Primary constituent elements within critical habitat boundaries are essential for the conservation of these DPSs – these are sites and habitat components that support one or more life stages, and include:

(1) Freshwater spawning sites with water quantity and quality conditions and substrate supporting spawning, incubation and larval development;

​

(2) Freshwater rearing sites with:

(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.

​

(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.

(d) Exclusion of Indian lands. Critical habitat does not include habitat areas on Indian lands. The Indian lands specifically excluded from critical habitat are those defined in the Secretarial Order, including:

(1) Lands held in trust by the United States for the benefit of any Indian tribe;

(2) Land held in trust by the United States for any Indian Tribe or individual subject to restrictions by the United States against alienation;

(3) Fee lands, either within or outside the reservation boundaries, owned by the tribal government; and

(4) Fee lands within the reservation boundaries owned by individual Indians.

(e) Land owned or controlled by the Department of Defense. Critical habitat does not include any areas subject to an approved Integrated Natural Resource Management Plan or associated with Department of Defense easements or right-of-ways. In areas within Navy security zones identified at 33 CFR 334 that are outside the areas described above, critical habitat is only designated within a narrow nearshore zone from the line of extreme high tide down to the line of mean lower low water.

(f) Land covered by an approved Habitat Conservation Plan. Critical habitat does not include any areas subject to an approved incidental take permit issued by NMFS under section 10(a)(1)(B) of the ESA. The specific sites addressed include those associated with the following Habitat Conservation Plans:

(1) Washington Department of Natural Resources—West of Cascades

(2) Washington State Forest Practices, except those lands on the Kitsap Peninsula overlapping with areas occupied by Puget Sound steelhead and not classified as being in an approved or renewed status by the Washington Department of Natural Resources as of September 2015.

(3) Green Diamond Company.

(4) West Fork Timber Company.

(5) City of Kent.

(6) J.L. Storedahl and Sons.

Toxicity of potash (KCl) to fish: The efficacy of potassium chloride treatment is inhibited in the presence of elevated dissolved solids, in particular the concentration of sodium (Moffitt et al. 2016, Stockton-Fiti and Moffitt 2017).

​

Based upon the acute toxicity testing of KCl using both juvenile Brook Trout (Salvelinus fontinalis) and juvenile Chinook Salmon (Oncorhynchus tshawytscha), acute lethal effects of potash on these salmonids at these life stages are not expected at concentrations commonly used to control invasive dreissenid mussels (100 mg/L) (Densmore et al. 2018). Exposure concentrations of as much as 800 mg/L KCl, eight times greater than the dose of KCl used as a molluscicide, were applied to these fish species in static systems for 96 hours; there was no evidence of mortality attributable to KCl exposure among either species (Densmore et al. 2018). Behavioral or gross morphological effects on these fish from KCl-based molluscicide applications at levels up to 800 mg/L were also not indicated (Densmore et al. 2018). Hillard et al. (2019) exposed Fall Chinook Salmon eggs within 24 hours of fertilization to 750 mg/L potassium chloride for one hour followed by 20 mg/L formalin for three hours to disinfect the eggs and kill any zebra mussel larvae. There was no effect on salmon egg viability.

​

Several listed fish species forage on invertebrates, particularly during juvenile life stages. The ecotoxicity of muriate of potash on invertebrates is 48 hours @ EC50 @ 337–825 mg/L (Daphnia magna), and 96 hours @ LC50 @ 940 mg Physa heterostropha (Mosaic 2004). Fathead minnow (Pimephales promelas) trials - LC50 @ 880 mg/L KCl for 96 hours (Mount et al. 2007). Daphniid exposure trials – LC50 @ 196 mg/L for 48 hours; significant mortality of sensitive aquatic invertebrates is not expected at the KCl concentrations used to control dreissenids (Densmore et al. 2018). Crayfish (Cambarus spp.) exposure trials resulted in mortality and temporary paralysis at concentrations of 800 and 1,600 mg/L for at least 24 hours (Densmore et al. 2018). Other ecotoxicology studies: Bluegill sunfish (Lepomis macrochirus) – LC50 – 2010 mg/L (Mosaic 2014). Substantial differences exist in the accuracy of models to predict organism survival to introduced toxins, such as potassium, calcium, and magnesium (Pillard et al. 2000). Brook Trout are less sensitive to potassium chloride than Rainbow Trout (Oncorhynchus mykiss), and Fathead Minnows are two to three times more sensitive to KCl than Brook Trout or Rainbow Trout (Lazorchak 2007).

