US Wind would bring up to four offshore export cables through Indian River Bay to shore to landing locations at 3Rs Beach or Tower Road within the Delaware Seashore State Park (figure 1). The US Wind export cable would be connected to the onshore transmission cable at the landfall locations using horizontal directional drilling (HDD) and a jet plow. Cables would be pulled into cable ducts that would route the cables under the beach to subterranean transition vaults, located in existing developed areas such as parking lots. US Wind evaluated cofferdams at the HDD locations and determined that the use of a gravity cell would be more appropriate for soil conditions as well as avoid the use of a vibratory hammer that would create additional underwater sound. The gravity cell would be lowered onto the seafloor and would not require the walls of the cell to be driven into the seabed ( i.e., no pile driving would occur). The HDD drill rig would be set up onshore in an excavated area and the drill would advance to the offshore exit point. The offshore cable would be pulled in through the HDD ducts into the cable jointing/transition vault at the landfall location. The cable installation vessel would then begin laying the cable on the seabed as described in the Cable Laying and Installation section below. Given the work is not expected to produce noise levels that could result in harassment to marine mammals, HDD and gravity cell installation is not expected to result in the take of marine mammals. US Wind did not request, and NMFS is not proposing to authorize, take associated with cable landfall construction; therefore, this activity is not discussed further.

Cable burial operations would occur both in the Lease Area and ECCs from the Lease Area to shore. The inter-array cables would connect the WTGs to any one of the OSSs. All WTGs would connect to an OSS in strings of 4-6 WTGs via the inter-array cables. Cables within the ECCs would carry power from the OSSs to shore at the landfall location(s) within Delaware Seashore State Park. The offshore export cables would be buried in the seabed at a target depth of up to 1 m (3.3 ft) to 3 m (9.8 ft), although the exact depth would not exceed 4 m (13.1 ft). Inter-array cable burial operations would be installed to a target depth of 1 m (3.3 ft) to 2 m (6.6 ft), not to exceed 4 m (13.1 ft) in depth and would follow installation of the WTG and OSS foundations as the foundations must be in place to provide connection points. Offshore cable installation may occur concurrently with foundation installation.

Cable laying, cable installation, and cable burial activities planned to occur during the construction of the Project would include the following methods: offshore export cable pull through the HDD duct, simultaneous lay and burial for cable installation through the use of a jet plow, and post-lay burial for cables, as needed. Offshore export cables would be pulled through the HDD duct, as described in the Cable Landfall Construction section above. The inter-array cables would be installed from a dynamically positioned cable installation vessel. US Wind plans to use a jet plow to achieve the target inter-array and offshore cable burial depth. If necessary, post-lay cable burial would be completed through the use of a cable installation support vessel and remotely operated vehicle (ROV) system (US Wind, Inc., 2023a). Areas with cable crossings or hard bottoms may require additional protection measures, such as mattresses, rock placement, or cable protection systems. In shallow areas of cable installation, dredging may be necessary to allow access by the cable lay barge. As the noise levels generated from cable laying and installation work are low, the potential for take of marine mammals to result is discountable. US Wind is not requesting, and NMFS is not proposing, to authorize take associated with cable laying activities. Therefore, cable laying activities are not analyzed further in this document.

Site preparation typically includes sand bedform leveling, boulder clearance, pre-lay grapnel runs, and a pre-lay survey to prepare the area for export cable installation. Route clearance activities would be conducted prior to offshore export cable installation. Project activities would include a pre-installation survey and grapnel run along the offshore export cable corridor to remove debris that could impact the cable lay and burial. US Wind does not expect pre-installation seabed preparation, such as leveling, pre-trenching, to be necessary. A pre-lay grapnel run would be conducted along the cable route to remove debris that could impact cable lay and burial.

US Wind would also deposit rock around each foundation as scour protection. Prior to or following the installation of a monopile or jacket foundation for the OSS, a first layer of scour protection rocks will be deployed in a circle around the pile location to stabilize the seabed (US Wind, Inc., 2023a). If suction bucket foundations are selected for OSSs, scour protection would be deployed after buckets reach target penetration depth. A 1-2 m (2-7 ft) thick second layer of larger rocks would be placed for stabilization once the inter-array cables have been pulled into the monopile. Scour protection may also be applied as additional protection for cables after burial.

NMFS does not expect scour protection placement or site preparation work, including pre-lay grapnel runs and pre-lay surveys, to generate noise levels that would cause take of marine mammals. Although not anticipated, any necessary dredging, bedform leveling, or boulder clearance would be extremely localized at any given time, and NMFS expects that any marine mammals would not be exposed at levels or durations likely to disrupt behavioral patterns ( i.e., migrating, foraging, calving, etc. ). Therefore, the potential for the take of marine mammals to result from these activities is so low as to be discountable. US Wind did not request, and NMFS is not proposing, to authorize any takes associated with site preparation and scour protection activities; therefore, they are not analyzed further in this document.

US Wind will utilize various types of vessels over the course of the 5-year proposed regulations for surveying, foundation installation, cable installation, WTG and OSS installation, and support activities. US Wind has identified several existing port facilities located in Maryland, Virginia, Delaware, and New Jersey to support offshore construction, assembly and fabrication, crew transfer and logistics, and other operational activities. In addition, some components, materials, and vessels could come from Canadian and European ports. A variety of vessels would be used throughout the construction activities. These range from crew transportation vessels, tugboats, jack-up vessels, cargo ships, and various support vessels (table 5). Details on the vessels, related work, operational speeds, and general trip behavior can be found in table 1-2 of the ITA application and table 4-1 in the COP volume 1.

As part of various vessel-based construction activities, including cable laying and construction material delivery, dynamic positioning thrusters may be utilized to hold vessels in position or move slowly. Sound produced through use of dynamic positioning thrusters is similar to that produced by transiting vessels, and dynamic positioning thrusters are typically operated either in a similarly predictable manner or used for short durations around stationary activities. Fall pipe vessels may use dynamic positioning thrusters during the installation of scour protection up to 24 hours per day. Jack-up cranes or floating cranes may use dynamic positioning thrusters for up to 4 hours per WTG or OSS installation. Heavy lift and general cargo vessels may use dynamic positioning thrusters for the delivery of Project components from the manufacturing location to the staging/assembly port only while maneuvering in port. Multipurpose offshore supply vessels may also use dynamic positioning thrusters throughout the day during the pre-lay grapnel run boulder clearance and cable burial. Jack-up or accommodation vessels may use dynamic positioning thrusters while constructing housing for offshore works, yet only while maneuvering to the site, which would last approximately 2 hours per WTG or OSS. Dynamic positioning thrusters may also be used by vessels throughout the day for pre-installation, geophysical and geotechnical verification surveys, cable installation, placement of scour protection and concrete mattresses, seabed preparation and leveling, and commissioning activities. Sound produced by dynamic positioning thrusters would be preceded by, and associated with, sound from ongoing vessel noise and would be similar in nature; thus, any marine mammals in the vicinity of the activity would be aware of the vessel's presence. Construction-related vessel activity, including the use of dynamic positioning thrusters, is not expected to result in take of marine mammals. US Wind did not request, and NMFS does not propose to authorize, any take associated with vessel activity.

The total vessels expected for use during the Project are provided in table 5; more details can be found in table 1-2 of the ITA application. Assuming the maximum design scenario, approximately 458 total vessel round trips are expected to occur during the MarWin construction campaign (2025), approximately 1,944 total vessel round trips are expected to occur during the Momentum Wind construction campaign (2026), and approximately 1,587 total vessel round trips are expected to occur during the Future Development construction campaign (2027). Vessels would remain on site during construction activities each year to reduce the number of transits between the Project Area and ports.

For operations and maintenance, US Wind anticipates that up to 10 vessels could be used, although not all vessels would operate at the same time or every year. A fall pipe vessel, jack-up vessel, and multi-role survey vessel only be used for non-routine maintenance activities (table 5). Crew transfer vessels would not be likely to operate on a daily basis year-round, however, to be conservative, US Wind assumed that these vessels would operate on a daily basis (table 5).

While a vessel strike could cause injury or mortality of a marine mammal, NMFS is proposing to require extensive vessel strike avoidance measures that would avoid vessel strikes from occurring (see Proposed Mitigation section). US Wind has not requested, and NMFS is not proposing to authorize, take from vessel strikes.

Fisheries and benthic monitoring surveys are being designed for the project in collaboration with UMCES. UMCES and US Wind would conduct pot surveys and recreational fishing surveys focusing on evaluating the extent that commercial and recreational fisheries would be impacted due to changes in black sea bass aggregation behaviors during and after Project construction activities. The program includes a trial baseline year to test deployments and collect baseline data in the Project Area as well as a data synthesis year before construction activities would begin. UMCES and US Wind would conduct additional passive acoustic monitoring research for marine mammals.

