distribution and demographic role (i.e., functionally 

independent, potentially independent, or dependent). Functionally independent (independent) 

populations were sufficiently large to be viable in isolation, (i.e., a negligible extinction risk). 

Potentially independent populations were potentially viable in isolation, but were likely 

influenced by immigrants from adjacent populations. Dependent populations were unlikely to 

persist over a 100-year time period in isolation. However, immigration from other nearby 

populations reduced the extinction risk for dependent populations. The historical conditions of 

the populations for each ESU/DPS serve as a point of reference for evaluating the current 

viability of populations10 and the status of the species. The current viability is used as the base 

condition from which the effects of the proposed action on individuals are evaluated to determine 

whether these effects are likely to increase the probability of extinction of the populations those 

individuals represent.

 

10 The TRTs did not propose that historical conditions are the criteria or benchmark for evaluating population or 

ESU viability (extinction risk).
48

In our Approach to the Assessment section, NMFS introduced the VSP concept and its four 

criteria. We restate that a VSP is an independent population (a population of which extinction 

probability is not substantially affected by exchanges of individuals with other populations) with 

a negligible risk of extinction, over a 100-year period, when threats from random catastrophic 

events, local environmental variation, demographic variation, and genetic diversity changes are 

taken into account (McElhany et al. 2000). The four factors defining a viable population are a 

population’s: (1) spatial structure; (2) abundance; (3) annual growth rate, including trends and 

variability of annual growth rates; and (4) diversity (McElhany et al. 2000). 

A population’s tendency to increase in abundance (i.e., lambda (λ) > 1.000) and its variation in 

annual population growth defines a viable population (McElhany et al. 2000, Morris and Doak 

2002). A negative long-term trend in average annual population growth rate (i.e., λ < 1.000) will 

eventually result in extinction. Further, a weak positive long-term growth rate will increase the 

risk of extinction as it maintains a small population at low abundances over a longer time frame. 

A large variation in the growth rates also increases the likelihood of extinction (Lande 1993, 

Morris and Doak 2002). 

Thus, in our status reviews of each listed salmonid species, we provide information on 

population abundance and annual growth rate of extant populations. We use the median annual 

population growth rate (denoted as lambda, λ) from available time series of abundance for 

independent populations (Good et al. 2005). Several publications provide a detailed description 

of the calculation of lambda (Good et al. 2005, McClure et al. 2003). The lambda values for 

salmonid populations of concern in this Opinion are summarized in Table 6.

Table 6. Species and population annual growth rates (Good et al., 2005)

ESU Population λ– H=0 95% CI – lower 95% CI -upper

CV spring-run Chinook Butte Creek

Deer Creek

Mill Creek

1.300

1.170

1.190

1.060

1.040

1.000

1.600

1.350

1.470

Sac. River winter-run 

Chinook Sacramento River 0.970 0.870 1.090

California CV steelhead Sacramento River 0.950 0.900 1.020
49

6.2. Conservation Role of Critical Habitat for the Species

The action area for this consultation contains designated critical habitat. Critical habitat is 

defined as the specific areas within the geographical area occupied by the species, on which are 

found those physical or biological features that are essential to the conservation of the species, 

and which may require special management considerations or protection. Critical habitat can 

also include specific areas outside the geographical area occupied by the species that are 

determined by the Secretary to be essential for the conservation of the species (ESA of 1973, as 

amended, section 3(5)(A)). 

The primary purpose in evaluating the status of critical habitat is to identify for each ESU or 

DPS the function of the critical habitat to support the intended conservation role for each species. 

Such information is important for an adverse modification analysis as it establishes the context 

for evaluating whether the proposed action results in negative changes in the function and role of 

the critical habitat for species conservation. NMFS bases its critical habitat analysis on the areas 

of the critical habitat that are affected by the proposed action and the area’s physical or 

biological features that are essential to the conservation of a given species, and not on how 

individuals of the species will respond to changes in habitat quantity and quality.

In evaluating the status of designated critical habitat, we consider the current quantity, quality, 

and distribution of those primary constituent elements or PCEs that are essential to the 

conservation of the species [50 CFR 424.12(b)]. NMFS has identified PCEs of critical habitat 

for each life stage (e.g., migration, spawning, rearing, and estuary) common for each species. To 

fully understand the conservation role of these habitats, specific physical and biological habitat 

attributes (e.g., water temperature, water quality, forage, etc.) were identified for each life stage. 

Specifically, during all freshwater life stages, salmonids require cool water that is free of 

contaminants. During the juvenile life stage, salmonids also require stream habitat that provides 

excess forage (i.e., prey abundance). Besides potential toxicity, water free of contaminants is 

important as contaminants can disrupt normal behavior necessary for successful migration, 

spawning, and juvenile rearing. Sufficient forage is necessary for juveniles to maintain growth 
50

that reduces freshwater predation mortality, increases overwintering success, initiates 

smoltification, and increases ocean survival. A description of the past, ongoing, and continuing 

activities that threaten the functional condition of PCEs and their attributes are described in the 

Environmental Baseline section of this Opinion.

NMFS has identified six common PCEs for 7 California listed Chinook salmon (including the 

Sacramento winter-run and Central Valley spring-run) and steelhead (including the Central 

Valley DPS) (70 FR 52488, Sept. 2, 2005), 12 ESUs of Oregon, Washington, and Idaho salmon 

(chum, sockeye, Chinook) and steelhead (70 FR 52630, Sept. 2, 2005), and for the Oregon Coast 

coho salmon (73 FR 7816, Feb. 11, 2008). They are: 

(1) Freshwater spawning sites with water quantity and quality, and suitable substrate size as 

attributes necessary to support spawning, incubation and larval development; 

(2) Freshwater rearing sites with the following attributes: (i) Water quantity and floodplain 

connectivity to form and maintain physical habitat conditions and support juvenile growth and 

mobility; (ii) Water quality and forage supporting juvenile development; and (iii) Natural cover 

such as shade, submerged and overhanging large wood, log jams and beaver dams, aquatic 

vegetation, large rocks and boulders, side channels, and undercut banks. 

(3) Freshwater migration corridors free of obstruction and excessive predation with water 

quantity and quality conditions and natural cover such as submerged and overhanging large 

wood, aquatic vegetation, large rocks and boulders, side channels, and undercut banks supporting 

juvenile and adult mobility and survival.

(4) Estuarine areas free of obstruction and excessive predation with:

(i) Water quality, water quantity, and salinity conditions supporting juvenile and adult 

physiological transitions between fresh- and saltwater; (ii) Natural cover such as submerged and 

overhanging large wood, aquatic vegetation, large rocks and boulders, side channels; and (iii) 

Juvenile and adult forage, including aquatic invertebrates and fishes, supporting growth and 

maturation.
51

(5) Nearshore marine areas free of obstruction and excessive predation with:

(i) Water quality and quantity conditions and forage, including aquatic invertebrates and fishes, 

supporting growth and maturation; and (ii) Natural cover such as submerged and overhanging 

large wood, aquatic vegetation, large rocks and boulders, and side channels.

(6) Offshore marine areas with water quality conditions and forage, including aquatic 

invertebrates and fishes, supporting growth and maturation.

In this section, we also identify the conservation values of watersheds located within the critical 

habitat designated for a species. If the effects on PCEs are important at the watershed scale, then 

the conservation value for the watershed is used to assess the conservation role of that watershed 

in the context of range wide critical habitat. The conservation value of a particular watershed 

was determined by Critical Habitat Analytical Review Teams (CHARTs). These teams

considered the presence of PCEs within each occupied area of a watershed and the activities that 

potentially affect the PCEs, and assigned conservation values for watersheds within designated 

critical habitat. 

Each watershed was scored as low, moderate, or high conservation value. High value 

watersheds/areas have a high likelihood of promoting species conservation, while low value 

watersheds/areas are less important for species conservation. Scores were based on: (1) a 

comparison of current quantity of PCEs within a watershed relative to other watersheds and 

probable historic quantity of PCEs within the watershed; (2) existing quality of PCEs in 

watersheds; (3) the likelihood of achieving PCE potential in a watershed; (4) the PCEs’ support 

of rare genetic or life history characteristics or rare/important habitat types in the watershed; (5) 

consideration of the PCEs’ support of variable-sized populations relative to other watersheds and 

the probable historical levels in the watershed; and (6) considerations of the PCE support of 

spawning or rearing of varying numbers of populations.
52

6.3. Chinook Salmon

6.3.1.1. Description of the Species

Chinook salmon are the largest of the Pacific salmon and historically ranged from the Ventura 

River in California to Point Hope, Alaska in North America, and in northeastern Asia from 

Hokkaido, Japan to the Anadyr River in Russia (Healey 1991). Chinook salmon prefer streams 

that are deeper and larger than those used by other Pacific salmon species. We discuss the 

distribution, life history, status, and critical habitat of the two species of endangered and 

threatened Chinook salmon separately.

Chinook salmon are generally described as one of two races, within which there is substantial 

variation (Healey 1991, Groot and Margolis 1991). One race, the “stream-type,” resides in fresh

water for a year or more following emergence from gravel nests. Juveniles migrate to sea as 

yearlings. Stream-type Chinook salmon normally returns in late winter and early spring (springrun) as immature adults and reside in deep pools during summer before spawning in fall. The 

other race, the “ocean-type,” migrate to the ocean within their first year (sub-yearlings) and 

usually return as full mature adults in fall (fall-run). Fall-run adults spawn soon after river entry. 

The timing of return to fresh water, and ultimately spawning, often provides a temporal isolating 

mechanism for populations with different life histories. Return timing is often related to

spawning location. Thus, differences in the timing of spawning migration also serve as a 

geographic isolating mechanism. Fall-run Chinook salmon generally spawn in the mainstem of 

larger rivers and are less dependent on flow, although early autumn rains and a drop in water 

temperature often provide cues for movements to spawning areas. Spring-run Chinook salmon 

take advantage of high flows from snowmelt to access the upper reaches of rivers.

Successful incubation depends on several factors including dissolved oxygen (DO) levels, 

temperature, substrate size, amount of fine sediment, and water velocity. Chinook salmon egg 

incubation time is highly correlated with water temperature (McCullough 1999). Spawning sites 

have larger gravel and more water flow up through the gravel than the sites used by other Pacific 

salmon. Maximum survival of incubating eggs and the pre-emergent alevins occurs at water 
53

temperatures between about 5.5° and 13.5°C. Development time is influenced by degree days 

with fertilization to emergence taking up to 325 days at 2°C and about 50 days at 16°C

(McCullough 1999). Fry emergence commonly begins in December and continues into midApril (Leidy 1984). When emerging from the redd, fry move through the interstitial spaces in 

the redd substrate to escape the gravel. However, a high content of fines and sand in the redd 

substrate can severely hinder fry emergence and cause high mortality (Bjornn and Reiser 1991). 

Optimal temperatures for both Chinook salmon fry and fingerlings range from 12° to 14°C 

(Boles 1988). Temperatures above 15°C increase the risk of diseases and lower the tolerance to 

other stressors (McCullough 1999). At about 19°C, Chinook salmon cease to eat. In the 

laboratory, 50% mortality during a 24 hour period is observed at 24° to 25°C (Brett 1952, 

Hanson 1997) the exact lethal temperature being somewhat dependent on the temperature that 

the fish has been acclimated to.

Chinook salmon alevins, as is the case for other salmonids, rely on yolk for nutrition until the 

onset of active feeding. It is important that the young start feeding at the proper time since 

failure to start feeding can retard growth and lead to behavioral or developmental problems that 

reduce survival. In Chinook salmon, alevins may start feeding immediately upon emergence 

even if they have not yet absorbed all of the egg yolk (Linley 2001). During freshwater 

residence, Chinook salmon juveniles feed in the water column and from the water surface. Food 

items include a variety of small terrestrial and aquatic insects and aquatic crustaceans; the prey 

species of juveniles depend on availability (habitat and months), prey size distribution, and the 

size of the fish (Koehler et al. 2006, Rondorf et al. 1990). The coarse bottom substrate found in 

faster flowing riverine habitats supports drift of larger aquatic insects such as caddisflies 

(Trichoptera), mayflies (Ephemeroptera), stoneflies (Plecoptera), and other benthic organisms 

when they are present in the water column during high flow events. These taxa, when present, 

are important food items in terms of biomass for Chinook salmon juveniles. Terrestrial insects 

and midges (Diptera: Chironmidae) often dominate the diet in slower moving water with finer 

bottom substrate such as floodplains like the Yolo Bypass, off-channel ponds, sloughs, and in 

lakes/reservoirs (Miller and Simenstad 1997, Rondorf et al. 1990, Sommer et al. 2001, Tabor et 

al. 2006). In addition, copepods and daphnia may make up a high proportion of the diet in 

ponds, reservoirs and lakes, and in the mainstems of large rivers (Koehler et al. 2006, Rondorf et 
54

al. 1990, Sommer et al. 2001). At periods, swarming terrestrial insects such as ants can make up 

a substantial portion of the diet of Chinook salmon rearing in floodplains, ponds and reservoirs 

(Rondorf et al. 1990). In estuaries, scuds, mysids, and gammarid amphipods may be major prey 

(Miller and Simenstad 1997).

Studies of stream habitat use show that there are velocity thresholds for rearing fry and juveniles, 

that fish move to faster and deeper water as they grow, and that fish use substrate and cover as 

refuge from high velocities (Chapman and Bjornn 1969, Everest and Chapman 1972, Johnson et

al. 1992). In the mainstem of large rivers and in lakes, fry and juveniles rear along the river 

margins and in nearshore areas that are less than one meter deep and have low lateral bank slopes 

(Sergeant and Beauchamp 2006, Tiffan et al. 2006). Juveniles tend to avoid the elevated water 

velocities found in the thalweg of river channels. As they grow larger, their habitat preferences 

change; juveniles move away from stream margins and begin to use deeper water (Everest and 

Chapman 1972, Tabor et al. 2006). When the river channel is greater than 9- to 10-ft in depth, 

juvenile salmon tend to inhabit the surface waters (Healey 1982).

Chinook salmon fry may also move into non-natal tributaries (i.e., streams other than those 

where they incubated) to rear (Limm and Marchetti 2009, Teel et al. 2009). In the Sacramento 

River, California, fry and juveniles move into seasonally inundated floodplains (Yolo Bypass) 

and off-channel water bodies to rear as they move downstream (Limm and Marchetti 2009, 

Sommer et al. 2001, Teel et al. 2009). However, Chinook salmon use of floodplain and offchannel habitat depend on availability of these habitats, the life history of the race, time of year, 

flow, and temperatures. Up to a certain limit, distribution in floodplain habitat is positively 

correlated with water temperatures (Limm and Marchetti 2009, Sommer et al. 2001, Teel et al. 

2009). Floodplain wetlands and off-channel habitat also often have higher prey densities 

(Sommer et al. 2001). Several studies have shown that fry rearing on large floodplains 

experience a higher growth rate, and possibly higher survival, than fry remaining in the main 

channel (Jeffres et al. 2008, Limm and Marchetti 2003, Sommer et al. 2001). The increased 

growth rate is likely caused by the higher water temperatures as well as the higher prey densities 

in these habitats. Having sufficient growth during the juvenile stage is critical as some studies 

indicate that size at smolting influence survival during the first year in the ocean. As flow 
55

decreases and water temperature increases in summer, juveniles move out of the inundated 

floodplain habitat or succumb to lethal temperatures and stranding. 

Many Chinook salmon populations use the estuary intensively for rearing, and a downstream 

movement of large numbers of fry is typical for many populations (Reimers 1973, Sazaki 1966, 

Thorpe 1994). Estuaries can provide a productive environment and additional growth, refuge 

from predators, and a transition to marine waters; availability of unmodified estuaries is 

correlated with difference between rivers in survival of hatchery reared fish from smolt to 

maturity (Magnusson and Hilborn 2003). Ocean-type Chinook salmon migrate downstream as 

fry immediately after emerging from spawning beds (Healey 1991). These smaller fry and subyearlings extensively use shallow water habitat and sloughs within the estuary to rear to the 

smolt stage (Fresh et al. 2005). Yearling juveniles of the river-type life history enter the 

estuaries at the smolting stage; they usually spend less time in estuaries and use deeper water 

than fry or sub-yearlings (Fresh et al. 2005).

Upon entering the marine environment, immature Chinook salmon maintain close proximity to 

nearshore areas. The highest ocean mortality of immature Chinook salmon occurs during the 

first year after entering the ocean. Expected survival during this period depends both on the 

condition of the fish such as size and the physical conditions of the marine environment. Ocean 

condition such as coastal upwelling and atmospheric condition such as El Niño have a significant 

influence on returning run size. Because of the annual variability in ocean and climatic 

conditions, the stock-recruitment relationship in Chinook salmon is weak.

Immature Chinook salmon of the ocean- and river-type may have different dispersal and 

migration patterns during their first marine year (Healey 1991). The larger stream-type 

immature fish disappear from the surface waters of the Strait of Georgia in early summer. In 

contrast, during their first ocean year, ocean-type fish are abundant in the sheltered surface 

waters and estuaries of the Strait of Georgia and the Puget Sound from July through November 

and some continue to be present throughout winter. Estuaries provide the only shelter along the 

open coasts of Washington, Oregon, and California; in these areas, ocean-type fry remain longer 

in their native estuaries. After ocean entry, immature Chinook salmon may move into large 
56

estuaries and bays as they migrate along the coast. Chinook salmon remain at sea for one to six 

years (more commonly two to four years), with the exception of a small proportion of yearling 

males (called jack salmon) which mature in fresh water or return after two or three months in salt 

water.

6.3.1.2. Status and Trends 

Chinook salmon face natural threats from flooding, changes in ocean productivity, and predation. 

Chinook salmon have declined from over harvests, loss of genetic integrity by mixing with 

hatchery reared fish, reductions in their distribution by migration barriers such as dams, mortality 

and loss in water diversions, loss of rearing habitat from gravel mining, degradation of riparian 

habitat, and modified stream function and reduced water quality from land use practices 

(logging, agriculture, and urbanization).

Climate change also poses significant hazards to the survival and recovery of salmonids. They 

included elevated water temperature, earlier spring runoff and lower summer flows, and winter 

flooding.

6.3.2. Central Valley Spring-run Chinook Salmon

The Sacramento River has sole distinction among the salmon-producing rivers of western North 

America of supporting four runs of Chinook salmon – spring, fall, late-fall, and winter runs 

(Yoshiyama et al. 2001). The Central Valley (CV) Spring-run Chinook salmon includes all 

naturally spawned populations of spring-run Chinook salmon in the Sacramento River, 

California, and its tributaries (Figure 5). The Feather River Hatchery spring-run Chinook salmon 

is included in this ESU. This artificially propagated population is no more divergent relative to 

the local natural populations than would be expected between closely related populations within 

this ESU. Table 9 identifies populations within the CV Spring-run Chinook salmon ESU, their 

abundances, and hatchery input.

Historically, spring-run occupied the upper and middle reaches (1,000 to 6,000 feet in elevation) 

of the San Joaquin, American, Yuba, Feather, Sacramento, McCloud, and Pit rivers, with smaller 
57

populations in most tributaries with sufficient habitat for over-summering adults (Yoshiyama et 

al. 1998). 

Figure 5. Central Valley Spring-run Chinook salmon distribution.
58

6.3.2.1. Life History

The CV spring-run exhibit a stream-type life history. CV Spring-run Chinook salmon are highelevation mainstem spawners that migrate as 3-5 year old adults into the Sacramento River from 

March to May (Yoshiyama et al. 2001) during peak snow-melt flows. Typically, spring-run 

Chinook salmon spawn higher in the watershed than other stocks. They over-summer in cool 

temperature pools before migrating out beginning in late August (Lindley et al. 2004b), or early 

September (Marcotte 1984b) to spawn. Adult spring-run leave the ocean to begin their upstream 

migration in late January and early February (CDFG 1998b) and enter the Sacramento River 

between March and September, primarily in May and June (Table 7). When they enter fresh 

water, spring Chinook are immature. They typically utilize mid- to high-elevation streams that 

provide appropriate temperatures and sufficient flow, cover, and pool depth to allow oversummering while conserving energy and allowing their gonadal tissue to mature (Yoshiyama et 

al. 1998). In Deer and Mill creeks, spawning occurs from late August to mid-October. Embryos 

hatch following a three to six month incubation period (Marcotte 1984b), and the alevins (sacfry) remain in the gravel for another 2-3 weeks. Once their yolk sac is absorbed, juveniles 

emerge between November and March to immediately begin feeding (Moyle 2002). Marcotte 

(1984b) reported observations that juvenile spring-run Chinook in Deer and Mill creeks, during 

most years, spend 8 to 9 months in the higher elevation streams, where they feed on drift insects 

(Table 7). Newly emerged fry tend to school in calm, shallow water near shoreline areas 

(Marcotte 1984b). As they grow larger, juvenile salmon shift to faster deeper water (Groot and 

Margolis 1991). The fry growing season is typically April to September with growth rate 

varying depending on stream productivity and water temperature (Spence et al. 1996a, Groot and 

Margolis 1991). Adult Chinook salmon require cool fresh water while they mature over the 

summer. Adult upstream migration may be blocked by temperatures above 21oC (McCullough 

1999). Juvenile spring-run emigration in the Sacramento River is highly variable and they may 

migrate either as soon as they emerge from the gravel or as yearlings. The majority of spring-run 

fry emerging in the tributaries migrate downstream from December through February during 

high flows. Juvenile CV Spring-run Chinook salmon have been observed rearing in the lower 

reaches of non-natal tributaries and intermittent streams in the Sacramento Valley during the 

winter months. Peak fry/sub-yearling movements are observed farther downstream in lower 
59

Sacramento River (Knights Landing) and the Delta during March and April. Up to 25% of 

juveniles may remain in the tributaries to rear and outmigrate as yearlings the next fall, normally 

starting in December. 

Table 7. Temporal occurrence of adult and juvenile CV Spring-run Chinook salmon in the 

Sacramento River. Darker shades indicate months of greatest relative abundance. Note: Yearling 

rear in their natal streams through the first summer following their birth. Downstream migration 

generally occurs the following fall and winter. Young of year migrate during the first spring after 

they hatch.

Adult migrationa

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. River basina,b

Up Sac. R. mainstemc

Mill Creekd

Dear Creekd

Butte Creekd

Adult Holdingb

Adult Spawningc

Juvenile migrationd

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. R. Tribse

Upper Butte Ck.f

Mill, Deer, Butte Cks.d

Sac. R. at RBDDc

Sac. R. at KLg,h

North Deltai

Relative Abundance = High = Medium = Low

Sources: aYoshiyama et al. (1998); bMoyle (2002); cMyers et al. (1998); dLindley et al. (2007) and Marcotte 

(1984b); eCDFG (1998b) fMcReynolds et al. (2005) and Ward et al. (2002, 2003); gSnider and Titus (2000b); 

hBajjaliya and Vincik (2008b); iUSFWS(2010a).

Studies by California’s Department of Fish and Game have shown Juvenile CV Spring-run 

Chinook had migrated past their sample sight just downstream of Knights Landing by the end of 

April (Bajjaliya and Vincik 2008b). Spring-run juvenile Chinook salmon were caught in screw 
60

traps below Knights Landing from December through April. Bajjaliya and Vincik (2008b) found 

peak abundance of CV Spring-run Chinook during this sampling period was late December into 

early January. Another peak of juvenile Chinook showed up later in the spring which they 

determined to be fall-run based on coded-wire tags and size distribution of fall-run released from 

the Colman National Fish Hatchery (Bajjaliya and Vincik 2008b). However, estimating the race 

of juvenile Chinook in the field by size and date of capture has some uncertainty. Criteria for 

determining race was developed by the CDFG in 1992 (Fisher 1992) as a weekly model of 

juvenile salmonid growth. This was modified to a daily criterion by the California Department 

of Water Resources (Greene 1992). It is currently the only tool used by several salmon 

monitoring programs with the Central Valley to determine race of juvenile Chinook in the field. 

Until markers for genetic differentiation of races are developed, the race determinations reported 

cannot be considered definitive (USFWS 2012b). Some fish caught in the traps below Knights 

Landing later in the spring identified by biologists in the field as fall-run, could in fact have some 

of the tail end of the spring-run population intermixed.

Juvenile CV Spring-run Chinook are likely to be lingering in the lower river and SacramentoSan Joaquin Delta into June before entering into San Francisco Bay (USFWS 2012b, USFWS 

2010a). Sampling conducted in the North Delta of the Sacramento River from 1993 – 2009 

showed the abundance of CV Spring-run juvenile Chinook (based on catch per unit effort) in 

most years peaked either in February or March (Table 7). A combination of Spring- and Fall-run 

Chinook continue to migrate through the North Delta through June (USFWS 2012b, USFWS 

2010a). 
61

6.3.2.2. Status and Trends

NMFS originally listed CV Spring-run Chinook salmon as threatened on September 16, 1999 (64 FR 50393), 

and reaffirmed their threatened status on June 28, 2005 (70 FR 37160) and again on August 15, 2011. 

Historically, spring-run Chinook salmon were predominant throughout the Sacramento and San Joaquin River 

drainages. All runs within the San Joaquin River basin are now extirpated. The CV Technical Recovery Team 

(TRT) delineated 18 independent populations of CV spring-run Chinook, along with a number of smaller 

populations, and four geographically separated groups (Lindley et al. 2004b). Of these 18 populations, only 

three are extant on the upper Sacramento River and they represent only the Northern Sierra Nevada geographic

group.

Naturally spawning populations of CV Spring-run Chinook salmon currently are restricted to accessible reaches 

of the upper mainstem Sacramento River and its tributaries Butte, Deer, and Mill Creeks. Limited spawning 

occurs in the basins of smaller tributaries (CDFG 1998b, Lindley et al. 2004b). Using data through 2005 and 

the criteria in Table 8, Lindley et al. (2007) found that the Mill Creek, Deer Creek, and Butte Creek populations 

were at or near low risk of extinction. However, the ESU as a whole, could not be considered viable because 

there were no extant populations in the three other diversity groups. In addition, Mill, Deer, and Butte creeks 

are close together, decreasing the independence of their extinction risks due to catastrophic disturbance 

(Williams et al. 2011).
62

Table 8. Criteria for assessing the level of risk of extinction for populations of Pacific salmonids in the Central 

Valley of California. Overall risk is determined by the highest risk score for any category.

 Risk of Extinction

Criterion High Moderate Low

Extinction Risk from PVA

>20 % within 20 years >5% within 100 years <5% within 100 years

or any ONE of: or any ONE of: or any ONE of:

Population sizeaNe ≤ 50 50 < Ne ≤500 Ne > 500

-or- -or- -orN ≤ 250 250 < N ≤ 2500 N > 2500

Population decline Precipitous declinebChronic decline or 

depressionc

No decline apparent or 

probable

Catastrophe, rate and 

effectd

Order of magnitude

decline within one 

generation

Smaller but significant 

declineeNot apparent

Hatchery influence High Moderate Low

a Census size N can be used if direct estimates of effective size Ne are not available, assuming 

Ne / N = 0.2.

b Decline within last two generations to annual run size ≤ 500 spawners, or run size > 500 but declining at ≥ 

10% per year over the past 10 years. Historically small but stable population not included.

c Run size has declined to ≤ 500, but now stable.

d Catastrophes occurring within the last 10 years.

e Decline < 90% but biologically significant.
63

Table 9. Central Valley Spring-run Chinook salmon--preliminary population structure, historic and most recent 

natural production, spawner abundance, and hatchery contributions (Good et al. 2005, USFWS and Reclamation 

2007, Williams et al. 2011).

Population

Historic Natural 

Production (1967 –

1991)

Most Recent Natural 

Production1(2000 –

2006)

Most Recent 

Spawner Abundance2

(2000- 2006)

Hatchery Abundance 

Contributions

Butte Creek 1,000 6,516 – 19,809 4,118 – 10,625 < 1%3

Deer Creek 3,300 1,387 – 3,461 637 – 2,759 Unknown

Mill Creek 2,200 1,184 – 26,190 544 – 1594 Unknown

Sacramento River 29,000 0 – 1,134 0 – 394 Unknown

Total

Estimated historic 

abundance: 

~700,000 for all 

populations

11,403 – 26,190 5,370 – 14,044 Unknown

1

Includes catches

2

 i.e., escapement

3

(Williams et al. 2011)

The Central Valley drainage supported spring-run Chinook salmon runs as large as 700,000 fish between the 

late 1880s and the 1940s (Brown et al. 1994). Before construction of Friant Dam, nearly 50,000 adults were 

counted in the San Joaquin River alone (Fry 1961).

Median natural production of spring-run Chinook salmon from 1970 to 1989 was 30,220 fish. In the 1990s, the 

population experienced a substantial production failure with an estimated natural production ranging between 

3,863 and 7,806 fish (with the exception of 1995 which had a natural production of an estimated 35,640 adults)

during the years between 1991 and 1997 (USFWS and Reclamation 2007). Numbers of naturally produced fish 

increased significantly in 1998 to an estimated 48,755 adults and estimated natural production has remained 

above 10,000 fish since then (USFWS and Reclamation 2007). 

In previous Opinions, NMFS reported that the Sacramento River trends and lambda show a long- and shortterm negative trend and negative population growth (Good et al. 2005). NMFS also reported the median 

production of Sacramento River tributary populations increased from a low of 4,248 with only one year 

exceeding 10,000 fish before 1998 to a combined natural production of more than 10,000 spring-run Chinook in 

all years after 1998 (data from (USFWS and Reclamation 2007)). It was then stated that time series data for 

Mill, Deer, Butte, and Big Chico Creeks spring-run Chinook salmon (updated through 2006) showed that all 

three tributary spring-run Chinook populations have long-and short-term lambdas >1; indicating population 

growth (Good et al. 2005). Finally we reported that although the populations are small, CV spring-run Chinook 

salmon have some of the highest population growth rates in the Central Valley. 
64

Currently Williams et al. (2011) report that the escapement of CV spring-run Chinook salmon to various areas 

of the Central Valley, with few exceptions, have declined over the past 10 years, in particular since 2006. The 

recent declines in abundance place Mill and Deer Creek populations in the high risk category due to their rate of 

decline, and in the case of Deer Creek, also the level of escapement. Butte Creek continues to satisfy the 

criteria for low extinction risk, although the rate of decline is close to triggering the population decline criterion 

for high risk (Williams et al. 2011). Overall, the recent declines have been significant but not severe enough to 

qualify as a catastrophe under the criteria of Lindley et al (2007). On a positive note, spring-run Chinook 

salmon appear to be repopulating Battle Creek, home to an historical independent population in the Basalt and 

Porous Lava diversity group that was extirpated for many decades. This population has increased in abundance 

to levels that would qualify it for a moderate extinction risk score (Williams et al. 2011). Similarly, the springrun Chinook salmon population in Clear Creek has been increasing, although Lindley et al. (2004b) classified 

this population as a dependent population, and thus is not expected to exceed the low-risk population size 

threshold of 2500 fish.

In order to determine the current likelihood of the spring-run ESU becoming viable, we used the historical 

population structure of spring-run presented in Lindley et al. (2007) and McElhany et al.’s (2000) description of 

viable salmonid populations (VSP) for evaluating populations. While McElhany et al. (2000) introduced and 

described the concept of VSP, Lindley et al. (2007) applied the concept to the spring-run ESU. Lindley et al. 

(2004b) identified 26 historical populations within the spring-run ESU; 19 were independent populations, and 7 

were dependent populations. Of the 19 independent populations of spring-run that occurred historically, only 

three remain, in Deer, Mill, and Butte creeks. Extant dependent populations occur in Battle, Antelope, Big 

Chico, Clear, Beegum, and Thomes creeks, as well as in the Yuba River, the Feather River below Oroville 

Dam, and in the mainstem Sacramento River below Keswick Dam.

Table 8 provides various quantitative criteria to evaluate the risk of extinction. The following provides the 

evaluation of the likelihood of the threatened spring-run ESU becoming viable based on the VSP parameters of 

population size, population growth rate, spatial structure, and diversity.
65

6.3.2.3. Population Size

As discussed above, Spring-run Chinook numbers declined drastically in the mid to late 1980s before stabilizing 

at very low levels in the early to mid 1990s. Since the late 1990s, there does not appear to be a trend in basinwide abundance, having fluctuated from approximately 25,000 fish in 1999 to slightly more than 10,000 fish in 

2008. Abundance is generally dominated by the Butte Creek population. Other independent and dependent 

populations are smaller. The cohort replacement rate behaved similarly, falling below 1.0 in the 3 of the 

previous 4 years, in parallel with the reduced escapement numbers. The 5-year moving average cohort 

replacement rate, however, has remained above 1.0 since 1995.

6.3.2.4. Population Growth Rate

Cohort replacement rates are indications of whether a cohort is replacing itself in the next generation. As 

mentioned in the previous subsection, the cohort replacement rate since the late 1990s has fluctuated, and does 

not appear to have a pattern. Since the cohort replacement rate is a reflection of population growth rate, there 

does not appear to be an increasing or decreasing trend. The 5-year moving average of population estimate 

indicated an increasing population trend since the mid-1990s until very recently (2006), at which point the 

population has decreased in two consecutive years. Good et al., (2005)

report lambda (λ) values slightly above replacement (1.000) for each of the three main spawning areas (Table 

6).

6.3.2.5. Spatial Structure

As stated above, Lindley et al. (2007) indicated that of the 19 independent populations of spring-run that 

occurred historically, only three (Butte, Mill, and Deer creeks) remain, and their current distribution makes the 

spring-run ESU vulnerable to catastrophic disturbance. Butte, Mill, and Deer Creeks all occur in the same 

biogeographic region (diversity group), whereas historically, independent spring-run populations were 

distributed throughout the CV among at least three diversity groups (i.e., basalt and porous lava, northern Sierra 

Nevada, and southern Sierra Nevada). In addition, dependent spring-run populations historically persisted in 

the Northwestern California diversity group (Lindley et al. 2004b). Currently, there are dependent populations 

of spring-run in the Big Chico, Antelope, Clear, Thomes, Battle, and Beegum creeks, and in the Sacramento, 
66

Feather, and Yuba rivers. The extant Feather River and mainstem Sacramento River populations probably do 

not represent historical populations (Lindley et al. 2007).

6.3.2.6. Diversity

Diversity, both genetic and behavioral, provides a species the ability to react to sudden environmental changes. 

As a species’ abundance decreases, and spatial structure of the ESU is reduced, a species has less flexibility to 

respond to changes in the environment. Spring-run have been entirely extirpated from the basalt and porous 

lava region and the southern Sierra Nevada region. The only viable and independent populations (i.e., Mill, 

Deer, and Butte creeks) of spring-run are limited to the northern Sierra Nevada region, and a few ephemeral or 

dependent populations are found in the Northwestern California region. A single catastrophe, for example, the 

eruption of Mount Lassen, a large wildland fire at the headwaters of Mill, Deer, and Butte creeks, or a drought, 

poses a significant threat to the extinction risk of the ESU that otherwise would not be there if the ESU’s 

spatial structure and diversity were greater. Spring-run do reserve some genetic and behavioral variation in that 

in any given year, at least two cohorts are in the marine environment, and therefore, not exposed to the same 

environmental stressors as their freshwater cohorts. 

Spring-run Chinook produced at the FRFH are part of the spring-run ESU (June 28, 2005, 70 FR 37160), and 

they compromise the genetic diversity of naturally-spawned spring-run. More than 523,000 FRFH spring-run 

fry were planted at the base of Whiskeytown Dam during the 3-year period 1991−1993 (CDFG 1998b). The 

fact that these hatchery fish behave more like fall-run (spawn later than spring-run in Deer, Mill, and Butte 

creeks), likely increases introgression of the spring- and fall- runs, and reduces diversity.

Until recently NMFS was unaware of any current reports of hatchery-origin fish spawning in the higher 

elevation areas of Butte, Deer, or Mill creeks utilized by spring-run Chinook. However, in 2010, McReynolds, 

CDFG, reported 10 coded-wire tags of Feather River spring Chinook Salmon were recovered from a sample of 

1,113 carcasses in the upper reach of Butte Creek (Williams et al. 2011). As 100% of Feather River hatchery 

spring-run Chinook are marked and tagged, this translates into slightly less than 1% of the Butte Creek returns 

(Table 9) being compromised with hatchery strays. This is well below the 10% allowable stray rate for out-ofdiversity-group-origin fish within one generation as described in Williams et al. (2011). 
67

6.3.2.7. Summary of Current Viability

As reported in Williams et al. (2011), the status of CV spring-run Chinook salmon has probably deteriorated on 

balance since the 2005 status review (Good et al. 2005), and Lindley et al.’s (2007) assessment, with two of the 

three extant populations slipping from low or moderate extinction risk to high extinction risk. Butte Creek and 

Deer Creek spring-run are at low risk of extinction, satisfying both the population viability analysis (PVA) and 

other viability criteria. However, continued documentation of hatchery strays to Butte Creek could put it on the 

verge of being moved to the high risk category. Mill Creek is at moderate extinction risk according to the PVA, 

but appear to satisfy the other viability criteria for low-risk status (Lindley et al. 2007). As a whole, the springrun fail the representation and redundancy rule for ESU viability. The current distribution of independent 

populations has been severely constricted to only one of their former geographic diversity groups. Therefore, 

the spring-run ESU is at moderate risk of extinction in 100 years.

6.3.2.8. Critical Habitat

NMFS designated critical habitat for this species on September 2, 2005 (70 FR 52488).

The critical habitat boundary includes the Sacramento River and several tributaries from the Big Chico tributary 

with Sacramento River upstream to Shasta Dam (Table 10).

There are 38 occupied HSA watersheds within the freshwater and estuarine range of this ESU. As shown in 

Figure 6, seven watersheds received a low rating, 3 received a medium rating, and 27 received a high rating of 

conservation value to the ESU (NMFS 2005). Four of these HSA watersheds comprise portions of the San 

Francisco-San Pablo-Suisun Bay estuarine complex which provides rearing and migratory habitat for this ESU.

The current condition of PCEs of the CV Spring-run Chinook salmon critical habitat indicates that PCEs are not 

currently functioning or are degraded; their conditions are likely to maintain a low population abundance across 

the ESU. Spawning and rearing PCEs are degraded by high water temperature caused by the loss of access to 

historic spawning areas in the upper watersheds which maintained cool and clean water throughout the summer. 

The rearing PCE is degraded by floodplain habitat being disconnected from the mainstem of larger rivers 

throughout the Sacramento River watershed, thereby reducing effective foraging. Migration PCE is degraded 

by lack of natural cover along the migration corridors. Juvenile migration is obstructed by water diversions 

along Sacramento River and by two large state and federal water-export facilities in the Sacramento-San 

Joaquin Delta.
68

Table 10. CV Spring-run Chinook salmon CALWATER HSA watersheds with conservation values.

HUC 4 Subbasin

HUC 5 Watershed conservation Value (CV)

High CV PCE(s) 1 Medium 

CV PCE(s) 1

Low CV PCE(s) 1

San Francisco Bay

San 

Francisco 

Bay

Estuary 

PCEs 0 0 1

Estuary 

PCEs

Suisun Bay Suisun Bay 1 0 0 0

Tehama 1 (1, 2, 3) 1 (1, 2, 3) 0

Whitmore 1 (1, 2, 3) 0 2 (1, 2, 3)

Redding 2 (1, 2, 3) 0 0

Eastern Tehama 4 (1, 2, 3) 0 0

Sacramento Delta 1 (2, 3, 1) 0 0

Valley PutahCache 1 (1, 2, 3) 0 0

Marysville 3 (1, 2, 3) 0 0

Yuba River 2 (1, 2, 3) 1 (1, 2, 3) 1 (1, 2, 3)

Valley-American 2 (1, 2, 3) 0 0

Colusa Basin 4 (1, 2, 3) 0 0

Butte Creek 1 (1, 2, 3) 0 0

Ball Mountain 0 0 1 (1, 2, 3)

Shasta Bally 3 (1, 2, 3) 0 1 (1, 2, 3)

North Diablo 

Range 0 1 (1, 2, 3) 0

San Joaquin Delta 0 0 1 (1, 2, 3)

Total 28 3 7

1 Numbers in parenthesis refers to the dominant (in river miles) PCE(s) within the HUC 5 watersheds. PCE 1 is spawning and 

rearing, 2 is rearing and migration, and 3 is migration and presence. PCEs with < means that the number of river miles of the PCE is 

much less than river miles of the other PCE.

Contaminants from agriculture and urban areas have degraded rearing and migration PCEs to the extent that 

they have lost their functions necessary to serve their intended role to conserve the species. Water quality 

impairments in the designated critical habitat of this ESU include inputs from fertilizers, insecticides, 

fungicides, herbicides, surfactants, heavy metals, petroleum products, animal and human sewage, sediment in 

the form of turbidity, and other anthropogenic pollutants. Pollutants enter the surface waters and riverine 

sediments as contaminated stormwater runoff, aerial drift and deposition, and via point source discharges. 

Some contaminants such as mercury and pentachlorophenol enter the aquatic food web after reaching water and 

may be concentrated or even biomagnified in salmon tissue. 
69

Figure 6. Central Valley Spring-run Chinook salmon Conservation Values per Sub-Area.

6.3.3. Sacramento River Winter-run Chinook Salmon

The ESU includes all winter-run Chinook salmon entering and using the Sacramento River system in the 

Central Valley, California. The ESU boundary extends from the Carquinez Strait by the City of Vallejo and 
70

Benicia upstream to the Sacramento River, including all its tributaries, to below Keswick Dam (Figure 7). The 

ESU now consists of a single spawning population.

6.3.3.1. Life History

The distribution of winter-run spawning and most rearing is limited to the upper Sacramento River and its 

tributaries, where spring-fed streams provide cold water throughout the summer, allowing for spawning, egg 

incubation, and rearing during the mid-summer period (Yoshiyama et al. 1998). The headwaters of the 

McCloud, Pit, and Little Sacramento rivers, and Hat and Battle creeks, historically provided clean, loose gravel; 

cold, well-oxygenated water; and optimal stream flow in riffle habitats for spawning and incubation. These 

areas also provided the cold, productive waters for egg and fry development and survival, and juvenile rearing 

over the summer.

The upper Sacramento River is the only spawning area used by winter-run, although occasional strays have 

been reported in Battle and Clear Creek. In recent years, the majority of winter-run (i.e., > 50 percent since 

2007) spawn in the area from Keswick Dam downstream to the ACID Dam (approximately 5 miles). Keswick 

Dam re-regulates flow from Shasta Dam and mixes it with water diverted from the Trinity River through the 

Spring Creek tunnel. When the gates are down at RBDD, or flashboards in at the ACID Dam, access to the 

upper Sacramento River basin, including tributaries, can only be achieved through the RBDD and ACID Dam 

fish ladders.

The winter-run Chinook salmon have characteristics of both stream- and ocean-type races (Healey 1991). 

Adults enter fresh water in winter or early spring but delay spawning most years until May and June. During 

the 4 to 6 week period when alevins remain in the gravel, they utilize their yolk-sac to nourish their bodies. As 

their yolk-sac is depleted, fry begin to emerge from the gravel to begin exogenous feeding in their natal stream. 

Fry typically range from 25 mm to 40 mm at this stage (Groot and Margolis 1991, Healey 1991). Upon 

emergence, fry swim or are displaced downstream (Healey 1991). The post-emergent fry disperse to the 

margins of their natal stream, seeking out shallow waters with slower currents, finer sediments, and bank cover 

such as overhanging and submerged vegetation, root wads, and fallen woody debris, and begin feeding on 

zooplankton, small insects, and other micro-crustaceans. Some fry may take up residence in their natal stream 

for several weeks to a year or more, while others are displaced downstream by the stream’s current. Once 

started downstream, fry may continue downstream to the estuary and rear there, or may take up residence in 
71

river reaches farther downstream for a period of time ranging from weeks to a year (Healey 1991). Fry seek 

nearshore habitats containing riparian vegetation and associated substrates important for providing aquatic and 

terrestrial invertebrates, predator avoidance, and slower velocities for resting. The benefits of shallow water 

habitats for salmonid rearing have been found to be more productive than the main river channels, supporting 

higher growth rates, partially due to higher prey consumption rates, as well as favorable environmental 

temperatures (Sommer et al. 2001). Fry emerge from the gravel in late June to early July and continue through 

October (Fisher 1994). As stated above, some offspring move downstream as fry (ocean-type) while others rear 

in the upper Sacramento River and move down as smolt (river-type). Monitoring programs have shown that 

young winter-run Chinook salmon fry start migrating to sea and may migrate past the Red Bluff Diversion Dam 

(RBDD) as early as mid-July with a peak movement over the RBDD in September (Vogel and Marine 1991, 

NMFS 1997, Bajjaliya and Vincik 2008b, 2010a, USFWS and Reclamation 2007). Normally fry have passed 

the RBDD by October, while smolts may pass over the RBDD until March (Table 11). Juvenile winter-runs 

occur in the Delta beginning in October, primarily from November through February, but can continue until 

early May. Winter-run juveniles remain in the Delta until they are from 5 to 10 months of age, and then begin 

emigrating to the ocean as early as November and continue through May (Fisher 1994, Myers et al. 1998). As 

with all runs, the timing of migration each year may vary somewhat due to changes in river flows, dam 

operations, and type of water year. The winter-run race matures between two and six years of age with the 

majority returning as three-year olds. 

Table 11. Temporal occurrence of adult and juvenile winter-run Chinook salmon in the Sacramento River. Darker 

shades indicate months of greatest relative abundance.

Adult migration

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. River basina

Upper Sac. R. mainstemb

Juvenile migration

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. River at RBDDc

Sac. River at KLd

Lower Sac. Rivere

North Deltae

RBDD=Red Bluff Diversion Dam KL = Knights Landing

Relative Abundance: = High = Medium = Low

a Yoshiyama et al. (1998); bMoyle (2002); c Martin et al. (2001); d

Snider and Titus (2000b) and Bajjaliya and Vincik (2008b); e 

USFWS (2010a, 2012b)
72

6.3.3.2. Status and Trends

NMFS listed Sacramento River winter-run Chinook salmon as endangered on January 4, 1994 (59 FR 440), and 

reaffirmed their endangered status on June 28, 2005 (70 FR 37160). The winter-run Chinook salmon spawned 

and reared in the upper Sacramento River and its tributaries (Slater 1963b, Yoshiyama et al. 1998). Today the 

Shasta Dam eliminates access to the historic spawning habitat. Cold water releases from Shasta Dam have also 

created conditions suitable for winter-run spawning and rearing in a 60- to 100-mile long portion of the 

Sacramento River downstream of the dam. However, spawning only occurs from Keswick Dam to 

approximately 44 miles downstream (Yoshiyama et al. 1998). As a result, the Sacramento River winter-run 

Chinook salmon has been reduced to a single spawning population confined to a portion of the main stem 

Sacramento River.
73

Figure 7. Sacramento River Winter-run Chinook salmon distribution.
74

Winter-runs may have been as large as 200,000 fish based upon commercial fishery records from the 1870s

(Fisher 1994). During the first three years of operation of the counting facility at the RBDD (From 1967 to 

1969), an average of 86,500 winter-run Chinook salmon were counted with a high of over 230,000 adults 

(CDFG 2008, Good et al. 2005). A rapid decline occurred from 1969 to 1979 after completion of the RBDD. 

Over the next 20 years, the population eventually reached a low point of only 186 adults in 1994. At that point, 

winter-run was at a high risk of extinction, as defined in the guideline for recovery of CV salmonids (Lindley et 

al. 2007). If not for a very successful captive broodstock program, construction of a temperature control device 

on Shasta Dam, changing operational procedures at the RBDD (keeping the gates up for most of the year), and 

restrictions in ocean harvest, the population would have likely failed to exist in the wild.

The population grew rapidly from the early 1990s to mid-2005. Mean run size increased from 1,363 before 

2000 with all runs estimated to less than 10,000 fish to an average run of 8,470 adults between 2000 and 2006 

with two runs estimated to more than 10,000 fish (USFWS and Reclamation 2007). However, the natural 

produced winter-run Chinook salmon plunged in 2007 and 2008, with 4,461 adults estimated for 2007 and a 

preliminary estimate between of 2,600-2,850 adults for 2008 (USFWS 2008). 

The Sacramento River winter-run Chinook salmon is expected to have lost some genetic diversity through 

bottleneck effects in the late 1980s and early 1990s. Hatchery releases may also have affected population 

genetics. The loss of natural spawning habitat and hydrological conditions has further removed the natural 

evolutionary processes that maintained the unique winter-run life history.

The status of winter-run is typical of most endangered species populations, that is, a sharp downward decline 

followed by years of low abundance (Figure 8). Lindley et al. (2004b) identified four historical populations 

within the winter-run ESU, all independent populations, defined as those sufficiently large to be historically 

viable in isolation and whose demographics and extinction risk were minimally influenced by immigrants from 

adjacent populations (McElhany et al. 2000). All four independent populations, however, are extinct in their 

historical spawning ranges. Three (Little Sacramento; Pit, Fall, Hat; and McCloud River) are blocked by the 

impassable Keswick and Shasta dams (Lindley et al. 2004b), and Battle Creek independent population is no 

longer self-sustaining (Lindley et al. 2007). Lindley et al. (2007) provided various quantitative criteria to 

evaluate the risk of extinction (Table 8). A population must meet all the low-risk thresholds to be considered 

viable. The following provides the evaluation of the likelihood of winter-run becoming viable based on the 
75

VSP parameters of size, population growth rate, spatial structure, and diversity. These specific parameters are 

important to evaluate because they are predictors of extinction risk, and the parameters reflect general biological 

and ecological processes that are critical to the growth and survival of salmon (McElhany et al. 2000). 

Figure 8. Estimated yearly adult natural production and in-river adult escapement of winter-run from 1967 - 2007 

based on RBDD ladder counts.

In a recent Biological Opinion on the long-term operations of the Central Valley Project and State Water Project 

(NMFS 2009a), NMFS concluded for all four VSP parameters that winter-run Chinook salmon are at a high risk 

of extinction. These parameters are summarized below.

6.3.3.3. Population Size

Information about population size provides an indication of the type of extinction risk that a population faces. 

For instance, smaller populations are at a greater risk of extinction than large populations because the processes 

that affect populations operate differently in small populations than in large populations (McElhany et al. 2000). 

One risk of low population sizes is depensation. Depensation occurs when populations are reduced to very low 
76

densities and per capita growth rates decrease as a result of a variety of mechanisms [e.g., failure to find mates 

and therefore reduced probability of fertilization, failure to saturate predator populations (Liermann and Hilborn 

2001). The winter-run population as represented by the 5-year moving average from adult escapement was 

following an increasing trend from the mid-1990s until 2006. In 2007, the winter-run population declined 

precipitously. Low adult escapement was repeated in 2008. Likewise, the 5-year moving average cohort 

replacement rate was relatively stable since the late 1990s, with each cohort approximately doubling in size. 

However, the cohort replacement of 6.08 in 2003 buffered the effect of the significant decline in the cohort 

replacement rate of 0.32 in 2007. In consideration of the almost 7-fold decrease in population in 2007, and the 

dry years that have followed, NMFS concludes that the winter-run are at high risk of extinction based on 

population size.

6.3.3.4. Population Growth Rate

The productivity of a population (i.e., production over the entire life cycle) can reflect conditions (e.g., 

environmental conditions) that influence the dynamics of a population and determine abundance. In turn, the 

productivity of a population allows an understanding of the performance of a population across the landscape 

and habitats in which it exists and its response to those habitats (McElhany et al. 2000). In general, declining 

productivity equates to declining population abundance. McElhany et al.(2000) suggested a population’s 

natural productivity should be sufficient to maintain its abundance above the viable level (a stable or increasing 

population growth rate). This guideline seems reasonable in the absence of numeric abundance targets. Winterrun Chinook salmon have declined substantially from historic levels. The one remaining population of winterrun on the mainstem Sacramento River is also the entire current ESU. Good et al. (2005) reported the medium 

population growth rate at slightly under 1.000 (Table 6). Although the population growth rate (indicated by the 

cohort replacement rate λ) increased since the late 1990s, it drastically decreased in 2007 and 2008, indicating 

that the population is not replacing itself, and is at a high risk of extinction in the foreseeable future. 

6.3.3.5. Spatial Structure

In general, there is less information available on how spatial processes relate to salmonid viability than there is 

for the other VSP parameters (McElhany et al. 2000). Understanding the spatial structure of a population is 

important because the population structure can affect evolutionary processes and, therefore, alter the ability of a 
77

population to adapt to spatial or temporal changes in the species’ environment (McElhany et al. 2000). The 

spatial structure of Sacramento River winter-run Chinook salmon resembles that of a panmictic population, 

where there are no subpopulations, and every mature male is equally likely to mate with every other mature 

female. The four historical independent populations of winter-run have been reduced to one population, 

resulting in a significant reduction in their spatial diversity. An ESU comprised of one population is not viable 

because it is unlikely to be able to adapt to significant environmental changes. A single catastrophe (e.g., 

volcanic eruption of Lassen Peak, prolonged drought which depletes the cold water pool at Lake Shasta, or 

some related failure to manage cold water storage, spill of toxic materials, or a disease outbreak) could extirpate 

the entire winter-run ESU if its effects persisted for 3 or more years. The majority of winter-run return to 

spawn in 3 years, so a single catastrophe with effects that persist for at least 3 years would affect all of the 

winter-run cohorts. Therefore, NMFS concludes that winter-run are at a high risk of extinction based on spatial 

structure.

6.3.3.6. Diversity

Diversity, both genetic and behavioral, is critical to success in a changing environment. Salmonids express 

variation in a suite of traits, such as anadromy, morphology, fecundity, run timing, spawn timing, juvenile 

behavior, age at smolting, age at maturity, egg size, developmental rate, ocean distribution patterns, male and 

female spawning behavior, and physiology and molecular genetic characteristics. The more diverse these traits 

(or the more these traits are not restricted), the more adaptable a population is, and the more likely that 

individuals, and therefore the species, would survive and reproduce in the face of environmental variation 

(McElhany et al. 2000). However, when this diversity is reduced due to loss of entire life history strategies or 

to loss of habitat used by fish exhibiting variation in life history traits, the species is in all probability less able 

to survive and reproduce given environmental variation. 

The primary factor affecting the diversity of winter-run is the limited area of spawning habitat available on the 

main stem Sacramento River downstream of Keswick Dam. This specific and narrow spawning habitat limits 

the flexibility and variation in spawning locations for winter-run to tolerate environmental variation. For 

example, a catastrophe on the main stem Sacramento River could affect the entire population, and therefore, 

ESU. However, with the majority of spawners being 3 years old, winter-run do reserve some genetic and 

behavioral variation in that in any given year, two cohorts are in the marine environment, and therefore, not 

exposed to the same environmental stressors as their freshwater cohorts. Although the Livingston Stone 
78

National Fish Hatchery (LSNFH) is characterized as one of the best examples of a conservation hatchery 

operated to maximize genetic diversity and minimize domestication of the offspring produced in the hatchery, it 

still faces some of the same diversity issues as other hatcheries in reducing the diversity of the naturallyspawning population. Therefore, Lindley et al. (2000) characterizes hatchery influence as a looming concern 

with regard to diversity. Even with a small contribution of hatchery fish to the natural spawning population, 

hatchery contributions could compromise the long term viability and extinction risk of winter-run. NMFS 

concludes that the current diversity in this ESU is much reduced compared to historic levels, and that winter-run 

are at a high risk of extinction based on the diversity VSP parameter. 

6.3.3.7. Critical Habitat

NMFS designated critical habitat for this species on June 16, 1993 (58 FR 33212). The designated critical 

habitat for winter-run includes the Sacramento River from Keswick Dam (RM 302 to Chipps Island (RM 0) at 

the westward margin of the Delta; all waters from Chipps Island westward to Carquinez Bridge, including 

Honker Bay, Grizzly Bay, Suisun Bay, and Carquinez Strait; all waters of San Pablo Bay westward of the 

Carquinez Bidge; and all waters of San Francisco Estuary to the Golden Gate Bridge north of the San 

Francisco/Oakland Bay Bridge. In the Sacramento River, critical habitat includes the river water column, river 

bottom, and adjacent riparian zone (limited to those areas above a streambank that provide cover and shade to 

the nearshore aquatic areas) used by fry and juveniles for rearing. In the area westward of Chipps Island, 

critical habitat includes the estuarine water column and essential foraging habitat and food resources used by 

winter-run as part of their juvenile emigration or adult spawning migration.

In designating critical habitat, NMFS considers the following requirements of the species: (1) space for 

individual and population growth, and for normal behavior; (2) food, water, air, light, minerals, or other 

nutritional or physiological requirements; (3) cover or shelter; (4) sites for breeding, reproduction, or rearing 

offspring; and generally, (5) habitats that are protected from disturbance or are representative of the historic 

geographical and ecological distributions of a species [see 50 CFR 424.12(b)]. In addition to these factors, 

NMFS also focuses on the known physical and biological features (essential features) within the designated area 

that are essential to the conservation of the species and that may require special management considerations or 

protection. These essential features may include, but are not limited to, spawning sites, food resources, water 

quality and quantity, and riparian vegetation and other natural cover.
79

NMFS identified specific water temperature criteria, minimum instream flow criteria, and water quality 

standards as essential physical features (PCEs) of the ESU’s habitat for species conservation. In addition, 

biological features vital for the Sacramento River winter-run Chinook salmon include unimpeded adult 

upstream migration routes, spawning habitat, egg incubation and fry emergence areas, rearing areas for 

juveniles, and unimpeded downstream migration routes for juveniles.

The designated critical habitat of this ESU has not been evaluated for the conservation value of individual 

subbasins or river sections. However, since spawning, rearing, and migration of the winter-run race is restricted 

to the mainstem of the Sacramento River, the entire Sacramento River is considered of high conservation value. 

The Delta is similarly considered of high conservation value for rearing and migration.

As there is overlap in designated critical habitat for both the Sacramento River winter-run Chinook salmon and 

the spring-run Chinook salmon, the conditions of PCEs for both ESUs are similar. The current condition of

PCEs for the Sacramento River winter-run Chinook salmon indicates that they are not currently functioning or 

are degraded. Their conditions are likely to maintain low population abundances across the ESU. Spawning 

and rearing PCEs are especially degraded by high water temperature caused by the loss of access to historic 

spawning areas in the upper watersheds where water maintain lower temperatures. The rearing PCE is further 

degraded by floodplain habitat disconnected (with the exception of the Yolo Bypass floodplain) from the 

mainstems of larger rivers throughout the Sacramento River watershed. The migration PCE is also degraded by 

the lack of natural cover along the migration corridors. Rearing and migration PCEs are further affected by 

pollutants entering the surface waters and riverine sediments as contaminated stormwater runoff, aerial drift and 

deposition, and via point source discharges. Juvenile migration is obstructed by water diversions along 

Sacramento River and by two large state and federal water-export facilities in the Sacramento-San Joaquin 

Delta.

6.4. California Central Valley Steelhead

The California Central Valley (CCV) steelhead DPS includes all naturally spawned steelhead populations below 

natural and manmade impassable barriers in the Sacramento and San Joaquin Rivers and their tributaries, 
80

excluding steelhead from San Francisco and San Pablo Bays and their tributaries, as well as two artificial 

propagation programs: the Coleman NFH, and Feather River Hatchery steelhead hatchery programs (

Figure 9).

6.4.1.1. Life History

Steelhead can be divided into two life history types, summer-run steelhead and winter-run steelhead, based on 

their state of sexual maturity at the time of river entry and the duration of their spawning migration, streammaturing and ocean-maturing. Only winter steelhead are currently found in Central Valley rivers and streams 

(McEwan and Jackson 1996a). At present, summer steelhead are found only in northern California coast 

drainages, mostly in tributaries of the Eel, Klamath, and Trinity River systems (McEwan and Jackson 1996a).

The CCV steelhead winter steelhead have the longest freshwater migration of any population of winter 

steelhead. CCV steelhead generally leave the ocean from August through April (Busby et al. 1996), and spawn 

from December through April, with peaks from January through March, in small streams and tributaries where 

cool, well oxygenated water is available year-round (Hallock et al. 1961, McEwan and Jackson 1996a). Most 

spawning habitat for steelhead in the Central Valley is located in areas directly downstream of dams containing 

suitable environmental conditions for spawning and incubation. The length of time it takes for eggs to hatch 

depends mostly on water temperature. Hatching of steelhead eggs in hatcheries takes about 30 days at 51°F. 

Fry emerge from the gravel usually about 4 to 6 weeks after hatching, but factors such as redd depth, gravel 

size, siltation, and temperature can affect emergence timing (Shapovalov and Taft 1954). Newly emerged fry 

move to the shallow, protected areas associated with the stream margin (McEwan and Jackson 1996a). 

Steelhead rearing during the summer occurs primarily in higher velocity areas in pools, although young of the 

year also are abundant in glides and riffles. Both spawning areas and migratory corridors comprise rearing 

habitat for juveniles, which feed and grow before and during their outmigration. Non-natal, intermittent 

tributaries, and inundated floodplains also may be used for juvenile rearing. On flood years when the Yolo 

Bypass (floodplain to the Sacramento River) is inundated, steelhead find exceptional rearing and foraging 

opportunities (Sommer et al. 2001). Steelhead migratory corridors are downstream of the spawning areas and 

include the lower mainstems of the Sacramento and San Joaquin rivers and the Delta. 
81

Productive steelhead habitat is characterized by complexity, primarily in the form of large and small woody 

debris. Cover is an important habitat component for juvenile steelhead both as velocity refugia and as a means 

of avoiding predation (Meehan and Bjornn 1991).

Juvenile steelhead emigrate episodically from natal streams during fall, winter, and spring high flows (Table 

12). Emigrating CV steelhead use the lower reaches of the Sacramento River and the Delta for rearing and as a 

migration corridor to the ocean. Juvenile CV steelhead feed mostly on drifting aquatic organisms and terrestrial 

insects and will also take active bottom invertebrates (Moyle 2002). 

Hallock et al. (1961) found that juvenile steelhead in the Sacramento River basin migrate downstream during 

most months of the year, but the peak period of emigration occurred in March and April, with a much smaller 

peak in the fall. Emigrating CCV steelhead use the lower reaches of the Sacramento River and the Delta for 

rearing and as a migration corridor to the ocean. Some juvenile steelhead may use tidal marsh areas, non-tidal 

freshwater marshes, and other shallow water areas in the Delta as rearing areas for short periods prior to their 

final emigration to the sea (Hallock et al. 1961). Nobriga and Cadrett (2003) also have verified these temporal 

findings based on analysis of captures at Chipps Island, Suisun Bay.

Table 12. Temporal occurrence of adult and juvenile CCV steelhead in the Central Valley. Darker shades indicate 

months of greatest relative abundance.

Adult migration

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. River basina,c

Sac. R. RBDDb,c

Mill, Deer creeksd

Sac. R. Freemont Weire

San Joaquin Riverf

Juvenile migration

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. River at RBDD a,b

Sac. River at KLb,g,h

Chipps Islandi

Mossdaleg

Woodbridge Damg

Sac. R. at Hoodj
82

RBDD=Red Bluff Diversion Dam KL = Knights Landing

Relative Abundance: = High = Medium = Low

a Hallock et al. (1961); bMcEwan (2001b); c USFWS (unpublished (1995) data); d CDFG (1995); e Bailey (1954); f CDFG Steelhead 

Report Card Data; g CDFG (unpublished data); h Snider and Titus (2000b); i Nobriga and Cadrett (2003);

j Schaffter (1980, 1997).
83

Figure 9. CCV steelhead distribution.
84

6.4.1.2. Status and Trends

NMFS originally listed CCV steelhead as threatened on March 19, 1998, and reaffirmed their threatened status 

on January 5, 2006 (71 FR 834). The CCV steelhead DPS may have consisted of 81 historical and independent 

populations (Lindley et al. 2006b). Spatial structure and patchiness strongly influenced suitable habitats being 

isolated due largely to high summer temperatures on the valley floor.

Over the past 30 years, the naturally-spawned steelhead populations in the upper Sacramento River have 

declined substantially. Hallock et al.(1961) estimated an average of 20,540 adult steelhead through the 1960s in 

the Sacramento River, upstream of the Feather River. Steelhead counts at the RBDD declined from an average 

of approximately 8,000 for the period of 1967 to 1977, to an average of approximately 2,000 through the early 

1990s, with an estimated total annual run size for the entire Sacramento-San Joaquin system, based on RBDD 

counts, to be no more than 10,000 adults (McEwan and Jackson 1996a, McEwan 2001b). Steelhead escapement 

surveys at RBDD ended in 1993 due to changes in dam operations. 

Nobriga and Cadrett (2003) compared CWT and untagged (wild) steelhead smolt catch ratios at Chipps Island 

trawl from 1998 through 2001 to estimate that about 100,000 to 300,000 steelhead juveniles are produced 

naturally each year in the Central Valley. Good et al. (2005) made the following conclusion based on the 

Chipps Island data: 

"If we make the fairly generous assumptions (in the sense of generating large estimates of spawners) 

that average fecundity is 5,000 eggs per female, 1 percent of eggs survive to reach Chipps Island, and 

181,000 smolts are produced (the 1998-2000 average), about 3,628 female steelhead spawn naturally in 

the entire Central Valley. This can be compared with McEwan's (2001b) estimate of 1 million to 2 

million spawners before 1850, and 40,000 spawners in the 1960s."

The species’ present distribution has been greatly reduced with about 80% of historic habitat lost behind dams 

and about 38% of habitat patches that supported independent populations are no longer accessible to steelhead 

(Lindley et al. 2006b). Existing wild steelhead stocks in the Central Valley are mostly confined to the upper 

Sacramento River and its tributaries, including Antelope, Deer, and Mill Creeks and the Yuba River. 

Populations may exist in Big Chico and Butte Creeks. A few wild steelhead are produced in the American and 
85

Feather Rivers (Good et al. 2005, McEwan and Jackson 1996a). Steelhead have also been observed in Clear 

Creek and Stanislaus River (Good et al. 2005, Demko and Cramer 2000). Until recently, steelhead were 

considered extirpated from the San Joaquin River system. Recent monitoring has detected small self-sustaining 

populations of steelhead in the Stanislaus, Mokelumne, Calaveras, and other streams previously thought to be 

void of steelhead (Good et al. 2005). In 2004, a total of 12 steelhead smolts were collected in monitoring trawls 

at the Mossdale station in the lower San Joaquin River (CDFG unpublished data).