​

Potash solutions in water can have varied concentrations throughout the water column because of the density of potash relative to the density of water in cold-water conditions (Lund et al. 2017). Hotspots, or accumulations of potash in higher concentrations, can accumulate at deeper depths (Lund et al. 2017).

As a result of these toxicity studies, in basin locations in which ESA-listed salmonids and their critical habitat exists, potash may be the most likely product to be used, based on least toxicity to aquatic life, and in particular trout and salmonids, as well as cost.

Copper Products (general information) Copper is one of the most toxic heavy metals to fish (Nowak and Duda 1996), and is broadly toxic to the salmon olfactory system (Baldwin et al. 2003), potentially inhibiting the ability of juveniles to avoid predators or to find food (DeForest et al. 2010). Hara et al. (1976) documented that copper is a neurobehavioral toxicant in fish that disrupts the normal function of the fish olfactory system. Dissolved copper damages the olfactory sensory epithelium (Moran et al. 1992; Julliard et al. 1996; Hansen et al. 1999a; Beyers and Farmer 2001). Copper interferes with the ability of fish to detect and respond to chemical signals in aquatic environments (Hara et al. 1976; Saucier et al. 1994; Hansen et al. 1999b; Beyers and Farmer 2001; Baldwin et al. 2003; Sandahl et al. 2004; Carreau and Pyle 2005).

​

Copper impairs gill gas exchange, upsets salt balance, negatively affects reproductive output and the immune system, affects glucose metabolism and the cellular structure of fish, and negatively affects liver and kidney function (Ferguson and Sandoval 2020). The proposed low and high application rates are well above the range of salmonids and their prey LC50 (96 hour), and the LC50 (96 hour) for pond snails falls at the lowest proposed application rate (TOXNET 1975–1986). For salmonids, the upper recommended limit is < 0.03 mg/l in hard water (>100 mg/l CaCO3) whereas in soft water, it is <0.0006 mg/l (Ferguson and Sandoval 2020).

Freshwater fish acutely exposed to copper can experience disturbances of ion regulation and ammonia excretion (Macpherson and Cremazy 2024). Acute copper exposure decreased fish survival in Brook Char, inhibited Na+ influx, transiently promoted Cl- loss, and transient inhibition of ammonia clearance (Macpherson and Cremazy 2024).

Baldwin et al. (2011) documented that naturally reared steelhead (O. mykiss) exposed to nominal copper concentrations of 5 and 20 µg/L for 3 hours exhibited a concentration-dependent loss in olfactory function.

The sensory physiology and predator avoidance behaviors of juvenile coho were both significantly impaired by copper at concentrations as low as 2 µg/L (Sandahl et al. 2007).

Juvenile coho exposed to low levels of dissolved copper (5-20 µg/L) and then presented with cues signaling proximity of a predator exhibited no alarm response when exposed to copper, and cutthroat trout were more effective predators on copper-exposed coho during predation trials, as measured by attack latency, survival time, and capture success rate. Copper-exposed coho are unresponsive to their chemosensory environment, unprepared to evade nearby predators, and significantly less likely to survive an attack sequence (McIntyre et al. 2012).

In a study to assess susceptibility of Chinook Salmon and Rainbow Trout to infection with Vibrio anguillarum following sublethal copper exposure (Baker and Knittel 1983), higher concentrations of copper produced peak susceptibility to infection in shorter time periods. Rainbow Trout stressed by copper required 50% fewer pathogens to induce a fatal infection than non-copper exposed fish (Baker and Knittel 1983).