Pot surveys offshore Ocean City would be conducted monthly from March through November using ropeless fishing gear to collect data on black sea bass relative abundance in the vicinity of the proposed turbine areas. Catches and sizes of other fauna would be assessed as well. US Wind would set strings of 15 pots (six strings, up to 90 pots total) from a commercial fishing vessel, each string with a 1-day duration set period. EdgeTech ropeless gear (EdgeTech, 2023) would allow sets (trawls) of 15 pots without any rope in the water column. Approximately 300-355 m (984-1,165 ft) of 7/16 inch (in) main-line rope would lie on the bottom during the survey. There would also be approximately 1.5 m of 7/16 in line that would form the bridle connecting each pot to the main line. Each string of pots would consist of 15 black sea bass pots, an EdgeTech pot, and an anchor. The EdgeTech pot would be the release pot attached at the end of each trawl. Each survey would consist of six strings deployed for a 1-day soak time (see diagram in Proposed Rule Comment Responses Memo, October 12, 2023). After the 1-day set period, UMCES and US Wind would retrieve the pot trawls by sending a release command from the on-site research vessel to activate an acoustic release on the release pot. Upon activation, the flotation with the attached rope would ascend to the water surface. UMCES and US Wind would recover the floatation connected to the release pot as well as the rest of the pots for that trawl. The pot survey would be conducted under a NMFS Scientific LOA for black sea bass collection research, of which a similar letter was received by UMCES from NMFS Greater Atlantic Regional Fisheries Office (GARFO) for the initial trial baseline year.

UMCES and US Wind would operate the recreational fishing survey off a recreational charter vessel based in Ocean City to compare data on black sea bass and other fauna between two artificial reef/wreck sites and two turbine sites using a Before-After-Control-Impact (BACI) study design. Angling techniques, such as drop bottom fishing and jigging, would be used to collect catch data on black sea bass and other fauna. Six monthly recreational surveys spanning a 2-day window each, would be conducted annually from May through October.

Passive acoustic monitoring research would focus on using rockhopper recorders to determine occurrence and position of large whales and dolphins as well as F-POD (full waveform capture Pod) devices to detect tonal echolocation clicks of small cetaceans in the Lease Area. The goal of the research would be to distinguish changes in marine mammal behavior due to natural inter-annual variation versus behaviors influenced by wind facility operations. US Wind and UMCES would use a before-during-after gradient design involving 2 years of monitoring in each period before, during, and after Project construction, from 2023 to 2029. The Rockhopper recorder would sample at 200 kHz for baleen whales and dolphins while the F-POD would detect echolocation clicks of small cetaceans. Rockhopper recorders would include a localization array with the Lease Area to allow the positions of calling North Atlantic right whales, humpback whales, and dolphins to be detected. Innovasea receivers would also be attached at up to four mooring sites within the Lease Area to examine spatiotemporal patterns of previously tagged fish, such as Atlantic sturgeon, white sharks, and sand tiger sharks.

Given the gear used (ropeless pot and hook and line), the fishery surveys present little risk to marine mammals (although some hook and line entanglement has been documented in marine mammals). To further minimize this already low risk of interaction, US Wind has proposed, and NMFS has included in the proposed rule, mitigation and monitoring measures to avoid taking marine mammals, including, but not limited to, monitoring for marine mammals before and during fishing/survey activities, not deploying, pulling gear, or fishing in certain circumstances, limiting tow times, and fully repairing nets and lines. All vessel captains and crew would also abide by the vessel strike avoidance measures outlined in § 217.344(b) of this rule. A full description of mitigation measures can be found in the Proposed Mitigation section.

With the implementation of these measures, US Wind does not anticipate, and NMFS is not proposing to authorize, take of marine mammals incidental to research pot and recreational surveys. Given no take is anticipated from these surveys, impacts from fishery surveys will not be discussed further in this document (with the exception of the description of measures in the Proposed Mitigation section).

Thirty-eight marine mammal species under NMFS' jurisdiction have geographic ranges within the western North Atlantic OCS (Hayes et al., 2023). However, for reasons described below, US Wind has requested, and NMFS proposes to authorize, take of only 19 species (comprising 20 stocks) of marine mammals. Sections 3 and 4 of US Wind's ITA application summarize available information regarding status and trends, distribution and habitat preferences, and behavior and life history of the potentially affected species. NMFS fully considered all of this information, and we refer the reader to these descriptions in the application instead of reprinting the information.

Additional information regarding population trends and threats may be found in NMFS' Stock Assessment Reports (SARs; https://www.fisheries.noaa.gov/national/marine-mammal-protection/draft-marine-mammal-stock-assessment-reports ) and more general information about these species ( e.g., physical and behavioral descriptions) may be found on NMFS' website ( https://www.fisheries.noaa.gov/find-species ).

Of the 38 marine mammal species and/or stocks with geographic ranges that include the Project Area ( i.e., found in the coastal and offshore waters of Maryland), 19 species are not expected to be present or are considered rare or unexpected in the Project Area based on sighting and distribution data (see table 3-1 in US Wind's ITA application). Specifically, the following cetacean species are known to occur off of Maryland but are not expected to occur in the Project Area due to the location of preferred habitat outside the Lease Area and ECCs, based on the best available information, and therefore US Wind did not request, and NMFS is not proposing to authorize take, of these species: Blue whale ( Balaenoptera musculus ), Cuvier's beaked whale ( Ziphius cavirostris ), four species of Mesoplodont beaked whales ( Mesoplodon densitostris, M. europaeus, M. mirus, and M. bidens ), Atlantic white-sided dolphin ( Lagenorhynchus acutus ), Clymene dolphin ( Stenella clymene ), dwarf sperm whale ( Kogia sima ), false killer whale ( Pseudorca crassidens ), Fraser's dolphin ( Lagenodelphis hosei ), melon-headed whale ( Peponocephala electra ), northern bottlenose whale ( Hyperoodon ampullatus ), pygmy killer whale ( Feresa attenuata ), pygmy sperm whale ( Kogia breviceps ), sperm whale ( Physeter macrocephalus ), spinner dolphin ( Stenella longirostris ), and white-beaked dolphin ( Lagenorhynchus albirostris ). Two species of phocid pinnipeds are also uncommon in the Project Area, including: harp seals ( Pagophilus groenlandica ) and hooded seals ( Cystophora cristata ). However, harp seals are known to strand in coastal Maryland. Therefore, NMFS is proposing to authorize take of harp seals.

In addition, the Florida manatee ( Trichechus manatus, a sub-species of the West Indian manatee) has been previously documented as an occasional visitor to the Mid-Atlantic region during summer months (Morgan et al., 2002; Cummings et al., 2014). However, manatees are managed by the U.S. Fish and Wildlife Service (USFWS) and are not considered further in this document.

Table 6 lists all species or stocks for which take is expected and proposed to be authorized for this action and summarizes information related to the population or stock, including regulatory status under the MMPA and Endangered Species Act (ESA) and potential biological removal (PBR), where known. PBR is defined as “the maximum number of animals, not including natural mortalities, that may be removed from a marine mammal stock while allowing that stock to reach or maintain its optimum sustainable population” (16 U.S.C. 1362(20)). While no mortality is anticipated or proposed to be authorized, PBR and annual serious injury and mortality from anthropogenic sources are included here as gross indicators of the status of the species or stocks and other threats. Take for 19 species (20 stocks) in table 6 is expected and proposed to be authorized for this activity.

Marine mammal abundance estimates presented in this document represent the total number of individuals that make up a given stock, or the total number estimated within a particular study or survey area. NMFS' stock abundance estimates for most species represent the total estimate of individuals within the geographic area, if known, that comprises that stock. For some species, this geographic area may extend beyond U.S. waters. All managed stocks in this region are assessed in NMFS' U.S. Atlantic and Gulf of Mexico SARs. All values presented in table 6 are the most recent available at the time of publication and, unless noted otherwise, use NMFS' final 2022 SARs (Hayes et al., 2023) available online at https://www.fisheries.noaa.gov/national/marine-mammal-protection/draft-marine-mammal-stock-assessment-reports .

As indicated above, all 19 species and 20 stocks in table 6 temporally and spatially co-occur with the activity to the degree that take is reasonably likely to occur. Three of the marine mammal species for which take is requested are listed as endangered under the ESA, including North Atlantic right, fin, and sei whales. In addition to what is included in sections 3 and 4 of US Wind's ITA application ( https://www.fisheries.noaa.gov/action/incidental-take-authorization-us-wind-inc-construction-and-operation-maryland-offshore-wind ), the SARs ( https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments ), and NMFS' website ( https://www.fisheries.noaa.gov/species-directory/marine-mammals ), we provide further detail below informing the baseline for select species ( e.g., information regarding current UME and known important habitat areas, such as Biologically Important Areas (BIAs; https://oceannoise.noaa.gov/biologically-important-areas ) (Van Parijs, 2015)). There are no ESA-designated critical habitats for any species within the project area ( https://www.fisheries.noaa.gov/resource/map/national-esa-critical-habitat-mapper ).

Under the MMPA, a UME is defined as “a stranding that is unexpected; involves a significant die-off of any marine mammal population; and demands immediate response” (16 U.S.C. 1421h(6)). As of July 2023, five UMEs are active. Four of these UMEs are occurring along the U.S. Atlantic coast for various marine mammal species. Of these, the most relevant to the project area are the North Atlantic right whale, humpback whale, and harbor and gray seal UMEs given the prevalence of these species in the project area. More information on UMEs, including all active, closed, or pending, can be found on NMFS' website at https://www.fisheries.noaa.gov/national/marine-life-distress/active-and-closed-unusual-mortality-events.

Below, we include information for a subset of the species that presently have an active or recently closed UME occurring along the Atlantic coast or for which there is information available related to areas of biological significance. For the majority of species potentially present in the specific geographic region, NMFS has designated only a single generic stock ( e.g., “western North Atlantic”) for management purposes. This includes the “Canadian east coast” stock of minke whales, which includes all minke whales found in U.S. waters and is also a generic stock for management purposes. For humpback and sei whales, NMFS defines stocks on the basis of feeding locations ( i.e., Gulf of Maine and Nova Scotia, respectively). However, references to humpback whales and sei whales in this document refer to any individuals of the species that are found in the project area. Any areas of known biological importance (including the BIAs identified in LaBrecque et al., 2015) that overlap spatially (or are adjacent) with the project area are addressed in the species sections below.