Historic CCV steelhead run size may have approached one to two million adults annually (McEwan 2001b). By 

the early 1960s, the steelhead run size had declined to about 40,000 adults (McEwan 2001b). Steelhead were 

counted at the Red Bluff Diversion Dam (RBDD) up until 1993. Counts at the dam declined from an average of 

11,187 for the period of 1967 to 1977, to an average of approximately 2,000 through the early 1990s. An 

estimated total annual run size for the entire Sacramento-San Joaquin system was no more than 10,000 adults

during the early 1990s (McEwan and Jackson 1996a, McEwan 2001b). Based on catch ratios at Chipps Island 

in the Delta and using some generous assumptions regarding survival, the average number of CV steelhead 

females spawning naturally in the entire Central Valley during the years 1980 to 2000 was estimated at about 

3,600 (Good et al. 2005).

CCV steelhead lack annual monitoring data for calculating trends and lambda. However, the RBDD counts and 

redd counts up to 1993 and later sporadic data show that the DPS has had a significant long-term downward 

trend in abundance (NMFS 2009a).

The CCV steelhead distribution ranged over a wide variety of environmental conditions and likely contained 

biologically significant amounts of spatially structured genetic diversity (Lindley et al. 2006b). Thus, the loss 

of populations and reduction in abundances have reduced the large diversity that existed within the DPS. The 

genetic diversity of the majority of CCV steelhead spawning runs is also compromised by hatchery-origin fish. 

In order to determine the current likelihood of the CV steelhead DPS becoming viable, we used the historical 

population structure of CV steelhead presented in Lindley et al. (2006b, 2007) and the concept of VSP for 

evaluating populations described by McElhany et al. (2000), and applied by Lindley et al (2007) to the CV 

steelhead DPS.
86

Table 8 provides various quantitative criteria to evaluate the risk of extinction. The following provides the 

evaluation of the likelihood of the threatened CV steelhead DPS becoming viable based on the VSP parameters 

of population size, population growth rate, spatial structure, and diversity.

6.4.1.3. Population Size

Estimated natural CV steelhead escapement in the upper Sacramento River has declined substantially from 1967 

through 1993. There is still a nearly complete lack of steelhead monitoring in the Central Valley (Good et al. 

2005), and therefore, data are lacking regarding a definitive population size for CV steelhead. However, the 

little data that exist indicate that the CV steelhead population continues to decline (Good et al. 2005).

6.4.1.4. Population Growth Rate

CV steelhead has shown a pattern of a negative growth rate since the late 1960s. Good et al. (2005) provided no 

indication that this trend has changed since the last CV steelhead population census in 1993. The λ value for 

CV steelhead is less than 1.000 (Table 6; (Good et al. 2005).

6.4.1.5. Spatial Structure

Lindley et al. (2006b) identified 81 historical and independent populations within the CV steelhead DPS. These 

populations form 8 clusters, or diversity groups, based on the similarity of the habitats they occupied for 

spawning and rearing. About 80 percent of the habitat that was historically available to CV steelhead is now 

behind impassable dams, and 38 percent of the populations have lost all of their habitats. Although much of the 

habitat has been blocked by impassable dams, or degraded, small populations of CV steelhead are still found 

throughout habitat available in the Sacramento River and many of the tributaries, and some of the tributaries to 

the San Joaquin River.

6.4.1.6. Diversity

Diversity, both genetic and behavioral, provides a species the opportunity to respond to sudden environmental 

changes. CV steelhead naturally experience the most diverse life history strategies of the listed Central Valley 

anadromous salmonid species. In addition to being iteroparous, they reside in freshwater for 2-4 years before 
87

emigrating to the ocean. However, as the species’ abundance decreases, and spatial structure of the DPS is 

reduced, it has less flexibility to respond to changes in the environment. CV steelhead abundance and growth 

rate continue to decline, largely the result of a significant reduction in the diversity of habitats available to CV 

steelhead (Lindley et al. 2006b). The genetic diversity of CV steelhead is also compromised by hatchery-origin 

fish, which likely comprise the majority of the natural spawning run, placing the natural populations at high risk 

of extinction (Lindley et al. 2007). Consistent with the life history strategy of winter-run and spring-run

Chinook salmon, some genetic and behavioral variation is conserved in that in any given year, there are 

additional cohorts in the marine environment, and therefore, not exposed to the same environmental stressors as 

their freshwater cohorts.

6.4.1.7. Critical Habitat

NMFS designated critical habitat for CCV steelhead on September 2, 2005 (70 FR 52488). Critical habitat 

includes stream reaches such as those of the Sacramento, Feather, and Yuba Rivers, and Deer, Mill, Battle, and 

Antelope creeks in the Sacramento River basin; the lower San Joaquin River to the confluence with the Merced 

River, including its tributaries, and the waterways of the Delta (Figure 10). The total area of critical habitat 

includes about 2,300 miles of stream habitat and about 250 square miles of estuarine habitat in the San 

Francisco-San Pablo-Suisan Bay estuarine complex.

 

There are 67 occupied HAS watersheds within the freshwater and estuarine range of this DPS. Twelve 

watersheds received a low rating, 18 received a medium rating, and 37 received a high rating of conservation 

value to the ESU (NMFS 2005). Four of these HSA watersheds comprise portions of the San Francisco-San 

Pablo-Suisun Bay estuarine complex which provides rearing and migratory habitat for this ESU.

Table 13. CCV spring-run Chinook salmon CALWATER HSA watersheds with conservation values.

HUC 4 Subbasin

HUC 5 Watershed conservation Value (CV)

High CV PCE(s) 1 Medium 

CV PCE(s) 1

Low CV PCE(s) 1

San Francisco Bay 1 2 0 0

South Bay 0 0 1 2

San Pablo 1 2 0 0

Suisun Bay 1 2 0 0

Tehama 1 1, 2, 3 1 1, 2, 3 0

Whitmore 3 1, 2, 3 2 1, 2, 3 2 1, 2, 3

Redding 2 1, 2, 3 0 0

Eastern Tehama 4 1, 2, 3 1 1, 2, 3 1 1, 2, 3
88

HUC 4 Subbasin

HUC 5 Watershed conservation Value (CV)

High CV PCE(s) 1 Medium 

CV PCE(s) 1

Low CV PCE(s) 1

Sacramento Delta 1 1, 2, 3 0 0

Valley Putah-Cache 0 2 1, 2, 3 0

American River 0 1 1, 2, 3 0

Marysville 2 1, 2, 3 1 1, 2, 3 0

Yuba River 2 1, 2, 3 0 2 1, 2, 3

Valley-American 2 1, 2, 3 0 0

Colusa Basin 4 1, 2, 3 0 0

Butte Creek 1 1, 2, 3 1 1, 2, 3 1 1, 2, 3

Ball Mountain 1 1, 2, 3 0 0

Shasta Bally 2 1, 2, 3 3 1, 2, 3 0

North Valley Floor 1 1, 2, 3 1 1, 2, 3 1 1, 2, 3

Middle Sierra 0 0 4 1, 2, 3

Upper Calaveras 1 1, 2, 3 0 0

Stanislaus River 1 1, 2, 3 0 0

San Joaquin Valley 

Floor 4

1, 2, 3 31, 2, 3 0

Delta-Mendota 

Canal 1

1, 2, 3 11, 2, 3 0

North Diablo Range 0 1 0

San Joaquin Delta 1 1, 2, 3 0 0

Total 37 18 12

1 Numbers in parenthesis refers to the dominant (in river miles) PCE(s) within the HUC 5 watersheds. PCE 1 is spawning 

and rearing, 2 is rearing and migration, and 3 is migration and presence. PCEs with < means that the number of river 

miles of the PCE is much less than river miles of the other PCE.

The current condition of CCV steelhead critical habitat is degraded, and does not provide the conservation value 

necessary for species recovery (Table 13). In addition, the Sacramento-San Joaquin River Delta, as part of 

CCV steelhead designated critical habitat, provides very little function necessary for juvenile CCV steelhead 

rearing and physiological transition to salt water. 

The spawning PCE is subject to variations in flows and temperatures, particularly over the summer months. 

Some complex, productive habitats with floodplains remain in the system and flood bypasses (e.g., Yolo and 

Sutter bypasses). However, the rearing PCE is degraded by the channelized, leveed, and riprapped river reaches 

and sloughs that are common in the Sacramento-San Joaquin system and which typically have low habitat 

complexity, low abundance of food organisms, and offer little protection from either fish or avian predators. 

Stream channels commonly have elevated temperatures.

The current conditions of migration corridors are substantially degraded. Both migration and rearing PCEs are 

affected by dense urbanization and agriculture along the mainstems and in the Delta which contribute to 
89

reduced water quality by introducing several contaminants. In the Sacramento River, the migration corridor for 

both juveniles and adults is obstructed by the RBDD gates which are down from May 15 through September 15. 

The migration PCE is also obstructed by complex channel configuration making it more difficult for CCV 

steelhead to migrate successfully to the western Delta and the ocean. In addition, the state and federal 

government pumps and associated fish facilities change flows in the Delta which impede and obstruct for a 

functioning migration corridor that enhance migration. The estuarine PCE, which is present in the Delta, is 

affected by contaminants from agricultural and urban runoff and release of wastewater treatment plants effluent.
90

Figure 10. California Central Valley Steelhead Conservation Value per Sub-area.
91

6.4.1.8. The Yolo Bypass floodplain habitat

Sommer et al. (2001, 2004, 2005b) have documented the importance of the Yolo Bypass as beneficial rearing 

habitat for juvenile salmonids relative to the riverine habitat in the near-by Sacramento River. Indeed, the 

24,000-ha Yolo Bypass floodplain is the primary floodplain of the Sacramento River, and of the combined 

Sacramento-San Joaquin delta. When inundated the Yolo provides significant rearing habitat for all species of 

Chinook, for steelhead and other fishes. Sommer et al. (2001) estimated that complete inundation of the Yolo 

Bypass creates a wetted area approximately 10 times larger than the adjacent reach of the Sacramento River, 

and is equivalent to a doubling of the wetted area of the entire delta portion of the San Francisco Estuary. Much 

of the floodplain habitat consists of broad shoals composed of soil and vegetation that are typical of the lowvelocity conditions selected by young salmon (Everest and Chapman 1972, Healey 1982). Sommer et al. 

(2001) demonstrated Chinook salmon rearing in the Yolo Bypass floodplain have higher apparent growth rates 

than those that remain in the Sacramento River channel. The Yolo has higher temperatures and a more diverse 

and abundant food supply (Sommer et al. 2001).

The Yolo Bypass floods an average of every other year, typically under high flow periods in winter and spring. 

The Yolo Bypass has a complex hydrology, with inundation possible from several different sources. The 

floodplain typically has a peak inundation period during January – March but can flood as early as October and 

as late as June. The primary input to the Yolo Bypass is through the Fremont Weir in the north, which conveys 

floodwaters from the Sacramento and Feather rivers. During major storm events, additional water enters from 

the east via the Sacramento Weir, adding flow from the American and Sacramento rivers. Flow also enters the 

Yolo Bypass from several small streams on its western margin, including Knights Landing, Ridge Cut, Cache 

Creek, and Putah Creek. Hydraulic residence times are typically longer in the Yolo Bypass than in the 

Sacramento River (Sommer et al. 2004). 

Floodwaters recede from the northern and western portions of the Bypass along relatively even elevation 

gradients of 0.09 percent west – east and 0.01 percent north – south into a perennial channel on the eastern edge 

(the “toe-drain”). The receding waters then join the Sacramento River near Rio Vista. The majority of the Yolo 

Bypass is managed for wildlife in a mosaic that includes riparian, wetland, upland, and perennial pond habitats; 

however a dominant land use is agriculture, of which rice growing is a significant component.
92

During typical flood events between January and March, all three listed salmonid species would benefit from 

the enhanced rearing and growing conditions (as described in Sommer et al. (2001)) the Yolo Bypass provides. 

Juvenile spring- and winter-run Chinook would be passing through in relatively high abundance in January and 

medium to low abundance February to March (Table 7 and Table 11). Steelhead would be passing through in 

medium abundance January – February but would be in relatively high abundance during March (Table 12). 

Sommer et al. (2001)found salmon in all regions of the floodplain and on all substrate types. Juvenile Chinook 

were in a high percentage of the samples in each region of the floodplain: (1) Fremont Weir (100 percent, n = 

13 samples), (2) Cache Creek Sinks (50 percent, n = 16 samples), (3) Yolo Bypass Wildlife Area (77 percent, n 

= 22 samples), 
93

7. Environmental Baseline

By regulation, environmental baselines for Opinions include the past and present impacts of all state, federal or 

private actions and other human activities in the action area, the anticipated impacts of all proposed federal 

projects in the action area that have already undergone formal or early section 7 consultation, and the impact of 

state or private actions which are contemporaneous with the consultation in process (50 CFR §402.02). The 

environmental baseline for this Opinion includes a general description of the natural and anthropogenic factors 

influencing the current status of the three listed Pacific salmonids and the environment within the action area.

Our summary of the environmental baseline complements the information provided in the Status of Listed 

Resources section of this Opinion, and provides the background necessary to understand information presented 

in the Effects of the Proposed Action, and Cumulative Effects sections of this Opinion. We then evaluate the 

consequences of these activities in combination with the environmental baseline to determine the likelihood of 

jeopardy or adverse modification of designated critical habitat.

The proposed action under consultation is focused geographically on the aquatic ecosystems in California’s 

Central Valley. Accordingly, the environmental baseline for this consultation focuses on the general status and 

trends of the aquatic ecosystem there and the consequences of that status for the listed salmonids under NMFS’ 

jurisdiction. We describe the principal natural phenomena affecting all listed Pacific salmonids under NMFS 

jurisdiction in the action area. The action area for the proposed action encompasses the entire freshwater range 

or a large portion of the freshwater range of the listed fish species and their proposed or designated critical 

habitat in this consultation. Therefore, we refer the reader to the Status of the Species section for general 

information on the species’ biology, ecology, status, and population trends at the species scale.

We further describe anthropogenic factors through the predominant land and water uses within the region. 

Background information on pesticides in the aquatic environment is also provided. This context illustrates how 

the physical and chemical health of regional waters and the impact of human activities have contributed to the 

current status of listed resources in the action area.

Much of the freshwater habitat overlaps geographically between Sacramento River winter-run Chinook, CV 

spring-run Chinook, and CV steelhead. Therefore, most of the baseline factors responsible for their current 

statuses are similar. Therefore, each of the following factors applies to Sacramento River winter-run Chinook, 
94

CV spring-run Chinook, and CV steelhead unless specified.

7.1. Natural Mortality Factors

Available data indicate high natural mortality rates for salmonids, especially in the open ocean/marine 

environment. According to Bradford et al. (1997), salmonid mortality rates range from 90 to 99%, depending 

on the species, the size at ocean entry, and the length of time spent in the ocean. Predation, inter- and 

intraspecific competition, food availability, smolt quality and health, and physical ocean conditions likely 

influence the survival of salmon in the marine environment (Brodeur et al. 2004, Bradford et al. 1997). In 

general, in freshwater rearing habitats, the natural mortality rate averages about 70% for all salmonid species 

(Bradford et al. 1997). For example, past studies in the Pacific Northwest suggest that the average freshwater 

survival rate (from egg to smolt) is 2 to 3% throughout the region (Bradford et al. 1997, Marshall and Britton 

1990). A number of suspected causes contributing to natural mortality include parasites and/or disease, 

predation, water temperature, low water flow, wildland fire, and oceanographic features and climatic variability. 

7.1.1. Parasites and/or Disease 

Most young fish are highly susceptible to disease during the first two months of life. The cumulative mortality 

in young animals can reach 90 to 95%. Although fish disease organisms occur naturally in the water, native 

fish have co-evolved with them. Fish can carry these diseases at less than lethal levels (Walker and Foott 1993, 

Kier Associates 1991, Foott et al. 2003). However, disease outbreaks may occur when water quality is 

diminished and fish are stressed from crowding and diminished flows (Spence et al. 1996b, Guillen 2003). 

Young salmonid species may become stressed and lose their resistance in higher temperatures (Spence et al. 

1996b). Consequently, diseased fish become more susceptible to predation and are less able to perform 

essential functions, such as feeding, swimming, and defending territories (McCullough 1999). Examples of 

parasites and disease for salmonids include whirling disease, infectious hematopoietic necrosis (IHN), sea-lice 

(Lepeophtheirus salmonis), Henneguya salminicola, Ichthyopthirius multifiliis or Ich, and Columnaris

(Flavobacterium columnare).
95

Whirling disease is a parasitic infection caused by the microscopic parasite Myxobolus cerebrali. Infected fish 

continually swim in circular motions and eventually expire from exhaustion. The disease occurs in the wild and 

in hatcheries and results in losses to fry and fingerling salmonids, especially rainbow trout. The disease is 

transmitted by infected fish and fish parts and birds. 

IHN is a viral disease in many wild and farmed salmonid stocks in the Pacific Northwest. This disease affects 

rainbow/steelhead trout, cutthroat trout (Salmo clarki), brown trout (Salmo trutta), Atlantic salmon (Salmo 

salar), and Pacific salmon including Chinook, sockeye, chum, and coho salmon. The virus is triggered by low 

water temperatures and is shed in the feces, urine, sexual fluids, and external mucus of salmonids. 

Transmission is mainly from fish to fish, primarily by direct contact and through the water.

Additionally, ich (a protozoan) and Columnaris (a bacterium) are two common fish diseases that were 

implicated in the massive kill of adult salmon in the Lower Klamath River in September 2002 (Guillen 2003, 

CDFG 2003). 

7.1.2. Predation

Salmonids are exposed to high rates of natural predation, during freshwater rearing and migration stages, as 

well as during ocean migration. Salmon along the U.S. west coast are prey for marine mammals, birds, sharks, 

and other fishes. Concentrations of juvenile salmon in the coastal zone experience high rates of predation. In 

the Pacific Northwest, the increasing size of tern, seal, and sea lion populations may have reduced the survival 

of some salmon ESUs/DPSs. 

7.1.2.1. Marine Mammal Predation 

Marine mammals are known to attack and eat salmonids. Harbor seals (Phoca vitulina), California sea lions 

(Zalophus californianus), and killer whales (Orcinus orca) prey on juvenile or adult salmon. Killer whales have 

a strong preference for Chinook salmon (up to 78% of identified prey) during late spring to fall (Hanson et al. 

2005, Hard et al. 1992, Ford and Ellis 2006). Generally, harbor seals do not feed on salmonids as frequently as 

California sea lions (Pearcy 1997). California sea lions from the Ballard Locks in Seattle, Washington have 

been estimated to consume about 40% of the steelhead runs since 1985/1986 (Gustafson et al. 1997). In the 

Columbia River, salmonids may contribute substantially to sea lion diet at specific times and locations (Pearcy 

1997). Spring Chinook salmon and steelhead are subject to pinniped predation when they return to the estuary 

as adults (NMFS 2006). Adult Chinook salmon in the Columbia River immediately downstream of Bonneville 
96

Dam have also experienced increased predation by California sea lions. In recent years, sea lion predation of 

adult Lower Columbia River winter steelhead in the Bonneville tailrace has increased. This prompted ongoing 

actions to reduce predation effects. They include the exclusion, hazing, and in some cases, lethal take of marine 

mammals near Bonneville Dam.

7.1.2.2. Avian Predation

Large numbers of fry and juveniles are eaten by birds such as mergansers (Mergus spp.), common murre (Uria 

aalage), gulls (Larus spp.), and belted kingfishers (Megaceryle alcyon). Avian predators of adult salmonids 

include bald eagles (Haliaeetus leucocephalus) and osprey (Pandion haliaetus) (Pearcy 1997). Caspian terns 

(Sterna caspia) and cormorants (Phalacrocorax spp.) also take significant numbers of juvenile or adult salmon. 

Stream-type juveniles, especially yearling smolts from spring-run populations, are vulnerable to bird predation 

in the estuary. This vulnerability is due to salmonid use of the deeper, less turbid water over the channel, which 

is located near habitat preferred by piscivorous birds (Binelli et al. 2005). For example, research shows that 

subyearlings from the LCR Chinook salmon ESU are also subject to tern predation. This may be due to the 

long estuarine residence time of the LCR Chinook salmon (Ryan et al. 2006). Caspian terns and cormorants 

may be responsible for the mortality of up to 6% of the outmigrating stream-type juveniles in the Columbia 

River basin (Roby et al. 2006, Collis 2007). 

Antolos et al. (2005) quantified predation on juvenile salmonids by Caspian terns nesting on Crescent Island in 

the mid-Columbia reach. Between 1,000 and 1,300 adult terns were associated with the colony during 2000 and 

2001, respectively. These birds consumed about 465,000 juvenile salmonids in the first and approximately 

679,000 salmonids in the second year. However, caspian tern predation in the estuary was reduced from a total 

of 13,790,000 smolts to 8,201,000 smolts after relocation of the colony from Rice to East Sand Island in 1999. 

Based on PIT-tag recoveries at the colony, these were primarily steelhead for Upper Columbia River stocks. 

Less than 0.1% of the inriver migrating yearling Chinook salmon from the Snake River and less than 1% of the 

yearling Chinook salmon from the Upper Columbia were consumed. PIT-tagged coho smolts (originating 

above Bonneville Dam) were second only to steelhead in predation rates at the East Sand Island colony in 2007 

(Roby et al. 2008). There are few quantitative data on avian predation rates on Snake River sockeye salmon. 

Based on the above, avian predators are assumed to have a minimal effect on the long-term survival of Pacific 

salmon (NMFS 2008b).
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7.1.2.3. Fish Predation 

Predation is a threat to salmonids within this ESU/DPS, especially in the lower Feather River, the Sacramento 

River, and in the Delta where there are high densities of non-native species such as striped bass (Morone 

saxatilis), largemouth bass (Micropterus salmoides), and sunfishes (Centrarchidae spp.) and native fish species 

(e.g., pikeminnows (Ptychocheilus oregonensis)) that prey on outmigrating juveniles (NMFS 2011a). 

The primary fish predators in estuaries are probably adult salmonids or juvenile salmonids which emigrate at 

older and larger sizes than others. They include steelhead smolts preying on young Chinook smolts. Outside 

estuaries, many large non-salmonid populations reside just offshore and may consume large numbers of smolts. 

These fishes include Pacific hake (Merluccius productus), Pacific mackerel (Scomber japonicus), lingcod 

(Ophiodon elongates), spiny dogfish (Squalus acanthias), various rock fish, and lamprey (Pearcy 1992, 

Beamish and Neville 1995, Beamish et al. 1992).

7.1.3. Wildland Fire 

Wildland fires that are allowed to burn naturally in riparian or upland areas may benefit or harm aquatic species, 

depending on the degree of departure from natural fire regimes. Although most fires are small in size, large size 

fires increase the chances of adverse effects on aquatic species. Large fires that burn near the shores of streams 

and rivers can have biologically significant short-term effects. They include increased water temperatures, ash, 

nutrients, pH, sediment, toxic chemicals, and large woody debris (Rinne 2004, Buchwalter et al. 2004). 

Nevertheless, fire is also one of the dominant habitat-forming processes in mountain streams (Bisson et al. 

2003). As a result, many large fires burning near streams can result in fish kills with the survivors actively 

moving downstream to avoid poor water quality conditions (Rinne 2004, Greswell 1999). The patchy, mosaic 

pattern burned by fires provides a refuge for those fish and invertebrates that leave a burning area or simply 

spares some fish that were in a different location at the time of the fire (USFS 2000). Small fires or fires that 

burn entirely in upland areas also cause ash to enter rivers and increase smoke in the atmosphere, contributing 

to ammonia concentrations in rivers as the smoke adsorbs into the water (Greswell 1999). 

The presence of ash also has indirect effects on aquatic species depending on the amount of ash entry into the 

water. All ESA-listed salmonids rely on macroinvertebrates as a food source for at least a portion of their life 

histories. When small amounts of ash enter the water, there are usually no noticeable changes to the 

macroinvertebrate community or the water quality (Bowman and Minshall 2000). When significant amounts of 
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ash are deposited into rivers, the macroinvertebrate community density and composition may be moderately to 

drastically reduced for a full year with long-term effects lasting 10 years or more (Buchwalter et al. 2003, 

Buchwalter et al. 2004, Minshall et al. 2001). Larger fires can also indirectly affect fish by altering water 

quality. Ash and smoke contribute to elevated ammonium, nitrate, phosphorous, potassium, and pH, which can 

remain elevated for up to four months after forest fires (Buchwalter et al. 2003).

7.1.4. Oceanographic Features, Climatic Variability and Climate Change

Oceanographic features of the action area may influence prey availability and habitat for Pacific salmonids. 

These features comprise climate regimes which may suffer regime shifts due to climate changes or other 

unknown influences. The action area includes important spawning and rearing grounds and physical and 

biological features essential to the conservation of listed Pacific salmonids - i.e., water quality, prey, and 

passage conditions. These Pacific oceanographic conditions, climatic variability, and climate change may affect 

salmonids in the action area.

There is evidence that Pacific salmon abundance may have fluctuated for centuries as a consequence of 

dynamic oceanographic conditions (Beamish and Bouillon 1993, Beamish et al. 2009, Finney et al. 2002). 

Sediment cores reconstructed for 2,200-year records have shown that Northeastern Pacific fish stocks have 

historically been regulated by these climate regimes (Finney et al. 2002). The long-term pattern of the Aleutian 

Low pressure system has corresponded to the trends in salmon catch, to copepod production, and to other 

climate indices, indicating that climate and the marine environment may play an important role in salmon 

production. Pacific salmon abundance and corresponding worldwide catches tend to be large during naturallyoccurring periods of strong Aleutian low pressure causing stormier winters and upwelling, positive Pacific 

Decadal Oscillation (PDO), and an above average Pacific circulation index (Beamish et al. 2009). A trend of 

an increasing Aleutian Low pressure indicates high pink and chum salmon production and low production of 

coho and Chinook salmon (Beamish et al. 2009). The abundance and distribution of salmon and zooplankton 

also relate to shifts in North Pacific atmosphere and ocean climate (Francis and Hare 1994).

Over the past century, regime shifts have occurred as a result of the North Pacific’s natural climate regime. 

Reversals in the prevailing polarity of the PDO occurred around 1925, 1947, 1977, and 1989 (Hare and Mantua. 

2000, Mantua et al. 1997). The reversals in 1947 and 1977 correspond to dramatic shifts in salmon production 

regimes in the North Pacific Ocean (Mantua et al. 1997). During the pre-1977 climate regime, the productivity 

of salmon populations from the Snake River exceeded expectations (residuals were positive) when values of the 
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PDO were negative (Levin 2003). During the post-1977 regime when ocean productivity was generally lower 

(residuals were negative), the PDO was negative (Levin 2003).

A smaller, less pervasive regime shift occurred in 1989 (Hare and Mantua. 2000). Beamish et al.(2000) 

analyzed this shift and found a decrease in marine survival of coho salmon in Puget Sound and off the coast of 

California to Washington. Trends in coho salmon survival were linked over the southern area of their 

distribution in the Northeast Pacific to a common climatic event. The Aleutian Low Pressure Index and the 

April flows from the Fraser River also changed abruptly about this time (Beamish et al. 2000).

The Intergovernmental Panel on Climate Change (IPCC) has high confidence that some hydrological systems 

have been affected through increased runoff and earlier spring peak discharge in glacier- and snow-fed rivers 

and through effects on thermal structure and water quality of warming rivers and lakes (IPCC 2007). 

Oceanographic models project a weakening of the thermohaline circulation resulting in a reduction of heat 

transport into high latitudes of Europe, an increase in the mass of the Antarctic ice sheet, and a decrease in the 

Greenland ice sheet (IPCC 2001). These changes, coupled with increased acidification of ocean waters, are 

expected to have substantial effects on marine and hydrological productivity and food webs, including 

populations of salmon and other salmonid prey (Hard et al. 1992).

Carbon dioxide emissions are also predicted to have major environmental impacts along the west coast of North 

America during the 21st century and beyond (Climate Impacts Group (CIG) 2004, IPCC 2001). Eleven of the 

past 12 years (1995 - 2006) rank among the 12 warmest years in the instrumental record of global surface 

temperature since 1850 (IPCC 2007). The IPCC predicts that, for the next two decades, a warming of about 

0.2oC per decade will occur for a range of predicted carbon dioxide emissions scenarios (IPCC 2007). This 

warming trend continues in both water and air. Global average sea level has risen since 1961 at an average rate 

of 1.8 mm/year and since 1993 at 3.1 mm/year, with contributions from thermal expansion, melting glaciers and 

ice caps, and the polar ice sheets (IPCC 2007).

Poor environmental conditions for salmon survival and growth may be more prevalent with projected warming 

increases. Increasing climate temperatures can influence smolt development which is limited by time and 

temperature (McCormick et al. 2009). Food availability and water temperature may affect proper maturation 

and smoltification and feeding behavior (Mangel 1994). Climate change may also have profound effects on 
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seawater entry and marine performance of anadromous fish, including increased salinity intrusion in estuaries 

due to higher sea levels, as well as a projected decrease of seawater pH (Orr et al. 2005). There is evidence that 

Chinook salmon survival in the Pacific during climate anomalies and El Nino events changes as a result of a 

shift from predation- to competition-based mortality in response to declines in predator and prey abundances 

and increases in pink salmon abundance (Ruggerone and Goetz 2004). If climate change leads to an overall 

decrease in the availability of food, then returning fish will likely be smaller (Mangel 1994). Finally, future 

climatic warming could lead to alterations of river temperature regimes, which could further reduce available 

fish habitat (Yates et al. 2008).

Although the impacts of global climate change are less clear in the ocean environment, early modeling efforts 

suggest that increased temperatures will likely increase ocean stratification. This stratification coincides with 

relatively poor ocean habitat for most Pacific Northwest salmon populations (Climate Impacts Group (CIG) 

2004, IPCC 2001).

We expect changing weather and oceanographic conditions may affect prey availability, temperature and water 

flow in habitat conditions, and growth for all 28 ESUs/DPSs. Consequently, we expect the long-term survival 

and reproductive success for listed salmonids to be greatly affected by global climate change.

In addition to changes in hydrological regimes that will affect salmon, climate change will affect agriculture as 

rainfall and temperature patterns shift. Some crops currently well-suited for particular regions may instead be 

grown in alternate locations. Agricultural pest pressures are also likely to change over time. Both the shifts in 

crop location and pest pressure are likely to change pesticide use patterns.

7.2. Anthropogenic Mortality Factors

In this section we address anthropogenic threats in the geographic regions across the action area. Land use 

activities associated with logging, road construction, urban development, mining, agriculture, and recreation 

have significantly altered fish habitat quantity and quality. Impacts associated with these activities include: (1) 

alteration of migration corridors and streambank and channel morphology; (2) alteration of ambient stream 

temperatures; (3) degradation of water quality; (4) elimination or degradation of spawning and rearing habitat; 

(5) fragmentation of available habitats: and (6) removal or impairment of riparian vegetation – resulting in 

increased water temperatures and streambank erosion. 
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In the following discussion we provide information on habitat access, water development, land use, water 

quality, and hatchery operations. This is followed with a discussion on pesticide detections in the aquatic 

environment and highlights their background levels from past and ongoing anthropogenic activities. This 

information is pertinent to EPA’s proposed registration of thiobencarb in the U.S. and its territories. 

Thiobencarb has been in use for multiple decades, it has documented presence in our nation’s rivers, and thus 

over the years have contributing effects to the environmental baseline. In this section, we cover California’s 

special requirements for thiobencarb use in this state. As water temperature plays such a strong role in 

salmonid distribution, we also provide a general discussion of anthropogenic temperature impacts. Next, we 

discuss the health of riparian systems and floodplain connectivity, as this habitat is vital to salmonid survival. 

Finally, we provide a brief overview of the results of section 7 consultations relevant to this analysis.

7.2.1. Habitat Blockages

Hydropower, flood control, and water supply dams have permanently blocked or hindered salmonid access to 

historical spawning and rearing grounds. Yoshiyama et al. (1996)calculated that roughly 2,000 linear miles of 

salmon habitat was actually available before dam construction and mining, and concluded that 82 percent is not 

accessible today. The percentage of habitat loss for steelhead is presumable greater, because steelhead were 

more extensively distributed upstream than Chinook salmon. See Figure 15 for location of dams within 

ESU/DPSs in California.

As a result of migrational barriers, winter-run, spring-run, and steelhead populations have been confined to 

lower elevation mainstems that historically only were used for migration and rearing. Population abundances 

have declined in these streams due to decreased quantity, quality, and spatial distribution of spawning and 

rearing habitat (Lindley et al. 2009). Higher temperatures at these lower elevations during late-summer and fall 

are also a major stressor to adult and juvenile salmonids. According to Lindley et al.(2004a), of the four 

independent populations of winter-run that occurred historically, only one mixed stock of winter-run remains 

below Keswick Dam. 

Similarly, of the 19 independent populations of spring-run that occurred historically, only three independent 

populations remain in Deer, Mill, and Butte Creeks. Dependent populations of spring-run continue to occur in 

Big Chico, Antelope, Clear, Thomes, and Beegum creeks and the Yuba River, but rely on the extant 

independent populations for their continued survival. CV steelhead historically had at least 81 independent 
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populations based on Lindley et al.’s (2006a) analysis of potential habitat in the Central Valley. However, due 

to dam construction, access to 38 percent of all spawning habitat has been lost, as well as access to 80 percent 

of the historically available habitat. 

Juvenile downstream migration patterns have been altered by the presence of dams. Juvenile winter-run, and 

spring-run on the mainstem Sacramento River, arrive at any given location downstream of Keswick Dam earlier 

than historical, since they are hatched much further downstream and have less distance to travel. Therefore, in 

order to smolt at the same size and time as historical, they must rear longer within the Sacramento River. 

However, as will be discussed below, the mainstem Sacramento River is not conducive to the necessary habitat 

features that provide suitable rearing habitat for listed anadromous fish species, especially for an extended 

duration of time. 

The California Department of Water Resources (DWR) Suisun Marsh Salinity Control Gates (SMSCG) located 

in Montezuma Slough were installed in 1988, and are operated with gates and flashboards to decrease the 

salinity levels of managed wetlands in Suisun Marsh. The SMSCG have delayed or blocked passage of adult 

Chinook salmon migrating upstream (Edwards et al. 1996, Tillman et al. 1996, CDWR 2002). As a result of the 

SMSCG fish passage study and a term and condition in NMFS’ 2004 CVP/SWP operations Opinion, the boat 

lock has remained open since the 2001-2002 control season (CVP/SWP operations BA), and adult fish passage 

has improved. 

The Red Bluff Diversion Dam (RBDD) impedes adult salmonid passage throughout its May 15 through 

September 15 “gates in” period. Although there are fish ladders at the right and left banks, and a temporary 

ladder in the middle of the dam, they are not very efficient at passing fish. The range of effects resulting from 

delays in upstream migration at RBDD include delayed, but eventually successful spawning, to prespawn 

mortality and the complete loss of spawning potential in that fraction of the population (NMFS 2009a). 

7.2.2. Water Development

The diversion and storage of natural flows by dams and diversion structures on Central Valley waterways have 

depleted streamflows and altered the natural cycles by which juvenile and adult salmonids base their migrations. 

As much as 60 percent of the natural historical inflow to Central Valley watersheds and the Delta have been 

diverted for human uses. Depleted flows have contributed to higher temperatures, lower dissolved oxygen 
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(DO) levels, and decreased recruitment of gravel and large woody debris (LWD). More uniform flows year 

round have resulted in diminished natural channel formation, altered food web processes, and slower 

regeneration of riparian vegetation. These stable flow patterns have reduced bedload movement (Mount 1995, 

Associates 2001), caused spawning gravels to become embedded, and decreased channel widths due to channel 

incision, all of which has decreased the available spawning and rearing habitat below dams. The storage of 

unimpeded runoff in these large reservoirs also has altered the normal hydrograph for the Sacramento and San 

Joaquin River watersheds. Rather than seeing peak flows in these river systems following winter rain events 

(Sacramento River) or spring snow melt (San Joaquin River), the current hydrology has truncated peaks with a 

prolonged period of elevated flows (compared to historical levels) continuing into the summer dry season. 

Water withdrawals, for agricultural and municipal purposes, have reduced river flows and increased water 

temperatures during the critical summer months, and in some cases, have been of a sufficient magnitude to 

result in reverse flows in the lower San Joaquin River (Reynolds et al. 1993). Direct relationships exist between 

water temperature, water flow, and juvenile salmonid survival (Brandes and McLain 2001). Elevated water 

temperatures in the Sacramento River have limited the survival of young salmon in those waters. Juvenile fallrun survival in the Sacramento River is also directly related to June streamflow and June and July Delta outflow 

(Dettman et al. 1987). 

Water diversions for irrigated agriculture, municipal and industrial use, and managed wetlands are found 

throughout the Central Valley. Thousands of small and medium-size water diversions exist along the 

Sacramento River, San Joaquin River, and their tributaries. Although efforts have been made in recent years to 

screen some of these diversions, many remain unscreened. Depending on the size, location, and season of 

operation, these unscreened diversions entrain and kill many life stages of aquatic species, including juvenile 

salmonids. For example, as of 1997, 98.5 percent of the 3,356 diversions included in a Central Valley database 

were either unscreened or screened insufficiently to prevent fish entrainment (Herren and Kawasaki 2001). 

Most of the 370 water diversions operating in Suisun Marsh are unscreened (Herren and Kawasaki 2001). 

Outmigrant juvenile salmonids in the Delta have been subjected to adverse environmental conditions created by 

water export operations at the Central Valley Project (CVP) and State Water Project (SWP) facilities.

Specifically, juvenile salmonid survival has been reduced by: (1) water diversion from the mainstem 

Sacramento River into the Central Delta via the Delta Cross Channel (DCC); (2) upstream or reverse flows of 
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water in the lower San Joaquin River and southern Delta waterways; (3) entrainment at the Central Valley 

Project and the State Water Project (CVP & SWP) export facilities and associated problems at the Clifton Court 

Forebay; and (4) as discussed above, increased exposure to introduced, non-native predators such as striped 

bass (Morone saxatilis), largemouth bass (Micropterus salmoides), and sunfishes (Centrarchidae spp.) within 

the waterways of the Delta while moving through the Delta under the influence of CVP/SWP pumping. 

The Anderson-Cottonwood Irrigation District (ACID) Dam operates a diversion dam across the Sacramento 

River located 5 miles downstream from Keswick Dam. ACID is one of the 3 largest diversions on the 

Sacramento River and has senior water rights of 128 thousand acre feet (TAF) of water since 1916 for irrigation 

along the west side of the Sacramento River. The installation and removal of the diversion dam flashboards 

requires close coordination between the Bureau of Reclamation and ACID. The diversion dam is operated from 

April through October. Substantial reductions in Keswick releases to install or remove the flashboards have 

resulted in dewatered redds, stranded juveniles, and higher water temperatures. Based on generalized run 

timing (Table 14), the diversion dam operations could impact winter- and spring-run Chinook. Redd 

dewatering would most likely affect spring-run Chinook in October; however, the reductions in flows are 

usually short-term, lasting less than 8 hours. Such short-term reductions in flows may cause some mortality of 

incubating eggs and loss of stranded juveniles. Reductions in Keswick releases are limited to 15 percent in a 

24-hour period and 2.5 percent in any 1 hour. Past operations have shown that the most significant reductions 

occur during wet years when Shasta releases are higher than 10,000 cubic feet per second (cfs). Average April 

releases from Keswick are 6,000 to 7,000 cfs. The likelihood of a flow fluctuation occurring (when Shasta 

storage > 4.5 MAF in April) is 17 percent, or 14 out of the 82-year historical record. During wet years, flows 

released from Shasta Dam are typically higher than in drier water year types. The amount of flow that needs to 

be reduced to get to safe operating levels for the installation of the flashboards at the ACID dam is therefore 

greater and the wetted area reduction downstream of Keswick Dam is thus greater. The likelihood of an 

October reduction in flows that could dewater redds is even lower, since average releases are 6,000 cfs in all 

water year types. 

Table 14. Generalized life-history timing for listed salmonids in the upper Sacramento River.

Species Adult 

Immigration Adult Holding Typical Spawning Egg Incubation Juvenile Rearing Juvenile Emigration

Winter-run Nov-Jul Jan-May Apr-Jul Apr-Oct Jun-Mar Jun-Mar

Spring-run Mar-Sep May-Sep Aug-Oct Aug-Mar Year round Oct-May

Steelhead Oct-May Sep-Dec Dec-Apr Dec-May Year round Jan-Jul
105

The ACID diversion dam was improved in 2001 with the addition of new fish ladders and fish screens around 

the diversion. Since upstream passage was improved a substantial shift in winter-run spawning has occurred. In 

recent years, more than half of the winter-run redds have typically been observed above the ACID diversion 

dam. This makes flow fluctuations more a concern since such a large proportion of the run is spawning so close 

to Keswick Dam. 

The Red Bluff Diversion Dam (RBDD) is owned and operated by the Bureau of Reclamation. The TehamaColusa Canal Authority (TCCA) operates the Corning Canal and Tehama-Colusa Canal, which divert up to 328 

TAF from the Sacramento River. RBDD is located 59 miles downstream of Keswick Dam. It blocks or delays 

adult salmonid migrating upstream to various degrees, depending on run timing. Based on various studies 

(Vogel et al. 1988, Hallock 1989, CDFG 1998a), problems in salmonid passage at RBDD provide a welldocumented example of a diversion facility impairing salmon migration. A portion of the winter-run adults 

encounter the gates down and are forced to use the fish ladders. There are 3 fish ladders on RBDD, one on each 

side and one temporary ladder in the middle of the dam. The RBDD fish ladders are not efficient at passing 

adult salmonids due to the inability of salmon to find the entrances. Water released from RBDD flows through a 

small opening under 11 gates across the river, causing turbulent flows that confuse fish and keep them from 

finding the ladders. The fish ladders are not designed to allow enough water through them to attract adult 

salmonids towards them. Previous studies (Vogel 2008, USFWS 2001) have shown that salmon can be delayed 

up to 20 days in passing the dam. These delays can reduce the fitness of adults that expend their energy 

reserves fighting the flows beneath the gates, and increase the chance of prespawn mortality. Run timing is 

critical to salmon, as it is what distinguishes one race from another. Delays of a week or even days in passage 

likely prevents some spring-run adults (those that encounter gates down in May and June) from entering 

tributaries above RBDD that dry up or warm up in the spring (e.g., Cottonwood Creek, Cow Creek). These 

delays have the potential of preventing these fish from accessing summer holding pools in the upper areas of the 

creeks. Delays could allow for the onset of temperature related diseases that could affect spawning success or 

cause pre-spawn mortalities.

7.2.3. Water Conveyance and Flood Control

The development of the water conveyance system in the Delta has resulted in the construction of armored, riprapped levees on more than 1,100 miles of channels and diversions to increase channel elevations and flow 
106

capacity of the channels (Mount 1995). Levee development in the Central Valley affects spawning habitat, 

freshwater rearing habitat, freshwater migration corridors, and estuarine habitat PCEs. As Mount (1995)

indicates, there is an “underlying, fundamental conflict inherent in this channelization.” Natural rivers strive to 

achieve dynamic equilibrium to handle a watershed’s supply of discharge and sediment (Mount 1995). The 

construction of levees disrupts the natural processes of the river, resulting in a multitude of habitat-related 

effects, including isolation of the watershed’s natural floodplain behind the levee from the active river channel 

and its fluctuating hydrology. 

Many of these levees use angular rock (riprap) to armor the bank from erosive forces. The effects of 

channelization, and rip-rapping, include the alteration of river hydraulics and cover along the bank as a result of 

changes in bank configuration and structural features. These changes affect the quantity and quality of 

nearshore habitat for juvenile salmonids and have been thoroughly studied (USFWS 2000, Garland et al. 2002, 

Schmetterling et al. 2001). Simple slopes protected with rock revetment generally create nearshore hydraulic 

conditions characterized by greater depths and faster, more homogeneous water velocities than occur along 

natural banks. Higher water velocities typically inhibit deposition and retention of sediment and woody debris. 

These changes generally reduce the range of habitat conditions typically found along natural shorelines, 

especially by eliminating the shallow, slow-velocity river margins used by juvenile fish as refuge and to escape 

from fast currents, deep water, and predators (USFWS 2000). 

Prior to the 1970s, there was so much debris resulting from poor logging practices that many streams were 

completely clogged and were thought to have been total barriers to fish migration. As a result, in the 1960s and 

early 1970s it was common practice among fishery management agencies to remove woody debris thought to be 

a barrier to fish migration (NMFS 1996b). However, it is now recognized that too much LWD was removed 

from the streams resulting in a loss of salmonid habitat and it is thought that the large scale removal of woody 

debris prior to 1980 had major, long-term negative effects on rearing habitats for salmonids in northern 

California (NMFS 1996b). Areas that were subjected to this removal of LWD are still limited in the recovery of 

salmonid stocks; this limitation could be expected to persist for 50 to 100 years following removal of the debris. 

Large quantities of downed trees are a functionally important component of many streams (NMFS 1996b). 

LWD influences stream morphology by affecting channel pattern, position, and geometry, as well as pool 

formation (Keller and Swanson 1979, Bilby 1984, Robison and Beschta 1990). Reduction of wood in the 
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stream channel, either from past or present activities, generally reduces pool quantity and quality, alters stream 

shading which can affect water temperature regimes and nutrient input, and can eliminate critical stream habitat 

needed for both vertebrate and invertebrate populations. Removal of vegetation also can destabilize marginally 

stable slopes by increasing the subsurface water load, lowering root strength, and altering water flow patterns in 

the slope. 

In addition, the armoring and revetment of stream banks tends to narrow rivers, reducing the amount of habitat 

per unit channel length (Sweeney et al. 2004). As a result of river narrowing, benthic habitat decreases and the 

number of macroinvertebrates, such as stoneflies and mayflies, per unit channel length decreases, affecting 

salmonid food supply. 

7.2.4. Land Use Activities

Table 15 summarizes general land use categories in the CV by percent and provides population densities. Figure 

11 depicts general land use in California. In the pages that follow, Figure 11 is zoomed in to just the Central 

Valley to show the overlay of the CV spring-run Chinook ESU (Figure 12), the Sacramento River winter-run 

Chinook ESU (

Figure 13), and CV steelhead (Figure 14). Land use activities continue to have large impacts on salmonid 

habitat in the Central Valley watershed. Until about 150 years ago, the Sacramento River was bordered by up 

to 500,000 acres of riparian forest, with bands of vegetation extending outward for 4 or 5 miles (CRA 1989). 

Starting with the gold rush, these vast riparian forests were cleared for building materials, fuel, and to clear land 

for farms on the raised natural levee banks. The degradation and fragmentation of riparian habitat continued 

with extensive flood control and bank protection projects, together with the conversion of the fertile riparian 

lands to agriculture outside of the natural levee belt. The dominant land use in the CV is agriculture. By 1979, 

riparian habitat along the Sacramento River diminished to 11,000 to 12,000 acres, or about 2 percent of historic 

levels (McGill 1987). The clearing of the riparian forests removed a vital source of snags and driftwood in the 

Sacramento and San Joaquin River basins. This has reduced the volume of LWD input needed to form and 

maintain stream habitat that salmon depend on in their various life stages. In addition to this loss of LWD 

sources, removal of snags and obstructions from the active river channel for navigational safety has further 

reduced the presence of LWD in the Sacramento and San Joaquin Rivers, as well as the Delta.

Table 15. Land uses and population density in the Sacramento and San Joaquin watersheds (Carter and Resh 

2005).

Watershed Land Use Categories (Percent) Density 

(people/mi2

Agriculture Forest Urban Other )
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Sacramento River 15 49 2 30 grass & shrub 61

San Joaquin River 30 27 2 36 grass & shrub 76

Increased sedimentation resulting from agricultural and urban practices within the Central Valley is one of the 

primary causes of salmonid habitat degradation (NMFS 1996a). Sedimentation can adversely affect salmonids 

during all freshwater life stages by: clogging or abrading gill surfaces, adhering to eggs, hampering fry 

emergence (Phillips and Campbell 1961), burying eggs or alevins, scouring and filling in pools and riffles, 

reducing primary productivity and photosynthesis activity (Cordone and Kelley 1961), and affecting intergravel 

permeability and DO levels. Excessive sedimentation over time can cause substrates to become embedded, 

which reduces successful salmonid spawning and egg and fry survival (Waters 1995). 

Land use categories are presented in Table 16 below. Land use activities associated with road construction, 

urban development, logging, mining, agriculture, and recreation have significantly altered fish habitat quantity 

and quality through the alteration of streambank and channel morphology; alteration of ambient water 

temperatures; degradation of water quality; elimination of spawning and rearing habitat; fragmentation of 

available habitats; elimination of downstream recruitment of LWD; and removal of riparian vegetation, 

resulting in increased streambank erosion (Meehan 1991). Urban stormwater and agricultural runoff are 

contaminated with pesticides, petroleum products, sediment, etc. Agricultural practices in the Central Valley 

have eliminated large trees and logs and other woody debris that would otherwise be recruited into the stream 

channel (NMFS 1998). Rice cultivation accounts for approximately half of all cultivated crops in the 

Sacramento Valley, and a quarter of all cultivated crops after combining the San Joaquin Valley with the 

Sacramento (Table 16).
109

Figure 11. Landuse in California (National Land Cover Database 2006).
110

Figure 12. CV spring-run Chinook ESU with associated land-use patterns.
111

Figure 13. Sacramento River winter-run Chinook ESU with associated land use patterns.

.
112

Figure 14. CV steelhead DPS with associated land use patterns.
113

Since the 1850s, wetlands reclamation for urban and agricultural development has caused 

the cumulative loss of 79 and 94 percent of the tidal marsh habitat in the Delta 

downstream and upstream of Chipps Island, respectively (Conomos et al. 1985, Nichols 

et al. 1986, Wright and Phillips 1988, Monroe et al. 1992, Project 1999). Prior to 1850, 

approximately 1400 km2 of freshwater marsh surrounded the confluence of the 

Sacramento and San Joaquin Rivers, and another 800 km2 of saltwater marsh fringed San 

Francisco Bay’s margins. Of the original 2,200 km2 of tidally influenced marsh, only 

about 125 km2 of undiked marsh remains today. In Suisun Marsh, saltwater intrusion 

and land subsidence gradually has led to the decline of agricultural production. 

Presently, Suisun Marsh consists largely of tidal sloughs and managed wetlands for duck 

clubs, which first were established in the 1870s in western Suisun Marsh (Goals Project 

1999). Even more extensive losses of wetland marshes occurred in the Sacramento and 

San Joaquin River Basins. Little of the extensive tracts of wetland marshes that existed 

prior to 1850 along the valley’s river systems and within the natural flood basins exist 

today. Most has been “reclaimed” for agricultural purposes, leaving only small remnant 

patches. 
114

Table 16. Area of land use categories within the range of listed salmonids in the Central 

Valley (km2). Total area for each category is given in bold. Land cover was determined 

via the National Land Cover Database 2006, developed by the Multi-Resolution Land 

Characteristics (MRLC) Consortium, comprised of nine federal agencies (USGS, EPA, 

USFS, NOAA, NASA, BLM, NPS, NRCS, and USFWS). Land cover class definitions are 

available at: http://www.mrlc.gov/nlcd_definitions.php . Rice information was obtained 

from the California Rice Commission (2010 values) and from the Cal. DPR pesticide use 

reporting data base (2010 county reports).

Land Cover Chinook Salmon Steelhead

Sub Category code Central Valley Spring-run 

Sacramento Winter-run

Central Valley

Water

Open Water 11

Perennial Snow/ice 12

Developed Land

Open Space 21

Low Intensity 22

Medium Intensity 23

High Intensity 24

Barren Land 31

Undeveloped Land

Deciduous Forest 41

Evergreen Forest 42

Mixed Forest 43

Shrub/Scrub 52

Herbaceous 71

Woody Wetlands 90

Emergent Wetlands 95

Agriculture

Hay/Pasture 81

Cultivated Crops 82

(Rice)

Total (including open water)

Total (without open water) 

367

 346

 0

2,755

1,174

 635

 616

 153

 178

15,063

 657

3,707

 476

3,245

6,261

 189

 527

5,796

 754

5,043

(2,344)

23,982

23,615

422

422

 0

3,534

1,472

 792

 837

 211

 222

19,138

 744

3,942

 583

3,786

9,396

 245

 431

10,507

1,640

8,867

(2,353)

33,601

33,179
115

Dredging of river channels to enhance inland maritime trade and to provide raw material 

for levee construction has significantly and detrimentally altered the natural hydrology 

and function of the river systems in the Central Valley. Starting in the mid-1800s, the 

Corps and private consortiums began straightening river channels and artificially 

deepening them to enhance shipping commerce. This has led to declines in the natural 

meandering of river channels and the formation of pool and riffle segments. The 

deepening of channels beyond their natural depth also has led to a significant alteration in 

the transport of bedload in the riverine system as well as the local flow velocity in the 

channel (Mount 1995). The Sacramento Flood Control Project at the turn of the 

nineteenth century ushered in the start of large scale Corps actions in the Delta and along 

the rivers of California for reclamation and flood control. The creation of levees and the 

deep shipping channels reduced the natural tendency of the San Joaquin and Sacramento 

Rivers to create floodplains along their banks with seasonal inundations during the wet 

winter season and the spring snow melt periods. These annual inundations provided 

necessary habitat for rearing and foraging of juvenile native fish that evolved with this 

flooding process. The armored rip-rapped levee banks and active maintenance actions of 

Reclamation Districts precluded the establishment of ecologically important riparian 

vegetation, introduction of valuable LWD from these riparian corridors, and the 

productive intertidal mudflats characteristic of the undisturbed Delta habitat. 

Urban stormwater and agricultural runoff may be contaminated with pesticides, oil, 

grease, heavy metals, polycyclic aromatic hydrocarbons (PAHs), and other organics and 

nutrients (CRWQCB-CVR 1998) that can destroy aquatic life necessary for salmonid 

survival (NMFS 1996a, NMFS 1996b). Point source (PS) and non-point source (NPS) 

pollution occurs at almost every point that urbanization activity influences the watershed. 

Impervious surfaces (i.e., concrete, asphalt, and buildings) reduce water infiltration and 

increase runoff, thus creating greater flood hazard (NMFS 1996a, NMFS 1996b). Flood 

control and land drainage schemes may increase the flood risk downstream by 

concentrating runoff. A flashy discharge pattern results in increased bank erosion with 

subsequent loss of riparian vegetation, undercut banks and stream channel widening. In 

addition to the PS and NPS inputs from urban runoff, juvenile salmonids are exposed to 
116

increased water temperatures as a result of thermal inputs from municipal, industrial, and 

agricultural discharges. 

Past mining activities routinely resulted in the removal of spawning gravels from streams, 

the straightening and channelization of the stream corridor from dredging activities, and 

the leaching of toxic effluents into streams from mining operations. Many of the effects 

of past mining operations continue to impact salmonid habitat today. Current mining 

practices include suction dredging (sand and gravel mining), placer mining, lode mining 

and gravel mining. Present day mining practices are typically less intrusive than historic 

operations (hydraulic mining); however, adverse impacts to salmonid habitat still occur 

as a result of present-day mining activities. Sand and gravel are used for a large variety 

of construction activities including base material and asphalt, road bedding, drain rock for 

leach fields, and aggregate mix for concrete to construct buildings and highways. 

Most aggregate is derived principally from pits in active floodplains, pits in inactive river 

terrace deposits, or directly from the active channel. Other sources include hard rock 

quarries and mining from deposits within reservoirs. Extraction sites located along or in 

active floodplains present particular problems for anadromous salmonids. Physical 

alteration of the stream channel may result in the destruction of existing riparian 

vegetation and the reduction of available area for seedling establishment. Loss of 

vegetation impacts riparian and aquatic habitat by causing a loss of the temperature 

moderating effects of shade and cover, and habitat diversity. Extensive degradation may 

induce a decline in the alluvial water table, as the banks are effectively drained to a 

lowered level, affecting riparian vegetation and water supply (NMFS 1996b). Altering 

the natural channel configuration will reduce salmonid habitat diversity by creating a 

wide, shallow channel lacking in the pools and cover necessary for all life stages of 

anadromous salmonids. In addition, waste products resulting from past and present 

mining activities, include cyanide (an agent used to extract gold from ore), copper, zinc, 

cadmium, mercury, asbestos, nickel, chromium, and lead. 

Juvenile salmonids are exposed to increased water temperatures in the Delta during the 

late spring and summer due to the loss of riparian shading, and by thermal inputs from 
117

municipal, industrial, and agricultural discharges. Studies by DWR on water quality in 

the Delta over the last 30 years show a steady decline in the food sources available for 

juvenile salmonids and an increase in the clarity of the water due to a reduction in 

phytoplankton and zooplankton. These conditions have contributed to increased 

mortality of juvenile Chinook salmon and steelhead as they move through the Delta. 

The following are excerpts from Lindley et al. (2009): 

“The long-standing and ongoing degradation of freshwater and estuarine habitats and 

the subsequent heavy reliance on hatchery production were also likely contributors to 

the collapse of the [fall-run] stock. Degradation and simplification of freshwater and 

estuary habitats over a century and a half of development have changed the Central 

Valley Chinook salmon complex from a highly diverse collection of numerous wild 

populations to one dominated by fall Chinook salmon from four large hatcheries.” 

“In conclusion, the development of the Sacramento-San Joaquin watershed has 

greatly simplified and truncated the once-diverse habitats that historically supported a 

highly diverse assemblage of populations. The life history diversity of this historical 

assemblage would have buffered the overall abundance of Chinook salmon in the 

Central Valley under varying climate conditions.” 

7.2.5. Water Quality

The water quality of the Central Valley has been negatively impacted over the last 150 

years. Increased water temperatures, decreased DO levels, and increased turbidity and 

contaminant loads have degraded the quality of the aquatic habitat for the rearing and 

migration of salmonids. Some common pollutants include effluent from wastewater 

treatment plants and chemical discharges such as dioxin from San Francisco bay 

petroleum refineries (McEwan and Jackson 1996b). In addition, agricultural drain water, 

another possible source of contaminants, can contribute up to 30 percent of the total 

inflow into the Sacramento River during the low-flow period of a dry year (NMFS 

2009c). 
118

7.2.6. Pesticides

The Regional Board, in its 1998 Clean Water Act §303(d) list characterized the Delta as 

an impaired waterbody having elevated levels of chlorpyrifos, dichlorodiphenyltrichlor 

(i.e. DDT), diazinon, electrical conductivity, Group A pesticides [aldrin, dieldrin, 

chlordane, endrin, heptachlor, heptachlor epoxide, hexachlorocyclohexanes (including 

lindane), endosulfan and toxaphene], mercury, low DO, organic enrichment, and 

unknown toxicities (CRWQCB-CVR 2010). Figure 16 shows numerous reaches where 

streams and rivers within the Central Valley are listed as impaired due to pesticides.

In general, water degradation or contamination can lead to either acute toxicity, resulting 

in death when concentrations are sufficiently elevated, or more typically, when 

concentrations are lower, to chronic or sublethal effects that reduce the physical health of 

the organism, and lessens its survival over an extended period of time. Mortality may 

become a secondary effect due to compromised physiology or behavioral changes that 

lessen the organism's ability to carry out its normal activities. For example, increased 

levels of heavy metals are detrimental to the health of an organism because they interfere 

with metabolic functions by inhibiting key enzyme activity in metabolic pathways, 

decrease neurological function, degrade cardiovascular output, and act as mutagens, 

teratogens or carcinogens in exposed organisms (Rand et al. 1995, Goyer 1996). For 

listed species, these effects may occur directly to the listed fish or to its prey base, which 

reduces the forage base available to the listed species. In the aquatic environment, most 

anthropogenic chemicals and waste materials, including toxic organic and inorganic 

chemicals eventually accumulate in sediment (Ingersoll 1995). Direct exposure to 

contaminated sediments may cause deleterious effects to listed salmonids. This may 

occur if a fish swims through a plume of the resuspended sediments or rests on 

contaminated substrate and absorbs the toxic compounds through one of several routes: 

dermal contact, ingestion, or uptake across the gills. Elevated contaminant levels may be 

found in localized “hot spots” where discharge occurs or where river currents deposit 

sediment loads. Sediment contaminant levels can thus be significantly higher than the 

overlying water column concentrations (EPA 1994). However, the more likely route of 

exposure to salmonids is through the food chain, when the fish feed on organisms that are 
119

contaminated with toxic compounds. Prey species become contaminated either by 

feeding on the detritus associated with the sediments or dwelling in the sediment itself. 

Therefore, the degree of exposure to the salmonids depends on their trophic level and the 

amount of contaminated forage base they consume. Response of salmonids to 

contaminated sediments is similar to water borne exposures once the contaminant has 

entered the body of the fish (Heath 1995).

7.2.6.1. National Pollutant Discharge Elimination System

Pollution originating from a discrete location such as a pipe discharge or wastewater 

treatment outfall is known as a point source. Point sources of pollution require a National 

Pollutant Discharge Elimination System (NPDES) permit. These permits are issued for, 

among other things, aquaculture, concentrated animal feeding operations, industrial 

wastewater treatment plants, biosolids (sewer/sludge), pre-treatment and stormwater 

overflows. The EPA administers the NPDES permit program and states certify that 

NPDES permit holders comply with state water quality standards. Nonpoint source 

discharges do not originate from discrete points; thus, nonpoint sources are difficult to 

identify and quantify. Examples of nonpoint source pollution include, but are not limited 

to, urban runoff from impervious surfaces, areas of fertilizer and pesticide application, 

sedimentation, and manure. 

According to EPA’s database of NPDES permits, about 243 NPDES individual permits 

are co-located with listed Pacific salmonids in California. See Figure 15 for NPDES 

permits located within listed salmonid ESUs/DPSs in California.

On November 27, 2006, EPA issued a final rule which exempted pesticides from the 

NPDES permit process, provided that application was approved under FIFRA. On 

January 7, 2009, the Sixth Circuit Court of Appeals vacated this rule (National Cotton 

Council v. EPA, 553 F.3d 927 (6th Cir. 2009)). The result of the vacature, according to 

the Sixth Circuit, is that “discharges of pesticide pollutants are subject to the NPDES 

permitting program” under the CWA. In response, EPA has developed a Pesticide 

General Permit through the NPDES permitting program to regulate such discharges. The 
120

permit has recently gone through Section 7(a)(2) consultation (NMFS 2011b) and was 

issued on October 31, 2011. The Pesticides General Permit generally authorizes 

discharge of pesticides into waterways, but provides for additional review of discharges 

of pesticides to waters containing NMFS listed resources.

7.2.7. Baseline Water Temperature - Clean Water Act

Elevated temperature is considered a pollutant in most states with approved Water 

Quality Standards under the federal Clean Water Act (CWA) of 1972. Under the 

authority of the CWA, states periodically prepare a list of all surface waters in the state 

for which beneficial uses - such as drinking, recreation, aquatic habitat, and industrial use 

– are impaired by pollutants. This process is in accordance with section 303(d) of the 

CWA. Estuaries, lakes, and streams listed under 303(d) are those that are considered 

impaired or threatened by pollution. They are water quality limited, do not meet state 

surface water quality standards, and are not expected to improve within the next two 

years. 

Each state has separate and different 303(d) listing criteria and processes. Generally a 

water body is listed separately for each standard it exceeds, so it may appear on the list 

more than once. If a water body is not on the 303(d) list, it is not necessarily 

contaminant-free; rather it may not have been tested. Therefore, the 303(d) list is a 

minimum list for the each state regarding polluted water bodies by parameter (Figure 16).

After states develop their lists of impaired waters, they are required to prioritize and 

submit their lists to EPA for review and approval. Each state establishes a priority 

ranking for such waters, considering the severity of the pollution and the uses to be made 

of such waters. States are expected to identify high priority waters targeted for Total 

Maximum Daily Load (TMDL) development for specific pollutants within two years of 

the 303(d) listing process. TMDLs once developed are considered during permitting 

processes.
121

Figure 15. California dams and NPDES permit sites.
122

Temperature is significant for the health of aquatic life. Water temperatures affect the 

distribution, health, and survival of native cold-blooded salmonids in the Pacific 

Northwest. These fish will experience adverse health effects when exposed to 

temperatures outside their optimal range. For listed Pacific salmonids, water temperature 

Figure 16. California 303(d) list: water bodies and stream segments included in the 2010 

Integrated Report.
123

tolerance varies between species and life stages. Optimal temperatures for rearing 

salmonids range from 10oC to 16oC. In general, the increased exposure to stressful water 

temperatures and the reduction of suitable habitat caused by drought conditions reduce 

the abundance of salmon. Warm temperatures can reduce fecundity, reduce egg survival, 

retard growth of fry and smolts, reduce rearing densities, increase susceptibility to 

disease, decrease the ability of young salmon and trout to compete with other species for 

food, and to avoid predation (Spence et al. 1996b, McCullough 1999). Migrating adult 

salmonids and upstream migration can be delayed by excessively warm stream 

temperatures. Excessive stream temperatures may also negatively affect incubating and 

rearing salmonids (Gregory and Bisson 1997). 

Sublethal temperatures (above 24oC) could be detrimental to salmon by increasing 

susceptibility to disease (Colgrove and Wood 1966) or elevating metabolic demand (Brett 

1995). Substantial research demonstrates that many fish diseases become more virulent 

at temperatures over 15.6oC (McCullough 1999). Due to the sensitivity of salmonids to 

temperature, states have established lower temperature thresholds for salmonid habitat as 

part of their water quality standards. A water body is listed for temperature on the 303(d) 

list if the 7-day average of the daily maximum temperatures 

(7-DADMax) exceeds the temperature threshold.

Water bodies that are not designated salmonid habitat are also listed if they have a oneday maximum over a given background temperature. Using publicly available 

Geographic Information System (GIS) layers, we determined the number of km on the 

303(d) list for exceeding temperature thresholds within the boundaries of each ESU/DPS 

(Table 17). Because the 303(d) list is limited to the subset of rivers tested, the chart 

values should be regarded as lower-end estimates. 
124

Table 17. Kilometers of river, stream, and estuary included in state 303(d) lists due to 

temperature. Data from most recent GIS layers available from CA 2010 

Species ESU or DPS 303(d) listed 

water (Km)

Chinook

Sacramento 

River Winter-run

29.9

Central Valley 

Spring-run

29.9

Steelhead Central Valley 367.8

CA 2010 (California EPA TMDL Program 2011)

While some ESU/DPS ranges do not contain any 303(d) rivers listed for temperature, 

others show considerable overlap. These comparisons demonstrate the relative 

significance of elevated temperature among ESUs/DPSs. Increased water temperature 

may result from wastewater discharge, decreased water flow, minimal shading by 

riparian areas, and climatic variation.