​

Fish eggs are more resistant than young fish fry to the toxic effects of copper sulfate (Gangstad 1986).

• Juvenile Rainbow Trout were exposed to either hard water, or soft water, spiked with copper for 30 days (Taylor et al. 2000). Fish in the hard-water, high dose (60 µg/L) treatment groups showed an increased sensitivity to copper.

• The mean 96-hour LC50 (with 95% confidence limits) for copper exposure in alevin, swim-up, parr and smolt Steelhead (Salmo gairdneri) are 28 (27–30), 17 (15–19), 18 (15–22), and 29 (>20) µg/L of copper respectively (Chen and Lin 2001). The mean 96-hour LC50 for copper exposure in alevin, swim-up, parr and smolt Chinook salmon are 26 (24–33), 19 (18–21), 38 (35–44), and 26 (23–35) µg/L of copper respectively. The experiments were done by adding copper as CuCl2.

• The 48-hour LC50 for Fathead Minnow is 19.2 + 3.1 (mean + SD) mcg/L Cu (Mastin and Rodgers 2000).

​

EarthTec QZ

According to the label for EarthTec QZ, “this pesticide is toxic to fish and aquatic invertebrates. Waters treated with this product may be hazardous to aquatic organisms. Treatment of aquatic weeds and algae can result in oxygen loss from decomposition of dead algae and weeds. This oxygen loss can cause fish and invertebrate suffocation. Do not use this product in waters with cyprinid and salmonid fish.” EarthTec is produced in the cupric ion form, the most toxic form of copper (Ferguson and Sandoval 2020).Direct bioassay of Rainbow Trout (assumed adult) subject to EarthTec QZTM resulted in a NOEC of 0.240 mg/L copper, and LC50 of 0.294 mg/L copper. The EPA maximum allowable dose is 1 mg/L. Fish kills have been reported after copper sulfate applications for algae control in ponds and lakes, however, oxygen depletion and dead organisms clogging the gills have been cited as the cause of fish deaths, resulting from massive and sudden plant death and decomposition in the water body (Bartsch 1954, Hanson and Stefan 1984, Masser et al. 2006). Copper can either temporarily, or permanently, disrupt olfaction in fish (Solomon 2009, Ferguson and Sandoval 2020), possibly interfering with their ability to locate food, predators, and spawning streams (Chapman 1978, Jaensson and Olsen 2010). It is unknown if there are any bioaccumulation effects of EarthTec QZTM; one recent study suggested the fate of copper in the environment is not fully known and should be considered (Lake-Thompson and Hofmann 2019).

Copper Ethanolamine Complex (e.g., Natrix, Cutrine Plus)

Natrix

The active ingredient in SePRO's Natrix, registered for the control of mollusks in still or flowing waters as well as crop and non-crop irrigation and drainage systems and chemigation systems, is Copper Ethanolamine Complex (28.2%), EPA Reg. # 67690-81. Per its label, "Unlike most organic pesticides, copper is an element and will not break down in the environment and will therefore accumulate with repeated applications. Copper is a micronutrient, but its pesticidal application rate exceeds the amount of copper needed as a nutrient."

Label environmental hazards: "This pesticide is toxic to fish and aquatic invertebrates. Waters treated with this product may be hazardous to aquatic organisms. Treatment in areas with dense aquatic weeds and algae can result in oxygen loss from decomposition of dead algae and weeds. This oxygen loss can cause fish and invertebrate suffocation. To minimize this hazard, do not treat more than ½ of the water body to avoid depletion of oxygen due to decaying vegetation. Do not make applications less than 14 days apart." "Certain water conditions including low pH (<6.5), low dissolved organic carbon (DOC) levels (3.0 mg/L or lower), and “soft” waters (i.e. alkalinity less than 50 mg/L), increases the potential acute toxicity to non-target aquatic organisms. Fish toxicity generally decreases when the hardness of water increases. Do not use in waters containing trout or other fish species that are highly sensitive to copper if the alkalinity is less than 50 ppm, pH values are <6.5, and DOC levels >3.0."