The North Atlantic right whale has been listed as Endangered since the ESA's enactment in 1973. The species was recently uplisted from Endangered to Critically Endangered on the International Union for Conservation of Nature (IUCN) Red List of Threatened Species (Cooke, 2020). The uplisting was due to a decrease in population size (Pace et al., 2017), an increase in vessel strikes and entanglements in fixed fishing gear (Daoust et al., 2017; Davis & Brillant, 2019; Knowlton et al., 2012; Knowlton et al., 2022; Moore et al., 2021; Sharp et al., 2019), and a decrease in birth rate (Pettis et al., 2022; Reed et al., 2022). The western Atlantic stock is considered depleted under the MMPA (Hayes et al., 2023). There is a recovery plan (NMFS, 2005) for the North Atlantic right whale, and NMFS completed 5-year reviews of the species in 2012, 2017, and 2022 which concluded no change to the listing status is warranted.

Designated by NMFS as a Species in the Spotlight, the North Atlantic right whale is considered among the species with the greatest risk of extinction in the near future ( https://www.fisheries.noaa.gov/topic/endangered-species-conservation/species-in-the-spotlight ).

The North Atlantic right whale population had only a 2.8-percent recovery rate between 1990 and 2011 and an overall abundance decline of 23.5 percent from 2011 to 2019 (Hayes et al., 2023). Since 2011, the North Atlantic right whale population has been in decline; however, the sharp decrease observed from 2015 to 2020 appears to have slowed, though the right whale population continues to experience annual mortalities above recovery thresholds (Pace et al., 2017; Pace et al., 2021; Linden, 2023). North Atlantic right whale calving rates dropped from 2017 to 2020 with zero births recorded during the 2017-2018 season. The 2020-2021 calving season had the first substantial calving increase in 5 years with 20 calves born (including 2 mortalities) followed by 15 calves during the 2021-2022 calving season and 12 births (including 1 mortality) in 2022-2023 calving season. These data demonstrate that birth rates are increasing. However, mortalities continue to outpace births. Best estimates indicate fewer than 70 reproductively active females remain in the population and adult females experience a lower average survival rate than males (Linden, 2023). In 2023, the total annual average observed North Atlantic right whale mortality increased from 8.1 (which represents 2016-2020) to 31.2 (which represents 2015-2019), however, this updated estimate also accounts for undetected mortality and serious injury (Hayes et al., 2023). Although the predicted number of deaths from the population are lower in recent years (2021-2022) when compared to the high number of deaths from 2014 to 2020 suggesting a short-term increase in survival, annual mortality rates still exceed PBR (Linden, 2023).

Critical habitat for North Atlantic right whales is not present in the Project Area. However, the Project Area both spatially and temporally overlaps a portion of the migratory corridor BIA within which North Atlantic right whales migrate south to calving grounds generally in November and December, followed by a northward migration (primarily moms with young calves) into feeding areas far north of the Project Area in March and April (LaBrecque et al., 2015; Van Parijs, 2015). North Atlantic right whale foraging may rarely opportunistically occur around the Project Area, yet the region is not considered primary foraging habitat. Engelhaupt et al. (2023) documented feeding and socializing behavior off Virginia and North Carolina, just south of the Project Area, suggesting that North Atlantic right whales may use the mid-Atlantic migratory corridor for more than just migration.

NMFS' regulations at 50 CFR 224.105 designated Seasonal Management Areas (SMAs) for North Atlantic right whales in 2008 (73 FR 60173, October 10, 2008). SMAs were developed to reduce the threat of collisions between ships and North Atlantic right whales around their migratory route and calving grounds. The Delaware Bay SMA overlaps with the export cable corridor of the proposed project. This SMA is currently active from November 1 through April 30 of each year and may be used by North Atlantic right whales for migrating and/or feeding. As noted above, NMFS is proposing changes to the North Atlantic right whale speed rule (87 FR 46921, August 1, 2022). Due to the current status of North Atlantic right whales and the spatial proximity overlap of the proposed project with areas of biological significance, ( i.e., a migratory corridor, SMA), the potential impacts of the proposed project on North Atlantic right whales warrant particular attention.

During the spring, North Atlantic right whales use the migratory corridor BIA to move north from calving grounds off Georgia and Florida to feeding grounds in New England and Canadian waters (Hayes et al., 2023). Right whales feed primarily on the copepod, Calanus finmarchicus, a species whose availability and distribution has changed both spatially and temporally over the last decade due to an oceanographic regime shift that has been ultimately linked to climate change (Meyer-Gutbrod et al., 2021; Record et al., 2019; Sorochan et al., 2019). This distribution change in prey availability has led to shifts in right whale habitat-use patterns over the same time period (Davis et al., 2020; Meyer-Gutbrod et al., 2022; Quintano-Rizzo et al., 2021; O'Brien et al., 2022; Van Parijs et al., 2023) with reduced use of foraging habitats in the Great South Channel and Bay of Fundy and increased use of habitats within Cape Cod Bay and a region south of Martha's Vineyard and Nantucket Islands (Stone et al., 2017; Mayo et al., 2018; Ganley et al., 2019; Record et al., 2019; Meyer-Gutbrod et al., 2021; Van Parijs et al., 2023); these foraging habitats are all located several hundred kilometers north of the project area. In late fall ( i.e., November), a portion of the right whale population (including pregnant females) typically departs the feeding grounds in the North Atlantic, moves south along the migratory corridor BIA, including through the Project Area, to right whale calving grounds off Georgia and Florida. Observations of these transitions in right whale habitat use, variability in seasonal presence in identified core habitats, and utilization of habitat outside of previously focused survey effort prompted the formation of a NMFS' Expert Working Group, which identified current data collection efforts, data gaps, and provided recommendations for future survey and research efforts (Oleson et al., 2020). Recent research indicates understanding of their movement patterns remains incomplete and not all of the population undergoes a consistent annual migration (Davis et al., 2017; Gowan et al., 2019; Krzystan et al., 2018). Non-calving females may remain in the feeding grounds, during the winter in the years preceding and following the birth of a calf to increase their energy stores (Gowen et al., 2019).

Although North Atlantic right whales move seasonally between foraging and calving grounds, Davis et al. (2017) acoustically detected right whales along the coast from Cape Hatteras, NC, United States to Nova Scotia, Canada year-round, suggesting that North Atlantic right whale use of the mid-Atlantic and southeast has increased since 2010 (Davis et al., 2017). North Atlantic right whale presence in the Project Area is predominately seasonal with individuals likely to be transient and migrating through the area. Bailey et al. (2018) acoustically detected the year-round presence of North Atlantic right whales in the vicinity of the Project Area, with a maximum abundance during the late winter and early spring. In addition, a monitoring buoy, deployed by UMCES offshore of Ocean City Maryland in 2022, acoustically detected the presence of North Atlantic right whales in the lease area from November through January, with the highest frequency of confirmed detections occurring during the months of December and January (Woods Hole Oceanographic Institute, 2022). Visual surveys also confirm a maximum abundance of North Atlantic right whales in the vicinity of the Lease Area during the winter (Barco et al., 2015; Williams et al., 2015). As part of the Mid-Atlantic Baseline Studies Project and Maryland Project, Williams et al. (2015) conducted standardized aerial and boat-based surveys of the Delaware, Maryland, Virginia Wind Energy Areas (WEAs), and visually observed North Atlantic right whales in the lease area during the months of February and March. Based upon year-round aerial surveys conducted from 2013 to 2015, Barco et al. (2015) observed the largest numbers of North Atlantic right whales in the Maryland WEA during the month of January, suggesting that the area may be a destination for non-breeding individuals and pulses of North Atlantic right whales may travel through the region. Barco et al. (2015) also documented North Atlantic right whale open mouth behavior, which is consistent with, though not necessarily indicative of, feeding. As part of the U.S. Navy's Marine Species Monitoring Program, HDR has conducted aerial and vessel-based surveys for large whales off Virginia and North Carolina since 2015. The majority of North Atlantic right whale sightings have occurred in these areas, just south of the Project Area, during the months of January-March (Aschettino et al., 2023). The highest density month for North Atlantic right whales in the vicinity of the lease area is February (0.00076 individuals/km (0.54 nmi grid square)) (Roberts et al., 2023).

Since 2017, 98 dead, seriously injured, or sublethally injured or ill North Atlantic right whales along the United States and Canadian coasts have been documented, necessitating a UME declaration and investigation. The leading category for the cause of death for this ongoing UME is “human interaction,” specifically from entanglements or vessel strikes. As of October 30, 2023, there have been 36 confirmed mortalities (dead, stranded, or floaters) and 34 seriously injured free-swimming whales for a total of 70 whales. Beginning on October 14, 2022, the UME also considers animals with sublethal injury or illness bringing the total number of whales in the UME to 115. Approximately 42 percent of the population is known to be in reduced health (Hamilton et al., 2021) likely contributing to smaller body sizes at maturation, making them more susceptible to threats and reducing fecundity (Moore et al., 2021; Reed et al., 2022; Stewart et al., 2022). More information about the North Atlantic right whale UME is available online at https://www.fisheries.noaa.gov/national/marine-life-distress/2017-2023-north-atlantic-right-whale-unusual-mortality-event.