7.2.8. Hatchery Operations and Practices

Five hatcheries currently produce Chinook salmon in the Central Valley, and four of 

these also produce steelhead. Releasing large numbers of hatchery fish can pose a threat 

to wild Chinook salmon and steelhead stocks through genetic impacts, competition for 

food and other resources between hatchery and wild fish, predation of hatchery fish on 

wild fish, and increased fishing pressure on wild stocks as a result of hatchery production 

(Waples 1991). The genetic impacts of artificial propagation programs in the Central 

Valley are primarily caused by straying of hatchery fish and the subsequent interbreeding 

of hatchery fish with wild fish. In the Central Valley, practices such as transferring eggs 

between hatcheries and trucking smolts to distant sites for release contribute to elevated 

straying levels (DOI 1999). For example, Nimbus Hatchery on the American River rears 

Eel River steelhead stock and releases these fish in the Sacramento River basin. One of 

the recommendations in the Joint Hatchery Review Report (NMFS/CDFG 2001) was to 
125

identify and designate new sources of steelhead brood stock to replace the current Eel 

River origin brood stock. 

Hatchery practices as well as spatial and temporal overlaps of habitat use and spawning 

activity between spring- and fall-run fish have led to the hybridization and 

homogenization of some subpopulations (CDFG 1998a). As early as the 1960s, Slater 

(1963a) observed that spring-run and early fall-run were competing for spawning sites in 

the Sacramento River below Keswick Dam, and speculated that the two runs may have 

hybridized. Spring-run from the FRFH have been documented as straying throughout the 

Central Valley for many years (CDFG 1998a), and in many cases have been recovered 

from the spawning grounds of fall-run, an indication that FRFH spring-run may exhibit 

fall-run life history characteristics. Although the degree of hybridization has not been 

comprehensively determined, it is clear that the populations of spring-run spawning in the 

Feather River and counted at RBDD contain hybridized fish. 

The management of hatcheries, such as Nimbus Fish Hatchery and FRFH, can directly 

impact spring-run and steelhead populations by over-saturating the natural carrying 

capacity of the limited habitat available below dams. In the case of the Feather River, 

significant redd superimposition occurs in-river due to hatchery overproduction and the 

inability to physically separate spring-run and fall-run adults. This concurrent spawning 

has led to hybridization between the spring-run and fall-run in the Feather River. At 

Nimbus Hatchery, operating Folsom Dam to meet temperature requirements for returning 

hatchery fall-run often limits the amount of water available for steelhead spawning and 

rearing the rest of the year. 

The increase in Central Valley hatchery production has reversed the composition of the 

steelhead population, from 88 percent naturally-produced fish in the 1950s (McEwan 

2001a) to an estimated 23 to 37 percent naturally-produced fish currently (Nobriga and 

Cadrett 2003). The increase in hatchery steelhead production proportionate to the wild 

population has reduced the viability of the wild steelhead populations, increased the use 

of out-of-basin stocks for hatchery production, and increased straying (NMFS/CDFG 
126

2001). Thus, the ability of natural populations to successfully reproduce and continue

their genetic integrity likely has been diminished. 

The relatively low number of spawners needed to sustain a hatchery population can result 

in high harvest-to-escapements ratios in waters where fishing regulations are set 

according to hatchery population. This can lead to over-exploitation and reduction in the 

size of wild populations existing in the same system as hatchery populations due to 

incidental bycatch (McEwan 2001a). 

Hatcheries also can have some positive effects on salmonid populations. Winter-run 

produced in the LSNFH are considered part of the winter-run ESU. Spring-run produced 

in the FRFH are considered part of the spring-run ESU. Artificial propagation has been 

shown to be effective in bolstering the numbers of naturally spawning fish in the short 

term under specific scenarios. Artificial propagation programs can also aid in conserving 

genetic resources and guarding against catastrophic loss of naturally spawned populations 

at critically low abundance levels, as was the case with the winter-run population during 

the 1990s. However, relative abundance is only one component of a viable salmonid 

population. 

7.3. Baseline Habitat Condition

As noted above in the discussion on land use, and in the Status of the Species section, the 

riparian zones for much of the ESUs/DPSs are degraded. Riparian zones are the areas of 

land adjacent to rivers and streams. These systems serve as the interface between the 

aquatic and terrestrial environments. Riparian vegetation is characterized by emergent 

aquatic plants and species that thrive on close proximity to water, such as willows. This 

vegetation maintains a healthy river system by reducing erosion, stabilizing main 

channels, and providing shade. Leaf litter that enters the river becomes an important 

source of nutrients for invertebrates (Bisson and Bilby 2001). Riparian zones are also the 

major source of large woody debris (LWD). When trees fall and enter the water, they 

become an important part of the ecosystem. The LWD alters the flow, creating the pools 

of slower moving water preferred by salmon (Bilby et al. 2001). While not necessary for 
127

pool formation, LWD is associated with around 80% of pools in northern California, 

Washington, and the Idaho pan-handle (Bilby and Bisson 2001). 

Bilby and Bisson (2001) discuss several studies that associate increased LWD with 

increased pools, and both pools and LWD with salmonid productivity. Their review also 

includes documented decreases in salmonid productivity following the removal of LWD. 

Other benefits of LWD include deeper pools, increased sediment retention, and channel 

stabilization. 

Floodplains are relatively flat areas adjacent to larger streams and rivers (e.g., the Yolo 

Bypass). They allow for the lateral movement of the main channel and provide storage 

for floodwaters during periods of high flow. Water stored in the floodplain (i.e., 

hyporheic flow) is later released during periods of low flow. This process ensures 

adequate flows for salmonids during the summer months, and reduces the possibility of 

high-energy flood events destroying salmonid redds (Smith 2005).

Periodic flooding of these areas creates habitat used by salmonids. Thus, floodplain areas 

vary in depth and widths and may be intermittent or seasonal. Storms also wash sediment 

and LWD into the main stem river, often resulting in blockages. These blockages may 

force the water to take an alternate path and result in the formation of side channels and 

sloughs (Benda et al. 2001). Side channels and sloughs are important spawning and 

rearing habitat for salmonids. The degree to which these off-channel habitats are linked 

to the main channel via surface water connections is referred to as connectivity (PNERC 

2002). As river height increases with heavier flows, more side channels form and 

connectivity increases. Juvenile salmonids migrate to and rear in these channels for a 

certain period of time before swimming out to the open sea.

Healthy riparian habitat and floodplain connectivity are vital for supporting a salmonid 

population. Chinook salmon and steelhead have life history strategies that rely on 

floodplains during their juvenile life stages. Sommer et al. (2001) was able to show that 

the Yolo Bypass, floodplain to the Sacramento River, provides better rearing and 
128

migration habitat for juvenile Chinook than adjacent river channels. During a 1998 and 

1999 study, salmon increased in size substantially faster in the seasonally inundated 

floodplain than in the river.

Once floodplain areas have been disturbed, it can take decades for their recovery (Smith 

2005). Consequently, most land use practices cause some degree of impairment. 

Development leads to construction of levees and dikes, which isolate the main stem river 

from the floodplain. Agricultural development and grazing in riparian areas also 

significantly change the landscape. Riparian areas managed for logging, or logged in the 

past, are often impaired by a change in species composition. Most areas in the northwest 

were historically dominated by conifers. Logging results in recruitment of deciduous 

trees, decreasing the quality of LWD in the rivers. Deciduous trees have smaller 

diameters than conifers; they decompose faster and are more likely to be displaced 

(Smith 2005). 

Without a properly functioning riparian zone, salmonids contend with a number of 

limiting factors. They face reductions in quantity and quality of both off-channel and 

pool habitats. Also, when seasonal flows are not moderated, both higher and lower flow 

conditions exist. Higher flows can displace fish and destroy redds, while lower flows cut 

off access to parts of their habitat. Finally, decreased vegetation limits the available 

shade and cover, exposing individuals to higher temperatures and increased predation.

7.4. Baseline Pesticide Detections in Aquatic Environments

In the environmental baseline, we address pesticide detections reported as part of the U.S. 

Geological Survey (USGS) National Water-Quality Assessment Program’s (NAWQA) 

national assessment (Gilliom et al. 2006b). We chose this approach because the 

NAWQA studies present the same level of analysis for each area. Further, given the lack 

of uniform reporting standards, we are unable to present a comprehensive basin-specific 

analysis of detections from other sources. 

According to Gilliom et al. (2006b), the distributions of the most prevalent pesticides in 
129

streams and ground water correlate with land use patterns and associated present or past 

pesticide use. When pesticides are released into the environment, they frequently end up 

as contaminants in aquatic environments. Depending on their physical properties some 

are rapidly transformed via chemical, photochemical, and biologically mediated reactions 

into other compounds, known as degradates. These degradates may become as prevalent 

as the parent pesticides depending on their rate of formation and their relative persistence.

In the Exposure section of the Effects of the Proposed Action we present a more 

comprehensive discussion of available monitoring data from the NAWQA program, state 

databases maintained by California, and other targeted monitoring studies. 

7.4.1.1. National Water-Quality Assessment Program 

From 1992 - 2001, the USGS sampled water from 186 stream sites within 51 study units; 

bed-sediment samples from 1,052 stream sites, and fish from 700 stream sites across the 

continental U.S. Concentrations of pesticides were detected in streams and groundwater 

within most areas sampled with substantial agricultural or urban land uses. NAWQA 

results further detected at least one pesticide or degradate more than 90% of the time in 

water, in more than 80% in fish samples, and greater than 50% of bed-sediment samples 

from streams in watersheds with agricultural, urban, and mixed land use (Gilliom et al. 

2006b).

Twenty-four pesticides and one degradate were each detected in over 10% of streams in 

agricultural, urban, or mixed land use areas. These 25 compounds include 11 agricultureuse herbicides and the atrazine degradate deethylatrazine; 7 urban-use herbicides; and 6 

insecticides used in both agricultural and urban areas. Two of the herbicides used 

primarily in urban areas are 2,4-D and diuron. Both herbicides were detected roughly 

12% of the time in agricultural streams and between 20% and 25% of the time in urban 

streams. In a previous Opinion, NMFS assessed the effects of these two herbicides on 

salmonids (http://www.nmfs.noaa.gov/pr/pdfs/consultations/pesticide_opinion4.pdf). 

Five of the insecticides were carbaryl, carbofuran, chlorpyrifos, diazinon, and malathion. 

NMFS assessed the effects of these five insecticides on listed salmonids in its 2008 and 
130

2009 Opinions (NMFS 2008c, NMFS 2009e). 

Another dimension of pesticides and their degradates in the aquatic environment is their 

simultaneous occurrence as mixtures (Gilliom et al. 2006b). Mixtures result from the use 

of different pesticides for multiple purposes within a watershed or groundwater recharge 

area. Pesticides generally occur more often in natural waterbodies as mixtures than as 

individual compounds. Mixtures of pesticides were detected more often in streams than 

in ground water and at relatively similar frequencies in streams draining areas of 

agricultural, urban, and mixed land use. More than 90% of the time, water from streams 

in these developed land use settings had detections of two or more pesticides or 

degradates. About 70% and 20% of the time, streams had five or more and ten or more 

pesticides or degradates, respectively (Gilliom et al. 2006b). Fish exposed to multiple 

pesticides at once may also experience additive and synergistic effects. If the effects on a 

biological endpoint from concurrent exposure to multiple pesticides can be predicted by 

adding the potency of the pesticides involved, the effects are said to be additive. If, 

however, the response to a mixture leads to a greater than expected effect on the 

endpoint, and the pesticides within the mixture enhance the toxicity of one another, the 

effects are characterized as synergistic. These effects are of particular concern when the 

pesticides share a mode of action. NAWQA analysis of all detections indicates that more 

than 6,000 unique mixtures of 5 pesticides were detected in agricultural streams (Gilliom 

et al. 2006b). The number of unique mixtures varied with land use. 

More than half of all agricultural streams sampled and more than three-quarters of all 

urban streams had concentrations of pesticides in water that exceeded one or more 

benchmarks for aquatic life. Aquatic life criteria are EPA water-quality guidelines for 

protection of aquatic life. Exceedance of an aquatic life benchmark level indicates a 

strong probability that aquatic species are being adversely affected. However, aquatic 

species may also be affected at levels below criteria. In agricultural streams, most 

concentrations that exceeded an aquatic life benchmark involved chlorpyrifos (21%), 

azinphos methyl (19%), atrazine (18%), p,p’-DDE (16%), and alachlor (15%) (Gilliom et 

al. 2006b). Finally, organochlorine pesticides that were discontinued 15 to 30 years ago 
131

still exceeded benchmarks for aquatic life and fish-eating wildlife in bed sediment or fish 

tissue samples from many streams. 

7.5. Baseline Pesticide Consultations

NMFS has consulted with EPA on the registration of several pesticides (Table 18). 

NMFS (NMFS 2008a) determined that current use of chlorpyrifos, diazinon, and 

malathion is likely to jeopardize the continued existence of 27 listed salmonid 

ESUs/DPSs – including CV spring-run Chinook, Sacramento winter-run Chinook, and 

CCV steelhead (CV listed species). NMFS (NMFS 2009b) further determined that 

current use of carbaryl and carbofuran is likely to jeopardize the continued existence of 

22 ESUs/DPSs – including the CV listed species; and the current use of methomyl is 

likely to jeopardize the continued existence of 18 ESUs/DPSs of listed salmonids –

including the CV listed species. NMFS also published conclusions regarding the 

registration of 12 different a.i.s (NMFS 2010b). NMFS concluded that pesticide products 

containing azinphos methyl, disulfoton, fenamiphos, methamidophos, or methyl 

parathion are not likely to jeopardize the continuing existence of any listed Pacific 

Salmon or destroy or adversely modify designated critical habitat. NMFS also concluded 

that the effects of products containing bensulide, dimethoate, ethoprop, methidathion, 

naled, phorate, or phosmet are likely to jeopardize the continued existence of some listed 

Pacific Salmonids and to destroy or adversely modify designated habitat of some listed 

salmonids – including the CV listed species. In June 2011, NMFS issued a biological 

opinion on the effects of four herbicides and two fungicides (Table 18). NMFS 

concluded that products containing 2,4-D are likely to jeopardize the existence of all 

listed salmonids, and adversely modify or destroy the critical habitat of some ESU / 

DPSs. Products containing chlorothalonil or diuron were also likely to adversely modify 

or destroy critical habitat – including CV listed species but not likely to jeopardize listed 

salmonids. NMFS also concluded that products containing captan, linuron, or triclopyr 

BEE do not jeopardize the continued existence of any ESUs/DPSs of listed Pacific 

salmonids or adversely modify designated critical habitat (NMFS 2011) (Table 18). In 

May 2012, NMFS issued a biological opinion on the effects of three herbicides, oryzalin, 

pendimethalin, and trifluralin concluding registration of all three herbicides are likely to 
132

jeopardize listed species in the Central Valley and destroy or adversely modify their 

designated critical habitat (Table 18).

Table 18. Consultations with EPA's proposed FIFRA actions. Past Opinion's jeopardy (J) 

and adverse modification (AM) determinations pertaining to California Central Valley ESA 

listed Pacific salmonids.

Opinion 

number

Pesticide

ESU or DPS

CV Spring-run Chinook Sac R. Winter-run Chinook CCV Steelhead

1

Chlorpyrifos J, AM J, AM J, AM

Diazinon J, AM J, AM J, AM

Malathion J, AM J, AM J, AM

2

Carbaryl J, AM J, AM J, AM

Carbofuran J, AM J, AM J, AM

Methomyl J, AM J, AM J, AM

3

Dimethoate J, AM J, AM J, AM

Naled J, AM J, AM J, AM

Phosmet J, AM J, AM J, AM

Ethroprop No J, No AM J, AM No J, No AM

Phorate J, AM J, AM J, AM

Methidathion J, AM J, AM J, AM

Bensulide J, AM J, AM No J, No AM

4

2,4-D J, AM J, AM J, AM

Triclopyr BEE No J, No AM No J, No AM No J, No AM

Diuron AM, No J AM, No J AM, No J

Linuron No J, No AM No J, No AM No J, No AM

Captan No J, No AM No J, No AM No J, No AM

Chlorthalonil AM AM AM

5

Oryzalin J, AM J, AM J, AM

Pendimethalin J, AM J, AM J, AM

Trifluralin J, AM J, AM J, AM
133

7.6. Rice Production and the Use of Thiobencarb in California

Rice growing is of interest because of its overlap with the distribution of listed salmonids. 

Figure 18 shows this overlap between rice and CV spring-run Chinook, Figure 19 shows the 

overlap with Sacramento River winter-run Chinook, and Figure 20 shows the rice overlap with 

CV steelhead. California is the nation’s second largest rice producing state, with 2011 

production totaling 4 1⁄2 billion pounds on over one-half million acres (USDA 2012). Current 

rice production is mainly in nine Sacramento Valley counties (Butte, Colusa, Glenn, Placer, 

Sacramento, Sutter, Tehama, Yolo, and Yuba) as displayed in Figure 17 (CRC 2011). Rice is 

also farmed in counties outside the Sacramento Valley; however currently the acreages in the 

San Joaquin Valley are generally small (CRC 2011). Currently, more than 95 percent of the 

state’s rice crop is grown within 100 miles of the city of Sacramento (CRC 2011). 

Figure 17. 2011 Sacramento Valley Rice Acres by County (CRC 2011). 
134

Figure 18. Central Valley rice growing overlay with listed CV spring-run Chinook ESU.
135

Figure 19. Central Valley rice growing overlay with Sacramento River winter-run Chinook ESU.
136

Figure 20. Central Valley rice growing overlay with CV steelhead DPS (note change in scale from 

previous figures).
137

Most California rice is produced by direct seeding into standing water (wet seeded), and a 

continuous flood is maintained for most of the season. Limited acreage is dry seeded, which also 

uses permanent flood after stand establishment. Seasonal rainfall and weather conditions 

influence rice planting and rice pesticide application. On a typical year, fields are planted from 

mid-April to the end of May, and field drainage occurs during August and September. 

Thiobencarb is typically applied in May (CRC 2011).

Reported annual use of thiobencarb in California declined from over one million lbs in 2000 to 

approximately 300 thousand lbs per year during the most recent surveys, 2006-2009 (CDPR 

2010a). Use of thiobencarb products within the state is reflective of the distribution of rice, the 

only registered use site for thiobencarb. Although rice acreage and thiobencarb use has remained 

relatively stable in California in recent years, differences have been observed among the rice 

producing counties in both the amount of thiobencarb used, and the selection of thiobencarb 

products (CDPR 2010a). As reported by California Department of Pesticide Regulation, from 

1980 to 2010, active ingredient thiobencarb has been typically applied at or near the maximum 

application rate of 4 lbs/A. Thiobencarb has liquid and granular (dry) formulations. The 

common liquid formulation in California is Abolish 8 EC. The most common granular 

formulation used in California is Bolero 15 G with a transition to Bolero UltraMax beginning in 

2008 (both are 15 percent thiobencarb products). According to the CDPR Pesticide Use Report 

(PUR), the relative proportions of active ingredient thiobencarb used in California has been 

running approximately 25 percent as Abolish 8 EC and 75 percent as Bolero UltraMax (PUR 

database, January 26, 2012).

California has set a drinking water standard known as a maximum contaminant level (MCL) for 

thiobencarb at 1.0 μg/L as a nuisance for taste at the intakes for Sacramento and West 

Sacramento. California’s Department of Pesticide Regulation (DPR), in cooperation with the 

Central Valley Regional Water Quality Control Board (CVRWQCB), developed recommended 

permit conditions to meet this water quality objective for thiobencarb. Under the CVRWQCB 

Water Quality Control Plan (Basin Plan) for the Sacramento River and San Joaquin River 

Basins, the performance goal in the agricultural drains is 1.5 μg/L. The CVRWQCB approved 

resolution can be reviewed at:
138

www.waterboards.ca.gov/centralvalley/board_decisions/adopted_orders/.

The rice industry, via the California Rice Commission (CRC), is responsible for water 

monitoring, annual reporting to the CVRWQCB, and coordinating the participation of all 

program stakeholders. The rice industry is ultimately responsible for meeting the water quality 

objectives. DPR as a co-regulator with the water boards continues to use their authority to 

regulate the sales and use of pesticides to address water quality issues involving pesticides. DPR 

works with CVRWQCB, the CRC, and the rice industry to address all rice pesticide issues and 

have developed criteria for water management, measures to minimize drift, and set minimum 

educational requirements for growers.11 

7.6.1. Water Management

Rice growing requires water to be held on fields for extended periods of time. Water must be 

held on the fields after thiobencarb is applied. The duration period varies depending on 

formulation. According to the labels EPA is considering for re-registration, liquid formulations 

(e.g., Abolish) require a 14-day hold period after application. Granular formulations (e.g., 

Bolero) specify field water is not to be released within 30 days of application. However, DPR 

has additional hold time restrictions that are presented in Table 19 (CDPR 2011). 

Table 19. California DPR thiobencarb water management requirements summary.

Water must be held for the 

indicated number of 24-hour 

periods on site or containment 

before release to State Waters.

Bolero 15-G Bolero 

UltraMax

Abolish 8EC

Hold Hold Hold

Single Field(e) 30 30 19

Single field Southern areas only 

(a).

19 19

Single permitted release into 

tailwater recovery system or 

pond onto fallow field [Except 

Southern area (a)].

14(b) 14(c) 14(b)

Multi-growers & district release 

onto closed recirculating 

systems.

6 6 6

 

11

With Resolution No. R5-2010-9001, the CRWQCB - Central Valley Region approved management practices for thiobencarb, that include the 

formation of a Storm Event Work Group, increased monitoring of thiobencarb, increased focus on seepage, restricting the use of thiobencarb near 

rivers, and increased education efforts including CRC-hosted preseason mandatory stewardship meetings.
139

Multi-growers & district release 

onto closed recirculating systems 

in Southern area(a).

6 6

Release from closed 

recirculating system. 19 19 19

Release into area that discharge 

negligible amount into perennial 

streams

19 19 6(d)

Emergency Release of tailwater 19 19 19

Commissioner verifies the 

hydrologic isolation of the fields 6 6 6

a – Sacramento/San Joaquin Valley defined as: South of the line defined by Roads E10 and 116 in Yolo County and the American River in 

Sacramento County.

b – Thiobencarb permit condition allows Bolereo 15G label hold period of 14 days.

c – Thiobencarb permit condition allows Bolero UltraMax label hold period of 14 days.

d – See hydrologic isolation fields.

e – When drainage begins after 30 day hold, discharge must not exceed two inches of water over a drain box weir for seven additional days. 

Unregulated discharges from these fields may then begin after 37 days.

Currently Bolero UltraMax is primarily used on water seeded fields. Abolish 8EC is primarily 

used on dry-seeded fields or fields that must be drained prior to application. 

Seepage through berms is a concern and sometimes an issue where treated fields can release 

thiobencarb earlier than the prescribed holding times. County agricultural commissioners 

(CACs) monitor for seepage when inspecting for water-holding compliance. Any visible 

seepage moving offsite during the water-holding period draining into waters of the State is 

considered an early release and is a violation of California’s water-holding requirements. Over a 

recent 5 year span (2006-2010), on average CACs inspected 773 fields each year. On average 

1.8 enforcement actions were taken; CRC annual reports may be found at: 

http://www.waterboards.ca.gov/centralvalley/water_issues/irrigated_lands/monitoring_plans_rep

orts_reviews/monitoring_report_reviews/coalitions/california_rice_commision/. 

7.6.1.1. Water Monitoring – the Rice Pesticide Program

The Rice Pesticide Program (RPP) is a longstanding watershed effort whereby rice growers 

follow Regional Water Board-approved management practices contained in use permits obtained 

from CACs when applying rice pesticides. The RPP includes monitoring of Sacramento Valley 

agricultural drains and the Sacramento River by the CRC (Figure 21 and Table 20). The Cities 

of Sacramento and West Sacramento also conduct monitoring at their drinking water intakes on 
140

the Sacramento River for thiobencarb. Monitoring is conducted twice weekly during the peak 

discharge period for six weeks from May until early June.

Figure 21. Rice Pesticide Program (RPP) Monitoring Sites
141

Table 20. RPP Monitoring Sites (CRWQCB-CVR 2005)

Abbreviation Name (County Location) Type

CBD5 Colusa Basin Drain (CBD) at Hwy 20 (Colusa County) Ag drain

CBD1 CBD at Road 99E (Yolo County) Ag drain

BS1 Butte Slough at Lower Pass Rd (Sutter County) Ag drain

SS1 Sacramento Slough at DWR gauging station (Sutter County) Ag drain

SR1 Sacramento River at Village Marina (Sacramento County) River

Municipal Intake Sites

SSR City of Sacramento Intake, Sacramento River 0.3 Km downstream of the American 

River (Sacramento County)

River

WSR City of West Sacramento Intake at Bryte Bend (Yolo County) River

The CRC maintains close communication with the CACs during the thiobencarb use season. 

California spring weather conditions can result in occasional storms, which cause concern during 

the thiobencarb water-holding period. The CRC implemented a Storm Event Work Group to 

facilitate communication with the CACs, DPR, water boards, registrants, city stakeholders and 

rice growers during storm events, and to determine an increase in thiobencarb monitoring based 

on information gathered from the group.

CACs also monitor compliance with prescribed holding times (Table 19). Over the same fiveyear period, CACs conducted 4,569 inspections and issued 11 civil penalties for non-compliance 

(average 2.2 penalties per year).

7.6.2. Drift Minimization

California has stricter standards for minimizing drift from aerial applications then those required 

on EPA approved labels. These are summarized below:

No aerial applications shall be made or continued within 1⁄2 mile of the Sacramento or 

Feather Rivers in the Sacramento Valley rice growing counties of Butte, Colusa, Glenn, 

Placer, Sacramento, Sutter, Tehama, Yolo, and Yuba unless there is a continuous positive 

airflow away from the river.

In the Sacramento Valley rice growing counties of Butte, Colusa, Glenn, Placer, 
142

Sacramento, Sutter, Tehama, Yolo, and Yuba, no aerial application shall be made or 

continued within 1⁄2 mile of the Sacramento or Feather Rivers when the wind speed 

exceeds 7 miles per hour.

In Sacramento and Yolo Counties, no aerial applications shall be made or continued 

within 1⁄4 mile of the Sacramento River unless they are made under the direct supervision

of the county agricultural commissioner’s representative.

In Sacramento and Yolo Counties, the maximum acres treated by air each day within 1⁄4 

mile of the Sacramento River shall not exceed 33 percent of the average acres treated per 

day by air within this area in each county during 2002. 

In addition to the above, DPR provides focused oversight inspections of thiobencarb aerial 

applications to monitor drift mitigation requirements. From 2006-2010, there were 144 

inspections. During these inspections, six compliance violations were issued (CRC Annual 

Reports).

7.6.3. Applicator Education 

The CVRQWCB resolution approving the use of thiobencarb under the Rice Pesticide Program,

requires the CRC to host annual thiobencarb mandatory stewardship meetings. The pre-season 

meetings, in collaboration with Valent, the CACs, DPR and the CVRWQCB, provides 

stewardship to the permit applicant and/or his/her authorized representative. In addition, the 

CRC extends stewardship to the pest control advisors (PCAs) and pilots through collaboration 

with the California Aerial Applicators Association (CAAA). The permit applicant must have a 

certification of completion, issued by the CRC, in order to obtain a restricted materials permit 

from the CAC. 

Thiobencarb is a restricted material in California requiring a permit is issued by the CAC in the 

county the rice is grown. The DPR restricted materials permit conditions require growers to add 

thiobencarb to the permit on an annual basis. Under the restricted materials program, the grower, 

or his/her representative must file a Notice of Intent (NOI) within 24-hours prior to commencing 
143

the use of a pesticide requiring a permit. The CACs retain the right to deny permits and 

thiobencarb NOIs if the field conditions are not adequate for a thiobencarb application. 

All applications require a written recommendation from a licensed PCA prior to the submittal of 

the NOI. The permit is also required for the use and sale of the restricted material through a 

licensed dealer, or business. A licensed private applicator, licensed pilot, or qualified applicator 

must apply the restricted material as authorized by DPR. 

The CDPR publishes voluntary interim measures for mitigating the potential impacts of pesticide 

usage to listed species. These measures are available online as county bulletins 

(http://www.cdpr.ca.gov/docs/endspec/colist.htm). Measures from these county bulletins that 

apply to thiobencarb use in salmonid habitat are:

Do not use in currently occupied habitat except as specified in Habitat Descriptors, in 

organized habitat recovery programs, or for selective control of exotic plants.

For sprayable or dust formulations: when the air is calm or moving away from 

habitat, commence applications on the side nearest the habitat and proceed away from 

the habitat. When air currents are moving toward habitat, do not make applications 

within 200 yards by air or 40 yards by ground upwind from occupied habitat. The 

CAC may reduce or waive buffer zones following a site inspection, if there is an 

adequate hedgerow, windbreak, riparian corridor or other physical barrier that 

substantially reduces the probability of drift. 
144

8. Effects of the Proposed Action to Threatened and Endangered Pacific 

Salmonids

The analysis includes three primary components: exposure, response, and risk characterization. 

We analyze exposure and response, and integrate the two in the risk characterization phase 

where we address support for risk hypotheses. These risk hypotheses are predicated on effects to 

salmonids. Designated critical habitat is analyzed separately and predicated on effects to 

salmonid-supporting habitats (see Effects of the Proposed Action to Designated Critical Habitat 

and Integration and Synthesis for Designated Critical Habitat).

8.1. Exposure Analysis

In this section, we identify and evaluate potential exposure of salmonids to the stressors of the 

action (Figure 22). We begin by presenting general life history information of vulnerable life 

stages of Pacific salmon and steelhead. Next, we discuss the physical and chemical properties of 

thiobencarb and its degradation products that influence exposure of listed species and designated 

critical habitat to these stressors of the action. We then evaluate co-occurrence of salmon habitat 

with the stressors of the action by comparing the distribution of sites authorized for thiobencarb

use by product labeling to the distribution of each species and their designated critical habitat. 

To further characterize exposure where co-occurrence exists, we summarize EPA exposure 

estimates presented in BEs, present additional exposure estimates for shallow floodplain habitats 

utilized by salmonids, and summarize the available water quality monitoring data. Finally, we 

conclude with a summary of anticipated ranges of exposure when pesticide use is proximate to 

salmon habitats, and characterize the uncertainty contained in this analysis. Because the ESA 

section 7 consultation process is intended to insure that the agency action is not likely to 

jeopardize listed species or destroy or adversely modify critical habitat, NMFS considers a 

variety of exposure scenarios in addition to those presented in EPA’s BEs. These scenarios 

provide thiobencarb estimates for the range of habitats used by listed salmonids. 
145

Figure 22. Exposure analysis

8.1.1. Threatened and Endangered Pacific Salmonids use of Aquatic Habitats

Within the Status section we discussed salmonid lifecycles, life histories, and the use and 

significance of aquatic habitats. Listed salmonids occupy a variety of aquatic habitats that range 

from shallow, low-flow freshwaters to open reaches of the Pacific Ocean. All listed Pacific 

salmonid species use freshwater, estuarine, and marine habitats at some point during their life. 

The temporal and spatial use of habitats by salmonids depends on the species and the 

individuals’ life history and life stage as well as environmental factors such as river flows. 

In this section we describe the habitats used by the three species of listed salmonids that occur in 

California’s Central Valley rice growing region: Central Valley Spring-run Chinook salmon, 

Sacramento River Winter-run Chinook salmon, and California Central Valley Steelhead. 

General life history descriptions illustrating the use of aquatic habitats by Chinook salmon and 

steelhead are provided below in Table 21. Additionally, we describe species temporal use of 

aquatic habitats of the three species within California’s Central Valley to determine potential 

spatial and temporal overlap with thiobencarb.

Co-occurrence of action stressors 

and listed species

Distribution of 

individuals

Exposure Profile

Distribution of 

habitat
146

Table 21. General life histories of Pacific salmonids which utilize habitat that overlaps with 

thiobencarb spatial use patterns.

Species General Life History Descriptions

(number of 

listed 

ESUs or 

DPSs)

Spawning Migration Spawning Habitat Juvenile Rearing and Migration

Chinook 

(2)

Mature adults (usually 

three to five years old) 

enter rivers (spring 

through fall, depending on 

run). Adults migrate and 

spawn in river reaches 

extending from above the 

tidewater inland hundreds 

of miles from the Pacific. 

Migrating adults typically 

follow the thalweg.

Chinook salmon migrate 

and spawn in four distinct 

runs (spring, fall, summer, 

and winter). Chinook 

salmon are semelparous1.

Generally spawn in 

the middle and 

upper reaches of 

main stem rivers 

and larger tributary 

streams.

The alevin life stage primarily 

resides just below the gravel 

surface until they approach or 

reach the fry stage. Immediately 

after leaving the gravel, fry 

distribute to habitats that provide 

refuge from fast currents and 

predators. Juveniles exhibit two 

general life history types:

Ocean-type fish migrate to sea in 

their first year, usually within six 

months of hatching. Ocean-type 

juveniles may rear in the estuary 

for extended periods. Streamtype fish migrate to the sea in the 

spring of their second year. 

Steelhead 

(1)

Mature adults (typically 

three to five years old) 

may enter rivers any 

month of the year, and 

spawn in late winter or 

spring. Migrating adults 

typically follow the 

thalweg. Steelhead are 

iteroparous2.

Usually spawn in 

fine gravel in a 

riffle above a pool. 

The alevin life stage primarily 

resides just below the gravel 

surface until they approach or 

reach the fry stage. Immediately 

after leaving the gravel, swim-up 

fry usually inhabit shallow water 

along banks of stream or aquatic 

habitats on streams margins. 

Steelhead rear in a wide variety 

of freshwater habitats, generally 

for two to three years, but up to 

six or seven years is possible. 

They smolt and migrate to sea in 

the spring. 

1 spawn only once

2 may spawn more than once

Freshwater, estuarine, and marine near-shore habitats are areas subject to pesticide loading from 

runoff and drift given their proximity to pesticide application sites. Small streams and many 

floodplain habitats are more susceptible to higher pesticide concentrations than other aquatic 

habitats used by salmon because their physical characteristics provide less dilution and 

dissipation. Examples of floodplain habitats include alcoves, channel edge sloughs, overflow 

channels, backwaters, terrace tributaries, off-channel dredge ponds, off-channel ponds, and 
147

braids (Anderson 1999, Beechie and Bolton 1999, Swift III 1979). Though floodplain habitats 

typically vary in surface area, volume, and flow, they are frequently shallow, low to no-flow 

systems protected from a river’s, or a stream’s, primary flow. 

Rearing and migrating juvenile salmonids use floodplain habitats extensively (Beechie and 

Bolton 1999, Beechie et al. 2005, Caffrey 1996, Henning et al. 2006, Montgomery 1999, Morley 

et al. 2005, Opperman and Merenlender 2004, Roni 2002). Diverse, abundant communities of 

invertebrates also populate floodplain habitats and many species represent an important food 

source for salmon. The presence of abundant food resources is partially responsible for juvenile 

salmonids reliance on these habitats. Both stream-type juvenile Chinook salmon and steelhead 

use floodplain habitats for extended durations (several months). 

8.1.2. Chemical Exposure Pathways to Salmonids Habitats

Pesticides can contaminate surface waters via runoff, erosion, leaching, spray drift from 

application at terrestrial sites or direct application to aquatic habitats, and atmospheric 

deposition. Aquatic habitats can be contaminated by pesticides applied to terrestrial target sites 

through several pathways and by direct application to surface waters for control of plants, 

mosquitoes, and other aquatic pests. For example, spray-applied pesticides may result in offtarget deposition of droplets at the time of application. The likelihood of spray drift to an aquatic 

habitat is determined by the application method, the proximity to the habitat, and meteorological 

conditions at the time of application. 

Surface water contamination is also influenced by the environmental fate properties of the 

chemical. For example, secondary drift or vapor drift is dependent on a chemical’s volatility and 

refers to the redistribution of pesticides from plant and soil surfaces through volatilization and 

subsequent atmospheric deposition. Runoff and leaching, the horizontal and vertical movement 

of pesticides with rainwater or irrigation water, are influenced by chemical-specific properties 

that determine the compound’s persistence and mobility in soil and water. Standardized tests are 

typically used to characterize mobility (e.g., solubility, Kd and Koc) and persistence under 

different environmental conditions (e.g., hydrolysis, photolysis, and metabolism half-lives in 

aerobic and anaerobic environments). Below we present environmental fate properties of 
148

thiobencarb to characterize the relative importance of different exposure pathways in terms of 

the potential for the active ingredient and its degradates to contaminate salmonid bearing habitats 

and designated critical habitats.

8.1.3. Summary of Chemical Fate of Thiobencarb 

Thiobencarb can be applied to rice either before or after rice fields are flooded. After a specified 

holding period, rice irrigation water can be discharged to salmonid habitats via a system of drains 

within the Central Valley of California. Salmonids may be exposed to thiobencarb when present 

in the water column where the chemicals cross gill surfaces during respiration, or where fish 

sensory systems come in direct contact with contaminated water (i.e., olfactory sensory neurons). 

Other routes may contribute to overall exposure including incidental ingestion of the chemical in 

sediment or ingestion of the chemical in food items. Below we summarize chemical fate 

properties of thiobencarb reported by EPA. Where discrepancies existed between EPA 

assessments, we deferred to the more recent document. 

8.1.3.1.

Figure 23. Chemical structure of thiobencarb

Thiobencarb (Figure 23) is a thiocarbamate herbicide relatively stable to abiotic and biotic 

degradation in water (Table 24). However, it is not expected to persist in surface water given its 

affinity for soil and sediment. Two field studies reported median partitioning of thiobencarb 

residues of 5.6:1 and 6.6:1 for the soil:water ratio (EPA 2009b). Aqueous photolysis occurs 

slowly in surface waters with a half-life of 190 d. However, it may occur more quickly in the 

presence of photo-sensitizers based on its degradation in acetone (half-life of 12 d). The 

photodegradation rate of thiobencarb in the presence of relevant photo-sensitizers, such as humic 

substances that occur in surface waters, is uncertain. Degradation in soils also occurs slowly, as 
149

evidenced by soil photolysis half-lives of 168-280 d and aerobic soil metabolism half-lives of 27-

58 d. The Koc’s of 384-1438 indicate thiobencarb is moderately to slightly mobile in the soil. 

The vapor pressure indicated that thiobencarb has an intermediate volatility from dry surfaces 

and the Henry’s Law Constant suggests it will not volatilize from water (EPA 2009b).

Thiobencarb has been found in trace concentrations in air and precipitation and is expected to 

travel up to 5 km from the site of application. However, thiobencarb has an atmospheric 

degradation half-life of 0.421 days suggesting a limited potential for long-range transport (EPA 

2009b). Potential transport mechanisms include primary and secondary spray drift, surface water 

runoff, and rice paddy discharge and seepage. Thiobencarb moderately accumulates in fish 

tissue but depuration occurs rapidly with 93-95% of residues eliminated in three days (EPA 

2009b).

Table 22. Environmental fate characteristics of thiobencarb1.

Parameter Value

Water solubility 30 mg/L at 20 °C

Vapor pressure 1.476 x 10-6- 2.2 x 10-5mm Hg

Henry's law constant 2.49 x 10-7atm m3 mol -1

Octanol/Water partition coefficient Log Kow = 1.3 – 3.42

Hydrolysis (t1⁄2) pH 5, pH 7, & pH 9 Stable

Aqueous photolysis (t1⁄2) 190 d ; 12 d in acetone 

Soil photolysis (t1⁄2) 168 - 280 d

Aerobic soil metabolism (t1⁄2) 27 - 58 d

Anaerobic soil metabolism (t1⁄2) Stable

Aerobic aquatic metabolism (t1⁄2) Stable

Anaerobic aquatic metabolism (t1⁄2) Stable

Soil partition coefficient Koc = 384-1435 L/kgsoil

Fish Bioconcentration Factor (BCF) 128x (edible)

639x (non-edible)

411x (whole fish)

1- (EPA 2009b)

8.1.3.2.

8.1.3.3. Degradates of thiobencarb 

The molecular structure of a pesticide may be modified by biotic (e.g., microbial metabolism) or 

abiotic (e.g., photolysis and hydrolysis) processes. The products of these processes typically have 

different toxicities, environmental fate characteristics, and risks compared to the parent pesticide. 

EPA indicated that the main transformation products of thiobencarb found in laboratory 

photolysis and metabolism studies were 4-chlorobenzoic acid, 4-chlorobenzaldehyde, and carbon 
150

dioxide (EPA 2009b). Two additional microbial metabolites were identified as thiobencarb 

transformation products monitored in field dissipation studies (Table 23). 

Table 23. Degradates of thiobencarb (EPA 2009b).

Transformation product Percent of 

applied parent

4-chlorobenzoic acid 561

5

2

4-chlorobenzaldehyde 29.43

Carbon dioxide 23 - 772

Thiobencarb sulfoxide Not specified4

4-chlorobenzylmethylsulfone Not specified4

1- Aqueous photolysis

2- Soil metabolism (aerobic or anaerobic conditions not specified)

3- Photolysis (aqueous or soil matrix not specified)

4- Aquatic field dissipation studies

Existing laboratory studies provided insufficient information to characterize aquatic exposure. 

Field dissipation studies are relatively more informative than laboratory studies because they 

provide information on the concentrations and persistence of thiobencarb and two metabolites in 

flooded rice paddies (Table 24).

Table 24. Concentration of thiobencarb and metabolites reported in field dissipation studies

Study Results

Field dissipation of 

Bolero 8EC in rice

(Lai 1991)(MRID#42003404)

Port Barre, Louisiana field site

Aerial application of liquid formulation at 

4 lbs a.i./acre to dry-seeded rice

A “flush irrigation” was conducted 3 days 

post application. The plots were 

permanently flooded to 4.5 inches at 7 

days post-application 

Parent Thiobencarb

Mean concentration (μg/L)

Day 7: 8.2

Day 8: 11.9

Day 10: 13.0 (highest mean concentration 

3 days post flooding)

Day 12: 10.3

Day 28: 0.58

Thiobencarb sulfoxide 

Highest mean concentration 8.9 μg/L, Day 8 

(1 day post-flood)

4-chlorobenzylmethylsulfone 

Highest mean concentration 5.2 μg/L, Day 12
151

Study Results

(5 days post-flood)

Dissipation half-life:

Parent thiobencarb 5.8 d

Thiobencarb sulfoxide 3.4 d

4-chlorobenzylmethylsulfone 6.0 d

Aquatic field dissipation of 

Bolero 10G in rice

(Ho 1990) (MRID#43404005)

Nelson, California field site

Aerial application of granular formulation 

to flooded rice at a rate of 4 lbs a.i./acre

Flood water maintained for 6 days then 

released to fallow check where it was 

held for 8 days then re-circulated 

Parent Thiobencarb 

Mean concentration μg/L

Day 0: 267 

Day 1: 252

Day 2: 341

Day 3: 438

Day 5: 159

----------- drained-----------

Day 14: 44

Day 33: 3

Day 92: 1

Peak mean concentration 438 μg/L

observed 3-d post application

Thiobencarb sulfoxide 

Day 3 highest mean concentration: 22 μg/L

4-chlorobenzylmethylsulfone

Day 10 highest mean concentration: 8.3 μg/L

Dissipation half-life:

Parent thiobencarb 8.7 d

Thiobencarb sulfoxide 2.8 d

4-chlorobenzylmethylsulfone 10.4 d

Fate of thiobencarb and 

molinate in rice fields

(Ross and Sava 1986)

Aerial application of granular formulation 

Parent Thiobencarb (No degradates measured)

Mean concentration μg/L

Day 0: 79

Day 2: 567

Day 4: 576
152

Study Results

to flooded rice at a rate of 4 lbs a.i./acre

Water depth held at 10 inches ± 2 inches 

for 6 day holding period, then all water 

drained from field. Water depth was then 

maintained at 7 inches ± 2 inches with 

inflow/outflow during post-holding

Day 6: 515

----------- drained-----------

Day 8: 367

Day 16: 56

Day 32: 8

Peak mean concentration 576 μg/L

observed 4-d post application

Dissipation half-life > 6 days

Thiobencarb did not decline significantly 

during the 6-d holding period 

8.1.4. Exposure of salmonid habitats to the stressors of the action 

8.1.4.1. Co-occurrence associated with thiobencarb use. 

We evaluated co-occurrence of listed salmonids with stressors of the action by comparing the 

spatial and temporal distribution of salmon with potential use of thiobencarb based on label 

specifications. To evaluate the areal extent of application sites near salmon-bearing waters, we

used a GIS overlay containing land use classifications and salmon distributions12 to determine 

overlap with the three California Central Valley ESUs/DPSs (Table 16, Figure 11). Based on the 

2006 National Land Cover Database, the spatial coverage of cropland accounts for 

approximately 21-26% of the total area within the freshwater distribution of the three species (

Table 25). Because cropping patterns may change over time, the land use classification for 

cropland was used for the initial evaluation of registered uses. Statistics from CDPR indicate 

that more than a half million acres have been planted annually in rice in the Sacramento Valley 

since 2006 (CDPR 2011). Rice is the dominant crop in the Sacramento Valley and accounts for 

more than 6% of the total area with the distribution of the California Central Valley Steelhead 

and more than 8% of the total areas within the distribution of spring-run and winter-run Chinook. 

 

12 http://www.nwr.noaa.gov/ESA-Salmon-Listings/Salmon-Populations/Maps/
153

Table 25. Co-occurrence of listed Pacific salmonids with potential application of pesticides to use 

sites within the salmonids' freshwater distribution.

ESU / DPS Spatial coverage of 

cropland within species 

freshwater distribution

Temporal overlap of ESU 

and labeled use of 

Thiobencarb

Central Valley Spring-Run Chinook Salmon 21.0% Yes

Sacramento River Winter-Run Chinook 21.0% Yes

California Central Valley Steelhead 26.4% Yes

There is temporal overlap of approved labeled uses with species presence in freshwater habitats 

because most of the listed Pacific salmonids in California occur in freshwater year-round in some 

life stage, and thiobencarb labels place no restrictions on the timing of application. Historically 

thiobencarb has been applied most frequently in the months of May and June (Figure 24). A 

considerable number of applications (20 or more during the month in a given year) have 

occurred as early as April and as late as July (Figure 24). Application of thiobencarb during 

other months has been extremely rare (average 0 - <1 application/month). Peak aquatic 

concentrations associated with the drift pathway are expected to occur on the day of application. 

Exposure associated with thiobencarb discharge from treated fields is expected to occur after the 

14 and 30 day holding requirements have been met for liquid and granular formulations, 

respectively. However, exposure associated with discharge from treated fields may also occur 

earlier due to thiobencarb seepage through berms or checks or through emergency releases due to 

flooding or salinity. 

After reviewing thiobencarb’s label authorized use and historical use in California, we expect 

designated critical habitats of the listed Central Valley salmonids to be exposed to thiobencarb 

from April – July, and particularly during May and June (Figure 24). The application and 

discharge period for thiobencarb overlaps with the general timing of freshwater residence of 

several salmonid life stages (Table 26). For example, the peak May – June application period 

overlaps with adult freshwater migration and juvenile freshwater rearing and migration in all 

three Central Valley species. Sacramento River winter-run Chinook are also actively spawning, 

and are present in freshwater as eggs, alevin, and fry during the application and discharge period 

(April-July). While Central Valley steelhead are likely to have completed spawning activities 

prior to applications in May and June, eggs, alevin and fry are all present in freshwater habitats 
154

during this timeframe. From these patterns, we expect all three species and their designated 

critical habitat to be exposed to thiobencarb and other ingredients in thiobencarb products. . 

Figure 24. Number of thiobencarb applications in California by month from 1995-2010 (source: 

CDPR PUR database). Points in figure represent the number of application for a year. Months with 

zero applications are not plotted. The solid horizontal line represents the monthly mean number of 

applications over all of the years.
155

Table 26. Generalized run-timing of ESA-listed Central Valley Pacific Salmonids by life stage.

ESU/DPS Life Stage Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Central 

Valley

Spring-run 

Chinook

Adult prespawn migration

Spawning

Incubation

Alevin to fry emergence

Rearing and migration

Sacramento 

River

Winter-run 

Chinook

Adult prespawn migration

Spawning

Incubation

Alevin to fry emergence

Rearing and migration

Central 

Valley

Steelhead

Adult prespawn migration

Spawning

Incubation

Alevin to fry emergence

Rearing and migration

8.1.4.2. Species-specific temporal and spatial considerations

Because thiobencarb is registered for use on only one crop that has a relatively localized and 

predictable distribution, we were able to take a more in-depth approach to assess temporal and 

spatial co-occurrence than was possible in previous NMFS Opinions on EPA registration actions. 

This approach assumes that within the distribution of listed species, (1) rice production will 

continue to be confined to the Central Valley of California, and (2) thiobencarb will not be 

approved for other uses within the species range. Reinitiation of formal consultation will be 

required if there are changes in these use patterns or Washington, Oregon, or Idaho authorize use 

of thiobencarb (see Reinitiation Notice below). We expect co-occurrence of thiobencarb use 

with several different life-history stages of each listed species (Figure 24, Table 26). To 

understand the extent of potential spatial and temporal overlap between thiobencarb use (April –
156

July) and species presence we discuss each species’ adult and juvenile migration through rice 

growing areas. Figure 25 shows the distribution of rice in the Sacramento and upper San Joaquin 

Valleys according to the 2010 NASS Crop Data Layer. 
157

Figure 25. Distribution of rice in the Sacramento Valley based on the 2010 NASS Crop Data Layer. 

North Delta, Knights Landing (KL) and Red Bluff Diversion Dam (RBDD) fish sampling sites and 

distribution of the Central Valley Spring Run Chinnok are also shown.
158

8.1.4.3. Abundance of the Three Species in CA Central Valley

The relative abundance of the listed salmonids at various sampling stations is reported in Table 

27, Table 29, and Table 31. We scored relative abundance as “low,” “medium,” or “high” at the 

different sampling locations over two-week intervals qualitatively by comparing capture rates at 

sample stations to overall abundance of each species observed at that station during the year. For 

example, several thousand spring-run Chinook are monitored annually at the Knights Landing

(KL) station (Snider and Titus 2000a, Bajjaliya and Vincik 2008a). The relative abundance 

ratings in Table 27 of low, medium, and high correspond to capture rates in the 10’s, 100’s, and 

1,000s, respectively, for that species. We used this information to determine the likely extent of

co-occurrence of the three species with thiobencarb (Table 28, Table 30, and Table 32). This 

qualitative analysis is based on two primary assumptions: first, the spatial pattern of current rice 

growing area as defined by the 2010 NASS Crop Data Layer is a reasonable surrogate for future 

thiobencarb use sites, and, second, that the timing of thiobencarb applications and release of 

water from thiobencarb treated fields will remain relatively comparable for the next 15 years. 

We find support for these assumptions from the conditions needed for successful rice growing in 

California’s CV including temperature, water, and soil. These necessary conditions are not met 

by other geographic areas within CA, therefore, the geographical expansion outside of the 

Central Valley is unlikely (Hill et al. 1992, California 1993)

Using fish collection data from CDFG and USFWS, we evaluated typical species presence at 

sampling stations on the Sacramento River (at Red Bluff Diversion Dam (RBDD) and Knights 

Landing (KL)), in the North Delta of the Sacramento River (at Chipps Island and Hood), on the 

San Joaquin River (at Mossdale), and on the Mokelumne River (at Woodbridge dam) (Figure 25, 

(USFWS 2010a, USFWS 2012a, Marcotte 1984a, Bajjaliya and Vincik 2008a, Snider and Titus 

2000a). We used these monitoring locations to help us understand the movement of the species 

upstream as adults and downstream as juveniles as they entered and left current rice growing 

areas. The current distribution of rice in California is predominately located in the Sacramento 

Valley downstream of the RBDD and upstream of the North Delta sampling stations. The KL 

sampling station is within current rice growing areas. Steelhead monitored at Mossdale and 
159

Woodbridge stations also represent occurrence in rice growing areas, although rice is grown to a 

lesser extent in these locations compared to the KL area on the Sacramento River. 

8.1.5. Analysis of Species Presence in Rice Growing Areas

8.1.5.1. Co-occurrence of Central Valley spring-run Chinook salmon with thiobencarb

It is likely that adult Central Valley (CV) spring-run Chinook will be exposed to thiobencarb as 

they travel through rice growing areas while migrating up-stream to cooler, higher elevation 

waters because their abundance is high in rice growing areas when thiobencarb is most 

frequently applied (Table 7). Since peak aquatic concentrations associated with the drift 

pathway are expected to occur on the day thiobencarb is applied, individuals may be exposed 

multiple times during their upstream migration (i.e., they may be exposed to thiobencarb at 

multiple locations given the distribution of potential application sites along their migration 

corridor). Individuals could also be exposed during run-off events involving normal discharges, 

unintended seepage from the fields, or from approved emergency discharges. The occurrence of 

seepage has been relatively infrequent, occurring in less than 0.5% of rice fields inspected over a 

five year period (Table 43). 

Adult holding and spawning mostly occur at higher elevations in the upper main stem 

Sacramento River tributaries of Tehama and Butte counties (e.g., Mill Creek and Deer Creek, 

(Figure 26), while rice is grown primarily on the valley floor (Lindley et al. 2004a, Yoshiyama et 

al. 1998, Marcotte 1984a). In the main stem, most spawning is above RBDD; in Mill Creek 

spawning extends from above the Little Mill Creek confluence upstream to about one mile above 

the Highway 36 bridge; in Deer Creek most spawn from about the Ponderosa Way bridge 

upstream to upper Deer Creek falls; in Butte Creek, most spawning takes place about 40 miles 

upstream from the mouth. Some spring-run fish may spawn in the main stem below RBDD but 

estimates are not available and these fish are often considered fall-run Chinook (CDFG 1995, 

Marcotte 1984a).
160

Figure 26. CV spring-run Chinook probable and known spawning areas in relation to 2010 rice 

growing areas.
161

(Good et al. 2005). Spawning also occurs in the Feather River, but it begins in August and ends 

in October. Overall, we do not anticipate spawning, egg incubation, and fry emergence to cooccur with thiobencarb runoff in rice growing areas. 

Exposure to significant concentrations of thiobencarb via primary drift is also not anticipated

given what we understand about the timing and proximity of spawning to expected thiobencarb 

use sites (Table 28). A drift event during application of thiobencarb liquid formulations could 

lead to deposition into spawning areas, but we expect such deposition to be limited and more of 

an exception than the rule. Furthermore, should primary drift occur, it would likely happen prior 

to the onset of spawning.

After fry emergence, most juveniles rear in their natal streams (Mill, Deer, and Butte creeks) 

upstream of the rice growing areas (Lindley et al. 2004a, Marcotte 1984a). Some fry (young of 

the year) may begin migration downstream post emergence from the gravel, but this typically 

occurs in the fall (Lindley et al. 2004a). Based on fish monitoring programs in the Sacramento 

(Snider and Titus 2000a, Bajjaliya and Vincik 2008a, Marcotte 1984a, USFWS 2010b), juvenile 

migration out of natal streams and through rice growing areas is initiated in late summer and 

largely complete by May, when the majority of thiobencarb is typically applied (Figure 24 and 

Table 27). Juveniles are found below Knights Landing (KL in Figure 25 and Table 27) and in 

the North Delta by April (Figure 25 and Table 27). Some juveniles do occupy waters in rice 

growing areas into early May. We expect few juveniles to occupy water within, or adjacent to,

rice areas in June. Exposure from drift is expected to be limited as juvenile abundance in close 

proximity to rice fields is expected to be low during applications (Table 27 and Table 28). 

Juveniles are abundant in the San Francisco Bay estuary (i.e., in the North Delta; downstream of 

rice growing areas) during thiobencarb application and discharge periods. We expect exposure 

in the estuarine habitats will be at concentrations that are low compared to concentrations that 

are found in salmon habitats located closer to field discharge. For example, the highest 

concentration reported by Kuivila and Jennings (2007) in eastern Suisun Bay, at Mallard Island 

(approximately one-half mile from Chipps Island), was 66 ng/L, 1-2 orders of magnitude lower 

than values frequently detected at sample stations located near rice cultivation (summary of Rice 

Pesticide Monitoring Program reported below). 
162

The Yolo Bypass is a historical floodplain of the Sacramento River that is now utilized to 

manage flood waters and produce a variety of crops. It is seasonally inundated with water to 

varying degrees and is recognized as an important rearing area for salmonids in the Sacramento 

Valley. During high water years, when the Yolo Bypass floods, juveniles are expected to occupy 

the Yolo and linger there until flood waters recede (Sommer et al. 2001, Sommer et al. 2005a). 

Rice growing does occur in the Yolo Bypass, but is delayed until after flood waters recede. 

Therefore, during flood years when juveniles have access to the Yolo Bypass, co-occurrence 

with thiobencarb use is unlikely. 

Table 27. Temporal occurrence of adult and juvenile CV spring-run Chinook salmon in the 

Sacramento River. Darker shades indicate months of greatest relative abundance in each location 

along the river.

Adult upstream migration

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. River basin

Up Sac. R. mainstem

Juvenile rearing and downstream migration

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. R. at RBDD1

Sac. R. at KL2

North Delta3

Thiobencarb use/release

Relative Abundance = High = Medium = Low

1

upstream of current rice growing areas; 2 within current rice growing areas; 3downstream of current rice growing areas.
163

Table 28. CV spring-run Chinook relative abundance in rice areas (Figure 25) during thiobencarb 

use periods April - July.

Life Stage None Expected Low Medium High

Adult Migration X

Spawning X

Early rearing X

Juvenile 

migration X

8.1.5.2. Co-occurrence of Sacramento River winter-run Chinook with thiobencarb

Typically, adult Sacramento River winter-run Chinook salmon enter fresh water in winter or 

early spring. The adults migrate through rice growing areas of the Sacramento River Basin en 

route to spawning grounds in medium abundance during the time when thiobencarb is usually 

applied and discharged. Spawning usually occurs between May and June, the peak period of 

thiobencarb applications. Moderate numbers of adults moving up-river to reach spawning 

grounds are likely exposed to thiobencarb by drift or runoff from thiobencarb applications (Table 

30). As discussed in the Status of Listed Resources section, winter-run Chinook spawning is 

now restricted within roughly 44 miles of the main-stem upper Sacramento River immediately 

below Keswick Dam (Figure 27). A majority of the spawning (greater than 50 percent) is 

between Keswick Dam and the ACID Dam approximately 5 miles away. All of the spawning is

located upstream of current RBDD, above the areas where rice is currently cultivated (Figure 25

and Figure 27). Consequently, we do not expect spawning, egg incubation, and fry emergence to 

co-occur with thiobencarb based on the current location of spawning areas. 
164

Figure 27. Sacramento River winter-run Chinook spawning areas relative to rice growing areas. 

Spawning is confined to the 44 river miles immediately downstream of Keswick Dam (Yoshiyama 

et al. 1998). In California, Rice is primarily cultivated in the Sacramento Valley downstream of the 

Red Bluff Diversion Dam (RBDD).
165

Upon emergence, most fry take up residence in their natal stream for several weeks to a year or 

more, while others are displaced downstream by the stream’s current. Early hatched fry that do 

move down soon after emergence, instead of rearing in their natal waters, may be exposed to 

thiobencarb. Monitoring studies suggest juveniles are either absent, or in low abundance, in rice 

growing areas during the months of April – July when thiobencarb exposure is most likely 

(Table 29). Relative abundance estimates suggest most fish will have passed through the lower 

end of the rice growing area (i.e., the lower Sacramento River monitoring stations) prior to May. 

Early hatched fry start arriving at the Red Bluff Diversion Dam (RBDD) in low relative 

abundance in July. While the RBDD is above rice growing areas (Figure 25), some of these fish 

could migrate down into waters that may receive late discharge waters off rice fields following a 

30-day hold. However, according to fish monitoring studies, most juvenile winter-run Chinook 

migrate through the rice growing areas during the period when thiobencarb is not being applied,

August – April (Snider and Titus 2000a, Bajjaliya and Vincik 2008a, Marcotte 1984a, USFWS 

2010b). This suggests exposure from drift and runoff into habitats in close proximity to rice 

fields would occur to a relatively low number of individuals (Table 30). 

As with the spring-run Chinook, the Yolo Bypass and North Delta represent important rearing 

areas for winter-run Chinook. During high water years, when the Yolo Bypass floods early, 

juveniles may linger in floodplains habitats until flood waters recede. However, rice growing in 

the Yolo Bypass is delayed until after flood waters recede. Therefore, during flood years when 

juveniles have access to the Yolo Bypass, co-occurrence with thiobencarb use is unlikely. 

Juveniles rear in high abundance in the North Delta during the periods of thiobencarb application 

and discharge of thiobencarb-treated fields. Although these habitats are downstream of rice 

growing areas, they are generally not in close proximity to rice fields. Exposure to thiobencarb 

is expected to be significantly less than predicted by drift and runoff models for water bodies in 

close proximity to rice fields. 
166

Table 29. Temporal occurrence of adult and juvenile winter-run Chinook salmon in the 

Sacramento River. Darker shades indicate months of greatest relative abundance in each location 

along the river.

Adult migration

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. River basin

Upper Sacramento River

Juvenile migration

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. River at RBDD1

Sac. River at KL2

Lower Sac. River3

North Delta3

Thiobencarb use/release

RBDD=Red Bluff Diversion Dam KL = Knights Landing

Relative Abundance: = High = Medium = Low

1

upstream of current rice growing areas; 2 within current rice growing areas; 3downstream of current rice growing areas.

Table 30. Sacramento River winter-run Chinook relative abundance in rice areas (Figure 25) 

during thiobencarb use periods April – July.

Life Stage None expected Low Medium High

Adult migration X

Spawning X

Early rearing X

Juvenile 

migration X

8.1.5.3. Co-occurrence of Central Valley steelhead with thiobencarb

Generally, CV steelhead that utilize the Sacramento River Basin enter freshwater from late June

to April, but some individuals are present in freshwater year-round. Steelhead that utilize the 

San Joaquin and its tributaries generally begin to enter the river in July. Relatively few prespawn adults migrate through rice production areas (Figure 24) at the time thiobencarb is applied 
167

or when held water is released from rice fields post application (Table 32 and Table 31). Some 

co-occurrence is likely, but the number of individuals exposed during adult migrations is 

expected to be relatively low as most of the pre-spawning adults would be higher in the 

watershed (above RBDD) during the application and discharge period (April-July; Table 32). 

Unlike salmon, steelhead can spawn several times before they die and therefore spawned adults 

moving down river as kelts could be exposed to thiobencarb via mechanisms described above. 

We do not have data indicating relative abundance of down migrant kelts. 

CV Steelhead have the longest freshwater migration of any population of winter steelhead. 

Available information for natural populations of steelhead reveals considerable overlap in 

migration and spawn timing between populations of the same run type. Peak spawning is 

January through March, and takes place mostly high in the watersheds of the Sacramento, 

Mokelumne, and possibly the San Joaquin rivers (Figure 28). Most spawning habitat for CV 

steelhead is located downstream of dams in areas containing suitable environmental conditions 

(e.g., cold water, suitable gravels and stream flow) for spawning and incubation. We expect 

little, if any, spawning, egg incubation, and fry emergence to co-occur with thiobencarb use in 

rice production based on the current locations of known spawning areas (Figure 28). 

Furthermore, the primary thiobencarb application window (May and June) does not overlap with 

peak spawning.

As with adults, juveniles are present in the river year-round. After fry emergence, most juveniles 

rear in their natal streams at elevations above rice growing areas. Steelhead rearing takes place 

primarily in higher velocity areas of pools, although they are also abundant in glides and riffles. 

Juvenile steelhead emigrate episodically from natal streams during fall, winter, and spring high 

flows. Their relative abundance is low to moderate in rice growing areas during the thiobencarb 

application and discharge period (Table 32 and Table 31 ). Relative abundance peaks at the 

RBDD from August into November and at the Knights Landing screw traps in March. By 

March/April the bulk of the steelhead run has entered the estuary and individuals are present in 

monitoring data at Chipps Island (lower North Delta). There are many late arrivals from the 

natal spawning grounds that would migrate through rice growing areas (e.g. Knights Landing) 

well into May and perhaps early June (e.g., lower Sacramento River near Hood). However, fish 
168

monitoring data suggest that the majority of fish would have migrated through the rice growing 

areas prior to the thiobencarb application period. 

While most of the rice grown in the Central Valley is in the lower Sacramento River Basin, some 

rice is grown in the lower San Joaquin River and Mokelumne River watersheds. Fish monitoring 

on the San Joaquin (Mossdale) and Mokelumne (Woodbridge dam) indicate steelhead, in 

general, are present in these waters at very low abundance. Juveniles are present in moderate 

relative abundance in rice growing areas of both systems during times when risk of exposure to 

thiobencarb is highest. Therefore, while we expect thiobencarb exposure to juveniles rearing in 

the San Joaquin and Mokelumne Rivers, we expect the exposure to be limited given the 

relatively low abundance of both fish and rice in these areas (Table 31). Overall, drift and runoff 

into occupied steelhead rearing and migration habitats are expected infrequently. We expect low 

juvenile steelhead abundance in the Sacramento River and medium abundance in the San 

Joaquin River when thiobencarb is applied and rice holding waters are released (Table 32).
169

Figure 28. CV steelhead probable and known spawning areas in relation to 2010 rice growing 

areas.
170

Table 31. Temporal occurrence of adult and juvenile CV steelhead in the Central Valley. Darker 

shades indicate months of greatest relative abundance in each location.

Adult migration

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. River basin

San Joaquin River

Juvenile migration

Location Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Sac. River at RBDD1

Sac. River at KL2

Sac. R. at Hood3

Chipps Island3

Mossdale2

Woodbridge Dam2

Thiobencarb use/release

RBDD=Red Bluff Diversion Dam KL = Knights Landing

Relative Abundance: = High = Medium = Low

1

upstream of current rice growing areas; 2 within current rice growing areas; 3downstream of current rice growing areas.

Table 32. CV steelhead relative abundance in rice areas (Figure 25) during thiobencarb use 

periods April - July.

Life Stage None expected Low Medium High

Adult migration X

Spawning X

Early rearing X

Juvenile migration

– San Joaquin &

Mokelumne rivers

X

Juvenile migration 

– Sacramento R. X

 
171

8.2. Modeling: Estimates of exposure to thiobencarb

8.2.1. EPA exposure estimates

EPA did not utilize modeling to estimate potential exposure of listed Pacific salmonids to 

thiobencarb or other stressors of the action (EPA 2002a). However, EPA used the Rice Model 

and the AgDrift model to estimate thiobencarb exposure to California red-legged frog and Delta 

smelt (EPA 2009b). Below, we discuss the utility of these estimates for evaluating exposure to 

listed Pacific salmonids. 