According to the Safety Data Sheet for Natrix, toxicological effects on rabbits (dose >2000 mg/kg) = LD50 dermal and rats (dose 590 mg/kg) = LD50 oral. Natrix has been demonstrated to be a several irritant to the eyes of rabbits @ 0.1 ml exposure and a severe irritant to the skin of rabbits @ 0.5 ml.

Natrix is comprised of 14.9% copper triethanolamine complex, 13.3% copper monoethanolamine complex, and 4 proprietary ingredients (3 of them range from 10-30% and one ranges from 30-60%).

Proprietary ingredient 1:

Crustaceans - Ceriodaphnia dubia Neonate - exposed for 48 hours (acute EC50 609.98 mg/L fresh water), fish - Pimephales promelas - exposed for 96 hours (acute LC50 11800 mg/L fresh water), and daphnia - Daphnia magna - exposed for 21 days (chronic NOEC 16 mg/L fresh water).

Proprietary ingredient 2:

Algae - Isochrysis galbana - exposed for 96 hours (acute EC50 80 mg/L fresh water), adult crustaceans - Crangon crangon - exposed for 48 hours (acute LC50>100 mg/L marine water), fish - Carassius auratus - exposed for 96 hours (acute LC50 170 mg/L fresh water)

Proprietary ingredient 3:

Crustaceans - Ceriodaphnia dubia - Neonate - exposed for 48 hours (acute EC50 4.53 mg/L fresh water)

Cutrine Plus

The active ingredient in Cutrine Plus is copper ethanolamine complex. According to the product label, "This copper product is toxic to fish and aquatic organisms. Unlike most organic pesticides, copper is an element and will not break down in the environment and will therefore accumulate in sediment with repeated applications. Copper is a micronutrient, but its pesticidal application rate exceeds the amount of copper needed as a nutrient. Do not use in waters containing Koi and hybrid goldfish. Not intended for use in small volume, garden pond systems. Avoid treating waters with pH values <6.5, DOC levels >3.0, and alkalinity less than 50 ppm (e.g., soft or acid waters), as trout and other sensitive species of fish may be killed under such conditions if present."

According to the Safety Data Sheet for Cutrine Plus, acute oral toxicity in rats (LC50) is believed to be 1,000 mg/kg (acute toxicity estimate is >40 mg/l @ an exposure time of 4 hours. Acute dermal toxicity in rabbits (LD50) is >5,000 mg/kg.

Duke (2007) documented both laboratory and field responses of target and non-target species to algaecide exposures, including exposure to Cutrine Plus. Ceriodaphnia dubia exposed to Cutrine Plus @ a concentration of 54 µg acid-extractable Cu/L resulted in 96-h LC50 and @ a concentration of 70 µg acid-extractable Cu/L resulted in 96-h EC50. Murray-Gulde (2002) documented a concentration of 124 µg acid-extractable Cu/L resulted in 96-h LC50.

Pimephales promelas exposed to Cutrine Plus concentrations of 731 and 973 µg acid-extractable Cu/L resulted in 96-h LC50 (Duke 2007). Murray-Gulde (2002) documented Cutrine Plus concentrations of 863 µg acid-extractable Cu/L resulted in 96-h EC50. Mastin and Rodgers (2000) documented a Cutrine Plus concentration of 255.4 µg acid-extractable Cu/L resulted in 48-h EC50.