Humpback whales were listed as endangered under the Endangered Species Conservation Act (ESCA) in June 1970. In 1973, the ESA replaced the ESCA, and humpbacks continued to be listed as endangered. On September 8, 2016, NMFS divided the once single species into 14 distinct population segments (DPS), removed the species-level listing, and, in its place, listed four DPSs as endangered and one DPS as threatened (81 FR 62259, September 8, 2016). The remaining nine DPSs were not listed. The West Indies DPS, which is not listed under the ESA, is the only DPS of humpback whales that is expected to occur in the Project Area. Bettridge et al. (2015) estimated the size of the West Indies DPS population at 12,312 (95 percent confidence interval (CI) 8,688-15,954) whales in 2004-2005, which is consistent with previous population estimates of approximately 10,000-11,000 whales (Stevick et al., 2003; Smith et al., 1999) and the increasing trend for the West Indies DPS (Bettridge et al., 2015).

The Project Area does not overlap with any BIAs or other important areas for the humpback whales. A humpback whale feeding BIA extends throughout the Gulf of Maine, Stellwagen Bank, and Great South Channel from May through December, annually (LaBrecque et al., 2015). However, this BIA is located approximately 556.2 km (345.6 mi) north of the Project Area, and thus, would not be impacted by project activities.

Humpback whale presence in the mid-Atlantic varies seasonally. Humpback whales are most typically observed in this region during the winter months (Williams et al., 2015d; Barco et al., 2015) and are known to be migratory off coastal Maryland, moving seasonally between northern feeding grounds in New England and southern calving grounds in the West Indies (Hayes et al., 2023). However, not all humpback whales migrate to the Caribbean during the winter as individuals are sighted in mid- to high-latitude areas during this season (Swingle et al., 1993; Davis et al., 2020). In addition to a migratory pathway, the mid-Atlantic region also represents a supplemental winter feeding ground for juveniles and mature whales (Barco et al., 2002). Records of humpback whales off the U.S. mid-Atlantic coast (New Jersey south to North Carolina) suggest that these waters are used as a winter feeding ground from December through March (Mallette et al., 2017; Barco et al., 2002; LaBrecque et al., 2015) and represent important habitat for juveniles, in particular (Swingle et al., 1993; Wiley et al., 1995).

Acoustic monitoring in the vicinity of the lease area has detected the presence of humpback whales year-round, although detections exhibit similar seasonal trends as visual sightings. Humpback whale detections were lowest during the summer months (June through September), increased through the winter (January through March) and peaked in April (Bailey et al., 2018). Davis et al. (2020) also found detections of humpback whales off the mid-Atlantic (Virginia) to peak from January through May. Density modeling (Roberts et al., 2023) confirms April (0.00187 individuals per 1 km (0.54 nmi) grid cell) as the month of the highest average density of humpback whales in the vicinity of the Project Area.

Since January 2016, elevated humpback whale mortalities along the Atlantic coast from Maine to Florida led to the declaration of a UME. As of October 2, 2023, 209 humpback whales have stranded as part of this UME. Partial or full necropsy examinations have been conducted on approximately 90 of the known cases. Of the whales examined, about 40 percent had evidence of human interaction, either ship strike or entanglement. While a portion of the whales have shown evidence of pre-mortem vessel strike, this finding is not consistent across all whales examined and more research is needed. As the humpback whale population has grown, they are seen more often in the mid-Atlantic. Since January 2023, 34 humpbacks have stranded along the east coast of the United States (1 of these stranded in Maryland). These whales may have been following their prey (small fish) which were reportedly close to shore this past winter. These prey also attract fish that are targeted by recreational and commercial fishermen, which increases the number of boats in these areas. More information is available at https://www.fisheries.noaa.gov/national/marine-life-distress/active-and-closed-unusual-mortality-events.

Fin whales frequently occur in the waters of the U.S. Atlantic Exclusive Economic Zone (EEZ), principally from Cape Hatteras, North Carolina northward and are distributed in both continental shelf and deep-water habitats (Hayes et al., 2023). Although fin whales are present north of the 35-degree latitude region in every season and are broadly distributed throughout the western North Atlantic for most of the year, densities vary seasonally (Edwards et al., 2015; Hayes et al., 2023). Fin whales typically feed in the Gulf of Maine and the waters surrounding New England, but their mating and calving (and general wintering) areas are largely unknown (Hain et al., 1992; Hayes et al., 2023). Acoustic detections of fin whale singers augment and confirm these visual sighting conclusions for males. Recordings from Massachusetts Bay, New York Bight, and deep-ocean areas have detected some level of fin whale singing from September through June (Watkins et al., 1987; Clark and Gagnon, 2002; Morano et al., 2012). These acoustic observations from both coastal and deep-ocean regions support the conclusion that male fin whales are broadly distributed throughout the western North Atlantic for most of the year (Hayes et al., 2022).

Fin whale feeding BIAs occur offshore of Montauk Point, New York from March to October (2,933 km 2 ) (Hain et al., 1992; LaBrecque et al., 2015) and year-round in the southern Gulf of Maine (18,015 km 2 ). However, given the more southerly location of the Project Area (located approximately 364.8 km (226.7 mi) and 546.2 km (339.4 mi) away from these BIAs, respectively), there is no spatial overlap from with these BIAs.

Fin whales were among the most frequently observed baleen whale species during the Maryland Wind Energy Area aerial surveys conducted for the Maryland Department of Natural Resources (MD DNR) by the Virginia Aquarium and Marine Science Center Foundation (Barco et al., 2015), and the most commonly detected baleen whale species during acoustic monitoring surveys from 2014 to 2017 in the Maryland WEA, although the majority of detections were offshore of the WEA (Bailey et al., 2018a). Fin whale abundance in the vicinity of the Project Area peaked during the winter and early spring (Williams et al., 2015d; Barco et al., 2015), with the lowest occurrence documented during summer and early fall (Bailey et al., 2018). Consistent with visual sightings and acoustic detections, the highest average density of fin whales in the vicinity of the proposed Project Area occurs in January (0.00214 individuals per 1 km (0.54 nmi) grid cell) (Roberts et al., 2023). There is no active fin whale UME.

Minke whales are common and widely distributed throughout the U.S. Atlantic EEZ (Cetacean and Turtle Assessment Program (CETAP), 1982; Hayes et al., 2022), although their distribution has a strong seasonal component. Individuals have often been detected acoustically in shelf waters from spring to fall and more often detected in deeper offshore waters from winter to spring (Risch et al., 2013). Minke whales are abundant in New England waters from May through September (Pittman et al., 2006; Waring et al., 2014), yet largely absent from these areas during the winter, suggesting the possible existence of a migratory corridor (LaBrecque et al., 2015). A migratory route for minke whales transiting between northern feeding grounds and southern breeding areas may exist to the east of the Project Area, as minke whales may track warmer waters along the continental shelf while migrating (Risch et al., 2014). Risch et al. (2014) suggests the presence of a minke whale breeding ground offshore of the southeastern US during the winter.

There are two minke whale feeding BIAs identified in the southern and southwestern section of the Gulf of Maine, including Georges Bank, the Great South Channel, Cape Cod Bay and Massachusetts Bay, Stellwagen Bank, Cape Anne, and Jeffreys Ledge from March through November, annually (LaBrecque et al., 2015). However, these BIAs are approximately 512.1 km (318.2 mi) and 668.8 km (415.6 mi) northwest of the Project Area, respectively, and would not be impacted by the proposed project activities.

Overall, minke whale use of the Project Area is likely highest during fall, winter, and spring months based upon visual sightings and acoustic detections in the vicinity of the lease area during the months of November, January, February, and April (Bailey et al., 2018a; Barco et al., 2015; Williams et al., 2015b). The highest average density of minke whales in the vicinity of the lease area is expected to occur in May (0.00750 individuals per 1 km (0.54 nmi)).

From 2017 through 2022, elevated minke whale mortalities detected along the Atlantic coast from Maine through South Carolina resulted in the declaration of a UME. As of October 2, 2023, a total of 160 minke whale mortalities have occurred during this UME. Full or partial necropsy examinations were conducted on more than 60 percent of the whales. Preliminary findings in several of the whales have shown evidence of human interactions or infectious disease, but these findings are not consistent across all of the minke whales examined, so more research is needed. More information is available at https://www.fisheries.noaa.gov/national/marine-life-distress/2017-2022-minke-whale-unusual-mortality-event-along-atlantic-coast.

The Nova Scotia stock of sei whales can be found in deeper waters of the continental shelf edge of the eastern United States and northeastward to south of Newfoundland (Mitchell, 1975; Hain et al., 1985; Hayes et al., 2022). During spring and summer, the stock is mainly concentrated in northern feeding areas, including the Scotian Shelf (Mitchell and Chapman, 1977), the Gulf of Maine, Georges Bank, the Northeast Channel, and south of Nantucket (CETAP, 1982; Kraus et al., 2016; Roberts et al., 2016; Palka et al., 2017; Cholewiak et al., 2018; Hayes et al., 2022). Sei whales have been detected acoustically along the Atlantic Continental Shelf and Slope from south of Cape Hatteras, North Carolina to the Davis Strait, with acoustic occurrence increasing in the mid-Atlantic region since 2010 (Davis et al., 2020). Although their migratory movements are not well understood, sei whales are believed to migrate north in June and July to feeding areas and south in September and October to breeding areas (Mitchell, 1975; CETAP, 1982; Davis et al., 2020). Sei whales generally occur offshore; however, individuals may also move into shallower, more inshore waters (Payne et al., 1990; Halpin et al., 2009; Hayes et al., 2022).