8.2.2. Utility of EPA-derived EECs for defining exposure to Pacific salmonid habitats

As described in the Approach to the Assessment section, our exposure analysis begins at the 

organism (individual) level of biological organization. We consider the life stage and life 

histories of the individuals likely to be exposed. This scale of assessment is essential as adverse 

effects to individuals may result in population-level consequences, particularly for populations of 

extremely low abundance (i.e., threatened and endangered species). Characterization of impacts 

to an individual’s fitness is necessary to assess potential impacts to populations, and ultimately to 

the species. To assess risk to individuals, we must consider the range in concentrations to which 

individuals of the population may be exposed. The highest concentrations in aquatic habitats are 

typically associated with direct application to water, or off-target deposition of pesticides into 

shallow habitats in close proximity to the target site. Pacific salmonids utilize a variety of 

aquatic habitats (Table 21). All three listed salmonid species within CA’s Central Valley use 

shallow, low flow habitats at some point in their life cycle. 

EPA Rice Model. EPA estimated the peak concentrations of thiobencarb within a rice paddy to 

be 2018 μg/L using the Rice Model. This model calculates peak concentration based solely on 

the dilution volume of the paddy and an assumption of instantaneous partitioning of the pesticide 

between the rice field water and sediment. The simulation included direct application of 

thiobencarb at the labeled application rate (4 lbs a.i./A) to a 0.1 meter-deep rice field. This depth 

is consistent with the cultivation of rice. A first-order decay rate of 0.1252/day was then applied 

to the peak concentration to estimate concentrations within the rice field at various times post 

application, as well as time-weighted-average concentrations. This decay rate is a field 
172

dissipation rate and its use resulted in “double counting” of partitioning to sediment and 

potentially an underestimation of aquatic concentrations. 

Although we do not expect listed Pacific salmonids to use flooded rice fields, salmonids may be 

exposed to thiobencarb that is discharged to their habitats or manages to seep from rice fields 

into drainage canals and ultimately into salmonid habitats. Therefore, water concentration 

estimates in flooded rice fields, particularly at discharge, can be used as an upper bound for 

runoff concentrations for salmon and their designated critical habitat. FIFRA labels currently 

require treated flood waters to be held a minimum of 14 d before discharge (EPA Reg. No. 

59639-79). Some labels specify not to drain herbicide-treated fields for a minimum of 30 days 

after application (EPA Reg. No. 59639-112). Using the rice model and assuming a 5.5 day halflife for thiobencarb in water, EPA estimated discharge concentrations of 350 μg/L and 47 μg/L 

after 14 d and 30 d holding periods, respectively. The actual discharge concentrations are 

expected to vary somewhat depending on site-specific dissipation and may be either greater or 

less than predicted. Three aquatic dissipation studies were available for thiobencarb. The 

authors of these studies report half-lives of 5.8 d, > 6 d, and 8.7 d, slightly longer than the 

dissipation rate assumed by EPA modeling (Table 24). Concentrations observed in waters 

receiving discharge will be reduced to different degrees depending on flow rate and dilution 

capacity of the receiving water.

EPA AgDrift model estimates. EPA used the AgDrift model to determine distance to listed 

species habitats required to prevent adverse impacts to listed species through exposure caused by 

primary drift of the liquid formulation of thiobencarb. Model simulations assumed EPA default 

inputs and modeled drift to a 2-meter deep farm pond (EPA 2009b). The physical characteristics 

of EPA’s farm pond provide a reasonable representation of some of the habitats used by 

salmonids. NMFS used AgDrift to estimate thiobencarb concentrations in other important 

habitats, such as floodplain habitats with lower dilution capacity (NMFS exposure estimates for 

flood habitats).

EPA estimates of exposure to other action stressors. EPA did not provide model estimates for 

other ingredients, tank mixtures components, or degradates of thiobencarb. Only one 
173

thiobencarb product contains multiple a.i.s, and is formulated with propanil (EPA registration 

#71085-30). This end-use product is currently not registered for use in California. However, 

several thiobencarb labels recommend tank mixture applications with propanil or other 

herbicides to improve product efficacy. Estimates of exposure to identified metabolites and 

degradates were also absent (Table 23). The missing estimates of exposure to these stressors 

introduce substantial uncertainty to the exposure analysis. 

NMFS conclusions on utility of available EPA model estimates The Rice Model simulation

paired with field dissipation calculations provide a reasonable estimate of parent thiobencarb 

concentrations that may be discharged into habitats where listed salmonids reside. AgDrift 

modeling was useful in identifying that primary drift of spray applied thiobencarb represents a 

transportation pathway predicted to exceed EPA’s Endangered Species Levels of Concern for 

fish and their habitat (Page 111 in (EPA 2009b). NMFS provides additional modeling estimates 

below to characterize the potential range of exposure in salmonid habitats and to address 

exposure to other stressors of the action. 

8.2.3. NMFS exposure estimates for floodplain habitats

The “farm pond” scenario utilized by EPA is representative of some habitats used by listed 

salmonids. However, other habitats may be more or less susceptible to higher pesticide 

concentrations given their physical characteristics. For example, small streams and some 

floodplain habitats represent examples of habitats used by salmonids that can have a lower 

capacity to dilute pesticide inputs than the farm pond. NMFS derived estimates for a vulnerable 

floodplain habitat to capture the potential range of thiobencarb exposure in listed salmonids.

8.2.3.1. Application of thiobencarb to fields adjacent to salmonid habitat

Drift during application is a transport mechanism that can result in significant deposition of 

pesticides in aquatic habitats immediately adjacent to treated fields, including shallow floodplain 

habitats where juvenile salmonids rear and shelter. We derived exposure estimates for floodplain 

habitats using the AgDrift model to estimate downwind deposition from pesticide drift (Teske 

2001). AgDrift is a field-scale model that evaluates pesticide drift. The drift estimates derived 

represent average projected drift. Although AgDrift reasonably predicts drift, drift is highly 
174

variable and is influenced by site-specific conditions and application equipment (Bird et al. 

2002). 

Our model inputs incorporated application requirements from thiobencarb labels (Table 33). 

Some thiobencarb product labels include a set-back to aquatic habitats. For example, Bolero 8 

EC prohibits product application within 1000 feet of the endangered fat pocketbook pearly 

mussel (EPA Registration No. 59639-79). Additionally, under the Environmental Hazards 

section of the label, application of the product is not permitted south of the Intercoastal 

Waterway in Louisiana or within two miles from the shorelines of Matagorda Bay or Galveston 

Bay in Texas. Existing labels contain no setbacks to salmonid habitats. The label instructs 

applicators to apply liquid formulations of thiobencarb at up to 4 lbs of thiobencarb per acre with 

aerial or ground application equipment. Historic use data for California suggest the maximum 

labeled rate is the predominant rate used (Appendix 3). Our simulations assumed an aquatic 

habitat that was 2 m wide and of variable depths (0.1 – 2 m). These dimensions are consistent 

with some of the smaller, and potentially more vulnerable, floodplain habitats and small streams 

used by salmonids. 

Table 33. Estimated average initial thiobencarb concentrations in a floodplain habitat that is 2m

wide and of variable depths using AgDrift 2.0.05.

Application 

method Model inputs

Simulation:

Rate in lbs 

a.i./A

Habitat 

Depth (m)

Average Initial

Concentration in 

Surface Water 

( g/L)

Ground Tier 1 ground, Low ground boom

spray, ASAE fine to medium/coarse 

distribution, 50th percentile estimate

4 0.1

0.5

1

2

736

147

74

37

Aerial Tier 1 aerial spray, ASAE medium 

to coarse droplet distribution

4 0.1

0.5

1

2

1,910

382

191

96

We also used AgDrift to estimate potential exposure of riparian vegetation to drift of sprayapplied thiobencarb. The simulations evaluated application of thiobencarb at the 4 lbs a.i./A 

labeled rate. The terrestrial assessment tool in AgDrift was used to identify point-deposition at 

various distances downwind from the target treatment site (Table 34). Predicted deposition of 

thiobencarb decreased as distance from the target site increased. 
175

Table 34. AgDrift Estimates for point deposition of thiobencarb applied at the 4 lbs a.i./A rate.

Application method Distance (ft) from 

target spray area

Estimated drift of spray-applied thiobencarb

lbs a.i./A Percent of applied

Ground

Tier 1 ground, Low ground boom

spray, ASAE fine to 

medium/coarse distribution, 50th

percentile estimate

1

10

100

1000*

1.19

0.09

0.010

0.001

29.68

2.24

0.24

0.03

Aerial 

Tier 1 aerial spray, ASAE medium 

to coarse droplet distribution

1

10

100

1000*

1.91

1.21

0.22

0.02

47.76

30.34

5.61

0.55

*Simulation runs at buffer of 997 feet due to model limitation. 

8.2.4. NMFS exposure estimates for pesticide mixtures

Co-application of pesticides increases the likelihood of exposure to pesticide mixtures. Several 

thiobencarb labels recommend application with other pesticides to improve weed management 

(e.g., propanil, glyphosate, and paraquat; EPA Reg. No. 59639-79, 63588-6, 63588-14). 

Thiobencarb labels also make general recommendations for applications of other pesticides, such 

as avoiding applications of other ingredients that share a common mode of action to manage 

weed resistance (e.g., EPA Reg. No. 59639-112, 63588-14). We reviewed the available pesticide 

use data for California to determine the pesticides that are most commonly co-applied with 

thiobencarb. Co-application refers to situations where more than one active ingredient is applied 

to an agricultural field on the same day. This might be accomplished through tank mixture 

applications or separate applications to the same field (e.g. treatment of weeds on levees 

surrounding the rice). Over the last decade, thiobencarb has been co-applied with other 

pesticides 23% of the time (Appendix 4). During that timeframe there have been more than 150 

unique pesticide combinations involving the application of thiobencarb and at least one other 

pesticide applied to the same field on the same day (Appendix 4). 

It is not feasible to estimate exposure to all of the possible pesticide mixtures that are allowed to 

be used. Below we present exposure estimates associated with applications of thiobencarb, 

propanil, and copper sulfate (Table 35). According to labels, both propanil and copper sulfate 
176

products can legally be applied with thiobencarb. Propanil is the active ingredient most 

commonly recommended for co-application on thiobencarb labels. The Bolero® 8 EC label 

(EPA Reg. No. 59639-79), recommends applying thiobencarb at a rate of 2 – 2.5 lbs a.i./A in 

combination with propanil. Propanil may be applied to rice at rates of up to 6 lbs a.i./A (EPA 

Reg. No. 71085-2). Copper sulfate has been the most frequent pesticide co-applied with 

thiobencarb since 1999. Although copper sulfate products are not specifically recommended for 

tank mixture on thiobencarb product labels, these mixtures are authorized under FIFRA unless 

specifically prohibited on the product label. Copper can be applied to rice at rates up to 15 lbs 

a.i./A (EPA Reg. No. 56576-1). 

Table 35. Estimated average initial pesticide concentration in a floodplain habitat that is 2m wide 

and 0.1m deep using AgDrift 2.0.05.

Application 

method Model inputs1 Simulation:Rate in lbs a.i./A

Buffer

(feet)

Average Initial

Concentration in 

Surface Water 

( g/L)

Ground Tier 1 ground, Low ground boom

spray, ASAE fine to 

medium/coarse distribution, 50th

percentile estimate

2 (Thiobencarb)

6 (Propanil)

15 (Copper sulfate)

0

0

0

368

1,103

2,759

Aerial Tier 1 aerial spray, ASAE 

medium to coarse droplet 

distribution

2 (Thiobencarb)

6 (Propanil)

15 (Copper sulfate)

0

0

0

956

2,940

7,170

8.3. Monitoring Data: Measured Concentrations of Parent Compounds in Surface 

Waters

We evaluated data from three sources: USGS’ NAWQA database, a state database maintained by 

California Department of Pesticide Regulation, and monitoring studies that target specific 

applications of thiobencarb and monitor thiobencarb concentrations on or adjacent to the treated 

rice fields. Information provided by the two databases includes ambient monitoring data with 

sampling stations distributed across a range of land uses. They also include studies that 

investigate water quality impacts associated with specific pesticide uses, such as the use of 

pesticides for rice production. We also reviewed monitoring studies that target field-scale 
177

operations investigating thiobencarb concentrations in discharge from treated fields and in 

aquatic habitats immediately adjacent to treated fields. These studies were obtained from a 

variety of sources and include both published scientific literature and gray literature. In the 

following section we describe study design considerations for assessing the utility of monitoring 

data for evaluating exposure of pesticides to salmon.

8.3.1. Monitoring data considerations

Surface water monitoring can provide useful information regarding real-time exposure and the 

occurrence of environmental mixtures. A primary consideration in evaluating monitoring data is 

whether the study design is sufficient to address exposure in a qualitative, quantitative, or 

probabilistic manner. The available monitoring studies were conducted under a variety of 

protocols and for varying purposes. General water quality monitoring conducted in larger 

streams and rivers frequently does not capture “peak” concentrations because it is not correlated 

with applications and/or storm events following those applications and not all habitat types are 

sampled. This is one of the reasons NMFS did not use available monitoring data for 

probabilistic modeling (i.e., it likely does not contain the complete range of possible 

concentrations). 

Of the monitoring programs discussed, none were specifically designed to evaluate potential 

exposure of listed salmonids to pesticides in surface waters. Common aspects that limit the 

utility of the available monitoring data as accurate depictions of exposure within listed salmonid 

habitats include: (1) protocols were not designed to capture peak concentrations or durations of 

exposure in habitats occupied by listed species; (2) limited utility as a surrogate for other nonsampled surface waters; and (3) lack of representativeness of current and future pesticide uses 

and conditions.

Protocols not designed to capture peak exposure. None of the available monitoring studies were 

designed to evaluate peak exposure of listed salmonids in the Central Valley (or anywhere else 

salmonids reside) to thiobencarb in surface waters. The NAWQA program provides a 

considerable dataset that is useful for evaluating trends in water quality (Hirsch et al. 1988). The 

NAWQA design does not result in an unbiased representation of surface waters, which limits the 

ability to make statistical extrapolations to waters not sampled. Sampling by NAWQA and other 
178

studies was frequently not conducted in coordination with specific applications of thiobencarb at 

the field scale. Similarly, sampling was not designed with consideration of salmon distributions

or to target salmonid habitats most likely to contain the greatest concentrations of pesticides. 

Given the relatively rapid dissipation of pesticides in flowing water habitats, it is not surprising 

that pesticide concentrations from these datasets were generally much lower than those predicted 

by modeling efforts and those that monitored targeted pesticide applications at the field scale. 

Limited applicability to other locations. Pesticide runoff and drift are influenced by a variety of 

site-specific variables such as meteorological conditions, soil type, slope, and physical barriers to 

runoff and drift. Additionally, surface water variables such as volume, flow, and pH influence 

both initial concentrations and persistence of pesticides in aquatic habitats. Finally, cropping 

patterns and pesticide use have high spatial variability. Given these and other site-specific 

factors, caution should be used when extrapolating monitoring data to other sites.

Representativeness of current and future uses. Pesticide use varies annually depending on 

regulatory changes, market forces, cropping patterns, and pest pressure. Pesticide use patterns 

change annually and may result in either increases or decreases in use of pesticide products for 

specific uses. Since 1980, the acreage treated with thiobencarb has ranged between a low of 

1,973 acres in 1980, and a high of 252,506 acres in 2000 (Figure 29). Prediction of future use of 

pesticides is complicated by a number of factors including climate change that may affect 

agriculture uses and pest pressures. Additionally, regulatory changes specifying how 

thiobencarb can be used have changed over the years (e.g., changes in holding time requirements 

for thiobencarb-treated rice fields likely reduced thiobencarb concentrations in receiving waters). 

Such changes add to the difficulty of predicting future water quality conditions from historic 

monitoring data. 
179

Figure 29. Acres treated with thiobencarb in California. Data from the California Pesticide Use 

Reporting Database 1980 – 2010.

0

50,000

100,000

150,000

200,000

250,000

300,000

1980 1985 1990 1995 2000 2005 2010
180

8.3.2. USGS NAWQA Data

We obtained updated data from the USGS NAWQA database to evaluate the occurrence of 

thiobencarb in surface waters monitored in California from 1992-2011. The database query 

resulted in approximately two thousand surface water samples obtained from 74 unique locations 

in which thiobencarb was an analyte (Figure 30). 

Figure 30. Distribution of NAWQA monitoring sites that have sampled for the presence of 

thiobencarb relative to the range of threatened and endangered Pacific salmonids in California.
181

Many of the locations sampled for thiobencarb do not contain listed salmonids. Of the 74 

locations sampled, five were within the distribution of Central Valley spring-run Chinook and 

Sacramento River winter-run Chinook. Forty sites located within the distribution of the 

California Central Valley steelhead have been sampled since 1992 (Table 36). Many of the 

sample locations were outside of the rice growing region of California including the Sacramento 

Valley and the north to central San Joaquin Valley (Hill et al. 1992). 

Table 36. Number of NAWQA sample sites within the freshwater distribution of listed Pacific 

salmonids in the Central Valley of California, as determined through GIS analysis

ESU / DPS Kilometers of Stream Inhabited

Sample Sites in 

Freshwater Habitat

(Sites with Detections)

Central Valley Spring – Run Chinook 2,212.94 5 (4)

Sacramento River Winter – Run Chinook 546.84 5 (4)

California Central Valley Steelhead 4,273.66 40 (13)

Sampling effort varied considerably among the sample locations. Approximately 54% of the 

sites screened for thiobencarb were sampled five or fewer times during the span of 19 years. A 

small number of sites accounted for the majority of the data; four of the 74 sites accounted for 

more than 50% of the sampling data for thiobencarb. Eighty-six percent of the data was 

collected from 18 sites, including 4 sites that fell outside the distribution of listed salmonids. 

Overall, the sampling effort does not correspond well with the distribution of listed salmon or the 

co-occurrence of listed salmon and rice growing areas. Consequently, we do not expect the data 

set to be representative of exposure distributions for listed salmonids. 

The USGS monitoring program does not generally coordinate sampling efforts with specific 

pesticide applications or runoff events, detected concentrations are likely to be lower than actual 

peak concentrations that occur in surface waters proximate to application sites immediately 

following drift or runoff events. Summary information for quantifiable concentrations of 

thiobencarb addressed in this Opinion is presented in Table 37. Non-detects are reported as less 

than (“<”) the laboratory reporting level (LRL) for that sample. Other than total number of 

samples (n), summary statistics were calculated on samples not designated as (“<”). The LRL 

ranges reported were estimated based on “<”-qualified data. Nearly all of the concentrations that 

could be quantified were designated as “E,” meaning the concentrations were estimated. 
182

Thiobencarb was detected in 10.6% of samples screened, broadly ranging from 0.001 to 4.38 

μg/L (median, 0.01 ug/L). 

Table 37. Detections and concentrations of thiobencarb reported in the NAWQA database.

Statistic Thiobencarb

Number of Stations 74

Number of Observations 2256

Detects 241

Percent Detections 10.6%

Median (ug/L) 0.0102

Range (ug/L) 0.001-4.38

LRL (ug/L) 0.002-0.019

Year range 1992-2011

U.S. Geological Survey conducted a study during the 2002 and 2003 growing season to 

determine if changes in pesticide use on rice resulted in corresponding changes in pesticide 

concentrations in surface waters (Orlando and Kuivila 2004). This study is not included in the 

NAWQA database. Five surface water sites within the Sacramento Valley were analyzed for 

five rice pesticides, including thiobencarb. Samples were collected weekly from May through 

July. Maximum concentrations of all pesticides in 2003 were less than half of the 2002 levels. 

Higher peak pesticide concentrations and greater duration of exceedance of water quality 

protection goals observed in 2002 were attributed to differences in weather and pesticide use. 

Typically, little precipitation falls during the rice planting and growing season in the Sacramento 

Valley. However, in 2002 a significant storm in mid-May forced some early releases of field 

water including documented releases of thiobencarb that likely contributed to elevated 

concentrations in the Colusa Basin Drain and the Sacramento Slough. Additionally, spring rains 

in 2003 delayed planting and may have resulted in the observed decrease in thiobencarb use. 

The peak thiobencarb concentration observed during the study was 7.16 μg/L. Samples collected 

from the same locations in earlier years have revealed concentrations up to 170 μg/L in 1982. 

Regulations have since been implemented to reduce concentrations of rice pesticides in irrigation 

returns. Implementing holding times for pesticide-treated rice field water in the 1980s was 
183

credited in decreasing concentrations of thiobencarb and other pesticides in surface water in the 

Sacramento Valley (Orlando and Kuivila 2004).

We evaluated monitoring data from the California Department of Pesticide Regulation (CDPR), 

public database of pesticide monitoring data for surface waters in California13. Values in this 

database originate from monitoring studies conducted by CDPR, USGS, state, city and county 

water resource agencies along with some studies conducted by non-governmental or intergovernmental groups such as Deltakeeper. To avoid redundant use of these data, USGS data 

found in the CDPR database are excluded from the following data summary.

 

13 (www.cdpr.ca.gov/docs/emon/surfwtr/surfdata.htm)
184

The database provided information on more than 2,800 sampling events for thiobencarb in 

California since 1994. Sampling has been conducted at 393 unique locations (Figure 31).

Figure 31. Distribution of CDPR monitoring sites that have sampled for the presence of 

thiobencarb relative to the range of threatened and endangered Pacific salmonids in California. 

Since 1994, 28 unique locations where sampled within the range of the Central Valley spring-run 

Chinook and Sacramento River winter-run Chinook, and 66 locations were sampled within the 

range of the California Central Valley Steelhead (Table 38). Sample locations do not correspond 

well with the distribution of potential thiobencarb use sites. Rice is the only authorized use of 

thiobencarb in California and is grown almost exclusively in the Sacramento Valley and the 

central San Joaquin Valley. Detections of thiobencarb in southern California and in coastal areas 
185

of California are unexpected (Figure 31). These detections are presumed to represent 

misapplications or database entry errors since rice is not grown in those areas. 

Table 38. Number of CDPR database sample sites within the distribution of listed Pacific 

salmonids in California, as determined through GIS analysis.

ESU / DPS

Kilometers of 

Stream 

Inhabited

Sample Sites in 

Freshwater Habitat

(Sites with Detections)

Central Valley Spring – Run 

Chinook 2,212.94 28 (7)

Sacramento River Winter – Run 

Chinook 546.84 28 (7)

California Central Valley Steelhead 4,273.66 66 (10)

We conducted a simple GIS analysis to evaluate the spatial relationship between monitoring 

stations and rice fields. The thiobencarb monitoring sites were frequently located long distances 

from rice fields. A 1-km buffer was placed around each monitoring station and compared to the 

most recent crop data layer for rice (USDA-NASS Cropland Data Layer 2009). We found that 

89% of the thiobencarb monitoring sites were more than 1 km (3281 ft) from the nearest rice 

field. This suggests that in general, the dataset would not be useful for evaluating peak 

concentrations that could occur in surface waters from near-field (<1000 ft) applications of 

thiobencarb. 

Sampling effort was not equivalent among monitoring sites. More than half of the monitoring 

stations were sampled three times or less during the 17-year period. Five sample locations, 

roughly 1% of the monitoring stations, were sampled 51 to 203 times representing 22% of all the 

thiobencarb data collected. These stations and two others located within the Sacramento Valley 

rice-growing Region accounted for 95% of the thiobencarb detections (Figure 32). This region is 

within the range of the Central Valley Spring-Run Chinook, the Sacramento River Winter-run 

Chinook, and the California Central Valley Steelhead.
186

Figure 32. The location of seven sample sites accounting for 95% of the detections in thiobencarb

in the CDPR monitoring database.
187

The contents of the database for thiobencarb are summarized in Table 39 while individual 

programs and studies are described below in the text. The CDPR requires a formal QA/QC 

protocol for data submitted or does a separate QA/QC review, thus only data subject to 

appropriate QA/QC procedures are included in the surface water database. Unlike the USGS 

NAWQA data set, the CDPR database may contain whole water samples as well as filtered 

samples. If whole water concentrations are reported for compounds that sorb significantly to the 

particulate phase, concentrations would appear higher than in a filtered sample, which represents 

only the dissolved phase. The database, last updated in June 2011, consists of approximately 

378,000 data records. Each record reports a specific sampling site, date, and analyte. In this 

database, detections below the LRL are reported as 0 μg/L. Our summary statistics for the 

datasets were calculated on samples with values above the LRL. The number of observations is 

indicative of monitoring intensity rather than actual occurrence in surface waters. 

Table 39. Detections and concentrations of thiobencarb from the CDPR database

Statistic Thiobencarb

Number of Studies 25

Number of Stations 393

Number of Observations 2836

Detects 180

Percent Detections 6%

Median (ug/L) 1.275

Range (ug/L) 0.008-16.55

LRL (ug/L) 0.005-2

Year range 1994-2011

The CDPR database includes 2836 thiobencarb observations from 25 studies monitoring a total 

of 393 stations. Thiobencarb was detected in approximately six percent of samples screened for 

the pesticide (median 1.28 μg/L, range 0.008-16.55 μg/L, LRL=0.05-2 μg/L). Some studies 

reported in the database are annual monitoring results for multiyear monitoring programs, while 

other studies encompass several years. The Sacramento River Watershed Program was a single 

year monitoring effort that monitored for thiobencarb in 2002. Monitoring in this program 

screened for thiobencarb at five stations and detected thiobencarb in one of the five samples 
188

collected (2.1 ug/L, LRL=0.5 ug/L). The Central Valley Regional Water Quality Control Board 

included thiobencarb among the analytes screened for during its investigation into the sources 

and concentrations of diazinon in the Sacramento watershed during the 1994 spray season. 

Thiobencarb was detected in 35 of 64 samples taken from eight stations during this study 

(median 0.012 μg/L, range 0.008-0.042 μg/L, LRL=0.008 μg/L). The California State Water 

Resources Control Board Surface Water Ambient Monitoring Program (SWAMP) studies cover 

the years 2001 to 2008 and include 943 observations from 271 stations monitored for 

thiobencarb. Thiobencarb was detected nine times with a median concentration of 0.93 ug/L 

(range=0.109-11.8 μg/L, LRL=0.005-0.2 μg/L). The highest value at 11.8 μg/L was detected in 

a sample collected in May of 2006 from the New River at Boundary. The highest thiobencarb 

concentrations are reported in two monitoring studies targeting areas of high pesticide use which 

were sampled multiple times over the growing season or during peak irrigation. Such studies are 

more likely to capture peak or near peak pesticide concentrations. These studies are discussed in 

more detail in the following section on targeted monitoring. 

The CDPR database includes data from two notable studies. The Central Valley Regional Water 

Quality Control Board monitoring project investigated water quality between 2004 and 2009 in 

agricultural drains of the Central Valley (hereafter "Irrigated Lands Study"), and CDPR 

monitored for pesticides used in rice cultivation between 1995 and 2002 (hereafter “Rice 

Pesticide Monitoring Program”). These two studies account for a majority of the thiobencarb 

data in the CDPR database; cumulatively they represent 64% of the observations that have 

screened for the presence of thiobencarb and 84% of thiobencarb detections. Although these 

studies do not target thiobencarb applications at the “field scale” or monitor exposure in 

salmonids, they are designed to evaluate surface water concentrations corresponding to regional 

agricultural practices. 

Irrigated Lands Study. The irrigated lands study monitoring stations are located near agricultural 

drainages flowing into creeks or rivers in catchments dominated by return flow from mixed row 

crops and/or alfalfa, or areas where the primary land use was rice culture. The study design 

focused sampling efforts on periods of peak irrigation, especially the first major irrigation of the 

season. The purpose of the study was to evaluate the magnitude and extent of water quality 
189

problems in waters that receive agricultural drainage. The majority of detects of thiobencarb 

occurred in May and June which corresponds to thiobencarb’s peak use (Table 40). Among 

1419 samples collected from 114 stations, thiobencarb occurred in 72 samples from 28 stations 

at concentrations as high as 7.58 ug/L (LRL = 0.05-2 ug/L). The high concentration was 

detected within the Natomas Central Mutual Water District in the Sacramento Valley. At the 

time of sample collection, thiobencarb could legally be discharged into irrigation drains after a 

minimum six-day hold period because the district maintained a “closed system.” The district 

was required to hold the water for the remainder of the 30-day hold period prior to discharge to 

creeks and rivers. 

Table 40. Summary of Irrigated Lands Study data from the Central Valley Regional Water Quality 

Control Board (2004-2006).

Statistic Thiobencarb

Number of Stations 114

Number of Observations 1419

Detects 72

Percent Detections

5%

Median (ug/L) 0.325

Range (ug/L) 0.055-7.58

LRL (ug/L) 0.05-2

Rice Pesticide Monitoring Program. CDPR implemented the Rice Pesticides Program (RPP) in 

1983 to reduce thiobencarb and molinate discharges into surface waters. In 1990, the Central 

Valley Regional Water Quality Control Board and CDPR established performance goals for 

thiobencarb and several other pesticides used in rice production. The 1.5 μg/L performance goal 

for thiobencarb was established to help meet California’s Maximum Concentration Limit of 1 

μg/L in drinking water. This secondary MCL was established based on a threshold for impaired 

taste to drinking water. Data from the RPP are used to verify compliance with this and other 

performance goals and implement adaptive management (Bennett et al. 1998, Moran 2012).

Some of the sample locations have varied over the years. Recent monitoring evaluated by the 

RPP includes samples collected from five primary monitoring stations in the Sacramento Valley 
190

(Figure 33). Additionally, samples collected at two water intake locations on the Sacramento 

River are provided by water treatment facilities. 

Figure 33. Sacramento Valley RPP monitoring locations (source: Moran 2012 modification of 

figure from CRC 2004 RPP Report) 
191

CDPR prepared reports for the RPP monitoring program between 1995-2002. We searched the 

CDPR database for three stations that have been systematically monitored for thiobencarb since 

1995. Water samples were collected yearly during the rice pesticide application period from 

three sites between 1995 and 2002 (LOQ 0.5-1 μg/L): Butte Slough at Lower Pass Road, Colusa 

Basin Drain #5, and from the Sacramento River at Village Marina/Crawdads Cantina. These 

areas receive irrigation returns from rice-growing areas (referenced in their metadata as studies 

17, 30, 34, 40, 53, 67, 73 and 75). These stations were also sampled in later years by the 

California Rice Commission, and additional sample locations have been added to the Rice 

Pesticide Monitoring Program. Prior to 2002, detections occurred most frequently and at the 

highest concentrations at the Colusa Basin Drain site, with a 73 to 100% detection rate over the 

years sampled (Table 41). A seasonal increase in concentrations occurred each May followed by 

a decline by late June or early July. This seasonal increase corresponds with the period of 

thiobencarb application in California (Figure 24). Over the eight years of monitoring, 

thiobencarb was detected in 13-71% of samples from the Butte Slough site (peak concentration 

range 1.3-3.6 ug/L) and in 0-23% of samples from the Sacramento River site (peak concentration 

range 0.5-0.6 ug/L). Thiobencarb detections at the Butte Slough site generally occurred from 

mid-May to early June each year, while four of the six detections at the Sacramento River Site 

occurred in samples collected between May 21 and May 30 of 2002. 
192

Table 41. Annual monitoring results for thiobencarb at stations associated with rice cultivation 

(1995-2002; LRL=0.5 ug/L) 

1995 1996 1997 1998 1999 2000 2001 2002

Sacramento 

River

# 

samples 16 17 17 17 8 7 12 26

% detects 0% 0% 0% 0% 13% 0% 8% 15%

Median nd nd nd nd 0.5 nd 0.5 0.52

Min

Max

0.29

0.6

Colusa Basin 

Drain no.5

# 

samples 22 23 24 23 15 14 23 24

% detects 73% 70% 75% 61% 93% 100% 78% 63%

Median 0.87 4.0 1.9 2.35 3.95 4.75 1.85 3.34

Min

Max

0.25

3.67

1

16.55

0.25

12.3

1.1

11

0.7

11.8

1.7

10.7

0.5

5.9

0.28

7.58

Butte Slough

# 

samples 15 19 16 17 8 7 12 23

% detects 13% 37% 50% 35% 50% 71% 25% 30%

Median 1.2 1.2 1.25 0.6 0.9 1.1 1.5 2

Min

Max

1.1

1.3

0.7

2.0

0.5

6

0.6

1.9

0.6

4.1

1

1.6

0.8

2.6

0.7

3.62

“nd” indicates median not determined because sample not detected above LRL

median and minimum values reflect statistics on sample detects

Monitoring by the California Rice Commission since 2003 indicates a pattern of lower detection 

frequencies and lower maximum concentrations compared to earlier monitoring (1995-2002). 

The peak concentration observed since 2003 was 3.6 μg/L in the Colusa Basin Drain no. 5 in 

2004 (Table 42). Thiobencarb has generally not been detected at the primarly monitoring station 

on the Sacramento River, and detections at drinking water intakes on the Sacramento River have 

been ≤ 0.68 μg/L. These sample stations are on the mainstem of the river and not in the 

immediate vicinity of rice fields (Figure 33). 

Table 42. Annual monitoring results for thiobencarb at stations associated with rice cultivation 

(2003-2011; Detection Limit 0.5 or 0.1 ug/L)

2003 2004 2005 2006 2007 2008 2009 2010 2011

Sacramento 

River

# 

samples 23 14 17 15 16 16 16 16 15

% 

detects 0% 0% 0% 0% 0% 6% 0% 0% 0%

Median nd nd nd nd nd 0.62 nd nd nd
193

2003 2004 2005 2006 2007 2008 2009 2010 2011

Min na

Max

Butte 

Slough

# 

samples 17 15 17 14 17 16 16 15 16

% 

detects 12% 0% 12% 28% 0% 13% 6% 7% 13%

Median 0.56 nd 0.26 0.61 nd 1.56 0.50 0.80 0.55

Min 0.51 0.22 0.56 1.2 na na 0.5

Max 0.6 0.30 0.70 1.99 0.6

Sacramento 

Slough

# 

samples 17 15 17 13 15 15 15 15 16

% 

detects 0% 27% 12% 0% 0% 0% 0% 0% 0%

Median nd 0.75 0.47 nd nd nd nd nd nd

Min 0.5 0.33

Max 0.9 0.6

Colusa 

Basin Drain 

No. 1 

(South end 

of Ag Drain)

# 

samples 23 15 15 15 16 15 15 15 14

% 

detects 30% 20% 35% 20% 33% 53% 33% 40% 29%

Median 0.9 1.1 0.59 0.86 0.53 0.81 1.75 65 0.9

Min 0.5 0.7 0.20 0.60 0.50 0.56 1.24 0.50 0.6

Max 2.3 1.6 0.67 0.90 0.76 1.80 1.84 1.8 1.2

Colusa 

Basin Drain 

No. 5 

(North end 

of Ag Drain)

# 

samples 18 16 18 16 15 15 15 15 15

% 

detects 28% 13% 17% 6% 7% 27% 40% 33% 13%

Median 0.62 2.56 0.24 0.97 0.54 0.83 0.82 0.85 1.25

Min 0.5 0.8 0.14 na na 0.75 0.54 0.61 1.1

Max 1.3 3.6 0.32 1.02 1.24 1.5 1.4

“nd” indicates median not determined because sample not detected above method detection limit

“na” indicates range not applicable because only one sample detected above method detection limit

median and minimum values reflect statistics on sample detects
194

A recent summary evaluated trends in thiobencarb detections in the Sacramento Valley (Moran 

2012). The highest concentrations of thiobencarb were monitored in the 1980s, and between 

1992 and 2002. The lowest levels occurred between 2003 and 2007. Increased concentrations of 

thiobencarb were noted starting in 2008. Additionally, while there were no documented incidents 

of exceeding the 1.5 μg/L performance goal between 2004 and 2007, the performance goal was 

exceeded on six occasions between 2008 and 2010. These exceedances were monitored at three 

sites (CBD1, CBD5, and BS1; Figure 33) at concentrations ranging from 1.5 – 2.0 μg/L. A 

number of factors were evaluated and considered unlikely sources of the increase including 

changes in thiobencarb use, management practices, rain events, river flows, wind, emergency 

releases, and violoations of water hold or seepage requirements. Moran (2012) concluded that 

transition from the 15G formulation of thiobencarb to the Bolero Ultramax formulation may 

have been the major factor in increased thiobencarb concentrations observed in the Sacramento 

Valley from 2008-2011. 

All federal thiobencarb labels currently require water from thiobencarb treated fields be held a 

minimum of 14 days prior to release. Thiobencarb products labeled for California require 

holding periods of granular formulations of up to 30 days. Additionally California state 

regulatory bodies have further instituted, 19-d holding periods for liquid formulation i.e., 

Abolish®, to reduce surface water concentrations, meet performance goals, and permit 

requirements (Kelley 2000). Monitoring results suggest that these holding requirements have 

resulted in lower concentrations of thiobencarb in surface waters. The monitoring program has 

also documented concentrations of thiobencarb exceeding the 1.5 μg/L target concentration at 

some locations prior to the onset of water release. In some cases, concentrations in surface 

waters near treated fields have remained elevated above 1.5 μg/L for weeks. These occurrences 

have been attributed to a number of factors including thiobencarb drift, unplanned as well as 

planned emergency releases, and seepage (Kelley 2000).

Emergency releases of water from treated fields may be authorized by the state on a conditional 

basis (e.g., to avoid crop damage associated with high salinity levels or due to heavy rainfall). 
195

Seepage refers to the lateral movement of thiobencarb from the rice field to adjacent waters prior 

to release through structurally compromised weir boxes and levees. In California, County 

Agricultural Commissioners serve as local enforcement for pesticide use requirements and 

conduct a variety of monitoring activities including monitoring for seepage and compliance with 

pesticide hold requirements. Recent monitoring indicates the occurrence of seepage (Table 43), 

emergency releases of water from fields treated with thiobencarb (Table 44), and violations of 

label requirements have been relatively infrequent. However, where seepage and other early 

releases of water from thiobencarb treated sites do occur, they are likely to contribute to greater 

surface water concentrations of thiobencarb and may result in greater exposure to listed 

salmonids. 

Table 43. Results of County Agricultural Commissioner's monitoring of seepage associated with 

thiobencarb applications and compliance with state regulations (2006-2010).

Year Seepage Inspections

Thiobencarb

Inspections

Seepage 

<5 gpm

Seepage

>5 gpm

Enforcement 

actions

2006 929 29 5 1

2007 839 8 3 2

2008 250 25 2 3

2009 883 30 0 0

2010 964 46 4 3

Total 3865 138 14 9

Average 773 27.6 2.8 1.8
196

Table 44. Results of County Agricultural Commissioner's monitoring of compliance with state 

regulations associated with the use of thiobencarb (2006-2010).

Year Emergency Releases Water Hold Application

Inquiries Releases Inspections Civil 

Penalties

Inspections Compliance 

Violations

2006 0 0 932 4 37 3 

2007 4 1 839 2 30 0

2008 1 1 750 2 27 3

2009 2 1 1,012 0 22 0

2010 2 0 1,036 3 28 0

Total 9 3 4,569 11 144 6

Average 1.8 0.6 913.8 2.2 28.8 1.2

8.3.3. Targeted monitoring of thiobencarb during applications

Below we discuss studies that evaluated specific applications of thiobencarb to determine aquatic 

concentrations associated with edge of field drift, runoff, and discharge of thiobencarb.

8.3.3.1. Lauck 1979 – MRID 25179

A series of field studies characterized off-site movement of thiobencarb from applications to rice 

paddies near estuarine systems in Texas (Lauck 1979). The applications were made to 1900 

acres of dry seeded rice in five rice growing areas in Brazoria County, Texas. Surface water, 

drift, and off-target vegetation were monitored to characterize off-site movement of thiobencarb 

following applications. Concentrations of thiobencarb were monitored from planned discharge 

(flushing) and from unanticipated releases associated with runoff from rain events. Thiobencarb 

was transported off-site by drift and runoff at all five study areas (Table 45). Concentrations in 

runoff were highest during unplanned releases immediately following applications. For 

example, a concentration of 8.9 mg/L was detected 2 hrs after application during a emergency 

release of rice water to a drainage ditch due to a heavy rain event. We consider this value as a 

worst case scenario. We note that current EPA labels do not authorize such a release. Although 

it may be possible to achieve comparable runoff from treated fields due to seepage and early 

releases, these events occur on an infrequent basis (Table 43 and Table 44). Therefore, exposure 
197

of salmonids and their designated critical habitat to concentrations of this magnitude are highly 

unlikely, occurring only rarely. 

Many of the runoff releases occurred within a few days of application and frequently attained 

concentrations above 200 ug/L. The longer the holding time before release of rice water, the 

lower the concentrations of thiobencarb. As expected, after each release, subsequent releases 

showed lower thiobencarb concentrations. Thiobencarb was frequently detected in bayous 

(receiving waters) ranging from non-detectable to 385 ug/L. The maxima detected at each site 

resulted from releases that occurred on the same day of application or within a few days after 

application. In Area 2, two sites showed concentrations of 83 and 64 ug/L 14 days after 

application whereas in Area 3, 7.0 and 140 ug/L detections were measured 13 days after 

application. Thiobencarb concentrations detected in bayous, 26 to 30 days post application from 

all areas ranged from non-detectable to 6 ug/L. 

Thiobencarb drifted off-site at each area and appeared dependent on wind speed (wind direction 

was not reported). Three areas were monitored for off-site drift (Table 45). As expected, the 

greatest drift occurred at distances nearest the center of the swath. Thiobencarb was detected at 

distances of up to 800 m, however most drift occurred within 50 m. 

Non-target vegetation was collected from one of the study areas before and after applications. 

Cattail and turtle grass were the predominant vegetation within drainage ditches and emergent 

aquatic vegetation, respectively. Vegetation was collected at 12.5, 50, 100, 200, and 400 meters 

downwind from application. Grasses and broadleaf plants were collected including a rush, 

nutgrass, and hedge parsley. The vegetation was not noticeably affected by drift according to the 

study results. All species of levee vegetation were directly exposed to thiobencarb at 4.0 

lbs/acre. Canary reed grass was the only plant that showed symptoms of injury (i.e., size and 

maturity were reduced by 55% relative to untreated plants). According to the study, “other 

resident winter annuals, perennials, and woody plants within levees and bayous showed no 

symptoms of injury.” The study did not report when the plants were assessed for damage, nor 

what type of effects were measured, how they were measured, or any raw data relating to plant 

responses. 
198

Table 45. Texas field studies that evaluated offsite drift and runoff with thiobencarb in three tidally 

influenced bayous (Lauck 1979) 

Bolero 8EC: 

84.6% 

thiobencarb

Area 1

(592 acres)

Area 2

(451 acres)

Area 3

(435 acres)

Area 4

(894)

Area 5

(206 acres)

Application date 3/25/1979 3/20/1979 3/20/1979 3/20; 3/24; 

4/08/1979

4/08/1979

Lbs/ Acre 

applied 

(nominal)

4 lbs/acre 4 lbs/acre 4 lbs/acre 4 lbs/acre 4 lbs/acre

Application type aerial aerial aerial aerial Aerial

Number of 

applications

1 1 1 1 1

Drainage ditch 

distance 

between rice 

field and bayous 

0.3 - 4 miles 200 - 2300 ft 750 -1650 ft 1650 - 2400 ft Not reported

Weather 

conditions day 

of application

- Air temp = 

16.1◦C

- 0-2 mph

-Air temp = 

24.7◦C

- 5-10 mph

-Air temp = 

29.2◦C

- 4-8.5 mph

na na

Spray drift card 

distance from 

application site 

(meters)

12, 25, 50, 

100, 200, 400, 

500, 700, 800, 

1200

12, 25, 50, 100, 

200, 400, 500, 

700, 800, 1200

12, 25, 50, 

100, 200, 400, 

500, 700, 800, 

1200

na na

Drift on 

application day 

(percent of 

applied)

39 ft = 5.71

82 ft = 2.51

164 ft = 0.44

328 ft = 0.21

656 ft = 0.32

1312 ft = 0.05

1640 ft = nd 

2297 ft = nd

2625 ft = nd

3937 ft = nd

39 ft = 7.06

82 ft = 2.30

164 ft = 1.48

328 ft = 0.29

656 ft = 0.05

1312 ft = 0.02

1640 ft = 0.01

2297 ft = 0.06

2625 ft = nd

3937 ft = nd

39 ft = 15.02

82 ft = 2.13

164 ft = 0.45

328 ft = 0.07

656 ft = 0.03

1312 ft = 0.02

1640 ft = nd

2297 ft = nd

2625 ft = 0.01

3937 ft= nd

na na

Surface water 

monitoring of 

thiobencarb 

Field outlets, 

drainage ditch:

Maximum = 

0.142 – 0.595 

mg/L

Receiving 

water = <2.0 – 

40 ug/L

Field outlets, 

drainage ditch:

Maximum = 

0.69 - 8.9 mg/L 

application day

Receiving 

water = 

<2.0 – 83 ug/L

Field outlets, 

drainage ditch:

Maximum = 

0.028 -0.405 

mg/L 

Receiving 

water = 

<2.0 – 385 

ug/L

Field outlets, 

drainage ditch:

Maximum = 

0.010 -0.313 

mg/L 

Receiving 

water = 

<2.0 – 55 ug/L

Field outlets, 

drainage ditch:

Maximum = 

0.271 – 0.287 

mg/L 

Receiving 

water = 

<2.0 – 64 ug/L

na denotes not applicable, spray drift not monitored

nd denotes not detected
199

8.3.3.2. Beiver 1994 – MRID 434040-03

Two studies evaluated off-site transport of thiobencarb and its degradate, thiobencarbsulfoxide, 

following a single aerial application of Bolero 8EC to rice fields, one in Texas and one in 

Arkansas. Study results are presented in MRID 434040-03 and summarized in EPA RED, CRLF 

BE, and salmonid BE. Table 46 provides information on the two field studies including results 

of drift and surface water monitoring. 

Drift was measured on application day with spray cards located along the fields perimeters. In 

the Arkansas study the greatest offsite drift occurred (3.9 -14.5% of the application amount) west 

of the rice field. A SE wind of 6 mph was measured during the application. Drift was less 

extensive east and south of the field. Drift cards alongside the bayou collected no detectable 

thiobencarb (<0.018 μg/L). In the Texas study, drift was less extensive, with a maximum of 

7.4%, likely due to reduced winds (3-4 mph). Four drift cards were placed alongside Bernard 

Creek, two of which resulted in 0.2% of the applied thiobencarb detected (i.e., 0.008 lbs 

thiobencarb per acre). The results supported the hypothesis that drift indeed occurs following 

aerial application of thiobencarb to rice and can be a significant pathway for offsite 

contamination. Wind speed and direction, application rate, distance to aquatic habitats, droplet 

size, and release height are all key determinants of drift. 

Runoff was monitored in both studies for thiobencarb and thiobencarbsulfoxide. Results show 

that thiobencarb moved offsite during rain events and planned flushing events. Highest 

concentrations occurred during the first runoff events in both studies (Arkansas 380 μg/L and 

Texas 2300 ug/L). The three day average of thiobencarb in the Texas drainage ditch was 1120 

μg/L (sampled multiple times per day), while the four day average in the Arkansas drainage 

ditch was 277 μg/L. Receiving waters (Bayou Bartholomew [AR] and West Bernard Creek 

[TX]), where aquatic life occurred, were also monitored in both experiments. No toxicity 

experiments were conducted with aquatic animals or plants in either experiment. Receiving 

water concentrations reached 42 μg/L in the Texas study and remained less than 1.0 μg/L in the 

Arkansas study, with the exception of one site which detected thiobencarb on day 2 and 3 post 
200

treatment at a concentration of 260 μg/L. Thiobencarbsulfoxide was consistently detected in 

both studies with maximum concentrations in the 60 μg/L range. 

In aggregate, the results identify drift and runoff as contributing pathways to aquatic 

contamination from field applications. Additionally, the data show that early release of flooded 

rice fields, whether due to rain or flushing practices, can contribute highly toxic concentrations to 

off-site aquatic habitats. Elevated concentrations within drainage systems and proximate aquatic 

habitats remain for multiple days. Current federal labels, and more restrictive labels on 

thiobencarb products sold in California, require that water from thiobencarb treated fields be held 

14 days or longer prior to release to aquatic habitats. Therefore, concentrations observed in 

receiving waters <14 days after application may not reflect likely exposure of listed salmonids 

and their designated critical habitat. Peak concentrations observed in receiving waters 14 days or 

more after applications were <30 ug/L.

Table 46. Registrant submitted field studies in Arkansas and Texas that evaluated off-site drift and 

runoff with thiobencarb and thiobencarbsulfoxide (MRID 434040-03)

Bolero 8EC: 84.6% thiobencarb Bayou Bartholomew, AR

April 30– June 27 1993

58 day experiment

East Bernard, TX

May 12- June 26 1993

45 day experiment

Application date May 1 May 14

Lbs/ Acre applied (nominal) 4 lbs/acre 4 lbs/acre

Application type aerial Aerial

Number of applications 1 1

Drainage ditch distance between 

rice field and receiving water 

0.5 miles 300 ft

Weather conditions day of 

application

-Air temp = 21◦C

-SE wind 9.7 km/hr (6 mph)

-100% cloud cover with light rain

-Air temp = 26.7◦C

-N-NE and N-NW wind 4.8-6.4 

km/hr (3-4 mph)

-Clear skies

Spray drift card distance from 

application site

100 ft, 12 cards deployed and 4 

alongside receiving water

100 ft, 11 cards deployed and 4 

alongside receiving water

Drift on application day (percent 

applied), drift cards placed 

(distance from field not reported)

Maximum= 14.5%

Range = <0.018- 14.5%

North side of field = 5.3-7.8%

West side of field = 3.9-14.5%

East side of field = none detected

South side of field = <0.0018-

0.032%

Receiving water= <0.018 ug/L

Maximum= 7.4%

Range = <0.017 – 7.4%

North side of field=no drift cards 

deployed

West side of field=6.8-7.4%

East side of field= <0.6% (pilot 

did not spray on East side border

South side of field= 0.89-5.2%

Receiving water= <0.017-0.2%

Rain and flushing events causing 

runoff of thiobencarb

Percent indicates % of runoff 

compared to total water volume

Day 1-4, 13% runoff

Day 10-13, 33% runoff

Day 25-28, 25% runoff

Day 29-30, 34% runoff

Day 3-5, 30% runoff

Day 9-11, 36% runoff

Day 12-16, 17% runoff
201

Bolero 8EC: 84.6% thiobencarb Bayou Bartholomew, AR

April 30– June 27 1993

58 day experiment

East Bernard, TX

May 12- June 26 1993

45 day experiment

Surface water monitoring 

Thiobencarb (ug/L)

Drainage ditch:

Maximum = 380 ug/L

4 day mean = 277 ug/l

1 day PTI: 330-380 ug/L

13 days PTI: 10 ug/L

28 days PTI: <1.2 ug/L

Receiving water = <0.05-260

ug/L

-Max (260 ug/L) on 3 day PTI

-26 ug/L on day 42 PTI

Drainage ditch:

Maximum = 2300 ug/L

3 day mean = 1120 ug/L 

3 days PTI: 1600-2300 ug/L

14 days PTI: 1.6-3.6 ug/L

30 days PTI: <0.8 ug/L

Receiving water = <0.5-42 ug/L

-Max (42 ug/L) on 12 days PTI

-29 ug/L on day 14 PTI

Surface water monitoring 

Thiobencarbsulfoxide (ug/L) 

Drainage ditch:

Maximum = 61 ug/L

4 day mean = 45 ug/L

Receiving water = <1 ug/L

Drainage ditch:

Maximum = 67 ug/L

3 day mean = 49 ug/L 

Receiving water = <0.5-5.1ug/L

8.3.3.3. Thiobencarbmonitoring published in the open literature

We located several studies from Japan on the use of thiobencarb in rice. In the first study, 

transport and partitioning of thiobencarb entering the inner Sugao marsh and mouth of the 

Iinuma River was monitored over the course of normal water discharge from surrounding rice 

paddies (Ibaraki Prefecture, Japan). Thiobencarb was detected in samples collected between late 

April to early May with suspended solids accounting for 3.5% of the estimated load of 

thiobencarb to the marsh, and fine suspended solids accounting for 59% of the total suspended 

solid fraction (Kawakami et al. 2006). 

Dissipation and runoff of pesticides used in rice cultivation was monitored during the 2003 

growing season at a rice paddy located in Higashi, Hiroshima City, Japan. Peak concentrations 

of thiobencarb (417.55-461.89 ug/L) were reached in paddy fields 3 days after application at a 

rate of 1.3 lbs a.i./A. Thiobencarb was detected downstream from the paddy up to 3 days after 

application at 0.02-0.08% of the applied concentration. Thiobencarb concentrations decreased to 

trace levels within one month after application (Parveen et al. 2005).

The concentrations, loadings and losses of pesticides used in rice fields were investigated 

between 1993 and 1997 in the Seta River, which is the only natural outlet of Lake Biwa, Japan. 
202

The Lake Biwa catchment area is 20% paddy fields with six rivers flowing into the lake. The 

detection of thiobencarb in the influent rivers began in the middle of May and declined to 

detection limits after July. Thiobencarb was detected in 27% of the samples collected between 

late May and early August at concentrations ranging from 0.026-0.057 μg/L (Sudo et al. 2002). 

A study evaluating continuous versus intermittent irrigation schemes as rice field water 

management practices indicated that the intermittent irrigation results in less losses of pesticide 

from rice paddies. The day after application at a rate of 1.3 lbs a.i./A, thiobencarb concentrations 

peaked in paddy drainage waters at 595 μg/L and rapidly declined over a period of two weeks 

(Watanabe et al. 2007). 

8.3.4. Exposure to Other Action Stressors

Stressors of the action also include the metabolites and degradates of the a.i.s, other active and 

inert ingredients included in their product formulations, and tank mixtures and adjuvants 

authorized on their product labels. Below we summarize information presented in the BEs and 

provide additional information to characterize exposure to these stressors. 

8.3.4.1. Metabolites and degradates of thiobencarb

EPA documents identified several degradates of thiobencarb (see previous section Summary of 

Chemical Fate of Thiobencarb). However, estimates quantifying potential exposure of listed 

salmonids and their habitat to these transformation products were limited and remain a 

considerable source of uncertainty. In general, failure to consider exposure to these breakdown 

products increases the likelihood that risk is underestimated.

8.3.4.2. Other ingredients in formulated products

NMFS reviewed all of the active labels of thiobencarb and found only one contained multiple 

a.i.s (EPA Reg. No. 71085-30). Active enduse products contain 10 – 84 % thiobencarb. Valent 

provided NMFS with Confidential Statements of Formula for Bolero Ultramax Herbicide (EPA 

Reg. No. 59639-79) and Bolero 15G (EPA Reg. No. 59639-112), the two thiobencarb products

currently marketed in California. The specific chemical constituents in Bolero Ultramax 

Herbicide and other thiobencarb end-use products have not been communicated to us (Table 47). 
203

Table 47. Examples of thiobencarb product ingredients.

EPA Product 

Registration 

Number

Active Ingredients % Other Ingredients %

59639-79* Thiobencarb 84 16

59639-80 Thiobencarb 10 90

59639-112* Thiobencarb 15 85

63588-4 Thiobencarb 97.4 2.6

63588-6 Thiobencarb 84 16

63588-14 Thiobencarb 15 85

71085-30 Propanil 35, Thiobencarb 31 34

CA-930003 Thiobencarb 84 16

*Currently registered in California

Nonylphenol (NP) and nonylphenol polyethoxylates are “other ingredients” that may be part of a 

pesticide product formulation and are common adjuvant ingredients added during pesticide 

applications. NP and nonylphenol polyethoxylates are also ingredients in detergents, cosmetics, 

and other industrial products and are a common wastewater contaminant from industrial and 

municipal sources (Koplin et al. 2002). NP has been linked to endocrine disrupting effects in 

aquatic systems (Arsenault et al. 2004, Brown et al. 1999, Brown et al. 2003, Brown et al. 2005, 

Madsen et al. 2004, Schoenfuss et al. 2008a). A national survey of streams found that NP was 

among the most ubiquitous organic wastewater contaminants in the U.S., detected in more than 

50% of the samples tested (Koplin et al. 2002). 

Table 48. Detection and concentrations of nonionic detergent degradates in streams of the U.S.

(Koplin et al 2002).

Chemical Frequency 

Detected Maximum ( g/L) Median

( g/L)

4-nonylphenol 50.6 40 0.8

4-nonylphenol monoethoxylate 45.9 20 1

4-nonylphenol diethoxylate 36.5 9 1

4-octylphenol monoethoxylate 43.5 2 0.2

4-octylphenol diethoxylate 23.5 1 0.1

We are uncertain to what degree NP and NP-ethoxylates may or may not occur in pesticide 

products that contain thiobencarb and/or are added prior to application. Inert ingredients are 

often not specified on product labels. Additionally, NP and NP-ethoxylates represent a very 
204

small portion of the more than 4,000 inert ingredients that EPA permits for use in pesticide 

formulations (Koplin et al. 2002). Many of these inerts are known to be hazardous in their own 

right (e.g., xylene is a neurotoxin and coal tar is a known carcinogen). Several permitted inerts 

are also registered a.i.s (e.g., copper, zinc, chloropicrin, chlorothalonil). Inerts can be more than 

50% of the mass of pesticide products, and millions of pounds of these products are applied to 

the landscape each year (Koplin et al. 2002). This equates to large contaminant loads of inerts 

that may adversely affect salmon or their habitat. Uncertainty regarding exposure to these 

ingredients will be qualitatively incorporated into our analysis. 

8.3.4.3. Tank Mixtures

Several pesticide labels authorize the co-application of other pesticide products and other 

materials in tank mixes, thereby increasing the likelihood of exposure to multiple chemical 

stressors (see previous section NMFS exposure estimates for pesticide mixtures). In some cases 

specific application of other pesticide products or adjuvants are recommended. In all cases, tank 

mixtures are authorized unless specifically prohibited on the product label. These ingredients 

and the other inert ingredients in these products are considered part of the action because they are 

authorized by EPA’s approval of the FIFRA label. Exposure to, and risk associated with,

potential ingredients in tank mixtures were not addressed in EPA’s BEs and remain a significant 

source of uncertainty. 

8.3.4.4. Environmental Mixtures

As described in the Approach to the Assessment, we analyze the status of listed species in 

conjunction with the Environmental Baseline in evaluating the likelihood that action stressors 

will reduce the viability of populations of listed salmonids. This involves considering 

interactions between the stressors of the action and the Environmental Baseline. For example, 

we consider that listed salmonids may be exposed to the wide array of chemical stressors that 

occur in the various marine, estuarine, and freshwater habitats they occupy throughout their life 

cycle. Exposure to multiple pesticide ingredients most likely occurs in freshwater habitats and 

nearshore environments adjacent to areas where pesticides are used. As of 1997, about 900 a.i.s 

were registered in the U.S. for use in more than 20,000 different pesticide products (Aspelin and 

Grube 1999). Typically 10 to 20 new a.i.s are registered each year (Aspelin and Grube 1999). In 
205

a typical year in the U.S., pesticides are applied at a rate of approximately five billion pounds of 

a.i. per year (Kiely et al. 2004). Pesticide contamination in the nation’s freshwater habitats is 

ubiquitous, and pesticides usually occur in the environment as mixtures (Gilliom et al. 2006a). 

Gilliom et al. (2006a) estimated that over “90% of the time, water from streams with agricultural, 

urban, or mixed-land-use watersheds had detections of two or more pesticides or degradates, and 

about 20% of the time they had detections of 10 or more.” The likelihood of exposure to 

multiple pesticides throughout a listed salmonid’s lifetime is great, considering the geographical 

range of their migration routes and habitats occupied during spawning and rearing. 

Studies have suggested that assessment of pesticide mixture toxicity to aquatic life is needed 

given the widespread and common occurrence of pesticide mixtures, particularly in streams 

where the total combined toxicity of pesticide mixtures may be greater than that of any single 

pesticide compound (Gilliom 2007, Gilliom et al. 2006a). Exposure to multiple pesticide 

ingredients can result in additive and synergistic responses as described in the Risk 

Characterization section. It is reasonable to conclude that compounds sharing a common mode 

of action cause additive effects and in some cases synergistic effects. Exposure to these 

compounds and other baseline stressors (e.g., thermal stress) was not a consideration in EPA’s

BEs, which only considered effects from single a.i.s. Therefore, risk to listed species may be 

underestimated in EPA’s assessments. 

8.4. Exposure Conclusions

Pacific salmon and steelhead use a wide range of freshwater, estuarine, and marine habitats and 

many migrate hundreds of miles to complete their life cycle. Thiobencarb is frequently applied 

to rice crops within the distribution of three listed salmonids that occur in the Central Valley of 

California: CV spring-run Chinook, Sacramento River winter-run Chinook, and CV steelhead. 

Thiobencarb and its degradates have been detected in habitats utilized by these species. Because 

the action proposes continued use on rice in California for the next 15 years, we expect 

thiobencarb will continue to be present within the freshwater distribution of these species, 

including times when salmon are present. Therefore we expect some individuals of each of these 
206

species, and their designated critical habitats, will be exposed to the thiobencarb and other 

stressors of the action. 

We considered several sources of information to characterize the likely range of exposure to 

thiobencarb and the other action stressors (Table 49). Inherent in the modeling estimates is the 

assumption that the pesticide is applied in a location next to or draining into salmon-bearing 

waters. Monitoring data may reflect pesticide applications proximate to the waterbody (i.e.,

targeted monitoring), or resulting from more distant uses in the watershed (ambient monitoring). 

The surface water monitoring data used were not designed to determine exposure to listed 

salmonids. Monitoring studies were designed for other purposes and sample design did not 

reflect the spatial and temporal distribution of the listed species. Therefore, caution should be 

exercised in using these data for that purpose, especially when conducting probabilistic 

assessments. Defining exposure of the stressors of the action to the listed species is also 

complicated by uncertainty associated with the following factors:

Product labels authorize the application of chemical mixtures that are not specified or 

clearly defined (e.g., the ingredients of pesticide formulations are not fully disclosed, 

labels recommend tank mixture applications with other products, and tank mixtures with 

other pesticides are permitted unless specifically stated otherwise); 

Historical use of thiobencarb, including frequency of use, locations of use, and the 

amount of thiobencarb products applied may not reflect future use.

Table 49. Chemical exposure data ranges in monitoring data and modeling.

Exposure pathway Value and Units

Runoff

Peak concentrations measured within rice fields (Table 24)

Thiobencarb 

Thiobencarb sulfoxide

4-chlorobenzylmethylsulfone

13 - 576 μg/L

8.9 - 22 μg/L

5.2 – 8.3 μg/L

Dissipation half-life (Table 24)

Thiobencarb

Thiobencarb sulfoxide

4-chlorobenzylmethylsulfone

5.8 – 8.7 d

2.8 - 3.4 d

6.0 – 10.4 d

Concentrations estimated at discharge (see EPA Rice Model)
207

Exposure pathway Value and Units

Day 0

Day 14 

Day 30 

2018 μg/L

350 μg/L

47 μg/L

Concentrations measured in discharge and receiving water

(see Targeted Monitoring)

Day 0 

Day 14

Day 30 

Non-detect - 8,900 μg/L

Non-detect -83 μg/L

Non-detect – 6 μg/L

Frequency of early releases in Sacramento Valley (Table 44)

Authorized

Unauthorized

Seepage < 5 gpm

Seepage > 5 gpm

0-3/year

0.2% of applications

3.6% of applications

0.4% of applications

Drift

AgDrift predictions to floodplain habitats without buffer (Table 33)

Ground application

Aerial Application

AgDrift Predictions to terrestrial habitats (Table 34) –

ground applications with buffer

1 ft

10 ft

100 ft

1000 ft

AgDrift Predictions to terrestrial habitats (Table 34) –

aerial applications with buffer

1 ft

10 ft

100 ft

1000 ft

Measured thiobencarb drift from aerial application with buffer 

(Table 45)

39 ft 

82 ft 

164 ft 

328 ft 

656 ft 

1312 ft 

1640 ft 

2297 ft 

2625 ft 

3937 ft 

37 – 736 μg/L

96-1,910 μg/L

1.19 lbs/A; 29.68% applied

0.09 lbs/A; 2.24% applied

0.010 lbs/A; 0.24% applied

0.001 lbs/A; 0.03 % applied

1.91 lbs/A; 47.76% applied

1.21 lbs/A; 30.34% applied

0.22 lbs/A; 5.61% applied

0.02 lbs/A; 0.55% applied

5.71 - 15.02 % applied

2.13 – 2.51 % applied

0.44 – 1.48 % applied

0.07 – 0.29 % applied

0.03 – 0.32 % applied

0.02 – 0.05 % applied

0.00 – 0.01 % applied

0.00 – 0.06 % applied

0.00 - 0.01 % applied

0.00 % applied

Ambient monitoring in California surface water

(see USGS NAWQA Data and Monitoring Data from California)

Prior to establishing irrigation water holds

Since establishing irrigation water holds

Not detected – 170 μg/L 

Not detected – 16.55 μg/L
208

Runoff of thiobencarb from rice fields, including planned discharge following label-specified 

holding periods of 14 and 30 days, will result in deposition of thiobencarb to salmonid habitats. 

Concentrations may attain levels as high as 350 μg/L at 14 days and 47 μg/L at 30 days in 

discharge waters (Table 49). Concentrations may be orders of magnitude higher in runoff if rice 

irrigation water is released within hours or days post application. This may occur due to 

emergency releases of irrigation water, and from unplanned and/or unauthorized discharges. 

Several field studies reported concentrations in the mg/L range following early releases, one 

attaining levels as high as 8.9 mg/L (Lauck 1979). We do not expect this to be a common 

occurrence based on reports over the last decade on the frequency of early releases. 

Drift of the liquid formulation of thiobencarb into salmonid habitats is expected during the 

application period, primarily May and June, and to a lesser extent April and July. Concentrations 

will depend on the proximity to application area, application method, droplet size, release height, 

wind speed/directions, receiving water volume/flow, and interception by riparian vegetation. 

Based on drift studies, modeling exercises, and surface water monitoring studies, aquatic 

concentrations from drift may range from non-detectable to as high as 1910 μg/L (Table 49). 