Mastin and Rodgers (2000) documented the toxicity and bioavailability of copper herbicides, including Cutrine-Plus and copper sulfate, to freshwater animals 7 days after initial herbicide applications. Aqueous 48-h toxicity experiments were performed to contrast responses of Daphnia magna Strauss, Hyalella azteca Saussure, Chironomus tentans Fabricius, and Pimephales promelas Rafinesque to Cutrine(R)-Plus. 48-h LC50s for organisms exposed to Cutrine-Plus was 11.3 µg Cu/L. Organisms exposed to Cutrine-Plus and copper sulfate had exposure-response slopes of 8.61 and 5.07% mortality/µg Cu/L, respectively. Bioavailability of Cutrine-Plus was determined by comparing survival data (LC50s) of test organisms exposed to herbicide concentrations during the first and last 48-h of a 7-day exposure period. Even in these relatively simplified water-only exposures, a transformation of copper to less bioavailable species over time was observed with a 100-200% decrease in toxicity (i.e., an increase in 48-h LC50s) for all four test animals. This series of laboratory experiments provides a worst-case scenario for determining the risk associated with the manufacturer's recommended application rates of Cutrine-Plus (200-1,000 microg Cu/L) and copper sulfate (100-500 microg Cu/L) in natural waters for four nontarget freshwater animals.

The EPA Ecotox database includes 7 research studies associated with copper ethanolamine complex:

• Water flea (Daphnia magna) less than 24 hours old were exposed to 0.82 (0.56-0.94) AI mg/L and 0.2 AI mg/L in the lab in flow-through fresh water conditions, and EC50 was 2 days and no observed effective level (NOEL) was observed after 2 days, respectively (US EPA 1992).

• Rainbow trout (Oncorhynchus mykiss) were exposed to 1.5 (1.3-1.7) AI mg/L in the lab in flow-through freshwater conditions, and LC50 was observed after 4 days (US EPA 1992).

• Bluegill (Lepomis macrochirus) were exposed to 2 AI mg/L and 4.2 (3.5-6) in the lab in flow-through fresh water conditions, and NOEL was observed after 4 days and mortality was observed after 4 days, respectively (US EPA 1992).

Zequanox requires containment within a barrier system when applied to surface waters for the management of dreissenid mussels; without containment, effective treatment concentrations cannot be maintained (Luoma et al. 2019).

​

No mortality from Zequanox® has been observed in Fathead Minnows, young-of-the-year Brown Trout (Salmo trutta), and juvenile Bluegill Sunfish (Bureau of Reclamation 2011). Fish trials conducted with dead bacteria have indicated that applications of killed cells were harmless to fish, yet were still highly lethal to Dreissena spp. mussels (Bureau of Reclamation 2011). Temporary, but substantial, reductions in dissolved oxygen were observed in treatment locations during the morning following Zequanox® treatment in two trials, likely due to the presence of the barriers that prevented well-oxygenated water from circulating into treatment zones from adjacent areas in the lake (Whitledge et al. 2015). During one of the few open water treatments using Zequanox®, dissolved oxygen levels declined as a result of using containment barriers (DO values of 0.1ppm were reached within 24 hours, and these hypoxic conditions continued for 7 days) (Lund et al. 2017).

​

A 2018 study evaluated the effects of Zequanox® on juvenile Lake Sturgeon (Acipenser fulvescens) and Lake Trout (Salvelinus namaycush) (Luoma et al. 2018). No acute mortality was observed in either species; however, significant latent mortality was observed in Lake Trout that were exposed to the highest dose of Zequanox®. Statistically significant, but biologically minimal, differences were observed in the weight (range 20.17 to 21.49 g) of surviving Lake Sturgeon at the termination of the 33 d post-exposure observation period. Survival was not impacted in the Lake Trout in the 100 mg/L treated group for the first 3 weeks; however, impacts were readily detectable 4 weeks (28 d) after Zequanox® exposure. Poor food consumption, emaciation, and abdominal hemorrhaging were observed about 3 to 4 weeks after exposure in some of the Lake Trout exposed to 100 mg/L A.I. of Zequanox®.