A sei whale feeding BIA occurs in New England waters from May through November (LaBrecque et al., 2015). However, this BIA is located approximately 501.5 km (311.6 mi) north of the Project Area and not likely to be impacted by the Project activities.

Sei whales were sighted infrequently during visual surveys (Williams et al., 2015d) and acoustic monitoring (WHOI, 2022; WHOI, 2023) of the Maryland WEA. The highest average density of sei whales in the vicinity of the lease area is expected to occur during the month of April (0.00061 individuals per 1 km (0.54 nmi) (Roberts et al., 2023). There is no active sei whale UME.

Since June 2022, elevated numbers of harbor seal and gray seal mortalities have occurred across the southern and central coast of Maine. This event has been declared a UME. Preliminary testing of samples has found some harbor and gray seals positive for highly pathogenic avian influenza. While the UME is not occurring in the Project Area, the populations affected by the UME are the same as those potentially affected by the project. Information on this UME is available online at https://www.fisheries.noaa.gov/2022-2023-pinniped-unusual-mortality-event-along-maine-coast.

The above event was preceded by a different UME, occurring from 2018 to 2020 (closure of the 2018-2020 UME is pending). Beginning in July 2018, elevated numbers of harbor seal and gray seal mortalities occurred across Maine, New Hampshire, and Massachusetts. Additionally, stranded seals have shown clinical signs as far south as Virginia, although not in elevated numbers, therefore the UME investigation encompassed all seal strandings from Maine to Virginia. A total of 3,152 reported strandings (of all species) occurred from July 1, 2018, through March 13, 2020. Full or partial necropsy examinations have been conducted on some of the seals and samples have been collected for testing. Based on tests conducted thus far, the main pathogen found in the seals is phocine distemper virus. NMFS is performing additional testing to identify any other factors that may be involved in this UME, which is pending closure. Information on this UME is available online at: https://www.fisheries.noaa.gov/new-england-mid-atlantic/marine-life-distress/2018-2020-pinniped-unusual-mortality-event-along.

Hearing is the most important sensory modality for marine mammals underwater, and exposure to anthropogenic sound can have deleterious effects. To appropriately assess the potential effects of exposure to sound, it is necessary to understand the frequency ranges marine mammals are able to hear. Not all marine mammal species have equal hearing capabilities ( e.g., Richardson et al., 1995; Wartzok and Ketten, 1999; Au and Hastings, 2008). To reflect this, Southall et al. (2007, 2019a) recommended that marine mammals be divided into hearing groups based on directly measured (behavioral or auditory evoked potential techniques) or estimated hearing ranges (behavioral response data, anatomical modeling, etc. ). Note that no direct measurements of hearing ability have been successfully completed for mysticetes ( i.e., low-frequency cetaceans). Subsequently, NMFS (2018) described generalized hearing ranges for these marine mammal hearing groups. Generalized hearing ranges were chosen based on the approximately 65-decibel (dB) threshold from the normalized composite audiograms, with the exception for lower limits for low-frequency cetaceans where the lower bound was deemed to be biologically implausible and the lower bound from Southall et al. (2007) retained. Marine mammal hearing groups and their associated hearing ranges are provided in table 7.

The pinniped functional hearing group was modified from Southall et al. (2007) on the basis of data indicating that phocid species have consistently demonstrated an extended frequency range of hearing compared to otariids, especially in the higher frequency range (Hemilä et al., 2006; Kastelein et al., 2009; Reichmuth and Holt, 2013). For more detail concerning these groups and associated frequency ranges, please see NMFS (2018) for a review of available information.

NMFS notes that in 2019a, Southall et al. recommended new names for hearing groups that are widely recognized. However, this new hearing group classification does not change the weighting functions or acoustic thresholds ( i.e., the weighting functions and thresholds in Southall et al. (2019a) are identical to NMFS 2018 Revised Technical Guidance). When NMFS updates our Technical Guidance, we will be adopting the updated Southall et al. (2019a) hearing group classification.

This section includes a summary and discussion of the ways that components of the specified activity may impact marine mammals and their habitat. The Estimated Take of Marine Mammals section later in this document includes a quantitative analysis of the number of individuals that are expected to be taken by this activity. The Negligible Impact Analysis and Determination section considers the content of this section, the Estimated Take of Marine Mammals section, and the Proposed Mitigation section, to draw conclusions regarding the likely impacts of these activities on the reproductive success or survivorship of individuals and how those impacts on individuals are likely to impact marine mammal species or stocks. General background information on marine mammal hearing was provided previously (see the Description of Marine Mammals in the Geographic Area section). Here, the potential effects of sound on marine mammals are discussed.

US Wind has requested, and NMFS proposes to authorize, the take of marine mammals incidental to the construction activities associated with the project area. In their application, US Wind presented their analyses of potential impacts to marine mammals from the acoustic sources. NMFS both carefully reviewed the information provided by US Wind, as well as independently reviewed applicable scientific research and literature and other information to evaluate the potential effects of the Project's activities on marine mammals.

The proposed activities would result in the construction and placement of up to 119 permanent foundations to support WTGs, OSSs, a Met tower, and seafloor mapping using HRG surveys. There are a variety of types and degrees of effects to marine mammals, prey species, and habitat that could occur as a result of the Project. Below we provide a brief description of the types of sound sources that would be generated by the project, the general impacts from these types of activities, and an analysis of the anticipated impacts on marine mammals from the project, with consideration of the proposed mitigation measures.

This section contains a brief technical background on sound, on the characteristics of certain sound types, and on metrics used in this proposal inasmuch as the information is relevant to the specified activity and to a discussion of the potential effects of the specified activity on marine mammals found later in this document. For general information on sound and its interaction with the marine environment, please see: Au and Hastings, 2008; Richardson et al., 1995; Urick, 1983; as well as the Discovery of Sound in the Sea (DOSITS) website at https://www.dosits.org. Sound is a vibration that travels as an acoustic wave through a medium such as a gas, liquid, or solid. Sound waves alternately compress and decompress the medium as the wave travels. These compressions and decompressions are detected as changes in pressure by aquatic life and man-made sound receptors such as hydrophones (underwater microphones). In water, sound waves radiate in a manner similar to ripples on the surface of a pond and may be either directed in a beam (narrow beam or directional sources) or sound beams may radiate in all directions (omnidirectional sources).

Sound travels in water more efficiently than almost any other form of energy, making the use of acoustics ideal for the aquatic environment and its inhabitants. In seawater, sound travels at roughly 1,500 meters per second (m/s). In-air, sound waves travel much more slowly, at about 340 m/s. However, the speed of sound can vary by a small amount based on characteristics of the transmission medium, such as water temperature and salinity. Sound travels in water more efficiently than almost any other form of energy, making the use of acoustics ideal for the aquatic environment and its inhabitants. In seawater, sound travels at roughly 1,500 m/s. In-air, sound waves travel much more slowly, at about 340 m/s. However, the speed of sound can vary by a small amount based on characteristics of the transmission medium, such as water temperature and salinity.

The basic components of a sound wave are frequency, wavelength, velocity, and amplitude. Frequency is the number of pressure waves that pass by a reference point per unit of time and is measured in hertz (Hz) or cycles per second. Wavelength is the distance between two peaks or corresponding points of a sound wave (length of one cycle). Higher frequency sounds have shorter wavelengths than lower frequency sounds, and typically attenuate (decrease) more rapidly, except in certain cases in shallower water.

The intensity (or amplitude) of sounds is measured in dB, which are a relative unit of measurement that is used to express the ratio of one value of a power or field to another. Decibels are measured on a logarithmic scale, so a small change in dB corresponds to large changes in sound pressure. For example, a 10-dB increase is a ten-fold increase in acoustic power. A 20-dB increase is then a hundred-fold increase in power and a 30-dB increase is a thousand-fold increase in power. However, a ten-fold increase in acoustic power does not mean that the sound is perceived as being 10 times louder. Decibels are a relative unit comparing two pressures; therefore, a reference pressure must always be indicated. For underwater sound, this is 1 microPascal (μPa). For in-air sound, the reference pressure is 20 microPascal (μPa). The amplitude of a sound can be presented in various ways; however, NMFS typically considers three metrics. In this proposed rule, all decibel levels are referenced to (re) 1μPa.

Sound exposure level (SEL) represents the total energy in a stated frequency band over a stated time interval or event and considers both amplitude and duration of exposure (represented as dB re 1 μPa 2 -s). SEL is a cumulative metric; it can be accumulated over a single pulse (for pile driving this is often referred to as single-strike SEL; SEL ss ) or calculated over periods containing multiple pulses (SEL cum ). Cumulative SEL represents the total energy accumulated by a receiver over a defined time window or during an event. The SEL metric is useful because it allows sound exposures of different durations to be related to one another in terms of total acoustic energy. The duration of a sound event and the number of pulses, however, should be specified as there is no accepted standard duration over which the summation of energy is measured.

Root mean square (rms) is the quadratic mean sound pressure over the duration of an impulse. Root mean square is calculated by squaring all of the sound amplitudes, averaging the squares, and then taking the square root of the average (Urick, 1983). Root mean square accounts for both positive and negative values; squaring the pressures makes all values positive so that they may be accounted for in the summation of pressure levels (Hastings and Popper, 2005). This measurement is often used in the context of discussing behavioral effects, in part because behavioral effects, which often result from auditory cues, may be better expressed through averaged units than by peak pressures.