We assume that the exposure estimates provided by EPA in the BEs, and additional modeling 

and monitoring information provided above, represent realistic exposure levels for some 

individuals of the three listed species. Further, we assume the distribution within the range of 

exposures is a function of pesticide use and the duration of time listed salmonids spend in these 

habitats. All listed Pacific salmon and steelhead occupy habitats that could contain high 

concentrations of these pesticides at one or more life stages. However, the time spent in these 

habitats varies among the species that occur in the rice growing region. We are unable to 

accurately define exposure distributions for thiobencarb and the other stressors of the action 

given limitations of the available information. We assume the highest probability of exposure 

occurs in freshwater habitats in close proximity to Central Valley rice fields where thiobencarb is 

applied. We evaluated several spatial and temporal relationships to qualitatively assess the 

relative abundance of the three species in aquatic habitats near rice (Table 28, Table 30, Table 

32, and Table 50). 
209

Table 50. Relative abundance of listed salmonids near rice fields during the period when 

thiobencarb is applied and water is discharged from thiobencarb treated fields (April - July).

Life Stage None Expected Low Medium High

CV spring-run Chinook

Adult Migration X

Spawning X

Early rearing X

Juvenile 

migration X

Sacramento River winter-run Chinook

Adult migration X

Spawning X

Early rearing X

Juvenile 

migration X

CV steelhead

Adult migration X

Spawning X

Early rearing X

Juvenile 

migration X

The relative abundance estimates suggest that exposure will vary depending on the species and 

life stage. Given the available information, we conclude that some adults of the three listed 

species are likely to be exposed to thiobencarb during their spawning migration. However, 

spawning locations for these species occur in the watershed above the locations where rice is 

grown. Consequently, we consider it unlikely that any significant exposure to thiobencarb will 

occur during spawning, egg incubation, and fry emergence. However, some post-emergence fry, 

parr and pre-smolts may be exposed to the stressors of the action as they are displaced or 

migrate downstream to habitats that are near locations where thiobencarb is applied. These 
210

observations are carried forward to the Risk Characterization section to develop risk hypotheses 

and evaluate potential effects to individuals and populations. Additionally, we carry forward the 

following conclusions:

The vast majority of exposure to thiobencarb drift occurs in May and June with less 

frequent exposures in April and July. Floodplain habitats alongside rice fields are at the 

greatest risk of receiving elevated concentrations from aerial and ground applications 

(i.e., 37 – 1910 μg/L); 

runoff containing thiobencarb typically occurs 14-30 days post application (after required 

holding periods), thus the greatest probability of exposure to the highest runoff 

concentrations likely occurs from mid-May through the end of July; 

riparian systems and multiple life stages of salmonids are likely exposed to thiobencarb 

from drift during applications and from runoff following release of rice irrigation water 

into drainage networks that ultimately return water to salmonid containing waters; 

migrating adults, rearing juveniles, and migrating juveniles overlap with peak 

thiobencarb applications in May and June; and

monitoring data and modeling results show that thiobencarb contaminates salmonid 

habitats from drift and runoff with peak values occurring during thiobencarb 

applications.

Substantial data gaps in EPA’s exposure estimates include estimates for “other ingredients” in 

pesticide formulations, other pesticide products authorized for co-application with thiobencarb, 

adjuvants, degradates, and metabolites. Although NMFS is unable to comprehensively quantify 

exposure to these chemical stressors, we are aware that exposure to these stressors is likely. We 

assume these chemical stressors may pose additional risk to listed Pacific salmonids. In order to 

ensure that EPA’s action is not likely to jeopardize listed species or destroy or adversely modify 

critical habitat, NMFS analyzes potential exposure based on all stressors that could result from 

all uses authorized by EPA’s action.
211

8.5. Response Analysis

In this section we evaluate toxicity information from the stressors of the action organized by 

assessment endpoints. The endpoints target potential effects to individual salmonids and their 

supporting habitats. The assessment endpoints represent biological and habitat attributes that, 

when adversely affected, lead to reduced fitness of individual salmonids or degrade PCEs such 

as prey abundance, water quality, and riparian vegetation (evaluated in risk characterization for 

designated critical habitat section). Uncertainties in the available toxicity information are 

discussed as they are encountered and summarized at the end of this section. Following the 

response analysis, we compare concentrations described in the exposure analysis with 

assessment endpoints to evaluate whether individual fitness or habitat endpoints might be 

compromised. Salmonid and designated critical habitat risk hypotheses are evaluated separately 

in the Risk Characterization sections.

 

Figure 34. Response Analysis Conceptual Model

We begin the response analysis by describing the toxic mode and mechanism of action of 

thiobencarb, which sets the stage for what biological endpoints are assessed. Next, we 

summarize the toxicity data presented in the salmonid BE, RED, IRED, California Red Legged 

Frog BE, EFED science chapters, and open literature (Table 51). The information is organized 

by assessment endpoints (e.g., survival, growth, migration, etc.). The information provided in 

Effects of pesticide products 

on ESA-listed species and 

their habitat 

Individual 

responses

Habitat 

responses

Response Profile
212

EPA documents primarily addressed aspects of survival, growth, and reproduction of some 

aquatic species following exposure to thiobencarb. Other information from selected field 

experiments on ecological endpoints was sometimes discussed within EPA documents. Since we 

found little information in EPA’s documents regarding formulation, other ingredients, and 

mixture toxicity, we conducted our own literature review.

Under the ESA and implementing regulations NMFS evaluates all direct and indirect effects of a

federal action. We therefore evaluate all aspects of an action that may reduce fitness of 

individuals or reduce PCEs of designated critical habitat. This includes toxicity information for 

thiobencarb, its degradates, other ingredients within thiobencarb formulations, and other 

pesticide active ingredients commonly combined in recommended tank mixtures. The evaluation 

includes information that EPA provided on survival, growth, or reproduction, and also 

encompasses a broader range of endpoints including behaviors, endocrine disruption, and other 

physiological alterations. The information we assessed is derived from published scientific 

journals, government agency reports, theses, books, applicant-submitted information, and 

independent reports. The most relevant study results are those that directly measure effects to 

identified assessment endpoints derived from studies with salmonids, preferably ESA-listed 

Pacific salmonids or hatchery surrogates. We also evaluate additional stressors that may 

influence the toxicity of the stressors of the action such as temperature.

8.5.1. Thiobencarb’s Mode and Mechanism of Action 

Thiobencarb belongs to a class of herbicides known as thiocarbamates which impair and kill 

plants by inhibiting the production of hormones called gibberellins. Thiobencarb also inhibits a 

plant’s fatty acid and lipid biosynthesis (Weed Science Society of America 2007). 

Thiocarbamates affect the functioning of acetyl-CoA elongases, preventing the extension of the 

fatty acid chains (Gronwald 1991). Carbamothioates “may be rapidly metabolized or poorly 

translocated” and effects on plant lipid biosynthesis generally occur at the plant surface 

(Gronwald 1991). Thiobencarb is currently registered as an herbicide.

Thiobencarb is also an acetylcholinesterase inhibitor. USEPA included thiobencarb in its 

analysis of common mechanism of action with six other thiocarbamates: molinate, EPTC, 
213

trillate, butylate, pebulate, and cycloate (USEPA 2001). EPA concluded that thiobencarb shares 

a common mechanism of action for AChE inhibition with other thiocarbamates. Thus, we use 

other thiocarbamates as chemical surrogates when information is unavailable for thiobencarb. 

Empirical research indicates that AChE is inhibited by thiobencarb in fish and mammals 

(Fernandez-Vega et al. 1999, Pentyala and Chetty 1993). For example, thiobencarb reduced 

European eel plasma AChE activity by 50% following 96 hours of exposure at 220 μg/L (1/60th

of the 96 hr LC50) and remained depressed (<50% activity) following five days of recovery in 

clean water (Fernandez-Vega et al. 1999). 

8.5.2. Temperature and toxicity

We located no information showing specific effects of temperature on thiobencarb’s toxicity. 

Elevated temperatures typically increase the magnitude of toxic effects in fish particularly for 

pesticides that are transformed in the fish to more toxic metabolites such as organophosphates 

(Mayer and Ellersieck 1986). Carbamates including thiocarbamates have also been shown to 

impart greater toxicity to organisms when exposed to elevated temperatures (Altinok et al. 2006). 

Differences in toxicity due to temperature have been attributed to differences in respiration rate, 

chemical absorption, and metabolism. As discussed in the Environmental Baseline, temperature 

is a recognized stressor to salmonids in the Central Valley (Myrick and Cech 2005). Water 

temperatures in the lower Sacramento River regularly exceed 20 °C by late spring, and statistical 

studies of coded wire-tagged juvenile Chinook show increased mortality as a function of 

temperature (Baker et al. 1995). Water temperatures higher than optimum levels can kill 

salmonids, increase physiological stress making them more susceptible to other stressors, 

increase predation, and affect salmonid’s prey base. Thus temperature directly affects survival, 

growth rates, distribution, and developmental rates. We therefore discuss Chinook and steelhead 

fitness implications in the context of elevated temperatures enhancing thiobencarb’s toxicity. 

8.5.3. pH and toxicity

We located no studies that tested pH’s effect on thiobencarb’s aquatic toxicity. However, pH 

influenced acute lethality of N-methyl carbamates aminocarb, carbofuran, and carbaryl (Mayer 

and Ellersieck 1986). We are uncertain whether thiobencarb, a thiocarbamate, will behave 
214

similarly as the three N-methylcarbamates. Thus this is a recognized data gap and we cannot 

definitively determine whether biological endpoints will be more or less affected by pH.

Studies with mixtures of AChE inhibiting herbicides. We located information on mixture toxicity 

indicating additive toxicity when fish are jointly exposed to thiobencarb and another 

thiocarbamate, molinate (Finlayson and Faggella 1986). Acute lethality tests (96-hr LC50s) of a 

thiobencarb and molinate mixture showed additive toxicity to juvenile steelhead, Chinook 

salmon, and channel catfish. Thiobencarb and molinate were twice as lethal when present 

together at 1:1 LC50-ratios than they were individually (Finlayson and Faggella 1986). 

Thiobencarb was 18-22 times more toxic than molinate based on individual LC50s for the three 

fish species. Thiobencarb concentrations used in the ratios were 360 (steelhead), 430 (Chinook), 

and 990 μg/L (channel catfish). These results corroborate additive mixture toxicity when 

pesticides share a common mechanism/mode of toxic action. Although use of molinate has been 

prohibited since 2009, other thiocarbamate pesticides are currently used in California and may 

co-occur in the environment with thiobencarb in the future (e.g., EPTC and cycloate; (CDPR 

2010b) . 

8.5.4. Herbicide effects to salmonids and their habitats

In previous Opinions, we have addressed organo-phosphorus (OP) and carbamate insecticides 

(NMFS 2008a, NMFS 2009e, NMFS 2010a). Although used to control insects, these pesticides

have a mode of action (cholinesterase inhibition) expected to directly affect salmon and other 

non-target organisms, such as aquatic and terrestrial invertebrates that provide a forage base for 

salmon. While thiobencarb is also a cholinesterase inhibitor that may have direct effects on fish 

and invertebrates, we also investigated the indirect effects caused by its herbicidal mode of 

action. We surveyed available literature regarding herbicide effects in the environment, 

considering a broad range of herbicides, including ones not addressed in this Opinion. A 

summary of this survey and conceptual models based on information gleaned from this survey 

are presented below. 
215

8.5.5. Importance of plants and other photosynthetic organisms in fueling secondary 

production within salmonid habitats

Secondary production within aquatic systems, including production of juvenile salmonids, is 

ultimately fueled by plants and other photosynthetic organisms (e.g., green algae, diatoms, 

cyanobacteria). In salmonid freshwater and estuarine habitats, this energy comes from two 

sources: (1) primary production within aquatic habitats (autochthonous inputs), as well as (2) 

inputs of organic matter from adjacent terrestrial ecosystems (allochthonous inputs) (Allan and 

Castillo 2008). Plants and other photosynthetic organisms are primary producers, and can be 

consumed (by “consumers”) as living tissue that is grazed from the benthos (periphyton), as 

living tissue collected from the water column (phytoplankton), or as dead tissue that is consumed 

after being colonized by microbial communities (detritus). Invertebrates and fish are specific 

with regard to their ability to feed on these various food resources, and these distinctions help 

define functional feeding groups that include grazers, shredders, and predators, among others. 

Therefore, although there is great diversity in the pathways energy takes in an aquatic system, 

much of the energy that fuels production in aquatic habitats derives ultimately from plants and 

other photosynthetic organisms.

Fish can consume a very high proportion of the invertebrate secondary production in aquatic 

habitats (Huryn 1996, Huryn 1998). Juvenile salmonids are predators that can consume a wide 

range of invertebrates, including those from all functional feeding groups. Changes in the 

production of any of these groups could change prey availability for these fish. For example, a 

reduction in periphyton production on rocks in a stream could reduce invertebrate grazer 

production. Likewise, a change in the quantity or quality of terrestrial leaf litter falling from a 

riparian buffer could alter the production of invertebrate shredders downstream. In addition to 

being the ultimate source of food for much of the invertebrate community, plants also provide 

habitat for invertebrates and fish, including but not limited to substrate for them to shelter on and 

under (e.g., macrophytes, root wads). Plants and other photosynthetic organisms within and 

adjacent to salmonid freshwater and estuarine habitats are therefore essential components of 

productive salmonid habitats. Actions that affect the diversity, biomass and/or the production of 

primary producers in and around salmonid habitats may limit or alter secondary production 

within those systems as well (Figure 35, Figure 36).
216

As food resources, living plants and other living photosynthetic organisms are especially 

nutritious for grazing invertebrates and herbivorous fish (Torres-Ruiz et al. 2007), and they often 

contribute more to overall secondary production within a system than would be expected simply 

by their standing stock at any one point in time (Allan and Castillo 2008). Because of this high 

nutritional value, the autochthonous production of plants and other photosynthetic organisms can 

be limited by grazers, though abiotic factors such as light, nutrients, and water velocity are also 

often limiting (Blanchet et al. 2008, Rosemond et al. 2000, Sanderson et al. 2009). The relative 

importance of these biotic and abiotic factors in limiting primary production varies by system 

and can change seasonally (Huryn 1998, Sanderson et al. 2009). 

When primary production is limited or low, consumer production can be limited. This has been 

demonstrated primarily by amending a limiting resource to the point at which it is no longer 

limiting. For example, when nutrients are added to nutrient-limited systems, primary production 

and consequently secondary production can increase (Harvey et al. 1998, Mundie et al. 1991). 

Fewer studies have examined explicitly how reductions in primary producers (or primary 

production) affect fish and invertebrates, as would potentially occur when sensitive 

photosynthetic organisms at the base of an aquatic food web are exposed to herbicides. In some 

cases when algal biomass is reduced by disturbances, invertebrate grazer growth and abundance 

decline. Higher trophic levels can be affected by these bottom-up effects, as Perry and 

colleagues (2003) observed in juvenile Chinook salmon. In small tributaries of the Yukon River, 

a fire and flood reduced the proportion of high quality autochthonously derived energy that 

salmon consumed, suggesting there may be direct and indirect effects of disturbances on energy 

transfer among trophic levels including salmon (Perry et al. 2003). 

The loss or reduction of inputs of organic matter (including leaf litter, woody debris, and 

terrestrial insects) from adjacent terrestrial ecosystems can also significantly reduce invertebrate 

secondary production and potentially fish production (Wallace et al. 1999, Allan et al. 2003). 

This was demonstrated by Wallace et al. (1999) when they excluded terrestrial leaf litter from a 

forest stream in the southeast for four years and found that invertebrate production in the affected 

habitats declined by 78%. Although there were no fish in these systems, they did observe 
217

reductions in the top invertebrate predators, illustrating that bottom-up effects of this exclusion 

of plant material permeated throughout the food web (Wallace et al. 1999). Similarly, Fischer et 

al. (2010) suggested that differences in food availability associated with the presence or absence 

of riparian buffers likely affected differences in observed fish growth. In systems where 

allochthonous inputs sustain secondary production (including shaded, forested streams that 

provide rearing habitat for some salmonids), a reduction in allochthonous inputs could reduce 

secondary production, and consequently affect fish production. In addition to organic inputs, 

riparian vegetation provides shade for aquatic habitats, increases bank stability, helps buffer 

aquatic habitats from contaminants present upland, and helps maintain natural flow dynamics of 

water, nutrients and sediment (Richardson et al. 2010). 

Numerous studies illustrate the trophic linkages among plants and other photosynthetic 

organisms and the secondary production of fish and their prey. While it is logical that reductions 

in autochthonous and/or allochthonous food resources could limit consumers and predators, 

including juvenile salmonids, there are often a number of factors that affect the magnitude and 

even the direction of change within complex aquatic food webs. These relationships may be 

directly or indirectly affected by herbicides. The following sections briefly review some of these 

impacts and discuss the challenges faced in predicting how herbicides may affect salmonids and 

their critical habitats. 

8.5.6. Effects of herbicides on non-target aquatic communities

Potential effects of herbicides on aquatic and riparian communities are illustrated in Figure 35

and in Figure 36. The range of effects includes direct effects (primarily negative) on 

photosynthetic organisms and water quality parameters, as well as indirect effects (positive and 

negative) on multiple trophic levels and water quality. Generally, if an herbicide exposure is 

great enough to reduce primary production within or adjacent to aquatic habitats, there may be 

effects on multiple trophic levels, including salmonids. A number of factors contribute to the 

magnitude and direction of effects, although it’s difficult to predict and identify patterns within 

and across aquatic systems. 
218

Numerous studies using standard toxicity tests have demonstrated that herbicides reduce the 

growth and biomass of photosynthetic organisms. Plants and photosynthetic organisms are 

typically more sensitive to herbicides than invertebrates and fish because of the herbicides’ 

various mechanisms of toxic action. For example, Brock and others (Brock et al. 2004)

determined HC5s (hazardous concentrations for 5% of the species) for two herbicides 

(metribuzin and metamitron) on a variety of taxa, and found, not surprisingly, that the algae and 

macrophytes were >100 to >1000x more sensitive than invertebrates and fish. Similarly, Van 

den Brink et al. (Van den Brink et al. 2006) found that herbicides varied in their toxicity, and that 

relative sensitivities (based on short-term toxicity growth tests) of the taxonomic groups included 

in the study were algae ≥ macrophytes > invertebrates > vertebrates. For some herbicides, algae 

and macrophytes were similar in their sensitivities, e.g., for atrazine and diquat, (Van den Brink 

et al. 2006), but for others, such as 2,4-D (an auxin simulator), macrophytes were significantly 

more sensitive than all of the algae taxa included in the analyses (Van den Brink et al. 2006). In 

their extensive review of herbicides, Brock et al. (2000) also concluded that auxin simulators like 

2,4-D were generally more toxic to macrophytes than other photosynthesis inhibitors. Neither 

paper reviewed studies on thiobencarb, however Brock et al. (2000) reviewed studies on triallate 

(a thiocarbamate) and remarked that the acute 48 h survival EC50 for Daphnia magna (57 μg/L ; 

(Johnson 1986)) is similar to that of a standard test algae, C. selenastrum (47 μg/L ; (Fairchild 

JF et al. 1997)). 

The direct effects of herbicides on diverse communities of aquatic primary producers can be 

highly variable. For instance, Gruessner and Watzin (Gruessner and Watzin 1996) exposed 

stream communities in microcosms to a low concentration of atrazine (5 μg/L) for 14 days, but 

found no effect on algal biomass. In other studies, the species composition of primary producers 

changes after exposure while abundance may increase or decrease. Wendt-Rasch et al. (WendtRasch et al. 2003) found that even though macrophyte root growth in mesocosms declined 

following exposure to metsulfuron methyl, the biomass of periphytic algae on those macrophytes 

actually increased. In addition, the algal species composition was significantly different in the 

mesocosm exposed to the highest dose (Wendt-Rasch et al. 2003). Hartgers et al. (Hartgers et al. 

1998) observed an initial decline in the abundance of some phytoplankton taxa following 
219

exposure to a mixture of herbicides (atrazine, diuron and metolachlor), but by 14 days postapplication several phytoplankton taxa had actually increased in abundance. 

In addition to direct effects on primary producers, there may be direct effects of herbicides on 

microbial communities. The processing of organic matter by microbial communities – which 

includes, in part, making leaf litter palatable to some invertebrates – is a critical energy pathway 

within aquatic food webs. Despite their importance, there are relatively few studies examining 

the effects of pesticides on microbial communities. Of the few studies regarding herbicides, it 

appears there may be some direct and indirect effects at relatively low concentrations. Microbial 

communities were altered following exposure to various concentrations of atrazine, with some 

taxa becoming more abundant and productive while others declined (DeLorenzo et al. 1999). In 

another study, the herbicide diuron limited algal growth in mesocosms, and the abundance, 

diversity, and activity of the associated microbial community was also limited (Pesce et al. 

2006). The authors suggested that diuron decreased the capacity of the microbial community to 

recover when favorable conditions were provided (as was the case in the control mesocosms), 

and this reduced the efficiency of the microbial food web (Pesce et al. 2006). Although it is 

difficult to extrapolate short-term mesocosm studies to potential longer-term effects in the 

natural environment, these studies suggest that exposure to herbicides can directly affect the 

structure as well as function of the diverse communities that are the base of aquatic food webs.

The effects of herbicides, either by reducing primary producers or by changing the processes and 

paths through which energy flows, can have significant effects on higher trophic levels. For 

example, herbicides are commonly found to reduce the abundance (or biomass or growth rates) 

of consumers. Interestingly, these indirect effects of herbicides are often reported at 

concentrations well below those found to have direct effects on consumers. The population 

growth rate of an aquatic oligochaete Lumbriculus variegatus was reduced by 50% after being 

exposed to only 6 μg/L of the herbicide terbutryn (Brust et al. 2001). This effect was attributed 

to the reduction of the food source of the oligochaete by the herbicide at a concentration three 

orders of magnitude lower than the concentration that caused acute toxicity to the oligochaete 

itself. Similarly, Dewey (Dewey 1986) found that multiple trophic levels within experimental 

ponds were impacted by atrazine, though effects on higher trophic levels were likely due to 
220

indirect effects (reduction in food resources). These effects throughout the food web were found 

at concentrations one order of magnitude lower than acute toxicity values for a common midge 

(Dewey 1986). Brock et al. (Brock et al. 2004) observed long-term (lasting >8 weeks) changes 

in the macroinvertebrate communities within mesocosms treated with metribuzin at 

concentrations 20x lower than the HC5s for aquatic invertebrates. In a similar study, predatory 

ciliates were relatively more affected by the reduction of their prey (phototrophic flagellates) due 

to exposure to the herbicide prometryn than by the direct toxicity (Liebig et al. 2008). Finally, a 

number of studies have documented declines in zooplankton densities due to reductions in their 

phytoplankton food sources following exposure to herbicides (DeNoyelles et al. 1982, Juttner et 

al. 1995, Kasai and Hanazato 1995).

These examples illustrate that reduced primary production due to herbicide exposure results in 

bottom-up effects. Alternatively, if an herbicide is directly toxic to consumers, primary 

production may actually increase as grazing pressure declines (Rohr and Crumrine 2005). In 

addition, sublethal effects of herbicides on invertebrates have also been found at environmentally 

relevant concentrations, and this may also have effects throughout the food web. For example, 

Cook and Moore (Cook 2008) found the herbicide metolachlor (at an environmentally relevant 

concentration of 80 g/L) altered agonistic behavior in crayfish. 

Effects on water quality are also often reported. These changes are due in part to changes in 

community metabolism (Brock et al. 2000). For example, if photosynthetic efficiency declines, 

it is expected that oxygen concentrations and pH decrease (Hartgers et al. 1998, Brock et al. 

2000). These effects have been shown to be dose-dependent by Pratt et al. (Pratt et al. 1997), as 

oxygen levels decreased most significantly in microcosms exposed to the highest doses of the 

herbicide diquat. Changes in water quality, especially significant declines in dissolved oxygen, 

may affect sensitive taxa, but it is unclear how often this may occur in salmonid habitats.

Brock et al. (Brock et al. 2000) concluded in their review of herbicides that indirect effects of 

photosynthetic inhibitors on consumers and predators occur at concentrations around the EC50

for standard algae taxa; these impacts on consumers and predators are likely due to reduced 

availability of food resources and the effects may be delayed relative to the exposure event. 
221

Other effects on the ecosystem (e.g., blooms of insensitive algae) can occur at lower 

concentrations (e.g., 0.1 of the EC50 of standard algae), and these effects may also be delayed. 

When macrophytes are impacted, organisms using those macrophytes as habitat are immediately 

impacted. Some studies published after the Brock et al. (2000) review noted indirect effects at 

surprisingly low concentrations, and, in general, papers published since their review corroborate 

their findings. As mentioned previously, Brock et al. (2000) reviewed studies on triallate (a 

thiocarbamate) and remarked that the acute 48 h survival EC50 for Daphnia magna (57 μg/L; 

(Johnson 1986)) is similar to that of a standard test algae, C. selenastrum (47 μg/L; (Fairchild JF 

et al. 1997)). It appears that triallate is an order of magnitude more toxic than thiobencarb based 

on D. magna survival, i.e., 10 μg/L vs. 101 μg/L, respectively. We present thiobencarb’s toxicity 

to algae and invertebrates in Table 51.

8.5.7. Challenges in scaling up effects and making predictions across salmonid habitats

The current literature describes a wide range of effects of herbicides. While it is difficult to 

generalize across these studies, it is clear that many studies illustrate that herbicides can have 

direct and indirect effects on multiple trophic levels within aquatic food webs, and often these 

effects occur at concentrations well below concentrations expected based on single-species acute 

toxicity tests. However, it is difficult to predict the magnitude, duration, and direction that these 

effects may have on juvenile salmonids and their habitat because multiple factors influence these 

effects. These factors include, but are not limited to, the composition and relative abundances of 

taxa at the time of exposure (Relyea 2009), the functional redundancy among taxa within the 

system, and the resilience of the various communities within the system (Brock et al. 2000). In 

addition, the abiotic conditions, the presence of other stressors, and the properties of the 

herbicides themselves (e.g., mode of action, persistence) can affect the magnitude, duration and 

direction of effects. 

Juvenile salmonids are generally opportunistic drift-feeders, and are therefore sensitive to factors 

that influence the general quantity and quality of invertebrate prey items. If, for instance, there 

were reductions in the production of invertebrate grazers or the inputs of invertebrate prey from 

riparian vegetation, salmonids may be forced to alter their foraging behavior (e.g., take more 
222

risks, select less energy-rich prey) (as shown in Figure 35 and in Figure 36). Alternatively, if 

there were shifts, rather than reductions, in the abundances and composition of the prey 

community within riparian and aquatic habitats, indirect impacts on salmonids may be minimal if 

foraging behaviors were not altered. Whether or not production of prey decreases or shifts (or 

increases) after exposure to herbicides will depend in part on the composition of the community 

(structure and function) and the relative sensitivities of those taxa. Multiple experiments 

conducted in mesocosms have demonstrated that the particular composition of the community at 

the time of exposure influences the magnitude of the impact as well as the trajectory of the 

recovery (Jenkins and Buikema 1998, Pesce et al. 2006, Relyea 2009, Rohr and Crumrine 2005),

and this would likely be the case as well in salmonid habitats.

Abiotic conditions may also affect how herbicides directly and indirectly affect salmonids and 

their habitats. For instance, herbicides can affect water quality parameters that may indirectly 

affect aquatic communities. Austin et al. (1991) suggest that increased algal production in 

oligotrophic systems after exposure to glyphosate may be due to the addition of phosphorous (in 

the glyphosate), and they suggest this could lead to eutrophication of salmonid habitats. 

Likewise, total phosphorous increased eightfold in earthen mesocosms treated with glyphosate 

(Perez et al. 2007). In forested watersheds in the southeastern United States, nitrogen 

concentrations were elevated in streams for two years after herbicides were applied (Neary et al. 

1993). This effect was likely due to the increased leaching from the terrestrial environment 

and/or reduction in uptake within the stream. Regardless of how nutrients become elevated 

(from the herbicide itself or from changes in biogeochemical cycles within the watershed) 

elevated nitrogen and phosphorous concentrations can stimulate periphyton growth in nutrientlimited systems and consequently affect higher trophic levels. Indirect effects from herbicides 

may also include an increase in stressful water temperatures due to reduced shading and longterm reductions in woody debris used for cover by salmonids from loss of riparian vegetation. If 

herbicides were used to reduce plant growth over a large area within a watershed, instream flow 

dynamics may be impacted enough to affect salmonids and their habitats (e.g., (Likens et al. 

1970). Finally, changes such as increased turbidity (due to reduced bank stability) or decreased

dissolved oxygen could have impacts on primary producers as well as consumers within 

salmonid habitats (Figure 35 and Figure 36).
223

In addition to the uncertainties associated with variable and diverse communities and the range 

of sensitivities they have to various abiotic conditions, there are uncertainties about how 

herbicides may affect aquatic systems affected by other stressors. When experiments are used to 

examine multiple stressors, the results are often variable and again (like simpler experiments) 

often depend on the abiotic and biotic conditions at the time. In a series of experiments, Rohr et 

al. (Rohr et al. 2004) found few interactions among food availability, drying conditions and 

atrazine (at 4 concentrations) on a streamside salamander, but they did find that the lethality of 

atrazine varied by year and may be condition dependent. Figure 35 and Figure 36 illustrate the 

direct and indirect effects stemming from herbicide exposure, but they do not attempt to capture 

the complex web of interactions that may arise when multiple stressors affect a system.
224

Herbicides

mortality of 

sensitive taxa

Trees

Shrubs

Grasses

change in competition

Trees

Shrubs

Grasses

reduction in bank stability may increase 

turbidity, sedimentation

increase in light availability 

stimulates primary producers

decrease 

in prey

increase 

in prey

top-down control of prey

release from predation

long-term reduction in inputs of woody debris and 

other organic matter; alteration of nutrient dynamics

change in competition

(-) Riparian vegetation (+)

(-) Riparian animal community (+)

Insects and other 

invertebrates

Amphibians

Other vertebrates

Insects and other 

invertebrates

Amphibians

Other vertebrates

change in competition

(-) Aquatic consumers (+)

Zooplankton

Insects

Amphibians

Fish

Zooplankton

Insects

Amphibians

Fish

(-) Aquatic primary producers and microbial community (+)

Photosynthetic 

organisms

Microbial 

community

Photosynthetic 

organisms

Microbial 

community

(-) Aquatic predators (+)

Insects

Crustaceans

Fish

Insects

Crustaceans

Fish

reduction in inputs of 

invertebrate prey

potential short-term increase in inputs of plant 

detritus; alteration of nutrient dynamics 

change in competition

change in competition

top-down control by consumers

release from consumption pressure

reduction 

in food 

and 

habitat

increase in 

food and 

habitat

long-term reduction in woody debris needed for shelter; increased frequency of 

stressful water temperatures; increased sedimentation

change in water flow dynamics

Figure 35. Part I of a conceptual model of potential effects of herbicides to aquatic communities. 

This figure focuses on potential effects of herbicides applied to riparian areas adjacent to 

salmonid habitats. Bolded arrows and text note those effects that are most likely to occur based 

on the frequency that they are reported in the literature.
225

Herbicides

reduction In habitat and substrate

(-) Aquatic consumers (+)

Zooplankton

Insects 

Amphibians

Fish

mortality of sensitive taxa

(-) Water quality (+)

DO

pH

Turbidity

Nutrients

Minerals

Alkalinity

DOC

Turbidity

Conductivity

(-) Aquatic microbial community (+)

cyanobacteria

ciliates

flagellates

(-) Aquatic primary producers (+)

Phytoplankton

Periphyton

Macrophytes

change in competition

Phytoplankton

Periphyton

Macrophytes

(-) Aquatic predators (+)

Insects

Crustaceans

Fish

reduction In food resources

reduction in primary 

productivity

change in quantity & quality of organic matter

source of 

nutrients

increase in food resources

change in quantity & quality of 

food resources

decrease 

in prey

stimulation by 

nutrients

increase 

in prey top-down 

control of prey

release from consumption by herbivores

release from 

predation

top-down control by herbivores

reduction in oxygen

cyanobacteria

ciliates

flagellates

Zooplankton

Insects 

Amphibians

Fish

change in competition

Insects

Crustaceans

Fish

change in competition

change in competition

Figure 36. Part II of a conceptual model of potential effects of herbicides to aquatic communities. 

This figure focuses on potential effects of herbicides that are applied to or otherwise reach 

salmonid habitats. Bolded arrows and text note those effects that are most likely to occur based 

on the frequency that they are reported in the literature.

Mixtures of pesticides present a particular challenge. Most of the experiments described above 

were conducted in mesocosms with a single exposure of a single herbicide. In field surveys in 

the Unites States as well as throughout Europe, herbicides are often among the most 
226

concentrated pesticides detected, but they are almost always found in mixtures with insecticides 

and fungicides (Gilliom 2007, Schafer et al. 2007). Although it is becoming more apparent that 

herbicides are often found in mixtures, the toxicity of the herbicides within those mixtures may 

depend on the composition of the mixture itself. When Van den Brink et al. (2009) examined the 

effects of a simple herbicide-insecticide mixture on mesocosm communities, they found the 

herbicide (atrazine) had fewer effects than expected, and they suggested that this effect may have 

been due to the reduced grazing pressure that resulted directly from the reduction in invertebrates 

caused by the insecticide (lindane). In a series of experiments comparing effects of single 

herbicides, single insecticides and mixtures of these, Relyea (2009) found that a mixture of five 

herbicides had relatively few effects on mesocosm communities compared to several individual 

insecticides, a mixture of 5 insecticides and a mixture of all 10 pesticides. One effect he did find 

was that chlorophyll concentrations in phytoplankton were similarly reduced after 16 days in 

both the acetochlor-alone treatment as well as the 5-herbicide mixture treatment. This suggests 

acetochlor alone, and not the other four herbicides, likely contributed to the overall toxicity of 

the mixture for this response variable. It is unclear, however, how other communities exposed to 

the numerous possible combinations of mixtures would respond. Finally, in addition to the 

composition of the mixture, the dose of the mixture may also be important in determining the 

direction of effect. In a study on eelgrass, low concentrations of a mixture of three herbicides 

(glyphosate, benzatone, and MCPA) were synergistic but high concentrations had an antagonistic 

effect (Nielsen and Dahllof 2007).

A final consideration and uncertainty in how herbicides may impact salmonids and their habitats 

is the question of resiliency of these aquatic ecosystems. The recovery of primary and secondary 

production – to rates observed prior to exposure – depends on the communities themselves and 

the exposure. For instance, if herbicides persist in the landscape, exposures may occur 

repeatedly (or continuously) depending on application rate, precipitation, and conditions in the 

watershed. Michael et al. (2006) found exposures of sulfometuron occurred repeatedly, due to 

wash off from the upstream forest after a single application (see also (Michael et al. 1999, 

Michael 2003). The persistence of an herbicide can affect the recovery of a community, as seen 

when the herbicide 3,4-dichloroaniline was added to mesocosms (Maund et al. 2009). This 

herbicide was initially added at a concentration equal to the median LC50 value of taxa in the 
227

mesocosms, but it persisted several months (median dissipation time was estimated at 30 d). The 

lack of recovery of populations within the mesocosms by 10 months and the delay of recovery 

even when colonists were added following exposure was attributed to the persistent toxicity 

(Maund et al. 2009). Generally, photosynthesis has been found to resume rapidly once exposure 

stops, while indirect effects on longer-lived taxa can persist much longer (Brock et al. 2000, 

Brock et al. 2004). This difference can lead to dynamics in trophic interactions (e.g., alterations 

between top-down and bottom-up control). These fluctuations have been found to stabilize in 

mesocosms within weeks to months, but for juvenile salmonids that require reliable food 

resources daily, this time period of recovery may be too long.

These uncertainties make it difficult to predict how herbicides will affect salmonids and their 

critical habitats, but they do not change NMFS’ assessment that there may be an adverse impact. 

8.5.8. Toxicity of Thiobencarb (Assessment Endpoints)

8.5.8.1. Direct Effects to Salmonids

We evaluate effects to salmonids based on toxicity information presented in the salmonid BE 

(EPA 2002a), the more recent California red-legged frog BE (EPA 2009b), REDs (EPA 1997a), 

ECOTOX database, and open literature. 

8.5.8.2. Survival

Individual survival is typically measured by incidences of death following 96-hour (h) exposures 

(acute test) and incidences of death following 21-day (d), 30-d, 32-d, and “full life cycle” 

exposures (chronic tests) to a subset of freshwater and marine fish species reared and exposed in 

laboratories under controlled conditions (temperature, pH, light, salinity, etc.) (EPA 2004).

Lethality of the pesticide is usually reported as the median lethal concentration (LC50), the 

statistically-derived concentration sufficient to kill 50% of the test population. It is derived from 

the number of surviving individuals at each concentration tested following a 96 h exposure and is 

usually estimated by probit or logit analysis and more recently by non-linear curve fitting 

techniques. Ideally, to maximize the utility of a given LC50 study, a slope, variability around the 

LC50, and a description of the experimental design, such as experimental concentrations tested, 
228

number of treatments and replicates used, solvent controls, etc., are needed. The slope of the 

observed dose-response relationship is particularly useful in interpolating incidences of death at 

concentrations below or above an estimated LC50. The variability of an LC50 is usually depicted 

by a confidence interval (95% CI) or error (standard deviation or standard error) and is 

illustrative of the degree of confidence associated with a given LC50 estimate (i.e., the smaller the 

range of uncertainty, the higher the confidence in the estimate). Without an estimate of the 

variability, it is difficult to infer the precision of the estimate. Furthermore, survival experiments 

are of most utility when conducted with the most sensitive life stage of the listed species or a 

representative surrogate. 

In the case of ESA-listed Pacific salmonids, there are several surrogates including hatchery 

reared coho salmon, Chinook salmon, steelhead, and chum salmon, as well as rainbow trout14. 

Unfortunately, slopes, estimates of variability for an LC50, and experimental concentrations 

frequently are not reported. In our review of the BEs, we did not locate any reported slopes of 

dose-response curves. Consequently, to insure that EPA’s action is not likely to jeopardize listed 

species, we must select LC50s from the lower range of available studies. We evaluate the 

likelihood of concentrations that are expected to kill fish and apply qualitative and quantitative 

methods to infer population-level responses of ESA-listed salmonids within the Risk 

Characterization section. Thiobencarb (in technical products and formulations) has been tested 

extensively in acute lethality toxicity tests with numerous fish species (Table 51). EPA reported 

some of these data in Appendix I of the California red-legged frog BE (EPA 2009b).

Data were typically from registrant submitted guideline studies and open literature studies. EPA 

constructed a species-sensitivity distribution for acute LC50s from studies they deemed 

acceptable in Appendix I, pg. 8 (EPA 2009b). 96 h LC50s ranged from 260 μg/L for white 

sturgeon (Acipenser transmontanus) to 13,200 μg/L for the European eel (Anguilla Anguilla). 

Most data were from 96 h tests, although some were from 24 h, 48 h, and 72 h tests. Rainbow 

trout (LC50 790 – 1,200 μg/L, n=5), steelhead (LC50 = 790, n=1), and Chinook salmon (LC50

 

14 Rainbow trout and steelhead are the same genus species (Oncorhynchus mykiss), with the key differentiation that 

steelhead migrate to the ocean while rainbow trout remain in freshwaters. Rainbow trout are therefore good 

toxicological surrogates for freshwater life stages of steelhead, but are less useful as surrogates for the life stages 

that use estuarine and ocean environments.
229

760, n=1) were among the more sensitive species tested based on lethality. Many LC50 assays on 

freshwater and estuarine fish species in California were conducted with thiobencarb-containing 

formulations to evaluate potential aquatic toxicity from rice growing operations in California’s 

Central Valley (Harrington 1990). Additionally, three species of atherinid fishes (coastal and 

estuarine species) were exposed to 96 h thiobencarb in static and flow through systems at day-ofhatch, 7 d, and 14 d (Borthwick et al. 1985). Treatment levels were verified analytically. Acute 

sensitivity was greatest for 7-d old fish where LC50s ranged from 204-464 μg/L flow through 

and from 396-483 μg/L static (Borthwick et al. 1985). Based on EPA toxicity categories for 

acute LC50s, EPA classified thiobencarb as highly toxic to salmonids.

Table 51 Thiobencarb toxicity values (μg/L) for aquatic organisms and plants reported in EPA 

salmonid BE, CRLF BE, RED, IRED, EFED science chapter, and ECOTOX. Abbreviations as 

follows: a.i. = active ingredient; NR = Not Reported; T= Technical grade; F = Formulated product; 

sw = estuarine/marine species; [ ] = 95% Confidence interval.

Assessment 

Endpoint

Concentration (μg/L aquatic tests; lbs a.i./acre terrestrial tests) 

Thiobencarb

Assessment 

measure

> 95% a.i.

(% a.i.)

< 95% a.i.

(% a.i.)

Survival

Source(s): 

EPA RLF 

2009

Appendix I; 

ECOTOX # 

12136;

ECOTOX #

15172

EPA RLF 

2009

Appendix I;

MRID #s 

139051, 

00080851, 

00080851, 

00050664

Registrant 

submitted

salmonid LC50

(96 h) Rainbow trout (O. mykiss) (95.5%; T) = 

1150, formulation IMC-3590

Rainbow trout (O. mykiss) (85.2%; F) = 1200, 

[700-1600], Bolero 8EC formulation

Rainbow trout (O. mykiss) (84%; F) = 1050, 

Bolero 8EC formulation

Rainbow trout (O. mykiss) (10%; F) = 1500, [1200-

1900] Bolero G (granular)

Rainbow trout, v. Donaldson trout (O. mykiss) (% 

a.i. not reported)= 1200

Chinook Salmon (O. tshawytscha) (% 85.2; F) = 

760; Bolero EC 

Steelhead (O. mykiss) (% 85.2; F = 790; Bolero EC

Chinook salmon (O. tshawytscha) (85.2%; F): 

survival of fry NOEC = 140, LOEC = 250; Bolero 

8EC; E0015472 (Faggella and Finlayson 1988)

Survival Non-salmonid 

freshwater, 

estuarine, and 

Bluegill sunfish (L. macrochirus) (95.5%; F) 

= 2480, IMC-3590

Bluegill sunfish (Lepomis macrochirus) (94%; T) 

= 2600 
230

Assessment 

Endpoint

Concentration (μg/L aquatic tests; lbs a.i./acre terrestrial tests) 

Thiobencarb

Assessment 

measure

> 95% a.i.

(% a.i.)

< 95% a.i.

(% a.i.)

marine fish LC50

(96 h) Bluegill sunfish (L. macrochirus) (10%; F) = 560 

[330-1200], Bolero 10G (granular)

Bluegill sunfish (L. macrochirus) (84%; F) = 1660, 

Bolero 8EC

Bluegill sunfish (L. macrochirus) (85.2%; F) = 

1700 [1200-2300], Bolero 8EC

Striped bass (Morone saxatilis) (sw) (85.2%; F):

Mean = 731; median = 760; range = 430-1000; 

n=17; Bolero 8EC, ages 7 – 45 post emergence; 

Bolero 8EC; Fujimura et al. 1991

White sturgeon (Acipenser transmontanus) (T) = 

260 [230-300], Bailey 1984 MRID #40651315

Reproduction 

or larval 

survival

NOEC/LOEC Fathead minnow (Pimephales promelas) 

(96.5%; T) = 53/110; 2 replicates/treatment 

only; lifecycle test (260 day exposure); 

Endpoints affected: survival, growth of F0 

generation, # eggs per spawn, hatching 

success, growth, survival F1 generation 

MRID 45695101

Sheepshead minnow (C. variegates) (sw) 

(95.2 %; T) = <150/150 28 d post hatch

survival

370/600 hatching success

MRID 00079112

 

Fish growth

and 

development

NOEC/LOEC Fathead minnow (Pimephales promelas) 

(96.5%; T) = 53/110; 2 replicates/ treatment; 

lifecycle test (260 day exposure); Endpoints 

affected: survival, growth of F0 generation, # 

eggs per spawn, hatching success, growth, 

survival F1 generation MRID 45695101

Zebrafish (Danio rerio)(96.5%; T) = 800 

LOEC, 96 h exposure, 100% abnormality and 

reduced length

Appendix 5, (NMFS 2011c)

Chinook salmon (O. tshawytscha) (85.2%; F) = 28/ 

49; survival of fry 140/ 250 Bolero 8EC; 88 day 

exposure 28 day pre hatch through 60 day post 

hatch. E0015472 Fujimura et al 1991 

Striped bass (M. saxatilis) (sw) (85.5%; F):

NOEC=21, LOEC=36; NOEC=<23, LOEC=23; 

NOEC=58, LOEC= 91 (n=3); Bolero 8EC, reduced 

dry weight, early life stage study.

Fujimura et al 1991 
231

Assessment 

Endpoint

Concentration (μg/L aquatic tests; lbs a.i./acre terrestrial tests) 

Thiobencarb

Assessment 

measure

> 95% a.i.

(% a.i.)

< 95% a.i.

(% a.i.)

Fish 

respiration

gill damage

(micropathological 

assessment)

Mosquitofish (Gambusia affinis), (Bolero 8EC15; 

formulation): @ 19, 90, 576, 1200, 1800 μg/L 

Bolero8EC inflamed gills, dose dependent 

response

Habitatsalmonid prey invertebrate 

survival

(48 h EC/LC50)

Amphipod (Gammarus pseudolimnaeus) 

(95.5%; F) = 720, IMC-3590

Midge (Chironomus tentans) (97.2%;T) = 

364 [322-413], probit slope = 3.42 [2.88-

4.16]; 

Survival NOEC = 135

Opossum shrimp (Neomysis mercedis) (sw)

(NR; T): 

96 h LC50 = 304; 7 d LC50 = 214; 

14 d LC50 = 91

(Bailey 1993)

Water flea (Daphnia magna) (94.4%) = 101.2 

[73.8 – 138.7]

Water flea (D. magna) (82.25%; F) = 210.7 [175.7 

– 252.7] Bolero 8EC

Water flea (D. magna) (85.2%; F) = 1200 [400-

3100] Bolero 8EC

Amphipod (G. pseudolimnaeus) (85.2%; F) = 1000 

[600-1700], Bolero 8EC

Amphipod (G. pseudolimnaeus) (84%; F) = 1000, 

Bolero 8EC

Midge (C. tepperi) (% not reported,; F) = 188 

(LOEC emergence); 375 (LOEC developmental 

time, wing length) (Burdett et al. 2001)

Invertebrate 

reproduction

NOEC/LOEC (21

d life-cycle test)

Water flea (D. magna) (96.9%; T) = 48.0/ 

90.0; # of offspring produced; MRID 

00079098

Water flea (D. magna) (96.2%; T) = na/ 38; 

# of offspring produced; MRID 4163610116

Water flea (D. magna) (95.2%; T) = 1.0/ 3.0; 

# of offspring produced; MRID 241483

Midge (Chironomus riparius) (97.2%; T) = 

180/ 420; 28 day sediment toxicity assay 

Opossum shrimp (N. mercedis) (sw) (NR; T): 

3.2/6.2 (Bailey 1993)

Mysid shrimp (Mysidopsis bahia) (sw) 

 

15 Thiobencarb per cent not reported in paper. Other Bolero 8EC studies report 85.2% thiobencarb.

16 MRID 41636101, McNamara, P. (1990) Bolero Technical: The Chronic Toxicity to Daphnia magna under FlowThrough Conditions: Lab Project Number: 90- 8-3444. Unpublished study prepared by Springborn Labs, Inc. 80 p. 

The experiment showed effects at the lowest concentration tested, 38 ug/L, therefore a NOEC cannot be determined.
232

Assessment 

Endpoint

Concentration (μg/L aquatic tests; lbs a.i./acre terrestrial tests) 

Thiobencarb

Assessment 

measure

> 95% a.i.

(% a.i.)

< 95% a.i.

(% a.i.)

(95.1%): NOEC not determined, EC05 = 9.8; 

survival of offspring17

Habitat: 

Riparian 

Vegetation

MRID 

41690902 

Hoberg, J.R. 

1990 

Sup 

Vegetative vigor

(lbs a.i./acre):

Monocots (96.6%; T) 

EC25: corn = 2.2, oat = 0.17, onion = 1.2, 

ryegrass = 0.073; 

NOEC: corn = 2.2, oat = 0.12, onion = 0.80, 

ryegrass = 0.02

Dicots (96.6%; T):

EC25: carrot = > 2.2, cabbage = 1.2, 

cucumber = na, lettuce = 1.3, soybean = 1.2, 

tomato = 1.8; 

NOEC: carrot = 2.2, cabbage = 1.4, 

cucumber = <0.12, lettuce = 0.80, soybean =

0.80, tomato = 2.2 

Habitat: 

Riparian 

Vegetation 

Sources:

MRID 

41690902 

Hoberg, J.R. 

1990; 

MRID 

44846201 

Chetram, R.S. 

1999 

Acceptable 

DER 

11/16/2002 

Seedling 

emergence 

(lbs a.i./acre): 

shoot length or 

mortality EC25 

Monocots (96.6%; T):

-Shoot length EC25: corn = >1.7, oat =

0.086, onion = 2.0;

NOEC: corn =1.7, oat = 0.055, onion 0.94;

-Mortality (rye grass)

EC25 = 0.019; NOEC = 0.0051 (17% 

mortality at 0.011)

Dicots (96/6%; T):

-shoot length EC25: carrot = >3.1, cabbage =

0.082, cucumber = >1.7, soybean = >1.7, 

tomato = 1.1;

NOEC: carrot = 2.1, cabbage = 0.071, 

cucumber = 0.16 soybean = 0.94, tomato =

0.94. 

-mortality (lettuce); EC25 = 0.27

Habitat: 

In-stream 

Primary 

Productivity

Source MRID 

41690901 

Giddings, 1990 

DER 

09/18/1995 

Aquatic plant 

growth: cell 

density EC50 & 

NOEC

Green algae (Selenastrum capricornutum) 

(96.6%; T) 120 hr EC50 = 17 [12-26], NOEC 

= 13

Freshwater diatom (Navicula pelliculosa) 

(96.6%; T) 120 hr EC50 = 380 [240-610]; 

NOEC = 65

Duckweed (Lemna gibba) (96.6%; T) 14 day 

frond production EC50 = 770 [380-1600]; 

NOEC = 140

 

17 EPA RED for thiobencarb concluded that a NOEC could not be determined because control had no replication. A 

nonlinear regression analysis was used by EPA to calculate an EC05 which was used in lieu of the NOEC.
233

Assessment 

Endpoint

Concentration (μg/L aquatic tests; lbs a.i./acre terrestrial tests) 

Thiobencarb

Assessment 

measure

> 95% a.i.

(% a.i.)

< 95% a.i.

(% a.i.)

Green algae (Scenedesmus acutus) (100%; 

analytical standard) 96 hr EC50 = 17 [16-19] 

(Sabater and Carrasco 1996)

Green algae (Chlorella saccharophila) 

(100%; analytical standard) 96 hr EC50 = 

4000 [3800-4100] (Sabater and Carrasco 

1996)

Algae (Pseudanabaena galeata) (100%; 

analytical standard) 96 hr EC50 = 370 [350-

400] (Sabater and Carrasco 1996)

A species sensitivity distribution based on reported 96 hr LC50s shows the position of salmonids 

in relation to other species tested
234

). The distribution contains those studies that EPA reviewed and ranked as acceptable, therefore 

not all fish LC50s are included. Salmonids are at the lower end of the distribution indicating they 

are more sensitive than many of the other species tested based on 96 hr lethality assays. Based 

on the confidence bands (shown in gray) around the means, substantial variation exists between 

and among species.
235

Figure 37. Species sensitivity distribution of freshwater fish 96 h LC50s derived from EPA 

denoted "acceptable" studies18

8.5.8.3. Growth and Reproduction 

Growth and reproduction endpoints are typically evaluated in FIFRA guideline tests conducted

by the chemical company registrants. In these tests, fish are exposed to the a.i. for variable 

durations depending on species tested. Fish are fed twice daily, ad libitum (i.e., an over 

abundance of food is available at time of feeding). The lowest concentration eliciting a 

statistically significant difference from controls (no treatment) to growth or reproductive 

endpoints is recorded (i.e., the Lowest Observable Effect Concentration (LOEC)), as well as the 

lowest exposure concentration tested that is not different than the control (i.e., the No Observable 

Effect Concentration (NOEC)). Many researchers have commented on the poor application of 

environmental statistics and laboratory testing regarding NOECS and LOECs (Laskowski 1995, 

Chapman 1996, Kooijman 1996), (Suter 1996) and (Landis and Chapman 2011). Prominent 

limitations include: (1) NOECs and LOECs are statistically derived, a function of the 

 

18 EPA developed this SSD using the Species Sensitivity Generator v.1.0, Causal Analysis/Diagnosis Decision 

Information System (CADDIS), EPA 2009a. Plot shows curve fit with bounds in gray, however no information was 

provided on how curve fit was done or what bounds represent.
236

concentrations selected by the experimenters, and are inconsistent between studies; and (2) 

NOECs and LOECs ignore the fundamental model of toxicology; (3) ignore critical data at other 

treatments; (4) use a lack of evidence as a no-effect; and (5) are limited to the concentrations 

tested. NOECs typically correspond to an EC10 to EC30 on an exposure-response curve (Moore 

and Caux 1997). A 30% affect rate within a population can be striking, particularly if the effect 

is on a critical biological endpoint such as reproduction, growth, migration, or olfaction. 

Previous salmonid population modeling suggests that when 14% mortality occurs to juveniles 

population growth rate is substantially affected (NMFS 2009e) . We therefore exercise caution in 

interpreting a NOEC as a true “no response” to an organism. 

Growth of individual organisms is an assessment endpoint derived from chronic fish and 

invertebrate toxicity tests summarized in the BE. Reproduction, at the scale of an individual, can 

be measured by the number of offspring per female (fecundity), and at the population scale by 

measuring the number of offspring per female in a population over multiple generations. The 

BEs summarized reproductive endpoints at the individual scale from chronic, freshwater fish 

experiments. Other assessment measures of reproduction include egg size, spawning success, 

sperm and egg viability, gonadal development, and hormone levels--most of which are rarely 

measured in standardized toxicity tests conducted pursuant to pesticide registration.

8.5.8.4. Respiratory Effects

Gills are a well known target of toxic insult in fish and respond quickly to degraded 

environmental conditions (Poleksic and Mitrovic-Tutundzic 1994). Following 7-d static 

exposures to Bolero 8EC at concentrations ranging from 19 – 1900 μg/L (~16.2-1620 μg/L 

thiobencarb; based on 85.2% by volume), gills of mosquitofish were inflamed (Persichino et al. 

1998). Upon examination by light microscopy, the inflammatory response was characterized by 

hypertrophy of respiratory epithelium, hyperplasia of mucous cells, and empty mucous cells 

(Persichino et al. 1998). At concentrations of 1200 μg/L and above, mosquitofish died. At 19 

μg/L Bolero 8EC, two of three fish perished due to combination of parasites and Bolero—no 

parasites were observed in control fish. Limitations in applying these experimental results 

include lack of verification of treatment concentrations by analytical chemistry methods, small 
237

treatment size, n=3, and no replication. Although the study’s limitations reduce its significance, 

the data do suggest adverse responses in fish gills which may be an additional mechanism of 

action.

8.5.8.5. Acetylcholinesterase Inhibiting Effects 

We located no salmonid studies that tested thiobencarb exposure on AChE activity, thus we 

broadened the search to other fish species. A suite of experiments evaluated the effects of 

thiobencarb on European eels, also referred to as yellow eels (Fernandez-Vega et al. 1999, 

Fernandez-Vega et al. 2002, Bretaud et al. 2001, Sancho et al. 2000, Babu et al. 1989).

Eels were fairly insensitive to thiobencarb’s capacity to kill them over 96 h, LC50s ranged from 

13.3 – 25 mg/L (Fernandez-Vega et al. 1999). Eel nervous systems were much more sensitive 

and experienced adverse effects at 1/60th of the LC50, a concentration of 220 μg/L. Juvenile eels

exposed for 96 h to thiobencarb in flow through conditions (continuous exposure) had more than 

a 50% reduction in their blood AChE activity. Statistically significant inhibition occurred at the 

first time point measured, 2 h, and continued through the 96 h. Normal activity returned at 72 h 

post-exposure indicating, similar to other carbamates, that AChE activity is reversible following 

exposure to carbamates. Eels also showed signs of lethargy and agitation during the experiment, 

common symptoms of nervous system toxicants. Brain, muscle, and gill AChE activities were 

inhibited by 30 – 50% following 96 h of continuous exposure (Fernandez-Vega et al. 2002). 

Recovery occurred within eight days for brain and muscle activities, while gill AChE activity 

had minimal recovery of < 29 %. Eels also exhibited behavioral impacts including agitation, loss 

of equilibrium, increased respiratory rhythm, uncoordinated swimming movements, and many 

remained at the bottom of exposure tanks. These aberrant behaviors were not observed in 

control exposures.

AChE activity in eel eyes was inhibited by thiobencarb at 220 μg/L (Sancho et al. 2000). 

Following 220 μg/L exposures (1/60th of the 96 hr LC50) at 2, 12, 24, 48, 72, and 96 h, total eye 

AChE and specific eye AChE were measured. Eye AChE was inhibited in a time-dependent 

manner with onset of inhibition occurring at 2 h and remaining throughout the exposure period. 

Inhibition occurred at 2 h and did not return to its normal levels over the six day recovery period 
238

in clean water. Maximal inhibition occurred at 12 h with more than 80% of activity inhibited. 

Eels were lethargic, hypoactive, and exhibited reduced opercular movements. Following a 

review of these findings, we found several limitations including the selection of the eye as a 

target for AChE inhibition, the methodology used to detect cholinesterase activity, the levels of 

activity within the eye, reported standard deviations of 0.00, and the apparent lack of a positive 

control. In the vertebrate retina, acetylcholine is used as the neurotransmitter for only a subset of 

one class of neurons, the amacrine cells. The majority of the neurons (e.g., rods, cones, 

horizontal, bipolar) use other neurotransmitters such as glutamate, gamma aminobutyric acid 

(GABA), and glycine. Acetylcholine is likely the neurotransmitter used to control the muscles 

that control the iris. Thus, only a small portion of the eye is likely to require AChE, making the 

consequence of AChE inhibition to overall eye function difficult to assess and likely subtle. The 

method used to quantify AChE detects multiple types of cholinesterase activity. For example, 

the method does not distinguish between AChE and butyrylcholinesterase. The authors do not 

discuss the possibility that much of the inhibition they detect by thiobencarb may be occurring to 

an enzyme other than AChE. Additionally, many of the mean values for specific activity in the 

eye have a standard deviation of 0.00. It is difficult to believe that there was zero variability 

around the means. The results of this study do show an inhibitory effect (reduction in AChE 

activity) of thiobencarb on eel eyes, however further research would be needed to confirm these 

results. Due to these limitations, we have low confidence in these results.

As no studies were found on thiobencarb’s effect on salmonids, we broadened our search to other 

thiocarbamates including EPTC, molinate, pebulate, trillate, butylate, and cycloate. 

Unfortunately, no additional empirical data were located on AChE inhibition in salmonids or 

other fish. 

These studies showed that (1) biochemical endpoints (AChE inhibition in blood, brain, muscle, 

gill and eyes) were inhibited in eels at concentrations well below the 96 hour LC50, at 1/60th of 

eel LC50; (2) adverse behavioral changes (lethargy) occurred at that concentration, (3) effects 

occurred within a couple of hours after exposure; and (4) recovery of AChE activity is tissue 

dependent (minimal recovery seen in whole eyes over a six day period and recovery in plasma 

activity occurring in three days). Although thiobencarb exposure to 220 μg/L would not kill eels
239

outright, it could cause disorientation, loss of visual competence, and loss of locomotion within a 

few hours of exposure, which may lead to predation, reduced foraging, and impaired migration. 

We recognize that listed salmonids’ AChE activity following exposure to thiobencarb is an 

uncertainty given that no experiments were identified that tested this response. We expect that 

thiobencarb will inhibit AChE, although the concentration that would induce this effect remains 

unknown. European eels, the only fish species we have AChE activity toxicity data for, are 

among the most tolerant of the various fish species for which LC50s are available. If the 1/60 

factor is presumed to apply to salmonid species, similar effects could be expected at roughly 12–

19 μg/L, based on the median LC50 values for Chinook salmon and rainbow trout (Table 51).

8.5.8.6. Swimming

Swimming is a critical function for anadromous salmonids necessary to complete their lifecycle. 

Impairment of swimming may affect feeding, migrating, predator avoidance, and spawning. It is 

the most frequently assessed behavioral response of toxicity investigations with fish (Little and 

Finger 1990). Swimming activity and swimming capacity of salmonids have been measured 

following exposures to a variety of AChE-inhibiting insecticides including the OPs chlorpyrifos, 

diazinon, and malathion (reviewed in NMFS 2008) and the carbamates carbaryl, methomyl, and 

carbofuran (reviewed in NMFS 2009). Swimming capacity is a measure of orientation to flow as 

well as the physical capacity to swim against it (Howard 1975, Dodson and Mayfield 1979). 

Swimming activity includes measurements of frequency and duration of movements, speed and 

distance traveled, frequency and angle of turns, position in the water column, and form and 

pattern of swimming. A review paper published in 1990 summarized many of the experimental 

swimming behavioral studies and concluded that effects to swimming activity generally occur at 

lower concentrations than effects to swimming capacity (Little and Finger 1990). Therefore, 

measurements of swimming activity are usually more sensitive than measurements of swimming 

capacity. A likely reason is that fishes with impaired swimming to the degree that they cannot 

orient to flow or maintain position in the water column are moribund (i.e., death is imminent). 

The authors of the review also concluded that swimming-mediated behaviors are frequently 

adversely affected at 0.3 – 5.0 % of reported fish LC50s, and that 75% of reported adverse effects 

to swimming occurred at concentrations lower than reported LC50s. Both swimming activity and 
240

swimming capacity are adversely affected by AChE inhibiting insecticides. We located no 

studies that measured impacts to salmonid swimming behaviors from exposure to thiobencarb. 

Several studies on eels (discussed previously) showed AChE inhibition at 1/60th of the LC50 (200 

μg/L) and observed that affected fish were lethargic, hypoactive, and exhibited reduced 

opercular movements, all classic symptoms of AChE poisoning.

8.5.9. Indirect Effects to Salmonids (Prey effects and Habitat Modifications)

Indirect effects on salmon from exposure to thiobencarb include reductions in prey base (aquatic 

invertebrates), disruptions in primary productivity in salmon habitats (phytoplankton and 

macrophytes), and effects on riparian vegetation (reduced cover and increased sedimentation). 

8.5.9.1. Effects to Salmonid prey

Prey Survival. Data from several freshwater and marine/estuarine aquatic invertebrates are 

available to determine their sensitivity to thiobencarb and presented in Table 51. Ranges in 

survival EC50s were observed for both freshwater species (~100-1200 μg/L, Table 51) and 

marine/estuarine species (150 – 4400 μg/L19). Several of the freshwater species tested are 

known salmonid prey including water fleas and amphipods. We located few studies that tested 

native macro-invertebrates to CA’s Central Valley. This is a notable data gap for effects to 

salmonid prey following exposure to thiobencarb.

Reproduction and Growth. We located four laboratory studies on the effects of thiobencarb to 

two freshwater species, a water flea and a chironomid, both of which are fed upon by young 

salmonids. The FIFRA guideline protocol for a 21-d chronic test on reproduction and growth 

was carried out in three separate studies with D. magna. The number of offspring was reduced 

by thiobencarb at 3, 38 and 90 μg/L (LOECs) with corresponding NOECs of 1, not determined,

and 48 μg/L in the three studies, respectively. In the study with a LOEC of 38 ug/L (MRID 

41636101), the NOEC could not be determined because the lowest concentration tested, 38 ug/L, 

 

19 Individual values obtained from EPA Registration Eligibility Decision, page 43, Table 17.EPA 1996. 'Ecological 

Effects Test Guidelines OPPTS 850.4000, Background -Nontarget Plant Testing.' in Ecological Effects Test 

Guidelines OPPTS 850.4000, Background -Nontarget Plant Testing, 15.
241

elicited a statistically significant effect. Growth endpoints were not reported in EPA documents. 

The limitations with NOECs were discussed previously and should be used with caution as 

typically they are not a true “no effect” concentration. Thus we expect some level of effect at 1

and 48 μg/L for water flea reproduction. In the chironomid study, sediments were spiked with 

thiobencarb and chironomids were exposed to contaminated sediments for 28 days. The 

treatment levels for aqueous thiobencarb representing the NOEC and LOEC were 180 and 420 

μg/L, respectively (Table 51). Other studies summarizing study designs and results with aquatic 

species can be found in Table 52.

Table 52 Study designs and results with aquatic species

Taxa/species Assessment 

measures

Concentrations 

tested (μg/L)

Exposure 

duration

Effects Data source

Freshwater 

pond 

invertebrate 

communities

Species richness 

and abundance 

of groups: 

Corixidae, 

Ostracoda, larval 

Hydrophilidae, 

larval Dytiscidae, 

adult Dytiscidae, 

chironomids, 

ceratopogonids, 

calanoids, 

cyclopoids

Sprayed once at 

3.2 lbs/acre 

(3600 g/HA) = 

2700 μg/L;

3 replicates

5 weeks; 

sampled 

at 1 and 5 

wks

@ 1 wk reduced 

abundance of 

chironomids, 

calanoids, and 

cyclopoids; 

@ 5 wks highest 

and lowest species 

richness and 

abundance in 

thiobencarb 

treatments

(Burdett et 

al. 2001)

Americamysis 

bahia (mysid)

Life cycle test: 

life stagespecific survival 

and 

reproduction;

Modeled 

population 

impacts

0, 22, 35, 76, 

181, 345, 734

28 days 22 μg/L = no 

effects;

35 μg/l = delayed 

reproduction;

76 μg/L=reduced 

overall 

reproduction; 

181 μg/L = 

reduced overall 

reproduction; 

345 μg/L = 

reduced overall 

reproduction and 

survival; 

734 μg/L= reduced 

overall 

reproduction and 

survival;

Population 

modeling results: 

@ 76 μg/L and 

above population 

declines 

anticipated 

(Raimondo 

and 

McKenney 

2006)
242

Taxa/species Assessment 

measures

Concentrations 

tested (μg/L)

Exposure 

duration

Effects Data source

Ceriodaphnia 

dubia (daphnid)

48 h survival 

tests with field 

collected water 

toxicity 

identification 

evaluation (TIE);

24 h & 48 h 

survival EC50 

with spiked 

laboratory water

No thiobencarb 

detected in fieldcollected water, 

TIE experiments; 

Methylparathion, 

carbofuran, 

molinate 

detected

Treatment 

concentration not 

reported

48 h

24 & 48 h 

exposures

No reported effects 

from thiobencarb 

as none detected 

in samples. 