​

Cold water, cool water, and warm water fish were tested for exposure-related effects to Pseudomonas fluorescens, Strain CL145A. (Luoma et al. 2015). Analyses of test animal condition factors and survival revealed that a 24-hour continuous dose of SDP affected all species. Calculated concentrations of SDP that would be lethal to 50 percent of the test animals (LC50) for the cold water species were 19.2 and 104.6 mg/L for Rainbow Trout and Brook Trout, respectively. The LC50 for the cool water species were 185.4, 176.9 and 8.9 mg/L for Yellow Perch (Perca flavescens), Walleye (Sander vitreus), and Lake Sturgeon, respectively. The LC50 for the warm water species were 173.6, 139.4, and 63.1 for the Largemouth Bass (Micropterus salmoides), Smallmouth Bass (Micropterus dolomieu), and Channel Catfish (Ictalurus punctatus), respectively.

Barbour et al. (2021) assessed avoidance behavior of fish to Zequanox® exposure, and determined that Brook Trout, Lake Trout, and Bluegill Sunfish avoided Zequanox®-treated water, and Lake Sturgeon and Fathead Minnow were attracted to Zequanox®-treated water. In the Barbour et al. (2021) study, one species of fish, Yellow Perch, was indifferent to Zequanox® exposure, likely because of their ability to function in both clear and turbid water and their tolerance to Zequanox® compared to other fishes. Sensitive fish species may avoid Zequanox® treatments, which can reduce overall risk to a species (Barbour et al. 2021).

​

Crain et al. (2008) assessed the effects of multiple stressors on aquatic environments and concluded effects were additive, or synergistic. Therefore, any of the treatments described above are assumed to adversely affect aquatic species to some degree and for some duration.

Zequanox® greatly increases turbidity (Barbour et al. 2021). Turbidity was 13x higher than in control plots, but there was no difference between treatment and control plots after 8 hours (Luoma et al. 2019). Fish behavioral response to turbidity varies with species and life stage, but has been associated with mortality (Kemp et al. 2011). Although salmonids have the ability to cope with some levels of turbidity during certain life stages (Gregory and Northcote 1993), salmonid populations not normally exposed to high levels of natural turbidity, or anthropogenic sources, may be negatively affected by low levels, e.g., 18-70 NTU (Gregory 1992). Effects of turbidity on salmonids include physiological effects (e.g., gill trauma, osmoregulation challenges, blood chemistry changes, and effects on reproduction and growth), behavioral effects (e.g., avoidance, territoriality, foraging and predation, and homing and migration), and habitat effects (e.g., reduction in spawning habitat, effect on hyporheic upwelling, reduction in BI habitat, and damage to redds (Bash et al. 2001).

​

Spawning, foraging, and seasonal migrations should be considered when planning Zequanox® treatments (Barbour et al. 2021); altering treatment timing and limiting the treatment area to provide access to refugia may be helpful mitigation steps.

Winter drawdowns can decrease taxonomic richness of macrophytes and benthic invertebrates and shift assemblage composition to favor taxa with r-selected life history strategies and with functional traits resistant to direct and indirect drawdown effects (Carmignani and Roy 2017). Fish assemblages, though less directly affected by winter drawdowns (except where there is critically low dissolved oxygen), can be indirectly negatively affected via decreased food resources and changes in spawning habitat (Carmignani and Roy 2017).

Drawdowns modify abiotic conditions, cause sediment dessication and freezing, place stress on vegetative root structures (Siver et al. 1986), displace plants as a result of erosion of frozen sediment during spring refills (Beard 1973, Mattson et al. 2004), and stifle species growth by increasing acidity and cations to toxic concentrations (Peverly and Kopka 1991). Annual winter drawdowns can, through time, coarsen sediment texture and remove nutrients in the exposure zone, making these sites unsuitable for macrophyte colonization and growth, particularly in steep-sided basins (Hellsten 1997).