Peak sound pressure (also referred to as zero-to-peak sound pressure or 0-pk) is the maximum instantaneous sound pressure measurable in the water at a specified distance from the source and is represented in the same units as the rms sound pressure. Along with SEL, this metric is used in evaluating the potential for PTS (permanent threshold shift) and TTS (temporary threshold shift).

Sounds can be either impulsive or non-impulsive. The distinction between these two sound types is important because they have differing potential to cause physical effects, particularly with regard to hearing ( e.g., Ward, 1997 in Southall et al., 2007). Please see NMFS et al. (2018) and Southall et al. (2007, 2019a) for an in-depth discussion of these concepts. Impulsive sound sources ( e.g., airguns, explosions, gunshots, sonic booms, impact pile driving) produce signals that are brief (typically considered to be less than 1 second), broadband, atonal transients (American National Standards Institute (ANSI), 1986; ANSI, 2005; Harris, 1998; National Institute for Occupational Safety and Health (NIOSH), 1998; International Organization for Standardization (ISO), 2003) and occur either as isolated events or repeated in some succession. Impulsive sounds are all characterized by a relatively rapid rise from ambient pressure to a maximal pressure value followed by a rapid decay period that may include a period of diminishing, oscillating maximal and minimal pressures, and generally have an increased capacity to induce physical injury as compared with sounds that lack these features. Impulsive sounds are typically intermittent in nature.

Non-impulsive sounds can be tonal, narrowband, or broadband, brief, or prolonged, and may be either continuous or intermittent (ANSI, 1995; NIOSH, 1998). Some of these non-impulsive sounds can be transient signals of short duration but without the essential properties of pulses ( e.g., rapid rise time). Examples of non-impulsive sounds include those produced by vessels, aircraft, machinery operations such as drilling or dredging, vibratory pile driving, and active sonar systems. Sounds are also characterized by their temporal component. Continuous sounds are those whose sound pressure level remains above that of the ambient sound with negligibly small fluctuations in level (NIOSH, 1998; ANSI, 2005) while intermittent sounds are defined as sounds with interrupted levels of low or no sound (NIOSH, 1998). NMFS identifies Level B harassment thresholds based on if a sound is continuous or intermittent.

Even in the absence of sound from the specified activity, the underwater environment is typically loud due to ambient sound, which is defined as environmental background sound levels lacking a single source or point (Richardson et al., 1995). The sound level of a region is defined by the total acoustical energy being generated by known and unknown sources. These sources may include physical ( e.g., wind and waves, earthquakes, ice, atmospheric sound), biological ( e.g., sounds produced by marine mammals, fish, and invertebrates), and anthropogenic ( e.g., vessels, dredging, construction) sound. A number of sources contribute to ambient sound, including wind and waves, which are a main source of naturally occurring ambient sound for frequencies between 200 Hz and 50 kHz (International Council for the Exploration of the Sea (ICES), 1995). In general, ambient sound levels tend to increase with increasing wind speed and wave height. Precipitation can become an important component of total sound at frequencies above 500 Hz and possibly down to 100 Hz during quiet times. Marine mammals can contribute significantly to ambient sound levels as can some fish and snapping shrimp. The frequency band for biological contributions is from approximately 12 Hz to over 100 kHz. Sources of ambient sound related to human activity include transportation (surface vessels), dredging and construction, oil and gas drilling and production, geophysical surveys, sonar, and explosions. Vessel noise typically dominates the total ambient sound for frequencies between 20 and 300 Hz. In general, the frequencies of anthropogenic sounds are below 1 kHz, and if higher frequency sound levels are created, they attenuate rapidly.

The sum of the various natural and anthropogenic sound sources that comprise ambient sound at any given location and time depends not only on the source levels (as determined by current weather conditions and levels of biological and human activity) but also on the ability of sound to propagate through the environment. In turn, sound propagation is dependent on the spatially and temporally varying properties of the water column and sea floor and is frequency-dependent. As a result of the dependence on a large number of varying factors, ambient sound levels can be expected to vary widely over both coarse and fine spatial and temporal scales. Sound levels at a given frequency and location can vary by 10-20 dB from day to day (Richardson et al., 1995). The result is that, depending on the source type and its intensity, sound from the specified activity may be a negligible addition to the local environment or could form a distinctive signal that may affect marine mammals. Human-generated sound is a significant contributor to the acoustic environment in the project location.

Anthropogenic sounds cover a broad range of frequencies and sound levels and can have a range of highly variable impacts on marine life from none or minor to potentially severe responses depending on received levels, duration of exposure, behavioral context, and various other factors. Broadly, underwater sound from active acoustic sources, such as those in the Project, can potentially result in one or more of the following: temporary or permanent hearing impairment, non-auditory physical or physiological effects, behavioral disturbance, stress, and masking (Richardson et al., 1995; Gordon et al., 2003; Nowacek et al., 2007; Southall et al., 2007; Götz et al., 2009). Non-auditory physiological effects or injuries that theoretically might occur in marine mammals exposed to high level underwater sound or as a secondary effect of extreme behavioral reactions ( e.g., change in dive profile as a result of an avoidance reaction) caused by exposure to sound include neurological effects, bubble formation, resonance effects, and other types of organ or tissue damage (Cox et al., 2006; Southall et al., 2007; Zimmer and Tyack, 2007; Tal et al., 2015).

In general, the degree of effect of an acoustic exposure is intrinsically related to the signal characteristics, received level, distance from the source, and duration of the sound exposure, in addition to the contextual factors of the receiver ( e.g., behavioral state at time of exposure, age class, etc. ). In general, sudden, high-level sounds can cause hearing loss as can longer exposures to lower-level sounds. Moreover, any temporary or permanent loss of hearing will occur almost exclusively for noise within an animal's hearing range. We describe below the specific manifestations of acoustic effects that may occur based on the activities proposed by US Wind. Richardson et al. (1995) described zones of increasing intensity of effect that might be expected to occur in relation to distance from a source and assuming that the signal is within an animal's hearing range. First (at the greatest distance) is the area within which the acoustic signal would be audible (potentially perceived) to the animal but not strong enough to elicit any overt behavioral or physiological response. The next zone (closer to the receiving animal) corresponds with the area where the signal is audible to the animal and of sufficient intensity to elicit behavioral or physiological responsiveness. The third is a zone within which, for signals of high intensity, the received level is sufficient to potentially cause discomfort or tissue damage to auditory or other systems. Overlaying these zones to a certain extent is the area within which masking ( i.e., when a sound interferes with or masks the ability of an animal to detect a signal of interest that is above the absolute hearing threshold) may occur; the masking zone may be highly variable in size.

Below, we provide additional detail regarding potential impacts on marine mammals and their habitat from noise in general, starting with hearing impairment, as well as from the specific activities US Wind plans to conduct, to the degree it is available (noting that there is limited information regarding the impacts of offshore wind construction on marine mammals).

Marine mammals exposed to high-intensity sound or to lower-intensity sound for prolonged periods can experience hearing threshold shift (TS), which NMFS defines as a change, usually an increase, in the threshold of audibility at a specified frequency or portion of an individual's hearing range above a previously established reference level expressed in decibels (NMFS, 2018). Threshold shifts can be permanent, in which case there is an irreversible increase in the threshold of audibility at a specified frequency or portion of an individual's hearing range or temporary, in which there is reversible increase in the threshold of audibility at a specified frequency or portion of an individual's hearing range and the animal's hearing threshold would fully recover over time (Southall et al., 2019a). Repeated sound exposure that leads to TTS could cause PTS.

When PTS occurs, there can be physical damage to the sound receptors in the ear ( i.e., tissue damage) whereas TTS represents primarily tissue fatigue and is reversible (Henderson et al., 2008). In addition, other investigators have suggested that TTS is within the normal bounds of physiological variability and tolerance and does not represent physical injury ( e.g., Ward, 1997; Southall et al., 2019a). Therefore, NMFS does not consider TTS to constitute auditory injury. Relationships between TTS and PTS thresholds have not been studied in marine mammals, and there is no PTS data for cetaceans. However, such relationships are assumed to be similar to those in humans and other terrestrial mammals. Noise exposure can result in either a permanent shift in hearing thresholds from baseline (a 40-dB threshold shift approximates a PTS onset; e.g., Kryter et al., 1966; Miller, 1974; Henderson et al., 2008) or a temporary, recoverable shift in hearing that returns to baseline (a 6-dB threshold shift approximates a TTS onset; e.g., Southall et al., 2019a). Based on data from terrestrial mammals, a precautionary assumption is that the PTS thresholds, expressed in the unweighted peak sound pressure level metric (PK), for impulsive sounds (such as impact pile driving pulses) are at least 6 dB higher than the TTS thresholds and the weighted PTS cumulative sound exposure level thresholds are 15 (impulsive sound) to 20 (non-impulsive sounds) dB higher than TTS cumulative sound exposure level thresholds (Southall et al., 2019a). Given the higher level of sound or longer exposure duration necessary to cause PTS as compared with TTS, PTS is less likely to occur as a result of these activities; however, it is possible, and a small amount has been proposed for authorization for several species.