Carbofuran and 

methyl parathion 

explained toxicity 

24 h survival EC50

= 580 μg/L (95%CI 

430-790)

48 h survival EC50

= 510 μg/L (95% 

CI 400-650)

(NorbergKing et al. 

1991)

8.5.9.2. Aquatic Plants (Phytoplankton and Vascular Plants)

As an herbicide, thiobencarb may impair growth or kill aquatic and riparian plants within salmon 

habitats. Table 51 summarizes toxicity data for aquatic primary producers including various 

types of algal species and vascular plants. Very few data were identified and discussed by EPA 

regarding the effects of thiobencarb on salmonid habitats. The available information is sparse, 

non-specific, and difficult to translate to habitat assessment endpoints. For example, 

phytoplankton provides energy to aquatic systems, while macrophytes provide structural 

components such as attachment sites for aquatic invertebrates and refugia for salmonids. Data 

from laboratory bioassays on one species of algae exposed to thiobencarb fail to encompass the 

complexity of aquatic systems where hundreds of algae make up but one component of a food 

web. We located five studies on five species of algae that report reductions in growth as 

measured by cell density. The EC50s (derived from 48-120 h exposures) ranged from 17- 4000 

μg/L, and include three species green algae, a cyanobacteria, and a freshwater diatom. 

Based on these data, species of green algae were the most and least sensitive (i.e., 17 μg/L and 

4000 μg/L). Algae serve as important food for zooplankton and early life stages of aquatic 

invertebrates. If these test species are representative of aquatic systems, then reductions in 

primary productivity or modifications in community structure via removal of sensitive species 
243

may result in “bottom-up” trophic cascades which may adversely affect salmonids. It is difficult 

to determine the level of effect as numerous site and species characteristics drive community 

responses.

Loss of structure provided by macrophytes may result in decreased populations of aquatic 

invertebrates or increased predation on juvenile salmonids. One experimental result is available 

with the aquatic vascular plant, duck weed (Lemna gibba), that addresses loss of macrophytes 

within aquatic systems. Thiobencarb is predicted to reduce frond production by 50% at 14 days 

following an exposure of 770 μg/L with a NOEC of 170 μg/L. We located no other studies with 

aquatic plants. This is a significant data gap given the importance of aquatic plants to healthy 

salmonid habitats.

8.5.9.3. Effects on Riparian Vegetation

Riparian vegetation is important for providing shade to the stream, stabilizing the stream banks, 

reducing sedimentation, and providing organic material inputs, both in terms of plant material 

and terrestrial insects. Riparian vegetation is a major focus of restoration efforts within 

California, and when present can reduce pesticide loading into aquatic resources. Riparian 

vegetation is an important assessment endpoint for herbicidal impacts on salmon habitats. 

Generally there are sparse data regarding the effects of herbicides on wild plants within riparian 

systems, other than weed species. EPA requires submission of crop effects data as part of the 

registration process (EPA 1996). This information currently provides the only basis for 

evaluating effects on herbaceous plants unless data are available from other sources. Registrantsubmitted guideline studies report vegetative vigor and seedling emergence EC25s for 

thiobencarb end-use products (EPA 1996). The overall assumption is that the sensitivity of plant 

species tested (typically plants used in agriculture) in the guideline studies will be representative 

of riparian species. There is no way to know this is the case, therefore a high degree of 

uncertainty regarding the toxicity of thiobencarb to riparian vegetation exists. We note that a 

Science Advisory Panel in 2001 (now more than 11 years ago) was convened by EPA to review 

non-target plant toxicity tests under the North American Free Trade Agreement. A host of 

recommendations were made as well as several research recommendations to fill identified data 
244

gaps. We were not able to confirm if any of the recommendations or research initiatives were 

undertaken although if implemented several would have informed the current consultations with 

effects on thiobencarb to aquatic and riparian plants (EPA 2002b).

We calculated medians for the range of crop data presented in the EPA assessments, and these

are presented in Table 51. Overall, tested monocots (grasses) appear more sensitive than dicots

(broadleaf plants) to thiobencarb. For monocots, the seedling emergence endpoint is more 

sensitive than the vegetative vigor based on NOECs, but based on EC25s, the sensitivity of the 

two endpoints is similar. Thiobencarb affected terrestrial crops at concentrations as low as 0.019 

lbs/acre indicating high sensitivity of some plants. Vegetative vigor, shoot length, mortality, and 

seedling emergence were all affected at application rates well below authorized label rates. In 

the tier II seedling emergence test, mortality of test plants occurred in the tests with ryegrass and 

lettuce. Mortality was the most sensitive toxic endpoint for these species (plants tended to die 

shortly after emerging). The most sensitive species was ryegrass, a monocot, for which the LC25

was 0.019 lb thiobencarb/acre. The most sensitive dicot was cabbage (EC25 based on shoot 

length was estimated to be 0.082 lb thiobencarb/acre). Based on the results of the terrestrial 

plant toxicity tests, it appears seedling emergence is the most sensitive endpoint for both dicots 

and monocots. No studies were located on the effects of thiobencarb on riparian vegetation, a 

significant uncertainty. Additionally, no information was located on effects to established 

riparian vegetation.

8.5.10. Degradate Toxicity 

Two “major”20 degradates were identified in EPA’s California red legged Frog BE, 4-

chlorobenzoic acid and 4-chlorobenzaldehyde, for which EPA located no toxicity data based on 

an ECOTOX database search. Therefore, EPA estimated potential effects using a structure 

activity relationship model, ECOSAR version 1.0. The model predicted aquatic toxicity 

concentrations for eight degradates: 4-chlorobenzoic acid, 4-chlorobenzaldehyde, thiobencarb 

 

20 degradates are considered “major” by EPA if the degradate makes up more than 10% of the parent active 

ingredient applied following standardized, registrant-submitted laboratory degradation studies.
245

sulfoxide, 2-hydroxythiobencarb, 4-chlorobenzylmethylsulfone, 4-chlorobenzylmethylsulfoxide, 

bencarb, and desmethyl thiobencarb. 

The ECOSAR model estimated that the two major degradates, 4-chlorobenzoic acid and 4-

chlorobenzaldehyde, were less toxic than thiobencarb. As none of the other degradates assessed 

by ECOSAR were considered “major” degradates, EPA assumed that degradate toxicity would 

be minimal. Overall EPA concluded that none of the degradates would alter its effect 

determinations for the California Red-legged Frog (EPA 2009b).

Thiobencarb’s lowest empirical 96-hr LC50 for freshwater fish (white sturgeon) is 260 μg/L,

while the ECOSAR model predicted LC50s of 2.9 mg/L (an order of magnitude less toxic)

suggesting that the model under predicts toxicity. The model was also used to predict freshwater 

fish 96-hr LC50s for the two major degradates, 421.7 mg/L for 4-chlorobenzoic acid, and 5.5 

mg/L for 4-chlorobenzaldehyde. The lowest predicted degradate 96-hr LC50 was 1.6 mg/L for 

bencarb. It is difficult to determine what significance should be placed on EPA’s toxicity 

modeling results given the lack of discussion of the non-major degradates and the absence of a 

discussion with in the BE. Bencarb, a carbamate ester, was consistently the most toxic degradate 

compared to the other degradates across aquatic invertebrate and fish endpoints that were 

modeled using ECOSAR (Appendix B, (EPA 2009b)). Bencarb achieved levels just below 10% 

in degradation studies, i.e., 8.3% at 21 days and 8.1% at thirty days in a photolysis study. 

Chemical fate and persistence data were not provided. These data indicate that bencarb may be 

of aquatic concern to salmonids and their habitat, however definitive conclusions on its toxicity 

is complicated by the lack of empirical information.

8.5.11. Tank Mixtures 

Thiobencarb is often co-applied with other pesticides, including products that contain propanil, 

copper sulfate, and many others (Appendix 4). Several thiobencarb labels specifically 

recommend co-application with propanil. Propanil is an herbicide with a wide range of acute 

LC50s for freshwater fish (2.1– 12.7 mg/L) and invertebrates (1.2 – 16 mg/L) (EPA 2009a). 

Rainbow trout are the most sensitive freshwater fish tested with an LC50 of 2.1 mg/L. Propanil is 

highly toxic to aquatic (EC50 0.016 -0.11mg/L) and terrestrial primary producers (EC25 0.09-12 
246

lbs propanil/ acre). Propanil’s degradate, 3,4-DCA, is 15-fold more toxic than propanil based on 

D. magna EC50 survival test (48-h EC50 = 0.528 μg/L parent; 0.035 mg/L degradate). We 

located no fish toxicity studies with 3,4-DCA and no studies on the potential combinatorial 

toxicity of thiobencarb and propanil.

While labels do not specifically recommend co-application of thiobencarb and copper sulfate 

products, such mixtures are not prohibited and they are co-applied as tank mixes. In California, 

pesticide use report data indicate approximately 7% of all sites that receive thiobencarb 

applications are also treated with copper sulfate on the same day (Appendix 4). Dissolved 

copper is highly toxic to fish and their invertebrate prey, particularly the olfactory sensory 

system of salmonids (Baldwin et al. 2003a, Baldwin et al. 2011, De Boeck et al. 1997, Hansen et 

al. 1999, Hansen et al. 2002, Linbo et al. 2006, Linbo et al. 2009, McIntyre et al. 2008, Mebane 

and Arthaud 2010, Sandahl et al. 2004, Tierney et al. 2006, van der Geest et al. 2000, van der 

Geest et al. 2002). Olfaction conveys critical environmental information that fishes use to mate, 

locate food, discriminate kin, avoid predators, and home. Any or all of these essential olfactorymediated behaviors may be affected by exposure to copper. Numerous studies spanning several 

species have shown ecologically relevant exposures to copper can interfere with fish olfaction, 

disrupting life history processes that determine individual survival and reproductive success. For 

example, copper impairs and destroys salmonid olfactory sensory neurons in a matter of minutes 

at low μg/L levels and effects persist for hours to weeks depending on exposure concentration 

and duration (Baldwin et al. 2003b). Measured behavioral effects in salmonids from impaired 

olfaction include compromised alarm response, loss of ability to avoid copper, interrupted 

spawning migrations, loss of homing ability, and delayed and reduced downstream migration of 

juveniles (Baldwin et al. 2003b, Baldwin et al. 2011, Hansen et al. 1999, McIntyre et al. 2008, 

Mebane and Arthaud 2010, Sandahl et al. 2004). Disruption of these essential behaviors reduces 

the likelihood of an individual salmonid completing its life cycle. If sufficient numbers of 

individuals are compromised, a population’s abundance and productivity can be reduced. 

The effect of environmental mixtures of thiobencarb with propanil and copper sulfate is 

uncertain as toxicity information assessing these mixtures was not located. Additionally the type 

of mixture toxicity is difficult to predict as these compounds do not share a mechanism of action. 
247

Given that propanil is an authorized tank mix and copper sulfate is frequently applied the same 

day to the same area, we conclude that impacts of such mixtures to aquatic plant communities 

will be greater than the impact of thiobencarb alone. Additionally, laboratory studies suggest 

propanil and copper are more toxic to a broader range of aquatic flora and fauna than thiobencarb 

suggesting that in combination, community effects may be increased compared to thiobencarb 

alone. 

8.5.12. Adjuvant Toxicity

Although no data were provided in the EPA available documents related to adjuvant toxicity, an 

abundance of toxicity information is available on the effects of the alkylphenol polyethoxylates, 

a family of non-ionic surfactants used extensively in combination with pesticides as dispersing 

agents, detergents, emulsifiers, adjuvants, and solubilizers (Xie et al. 2005). Two types of 

alkylphenol polyethoxylates, nonylphenol (NP) ethoxylates and octylphenol ethoxylates, degrade 

in aquatic environments to the more persistent, toxic, and bioaccumulative degradates NP and 

octylphenol, respectively. We did not receive information on the presence or absence of 

alkylphenol polyethoxylates in thiobencarb-containing formulations. Adjuvants are frequently 

mixed with formulations prior to applications, so although they may not be present in the 

formulations they could still be co-applied. Below we discuss NP’s toxicity as an example of 

potential adjuvant toxicity, as we received no information on adjuvant use or toxicity within the 

BEs provided by EPA.

We queried EPA’s ECOTOX online database and retrieved 707 records of NP’s acute toxicity to 

freshwater and saltwater species. The lowest reported LC50 for salmonids in ECOTOX was 130 

μg/L for Atlantic salmon. Aquatic invertebrates, particularly crustaceans, were killed at low 

concentrations of NP, with the lowest ECOTOX reported LC50 = 1 μg/L for Hyallela azteca. 

These data indicate that an array of aquatic species is killed by NP at low μg/L concentrations. 

We also queried EPA’s ECOTOX database for sublethal toxicity and retrieved 689 records of 

freshwater and saltwater species tested in chronic experiments. The lowest fish LOEC reported 

was 0.15 μg/L for fathead minnow reproduction. Numerous fish studies reported LOECs at or 

below 10 μg/L. 
248

Salmonid prey species appear highly sensitive to sublethal effects of NP at low concentrations. 

The amphipod, Corophium volutator, grew less and had disrupted sexual differentiation at 10 

μg/L (Brown et al. 1999). Multiple studies with fish indicated that NP disrupts fish endocrine 

systems by mimicking the female hormone 17β-estradiol (Arsenault et al. 2004, Brown et al. 

2003, Lerner et al. 2007a, Lerner et al. 2007b, Luo et al. 2005, Madsen et al. 2004, McCormick 

et al. 2005, Hutchinson et al. 2006, Jardine et al. 2005, Segner 2005). NP induced the production 

of vitellogenin in fish at concentrations ranging from 5-100 μg/L (Arukwe and Roe 2008, 

Hemmer et al. 2002, Ishibashi et al. 2006, Schoenfuss et al. 2008b). Vitellogenin is an egg yolk 

protein produced by mature females in response to 17β-estradiol, however immature male fish 

have the capacity to produce vitellogenin if exposed to estrogenic compounds. As such, 

vitellogenin is a robust biomarker of exposure. A retrospective analysis of an Atlantic salmon 

population crash suggested the crash was due to NP applied as an adjuvant in a series of 

pesticide applications in Canada (Brown and Fairchild 2003, Fairchild et al. 1999). Additionally, 

processes involved in sea water adaptation of salmonid smolts are impaired by NP (Lerner et al. 

2007a, Lerner et al. 2007b, Luo et al. 2005, Madsen et al. 2004, McCormick et al. 2005, Jardine 

et al. 2005).

These results demonstrate NP is of concern to aquatic life, particularly salmonid endocrine 

systems involved in reproduction and smoltification. This summary is for one of the more than 

4,000 inerts/other ingredients and adjuvants currently registered for use in pesticide formulations 

and there are likely others with equally deleterious effects. Unfortunately we received minimal 

information on the constituents found in thiobencarb formulations. Consequently, the effects 

that these other ingredients may have on listed salmonids and designated critical habitat remain 

an uncertainty and are a recognized data gap in EPA’s action under this consultation.
249

8.5.13. Uncertainties and Data Gaps Identified from Review of Available Toxicity Information 

for Thiobencarb21

1. No translation of aquatic primary producer toxicity data from single species to salmonid 

aquatic habitats;

2. No AChE activity measurements with salmonids;

3. No empirical data on effects of thiobencarb degradates on fish or habitat assessment 

endpoints;

4. No empirical data on impacts to salmonids’ olfaction, endocrine system, immune-suppression

capacity, migration, spawning, or smoltification;

5. No empirical data on effects to riparian plant species; and,

6. No information on other ingredient toxicity within thiobencarb-containing formulations.

8.5.14. Evaluation of Data Available for Response Analysis

We summarize the available toxicity information by assessment endpoint in Table 53. Data and 

information reviewed for each assessment endpoint was assigned a general qualitative ranking of 

either “low,” “moderate,” or “high.” To achieve a high confidence ranking, the information 

stemmed from direct measurements of an assessment endpoint, conducted with a listed species or 

appropriate surrogate, and was from a well-conducted experiment with stressors of the action or 

relevant chemical surrogates. A moderate ranking was assigned if one of these three general 

criteria was absent, and low ranking was assigned if two criteria were absent. Evidence of 

adverse effects to assessment endpoints for salmonids and their habitat from thiobencarb was 

available for acute lethality to salmonids and aquatic invertebrates, and highly variable for the 

other assessment endpoints. However, much less information was available for other 

ingredients, due in part to the lack of formulation information provided in EPA’s BEs as well as 

the statutory mandate under FIFRA for toxicity data on thiobencarb to support registration. We 

did locate a substantial amount of data on one group of adjuvants/surfactants, the NP ethoxylates. 

However, we received minimal information detailing tank mixes and other ingredients within 

formulations.

 

21 A finding of no information for biological assessment endpoints e.g., AChE inhibition in salmonids, does not 

mean that salmonid AChE is not affected by thiobencarb. 
250

Table 53. Summary of Toxicity Data for Thiobencarb

Assessment Endpoint

Evidence of 

adverse 

responses

(yes/no)

Concentration range of 

observed effect or 

concentrations tested showing 

absence of effect (μg/L)

Degree of 

confidence in 

effects

(low, moderate, 

high)

Salmonid survival (LC50) Yes

760-1200, n=7, lowest 

NOEC=140 (88 d exposure)

High

Growth (LOEC) Yes 23-800 n=6 Moderate

Reproduction (LOEC) Yes 110, n=1 High

Respiration Yes 19, n=1 Low

Swimming - - -

Olfactory-mediated behaviors - - -

Endocrine disruption - - -

AChE inhibition Yes

220 is 1/60th of eel LC50 (30-50% 

AChE inhibition), n=1; estimate for 

salmonids at 12-19

Low

Prey survival Yes 101-1200, n=8 High

Prey reproduction and growth 

(LOEC)

Yes 3-420, n=3 High

Aquatic primary production 

(EC50)

Yes 17-4000, n=6 High

Riparian vegetation (terrestrial 

EC25)

Yes

0.019-3.1 lbs thiobencarb/acre, 

n=18

High
251

Assessment Endpoint

Evidence of 

adverse 

responses

(yes/no)

Concentration range of 

observed effect or 

concentrations tested showing 

absence of effect (μg/L)

Degree of 

confidence in 

effects

(low, moderate, 

high)

Additive toxicity of 

thiocarbamates 

(molinate+thiobencarb)

Yes Multiple, n=2 High

Degradate Toxicity Yes

Toxicity responses modeled using 

EPA ECOSAR model

Low

Adjuvant toxicity: Nonylphenol

Survival

Reproduction

Smoltification

Endocrine disruption

Prey survival

Yes

Yes

Yes

Yes

Yes

130 - >1000

0.15 -10

5 -100

5 - 100

1 - >1000

High

High

Moderate

High

High

n indicates number of studies

- indicates no information found on assessment endpoint
252

9. Risk Characterization

In this section we integrate our exposure and response analyses to evaluate the likelihood of 

adverse effects to individuals and populations (Figure 38). We combined the exposure analysis 

with the response analysis to: (1) determine the likelihood of salmonid and habitat effects 

occurring from the stressors of the action; (2) evaluate the evidence presented in the exposure 

and response analyses to support or refute risk hypotheses; and (3) translate fitness level 

consequences of individual salmonids to population-level effects. The risk characterization 

section concludes with a general summary of species responses from population-level effects. 

We then evaluate the effects to specific ESUs (i.e., species), in the Integration and Synthesis 

section.

Figure 38. Schematic of Risk Characterization Phase.

Analyzed within the context 

of the Environmental 

Baseline (including multiple 

stressors such as temperature 

and environmental mixtures 

of pesticides); the Status of 

Listed Resources and 

Cumulative Effects

Effect on individuals

Effects on populations

Effects on species

(ESU or DPS)

Can EPA ensure that its action 

is not likely to jeopardize the 

continued existence of the 

species?

Effects on habitat

Effects on primary 

constituent elements

Effects on conservation 

value of designated 

habitat

Can EPA ensure that its action 

is not likely to adversely modify 

or destroy designated critical 

habitat?

Risk 

Characterization

Exposure Profile Response Profile

Addressed in the 

Integration and 

Synthesis Section
253

Field incidents of dead fish and crop damage reported in EPA incident database made available 

to NMFS

We also evaluate incident data supplied by EPA for thiobencarb as a separate line of evidence. 

No incidents of fish kills were submitted to us from EPA. 

Seven incidents of crop damage were identified. In most cases, no information was available to 

characterize drift. EPA ranked one incident as highly probable, three as probable, and three as 

possible. We are uncertain how to use these data other than adjacent crops have been affected by 

drifting thiobencarb.

Field studies with thiobencarb from rice irrigation water discharge

A series of studies in Halls Bayou were conducted in 1979 to test the effects of Bolero on aquatic 

organisms following releases of rice irrigation water (Harper et al. 1979). Fish, phytoplankton, 

and macroinvertebrates were evaluated. Much of the study reprint is illegible. What can be 

discerned is that caged fish studies with a variety of species (black drum, striped mullet, grass 

shrimp, and bullhead minnows) showed mortalities following release of thiobencarb 

contaminated water a few days after application due to heavy rains. Mortalities ranged from 0 to 

100% depending on the species and geographic area. No reference sites were identified in the 

study, thus the use of much of the data is questionable. We note that concentrations of 

thiobencarb may have achieved sufficient levels to kill sensitive fish and grass shrimp (2 – 83 

μg/L) at cage sites, however the authors note that other pesticides were also present. No 

laboratory acute toxicity data were available for the species tested. These unpublished data are 

of limited utility based on flawed experimental design.

Mixture analysis based on AChE inhibition

In previous biological opinions with AChE inhibiting insecticides, we analyzed the joint effects 

of multiple AChE inhibiting insecticides using simple, quantitative dose-response relationships 

based on empirical data (NMFS 2008a, NMFS 2009d, NMFS 2010a). We located no 

experimental data on thiobencarb’s capacity to reduce salmonid AChE activity and therefore lack 

the information necessary to conduct a mixture analysis. This does not mean that mixture 

toxicity does not occur, we expect it does given thiobencarb’s common mode of action. When 
254

thiobencarb co-occurs with other AChE inhibitors, we expect additivity, and potentially 

synergism with certain combinations, although we cannot quantify the effects. 

9.1. Exposure and Response Integration

In Figure 39 we show the overlap between exposure estimates for thiobencarb and 

concentrations that affect assessment endpoints. This portion of the analysis primarily focuses 

on thiobencarb, although we do present assessment endpoints for a common adjuvant, 

nonylphenol, mixture toxicity, and degradate information. The figure shows the exposure 

concentration ranges (minimum to maximum values) gleaned from the three sources of exposure 

data we analyzed: EPA’s estimates presented in the BEs that represent crop uses; NMFS’ 

modeling estimates for flood plain habitats; and surface water monitoring data from ambient 

monitoring programs and from targeted monitoring. In addition to the salmonid BEs submitted 

to NMFS, we also considered the exposure estimates developed by EPA in the BEs for the 

California red-legged frog. The effect concentrations are values taken from the toxicity data 

reviewed in the Response Analysis section. For the survival assessment endpoint, effect 

concentrations are LC50s, thus death of individuals occurring at concentrations below LC50s are 

not represented. Consequently, when LC50 effect concentrations are not exceeded by the 

exposure estimates, it does not mean there are no incidences of mortality. Thus for those 

instances where LC50s do not overlap or are not exceeded by exposure estimates, we discuss the 

difference in magnitude of the two metrics and apply best professional judgment on whether 

death of individuals is expected. 

This coarse analysis does not present temporal aspects of exposure nor does it show the 

distribution of toxicity values. It is also predicated primarily on standard toxicity endpoints as 

we located few ecologically relevant sublethal or field-collected data, a noted uncertainty with 

this analysis. However, the analysis does allow us to systematically address which assessment 

endpoints are likely to be affected by exposure to thiobencarb. Where significant uncertainty 

arises, we highlight the information and discuss its influence on our inferences and conclusions. 
255

Most of the concentration ranges overlap with assessment endpoints indicating that adverse 

effects are expected in salmonids and to their habitat if exposed for sufficient durations (Figure 

39). The maximum reported values from CDPR data (16.55 g/L) and NAWQA California data 

(4.38 μg/L) did not exceed effects concentrations for prey survival, fish reproduction, fish 

growth, or salmonid survival. Concentrations from targeted studies, EPA estimates, flood plain 

estimates, and early release of rice irrigation water exceeded effect concentrations for all 

assessment endpoints. Careful analysis is required to determine if exposure to thiobencarb is 

expected to exceed effect concentrations. For example, the targeted monitoring studies (n=10), 

reported a value of 8.9 mg/L which represents the highest concentration reported for thiobencarb 

in water. The value is the result of an early release of rice irrigation water 2 hrs after application 

due to a flood event. Thiobencarb labels generally require a holding period of at least 14 days 

prior to discharge for liquid formulations and up to 30 days for granular formulations, thus we 

did not rely heavily on 8.9 mg/L value to characterize exposure. We note that this value would 

only pertain to early releases of treated water and expect this to happen infrequently, which 

monitoring data indicate are reasonably certain to occur despite product labeling (Table 43). 

Similarly, the exposure numbers presented for floodplain habitats were generated using a fieldscale transport model to assess potential exposure among individuals and habitats. These 

estimates are extremely relevant given the species habitat requirements and the lack of EPA label 

requirements for set-backs (i.e., buffers) to salmonid habitats. However, we recognize that these 

estimates do not reflect the likely exposure of all individuals and all habitats, including exposure 

to individuals occupying deeper habitats and habitats that are not adjacent to thiobencarb 

applications. 

The most sensitive assessment endpoint for direct effects on salmonids was the AChE inhibition 

(EC50) which was estimated at 12-19 μg/L. However, we had low confidence in this study as the 

information stemmed from eels and was conducted with only one thiobencarb treatment at 220 

μg/L. It is uncertain whether salmonids would show the same proportional relationship between 

LC50 and AChE EC50 exhibited by eels. The next most sensitive assessment endpoint was 

aquatic primary production where 17 μg/L thiobencarb inhibited algal growth, while other algae 

were much more tolerant of concentrations up to 4000 μg/L. Survival based on LC50s of 

salmonids exposed to thiobencarb in various formulations was encompassed by the upper end of 
256

concentration ranges observed in targeted studies, EPA estimates, and flood plain habitat 

estimates. We do not disregard any of these effects at this point in the analysis as the evidence 

supports evaluation of whether an individual’s fitness or habitat endpoints are compromised. 

Figure 39. Thiobencarb exposure concentrations and effect concentrations for assessment 

endpoints

Key Findings of Exposure Analysis:

The vast majority of exposure to thiobencarb drift is expected to occur during the peak 

application period of May and June with less frequent exposures expected in April and 

July. Applications during other months are allowed but are extremely rare (state-wide 

average < one application per year). Floodplain habitats alongside rice fields are at the 

greatest risk of receiving elevated concentrations from aerial and ground applications, 

i.e., from 37 – 1910 μg/L. 

Salmonid survival 

LC50

Fish reproduction 

LOEC

Fish growth LOEC

Prey survival LC50

Respiration LOEC

Aquatic primary 

production EC50

EPA estimates

Flood plain habitat 

estimates

Targeted studies in 

receiving waters 

NAWQA & CALDPR

Early release of 

discharge water

1 10 100 1000

Thiobencarb concentration (μg/L)

Thiobencarb Exposure Concentrations

Effect Concentrations for Assessment Endpoints
257

Thiobencarb in runoff typically occurs 14-30 days post application (after required 

holding periods), thus the greatest probability of exposure to the highest runoff 

concentrations likely occurs from mid-May through the end of July. 

Riparian systems and multiple life stages of salmonids are likely exposed to thiobencarb 

from drift during applications and from runoff following release of rice irrigation water 

into drainage networks that ultimately return water to salmonid containing waters. 

Migrating adults, rearing juveniles, and migrating juveniles overlap with peak 

thiobencarb applications in May and June. 

Spawning adults, incubating eggs, and emerging fry exposure to thiobencarb is species 

dependent. 

Key Findings of Response Analysis:

Aquatic, laboratory toxicity tests indicate that thiobencarb is acutely toxic to fish 

(multiple endpoints; 19-1200 μg/L), invertebrates (survival EC50s; 100-1200 μg/L), and 

plants (EC25s; 17 – 4000 μg/L).

AChE inhibition in salmonids is likely, but unquantifiable as no studies with salmonids 

were available.

Riparian plants, particularly emergent vegetation, are highly sensitive to thiobencarb drift 

given agricultural plants sensitivity to thiobencarb. At the labeled rate of 4 lbs a.i./A, 

0.55% of the applied material is predicted to drift ~ 1000 ft from an aerial application. 

That amount of drift would be expected to reduce emergent vegetation seedling growth 

by 25%.

9.2. Evaluation of Risk Hypotheses 

In this phase of our analysis we examine the weight of evidence to determine whether it supports 

or refutes a given risk hypothesis (Table 54). This is not a statistical analysis, but rather a 

qualitative assessment of the available information presented in the response and exposure 

sections. We also highlight general uncertainties and data gaps associated with the data. In 
258

some instances there may be no information specifically related to a given hypothesis. In some 

cases, if information on a similar endpoint or chemical is available, and it is reasonable to do so, 

we extrapolate from the available data to fill gaps, recognizing that this may introduce additional 

uncertainty into the analysis. If the evidence supports the risk hypothesis, we determine whether 

it warrants an assessment at the population level or affects PCEs to a degree that warrants 

analysis of the potential to reduce the conservation value of designated critical habitat.

The information available to characterize thiobencarb exposure included surface water 

monitoring data and estimates from pesticide transport models (runoff and drift). We combine 

this information with the distribution and life-history characteristics of ESA-listed Pacific 

salmonids by ESU. As discussed in the Exposure Analysis section above, each source of 

information has inherent limitations and uncertainties. For example, the pesticide monitoring 

data were generally not designed to quantify peak exposure concentrations or distributions of 

exposure in salmonid habitats within the Central Valley. Consequently, pesticide exposure 

models were used to supplement monitoring data, and together this information was used to 

describe the potential range of pesticide concentrations. We conducted AgDrift model runs to 

provide estimates for concentrations resulting from drift to a shallow and narrow body of water, 

such as those found in floodplain habitats of the Central Valley. Small streams and many 

floodplain habitats such as the Yolo bypass are more susceptible to higher pesticide 

concentrations than larger, high flow systems as their physical characteristics provide less 

dilution. 

We recognize that pesticide concentrations will vary greatly among salmonid habitats in the 

Central Valley, and exposure duration and concentration will be reduced by higher water 

volumes and velocities. 

Standardized toxicity tests for pesticide registration are typically poor predictors of real world 

aquatic ecosystems as test organisms are exposed to constant pesticide concentrations for 

arbitrary durations (e.g., acute, 96-h and chronic, 21- or 28-d) that may poorly reflect field 

exposures, which tend to be repeated exposures Additionally, the tests do not account for 

exposure to stressors, pesticides and others, already in the water. The response of fish and their 
259

prey to different durations of exposure, and exposure representing different environmentally 

relevant dissipation patterns of thiobencarb is a prominent data gap. Exposure durations 

sufficient to elicit toxicological responses can occur at durations much shorter than standard 

toxicity tests. We therefore did not average exposure concentrations over time, i.e., timeweighted averages, because adverse responses to short term thiobencarb exposures such as 

pulses would likely be masked. 

Large spatial and temporal variability exists in the use of aquatic habitats by listed Pacific 

salmonids in the Central Valley. These differences occur at multiple scales of biological 

organization (i.e., individual, population, and species). Both an individual’s life stage and its life 

history are important considerations in its use of aquatic habitats. This natural variation is 

overlaid with the inherent variation of environmental factors including climate (e.g., precipitation 

patterns), habitat stressors, and land use. Given this biological and environmental variability, it 

is difficult to predict the precise exposure to thiobencarb for any one individual, let alone for a 

population or species. 

Consequently, we used the life history information of the three species to evaluate potential 

exposure in their myriad aquatic habitats (Table 28, Table 30, and Table 32). For example, we 

found that adult Chinook migrating upstream to reach spawning grounds and juvenile steelhead 

moving downstream were most at risk of thiobencarb exposure. 

We evaluated the potential for individual fitness consequences (i.e., assessment endpoints) by 

comparing the range in expected exposure concentrations with adverse effect levels in the 

context of aquatic habitat utilization. Given the geographic specificity of thiobencarb use in 

California, we focused on habitats used for rearing and migrating which range from first order 

streams to large rivers—the Sacramento River, San Joaquin River, and the Delta. These habitats 

are critical to successful adult migration, the development and growth of young fish, and for 

providing safe passage to and from the ocean. 

Salmonids also rely on the San Francisco Bay and the Pacific Ocean for migration and rearing 

prior to moving into open ocean areas. In general, thiobencarb exposure will be less intense in 
260

these areas compared to freshwater systems in the Central Valley. Exceptions may include the 

Sacramento-San Joaquin Delta where juveniles are rearing for extended periods (weeks-months) 

proximate to rice fields that may prematurely release thiobencarb-contaminated water due to 

flooding.

This framework allows us to evaluate risk hypotheses based on the spatial and temporal nature of 

exposure to thiobencarb across CA’s Central Valley. Below we make a determination of 

whether fitness of individuals is compromised warranting a subsequent analysis at the population 

level. 

DRIFT:

Drift of thiobencarb into salmonid habitats occurs during its application period, primarily May 

and June, and to a lesser extent in April and July. Concentrations depend on proximity to 

application area, application method, droplet size, release height, wind speed/direction, receiving 

water volume/flow, and interception by riparian vegetation. Based on drift studies, modeling 

exercises, and surface water monitoring studies, concentrations may range from non-detectable 

to 1910 μg/L (aerial application at edge of field). 

RUNOFF:

Runoff from rice fields, including planned discharge following holding periods from 14, 19, and 

30 days, will result in thiobencarb entering salmonid habitats. Concentrations may attain levels 

as high as 350 μg/l at 14 d and 47 μg/L at 30 d (both EPA modeling estimates) in discharge 

waters. Final concentrations within salmonid habitats receiving the discharge will likely be less 

than discharge concentrations and will be dependent on factors including receiving water flow 

and depth. Concentrations may be orders of magnitude higher in runoff if rice irrigation water is 

released within hours or days post application due to emergency releases. Several field studies 

reported concentrations in the mg/L range following early, emergency related releases, one 

attaining levels as high as 8.9 mg/L. We do not expect this to be a common occurrence based on 

reports over the last decade on frequency and measured concentrations from early releases

(Table 43).
261

Table 54. Risk Hypotheses

Effects to salmonids

1. Exposure to thiobencarb via drift or runoff is sufficient to:

a. Kill salmonids from direct exposure

b. Reduce salmonid survival through impacts to growth

c. Reduce reproduction

d. Impair swimming 

e. Impair respiration

f. Reduce salmonid growth through impacts on the availability and quantity of prey

Effects to salmonid habitats

2. Exposure to thiobencarb via drift or runoff is sufficient to:

a. Reduce numbers of aquatic primary producers, thereby affecting salmonid prey communities, 

salmonids, and salmonid instream cover.

b. Reduce riparian vegetation to such an extent that stream temperatures are elevated, erosion 

increases, and reductions in natural coverage results through reduced inputs of large woody debris and 

vegetation.

c. Reduce water quality and prey resources in estuarine environments.

Effects from other stressors of the action and contributing environmental factors

3. Exposure to degradates of thiobencarb will cause adverse effects to salmonids and their 

habitats.

4. Exposure to adjuvants, tank mixtures and other chemicals within pesticide products containing 

thiobencarb will cause adverse effects to salmonids and their habitats.

5. Exposure to other pesticides present in the action area can act in combination with thiobencarb 

to increase effects to salmonids and their habitats.

6. Exposure to elevated temperatures will enhance the toxicity of the stressors of the action.
262

9.2.1. Risk Hypotheses

1. Exposure to thiobencarb via drift or runoff is sufficient to:

A. Kill salmonids from direct exposure

A substantial body of laboratory toxicity data indicates that 50% of a test population of 

salmonids die following short term (<96 hrs) exposures to thiobencarb at levels at or above 760 

μg/L (lowest reported salmonid LC50). Thiobencarb kills salmonids at concentrations well below 

LC50s in 96-hr laboratory studies; however few test results reported the number of dead fish at 

treatment concentrations below an LC50. Exposure duration, health condition of individuals, and 

concentrations are key determinants of lethality in real world aquatic systems, as are the presence 

of other stressors of the action and stressors present in the environmental baseline. We located a 

single embryo-to-fry study that reported a survival NOEC of 140 μg/L for salmonid fry (28 days 

old) following 88 days of continuous flow through exposure from 28 days prehatch to 60 days 

post hatch (Fujimura et al. 1991). At 250 μg/L (LOEC), survival was reduced by 37% and at 

140 μg/L by 13% (no statistical difference from controls) (Fujimura et al. 1991). We don’t 

expect thiobencarb concentrations to persist at 140 μg/L for 88 days in Central Valley surface 

waters based on the monitoring data and pesticide exposure models.

We expect that adults and juveniles will die if exposed to thiobencarb-containing drift from 

applications proximate to salmonid bearing habitats. We expect that 14-d releases will kill 

sensitive individuals if they are located directly downstream from a discharge or are present 

within the drainage network. We also expect that early releases of thiobencarb-containing rice 

water into aquatic habitats are sufficient to kill juvenile and adult salmonids (Table 45). For 

these reasons, we discuss loss of juveniles and adults at the population scale.

B. Reduce salmonid survival through impacts to growth

Thiobencarb reduced Chinook salmon growth by 22% at 49 μg/L compared to unexposed fish 

following 88 days of continuous exposure in a laboratory experiment (Fujimura et al. 1991). It’s 

unlikely that concentrations of 50 μg/L or more will be attained and sustained based on available 

monitoring and pesticide modeling data. Chronic toxicity studies indicate that juvenile fathead 
263

minnow and striped bass growth are affected at concentrations ranging from 23 to 250 μg/L 

(LOECs). Exposure duration in these tests varied from 260 days in the fathead minnow tests to 

between 45 and 52 days for experiments with striped bass. We do not expect any juvenile 

salmon to be exposed to these concentrations for chronic durations that have been shown to 

cause impacts to growth (e.g., 45 days or more). Some juvenile salmon are expected to be 

exposed to these concentrations for shorter durations (e.g. 96 hrs or less), however we located no 

experiments that measured growth over these durations and this assessment endpoint remains 

highly uncertain. Zebrafish length measured following 96 h exposures to thiobencarb at 0.8, 8.0, 

80, and 800 μg/L showed statistically significant reductions at 800 μg/L (NMFS 2011d). 

Additionally, 100% of zebrafish exposed to 800 ug/L were observed as abnormal. Example 

abnormalities included edema, unhatched eggs, lethargic, deformed fins, and deformed tails 

(NMFS 2011d). 

Growth effects have been seen with other AChE inhibitors, such as organophosate and carbamate 

insecticides, however, thiobencarb has low potency compared to these compounds. For these 

reasons, we conclude that growth will not be sufficiently compromised to affect an individual’s

fitness.

C. Reduce reproduction

We located one study that reported effects to fish reproduction in freshwater fish (MRID 

45695101; (EPA 2009b). The LOEC was 110 μg/L (LOEC) following a 260-day exposure in 

flow through conditions. Concentrations from drift and early runoff may attain 110 μg/L, 

however durations of exposure will depend on the receiving habitat and factors such as flow, 

depth, etc. We also note that the majority of known spawning areas occur outside of rice 

growing areas. Although, significant reproduction responses such as reductions in the number of 

eggs spawned and hatching success were observed in this study, there is no information available 

to assess reproductive responses in salmonids exposed to thiobencarb for relevant exposure 

durations. Additionally, no adverse responses to reproduction were noted following exposure to 

53 μg/L for 260 days (LOEC). For these reasons, we do not expect reproduction to be affected 
264

and therefore do not anticipate an individual’s fitness to be compromised based on effects to 

reproductive endpoints.

D. Impair swimming 

Swimming is critical to salmonids at all life stages following hatching. We found no studies that 

tested swimming capacity or performance following exposure to thiobencarb. We found no 

direct information on the AChE response in salmonids following environmentally relevant 

exposures, which represents a significant data gap. Since thiobencarb is a likely AChE inhibitor, 

we do expect swimming to be compromised well below salmonid LC50. We note that if eels are 

an appropriate surrogate for AChE inhibition, then salmonids exposed to 1/60th of the LC50

would experience 40-50% inhibition of AChE. However, we located no evidence to refute or 

support this assertion. If eels are appropriate surrogates, then salmonids would experience 

neurotoxic effects such as reduced swimming ability at concentrations of 12-19 μg/L thiobencarb 

(i.e., 1/60th of the LC50) after a few hours of exposure. We conclude that salmonids likely 

experience AChE inhibition at concentrations below salmonid LC50s (i.e., 720 μg/L), however 

we have no empirical information to further refine effect thresholds. 

Given the targeted monitoring data, EPA’s model results, and NMFS flood plain habitat 

modeling results, swimming is likely affected at concentrations well below LC50s, less than 720 

μg/L, particularly for sensitive juveniles. We therefore conclude that adult and juvenile fitness 

are compromised from impaired swimming and discuss population level consequences.

E. Impair respiration

Functional breathing is critical to salmonid survival. We located one study that tested this 

hypothesis. The study showed that at 19 μg/L BoleroEC gills were abnormally inflamed 

compared to control fish. Study limitations included lack of verification of treatment 

concentrations, small treatment size (n = 3 fish), and no replication within treatments. Although 

the study’s limitations reduce its utility, the data do support adverse responses in fish gills and 

suggest that thiobencarb is a gill toxicant. If salmonids are as sensitive to BoleroEC as mosquito 

fish, respiration is likely impaired due to gill damage. Seven day static exposures maintaining an 

average concentration of ~20 μg/L is a potential scenario in lower flow, shallow, flood plain 
265

habitats as well as in salmonid habitats receiving continuous discharges of thiobencarb from rice 

growing operations. We therefore conclude that damaged gills reduce an individual’s fitness and 

discuss potential population level consequences.

F. Reduce salmonid growth through impacts on the availability and quantity of prey

Thiobencarb inhibits acetylcholinesterase activity in exposed invertebrates. Therefore, we 

evaluate toxic responses experienced by salmonid prey. The majority of data derive from 

laboratory bioassays with standard aquatic invertebrates (Table 51). Toxicity values for survival, 

growth, and reproduction were available and indicate that thiobencarb adversely affects aquatic 

invertebrates. Concentrations reducing survival ranged from 101-1200 μg/L following 48 - 96 hr 

exposures (EC50s). Reproduction and growth endpoints were affected at lower concentrations, 3-

420 μg/L following 21 d exposures. 

Salmonid habitats proximate to rice fields and particularly shallow, low-flow flood plain habitats 

are most at risk of receiving drift. In these areas, thiobencarb levels in drift may attain levels 

sufficient to kill aquatic invertebrates. Additionally, early releases of rice irrigation water 

containing thiobencarb may also kill salmonid prey. Therefore, we discuss the impact on 

populations of juveniles that rely on invertebrate communities for food.

2. Exposure to thiobencarb is sufficient to:

A. Reduce numbers of aquatic primary producers thereby affecting salmonid prey communities, 

salmonids, and salmonid instream cover;

Aquatic primary producers showed reduced growth (EC50s) at concentrations as low as 17 μg/L 

and as high as 4000 μg/L (Table 51). Both values were reported for species of green algae. No 

studies were available that tested aquatic community responses to thiobencarb in running water 

systems or evaluated salmonid habitats within CA’s Central Valley. The bioassay results 

indicate that following 48-hr continuous exposure to thiobencarb, reductions in aquatic primary 

producers can occur. Translating this information to real world systems is complicated by many 

factors including condition and composition of existing instream plant community, community 

responses to reductions in primary production, and prediction of bottom up effects in systems 
266

that are severely and continuously compromised by multiple stressors. We conclude that effects 

to primary producers may occur, particularly in salmonid habitats receiving drift and those 

exposed to early release of rice irrigation water. Therefore, we do not anticipate individual 

fitness of salmonids to be compromised by anticipated infrequent reductions in instream primary 

productivity.

B. Reduce riparian vegetation to such an extent that stream temperatures are elevated, erosions 

increases, and reductions in natural coverage results through reduced inputs of woody debris 

and vegetation.

Amount and frequency of thiobencarb drift are the predominant factors in evaluation of riparian 

vegetation responses. Thiobencarb adversely affects emerging plants, both monocots and dicots, 

at concentrations as low as 0.019 lb a.i./A, which represents < 1% of the labeled application rate 

and is comparable to the drift expected 1000 feet from an aerial application of thiobencarb. If 

riparian vegetation is as sensitive as the crops tested, they are extremely susceptible to 

thiobencarb’s herbicidal properties. Thiobencarb is applied both aerially and by ground 

throughout the application period and intensely in May and June. Approximately 25% of known 

thiobencarb application occur with liquid formulations. Riparian zones that are proximate to 

applications are at the greatest risk of exposure, but even zones that are 1000 ft away from 

application areas may receive toxic levels. Reductions in riparian systems likely compromise 

fish fitness by increasing temperatures, increasing erosion/sedimentation, and reducing natural 

cover.

We therefore conclude that loss of riparian vegetation affects salmonid habitats and salmonids 

and discuss potential effects to salmonid populations and to designated critical habitats.

C. Reduce water quality and prey resources in estuarine environments.

We located monitoring data within the San-Francisco Bay Estuary and Sacramento-San Joaquin 

Delta as well as toxicity information from a variety of estuarine and marine organisms. Based on 

our review of these data we expect that thiobencarb will reach CA’s estuaries; however measured 
267

levels attained typically are in the low ng/L range (maximum concentration of 66 ng/L) (Kuivila 

and Jennings 2007), orders of magnitude below adverse effect levels. Given these data we do not 

anticipate water quality in estuaries to be reduced by thiobencarb so we do not evaluate impacts 

to populations. 

3. Exposure to degradates of thiobencarb will cause adverse effects to salmonids and their 

habitats.

Many degradates are formed as thiobencarb breaks down in aquatic systems. We located no 

aquatic toxicity information for these degradates, even for those that represent more than 10% of 

thiobencarb applied. EPA applied structure activity models that identified several degradates as 

highly toxic to fish, invertebrates, and plants. Additionally, EPA predicted that some degradates, 

such as bencarb may impair salmonids and their habitats. We therefore conclude that thiobencarb 

degradates will cause adverse effects to salmonids and their habitats beyond those predicted with 

thiobencarb alone. 

4. Exposure to adjuvants, tank mixtures and other chemicals within pesticide products 

containing thiobencarb will cause adverse effects to salmonids and their habitats.

In addition to exposure to thiobencarb, which is currently the only stressor of the action

incorporated in the EPA’s risk assessment, salmonids and their habitat are likely exposed to

other stressors of the action, including additional active ingredients in formulated products and 

tank mixes. Salmonid habitats may also be exposed to a number of the approximately 4,000 

inert ingredients approved for use in end-use pesticide products by EPA, including adjuvants, 

surfactants, and other products that are applied as tank mixtures. Once the mixture (formulated 

pesticide or tank mix) is introduced into the environment, physico-chemical properties influence 

transport and resulting concentrations in sediment, water, and organisms. We expect some of 

these chemicals to reach salmonid habitats from spray drift deposition and from runoff events 

following application. Salmon and their habitats exposed to these multiple stressors are expected 

to show a greater adverse response than laboratory animals exposed only to thiobencarb, thus 
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available toxicity data generally underestimate the response from a field-applied pesticide 

mixture.

4.a. Additional Active Ingredients within Tank Mixes

Some of the active ingredients used in tank mixtures may be more toxic than the active 

ingredient, for example, copper sulfate. EPA has not consulted with NMFS on most of these 

active ingredients, and NMFS in this consultation is not evaluating the effects of each of the 

separate active ingredients with which thiobencarb could be mixed. This is a significant source 

of uncertainty in the analysis, and NMFS described some of its concerns in this Opinion.

Several labels recommend propanil as a tank mixture. Also, copper sulfate is frequently mixed 

with thiobencarb based on CDPR data. Propanil is an herbicide with a wide range of acute LC50s 

for freshwater fish (2.1– 12.7 mg/L) and invertebrates (1.2 – 16 mg/L) (EPA 2009a). Rainbow 

trout appear the most sensitive freshwater fish tested with an LC50 of 2.1 mg/L. Propanil is 

highly toxic to aquatic (EC50 16-110 μg/L) and terrestrial primary producers (EC25 0.09-12 lbs 

propanil/ acre). Propanil’s degradate, 3,4-DCA, is 15 fold more toxic than propanil based on D. 

magna EC50 survival test (48-h EC50 = 528 μg/L parent; 34.9 μg/L degradate). Drift of propanil 

into riparian areas and salmonid habitats will likely impair primary producers. 

Copper sulfate (and other copper organic complexes) readily disassociates in water once 

transported to aquatic habitats from either drift or runoff. Dissolved copper is acutely toxic at 

low μg/L concentrations to a wide range of salmonid assessment endpoints, salmonid prey, and 

primary producers (Fernandes and Henriques 1991, Baldwin et al. 2003b, Baldwin et al. 2011, 

Hansen et al. 1999, Mebane and Arthaud 2010). NMFS believes that for many of these 

endpoints copper could be significantly more toxic than thiobencarb. Given the levels of copper 

used for pest control in CA’s Central Valley, copper is potentially a major toxicant affecting 

listed salmonids. However, as noted above, EPA is not consulting on each active ingredient with 

which thiobencarb can be mixed, and NMFS is not in a position to evaluate this hypothesis with 

respect to the populations at issue in this opinion. Further, measures to keep thiobencarb out of 

the water should also reduce concentrations of other active ingredients.
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4. b. Inert/other ingredients

Labels for thiobencarb list the percentage of other ingredients, and can be as much as 90% of the 

total volume (Table 47). In past Opinions, other ingredients were referenced on labels and 

included highly toxic compounds such as aromatic solvents, xylene solvents, and petroleum 

distillates. PAHs (in some aromatic solvents) are known aquatic contaminants, and some have 

been linked to carcinogenic and immunogenic effects. Other aromatic solvents may have a 

narcotic effect. For example, EPA noted that mixtures of “xylenes and xylene isomers are 

moderately to highly toxic to aquatic species” (EPA 2005). Toxic effects vary dependent on the 

specific chemical. The likelihood of these compounds co-occurring in the water column is 

difficult to determine with any specificity, but can reasonably be presumed to occur in spray drift

deposition and runoff.

In addition to other/inert ingredients listed on the labels for the a.i.s considered in this

Opinion, thousands of other compounds are approved by EPA as additives to pesticide products 

without any specific requirement for the compound identity or amount to be listed on the labels. 

One example of these ingredients are nonylphenol polyethoxylates, which have been linked to 

endocrine disruption. These chemicals were addressed at length in previous Opinions on EPA 

pesticide registrations (NMFS 2008c, NMFS2009b). There are however, myriad others, some of 

which may increase the toxicity of thiobencarb. The majority of a pesticide formulation is often

composed of inert ingredients. For example, one thiobencarb label considered in this Opinion 

has more than >90% other ingredients. Consequently, salmonid exposure to these ingredients 

may be greater than exposure to thiobencarb. EPA currently has no specific method of 

accounting for this potential additional toxicity and risk, but it cannot be ignored. NMFS has 

opted to address the uncertainty associated with these ingredients qualitatively. 

Collectively, the available lines of evidence support the overall hypothesis that other stressors of 

the action cause adverse effects to salmonids and their habitat. From our review of the available 

information it is not possible to accurately quantify the contribution of other stressors of the 

action. These stressors include propanil and copper, as well as inert/other ingredients in 

pesticide formulations. These stressors of the action are an important consideration when
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assessing potential effects on listed salmonids and their habitat. Thus, we carry forward effects 

from these other stressors of the action when we discuss effects to salmonid populations.

5. Exposure to other pesticides present in the action area can act in combination with 

thiobencarb to increase effects to salmonids and their habitats.

The available toxicity and exposure data support this hypothesis. Other AChE inhibitors (e.g., 

carbamates and organophosphates) found in the action area likely result in additive or synergistic 

effects to exposed salmonids and aquatic invertebrates. The magnitude of effects will depend on 

the duration and concentrations of exposure. Effects of exposure to other pesticides are carried 

forward to the population-level analysis.

6. Exposure to elevated temperatures can enhance the toxicity of the stressors of the action.

We reviewed the available information to determine whether empirical data indicated enhanced 

toxicity at elevated temperatures for carbamates in general, and, in particular, for thiobencarb. 

Multiple experimental results indicated that increases in temperature resulted in lower LC50s for 

fish, including salmonids (Mayer and Ellersieck, 1988). Acute lethality bioassays with 

carbamates (aminocarb and carbaryl), showed a distinct relationship between toxicity (measured 

by 96 h LC50s) and temperature (Mayer and Ellersieck, 1988). These experiments were 

conducted with brook trout and yellow perch. We recognize that aminocarb and carbaryl are 

both N-methylcarbamates, a separate carbamate class than thiocarbamates (thiobencarb’s class), 

thus uncertainty exists whether thiocarbamates will respond to temperature in a similar fashion. 

These data support the hypothesis and we therefore carry forward temperature effects when we 

discuss effects to salmonid populations. 
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Table 55 summarizes the preceding discussion covering our risk hypotheses as to whether or not 

the stressors of the action may compromise individual fitness to the listed salmonids within 

California’s Central Valley.
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Table 55. Summary of determinations to risk hypotheses.

Effects to salmonids

Individual fitness 

compromised

1. Exposure to thiobencarb via drift or runoff is sufficient to:

a. Kill salmonids from direct exposure yes

b. Reduce salmonid survival through impacts to growth no

c. Reduce reproduction no

d. Impair swimming yes

e. Impair respiration yes

f. Reduce salmonid growth through impacts on the availability and quantity 

of prey

yes

Effects to salmonid habitats

2. Exposure to thiobencarb via drift or runoff is sufficient to:

a. Reduce aquatic primary producers thereby affecting salmonid prey 

communities, salmonids, and salmonid instream cover

no

b. Reduce riparian vegetation to such an extent that stream temperatures 

are elevated, erosion increases, and reductions in natural coverage results 

through reduced inputs of large wood and vegetation.

yes

c. Reduce water quality and prey resources in estuarine environments. no

Effects from other stressors of the action and contributing environmental factors

3. Exposure to degradates of thiobencarb will cause adverse effects to 

salmonids and their habitats. yes

4. Exposure to adjuvants, tank mixtures and other chemicals within 

pesticide products containing thiobencarb will cause adverse effects to 

salmonids and their habitats.

yes

5. Exposure to other pesticides present in the action area can act in 

combination with thiobencarb increase effects to salmonids and their 

habitats.

yes

6. Exposure to elevated temperatures will enhance the toxicity of the 

stressors of the action yes
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9.2.2. Population Level Responses

In this section, we discuss the likelihood of individual salmonid fitness consequences rising to 

the level at which population-level effects occur22. To forecast potential effects at the population 

level, first we determine the number and life stage of salmonids likely exposed to thiobencarb 

and the other stressors of the action. Since the overwhelming majority of thiobencarb 

applications are expected to occur between April and July (Figure 24), we overlay relative 

abundances of listed salmonid populations during these months in rice-growing areas of the 

Central Valley (Table 56). Next we evaluate the potential for population level consequences 

based on the relative abundance information. The more individuals within a population that 

overlap with high thiobencarb use, the greater the likelihood for population level consequences. 

Rearing salmonids showed low abundances in aquatic areas prone to thiobencarb drift and 

runoff. Migration timing and areas overlapped with thiobencarb use. Juvenile migration 

abundances were ranked as low for Chinook populations and medium for Central Valley 

steelhead populations. Migrating juvenile steelhead, therefore have a greater likelihood of 

exposure to thiobencarb than migrating Chinook. We do not expect spawning to co-occur within 

rice growing areas for Chinook populations, however upstream migrations of pre-spawn adults 

through rice growing areas showed high abundances for spring-run Chinook, moderate 

abundances for winter-run Chinook, and low abundance for steelhead between April and July. 

 

22 In previous Opinions we used population modeling to evaluate consequences to the population growth rate 

(lambda) from reduced survival of subyearling salmonids and reduced growth due to AChE effects. We do not 

employ these models in this Opinion because overlap of thiobencarb use areas with rearing subyearlings is not 

expected to a great extent. Therefore we expect much lower number of juveniles to be exposed. This was not the 

case for other pesticides in previous Opinions.
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Table 56. Relative abundance of listed salmonids in rice areas during thiobencarb use periods, 

April-July. 

Listed Species

Populations

Life-Stage 

Behaviors

Expected Abundance

 None Low Med High

Central Valley 

Steelhead: 

Adult migration X

1 population Spawning X

Early rearing X

Juvenile migration X

Sacramento River 

Winter-run Chinook

Adult migration X

1 population Spawning X

Early rearing X

Juvenile migration X

Central Valley Springrun ChinookAdult migration X

3 populations Spawning X

Early rearing X

Juvenile migration X

Based on the expected abundances of juveniles of each species co-occurring with thiobencarb 

application periods, we determined that population level consequences are not anticipated in the 

Central Valley from rice applications of thiobencarb (Table 57).

The two primary drivers for high concentrations of thiobencarb, drift into shallow flood plain 

habitats and early release of rice irrigation water are to likely occur infrequently, and exposure to 

such concentrations is expected to be infrequent given information on their occurrence and/or on 

spatial and temporal considerations. Most life stages of the three listed salmonids are not 

expected to occur in close proximity to thiobencarb use sites during the period of application and 

discharge, or are only expected to occur in relatively low abundance. The exceptions are 

migratory adult Chinook.
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A high proportion of spring-run Chinook adults migrate up the Sacramento River to spawn 

during thiobencarb application periods, and a moderate number of winter-run Chinook pass 

through these areas during peak application months. Adult migration through the rice growing 

areas occurs over a 7-8 month period compared to the relatively short application period of 

thiobencarb, which largely occurs during a span of 6-8 weeks. Migrating adults typically follow 

the thalweg where water is faster and deeper than other areas of the river possibly reducing 

exposure to edge of field concentrations and decreasing exposure duration, i.e., adults may cooccur with thiobencarb albeit where dilution is greatest (Table 21). Overall, we expect adults to 

be exposed to an array of thiobencarb concentrations, many of which would be in the ng/L range, 

and potentially some in the ug/L range. Duration of exposure will depend on individual adult 

behaviors, e.g., feeding, holding, and migrating, which modulate the toxicity experienced by an 

individual. Given the toxicity information, we expect some adults to show fitness level effects, 

however population level consequences are not anticipated.
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Table 57. Salmonid risk hypotheses and population level effects. 

Effects to salmonids

Individual fitness 

compromised

(yes/no)

Population 

level effects 

anticipated

(yes/no)

1. Exposure to thiobencarb via drift or runoff is sufficient to:

a. Kill salmonids outright yes no

b. Reduce salmonid survival through impacts to growth no --

c. Reduce reproduction no --

d. Impair swimming yes no

e. Impair respiration yes no

f. Reduce salmonid growth through impacts on the availability 

and quantity of prey

yes no

Effects to salmonid habitats

2. Exposure to thiobencarb via drift or runoff is sufficient to:

a. Reduce aquatic primary producers thereby affecting salmonid 

prey communities, salmonids, and salmonid instream cover

no --

b. Reduce riparian vegetation to such an extent that stream 

temperatures are elevated, erosion increases, and reductions in 

natural coverage results through reduced inputs of large wood 

and vegetation.

yes no

c. Reduce water quality and prey resources in estuarine 

environments.

no --

Effects from other stressors of the action and contributing environmental 

factors

3. Exposure to degradates of thiobencarb will cause adverse 

effects to salmonids and their habitats. yes no

4. Exposure to adjuvants, tank mixtures and other chemicals 

within pesticide products containing thiobencarb will cause 

adverse effects to salmonids and their habitats.

yes no

5. Exposure to other pesticides present in the action area can act 

in combination with thiobencarb increase effects to salmonids 

and their habitats.

yes no

6. Exposure to elevated temperatures will enhance the toxicity of 

the stressors of the action yes no
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10. Cumulative Effects

Cumulative effects include the effects of future state, tribal, local, or private actions that are 

reasonably certain to occur in the action area considered by this Opinion (50 CFR 402.02). 

Future federal actions that are unrelated to the proposed action are not considered in this section 

because they require separate consultation pursuant to section 7 of the ESA.

During this consultation, NMFS searched for information on future state, tribal, local, or private 

actions that were reasonably certain to occur in the action area. NMFS conducted electronic 

searches of business journals, trade journals, and newspapers using Google and other electronic 

search engines. Those searches produced reports on projected population growth, commercial 

and industrial growth, and global warming. Trends described below highlight the effects of 

population growth on existing populations and habitats for the Central Valley ESUs/DPSs. 

Changes in the near-term (five-years; 2017) are more likely to occur than longer-term projects 

(10-years; 2022). Projections are based upon recognized organizations producing best available 

information and reasonable rough-trend estimates of change stemming from these data. NMFS 

analysis provides a snapshot of the effects from these future trends on listed ESUs.

California’s Central Valley, like other regions on the west coast is projected to have some of the 

most rapid growth of any area in the U.S. within the next few decades. California, Idaho, 

Oregon, and Washington are forecasted to have double digit increases in population for each 

decade from 2000 to 2030 (USCB 2005). According to U.S. Census Bureau figures, California 

is in step with this prediction with a growth of 10 % between 2000 and 2010. The major rice 

growing counties are listed in Table 58. These counties surpassed the state average for 

population growth. The overall growth rate was 17.4 percent. Placer County grew the most 

rapidly among the major rice growing counties with a 40.3 percent growth rate (Census Bureau 

2010).
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Table 58. Population growth rates among major rice growing counties.

County 2000 Census 2010 Census Percent

Growth

Tehama 56,039 61,136 9.1

Glenn 26,453 28,229 6.7

Butte 203,171 220,577 8.6

Colusa 18,804 21,321 13.4

Sutter 78,930 92,614 17.3

Yuba 60,219 72,925 21.1

Yolo 168,660 199,407 18.2

Placer 248,399 348,552 40.3

Sacramento 1,223,499 1,400,949 14.5

All Combined 2,084,174 2,445,710 17.4

The Delta, East Bay, and Sacramento regions are expected to increase in population by nearly 

three million people by the year 2020 (California Commercial 2002). Increases in urbanization 

and housing developments can impact habitat by altering watershed characteristics, and changing 

both water use and stormwater runoff patterns. The anticipated growth will occur along both the 

I-5 and US-99 transit corridors in the east and Highway 205/120 in the south and west. Increased 

growth will place additional burdens on resource allocations, including natural gas, electricity, 

and water, as well as on infrastructure such as wastewater sanitation plants, roads and highways, 

and public utilities. Some of these actions, particularly those which are situated away from 

waterbodies, will not require Federal permits, and thus will not undergo review through the 

section 7 consultation process with NMFS.

Urban runoff from impervious surfaces and roadways may contain oil, heavy metals, PAHs, and 

other chemical pollutants and flow into state surface waters. Inputs of these point and non-point 

pollution sources into numerous rivers and their tributaries will affect water quality in available 

spawning and rearing habitat for salmon. Based on the increase in human population growth, we 

expect an associated increase in the number of NPDES permits issued and the potential listing of 

more 303(d) waters with high pollutant concentrations in state surface waters. Continued growth 

into forested and other natural areas will continue the cycle of altering landscapes to the 
279

detriment of salmon habitat. Altered landscapes adversely affect the delivery of sediment and 

gravel and significantly alter stream hydrology and water quality. 

Increased urbanization also is expected to result in increased recreational activities in the region. 

Among the activities expected to increase in volume and frequency is recreational boating. 

Boating activities typically result in increased wave action and propeller wash in waterways. 

This potentially will degrade riparian and wetland habitat by eroding channel banks and midchannel islands, thereby causing an increase in siltation and turbidity. Wakes and propeller wash 

also churn up benthic sediments thereby potentially resuspending contaminated sediments and 

degrading areas of submerged vegetation. This, in turn, would reduce habitat quality for the 

invertebrate forage base required for the survival of juvenile salmonids moving through the 

system. Increased recreational boat operation in the Delta is anticipated to result in more 

contamination from the operation of gasoline and diesel powered engines on watercraft entering 

the water bodies of the Delta. California is widely known for its scenic and natural beauty. 

Natural areas are used recreationally by residents and tourists. Increases in use could place 

additional strain on the natural state of park and nature areas that are also occupied by protected 

species. 

In the western states, which includes California, a 2.2 percent rise in agricultural output is 

anticipated (Figueroa and Woods 2007). Increases in agricultural practices may negatively affect 

riparian and wetland habitats through upland modifications that lead to increased siltation or 

reductions in water flow in stream channels flowing into the action area, including the 

Sacramento River and Delta. Grazing activities from dairy and cattle operations can degrade or 

reduce suitable critical habitat for listed salmonids by increasing erosion and sedimentation, as 

well as introducing nitrogen, ammonia, and other nutrients into the watershed, which then flow 

into receiving waters. Stormwater and irrigation discharges related to agricultural activities can 

contain a heavy nutrient load, numerous pesticides and herbicides that may negatively affect 

water quality, prey, salmonid reproductive success and survival rates (Dubrovsky et al. 1998). 

Increased output and water diversions for agriculture may also place greater demands upon 

limited water resources. Water diversions will reduce flow rates and alter habitat throughout 
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freshwater systems. As water is drawn off, contaminants will become more concentrated in 

these systems, exacerbating contamination issues in habitats for protected species. 

Mining has historically been a major component of California’s economy. With national output 

for metals projected to increase by 4.3% annually, output of western mines should increase 

markedly (Figueroa and Woods 2007). Increases in mining activity will add to existing 

significant levels of mining contaminants entering river basins. Given this trend, we expect 

existing water degradation in California streams that feed into or provide spawning habitat for 

threatened and endangered salmonid populations will be exacerbated. 

The above non-federal actions are likely to pose continuous unquantifiable negative effects on 

listed salmonids addressed in this Opinion. Each activity has negative effects on water quality. 

They include increases in sedimentation, increased point and non-point pollution discharges, 

decreased infiltration of rainwater (leading to decreases in shallow groundwater recharge, 

decreases in hyporrheic flow, and decreases in summer low flows).

Non-federal actions likely to occur in or near surface waters in the action area may also have 

beneficial effects on the species. They include implementation of riparian improvement 

measures, fish habitat restoration projects, and best management practices (e.g., associated with 

timber harvest, grazing, agricultural activities, urban development, road building, recreational 

activities, and other non-point source pollution controls).
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11. Integration and Synthesis for Listed Species

In this section we describe the potential for thiobencarb and its associated stressors to reduce the 

reproduction, numbers, or distribution of listed anadromous salmonids within California’s 

Central Valley, taking into account the status of the three species, the environmental baseline, 

and cumulative effects. 