​

Other adverse impacts of drawdowns (New Hampshire Department of Environmental Services 2019) may include:

• Large amounts of aquatic plants and organisms that succumb to the drawdown begin to decay shortly after drawdown, but nutrient release to the water body may not occur until full-pond level is achieved. Nutrients released from decayed material will quickly be used by algae and cyanobacteria, leading to increased cell production. Shallow lakes have shown shifts from clear, plant-dominated conditions to turbid, algal dominated systems.

• Algal or cyanobacteria blooms may follow.

• Aquatic food web changes may result in shifts in plant and animal structure.

• Oxygen concentrations throughout the water column may be impacted.

• Changes in the bottom sediment may also occur. Softer sediments may become compacted, or frozen segments that are lighter than water could loosen and float around in large masses, or as floating islands in the water body, only to settle once again in a new location.

Impacts to aquatic animal species can be significant. These impacts range from stranding animals to food chain modifications, or stressors associated with the drawdown. Fish, frogs, salamanders, turtles, aquatic insect larvae, mussels, and others can be affected by a drawdown. Agile and faster moving organisms may be able to move upstream or downstream to other unimpacted habitats, however, these fish may be confined to smaller, shallower areas where they become easy prey to consumers, or suffer from oxygen deprivation. Slower moving, more sedentary organisms have a greater risk to negative impacts. Freshwater mussels, snails, insects, and crayfish may not be able to find suitable habitat, and may succumb to the drawdown.

Fall and spring spawners, juvenile life stages in littoral zones, and insectivorous fish can be sensitive to drawdowns. Littoral spawning in the fall—Low water levels in spring can prevent fish access to spawning areas; the amount of fall to late spring drawdown is inversely correlated to year-class strengths of coregonid fishes (Gaboury and Patalas 1984). Fish that spawn on reservoir bottoms with winter drawdowns can experience dissolved oxygen deficiency in late winter, which affects survival of eggs and year-class strength (Sutela et al. 2002). Late winter drawdowns reduced lake whitefish abundance by more than 80% during three years of drawdowns because of reduced recruitment and decreased survival (Mills et al. 2002). Littoral spawning in the spring—Dewatered areas in early spring can limit the recruitment of spring spawners, such as northern pike (Kallemeyn 1987). Spring spawning could be negatively impacted by the effects of drawdowns that occur during years when winter and spring droughts occur (McDowell 2012). Littoral juvenile life stage—Different species of fish use differing behavioral strategies to address water fluctuations in natural and man-made lakes. One study tested fish behavior when lake level was decreased in the fall; larger burbot were more successful competing for suitable shelter than smaller burbot until a certain level, at which the largest fish abandoned shelter use while smaller fish persisted in sheltering behavior (Fischer and Öhl 2005). In contrast, stone loach showed no hierarchical order, or size-related shelter use (Fischer and Öhl 2005). Insectivorous fish—Hypolimnetic draws are associated with reduced abundance of aquatic fish and invertebrate communities and macroinvertebrates downstream of a dam (Haxton and Findlay 2008).

It is often desirable to remove any trapped fish from isolated treatment areas to avoid and minimize impacts to fish, particularly ESA-listed fish. However, the activity of salvaging fish, which may include fish collection, handling transportation and release away from export facilities can cause physiological stress to fish (Afentoulis et al. 2023). A study involving Delta Smelt (Hypomesus transpacificus) in the San Francisco Bay Estuary document stress response as a result of fish salvage (Afentoulis et al. 2023). Delta Smelt experienced stress in all phases of fish salvage, as evidenced by higher plasma cortisol concentrations. Wild Delta Smelt experienced higher levels of cortisol response and took longer to recover than did cultured Delta Smelt. Milston (2006) recommended fish not be exposed to successive stressors for at least 24 hours after the initial stressor. Afentoulis et al. (2023) recommends stress-reduction strategies be implemented, such as allowing salvaged fish to recover prior to release by placing them in commercial netpens for a period. Other strategies include written standard operating procedures for fish handling, enhanced operations staff training, reducing fish densities in the bucket and truck, and releasing fish at night (Afentoulis et al. 2023).