TTS is the mildest form of hearing impairment that can occur during exposure to sound, with a TTS of 6 dB considered the minimum threshold shift clearly larger than any day-to-day or session-to-session variation in a subject's normal hearing ability (Schlundt et al., 2000; Finneran et al., 2000; Finneran et al., 2002). While experiencing TTS, the hearing threshold rises, and a sound must be at a higher level in order to be heard. In terrestrial and marine mammals, TTS can last from minutes or hours to days (in cases of strong TTS). In many cases, hearing sensitivity recovers rapidly after exposure to the sound ends. There is data on sound levels and durations necessary to elicit mild TTS for marine mammals, but recovery is complicated to predict and dependent on multiple factors.

Marine mammal hearing plays a critical role in communication with conspecifics, and interpretation of environmental cues for purposes such as predator avoidance and prey capture. Depending on the degree (elevation of threshold in dB), duration ( i.e., recovery time), and frequency range of TTS, and the context in which it is experienced, TTS can have effects on marine mammals ranging from discountable to serious depending on the degree of interference of marine mammals hearing. For example, a marine mammal may be able to readily compensate for a brief, relatively small amount of TTS in a non-critical frequency range that occurs during a time where ambient noise is lower and there are not as many competing sounds present. Alternatively, a larger amount and longer duration of TTS sustained during time when communication is critical ( e.g., for successful mother/calf interactions, consistent detection of prey) could have more serious impacts.

Currently, TTS data only exist for four species of cetaceans (bottlenose dolphin, beluga whale ( Delphinapterus leucas ), harbor porpoise, and Yangtze finless porpoise ( Neophocaena asiaeorientalis )) and six species of pinnipeds (northern elephant seal ( Mirounga angustirostris ), harbor seal, ring seal, spotted seal, bearded seal, and California sea lion ( Zalophus californianus )) that were exposed to a limited number of sound sources ( i.e., mostly tones and octave-band noise with limited number of exposure to impulsive sources such as seismic airguns or impact pile driving) in laboratory settings (Southall et al., 2019a). There is currently no data available on noise-induced hearing loss for mysticetes. For summaries of data on TTS or PTS in marine mammals or for further discussion of TTS or PTS onset thresholds, please see Southall et al. (2019a) and NMFS (2018).

Recent studies with captive odontocete species (bottlenose dolphin, harbor porpoise, beluga, and false killer whale) have observed increases in hearing threshold levels when individuals received a warning sound prior to exposure to a relatively loud sound (Nachtigall and Supin, 2013; Nachtigall and Supin, 2015; Nachtigall et al., 2016a; Nachtigall et al., 2016b; Nachtigall et al., 2016c; Finneran, 2018; Nachtigall et al., 2018). These studies suggest that captive animals have a mechanism to reduce hearing sensitivity prior to impending loud sounds. Hearing change was observed to be frequency dependent and Finneran (2018) suggests hearing attenuation occurs within the cochlea or auditory nerve. Based on these observations on captive odontocetes, the authors suggest that wild animals may have a mechanism to self-mitigate the impacts of noise exposure by dampening their hearing during prolonged exposures of loud sound or if conditioned to anticipate intense sounds (Finneran, 2018; Nachtigall et al., 2018).

Exposure of marine mammals to sound sources can result in, but is not limited to, no response or any of the following observable responses: increased alertness; orientation or attraction to a sound source; vocal modifications; cessation of feeding; cessation of social interaction; alteration of movement or diving behavior; habitat abandonment (temporary or permanent); and in severe cases, panic, flight, stampede, or stranding, potentially resulting in death (Southall et al., 2007). A review of marine mammal responses to anthropogenic sound was first conducted by Richardson (1995). More recent reviews address studies conducted since 1995 and focused on observations where the received sound level of the exposed marine mammal(s) was known or could be estimated (Nowacek et al., 2007; DeRuiter et al., 2013; Ellison et al., 2012; Gomez et al., 2016). Gomez et al. (2016) conducted a review of the literature considering the contextual information of exposure in addition to received level and found that higher received levels were not always associated with more severe behavioral responses and vice versa. Southall et al. (2021) states that results demonstrate that some individuals of different species display clear yet varied responses, some of which have negative implications while others appear to tolerate high levels and that responses may not be fully predictable with simple acoustic exposure metrics ( e.g., received sound level). Rather, the authors state that differences among species and individuals along with contextual aspects of exposure ( e.g., behavioral state) appear to affect response probability.

Behavioral responses to sound are highly variable and context-specific. Many different variables can influence an animal's perception of and response to (nature and magnitude) an acoustic event. An animal's prior experience with a sound or sound source affects whether it is less likely (habituation) or more likely (sensitization) to respond to certain sounds in the future (animals can also be innately predisposed to respond to certain sounds in certain ways) (Southall et al., 2019a). Related to the sound itself, the perceived nearness of the sound, bearing of the sound (approaching vs. retreating), the similarity of a sound to biologically relevant sounds in the animal's environment ( i.e., calls of predators, prey, or conspecifics), and familiarity of the sound may affect the way an animal responds to the sound (Southall et al., 2007; DeRuiter et al., 2013). Individuals (of different age, gender, reproductive status, etc. ) among most populations will have variable hearing capabilities, and differing behavioral sensitivities to sounds that will be affected by prior conditioning, experience, and current activities of those individuals. Often, specific acoustic features of the sound and contextual variables ( i.e., proximity, duration, or recurrence of the sound or the current behavior that the marine mammal is engaged in or its prior experience), as well as entirely separate factors, such as the physical presence of a nearby vessel, may be more relevant to the animal's response than the received level alone.

Overall, the variability of responses to acoustic stimuli depends on the species receiving the sound, the sound source, and the social, behavioral, or environmental contexts of exposure ( e.g., DeRuiter and Doukara, 2012). For example, Goldbogen et al. (2013a) demonstrated that individual behavioral state was critically important in determining response of blue whales to sonar, noting that some individuals engaged in deep (greater than 50 m) feeding behavior had greater dive responses than those in shallow feeding or non-feeding conditions. Some blue whales in the Goldbogen et al. (2013a) study that were engaged in shallow feeding behavior demonstrated no clear changes in diving or movement even when received levels were high (~160 dB re 1µPa (microPascal)) for exposures to 3-4 kHz sonar signals, while deep feeding and non-feeding whales showed a clear response at exposures at lower received levels of sonar and pseudorandom noise. Southall et al. (2011) found that blue whales had a different response to sonar exposure depending on behavioral state, more pronounced when deep feeding/travel modes than when engaged in surface feeding.

With respect to distance influencing disturbance, DeRuiter et al. (2013) examined behavioral responses of Cuvier's beaked whales to mid-frequency sonar and found that whales responded strongly at low received levels (89-127 dB re 1µPa) by ceasing normal fluking and echolocation, swimming rapidly away, and extending both dive duration and subsequent non-foraging intervals when the sound source was 3.4-9.5 km away. Importantly, this study also showed that whales exposed to a similar range of received levels (78-106 dB re 1µPa) from distant sonar exercises (118 km away) did not elicit such responses, suggesting that context may moderate reactions. Thus, distance from the source is an important variable in influencing the type and degree of behavioral response and this variable is independent of the effect of received levels ( e.g., DeRuiter et al., 2013; Dunlop et al., 2017a; Dunlop et al., 2017b; Falcone et al., 2017; Dunlop et al., 2018; Southall et al., 2019a).

Ellison et al. (2012) outlined an approach to assessing the effects of sound on marine mammals that incorporates contextual-based factors. The authors recommend considering not just the received level of sound, but also the activity the animal is engaged in at the time the sound is received, the nature and novelty of the sound ( i.e., is this a new sound from the animal's perspective), and the distance between the sound source and the animal. They submit that this “exposure context,” as described, greatly influences the type of behavioral response exhibited by the animal. Forney et al. (2017) also point out that an apparent lack of response ( e.g., no displacement or avoidance of a sound source) may not necessarily mean there is no cost to the individual or population, as some resources or habitats may be of such high value that animals may choose to stay, even when experiencing stress or hearing loss. Forney et al. (2017) recommend considering both the costs of remaining in an area of noise exposure such as TTS, PTS, or masking, which could lead to an increased risk of predation or other threats or a decreased capability to forage, and the costs of displacement, including potential increased risk of vessel strike, increased risks of predation or competition for resources, or decreased habitat suitable for foraging, resting, or socializing. This sort of contextual information is challenging to predict with accuracy for ongoing activities that occur over large spatial and temporal expanses. However, distance is one contextual factor for which data exist to quantitatively inform a take estimate, and the method for predicting Level B harassment in this rule does consider distance to the source. Other factors are often considered qualitatively in the analysis of the likely consequences of sound exposure where supporting information is available.

Behavioral change, such as disturbance manifesting in lost foraging time, in response to anthropogenic activities is often assumed to indicate a biologically significant effect on a population of concern. However, individuals may be able to compensate for some types and degrees of shifts in behavior, preserving their health and thus their vital rates and population dynamics. For example, New et al. (2013) developed a model simulating the complex social, spatial, behavioral, and motivational interactions of coastal bottlenose dolphins in the Moray Firth, Scotland, to assess the biological significance of increased rate of behavioral disruptions caused by vessel traffic. Despite a modeled scenario in which vessel traffic increased from 70 to 470 vessels a year (a six-fold increase in vessel traffic) in response to the construction of a proposed offshore renewables' facility, the dolphins' behavioral time budget, spatial distribution, motivations, and social structure remained unchanged. Similarly, two bottlenose dolphin populations in Australia were also modeled over 5 years against a number of disturbances (Reed et al., 2020) and results indicate that habitat/noise disturbance had little overall impact on population abundances in either location, even in the most extreme impact scenarios modeled. Friedlaender et al. (2016) provided the first integration of direct measures of prey distribution and density variables incorporated into across-individual analyses of behavior responses of blue whales to sonar and demonstrated a fivefold increase in the ability to quantify variability in blue whale diving behavior. These results illustrate that responses evaluated without such measurements for foraging animals may be misleading, which again illustrates the context-dependent nature of the probability of response.