Our past Biological Opinions on pesticides and Pacific salmonids assessed pesticides authorized 

for use across the Pacific NW and California on multiple land types, including multiple crops, 

which could be applied throughout the year. Due to their broad spatial and temporal use we did 

not conduct single crop/use analyses. 

Thiobencarb’s authorized use is narrow. In the Western United States it is used solely on rice 

grown exclusively in the Central Valley of California where soil type, temperature, other 

climatic factors, and water availability make rice growing possible. Extensive use information 

for thiobencarb has been collected by CDPR which provides primary information on when, 

where, and how much has been applied to rice. For these reasons, we focused the analysis both 

temporally and spatially to California’s Central Valley which serves as habitat for three species 

of listed anadromous salmonids, Central Valley Spring-run Chinook, Sacramento River Winterrun Chinook, and California Central Valley Steelhead. 

Threats to all three species include loss of historical spawning habitat, degradation of remaining 

habitat, and threats to genetic viability. Each species requires cool, clean freshwater for 

successful development. In the Central Valley, summer water temperatures are suitable above 

150-500 m elevations; however access to these habitats is typically blocked by dams. In the case 

of steelhead, dams also have reduced expression of the anadromous life history form, greatly 

reduced the abundance of anadromous individuals, and prevented exchange of migrants among 

resident populations (i.e., populations located above dams).
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In addition to loss of habitat, each species must contend with widespread habitat degradation and 

modification of both rearing and migration habitats caused by water development and land-use 

practices. Natal streams do not have large impassable dams, like many CV streams, but they do 

have small hydropower dams and water diversions that have greatly reduced or eliminated instream flows during critical migration periods. Threats in freshwater migratory corridors include 

unscreened or inadequately screened water diversions, predation by native and non-native 

species, excessively high temperatures, and pollution from agricultural, urban, and mining landuse. 

Past hatchery practices have influenced the genetic viability of steelhead and spring-run 

Chinook. The Feather River spring-run populations depend on the Feather River Hatchery where 

production is likely a hybridization between spring-run and fall-run Chinook salmon. A 

declining proportion of wild CV steelhead spawners has resulted in reduced genetic viability, 

likely caused by increased interactions with hatchery fish. The winter-run Chinook population is 

the last surviving population of the species and is dependent on regulated flows of the 

Sacramento River. The recovery target for this species is 110,000 annual adult returns. The 

small number of winter-run Chinook (approximately 8000 annual natural adult returns based on 

the1992-2007 average) makes the species vulnerable to threats and reduces its capacity to 

recover following population perturbations. Sustained population declines over past decades 

raise concerns about the population’s genetic integrity and its capacity to recover from 

catastrophic events.

River and stream restoration programs, reductions of toxic materials into Central Valley waters, 

and improved water management are some examples of beneficial activities that have occurred 

in the Central Valley. Additionally, California’s regulatory agencies identified thiobencarb as an 

aquatic risk to both humans and aquatic life more than 20 years ago and have implemented a 

myriad of actions to reduce its potential to contaminate surface waters of the Central Valley. 

Measures include mandatory holding periods, application restrictions, a surface water monitoring 

program, an education program for applicators, and an enforcement program. In aggregate, these 

activities have been successful at reducing and tracking thiobencarb entering surface waters. 
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These activities have reduced exposure to toxic concentrations of thiobencarb to the three 

species.

We anticipate a variety of threats over the next 15 years (described in the Cumulative Effects 

Section). The human population of CA’s Central Valley grew by more than 17% between 2000 

and 2010 and is expected to increase annually. The more people in the Central Valley, the larger 

the urban and residential footprint. Listed salmonids will likely encounter altered stream and 

river flows, elevated water temperatures, and more pollutants. High intensity agricultural 

practices are expected to continue and, left unchecked, can, and do, impact riparian and wetland 

habitats, cause erosion and sedimentation, and contaminate surface waters with excess nutrients 

and pesticides. Increases in population and a continuation of high intensity agriculture will 

likely affect water quantity and quality, salmonid prey, and ultimately may affect the species’ 

recovery trajectories.

We present key findings of the effects analysis which we apply to the final assessment at the 

species level (Table 59). For each of the three listed species, we rank the potential of thiobencarb 

to reduce the reproduction, numbers, or distribution as Low, Medium, or High. For threatened 

species, a “high” designation equates to a jeopardy determination. For endangered species, 

which are more vulnerable to extinction, a medium or high designation equates to a jeopardy 

determination. In the Conclusion Section we present jeopardy or no jeopardy determinations.
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Table 59. Key Findings of Effects of the Proposed Action Section

Section of Biological 

Opinion

Summary Finding 

Exposure Individual salmonids from each of the three species are expected to be 

exposed to thiobencarb and its associated stressors.

Response Anticipated levels of thiobencarb and its associated stressors are sufficient to 

impair several endpoints linked to individual fitness of salmonids.

Risk Characterization Individual fitness consequences are likely, but not expected, to affect 

populations of the three species.

Environmental Baseline Pesticides, elevated temperature, water flow, other pollutants, and dams are 

some of the stressors affecting salmonid habitats within the Central Valley. 

Water quality is severely degraded throughout the Central Valley. California’s 

current risk reduction measures for thiobencarb reduced exposure to listed 

salmonids and their habitats.

Cumulative Effects Increases in human population and a continuation of high intensity agriculture 

are anticipated within the Central Valley. Some beneficial activities are also 

anticipated including river restoration projects and best management practices 

to improve water quality.

Juveniles and adults that encounter thiobencarb, its degradates, other ingredients within 

formulations, tank mixture chemicals within the Central Valley may or may not experience 

adverse effects. Whether a salmonid is affected is a function of how much it is exposed to and 

for how long. The longer the exposure period the greater the likelihood of experiencing reduced 

fitness. Similarly, the higher the concentrations of the stressors of the action the greater the 

likelihood of experiencing reduced fitness. 

We found that if salmonids rear during thiobencarb’s high use periods proximate to rice fields, 

they are at risk from drift. In rice growing areas, concentrations in flood plain habitats and small 

streams may attain levels that impair swimming, impair breathing, kill salmonid prey, and in 

some cases kill salmonids. We expect these occurrences to be infrequent, although not absent. 

We also expect that these effects to be magnified when other pesticides are present or when 

water temperatures are elevated. Waters that show elevated temperatures and/or contain other 

pesticides that share a mode of action with thiobencarb i.e., other AChE inhibitors, effects of the 

action will be more pronounced and dependent on potency of the chemicals as well as the 

magnitude of elevated temperature.
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The other high risk scenario involves early releases of thiobencarb-containing irrigation water. 

Rice water may be released into drainages prematurely during high rain events or from leaky 

dikes. High concentrations that would kill salmonids and their prey may occur if rice water is

released prior to the end of holding periods, particularly rice water released within a few days of 

application (Table 46). The magnitude of effects depends, in part, on how soon the water is 

released after application of thiobencarb, how far salmonid-bearing waters are located 

downstream from the release, how many salmonids are exposed, and how long they are exposed. 

We anticipate these occurrences as rare given the information reviewed and the current federal 

and state regulations already in place.

Central Valley Spring-run Chinook (Threatened species)

The ESU includes three populations in the upper Sacramento River and its tributaries as well as a 

population in the Feather River. The spatial distribution has been greatly reduced through 

extirpation of populations and dams blocking fish passage. Genetic diversity was similarly 

reduced with the extirpation of all San Joaquin runs. Abundance levels are severely depressed 

from historic estimates. Time series data show that the three tributary populations currently have 

growth rates just above replacement while the Feather River population is dependent on hatchery 

augmentation.

We rank the potential for thiobencarb exposure to cause species level consequences to CV 

Spring-run Chinook as low for the following reasons:

Exposure of migrating and spawning adults to thiobencarb is limited. Adults returning to 

their spawning habitats do migrate through rice areas of the Central Valley and are likely 

to be exposed to thiobencarb based on their high relative abundance during peak 

thiobencarb applications (Table 28). However, while the levels and durations of 

thiobencarb exposure experienced by migrating adults may adversely affect individuals, 

they are not expected to affect populations because exposures are expected to be confined 

to infrequent incidents of early release of water from treated fields and drift of liquid 

formulations to floodplain habitats (Table 36). 
286

Exposure of juveniles to thiobencarb is limited. While some individuals may experience 

adverse effects of exposure, the levels and durations of such exposure are not expected to 

affect juvenile populations because rearing and migrating juveniles have a low 

probability of exposure to thiobencarb applied near their habitat (reference table 7, table 

8). 

Anticipated effects from reduced riparian zone function will likely be limited based on 

the narrow application period of thiobencarb May-June and will primarily be affected 

where riparian areas are exposed to drift.

Sacramento River Winter-run Chinook (Endangered species)

The Sacramento River Winter-run Chinook salmon ESU is now comprised of a single 

population. As there is only one population, any population level effects will directly impact the 

species. This population spawns and rears in the main stem of the Sacramento River below 

Keswick Dam. Abundance and productivity have fluctuated greatly over the past two decades. 

The genetic diversity of this population has been reduced substantially through small population 

sizes and the influence of hatchery fish. The large fluctuations in productivity and abundance 

indicate that the species is highly vulnerable to extinction. 

We rank the potential for thiobencarb exposure to cause species level consequences to 

Sacramento River Winter-run Chinook as low for the following reasons:

Exposure of migrating and spawning adults to thiobencarb is limited. While some adults 

migrating through rice growing areas will be exposed to thiobencarb, all known spawning 

is located at higher elevations than rice fields (current Figure 6). We expect that some 

individuals will be exposed to thiobencarb drift and runoff during migration (current 

Tables 9, 10), possibly experiencing adverse effects, but, overall, we do not expect 

population level consequences for adult winter-run Chinook (Table 36). 

Exposure of juveniles to thiobencarb is limited. Spawning grounds are located up-stream 

of rice growing areas (Current Figure 6), and most juvenile migrations through Central 

Valley rice-growing areas do not occur during the time period when thiobencarb is 
287

applied (Snider and Titus 2000, Bajjaliya and Vincik, 2008, USFWS 2010, 2012). Thus, 

the probability of exposure to thiobencarb via drift and runoff into aquatic habitats in 

close proximity to rice fields is low.

Anticipated effects from reduced riparian zone function will likely be limited based on 

the narrow application period of thiobencarb May-June and will primarily be affected 

where riparian areas are exposed to drift.

California Central Valley Steelhead (Threatened species)

The DPS consists of 81 historical and independent populations. Its spatial structure has been 

greatly reduced by loss of habitat diversity and tributary access from dams. Available 

information shows a significant long-term downward trend in abundance (NMFS, 2009a). 

Population losses and reduction in abundance have reduced the genetic diversity that existed 

within the DPS.

We rank the potential for thiobencarb exposure to cause species level consequences to Central 

Valley steelhead as low for the following reasons:

Exposure of migrating and spawning adults to thiobencarb is limited. Co-occurrence of 

thiobencarb application with spawning adults is unlikely based on the location of known 

spawning areas relative to where rice is grown (current Figure 7). There is also little 

temporal overlap between thiobencarb application and adult migration through rice 

growing areas (currently Table 11). 

Exposure of rearing and migrating juveniles to thiobencarb is limited. Co-occurrence of 

thiobencarb with egg incubation, fry emergence, and early rearing is unlikely based on 

the location of known spawning areas relative to where rice is grown (current Figure 7). 

There may be some exposure of outmigrating juveniles to thiobencarb, especially in the 

San Joaquin and Mokelumne Rivers, however very little information is available on 

steelhead use of these rivers (current Tables 11, 12). Overall, while some individual 

juveniles may experience adverse effects during migratory periods, we do not expect 

population level consequences (Table 36). 
288

Anticipated effects from reduced riparian zone function will likely be limited based on 

the narrow application period of thiobencarb May-June and will primarily be affected 

where riparian areas are exposed to drift.

Table 60. Summary Findings for Species 

Species

Co-occurrence of action 

with rice growing areas 

in California

Potential for reduction in reproduction, numbers, 

distribution

Populations Species

Central Valley Springrun Chinook (T)

Yes Low Low

Sacramento River 

Winter-Run (E)

Yes Low Low

California Central Valley 

Steelhead (T)

Yes Low Low

T = threatened; E = endangered
289

12. Effects of the Proposed Action on Designated Critical Habitat

NMFS’ critical habitat analysis determines whether the proposed action will likely destroy or 

adversely modify critical habitat for ESA-listed species by examining potential reductions in the 

conservation value of the essential features of designated critical habitat. Our analysis does not 

rely on the regulatory definition of “adverse modification or destruction” of critical habitat. 

Instead, we rely on the statutory provisions of the ESA, including those in section 3 that define 

“critical habitat” and “conservation”, those in section 4 that describe the designation process, and 

those in section 7 setting forth the substantive protections and procedural aspects of consultation.

In this section NMFS evaluates the potential consequences to designated critical habitat from 

exposure to the stressors of the proposed action. The pathway of analysis is presented in Figure 

40. It is similar in structure to the jeopardy analysis, but focuses on whether the proposed action 

is likely to destroy or adversely modify designated critical habitat for listed Pacific salmonids. 

We first determine the potential for critical habitat to be exposed to the stressors of the proposed 

action. If we conclude that critical habitat is likely to be exposed, we assess the consequences of 

that exposure on the quality, quantity, or availability of one or more of those primary constituent 

elements (PCEs) that comprise critical habitat. Water quality, forage (prey availability), and 

natural cover (riparian plants) are key attributes of salmonid PCEs that are susceptible to the 

stressors of the action. Water quality encompasses a range of typically measured parameters, 

including dissolved oxygen, temperature, turbidity, and presence of chemical contaminants in 

sufficient concentrations to adversely affect aquatic organisms. 

Because the proposed action would degrade water quality by introducing thiobencarb and other 

chemicals, we evaluate concentrations of thiobencarb likely to adversely affect fish, prey, and 

terrestrial and aquatic plants as measures of degraded water quality. This analysis is conducted 

by comparing toxicity information reviewed earlier in the Response section with expected 

concentrations in salmonid habitats. Similarly, we evaluate adverse effects to salmonid prey to 

determine the effects of the action on prey availability and forage, a key attribute for many 

salmonid PCEs. 
290

Figure 40. Assessment Framework for Designated Critical Habitat Assessment Framework for 

Designated Critical Habitat.

Analyses based 

on the best

scientific and 

commercial data 

available on 

pesticide 

products use, 

transport, fate, 

toxicity, and 

species ecology

Habitat responses

Exposure Analysis Response Analysis

Response Profile

Effects of pesticide 

products on designated 

critical habitat 

Exposure Profile

Co-occurrence of pesticide 

products and geographic 

range of critical habitat

Risk 

Characterizatio

n

Stressors of the 

Action 

Effects on water quality 

and prey availability within 

Primary Constituent 

Elements (PCEs)

Effects on the function of 

the critical habitat to 

support the intended 

conservation role for the 

species

Can EPA ensure its action is 

not likely to adversely modify 

or destroy designated critical 

habitat?

Analyzed within the 

context of the 

Environmental Baseline 

(including multiple 

stressors such as 

temperature and 

environmental mixtures 

of pesticides); the 

Status of Listed 

Resources; and 

Cumulative Effects

Conducted in the 

Integration and 

Synthesis Section
291

We formulated several risk hypotheses to assess potential changes in PCEs of designated critical 

habitat based on: 1) the likely concentrations that would be observed where critical habitat is 

exposed to chemicals derived from pesticide applications; and 2) the response of PCEs to the 

anticipated concentrations.

NMFS used the assigned conservation values (high, medium, and low) of watersheds within each 

ESU/DPS for the PCEs of critical habitat identified for each life stage common to listed 

salmonids (described in the Status of Listed Resources section). Because watersheds with high 

conservation value are essential to the conservation of the species, reductions in the quantity, 

quality, or distribution of the PCEs supporting that watershed would be expected to adversely 

affect the function of critical habitat to support its intended conservation role. We assess these 

watersheds within the Integration and Synthesis for Designated Critical Habitat section.

NMFS has designated critical habitat for each of the three species. The action area for this 

Opinion encompasses all designated critical habitat for the three listed Pacific salmonids in the 

Central Valley. The PCEs for each listed species are described in the Status of Listed Resources

section of this Opinion. As the species of salmonids addressed in this Opinion have similar life 

history characteristics, they share many of the same PCEs. These PCEs include sites that support 

one or more life stages (sites for spawning, rearing, migration, and foraging) and contain 

physical or biological features essential to the conservation of the ESU/DPS, including: 

1. freshwater spawning sites; 

2. freshwater rearing sites; 

3. freshwater migration corridors;

4. estuarine areas; 

5. nearshore marine areas23; and

6. offshore marine areas.

 

23 Nearshore marine areas are free of obstruction and excessive predation with: (i) water quality and quantity 

conditions and forage, including aquatic invertebrates and fishes, supporting growth and maturation; and (ii) natural 

cover such as submerged and overhanging large wood, aquatic vegetation, large rocks and boulders, and side 

channels (70 FR 52488; 73 FR 7816).
292

Water quality, prey availability (forage), and natural cover in freshwater and estuarine areas are 

susceptible to the effects of the proposed action where they overlap with the stressors of the 

action. Effects to water quality and prey availability will be evaluated to determine the 

likelihood of reducing the quality of PCEs such as spawning and rearing sites, or migration 

corridors. Given the use and environmental fate profile of the pesticide formulations containing 

thiobencarb, we do not expect offshore marine areas and nearshore marine areas to be affected. 

Therefore, a risk hypothesis was not developed for offshore marine areas or for nearshore marine 

areas and further evaluation of these PCEs are unwarranted. 

Good water quality is a necessary attribute of all PCEs to support the conservation role of 

designated critical habitat. Water quality is clearly degraded when pesticides and other stressors 

of the action reach levels in salmonid habitat that are sufficient to affect aquatic organisms, 

including those that reduce individual fitness of exposed salmonids. Impacts to salmonid fitness 

were evaluated earlier in the document and these impacts are used as indicators of degraded 

water quality. We evaluate exposure and effect concentrations presented earlier in the Effects of 

the Proposed Action section to determine whether PCEs are affected. We re-evaluate the 

information to determine potential effects to PCEs. 

We also evaluate effects on salmonid prey because forage is an essential attribute of all PCEs 

except in spawning sites. Freshwater juvenile rearing and migratory habitats as well as some 

estuarine and nearshore marine areas must provide sufficient forage that support salmonid 

growth and development in the Central Valley. Reductions in the abundance of prey items can 

decrease the quality of rearing, migration, and estuarine PCEs, as they will support fewer 

individuals, especially during a salmonid’s first year of survival. Reductions in prey can reduce 

a PCE’s potential to support salmonids (juvenile development, growth, maturation, survival), 

thereby reducing the carrying capacity of critical habitat. 

We evaluated toxicity assessment endpoints including prey survival (EC50/LC50), prey growth, 

prey drift, prey reproduction, prey abundance, health condition of invertebrate aquatic 

communities (using indices of biological integrity), aquatic primary producers, terrestrial plants 

in riparian zones, and recovery of aquatic communities following thiobencarb exposure to 
293

determine whether expected concentrations are sufficient to affect PCEs for salmonid critical 

habitats. 

Exposure of designated critical habitats to the stressors of the action:

Designated critical habitat for the three species is located within the action area. Many 

freshwater areas overlap with the allowable uses of thiobencarb authorizations (Appendix 6). 

The stressors of the action contaminate these habitats primarily via drift and runoff (including 

from irrigation returns), and to a lesser extent from atmospheric deposition. Once in salmonid 

habitats of the Central Valley, thiobencarb persists for a period of time; the extent of which is 

dependent on the chemical, biological, and physical environment of the contaminated aquatic 

habitats. Expected concentrations of other/inert ingredients and adjuvants added to formulations 

prior to application remain unknown - a data gap. 

Table 49 (summary exposure section depicting concentrations) shows expected concentrations of 

thiobencarb and its degradates that were derived from EPA BEs, EPA incident data, surface 

water monitoring data, and NMFS exposure modeling estimates. These data will be discussed in 

the context of spawning, rearing, migrating, and estuarine PCEs. The vast majority of exposure 

information applies more readily to freshwater habitats compared to estuarine where much less 

information is available.

Responses of salmonid habitats to the stressors of the action

If PCEs are exposed to the stressors, we evaluate the level at which thiobencarb, any toxic 

degradates, and inert/other ingredients adversely affect water quality, prey availability, and 

natural cover (terrestrial and aquatic primary production). For many of the other ingredients, 

recommended tank mixtures, and degradates, not only was there no available exposure 

information, but also little to no toxicity information. In the Response and Risk Characterization

sections of the Effects of the Proposed Action, we showed that applications of thiobencarb can 

result in concentrations that may reduce salmonid survival, growth, swimming, respiration, and 

reduce riparian vegetation, all which independently translate to a degradation of water quality 

Table 53. These types of individual fitness consequences demonstrate a degradation of water 

quality in affected habitats. 
294

We summarized the available toxicity information in the Response Analysis (Table 53). It is 

important to note that the toxicity of thiobencarb is variable depending on the biological endpoint 

(e.g., acute lethality to fish and invertebrates), the levels expected in salmonid habitats, the 

presence of other AChE-inhibiting pesticides, and whether elevated water temperatures occur. 

12.1. Risk Characterization 

NMFS reviews the status of designated critical habitat affected by the proposed action separate 

from species effects by examining the condition and trends of primary constituent elements 

(PCEs) throughout the action area. PCEs are the physical and biological features identified as 

essential to the conservation of listed species. The PCEs for salmonids are identified in The 

Status of the Species section. We determined the PCEs at risk from the use of thiobencarb are 

freshwater rearing, freshwater migration, and time spent in estuarine areas. Spawning areas for 

each of the species are upstream of current rice growing areas. Therefore, the freshwater 

spawning PCE is not considered at risk from this action. Also, considering the use of 

thiobencarb is limited to rice which is more than 40 miles from the California coast, we do not 

consider nearshore or offshore marine PCEs at risk from this action. 

We use the toxicity information presented earlier in the Response section to evaluate the 

scientific lines of evidence that support or refute risk hypotheses developed for designated 

critical habitats. Freshwater rearing sites and migration corridors within designated critical 

habitats are likely to be exposed to the stressors of the action over the 15-year registration 

duration. We estimate expected concentrations and durations of exposure for these habitats 

based on pesticide use information, surface water monitoring data, EPA modeling estimates, and 

NMFS modeling estimates. For each risk hypothesis below, we qualitatively weigh the evidence 

to determine whether the PCE attributes of water quality and/or prey availability are affected. 

We ultimately determine whether the degradation of water quality, prey availability, and natural 

cover within freshwater rearing and migration corridors will rise to the level expected to reduce 

the intended conservation role of designated critical habitats, which is evaluated within the 
295

Integration and Synthesis for Designated Critical Habitat section. The final conclusion of 

whether EPA’s proposed action with thiobencarb containing end use products are likely to 

adversely modify or destroy a species’ designated critical habitat is provided in the Conclusion

section.

Risk hypothesis 1. Exposure to the stressors of the action is sufficient to degrade water quality, 

natural cover, and /or reduce prey availability in freshwater rearing and migration areas. 

Freshwater rearing and migration areas need to provide good water quality, abundant forage, and 

natural cover to support juvenile development, mobility and survival. Reductions in any of these 

attributes from the stressors of the action can limit the existing and potential carrying capacity of 

rearing and migration areas and subsequently reduce their conservation value. Recovery of listed 

salmonid populations in the Central Valley is tied closely to the success of juveniles to fully 

develop, mature, and grow during freshwater residency periods. All species of Pacific salmonids 

spend some amount of time in freshwater feeding and rearing areas. Chinook and steelhead 

spend much longer periods rearing in freshwater systems with steelhead trout spending up to 

several years before migrating to the ocean.

 

Freshwater rearing areas are diverse, extensive, and complex sites that can range from small, 

shallow, intermittent floodplain habitats to channel edges of large river systems such as the 

Sacramento River. Salmon recovery plans also call for floodplain connectivity and other 

restoration activities to improve rearing. Many freshwater salmonid rearing sites are located in 

floodplains where shallow, low flow habitats are at high risk of thiobencarb drift and runoff. The 

Yolo Bypass, when inundated, is one example. Habitats such as the Yolo provide abundant prey 

and are important foraging areas for developing juveniles (Sommer et al. 2001). 

Expected floodplain concentrations of thiobencarb are shown in the Exposure and Response 

Integration section above (Table 53). At these concentrations, water quality would be affected 

because concentrations exceed toxicity thresholds for fish and aquatic invertebrate survival as 

well as for plankton and aquatic macrophyte growth and survival. Additionally, we expect 

modification to riparian and floodplain vegetation will result in decreased prey production, 
296

decreased shading of these habitats, and an increase in the occurrence of elevated temperatures 

that are stressful to salmonids in freshwater rearing sites.

The habitats most likely affected would be those areas that receive primary drift from application 

or direct runoff from early releases of treated rice irrigation water. In these circumstances, 

thiobencarb may reduce forage availability because concentrations can reach levels that kill prey 

and affect their growth and reproduction (shown in the Exposure and Response Integration

section above (Figure 39).

Thiobencarb is applied primarily in May, and rice discharge waters are typically released in June 

or July after the required holding period. As previously stated, the Yolo Bypass floodplain is an 

important rearing area for juvenile salmonids when inundated by flood waters. Rice planting and 

subsequent thiobencarb application in the Yolo occurs after the water has receded and the 

salmonids are gone.24 During the rice growing period, the Yolo ceases to be an available rearing 

area. While drift or runoff of thiobencarb may affect some salmonid prey and natural cover on 

the Yolo, we do not know to what extent the vegetation and prey will rebound prior to 

subsequent flood events. 

The extent surface waters flowing through rice growing areas are used by salmonids for rearing 

(i.e., not merely a migratory corridor) is unknown. We presume rearing occurs within rice 

growing areas. Rearing in the Sacramento River for Chinook and steelhead is likely along

shallow edge habitats where emergent vascular plants and/or over-hanging vegetation provide 

shelter and prey resources. We would expect prey to be adversely affected in shallower areas 

receiving direct drift or early release of rice irrigation waters. Water quality would be adversely 

affected at those sites where runoff and drift occur. 

We also anticipate that in some situations, surface waters in the CA Central Valley receiving 

thiobencarb drift or early release of rice irrigation waters will alter characteristics of the plant 

 

24 Some Chinook may be entrained in perennial ponds within the Yolo floodplain. These fish would be exposed to 

excess temperatures and predation and are considered lost to the population Sommer, T., Harrell, W., and Nobriga, 

M. 2005a. Habitat use and stranding risk of juvenile Chinook salmon on a seasonal floodplain. North American 

Journal of Fisheries Management, 25(4): 1493-504.
297

communities in freshwater rearing and migration areas, thereby decreasing the availability of 

cover for salmonids and their prey. Aquatic plants provided natural cover for salmonids as well 

as prey habitat. Thiobencarb is acutely toxic to a range of terrestrial and primary producers. 

Aquatic primary producers showed reduced growth (EC50s) at concentrations as low as 17 μg/L 

and as high as 4000 μg/L (Table 51). Both values were reported for green algae. No studies 

were available that tested aquatic community responses to thiobencarb in running water systems 

and none that evaluated salmonid habitats within the Central Valley. The bioassay results 

indicate that following 48-h continuous exposure to thiobencarb, reductions in aquatic primary 

producers can occur. Aquatic vegetation provides cover to salmonids from avian predation as 

well as protection from larger fish. In addition, a diverse vascular plant community provides 

important substrate for an array of insect species upon which young salmon prey. We conclude 

that effects to primary producers are likely, particularly in salmonid habitats receiving 

thiobencarb drift from aerial applications and those areas exposed to early release of rice water. 

Collectively the data indicate that expected concentrations of thiobencarb from drift and early 

release of treated rice water are sufficient to adversely affect water quality, salmonid prey, and 

natural cover attributes of freshwater rearing and migration PCEs. Therefore, we evaluate these 

effects in order to determine whether the conservation value of species’ designated critical 

habitats will be reduced (see Integration and Synthesis for Designated Critical Habitat section 

below). The final conclusion of whether EPA’s proposed action with thiobencarb containing end 

use products are likely to adversely modify or destroy a species’ designated critical habitat is 

provided in the Conclusion section.

Risk hypothesis 2. Exposure to the stressors of the action is sufficient to degrade riparian areas 

adjacent to rearing and migration corridors.

Few studies were found on the effects of thiobencarb on riparian and aquatic (non-target) 

vegetation. Registrant submitted data indicated that thiobencarb is highly toxic to non-target 

plants where EC25s ranged from 0.019 – 3.1 lbs/A active ingredient. In a field study, tolerant and 

susceptible species of non-target bayou and levee vegetation adjacent to rice fields were recorded 

following treatment with Bolero 8 EC (Lauck 1979). All species of levee vegetation examined 
298

had emerged and were in a mature stage of development and new foliage of the bayou woody 

plants was rapidly expanding during the period of Bolero usage. The author did not evaluate 

long term effects to plants. All species had been directly exposed to Bolero by an aerial 

application at the label rate, 4 lbs/A. One plant showed symptoms of injury due to Bolero 

applications, reed canary grass. Growth and maturity was reduced by 55 percent relative to 

untreated plants (Lauck 1979). Based on these results, we conclude that grasses and sedges in 

riparian areas may be susceptible to adverse effects in areas receiving drift from thiobencarb 

applications. Vascular plants and other macrophytes are important structural elements that 

provide important cover to salmonids in their preferred freshwater rearing areas. 

It is unclear how established woody plants in the riparian corridor of the Sacramento and San 

Joaquin may respond to varying concentrations of thiobencarb containing drift. We can only 

presume that, at a minimum, new growth may be adversely affected. Drift may also adversely 

affect the growth performance of seedling trees. If this were the case, long-term health and 

function of those riparian zones receiving drift annually may be impaired. In extreme cases, 

where woody vegetation is eliminated from riparian zones to the extent that bank destabilization 

occurs, altered stream hydrology could affect the availability of other cover including rocks, side 

channels, and undercut banks. One extreme example suggested that the application of an 

herbicide to a riparian zone caused major long-term changes to the hydrology of a stream and 

degraded fish habitat (http://water.epa.gov/scitech/datait/tools/warsss/streamero.cfm). 

Herbicide-induced changes to vegetative communities in the riparian zone and aquatic habitat 

also have implications for the availability of prey as salmonids consume both terrestrial and 

aquatic insects. As plant communities are modified in riparian zones, the species that rely on 

them will also be affected. Habitat responses in these environments are expected to be variable 

and will depend on the sensitivity of existing plants to thiobencarb in conjunction with the other 

stressors in the environment. 

Collectively expected concentrations of thiobencarb from drift are sufficient to adversely affect 

riparian areas adjacent to rearing and migration corridors affecting water quality, salmonid prey, 

and natural cover attributes of freshwater rearing and migration PCEs. Therefore, we evaluate 

these effects in order to determine whether the conservation value of species’ designated critical 
299

habitats will be reduced (see Integration and Synthesis for Designated Critical Habitat section 

below). The final conclusion of whether EPA’s proposed action with thiobencarb-containing 

end use products are likely to adversely modify or destroy a species’ designated critical habitat is 

provided in the Conclusion section.

Risk hypothesis 3. Exposure to the stressors of the action is sufficient to degrade water quality, 

natural cover, and/or reduce prey availability in estuarine areas.

Estuarine areas require good water quality to support juvenile and adult physiological transitions 

between fresh water and salt water as well as to provide juvenile and adult cover and prey 

resources sufficient to support survival, growth, and maturation. Prey resources for Pacific 

salmonids within estuaries include a diverse group of organisms from aquatic invertebrates to 

small fishes. The allowable uses of the stressors of the action overlap with estuaries designated 

as critical habitat (Kuivila and Jennings 2007). 

Contamination of estuaries occurs via drift, runoff, and atmospheric deposition. Streams and 

rivers flowing into estuaries act as conveyor belts as they transport the stressors of the action 

from areas higher in watersheds (Johnson et al. 1997). EPA did not provide estimates for 

thiobencarb concentrations in estuarine habitats. A study to characterize the temporal inputs of 

thiobencarb (along with five other pesticides) into San Francisco Bay found thiobencarb in 28 

percent of the samples. The samples were taken 8 km downstream of the confluence of the 

Sacramento and San Joaquin rivers beginning in mid-May, following the release of rice-field 

water (Kuivila and Jennings 2007). Water samples were collected daily or twice-daily from midJanuary through mid-July. The maximum concentration detected was 66 ng/L in late May. It is 

impossible to determine how representative this value is relative to other areas in the estuary that 

may be occupied by listed salmonids, or relative to other values on subsequent years. However, 

the sample site selected receives the majority of flow from the Sacramento prior to entering the 

San Francisco Bay (Kuivila and Jennings 2007). The maximum value reported, 66 ng/L, is very 

low relative to freshwater toxicity endpoints we reviewed (Table 53). The occurrence of 

thiobencarb followed the application season (Figure 24), and the concentration of thiobencarb 

spiked following release of rice-field water.
300

The available toxicity information for estuarine and marine organisms for thiobencarb is 

presented in Table 61 as reported in EPA’s Red Legged Frog BE for thiobencarb. The estuarine 

aquatic toxicity data are from survival assays for the Atlantic silverside (fish) and mysid 

(estuarine invertebrate), and chronic results were reported for the Atlantic silverside and 

Opossum shrimp. The available studies indicate a similar range of sensitivity to thiobencarb. 

The toxicity values reported are three orders of magnitude greater than concentrations found in

the estuary (Kuivila and Jennings 2007). Without additional toxicity studies on specific prey 

organisms residing in the Sacramento-San Joaquin estuary, it is impossible to say for sure how 

thiobencarb exposures similar to those reported would affect salmonid prey. However, if 

responses are similar to those found for the opossum shrimp, concentrations would have to be 

100 times greater for an observable effect. It is unlikely that concentrations reported would be 

exceeded elsewhere in the estuary. Collectively, data do not suggest that estuarine PCEs are 

adversely affected, therefore we do not carry this hypothesis forward in our analysis of impacts 

to the conservation value of designated critical habitat.

Table 61. Thiobencarb’s assessment endpoint toxicity values (μg/L) for saltwater aquatic 

organisms presented in California Red-Legged Frog (CRLF) BEs.

Assessment 

endpoint

Test Species

Toxicity Value

(μg/L)

Citation

Estuarine/marine 

fish

Acute Atlantic silverside 

Menidia menidia

96-h LC50 = 204

(96 h)

(EPA 2009b)

Chronic Atlantic silverside 

Menidia menidia

NOEC = 6

(estimated)

(EPA 2009b)

Estuarine/marine 

invertebrates

Acute Mysid

Americamysis bahia

96-h LC50 = 150 

(110 -200)

(EPA 2009b)

Chronic Opossum Shrimp

Neomysis mercedis

NOEC = 3.2

LOEC = 6.2

(EPA 2009b)
301

Summary of the Effects of the Action on PCEs

We conclude that the available information on exposure and response of aquatic habitats to the 

stressors of the action supports two of the three risk hypotheses (Table 62). We expect essential 

physical and biological features will be reduced in rearing and migratory habitats, and that 

riparian vegetation will be negatively affected. Next, within the Integration and Synthesis of 

Effects to Designated Critical Habitat section, we evaluate whether these adverse changes to 

PCEs affect the conservation value of designated critical habitat.
302

Table 62. Risk to the primary constituent elements (PCEs) from stressors of the proposed action.

PCE

Essential Physical or Biological 

Feature

PCE at Risk from 

Stressors of the 

Action?

Freshwater Spawning Water Quality

Water Quantity

Substrate

No

No

No

Freshwater Rearing Water Quantity

Floodplain Connectivity

Water Quality

Forage

Natural Cover

No

No

Yes

Yes

Yes

Freshwater Migration Free of Obstructions

Water Quality

Water Quantity

Natural Cover

No

Yes

No

Yes

Estuarine Areas Free of Obstruction

Water Quality

Water Quantity

Salinity

Forage

Natural Cover

No

No

No

No

No

No

Nearshore Marine Free of Obstruction

Water Quality

Water Quantity

Natural Cover

Forage

No

No

No

No

No

Offshore Marine Water Quality

Forage

No

No
303

12.2. Integration and Synthesis for Designated Critical Habitat

This section describes NMFS’ assessment of the likelihood that EPA’s registration of 

thiobencarb will destroy or adversely modify designated critical habitat for three salmonids 

species covered in this Opinion. Each of the three species has similar PCEs, as described in the 

Effects to Designated Critical Habitat section. These PCEs are sites that support one or more 

life stages and include:

1. freshwater rearing sites; 

2. freshwater migration corridors;

3. estuarine areas; 

4. nearshore marine areas25; and

5. offshore marine areas.

We determined the PCEs at risk from thiobencarb are freshwater rearing and freshwater 

migration sites (Table 63). These sites contain physical or biological features essential to the 

conservation of the ESU/DPS. Physical features include cover, substrate, water temperature, and 

water quality. Biological attributes include forage i.e., prey. We do not expect freshwater 

spawning, estuarine areas, nearshore marine or offshore marine areas to be affected by the 

stressors of the action.

As noted in salmonid recovery plans, salmonids require cool water, free of contaminants. Water 

free of contaminants promotes normal fish behavior for successful migration, spawning, and 

juvenile rearing. In the juvenile life stage, salmonids also require stream habitat providing 

adequate forage. Sufficient forage is necessary for juveniles to maintain growth which 

subsequently reduces freshwater predation mortality, increases overwintering success, supports 

smoltification, and improves their ocean survival. 

 

25 Nearshore marine areas are free of obstruction and excessive predation with: (i) water quality and quantity 

conditions and forage, including aquatic invertebrates and fishes, supporting growth and maturation; and (ii) natural 

cover such as submerged and overhanging large wood, aquatic vegetation, large rocks and boulders, and side 

channels (70 FR 52488; 73 FR 7816).
304

We expect thiobencarb and its degradates, other ingredients within its formulations, tank mixture 

chemicals to affect natural cover (riparian plants), water quality, and forage (prey abundance). 

Destruction or adverse modification of designated critical habitat is evaluated based on whether 

the stressors of the action are expected to cause appreciable reductions in natural cover, water 

quality, or prey abundance at juvenile freshwater rearing and migration sites. The stressors of 

the action include thiobencarb, degradates, inert/other ingredients in formulations, surfactants, 

and tank mixtures; and their individual and collective interactions when applied to rice fields in 

California’s Central Valley. Data are not available for some of these other stressors, thus they 

are evaluated in a qualitative fashion. 

Table 63. Key Findings from the Effects of the Proposed Action Section to Critical Habitat 

Section of Biological 

Opinion

Summary Finding 

Exposure Habitats of the three species are expected to be exposed to thiobencarb and 

its associated stressors.

Response Anticipated levels of thiobencarb and its associated stressors are sufficient to 

impair several biological and physical attributes of primary constituent 

elements.

Risk Characterization Freshwater rearing and freshwater migration primary constituent elements are 

expected to be reduced. 

Environmental Baseline Pesticides, elevated temperature, water flow, other pollutants, and dams are 

some of the stressors affecting salmonid habitats within the Central Valley. 

Water quality is severely degraded throughout the Central Valley. California’s 

current risk reduction measures for thiobencarb reduced exposure to 

designated critical habitats.

Cumulative Effects Increases in human population and a continuation of high intensity agriculture 

are anticipated within the Central Valley. Some beneficial activities are also 

anticipated including river restoration projects and best management practices 

to improve water quality.

Responses of designated critical habitat to thiobencarb and its associated chemical stressors is a 

function of exposure concentrations and durations. The longer or more frequent the exposure the 

greater the likelihood of affecting PCEs. We found that designated critical habitats proximate to 

rice fields are at the greatest risk from thiobencarb containing drift. Drift is expected during 

application of the liquid formulations; however we expect minimal drift from granular 

formulations. Based on CDPR use statistics, granular applications comprise approximately 75% 

and liquid applications comprise the remaining 25%. In rice growing areas of the Central 

Valley, concentrations from spray drift in flood plain habitats and small streams may attain 
305

levels that impair swimming, impair breathing, kill salmonid prey, and in some cases kill 

salmonids. Additionally, we anticipate some riparian plants to die or experience reduced growth 

from drift. Overall, we expect these occurrences to be infrequent, although not absent. We also 

expect in stream effects to be magnified when other pesticides are present or when water 

temperatures are elevated. 

The other high risk scenario involves early releases of thiobencarb-containing irrigation water. 

Rice water may be released into drainages prematurely during high rain events or from leaky 

dikes. High concentrations that would kill salmonids, their prey, inhibit aquatic primary 

producers, all of which degrade water quality may occur if rice water is released prior to the end 

of holding periods, particularly rice water released within a few days of application (Table 53). 

The magnitude of effects depends, in part, on how soon the water is released after application of 

thiobencarb, how far salmonid-bearing waters are located downstream from the release, and how 

long they are exposed. We anticipate these occurrences as rare given the information reviewed 

and the current state regulations already in place (Table 43).

We used a GIS overlay to determine potential overlap of rice growing areas and designated 

critical habitat distributions for the three species to (Appendix 6). This overlay is a different 

approach compared to earlier Opinions on EPA’s registrations of pesticides where we used land 

use classifications i.e., agricultural, forestry, urban/developed, etc. were used rather than a 

specific crop. We selected the current approach because thiobencarb is used exclusively on rice 

in California’s Central Valley. In the Effects of the Proposed Action Section for critical habitat, 

we found that freshwater rearing and freshwater migration PCEs for the three species were at 

risk, particularly water quality, forage, and natural cover (Table 38: Risk to the PCEs). The GIS 

overlay shows that current rice fields overlap extensively with portions of each species’ 

designated critical habitats. The critical habitats that overlap rice areas have been ranked as 

having “high” conservation value (Appendix 6).26

 

26 At this point in time, conservation values within the designated critical habitat for Sacramento River winter-run 

Chinook have not been assigned. For this Opinion, we assumed each watershed to have a high conservation value. 
306

We evaluate whether anticipated effects to PCEs reduce appreciably the overall conservation 

value of the species’ designated critical habitat given the current status, environmental baseline, 

and cumulative effects. All three species designated critical habitats include the Sacramento 

River and its tributaries that overlap with rice growing in the Central Valley. Spawning, rearing 

and migration areas are currently degraded by elevated temperatures and lost access to historic 

spawning sites. Rearing PCEs in the Sacramento River are degraded by loss of floodplain 

habitats, thereby reducing foraging opportunities. Migration PCEs are degraded by lack of 

natural cover along the migration corridors and degraded water quality from a host of chemical 

and non-chemical stressors. Juvenile migration is further affected by water diversions and by 

two large state and federal water-export facilities along the Sacramento- San Joaquin Delta. In 

aggregate, many of the designated critical habitats within the Central Valley are in a degraded 

state and expected to remain so throughout the 15 years of the action.

Waters that show elevated temperatures and/or contain other pesticides that share a mode of 

action with thiobencarb i.e., other AChE inhibitors, effects of the action will be more pronounced 

and dependent on potency of the chemicals as well as the magnitude of elevated temperature.

Water quality, forage, and riparian plants were the assessment endpoints expected to be affected 

by use of thiobencarb on rice. We do not anticipate these endpoints to be degraded to the degree 

that would reduce appreciably the high conservation value of rearing and migratory PCEs (Table 

64). The high risk scenarios of spray drift and early release of thiobencarb-containing rice water 

are expected to be infrequent. Additionally, the toxic levels necessary to kill and impair plants 

are not expected in the vast majority of designated critical habitats. 

Table 64. Summary findings for Designated Critical Habitats 

Species

Co-occurrence of action 

with rice growing areas 

in California

Designated critical habitat 

PCEs affected Appreciably reduce 

conservation values

Central Valley Springrun Chinook (T) Yes Yes No

Sacramento River 

Winter-Run (E) Yes Yes No

California Central Valley 

Steelhead (T) Yes Yes No

T = Threatened; E = Endangered
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13. Conclusion

After reviewing the current status of Central Valley Spring-run Chinook, Sacramento River 

Winter-run Chinook, and California Central Valley Steelhead, the environmental baseline for the 

action area, the effects of the proposed action, and the cumulative effects, it is NMFS’ Opinion 

that the registration of thiobencarb is not likely to jeopardize the continued existence of these 

endangered or threatened species nor is it likely to destroy or adversely modify designated 

critical habitat of these three species (Table 65).

Table 65. Conclusions for EPA's Reregistration of Thiobencarb 

Species ESU Jeopardy

Adverse 

Modification/Destruction 

of Designated Critical 

Habitat

Chinook

Central Valley 

Spring-run

No No

Sacramento River 

Winter-run

No No

Steelhead California Central 

Valley No No
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14. Incidental Take Statement

Section 9(a)(1) of the ESA prohibits the taking of endangered species without a specific permit 

or exemption. Protective regulations adopted pursuant to section 4(d) of the ESA extend the 

prohibition to threatened species. Take is defined as to harass, harm, pursue, hunt, shoot, wound, 

kill, trap, capture or collect, or to attempt to engage in any such conduct (50 CFR 222.102). 

Harm is further defined by NMFS to include significant habitat modification or degradation that 

results in death or injury to listed species by significantly impairing essential behavioral patterns, 

including breeding, feeding, or sheltering. Incidental take is defined as take that is incidental to, 

and not the purpose of, the carrying out of an otherwise lawful activity conducted by the Federal 

agency or applicant (50 CFR 402.02). Under the terms of section 7(b)(4) and section 7(o)(2), 

taking that is incidental to and not intended as part of the agency action, whether implemented as 

proposed or as modified by reasonable and prudent alternatives, is not considered to be 

prohibited taking under the ESA provided that such taking is in compliance with the terms and 

conditions of this Incidental Take Statement. 

14.1. Amount or Extent of Take

As described earlier in this Opinion, this is a consultation on the EPA’s registration of pesticide 

products containing thiobencarb and their formulations as they are used in California’s Central 

Valley on rice and the effects of these applications on three listed ESUs/DPSs of Pacific 

salmonids. The EPA authorizes use of thiobencarb in the western U.S. in California on rice only 

as described in the Description of the Proposed Action and elsewhere in the document. The goal 

of this Opinion is to evaluate the impacts to NMFS’ listed resources from the EPA’s broad 

authorization of applied pesticide products. This Opinion is a partial consultation because 

pursuant to the court’s order, EPA sought consultation on only 26 listed Pacific salmonids under 

NMFS’ jurisdiction. Consultation with NMFS will be completed when EPA makes effect 

determinations on all remaining species under NMFS’ jurisdiction and consults with NMFS as 

necessary.
309

For this Opinion, NMFS anticipates the general direct and indirect effects that would occur from 

EPA’s registration of pesticide products California’s Central Valley to listed salmonids under 

NMFS’ jurisdiction during the 15-year duration of the proposed action. Recent and historical 

surveys indicate that listed salmonids occur in the action area, in places where they will be 

exposed to the stressors of the action. The RPM below provided in this Opinion is designed to 

reduce this exposure but not eliminate it. Pesticide runoff and drift of thiobencarb is likely to 

reach streams and other aquatic sites when they are applied to rice located adjacent to wetlands, 

riparian areas, ditches, off-channel habitats, perennial, intermittent, and ephemeral streams. 

These inputs into aquatic habitats are especially high during spray drift of liquid formulations 

and following early releases of thiobencarb-containing rice water. 

The range of effects of thiobencarb on salmonids includes direct and indirect toxicological 

effects. Within this range, effects include death of fish, impaired respiration and swimming. 

Adverse impacts to riparian vegetation could lead to increased water temperature, increased 

sedimentation from bank instability, reductions in cover, alterations to or decreases in prey 

production, and reduction in chemical and nutrient filtering from upland sources. Impacts to 

aquatic vegetation would reduce dissolved oxygen, natural cover, alter or reduce the prey base, 

affect growth, and lead to an increased susceptibility to predation. These results are not the 

purpose of the proposed action. Therefore, incidental take of listed salmonids is reasonably 

certain to occur over the 15-year duration of the proposed action.

Given the variability of real-life conditions, the broad nature and scope of the proposed action, 

and the migratory nature of salmon, the best scientific and commercial data available are not 

sufficient to enable NMFS to estimate a specific amount of incidental take associated with the 

proposed action. As explained in the Description of the Proposed Action and the Effects of the 

Proposed Action sections, NMFS identified multiple uncertainties associated with the proposed 

action. Areas of uncertainty include:

1. Unable to quantify effect of herbicides on salmon habitat due to variability in plant susceptibility 

to the herbicides and variability in species composition and density in the various locations.
310

2. Minimal information on exposure and toxicity for pesticide formulations, adjuvants, and 

other/inert ingredients within registered formulations;

3. Little information on permitted tank mixtures and associated exposure estimates;

4. Limited data on toxicity of environmental mixtures;

5. Responses from exposure to combinations of thiobencarb and other stressors in the baseline; 

6. Variable conditions of water bodies in which salmonids live.

NMFS therefore identifies, as a surrogate for the allowable extent of take, the ability of this 

action to proceed without any fish kills attributed to the legal use of thiobencarb, or any 

compounds, degradates, or mixtures in aquatic habitats containing individuals from any 

ESU/DPS. Because of the difficulty of detecting salmonid deaths, the fishes killed do not have 

to be listed salmonids. In general, salmonids appear to be more sensitive to thiobencarb than 

many other species of fish, so that if there are kills of other freshwater fishes attributed to use of 

pesticide products containing thiobencarb [or “to use of thiobencarb”], it is likely that salmonids 

have also died, even if no dead salmonids can be located. In addition, if stream conditions due to 

pesticide use kill less sensitive fishes in certain areas, the potential for lethal and non-lethal takes 

in downstream areas increases. A fish kill is considered attributable to thiobencarb, its 

metabolites, or degradates, if thiobencarb is known to have been applied in the vicinity, may 

reasonably be supposed to have run off or drifted into the affected area, and if surface water 

samples, or pathology indicate lethal levels of thiobencarb. NMFS notes that with increased 

monitoring and study of the impact of thiobencarb on water quality, particularly water quality in 

floodplain habitats, NMFS will be able to refine this incidental take statement, and future 

incidental take statements, to allow other measures of the extent of take. 

14.2. Reasonable and Prudent Measures

The measures described below are non-discretionary measures to avoid or minimize take that

must be undertaken by the EPA so that they become binding conditions of any grant or permit 

issued to the applicant(s), as appropriate, for the exemption in section 7(o)(2) to apply. The EPA 
311

has a continuing duty to regulate the activity covered by this incidental take statement. If the 

EPA (1) fails to assume and implement the terms and conditions or (2) fails to require the 

applicant(s) to adhere to the terms and conditions of the incidental take statement through 

enforceable terms that are added to the FIFRA label, the protective coverage of section 7(o)(2) 

may lapse. In order to monitor the impact of incidental take, the EPA must report a) the progress 

of the action and b) its impact on the species to NMFS OPR as specified in the incidental take 

statement [50 CFR§402.14(i)(3)].

To satisfy its obligations pursuant to section 7(a)(2) of the ESA, the EPA must monitor (a) the 

direct, indirect, and aggregate impacts to listed Pacific salmonids under NMFS’ jurisdiction of its 

long-term registration of pesticide products containing thiobencarb and (b) the direct, indirect, or 

aggregate impacts of pesticide misapplications in the aquatic habitats in which they occur. The 

purpose of the monitoring program is for the EPA to use the results of the monitoring data and 

modify the registration process in order to reduce exposure and minimize the effect of exposure 

where pesticides will occur in salmonid habitat. NMFS believes all measures described as part 

of the proposed action, together with use of the Reasonable and Prudent Measures and Terms 

and Conditions described below, are necessary and appropriate to minimize the likelihood of 

incidental take of listed species due to implementation of the proposed action.

The EPA shall:

1. Minimize the amount and extent of incidental take to the three species, CV spring run Chinook, 

Sacramento River winter run Chinook, and CV steelhead from use of pesticide products 

containing thiobencarb by reducing the potential of thiobencarb to reach salmon-bearing waters;

2. Monitor any incidental take or surrogate measure of take that occurs from the action.

14.3. Terms and Conditions

To be exempt from the prohibitions of section 9 of the ESA, within nine months following the 

date of issuance of this Opinion, the EPA must comply with the following terms and conditions. 

These terms and conditions implement the reasonable and prudent measures described above. 

These terms and conditions are non-discretionary.
312

1. To implement RPM 1, EPA shall include the following risk reduction measures on all 

thiobencarb labels registered in California or in the appropriate bulletins once the label 

refers to compliance with the bulletin. These reductions have a proven track record for 

reducing thiobencarb loading into surface waters, which is essential to minimize 

incidental take. The pesticide use limitations are adapted from California Department of 

Pesticide Regulation’s Prescribe Database and Pesticide Use Enforcement Standards 

Compendium, Volume 3 (Rev. 2-11), Appendix C, Sections C.1 and C.2.

Mandatory language follows:

Drift:

For sprayable or dust formulations: when the air is calm or moving away from salmon 

habitat, commence applications on the side nearest the habitat and proceed away from 

the habitat. When air currents are moving toward habitat, do not make applications 

within 200 yards by air or 40 yards by ground upwind from occupied habitat. The 

appropriate California county agricultural commissioner may reduce or waive buffer 

zones following a site inspection, if there is an adequate hedgerow, windbreak, 

riparian corridor or other physical barrier that substantially reduces the probability of 

drift.

Aircraft application equipment used to apply a pesticide spray solution shall be 

configured as follows: 

1. Functional boom length, measured from outboard nozzle to outboard nozzle, 

shall not exceed 75% of the overall wing span or rotor length. 

2. Boom pressure shall not exceed 40 pounds per square inch for the nozzles 

being used. 

3. The flow of liquid from each nozzle shall be controlled by a positive shutoff 

system. 

4. Nozzle orifices shall be directed backward, neutral to the airstream. 

5. Aircraft shall be equipped with: 

(a) Jet nozzles having an orifice of not less than one-sixteenth of an inch in 

diameter. Nozzles shall not be equipped with any device or mechanism 

which would cause a sheet, cone, fan, or similar type dispersion of the 

discharged material, except helicopters operating at 60 miles per hour or 

less may add a number 46 (or equivalent) or larger whirlplate; 

(b) Fan nozzles with a fan angle number not larger than 80 degrees and a 

flow rate not less than one gallon per minute at 40 pounds per square inch 

pressure (or equivalent), as an alternative to (a) for helicopters operating at 

60 miles per hour or less; or 

(c) Other nozzles for aircraft use, as authorized by the director of the 

California Department of Pesticide Regulation, after evaluation.
313

Aerial applications of a pesticide spray solution shall meet the following requirements: 

1. Apply only when there is a positive air flow. Wind speed shall not be more 

than ten miles per hour at the application site, as measured by an anemometer 

positioned four feet above the ground. 

2. Discharge shall start after entering the target site; discharge height shall not 

exceed ten feet above the crop or target; discharge shall be shut off whenever 

necessary to raise the equipment over obstacles; discharge shall be shut off before 

exiting the target site. 

Vehicle-mounted or towed ground equipment, other than handguns, used to make 

applications shall be equipped with: 

1. Nozzles having an orifice not less than one-sixteenth of an inch in diameter (or 

equivalent) and operated at a boom pressure not to exceed the manufacturer’s 

recommended pressure for the nozzles being used; or 

2. Low-pressure fan nozzles with a fan angle number not larger than 80 degrees 

and nozzle orifice not less than 0.2 gallon per minute flow rate (or equivalent) and 

operated at a boom pressure not to exceed 15 pounds per square inch. 

Applications of a pesticide spray solution made by vehicle-mounted or towed ground 

equipment shall meet the following requirements: 

1. Apply only when wind speed is ten miles per hour or less at the application site, 

as measured by an anemometer positioned four feet above the ground. 

2. Discharge shall start after entering the target site; discharge shall be shut off 

before exiting the target site.

The use of Bolero 10G formulation is prohibited in the Sacramento Valley rice growing

counties of Butte, Colusa, Glenn, Placer, Sacramento, Sutter, Tehama, Yolo, and Yuba.

In Butte, Colusa, Glenn, Placer, Sacramento, Sutter, Tehama, Yolo, and Yuba Counties, 

no aerial applications shall be made or continued within 1⁄2 mile of the Sacramento or

Feather Rivers unless there is a continuous positive airflow away from the river.

In Butte, Colusa, Glenn, Placer, Sacramento, Sutter, Tehama, Yolo, and Yuba Counties, 

no aerial application shall be made or continued within 1⁄2 mile of the Sacramento or 

Feather Rivers when the wind speed exceeds seven miles per hour.

In Sacramento and Yolo Counties, no aerial applications shall be made or continued 

within 1⁄4 mile of the Sacramento River unless they are made under the direct supervision 

of the appropriate California county agricultural commissioner or representative.

In Sacramento and Yolo Counties, the maximum acres treated by air each day within 1⁄4

mile of the Sacramento River shall not exceed 33 percent of the average acres treated per

day by air within this area in each county during 2002.

General Water-Holding:
314

Do not release water from the treated field during the water-holding period (Error! 

Reference source not found.).

Prevent seepage27from moving offsite during the water-holding period. 

1. Thiobencarb shall not be applied to rice fields exhibiting visible water seepage that 

moves offsite into drains that are considered state waters.

2. Borders surrounding each rice field shall be compacted before water is allowed to 

fill the field; the degree of compaction shall be sufficient to prevent water from 

seeping through the border. For example, compaction may be achieved by driving 

the tires or tracks of a tractor, or other heavy vehicle, on one side of the border.

3. This requirement (2) applies to new or reworked existing borders for the current 

rice season.

4. A common border between two existing rice fields does not need to be compacted.

 

27 Seepage is lateral movement of irrigation water through a rice field levee or border to 

an area outside the normally flooded production area. Seepage can occur through levees 

into adjacent dry fields or into adjacent drains and canals.
315

Rice pesticides water management requirments summary. Rice Pesticides Water Management 

Requirements Summary

Water must be held for the 

indicated number of 24-hour 

periods on site or containment 

before release to State Waters.

Granular 

(e.g., Bolero 15G 

and UltraMax)

Liquid

(e.g., Abolish 8EC)

Hold (days after 

application)

Hold (days after 

application)

Single Field(d) 30 19

Single field Southern areas only 

(a).

19

Single permitted release into 

tailwater recovery system or 

pond onto fallow field [Except 

Southern area (a)].

14(b) 14(b)

Multi-growers & district release 

onto closed recirculating 

systems.

6 6

Multi-growers & district release 

onto closed recirculating systems 

in Southern area(a).

6

Release from closed recirculating 

system. 19 19

Release into area that discharge 

negligible amount into perennial 

streams

19 6(c)

Emergency Release of tailwater 19 19

Commissioner verifies the 

hydrologic isolation of the fields 6 6

a – Sacramento/San Joaquin Valley defined as: South of the line defined by Roads E10 and 116 in Yolo County and 

the American River in Sacramento County.

b – Thiobencarb permit condition allows Bolereo 15G and Bolero UltraMax label hold periodd of 14 days.

c – See hydrologic isolation fields.

d – When drainage begins after 30 day hold, discharge must not exceed two inches of water over a drain box weir 

for seven additional days. Unregulated discharges from these fields may then begin after 37 days.

On rice fields treated with thiobencarb in the Sacramento Valley (north of the line 

defined by Roads E10 and 116 in Yolo County and the American River in Sacramento 

County), except those treated with liquid formulations e.g., Abolish® 8EC, all water 

must be retained on the treated fields for at least 30 days following application, except as 

described below. When drainage begins, discharge must not exceed two inches of water 

over a drain box weir for seven additional days. Unregulated discharges from these fields 

may then begin after 37 days.

When water is contained within a tailwater recovery system, ponded on 

fallow land, or contained in other systems appropriate for preventing 

discharge, the water must be retained in the system for 19 days, unless:
316

(a) The system is under the control of one permittee, then water may be 

discharged from the application after a 14-day water hold period.

(b) The system includes drainage from more than one permittee, then 

water must be retained on the site of application for six days before being 

discharged from the application site into the system.

(c) Water is on fields within the bounds of areas that discharge negligible 

amounts of rice field drainage into perennial streams until fields are 

drained for harvest.

(d) Water-hold may be reduced to six days if the appropriate California 

county agricultural commissioner evaluates such sites and verifies the 

hydrologic isolation of the fields.

On rice fields treated with thiobencarb in the Sacramento/San Joaquin Valley (south of 

the line defined by Roads E10 and 116 in Yolo County and the American River in 

Sacramento County), except those treated with liquid formulations e.g., Abolish® 8EC,

all water must be retained on the treated fields for at least 19 days following application, 

except as described below. When drainage begins, water discharge must not exceed two 

inches of water over a drain box weir for an additional seven days. Unregulated 

discharges from these fields may begin after 26 days.

When water is contained within a tailwater recovery system, ponded on 

fallow land, or contained in other systems appropriate for preventing 

discharge, the system may discharge 19 days following the last application 

of thiobencarb within the system unless:

(a) The system is under the control of one permittee, then water may be 

discharged from the application after a 14-day water-hold period.

(b) The system includes drainage from more than one permittee, then 

water must be retained on the site of application for six days before 

discharged from the application site into the system.

(c) Water is on fields within the bounds of areas that discharge negligible 

amounts of rice field drainage into perennial streams until fields are 

drained for harvest. Water-hold may be reduced to six days, if the 

appropriate California county agricultural commissioner evaluates such 

sites and verifies the hydrologic isolation of the fields.

On all areas and fields treated with liquid formulations e.g., Abolish® 8EC, all water 

must be retained on the treated fields for at least 19 days following application, except as 

described below. When drainage begins, water discharge must be released at a volume 

not to exceed two inches of water over a drain box weir for an additional seven days. 

Unregulated discharges from these fields may begin after 26 days.

For water contained within a tailwater recovery system, ponded on fallow 

land, or contained in other systems appropriate for preventing discharge, 

the system may discharge 19 days following the last application within the 

system unless:
317

(a) The system is under the control of one permittee, then water may be 

discharged from the application after a 14-day water-hold period.

(b) The system includes drainage from more than one permittee, then 

water must be retained on the site of application for six days before 

discharged from the application site into the system.

(c) Water is on fields within the bounds of areas that discharge negligible 

amounts of rice field drainage into perennial streams until fields are 

drained for harvest, then water-hold may be reduced to six days if the 

appropriate California county agricultural commissioner evaluates such 

sites and verifies the hydrologic isolation of the fields.

Training:

All thiobencarb applicators shall receive mandatory thiobencarb stewardship training.

This requirement is satisfied by:

1. Attending the Preseason Thiobencarb Stewardship Meeting, or receiving 

certification from the County Agricultural Commissioner after viewing a video 

of the Preseason Thiobencarb Stewardship Meeting. 

2. Inspection of thiobencarb aerial applications by California Department of 

Pesticide Regulation to monitor thiobencarb drift mitigation requirements.

2. 2.1 EPA shall include the following instructions requiring reporting of fish kills either 

on the labels or ESPP Bulletins:

NOTICE: Incidents where salmon appear injured or killed as a result of pesticide 

applications shall be reported to NMFS OPR at 301-427-8400 and to the National Pesticide 

Information Center (NPIC) at 1-800-858-7378. The finder should leave the fish alone, make 

note of any circumstances likely causing the death or injury, location and number of fish 

involved, and take photographs, if possible. Adult fish should generally not be disturbed 

unless circumstances arise where an adult fish is obviously injured or killed by pesticide 

exposure, or some unnatural cause. The finder may be asked to carry out instructions 

provided by NMFS OPR to collect specimens or take other measures to ensure that evidence 

intrinsic to the specimen is preserved.

2.2 EPA shall report to NMFS OPR annually any incidences regarding thiobencarb

effects on aquatic ecosystems added to its incident database that EPA has 

classified as “probable” or “highly probable.”
318

14.4 Conservation Recommendations

Section 7(a) (1) of the ESA directs federal agencies to use their authorities to further the 

purposes of the ESA by carrying out conservation programs for the benefit of endangered and 

threatened species. Conservation recommendations are discretionary agency activities to 

minimize or avoid adverse effects of a proposed action on listed species or critical habitat, to 

help implement recovery plans, or to develop information.

The following conservation recommendations would provide information for future 

consultations involving future authorizations of other pesticide a.i.s that may affect listed 

species: 

1. Conduct mixture toxicity analysis in screening-level and endangered species biological 

evaluations; 

2. Develop models to estimate pesticide concentrations in flood plain habitats; and

3. Collaborate with CDPR to review applicability of use limitations identified in the PRESCRIBE 

database to other pesticides and other listed species. 

4. Develop inspection program in coordination with USDA’ NRCS (or other entities) to evaluate 

riparian system condition and function. .This program could be used to inform future section 7 

pesticide consultations and could support structuring less restrictive reasonable and prudent 

alternatives and reasonable and prudent measures. The program ultimately could provide 

incentives to land owners for creating and maintaining riparian systems in listed species habitat.

In order for NMFS to be kept informed of actions minimizing or avoiding adverse effects or 

benefiting listed species or their habitats, the EPA should notify NMFS OPR of any conservation 

recommendations it implements in the final action. 

14.5 Reinitiation Notice

This concludes formal consultation on the EPA’s proposed registration of pesticide products 

containing thiobencarb and their formulations to ESA-listed Pacific salmonids under the 

jurisdiction of the NMFS. As provided in 50 CFR 402.16, reinitiation of formal consultation is 

required where discretionary federal agency involvement or control over the action has been 

retained (or is authorized by law) and if: (1) the extent of take specified in the Incidental Take 
319

Statement is exceeded; (2) new information reveals effects of this action that may affect listed 

species or designated critical habitat in a manner or to an extent not previously considered in this 

biological opinion; (3) the identified action is subsequently modified in a manner that causes an 

effect to the listed species or critical habitat that was not considered in this Opinion; or (4) a new 

species is listed or critical habitat designated that may be affected by the identified action. 

For this consultation, in addition to reinitiation for authorization by EPA for uses other than rice 

growing in these four states, reinitiation will also be required for authorization for use of 

thiobencarb by state or local governments in Washington, Oregon, or Idaho, or expansion of 

rice-growing in California beyond the counties in which rice is now grown. Should reinitiation 

occur because of a change in one state or one area only, the reiniation may review the impacts 

only in that geographic area. If reinitiation of consultation appears warranted due to one or 

more of the above circumstances, EPA must contact NMFS OPR. If none of these reinitiation 

triggers are met within the next 15 years, then reinitiation will be required because the Opinion 

only covers the action for 15 years.
320

15. References

Allan, J. D. and Castillo, M. M. 2008. Stream Ecology: Structure and Function of Running 

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16. Appendix 1 Abbreviations / Acronyms
351

7-DADMax 7-day average of the daily maximum

ACA Alternative Conservation Agreement

AChE acetylcholinesterase

a.i. active ingredient

APEs alkylphenol ethoxylates

APHIS U.S. Department of Agriculture Animal Plant and Health Inspection Service 

BE Biological Evaluation

BEAD Biological and Economic Analysis Division

BLM Bureau of Land Management

BMP Best Management Practices

BOR Bureau of Reclamation

BOR Bureau of Reclamation

BPA Bonneville Power Administration

BRT Biological Review Team (NOAA Fisheries)

BY Brood Years

CAISMP California Aquatic Invasive Species Management Plan

CALFED CALFED Bay-Delta Program (California Resource Agency)

CBFWA Columbia Basin Fish and Wildlife Authority

CBI Confidential Business Information

CC California Coastal

CCC Central California Coast

CCV Central California Valley

CDPR California Department of Pesticide Regulation

CHART Critical Habitat Assessment Review Team

CIDMP Comprehensive Irrigation District Management Plan

CFR Code of Federal Regulations

cfs cubic feet per second

CDFG California Department of Fish and Game

Corps U.S. Department of the Army Corps of Engineers 
352

CSOs combined sewer/stormwater overflows

CSWP California State Water Project

CURES Coalition for Urban/Rural Environmental Stewardship

CVP Central Valley Projects

CVRWQCB Central Valley Regional Water Quality Control Board

CWA Clean Water Act

d day

DCI Date Call-Ins

DDD Dichloro Diphenyl Dichloroethane

DDE Diphenyl Dichlorethylene

DDT Dichloro Diphenyl Trichloroethane

DER Data Evaluation Review

DEQ Oregon Department of Environmental Quality

DIP Demographically Independent Population

DOE Washington State Department of Ecology

DPS Distinct Population Segment

EC Emulsifiable Concentrate Pesticide Formulation

EC50 Median Effect Concentration

EEC Estimated Environmental Concentration

EFED Environmental Fate and Effects Division

EIM Environmental Information Management

EPA U.S. Environmental Protection Agency

ESPP Endangered Species Protection Program

ESA Endangered Species Act

ESU Evolutionarily Significant Unit 

EU European Union

EXAMS Tier II Surface Water Computer Model

FERC Federal Energy Regulatory Commission

FCRPS Federal Columbia River Power System

FFDCA Federal Food and Drug Cosmetic Act

FIFRA Federal Insecticide, Fungicide, and Rodenticide Act
353

FQPA Food Quality Protection Act

ft feet

GENEEC Generic Estimated Exposure Concentration

h hour

HCP Habitat Conservation Plan

HSRG Hatchery Scientific Review Group 

HUC Hydrological Unit Code

IBI Indices of Biological Integrity

ICTRT Interior Columbia Technical Recovery Team

ILWP Irrigated Lands Waiver Program

IPCC Intergovernmental Panel on Climate Change

IRED Interim Re-registration Decision

LCFRB Lower Columbia Fish Recovery Board

ISG Independent Science Group

ITS Incidental Take Statement

km kilometer

Lbs Pounds

LC50 Median Lethal Concentration.

LCR Lower Columbia River 

LOAEC Lowest Observed Adverse Effect Concentration.

LOEL Lowest Observed Adverse Effect level

LOC Level of Concern

LOEC Lowest Observed Effect Concentration

LOQ Limit of Quantification

LWD Large Woody Debris

m meter

MCR Middle Columbia River

mg/L milligrams per liter

MOA Memorandum of Agreement

MPG Major Population Group

MRID Master Record Identification Number
354

MTBE Methyl tert-butyl ether

NASA National Aeronautics and Space Administration

NAWQA U.S. Geological Survey National Water-Quality Assessment

NC Northern California

NEPA National Environmental Protection Agency

NLCD Natural Land Cover Data

NP Nonylphenol

NPDES National Pollutant Discharge Elimination System

NPS National Parks Services

NRCS Natural Resources Conservation Service

NWS National Weather Service

NEPA National Environmental Policy Act 

NMA National Mining Association

NMC N-methyl carbamates

NMFS National Marine Fisheries Service

NOAA National Oceanic and Atmospheric Administration

NOAEC No Observed Adverse Effect Concentration

NPDES National Pollution Discharge Eliminating System

NPIRS National Pesticide Information Retrieval System

NRC National Research Council

OC Oregon Coast

ODFW Oregon Division of Fish and Wildlife

OP Organophosphates

Opinion Biological Opinion

OPP EPA Office of Pesticide Program

PAH polyaromatic hydrocarbons

PBDEs polybrominated diphenyl ethers

PCBs polychlorinated biphenyls

PCEs primary constituent elements

POP Persistent Organic Pollutants

ppb Parts Per Billion
355

PPE Personal Protection Equipment

PSP Pesticide Stewardship Partnerships

PSAMP Puget Sound Assessment and Monitoring Program

PSAT Puget Sound Action Team

PRIA Pesticide Registration Improvement Act

PRZM Pesticide Root Zone Model

PUR Pesticide Use Reporting

QA/QC Quality Assurance/Quality Control

RED Re-registration Eligibility Decision

REI Restricted Entry Interval

RPA Reasonable and Prudent Alternatives

RPM reasonable and prudent measures

RQ Risk Quotient

SAP Scientific Advisory Panel

SAR smolt-to-adult return rate

SASSI Salmon and Steelhead Stock Inventory

SC Southern California

S-CCC South-Central California Coast

SONCC Southern Oregon Northern California Coast

SLN Special Local Need (Registrations under Section 24(c) of FIFRA)

SR Snake River

TCE Trichloroethylene

TCP 3,5,6-trichloro-2-pyridinal

TGAI Technical Grade Active Ingredient

TIE Toxicity Identification Evaluation

TMDL Total Maximum Daily Load

TRT Technical Recovery Team

UCR Upper Columbia River

USFS United States Forest Service

USC United States Code

USFWS United States Fish and Wildlife Service
356

USGS United States Geological Survey

UWR Upper Willamette River

VOC Volatile Organic Compounds

VSP Viable Salmonid Population

WDFW Washington Department of Fish and Wildlife

WLCRTRT Willamette/Lower Columbia River Technical Review Team

WQS Water Quality Standards

WWTIT Western Washington Treaty Indian Tribes

WWTP Wastewater Treatment Plant

YOY Young of year
357

17. Appendix 2: Glossary
358

303(d) waters Section 303 of the federal Clean Water Act requires states to prepare a list of all 

surface waters in the state for which beneficial uses – such as drinking, recreation, aquatic 

habitat, and industrial use - are impaired by pollutants. These are water quality limited estuaries, 

lakes, and streams that do not meet the state’s surface water quality standards and are not 

expected to improve within the next two years. After water bodies are put on the 303(d) list they 

enter into a Total Maximum Daily Load Clean Up Plan.

Active ingredient The component(s) that kills or otherwise affects the pest. A.i.s are always 

listed on the label (FIFRA 2(a)).

Adulticide A compound that kills the adult life stage of the pest insect.

Anadromous Fish Species that are hatched in freshwater migrate to and mature in salt water 

and return to freshwater to spawn.

Adjuvant A compound that aides the operation or improves the effectiveness of a 

pesticide.

Alevin Life history stage of a salmonid immediately after hatching and before the 

yolk-sac is absorbed. Alevins usually remain buried in the gravel in or 

near the egg nest (redd) until their yolk sac is absorbed when they swim up 

and enter the water column.

Anadromy The life history pattern that features egg incubation and early juvenile 

development in freshwater migration to sea water for adult development, 

and a return to freshwater for spawning.

Assessment Endpoint Explicit expression of the actual ecological value that is to be protected 

(e.g., growth of juvenile salmonids).
359

Bioaccumulation Accumulation through the food chain (i.e., consumption of food, 

water/sediment) or direct water and/or sediment exposure.

Bioconcentration Uptake of a chemical across membranes, generally used in reference to 

waterborne exposures.

Biomagnification Transfer of chemicals via the food chain through two or more trophic 

levels as a result of bioconcentration and bioaccumulation.

Degradates New compounds formed by the transformation of a pesticide by chemical 

or biological reactions. 

Distinct Population A listable entity under the ESA that meets tests of discreteness and

Segment significance according to USFWS and NMFS policy. A population is 

considered distinct (and hence a “species” for purposes of conservation 

under the ESA) if it is discrete from and significant to the remainder of its 

species based on factors such as physical, behavioral, or genetic 

characteristics, it occupies an unusual or unique ecological setting, or its 

loss would represent a significant gap in the species’ range.

Escapement The number of fish that survive to reach the spawning grounds or 

hatcheries. The escapement plus the number of fish removed by harvest 

form the total run size.

Evolutionarily A group of Pacific salmon or steelhead trout that is (1) 

Significant Unit substantially reproductively isolated from other conspecific

units and (2) represent an important component of the evolutionary legacy 

of the species.

Fall Chinook This salmon stock returns from the ocean in late summer and early 
360

Salmon fall to head upriver to its spawning grounds, distinguishing it from other 

stocks which migrate in different seasons.

Fate Dispersal of a material in various environmental compartments (sediment, 

water air, biota) as a result of transport, transformation, and degradation.

Flowable A pesticide formulation that can be mixed with water to form a suspension 

in a spray tank.

Fry Stage in salmonid life history when the juvenile has absorbed its yolk sac 

and leaves the gravel of the redd to swim up into the water column. The 

fry stage follows the alevin stage and in most salmonid species is followed 

by the parr, fingerling, and smolt stages. However, chum salmon 

juveniles share characteristics of both the fry and smolt stages and can 

enter sea water almost immediately after becoming fry. 

Half-pounder A life history trait of steelhead exhibited in the Rogue, Klamath, Mad, and 

Eel Rivers of southern Oregon and northern California. Following 

smoltification, half-pounders spend only 2-4 months in the ocean, then 

return to fresh water. They overwinter in fresh water and emigrate to salt 

water again the following spring. This is often termed a false spawning 

migration, as few half-pounders are sexually mature.

Hatchery Salmon hatcheries use artificial procedures to spawn adults and raise the 

resulting progeny in fresh water for release into the natural environment, 

either directly from the hatchery or by transfer into another area. In some 

cases, fertilized eggs are outplanted (usually in “hatch-boxes”), but it is 

more common to release fry or smolts.

Inert ingredients “an ingredient which is not active” (FIFRA 2(m)). It may be toxic or 

enhance the toxicity of the active ingredient.
361

Iteroparous Capable of spawning more than once before death

Jacks Male salmon that return from the ocean to spawn one or more years before 

full-sized adults return. For coho salmon in California, Oregon, 

Washington, and southern British Columbia, jacks are 2 years old, having 

spent only 6 months in the ocean, in contrast to adults, which are 3 years 

old after spending 1 1⁄2 years in the ocean.

Jills Female salmon that return from the ocean to spawn one or more years 

before full-sized adult returns. For sockeye salmon in Oregon, 

Washington, and southern British Columbia, jills are 3 years old (age 1.1), 

having spent only one winter in the ocean in contrast to more typical 

sockeye salmon that are age 1.2, 1.32.2, or 2.3 on return. 

Kokanee The self-perpetuating, non-anadromous form of O. nerka that occurs in 

balanced sex ration populations and whose parents, for several generations 

back, have spent their whole lives in freshwater.

Lambda Also known as Population growth rate, or the rate at which the abundance 

of fish in a population increases or decreases.

LRL Laboratory Reporting Level (USGS NAWQA data)- Generally equal to 

twice the yearly determined LT-MDL. The LRL controls false negative 

error. The probability of falsely reporting a non-detection for a sample that 

contained an analyte at a concentration equal to or greater that the LRL is 

predicted to be less than or equal to 1 percent.

Major Population A group of salmonid populations that are geographically and

Group (MPG) genetically cohesive. The MPG is a level of organization between 

demographically independent populations and the ESU.
362

Main channel The stream channel that includes the thalweg (longitudinal continuous 

deepest portion of the channel.

Metabolite A transformation product resulting from metabolism.

Mode of Action A series of key processes that begins with the interaction of a pesticide 

with a receptor site and proceeds through operational and anatomical 

changes in an organisms that result in sublethal or lethal effects.

Natural fish A fish that is produced by parents spawning in a stream or lake bed, as 

opposed to a controlled environment such as a hatchery.

Nonylphenols A type of APE and is an example of an adjuvant that may be present as an 

ingredient of a formulated product or added to a tank mix prior to 

application.

Off-channel habitat Water bodies and/or inundated areas that are connected (accessible to 

salmonid juveniles) seasonally or annually to the main channel of a stream 

including but not limited to features such as side channels, alcoves, ox 

bows, ditches, and floodplains.

Parr The stage in anadromous salmonid development between absorption of the 

yolk sac and transformation to smolt before migration seaward.

Persistence The tendency of a compound to remain in its original chemical form in the 

environment.

Pesticide Any substance or mixture of substances intended for preventing, 

destroying, repelling or mitigating any pest.
363

Reasonable and Recommended alternative actins identified during formal

Prudent Alternative consultation that can be implemented in a manner consistent

(RPA) with the scope of the Federal agency’s legal authority an jurisdiction, that 

are economically and technologically feasible, and that the Services 

believes would avoid the likelihood of jeopardizing the continued 

existence of the listed species or the destruction or adverse modification of 

designated critical habitat.

Redd A nest constructed by female salmonids in streambed gravels where eggs 

are deposited and fertilization occurs.

Riparian area Area with distinctive soils and vegetation between a stream or other body 

of water and the adjacent upland. It includes wetlands and those portions 

of flood plains an valley bottoms that support riparian vegetation.

Risk The probability of harm from actual or predicted concentrations of a 

chemical in the aquatic environment – a scientific judgment.

Salmonid Fish of the family Salmonidae, including salmon, trout, chars, grayling, 

and whitefish. In general usage, the term usually refers to salmon, trout, 

and chars.

SASSI A cooperative program by WDFW and WWTIT to inventory and evaluate 

the status of Pacific salmonids in Washington State. The SASSI report is 

a series of publications from this program.

Semelparous The condition in an individual organism of reproducing only once in a 

lifetime.
364

Smolt A juvenile salmon or steelhead migrating to the ocean and undergoing 

physiological changes to adapt from freshwater to a saltwater 

environment.

Sublethal Below the concentration that directly causes death. Exposure to sublethal 

concentrations of a material may produce less obvious effect on behavior, 

biochemical, and/or physiological function of the organism often leading 

to indirect death.

Surfactant A substance that reduces the interfacial or surface tension of a system or a 

surface-active substance.

Synergism A phenomenon in which the toxicity of a mixture of chemicals is greater 

than that which would be expected from a simple summation of the 

toxicities of the individual chemicals present in the mixture.

Technical Grade Pure or almost pure active ingredient. Available to formulators.

Active Ingredient Most toxicology data are developed with the TGAI. The percent 

(TGAI) AI is listed on all labels.

Technical Recovery Teams convened by NOAA Fisheries to develop technical products

Teams (TRT) related to recovery planning. TRTs are complemented by planning forums 

unique to specific states, tribes, or reigns, which use TRT and other 

technical products to identify recovery actions.

Teratogenic Effects produced during gestation that evidence themselves as altered 

structural or functional processes in offspring.

Total Maximum defines how much of a pollutant a water body can tolerate (absorb)
365

Daily Load (TMDL) daily and remain compliant with applicable water quality standards. All 

pollutant sources in the watershed combined, including non-point sources, 

are limited to discharging no more than the TMDL.

Viable Salmonid An independent population of Pacific salmon or steelhead trout

Population that has a negligible risk of extinction over a 100-year time frame. 

Viability at the independent population scale is evaluated based on the 

parameters of abundance, productivity, spatial structure, and diversity.

VSP Parameters Abundance, productivity, spatial structure, and diversity. These describe 

characteristics of salmonid populations that are useful in evaluating 

population viability. See NOAA Technical Memorandum NMFSNWFSC-, “Viable salmonid populations and the recovery of 

evolutionarily significant units,” McElhany et al., June 2000.

WDFW Washington Department of Fish and Wildlife is a co-manager of 

salmonids and salmonid fisheries in Washington State with WWTIT and 

other fisheries groups. The agency was formed in the early 1990s by the 

combination of the Washington Department of Fisheries and the 

Washington Department of Wildlife.

WWTIT Western Washington Treaty Indian Tribes is an organization of Native 

American tribes with treaty fishing rights recognized by the U.S. 

government. WWTIT is a co-manager of salmonids and salmonid 

fisheries in western Washington in cooperation with the WDFW and other 

fisheries groups.

WQS “A water quality standard defines the water quality goals of a waterbody, 

or portion thereof, by designating the use or uses to be made of the water 

and by setting criteria necessary to protect public health or welfare, 

enhance the quality of water and serve the purposes of the Clean Water
366

18. Appendix 3: Median and 95th percentile rate of thiobencarb 

application to rice in California (1980-2010)
367

Median and 95th percentile rate of thiobencarb application to rice in California (1980 - 2010). 

Source: California Department of Pesticide Regulation, Pesticide Use Report (PUR) Database.

Year Median Rate

(lbs a.i./A)

95th percentile Rate

(lbs a.i./A)

1980 4.00 4.03

1981 4.00 4.12

1982 4.00 4.08

1983 4.00 4.11

1984 4.00 4.08

1985 4.00 4.07

1986 4.00 4.08

1987 4.00 4.07

1988 4.00 4.08

1989 4.00 4.06

1990 4.00 4.07

1991 4.00 4.06

1992 4.00 4.05

1993 4.00 4.06

1994 4.00 4.03

1995 4.00 4.03

1996 4.00 4.02

1997 4.00 4.01

1998 4.00 4.01

1999 4.00 4.02

2000 4.00 4.02

2001 4.00 4.01

2002 4.00 4.03

2003 4.01 4.05

2004 4.01 4.05

2005 4.01 4.03

2006 4.01 4.02

2007 4.01 4.01

2008 3.99 4.01

2009 3.95 4.01

2010 3.50 3.95
368

19. Appendix 4: Co-Application of Thiobencarb with Other Pesticides in 

California (1999-2010) 
369

Co-application of thiobencarb with other pesticides in California.

This following tables summarize instances of co-application of thiobencarb with other active 

ingredients on rice in California from 1999 to 2010. These can include either tank mixture 

applications, or separate applications of different pesticides to the same field, on the same day. 

Either situation suggest the likely presence of environmental mixtures of pesticides. The 

statistics were provided by Larry Wilhoit, California Department of Pesticide Regulation 

(CDPR), and obtained from the CDPR Pesticide Use Report (PUR) Database. To obtain the 

information from the database, a co-application was defined as the use of more than one active 

ingredient (AI) on an agricultural field in the same day. In the PUR a field is identified by the 

combination of grower_id and site_loc_id after some cleaning of the data. The site_loc_id is 

cleaned by removing all spaces and assuming that site_loc_ids that are the same except for 

characters that could be confused (such as O and 0 and 2 and Z) and have the same MTRS refer 

to the same field. Each field is given a unique field_id. There are a few records that were not 

included in the summary information because of errors or missing data for the site_loc_id, 

application date, or some other field which made it impossible to determine the specific date and 

field.

Summary indicating yearly frequency of tank mixture applications of thiobencarb in California

YEAR

NUMBER OF 

THIOBENCARB 

APPLICATIONS 

WITH 

FIELD_ID

NUMBER OF 

COAPPLICATIONSPERCENT COAPPLICATIONSLBS 

THIOBENCARB

LBS 

OTHER 

AIS

1999 2,189 490 22 731,717 158,922

2000 2,989 620 21 1,006,965 267,879

2001 2,070 435 21 645,914 208,611

2002 2,656 582 22 843,773 274,187

2003 1,816 439 24 587,156 227,529

2004 1,476 359 24 521,556 157,158

2005 1,359 359 26 448,182 122,010

2006 908 180 20 310,346 73,926

2007 794 230 29 289,032 58,066

2008 820 155 19 263,499 37,479

2009 822 162 20 272,080 46,770

2010 834 237 28 258,402 16,467

Total 18,733 4,248 23 6,178,623 1,649,004
370

Summary of tank mixture application that included thiobencarb (California, 1999-2010)

LIST OF ACTIVE INGREDIENTS APPLIED WITH 

THIOBENCARB

NUMBER 

APPLICATIONS

LBS 

THIOBENCARB

LBS 

OTHER 

AIs

THIOBENCARB ALONE 14,474 4,828,157

COPPER SULFATE (PENTAHYDRATE) 1,128 393,884 1,278,042

BENSULFURON METHYL 660 219,889 3,417

LAMBDA-CYHALOTHRIN 470 147,278 483

PENOXSULAM 445 132,366 1,364

BISPYRIBAC-SODIUM 217 57,420 6,842

CARFENTRAZONE-ETHYL 201 78,718 3,943

MOLINATE 166 50,332 38,876

PROPANIL 136 34,605 43,839

PENDIMETHALIN 107 17,659 5,231

(S)-CYPERMETHRIN 66 20,497 82

CARBOFURAN 62 16,510 873

COPPER SULFATE (PENTAHYDRATE); 

BENSULFURON METHYL 43 16,904 53,050

PROPANIL; PENDIMETHALIN 30 5,408 8,048

BENSULFURON METHYL; LAMBDA-CYHALOTHRIN 29 8,013 159

COPPER SULFATE (PENTAHYDRATE); PENOXSULAM 29 11,378 35,647

METHYLATED SOYBEAN OIL; LAMBDACYHALOTHRIN; KEROSENE; DIMETHYL ALKYL 

TERTIARY AMINES; BENZOIC ACID 25 10,279 218

COPPER SULFATE (PENTAHYDRATE); MOLINATE 24 8,073 39,934

POLY-I-PARA-MENTHENE 21 6,278 1,454

PROPANIL; MOLINATE 21 2,016 6,577

MCPA, DIMETHYLAMINE SALT 18 6,186 7,486

COPPER SULFATE (PENTAHYDRATE); LAMBDACYHALOTHRIN 15 5,082 15,966

BENSULFURON METHYL; POLYACRYLAMIDE 

POLYMER 15 2,318 37

KEROSENE; LAMBDA-CYHALOTHRIN; SOYBEAN 

FATTY ACIDS, DIMETHYLAMINE SALT; BENZOIC 

ACID 14 5,030 123

ORTHOSULFAMURON 14 3,534 553

METHYLATED SOYBEAN OIL; (S)-CYPERMETHRIN; 

KEROSENE; DIMETHYL ALKYL TERTIARY AMINES; 

BENZOIC ACID 13 5,082 124

METHYLATED SOYBEAN OIL; BENSULFURON 

METHYL; KEROSENE; DIMETHYL ALKYL TERTIARY 

AMINES; BENZOIC ACID 12 4,935 315
371

LIST OF ACTIVE INGREDIENTS APPLIED WITH 

THIOBENCARB

NUMBER 

APPLICATIONS

LBS 

THIOBENCARB

LBS 

OTHER 

AIs

COPPER SULFATE (PENTAHYDRATE); 

CARFENTRAZONE-ETHYL 12 3,673 11,326

METHYLATED SOYBEAN OIL; KEROSENE; DIMETHYL 

ALKYL TERTIARY AMINES; BENZOIC ACID 11 3,292 325

MOLINATE; BENSULFURON METHYL 11 2,198 3,275

PENOXSULAM; (S)-CYPERMETHRIN 9 2,343 37

AZOXYSTROBIN; OLEIC ACID, METHYL ESTER; 2-(3-

HYDROXYPROPYL)-HEPTA-METHYL TRISILOXANE, 

ETHOXYLATED, ACETATE; BENSULFURON METHYL; 

BISPYRIBAC-SODIUM 8 555 84

CARFENTRAZONE-ETHYL; BENSULFURON METHYL 7 1,591 86

PROPANIL; LAMBDA-CYHALOTHRIN 7 2,388 3,543

PROPANIL; MOLINATE; CARBOFURAN 6 281 1,974

PENOXSULAM; LAMBDA-CYHALOTHRIN 6 3,128 49

BENSULFURON METHYL; POLYACRYLAMIDE 

POLYMER; POLYSACCHARIDE POLYMER 6 2,904 48

PROPANIL; PETROLEUM OIL, PARAFFIN BASED; 

POLYOXYETHYLENE SORBITAN MIXED FATTY ACID 

ESTERS; SORBITAN FATTY ACID ESTERS; ALPHA-

(PARA-NONYLPHENYL)-OMEGA-HYDROXYPOLY 

(OXYETHYLENE) SULFATE, AMMONIUM SALT 5 990 1,547

COPPER SULFATE (PENTAHYDRATE); 

POLYSILOXANE; LAMBDA-CYHALOTHRIN; 

DERIVATED NATURAL POLYMERS 5 298 996

MINERAL OIL; PETROLEUM DISTILLATES; LAMBDACYHALOTHRIN; TALL OIL FATTY ACIDS; DIMETHYL 

SOYA AMINE; BENZOIC ACID 5 2,173 43

SODIUM CARBONATE PEROXYHYDRATE 5 1,782 3,876

GLYPHOSATE, ISOPROPYLAMINE SALT 5 1,192 14

LAMBDA-CYHALOTHRIN; ALPHA-(PARANONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE); PHOSPHORIC 

ACID; ALPHA-ALKYLARYL-OMEGAHYDROXYPOLY(OXYETHYLENE); ISOPROPYL 

ALCOHOL; DIMETHYLPOLYSILOXANE 4 1,449 27

ALKYL (C8,C10) POLYGLYCOSIDE; LAMBDACYHALOTHRIN 4 3,036 26

CARFENTRAZONE-ETHYL; (S)-CYPERMETHRIN 4 1,164 59

PROPANIL; VEGETABLE OIL; PENDIMETHALIN; 3 295 1,040
372

LIST OF ACTIVE INGREDIENTS APPLIED WITH 

THIOBENCARB

NUMBER 

APPLICATIONS

LBS 

THIOBENCARB

LBS 

OTHER 

AIs

FATTY ACIDS, MIXED

DIFLUBENZURON; (S)-CYPERMETHRIN; 

POLYSILOXANE; DERIVATED NATURAL POLYMERS 3 1,105 22

PROPANIL; CYHALOFOP-BUTYL; TRICLOPYR, 

TRIETHYLAMINE SALT; BISPYRIBAC-SODIUM 3 1 3

PROPANIL; PETROLEUM DISTILLATES, ALIPHATIC; 

PENDIMETHALIN; CYHALOFOP-BUTYL 3 1,486 2,552

(S)-CYPERMETHRIN; PENOXSULAM 3 373 11

COPPER SULFATE (PENTAHYDRATE); PROPANIL 3 35 515

VINYL POLYMER 3 846 2

CARFENTRAZONE-ETHYL; BENSULFURON METHYL; 

LAMBDA-CYHALOTHRIN 3 720 38

COPPER SULFATE (PENTAHYDRATE); 

PENOXSULAM; LAMBDA-CYHALOTHRIN 3 1,134 5,107

CYHALOFOP-BUTYL 3 1,076 70

KEROSENE; GLYPHOSATE, ISOPROPYLAMINE SALT; 

SOYBEAN FATTY ACIDS, DIMETHYLAMINE SALT; 

BENZOIC ACID 3 621 349

LAMBDA-CYHALOTHRIN; POLYACRYLAMIDE 

POLYMER; POLYSACCHARIDE POLYMER 3 1,040 2

PENDIMETHALIN; ALPHA-PINENE BETA-PINENE 

COPOLYMER; MINERAL OIL; N,N-BIS-(2-OMEGAHYDROXYPOLY(OXYETHYLENE)ETHYL)ALKYLAMINE, 

ALKYL DERIVED FROM TALLOW FATTY ACIDS; TALL 

OIL FATTY ACIDS 3 1,352 533

PROPANIL; PENDIMETHALIN; POLY-I-PARAMENTHENE 3 633 1,278

CLOMAZONE 3 380 29

COPPER SULFATE (PENTAHYDRATE); 

BENSULFURON METHYL; POLYACRYLAMIDE 

POLYMER; POLYSACCHARIDE POLYMER 2 960 1,720

METHYLATED SOYBEAN OIL; BISPYRIBAC-SODIUM; 

KEROSENE; DIMETHYL ALKYL TERTIARY AMINES; 

BENZOIC ACID 2 788 34

OLEIC ACID, METHYL ESTER; BENSULFURON 

METHYL; 2-(3-HYDROXYPROPYL)-HEPTA-METHYL 

TRISILOXANE, ETHOXYLATED, ACETATE; 

BISPYRIBAC-SODIUM 2 89 8
373

LIST OF ACTIVE INGREDIENTS APPLIED WITH 

THIOBENCARB

NUMBER 

APPLICATIONS

LBS 

THIOBENCARB

LBS 

OTHER 

AIs

METHYLATED SOYBEAN OIL; BISPYRIBAC-SODIUM; 

ALPHA-(PARA-NONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE); POLYALKENE 

OXIDE MODIFIED HEPTAMETHYL TRISILOXANE 2 237 31

PENDIMETHALIN; BISPYRIBAC-SODIUM 2 306 91

PROPANIL; BISPYRIBAC-SODIUM 2 496 497

DIFLUBENZURON; LAMBDA-CYHALOTHRIN 2 500 5

MINERAL OIL; (S)-CYPERMETHRIN; PETROLEUM 

DISTILLATES; TALL OIL FATTY ACIDS; DIMETHYL 

SOYA AMINE; BENZOIC ACID 2 752 17

DIFLUBENZURON 2 525 6

LAMBDA-CYHALOTHRIN; VINYL POLYMER 2 1,324 4

BENSULFURON METHYL; METHYLATED SOYBEAN 

OIL; LAMBDA-CYHALOTHRIN; KEROSENE; 

DIMETHYL ALKYL TERTIARY AMINES; BENZOIC 

ACID 2 620 16

METHYLATED SOYBEAN OIL; PENOXSULAM; (S)-

CYPERMETHRIN; KEROSENE; DIMETHYL ALKYL 

TERTIARY AMINES; BENZOIC ACID 2 630 27

BENSULFURON METHYL; METHYLATED SOYBEAN 

OIL; ALPHA-(PARA-NONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE); LAMBDACYHALOTHRIN; POLYALKENE OXIDE MODIFIED 

HEPTAMETHYL TRISILOXANE 2 420 9

MOLINATE; LAMBDA-CYHALOTHRIN 2 590 257

VINYL POLYMER; BISPYRIBAC-SODIUM 2 120 9

PROPANIL; POLY-I-PARA-MENTHENE 2 389 868

CARBOFURAN; VINYL POLYMER 2 633 42

PENDIMETHALIN; BENSULFURON METHYL 2 415 106

CARFENTRAZONE-ETHYL; LAMBDA-CYHALOTHRIN 2 643 28

PENDIMETHALIN; VINYL POLYMER 2 506 164

PROPANIL; ORCHEX 796 OIL; OLEIC ACID; ALPHAALKYL (C9-C11)-OMEGAHYDROXYPOLY(OXYETHYLENE) 2 395 321

BISPYRIBAC-SODIUM; VINYL POLYMER 1 419 28,044

MOLINATE; PROPANIL; PETROLEUM OIL, PARAFFIN 

BASED; TRICLOPYR, TRIETHYLAMINE SALT; 

POLYOXYETHYLENE SORBITAN MIXED FATTY ACID 

ESTERS; SORBITAN FATTY ACID ESTERS; ALPHA- 1 176 98
374

LIST OF ACTIVE INGREDIENTS APPLIED WITH 

THIOBENCARB

NUMBER 

APPLICATIONS

LBS 

THIOBENCARB

LBS 

OTHER 

AIs

(PARA-NONYLPHENYL)-OMEGA-HYDROXYPOLY 

(OXYETHYLENE) SULFATE, AMMONIUM SALT

PROPANIL; MOLINATE; ISOPARAFFINIC 

HYDROCARBONS; ORCHEX 796 OIL; ALPHA-(PARANONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE); TALL OIL FATTY 

ACIDS; ALPHA-(PARA-NONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE), PHOSPHATE 

ESTER; N,N-BIS-(2-OMEGAHYDROXYPOLY(OXYETHYLENE)ETHYL)ALKYLAMINE, 

ALKYL DERIVED FROM TALLOW FATTY ACIDS 1 239 895

PETROLEUM OIL, PARAFFIN BASED; 

POLYOXYETHYLENE SORBITAN MIXED FATTY ACID 

ESTERS; BENSULFURON METHYL; SORBITAN FATTY 

ACID ESTERS; ALPHA-(PARA-NONYLPHENYL)-

OMEGA-HYDROXYPOLY (OXYETHYLENE) SULFATE, 

AMMONIUM SALT 1 260 113

PROPANIL; HYDROTREATED PARAFFINIC SOLVENT; 

TALL OIL FATTY ACIDS; ALPHA-(PARANONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE) 1 119 181

METHYLATED SOYBEAN OIL; KEROSENE; DIMETHYL 

ALKYL TERTIARY AMINES; BENZOIC ACID; LAMBDACYHALOTHRIN 1 278 5

PROPANIL; PHOSPHORIC ACID; ZINC SULFATE; 

MANGANESE SULFATE; FERROUS SULFATE; 

COCONUT IMIDAZOLINE SODIUM CARBOXYLATE 1 356 553

PROPANIL; MOLINATE; PETROLEUM OIL, PARAFFIN 

BASED; TRICLOPYR, TRIETHYLAMINE SALT; 

POLYOXYETHYLENE SORBITAN MIXED FATTY ACID 

ESTERS; SORBITAN FATTY ACID ESTERS; ALPHA-

(PARA-NONYLPHENYL)-OMEGA-HYDROXYPOLY 

(OXYETHYLENE) SULFATE, AMMONIUM SALT 1 265 685

PENDIMETHALIN; OLEIC ACID, METHYL ESTER; 

TRICLOPYR, TRIETHYLAMINE SALT; 2-(3-

HYDROXYPROPYL)-HEPTA-METHYL TRISILOXANE, 

ETHOXYLATED, ACETATE; BISPYRIBAC-SODIUM; 

CITRIC ACID; CALCIUM CHLORIDE 1 72 28
375

LIST OF ACTIVE INGREDIENTS APPLIED WITH 

THIOBENCARB

NUMBER 

APPLICATIONS

LBS 

THIOBENCARB

LBS 

OTHER 

AIs

BISPYRIBAC-SODIUM; ALPHA-OCTYLPHENYLOMEGA-HYDROXYPOLY(OXYETHYLENE); 

POLYACRYLAMIDE, POLYETHYLENE GLYCOL 

MIXTURE; PHOSPHORIC ACID 1 119 3

PROPANIL; BENSULFURON METHYL 1 474 605

ORTHOSULFAMURON; (S)-CYPERMETHRIN 1 390 623

COPPER SULFATE (PENTAHYDRATE); MOLINATE; 

BENSULFURON METHYL 1 400 2,882

PROPANIL; PENDIMETHALIN; CYHALOFOP-BUTYL; 

BISPYRIBAC-SODIUM; BENSULFURON METHYL 1 0 1

KEROSENE; BENSULFURON METHYL; LAMBDACYHALOTHRIN; SOYBEAN FATTY ACIDS, 

DIMETHYLAMINE SALT; BENZOIC ACID 1 290 12

COPPER SULFATE (PENTAHYDRATE); METHYLATED 

SOYBEAN OIL; DIFLUBENZURON; KEROSENE; 

DIMETHYL ALKYL TERTIARY AMINES; BENZOIC 

ACID 1 960 3,612

BENSULFURON METHYL; DIFLUBENZURON 1 808 24

METHYLATED SOYBEAN OIL; PENOXSULAM; 

LAMBDA-CYHALOTHRIN; KEROSENE; DIMETHYL 

ALKYL TERTIARY AMINES; BENZOIC ACID 1 428 16

MCPA, DIMETHYLAMINE SALT; BISPYRIBACSODIUM 1 573 73

DIFLUBENZURON; LAMBDA-CYHALOTHRIN; 

POLYSILOXANE; DERIVATED NATURAL POLYMERS 1 320 4

DERIVATED NATURAL POLYMERS; POLYSILOXANE; 

BENSULFURON METHYL; LAMBDA-CYHALOTHRIN 1 290 6

PETROLEUM DISTILLATES, ALIPHATIC; BISPYRIBACSODIUM 1 1,190 521

MINERAL OIL; BENSULFURON METHYL; 

PETROLEUM DISTILLATES; TALL OIL FATTY ACIDS; 

DIMETHYL SOYA AMINE; BENZOIC ACID 1 401 18

COPPER SULFATE (PENTAHYDRATE); 

DIFLUBENZURON; (S)-CYPERMETHRIN 1 405 353

METHYL PARATHION; BENSULFURON METHYL; 

METHYL PARATHION, OTHER RELATED; VINYL 

POLYMER 1 480 84

CITRIC ACID; PENOXSULAM; LAMBDACYHALOTHRIN; CALCIUM CHLORIDE 1 506 22

COPPER SULFATE (PENTAHYDRATE); 1 1,223 1,488
376

LIST OF ACTIVE INGREDIENTS APPLIED WITH 

THIOBENCARB

NUMBER 

APPLICATIONS

LBS 

THIOBENCARB

LBS 

OTHER 

AIs

ORTHOSULFAMURON

HYDROTREATED PARAFFINIC SOLVENT; 

CYHALOFOP-BUTYL; TALL OIL FATTY ACIDS; 

BISPYRIBAC-SODIUM; ALPHA-(PARANONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE); PENOXSULAM 1 119 16

CYHALOFOP-BUTYL; PROPANIL; VEGETABLE OIL; 

PENDIMETHALIN; FATTY ACIDS, MIXED 1 91 255

MINERAL OIL; PENOXSULAM; PETROLEUM 

DISTILLATES; LAMBDA-CYHALOTHRIN; TALL OIL 

FATTY ACIDS; DIMETHYL SOYA AMINE; BENZOIC 

ACID 1 177 7

MOLINATE; DIFLUBENZURON; BENSULFURON 

METHYL; LAMBDA-CYHALOTHRIN 1 390 92

CARFENTRAZONE-ETHYL; PENDIMETHALIN; 

PROPANIL 1 0 1

MOLINATE; PENOXSULAM; CARFENTRAZONEETHYL; BENSULFURON METHYL 1 7 0

BENSULFURON METHYL; (S)-CYPERMETHRIN 1 447 14

DIFLUBENZURON; POLYSILOXANE; (S)-

CYPERMETHRIN; DERIVATED NATURAL POLYMERS 1 476 4

AZOXYSTROBIN; FATTY ACIDS, C16-C18 AND C18-

UNSATURATED, METHYL ESTERS; 4-

NONYLPHENOL, FORMALDEHYDE RESIN, 

PROPOXYLATED; OLEIC ACID, METHYL ESTER; N,NBIS-(2-OMEGAHYDROXYPOLY(OXYETHYLENE)ETHYL)ALKYLAMINE, 

ALKYL DERIVED FROM TALLOW FATTY ACIDS; 

POLYBUTENES; 2-(3-HYDROXYPROPYL)-HEPTAMETHYL TRISILOXANE, ETHOXYLATED, ACETATE; 

BISPYRIBAC-SODIUM 1 24 14

METHYL PARATHION; KEROSENE; SOYBEAN FATTY 

ACIDS, DIMETHYLAMINE SALT; BENZOIC ACID 1 620 103

ALPHA-(PARA-NONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE); PHOSPHORIC 

ACID; ALPHA-ALKYLARYL-OMEGAHYDROXYPOLY(OXYETHYLENE); LAMBDACYHALOTHRIN; ISOPROPYL ALCOHOL; 

DIMETHYLPOLYSILOXANE 1 565 15

PROPANIL; PETROLEUM DISTILLATES, ALIPHATIC; 1 621 889
377

LIST OF ACTIVE INGREDIENTS APPLIED WITH 

THIOBENCARB

NUMBER 

APPLICATIONS

LBS 

THIOBENCARB

LBS 

OTHER 

AIs

PENDIMETHALIN

COPPER SULFATE (PENTAHYDRATE); (S)-

CYPERMETHRIN 1 531 1,970

PROPANIL; MINERAL OIL; TALL OIL FATTY ACIDS; 

ALPHA-(PARA-NONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE) 1 403 430

BISPYRIBAC-SODIUM; ALPHA-2,6,8-TRIMETHYL-4-

NONYLOXY-OMEGAHYDROXYPOLY(OXYETHYLENE); POLYETHYLENE 

GLYCOL MONO(3-(TETRAMETHYL-1-

(TRIMETHYLSILOXY)DISILOXANYL)PROPYL)ETHER; 

PROPYLENE GLYCOL; PHOSPHORIC ACID; 

TRISODIUM PHOSPHATE; 

DIMETHYLPOLYSILOXANE 1 119 5

PROPANIL; ORCHEX 796 OIL; OLEIC ACID; 

AZOXYSTROBIN; ALPHA-ALKYL (C9-C11)-OMEGAHYDROXYPOLY(OXYETHYLENE); BENSULFURON 

METHYL; CARFENTRAZONE-ETHYL 1 375 559

MOLINATE; PROPANIL 1 40 106

MOLINATE; POLY-I-PARA-MENTHENE 1 439 228

CYHALOFOP-BUTYL; VEGETABLE OIL; 

PENDIMETHALIN; FATTY ACIDS, MIXED 1 49 73

BENSULFURON METHYL; DIFLUBENZURON; 

LAMBDA-CYHALOTHRIN 1 465 19

MOLINATE; CARFENTRAZONE-ETHYL 1 600 218

PROPANIL; CYHALOFOP-BUTYL; BISPYRIBACSODIUM; FENOXAPROP-P-ETHYL 1 0 0

PETROLEUM OIL, PARAFFIN BASED; CYHALOFOPBUTYL; POLYOXYETHYLENE SORBITAN MIXED 

FATTY ACID ESTERS; AMMONIUM SULFATE; 

SORBITAN FATTY ACID ESTERS; PHOSPHORIC ACID; 

ALPHA-(PARA-NONYLPHENYL)-OMEGAHYDROXYPOLY (OXYETHYLENE) SULFATE, 

AMMONIUM SALT; CITRIC ACID; POLYACRYLIC 

POLYMER 1 847 326

PROPANIL; CYHALOFOP-BUTYL; PENOXSULAM; 

2,4-D; BISPYRIBAC-SODIUM 1 0 1

PROPANIL; ORCHEX 796 OIL; OLEIC ACID; 2,4-D, 

DIMETHYLAMINE SALT; ALPHA-ALKYL (C9-C11)-

OMEGA-HYDROXYPOLY(OXYETHYLENE); 1 59 677
378

LIST OF ACTIVE INGREDIENTS APPLIED WITH 

THIOBENCARB

NUMBER 

APPLICATIONS

LBS 

THIOBENCARB

LBS 

OTHER 

AIs

BISPYRIBAC-SODIUM

MINERAL OIL; CYHALOFOP-BUTYL; TALL OIL FATTY 

ACIDS; ALPHA-(PARA-NONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE); ALPHA-2,6,8-

TRIMETHYL-4-NONYLOXY-OMEGAHYDROXYPOLY(OXYETHYLENE); POLYETHYLENE 

GLYCOL MONO(3-(TETRAMETHYL-1-

(TRIMETHYLSILOXY)DISILOXANYL)PROPYL)ETHER; 

BISPYRIBAC-SODIUM; PROPYLENE GLYCOL; 

PENOXSULAM; CARFENTRAZONE-ETHYL; 

DIMETHYLPOLYSILOXANE 1 15 55

BENSULFURON METHYL; LAMBDA-CYHALOTHRIN; 

POLYSILOXANE; DERIVATED NATURAL POLYMERS 1 142 3

COPPER SULFATE (PENTAHYDRATE); LAMBDACYHALOTHRIN; POLYACRYLAMIDE POLYMER; 

POLYSACCHARIDE POLYMER 1 308 1,535

POLYACRYLAMIDE POLYMER; POLYSACCHARIDE 

POLYMER 1 119 0

PENDIMETHALIN; OLEIC ACID, METHYL ESTER; 

BISPYRIBAC-SODIUM; 2-(3-HYDROXYPROPYL)-

HEPTA-METHYL TRISILOXANE, ETHOXYLATED, 

ACETATE 1 118 26

PETROLEUM DISTILLATES; ALPHA-(PARANONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE); CYHALOFOPBUTYL 1 237 455

BENSULFURON METHYL; PROPANIL; PETROLEUM 

OIL, PARAFFIN BASED; POLYOXYETHYLENE 

SORBITAN MIXED FATTY ACID ESTERS; SORBITAN 

FATTY ACID ESTERS; ALPHA-(PARANONYLPHENYL)-OMEGA-HYDROXYPOLY 

(OXYETHYLENE) SULFATE, AMMONIUM SALT 1 1 3

LAMBDA-CYHALOTHRIN; POLYACRYLAMIDE 

POLYMER 1 182 2

DIFLUBENZURON; (S)-CYPERMETHRIN 1 320 2

HYDROTREATED PARAFFINIC SOLVENT; 

CYHALOFOP-BUTYL; TALL OIL FATTY ACIDS; ALPHA-

(PARA-NONYLPHENYL)-OMEGAHYDROXYPOLY(OXYETHYLENE); CITRIC ACID; 1 705 384
379

LIST OF ACTIVE INGREDIENTS APPLIED WITH 

THIOBENCARB

NUMBER 

APPLICATIONS

LBS 

THIOBENCARB

LBS 

OTHER 

AIs

CALCIUM CHLORIDE

MOLINATE; METHYLATED SOYBEAN OIL; LAMBDACYHALOTHRIN; KEROSENE; DIMETHYL ALKYL 

TERTIARY AMINES; BENZOIC ACID 1 456 108

COPPER SULFATE (PENTAHYDRATE); 

BENSULFURON METHYL; LAMBDA-CYHALOTHRIN; 

POLYSILOXANE; DERIVATED NATURAL POLYMERS 1 377 751

PROPANIL; ISOPARAFFINIC HYDROCARBONS; 

ORCHEX 796 OIL; ALPHA-(PARA-NONYLPHENYL)-

OMEGA-HYDROXYPOLY(OXYETHYLENE); TALL OIL 

FATTY ACIDS; ALPHA-(PARA-NONYLPHENYL)-

OMEGA-HYDROXYPOLY(OXYETHYLENE), 

PHOSPHATE ESTER; N,N-BIS-(2-OMEGAHYDROXYPOLY(OXYETHYLENE)ETHYL)ALKYLAMINE, 

ALKYL DERIVED FROM TALLOW FATTY ACIDS 1 623 934

OLEIC ACID, METHYL ESTER; 2-(3-

HYDROXYPROPYL)-HEPTA-METHYL TRISILOXANE, 

ETHOXYLATED, ACETATE; BISPYRIBAC-SODIUM 1 0 2

POLYALKENE OXIDE MODIFIED HEPTAMETHYL 

TRISILOXANE; BISPYRIBAC-SODIUM; PROPYLENE 

GLYCOL 1 15 2

COPPER SULFATE (PENTAHYDRATE); KEROSENE; 

LAMBDA-CYHALOTHRIN; SOYBEAN FATTY ACIDS, 

DIMETHYLAMINE SALT; BENZOIC ACID 1 173 647

MOLINATE; PROPANIL; BENSULFURON METHYL 1 1 1

SODIUM CARBONATE PEROXYHYDRATE; 

ORTHOSULFAMURON 1 1,080 1,278

CARFENTRAZONE-ETHYL; PENOXSULAM 1 141 11

TOTAL 18,733 6,178,623 1,649,004
380

20. Appendix 5: Toxicity of Eleven Pesticides to Embryonic Zebrafish
381

Toxicity of Eleven Pesticides to Embryonic Zebrafish

November 2011

Project Summary

The Northwest Fisheries Science Center conducted an experiment requested by NOAA’s 

Office of Protected Resources in support of a Biological Opinion regarding the toxicity of 

various pesticides to endangered salmon species. The experiment detailed here investigated the 

effects of eleven pesticides on developing zebrafish (Danio rerio), a species that is widely used 

as a toxicological model for other fish species. Zebrafish are a useful model species because the 

early ontogeny of zebrafish is rapid and well documented (Kimmel et al., 1995) and their 

features are easily observed through translucent chorions and bodies. In this experiment, 

embryonic zebrafish were exposed to oryzalin, trifluralin, prometryn, pendimethalin, fenbutatin 

oxide, thiobencarb, propargite, metolachlor, 1,3-dichloropropene, bromoxynil and diflubenzuron 

in 5-day static-renewal exposures. Toxicity endpoints included mortality, developmental 

abnormalities, and body length on the final day of the experiment. Three of the chemicals tested, 

prometryn, fenbutatin oxide, and diflubenzuron, did not produce an adverse effect on zebrafish 

survival, morphology or length at the tested concentrations. The pesticides trifluralin, 

pendimethalin and thiobencarb increased the rate of abnormality in developing zebrafish without 

appreciably increasing the rate of mortality at the concentrations tested. Fish lengths were 

significantly smaller following exposure to oryzalin, bromoxynil, trifluralin, pendimethalin, 

thiobencarb, propargite, metolachlor and 1,3-dichloropropene. 

Methods

Fish: Zebrafish (D. rerio) embryos were obtained from a colony maintained at the Northwest 

Fisheries Science Center according to standard operating procedures (Linbo, 2009). Male and 

female zebrafish were combined in spawning tanks and eggs were collected at the beginning of 

the next light cycle, approximately one hour after the spawning event. Embryos were housed in 

a temperature-controlled incubator at 28.5 C for the duration of the experiment. 
382

Pesticide stock solutions: Pesticides were obtained in pure form from Chem Service, Inc. (West 

Chester, Pennsylvania). Pesticide stock solutions were made in acetone and stored under dark 

conditions at 4 C. A working solution composed of stock solution and water from the zebrafish 

colony (system water) was mixed fresh at the start of each day, and subsequent exposure 

concentrations serially diluted. The maximum acetone concentration for any exposure was 

0.1%. The highest pesticide concentration of each compound tested was generally the reported 

rainbow trout or zebrafish 96-hr LC50 value (the concentration lethal to 50% of the test 

organisms). The highest exposure concentration of 1,3-dichloropropene was 100 times lower 

than the reported LC50 value because of observed developmental effects, while exposure 

concentrations of diflubenzuron were lower due to low solubility in acetone. 

Pesticide exposures: Normally developing zebrafish embryos at 1.5-2.5 hpf (hours postfertilization) were selected and placed in 60 mm acetone-washed glass Petri dishes with 10 ml of 

pesticide solution. Individual dishes contained 15 embryos and each exposure concentration was 

tested in triplicate. Exposures were conducted in batches comprised of one or two pesticides, 

water controls, and 0.1% acetone controls. Exposure solutions were renewed every 24 hours. 

Dead embryos were removed from the dishes each day to prevent fungal growth and 

contamination. 

Anatomical screening and measurement of fish body length: Embryos were scored every 24 hr 

for mortality and abnormalities through 5 dpf (days post-fertilization). See Table 2 for a 

description of the observed developmental abnormalities. Daily anatomical screenings were 

performed using a Nikon-SMZ-800 stereomicroscope with a diascopic base (Meridian 

Instruments, Seattle, Washington). Only surviving fish were screened for anatomical 

abnormalities. At 5 dpf, the embryos were anesthetized with tricaine methanesulfonate (MS222; Sigma-Aldrich, St. Louis, Missouri) to measure body length. All surviving embryos from 

each exposure dish were simultaneously photographed using a Spot RT digital camera 

(Diagnostic Instruments, Inc., Sterling Heights, Michigan) mounted on a stereomicroscope. 

Length was measured from the anterior tip of the mouth along the notochord to the posterior tip 
383

of the notochord, and quantified using ImageJ software (available online at 

http://rsbweb.nih.gov/ij/). 

Statistical tests: Length was the only parameter explicitly tested. Lengths of control fish were 

compared using a two-factor ANOVA comparing type (water and acetone) and batch, and 

showed a significant result of batch only. Subsequent analyses of exposures compared the 

average of three dishes (n = 3) to their corresponding batch controls. Differences in embryo 

lengths between concentrations of a given pesticide were tested using one-way ANOVAs with a 

Tukey HSD post hoc (Tables 3-13).

Results

Chemical-specific mortality and abnormality data, as well as their respective controls, are 

presented in Figures 1-11. Both water and acetone controls showed consistently low rates of 

both mortality and abnormality. We found that 3 pesticides (prometryn, fenbutatin oxide and 

diflubenzuron) showed no increases in mortality or abnormality as well as no significant 

differences in embryo length. Three additional chemicals (trifluralin, pendimethalin and 

thiobencarb) produced higher rates of abnormalities and significantly shorter embryos at the 

highest exposure concentration without increasing mortality. While the remaining pesticides 

(oryzalin, bromoxynil, propargite, metolachlor, and 1,3-dichloropropene) produced significantly 

shorter embryos at various exposure concentrations with no effect on mortality or abnormality, 

there was no clear dose-dependent trend. Whether there is a biological consequence to these 

shorter lengths at the concentrations tested here is a subject for further investigation. 
384

Table 1. Nominal concentrations of pesticides used in exposures and rainbow trout LC50 values.

Compound Name Type Exposure 

Concentrations ( g/l)

Rainbow Trout LC50

values ( g/l)

Oryzalin Herbicide 3, 30, 300, 3000 3260

Trifluralin Herbicide 0.05, 0.5, 5, 50 50

Prometryn Herbicide 0.9, 9, 90, 900 2900

Pendimethalin Herbicide 0.15, 1.5, 15, 150 138

Fenbutatin oxide Insecticide 0.01, 0.1, 1, 10 10

Thiobencarb Herbicide 0.8, 8, 80, 800 790

Propargite Insecticide 0.15, 1.5, 15, 150 <168

Metolachlor Herbicide 0.3, 3, 30, 300 300

1,3-Dichloropropene Insecticide 0.03, 0.3, 0.3, 3 270

Bromoxynil Herbicide 0.05, 0.5, 5, 50 41

Diflubenzuron Insecticide/Fungicide 2, 20, 200, 2000 72000

Table 2. Abnormalities observed during zebrafish embryo exposures.

Abnormality Description

Edema Accumulation of excess fluid in any one of the following cavities: heart, yolk sac, yolk 

extension, eyes.

Unhatched Failure to hatch at 5 dpf.

Curved Curvature of the tail dorsally in the sagittal plane such that a line drawn from the posterior tip 

of the notochord to the mouth of the fish would yield a gap between the line and body.

Lethargic An inability to maintain an upright posture and/or inactivity.

Deformed fins The absence or improper formation of fin tissue.

Deformed tail A notable shortening of the tail or improper notochord development.

Bent A bend in the body or tail of the embryo in the coronal plane.
385

Oryzalin

Oryzalin exposure did not impact developing zebrafish in a dose-dependent manner. 

Mortality was the highest (20%) at 30 g/l, but declined to 8.9% at 3000 g/l. Abnormality was 

the highest at 3000 g/l (17.1%), but was also elevated at 3 g/l (16.2%). The most common 

abnormality observed was edema. 

Figure 1: Percent mortality and abnormality observed in control and oryzalin-exposed zebrafish. 

Symbols are means (n = 3) SD. 
386

Table 3: Average length of fish exposed to oryzalin and controls (n = 3 dishes). There was a 

significant effect of oryzalin (One-way ANOVA, p < 0.0001). * Indicates treatment significantly 

different than controls (Tukey HSD, p < 0.01).

Bromoxynil

Bromoxynil exposure did not cause an increase in mortality or abnormality in developing 

zebrafish. The highest rate of abnormality (6.7%) was observed at 0.05 g/l and 50 g/l. 

Mortality occurred the most frequently at 0.5 g/l and 5 g/l at a rate of 2.2%. 

Treatment ( g/l) Average length ± SD (mm)

Water control 4.49 ± 0.02

0.1% acetone 4.50 ± 0.02

3 4.53 ± 0.05

30 4.48 ± 0.02

300 4.51 ± 0.05

3000 4.27 ± 0.02*
387

Figure 2. Percent mortality and abnormality in controls and zebrafish exposed to bromoxynil. 

Symbols are means (n = 3) SD. 

Table 4: Average length of fish exposed to bromoxynil and controls (n = 3 dishes). There was a 

significant effect of bromoxynil (One-way ANOVA, p < 0.0001). * Indicates treatment significantly 

different than controls (Tukey HSD, p < 0.05).

Treatment ( g/l) Average length ± SD (mm)

Water control 4.20 ± 0.04

0.1% acetone 4.06 ± 0.02

0.05 3.97 ± 0.03*

0.5 4.08 ± 0.01

5 4.05 ± 0.04

50 4.13 ± 0.06
388

Trifluralin

Exposure to trifluralin caused significant abnormalities at the highest dose tested (50 

g/l). The rate of abnormality at this dose was 95.3%, and the most common abnormality noted 

was lethargy, characterized by the absence of active swimming and a tendency to lose upright 

posture. Mortality was the greatest (22.2%) at 0.5 g/l. 

Figure 3. Percent mortality and abnormality of controls and zebrafish exposed to trifluralin. 

Symbols are means (n = 3) SD.
389

Table 5: Average lengths of fish exposed to trifluralin and controls (n = 3 dishes). There was a 

significant effect of trifluralin (One-way ANOVA, p < 0.0001). * Indicates treatment significantly 

different than controls (Tukey HSD, p < 0.01).

Treatment ( g/l) Average length ± 1 SD (mm)

Water control 4.02 ± 0.01

0.1% acetone 4.11 ± 0.07

0.05 4.05 ± 0.02

0.5 4.11 ± 0.07

5 4.01 ± 0.07

50 3.59 ± 0.03*
390

Prometryn

Prometryn exposure did not adversely affect either the rate of abnormality or mortality in 

developing zebrafish. The highest rate of morality observed was at 9 g/l (4.4%), and the 

highest rate of abnormality was at 0.9 g/l and 900 g/l (2.3%). 

Figure 4. Percent mortality and abnormality of controls and prometryn exposed fish. Symbols are 

means (n = 3) SD.

Table 6: Average lengths of fish exposed to prometryn and controls (n = 3 dishes). Exposure to 

prometryn did not significantly affect fish length (One-way ANOVA, p > 0.05).

Treatment ( g/l)

Average length ± SD (mm)

Water control 3.85 ± 0.06

0.1% acetone 3.96 ± 0.03

0.9 3.96 ± 0.06

9 3.95 ± 0.01

90 3.97 ± 0.03

900 3.85 ± 0.02
391

Pendimethalin

Embryos exposed to 150 g/l of pendimethalin developed a significant amount (100%) of 

abnormalities. Abnormal embryos were lethargic and struggled to swim. The highest rate of 

mortality (11.1%) was noted at 15 g/l.

Figure 5. Percent mortality and abnormality of controls and fish exposed to pendimethalin. 

Symbols are means (n = 3) SD.

Table 7: Average lengths of fish exposed to pendimethalin and controls (n = 3 dishes). 

Pendimethalin exposure significantly impacted the length of larvae (One-way ANOVA, p <0.001). * 

Indicates treatment significantly different than controls (Tukey HSD, p < 0.01). 

Treatment ( g/l) Average length ± SD (mm)

Water control 3.85 ± 0.06

0.1% acetone 3.96 ± 0.03

0.15 3.98 ± 0.04

1.5 3.97 ± 0.03

15 3.94 ± 0.06

150 3.59 ± 0.03*
392

Fenbutatin oxide

Fenbutatin oxide did not cause a dose-dependent change in mortality or abnormality. 

Mortality occurred the most frequently at 10 g/l (28.9%). Abnormality on the other hand was 

highest at 0.1 g/l (26.3%), and declined at higher concentrations. 

Figure 6. Percent mortality and abnormality of controls and fish exposed to fenbutatin oxide. 

Symbols are means (n = 3) SD. 

Table 8: Average lengths of fish exposed to fenbutatin oxide and controls (n = 3 dishes). 

Fenbutatin oxide exposure did not affect the length of fish (One-way ANOVA, p > 0.05).

Treatment ( g/l) Average length ± SD (mm)

Water control 3.90 ± 0.06

0.1% acetone 3.93 ± 0.01

0.01 3.91 ± 0.03

0.1 3.88 ± 0.06

1 3.91 ± 0.04

10 3.87 ± 0.02
393

Thiobencarb

Exposing developing zebrafish to thiobencarb produced abnormalities in 100% of the 

embryos at 800 g/l. The 5-dpf larvae behaved abnormally with erratic swimming patterns. 

Mortality at 800 g/l was 13.3%. 

Figure 7. Percent mortality and abnormality observed in controls and fish exposed to thiobencarb. 

Symbols are means (n = 3) SD.

Table 9: Average lengths of fish exposed to thiobencarb and controls (n = 3 dishes). There was a 

significant effect of thiobencarb (One-way ANOVA, p<0.0001). * Indicates treatment significantly 

different than controls (Tukey HSD, p < 0.01).

Treatment ( g/l) Average length ± SD (mm)

Water control 3.92 ± 0.04

0.1% acetone 3.99 ± 0.03

0.8 3.91 ± 0.03

8 3.87 ±0.04

80 3.91 ± 0.03

800 3.69 ± 0.07*
394

Propargite

Zebrafish embryos exposed to propargite did not show increased rates of mortality or 

abnormality. The highest rate of mortality (4.4 %) was observed at 0.15 g/l and 1.5 g/l. 

Embryos had the greatest number of abnormalities (13.6%) at 150 g/l.

Figure 8. Percent mortality and abnormality in controls and fish exposed to propargite. Symbols 

are means (n = 3) SD.

Table 10: Average lengths of fish exposed to propargite and controls (n = 3 dishes). Propargite 

produced significant effects (One-way ANOVA, p= 0.005). * Indicates treatment significantly 

different than controls (Tukey HSD, p < 0.01). 

Treatment ( g/l) Average length ± SD (mm)

Water control 3.92 ± 0.04

0.1% acetone 3.99 ± 0.03

0.15 3.95 ± 0.04

1.5 3.92 ± 0.04 

15 3.94 ± 0.02

150 3.83 ± 0.01*
395

Metolachlor

Exposure to metolachlor did not alter zebrafish mortality, although a higher rate (28.6%) 

of abnormality was observed at 300 g/l. The most frequent abnormality noted was a failure to 

hatch by 5 dpf. 

Figure 9. Percent mortality and abnormality of zebrafish exposed to metolachlor and controls. 

Symbols are means (n = 3) SD.

Table 11: Average lengths of fish exposed to metolachlor and controls (n = 3 dishes). There was a 

significant effect of metolachlor (One-way ANOVA, p < 0.0001). * Indicates treatment significantly 

different than controls (Tukey HSD, p < 0.01).

Treatment ( g/l) Average length ± SD (mm)

Water control 4.42 ± 0.03

0.1% acetone 4.37 ± 0.03

0.3 4.24 ± 0.06*

3 4.40 ± 0.05

30 4.23 ± 0.05*

300 4.18 ± 0.05*
396

1,3-Dichloropropene

Exposure to 1,3-dichloropropene caused an increase in abnormality and mortality in 

developing zebrafish, but not in a dose dependent manner. The highest rate of mortality (28.9%) 

occurred at 0.3 g/l, and declined at higher concentrations. The highest rate of abnormality 

(37.5%) was observed at 3 g/l. The rate of abnormality remained between 28.1% and 37.5% 

for all exposure concentrations and the most commonly observed abnormality was failure to

hatch by 5dpf. 

Figure 10. Percent mortality and abnormality observed in fish exposed to 1,3-dichloropropene and 

controls. Symbols are means (n = 3) SD.
397

Table 12. Average lengths of fish exposed to 1,3-dichloropropene and controls (n = 3 dishes). 

There was a significant effect of 1,3-dichloropropene (One-way ANOVA, p <0.001). * Indicates 

treatment significantly different than controls (Tukey HSD, p < 0.01).

Treatment ( g/l) Average length ± SD (mm)

Water control 4.46 ± 0.03

0.1% acetone 4.34 ± 0.03

0.03 4.32 ± 0.04

0.3 4.14 ± 0.05*

3 4.28 ± 0.08

30 4.27 ± 0.06

Diflubenzuron

Diflubenzuron did not influence zebrafish mortality or abnormality. The highest rate of 

abnormality (6.8%) was observed at 20 g/l, and the highest rate of morality (4.4%) was 

observed at 2 g/l. However, it is important to note that diflubenzuron was difficult to work with 

because of its low solubility in acetone (6.5 g/l). The most concentrated stock solution of 

diflubenzuron we were able to make was 2 g/l. Diflubenzuron appeared to remain in solution 

after dosing the exposure dishes, however after 24hrs, the highest exposure concentration dishes 

(2000 g/l) had visible floating particles. Thus, without using alternative methodologies (e.g. 

DMSO as the carrier), we are not confident about accurate dosing for this compound. 
398

Figure 11. Percent mortality and abnormality observed in control fish and fish exposed to 

diflubenzuron. Symbols are means (n = 3) SD.

Table 13: Average lengths of fish exposed to diflubenzuron and controls (n = 3 dishes). There was 

not a significant effect of diflubenzuron on fish length (One-way ANOVA, p > 0.05). 

Treatment ( g/l) Average length ± SD (mm)

Water control 4.20 ± 0.04

0.1% acetone 4.06 ± 0.02

2 4.06 ± 0.03

20 4.88 ± 0.06

200 4.98 ± 0.03

2000 4.08 ± 0.03
399

References

Kimmel, CB, WW Ballard, SR Kimmel, B Ullman and TF Schilling. 1995. Stages of 

embryonic development of the zebrafish. Developmental Dynamics 203(3):253-310.

Linbo, TL. 2009. Zebrafish (Danio rerio) husbandry and colony maintenance at the Northwest 

Fisheries Science Center. U.S. Dept. Commer., NOAA Tech. Memo. NMFS-NWFSC-100, 62p.
400

21. Appendix 6: Conservation Values of Designated Critical Habitat for 

Listed Salmonids in California’s Central Valley
401

Rice Areas and Conservation Values of Designated Critical Habitat for Listed Salmonids in 

California’s Central Valley

The maps below provide conservation values (High, Medium, Low) for each of the watershed 

units within the designated critical habitats for the three California Central Valley listed 

salmonids. Not all habitats have the same conservation value for an ESU. In determining what 

areas are critical habitat for listed salmonids, NMFS had to consider those physical or biological 

features that are essential to the conservation of a given species, called the primary constituent 

elements (PCEs). Designation of the conservation values required taking into account the 

quality, quantity, and distribution of PCEs within each watershed (50 CFR Part 226, September 

2, 2005). Using GIS, we overlaid the NASS CDL 2010 data layer for rice to get a spatially 

relevant distribution of rice growing areas with the watersheds and their conservation values. 

We evaluated the overlap between rice growing areas and designated critical habitat. 

At this point in time, conservation values within the designated critical habitat for Sacramento 

River winter-run Chinook have not been assigned. For this Opinion, we assumed each watershed 

to have a high conservation value. 
402
403
404