The following subsections provide examples of behavioral responses that give an idea of the variability in behavioral responses that would be expected given the differential sensitivities of marine mammal species to sound, contextual factors, and the wide range of potential acoustic sources to which a marine mammal may be exposed. Behavioral responses that could occur for a given sound exposure should be determined from the literature that is available for each species, or extrapolated from closely related species when no information exists, along with contextual factors.

Avoidance is the displacement of an individual from an area or migration path as a result of the presence of a sound or other stressors and is one of the most obvious manifestations of disturbance in marine mammals (Richardson et al., 1995). For example, gray whales ( Eschrichtius robustus ) and humpback whales are known to change direction—deflecting from customary migratory paths—in order to avoid noise from airgun surveys (Malme et al., 1984; Dunlop et al., 2018). Avoidance is qualitatively different from the flight response but also differs in the magnitude of the response ( i.e., directed movement, rate of travel, etc. ). Avoidance may be short-term with animals returning to the area once the noise has ceased ( e.g., Malme et al., 1984; Bowles et al., 1994; Goold, 1996; Stone et al., 2000; Morton and Symonds, 2002; Gailey et al., 2007; Dähne et al., 2013; Russel et al., 2016). Longer-term displacement is possible, however, which may lead to changes in abundance or distribution patterns of the affected species in the affected region if habituation to the presence of the sound does not occur ( e.g., Blackwell et al., 2004; Bejder et al., 2006; Teilmann et al., 2006; Forney et al., 2017). Avoidance of marine mammals during the construction of offshore wind facilities (specifically, impact pile driving) has been documented in the literature with some significant variation in the temporal and spatial degree of avoidance and with most studies focused on harbor porpoises as one of the most common marine mammals in European waters ( e.g., Tougaard et al., 2009; Dähne et al., 2013; Thompson et al., 2013; Russell et al., 2016; Brandt et al., 2018).

Available information on impacts to marine mammals from pile driving associated with offshore wind is limited to information on harbor porpoises and seals, as the vast majority of this research has occurred at European offshore wind projects where large whales and other odontocete species are uncommon. Harbor porpoises and harbor seals are considered to be behaviorally sensitive species ( e.g., Southall et al., 2007) and the effects of wind farm construction in Europe on these species have been well documented. These species have received particular attention in European waters due to their abundance in the North Sea (Hammond et al., 2002; Nachtsheim et al., 2021). A summary of the literature on documented effects of wind farm construction on harbor porpoise and harbor seals is described below.

Brandt et al. (2016) summarized the effects of the construction of eight offshore wind projects within the German North Sea ( i.e., Alpha Ventus, BARD Offshore I, Borkum West II, DanTysk, Global Tech I, Meerwind Süd/Ost, Nordsee Ost, and Riffgat) between 2009 and 2013 on harbor porpoises, combining passive acoustic monitoring (PAM) data from 2010 to 2013 and aerial surveys from 2009 to 2013 with data on noise levels associated with pile driving. Results of the analysis revealed significant declines in porpoise detections during pile driving when compared to 25-48 hours before pile driving began, with the magnitude of decline during pile driving clearly decreasing with increasing distances to the construction site. During the majority of projects, significant declines in detections (by at least 20 percent) were found within at least 5-10 km of the pile driving site, with declines at up to 20-30 km of the pile driving site documented in some cases. Similar results demonstrating the long-distance displacement of harbor porpoises (18-25 km) and harbor seals (up to 40 km) during impact pile driving have also been observed during the construction at multiple other European wind farms (Tougaard et al., 2009; Bailey et al., 2010; Dähne et al., 2013; Lucke et al., 2012; Haelters et al., 2015).

While harbor porpoises and seals tend to move several kilometers away from wind farm construction activities, the duration of displacement has been documented to be relatively temporary. In two studies at Horns Rev II using impact pile driving, harbor porpoise returned within 1 to 2 days following cessation of pile driving (Tougaard et al., 2009; Brandt et al., 2011). Similar recovery periods have been noted for harbor seals off England during the construction of four wind farms (Brasseur et al., 2012; Carroll et al., 2010; Hamre et al., 2011; Hastie et al., 2015; Russell et al., 2016). In some cases, an increase in harbor porpoise activity has been documented inside wind farm areas following construction ( e.g., Lindeboom et al., 2011). Other studies have noted longer term impacts after impact pile driving. Near Dogger Bank in Germany, harbor porpoises continued to avoid the area for over 2 years after construction began (Gilles et al., 2009). Approximately 10 years after construction of the Nysted wind farm, harbor porpoise abundance had not recovered to the original levels previously seen, although the echolocation activity was noted to have been increasing when compared to the previous monitoring period (Teilmann and Carstensen, 2012). However, overall, there are no indications for a population decline of harbor porpoises in European waters ( e.g., Brandt et al., 2016). Notably, where significant differences in displacement and return rates have been identified for these species, the occurrence of secondary project-specific influences such as use of mitigation measures ( e.g., bubble curtains, acoustic deterrent devices (ADDs)), or the manner in which species use the habitat in the project area, are likely the driving factors of this variation.

NMFS notes the aforementioned studies from Europe involve installing much smaller piles than US Wind proposes to install and, therefore, we anticipate noise levels from impact pile driving to be louder. For this reason, we anticipate that the greater distances of displacement observed in harbor porpoise and harbor seals documented in Europe are likely to occur off Maryland. However, we do not anticipate any greater severity of response due to harbor porpoise and harbor seal habitat use off Maryland or population-level consequences similar to European findings. In many cases, harbor porpoises and harbor seals are resident to the areas where European wind farms have been constructed. However, off Maryland, harbor porpoises are transient (with higher abundances in winter when foundation installation would not occur) and a very small percentage of the large harbor seal population are only seasonally present with no rookeries established. In summary, we anticipate that harbor porpoise and harbor seals will likely respond to pile driving by moving several kilometers away from the source but return to typical habitat use patterns when pile driving ceases.

Some avoidance behavior of other marine mammal species has been documented to be dependent on distance from the source. As described above, DeRuiter et al. (2013) noted that distance from a sound source may moderate marine mammal reactions in their study of Cuvier's beaked whales (an acoustically sensitive species), which showed the whales swimming rapidly and silently away when a sonar signal was 3.4-9.5 km away while showing no such reaction to the same signal when the signal was 118 km away even though the received levels were similar. Tyack et al. (1983) conducted playback studies of Surveillance Towed Array Sensor System (SURTASS) low-frequency active (LFA) sonar in a gray whale migratory corridor off California. Similar to North Atlantic right whales, gray whales migrate close to shore (approximately +2 km) and are low-frequency hearing specialists. The LFA sonar source was placed within the gray whale migratory corridor (approximately 2 km offshore) and offshore of most, but not all, migrating whales (approximately 4 km offshore). These locations influenced received levels and distance to the source. For the inshore playbacks, not unexpectedly, the louder the source level of the playback ( i.e., the louder the received level), whale avoided the source at greater distances. Specifically, when the source level was 170 dB rms and 178 dB rms, whales avoided the inshore source at ranges of several hundred meters, similar to avoidance responses reported by Malme et al. (1983, 1984). Whales exposed to source levels of 185 dB rms demonstrated avoidance levels at ranges of +1 km. Responses to the offshore source broadcasting at source levels of 185 and 200 dB, avoidance responses were greatly reduced. While there was observed deflection from course, in no case did a whale abandon its migratory behavior.

The signal context of the noise exposure has been shown to play an important role in avoidance responses. In a 2007-2008 Bahamas study, playback sounds of a potential predator—a killer whale—resulted in a similar but more pronounced reaction in beaked whales (an acoustically sensitive species), which included longer inter-dive intervals and a sustained straight-line departure of more than 20 km from the area (Boyd et al., 2008; Southall et al., 2009; Tyack et al., 2011). US Wind does not anticipate, and NMFS is not proposing to authorize take of beaked whales and, moreover, the sounds produced by US Wind do not have signal characteristics similar to predators. Therefore, we would not expect such extreme reactions to occur. Southall et al. (2011) found that blue whales had a different response to sonar exposure depending on behavioral state, more pronounced when deep feeding/travel modes than when engaged in surface feeding.

One potential consequence of behavioral avoidance is the altered energetic expenditure of marine mammals because energy is required to move and avoid surface vessels or the sound field associated with active sonar (Frid and Dill, 2002). Most animals can avoid that energetic cost by swimming away at slow speeds or speeds that minimize the cost of transport (Miksis-Olds, 2006), as has been demonstrated in Florida manatees (Miksis-Olds, 2006). Those energetic costs increase, however, when animals shift from a resting state, which is designed to conserve an animal's energy, to an active state that consumes energy the animal would have conserved had it not been disturbed. Marine mammals that have been disturbed by anthropogenic noise and vessel approaches are commonly reported to shift from resting to active behavioral states, which would imply that they incur an energy cost.

Forney et al. (2017) detailed the potential effects of noise on marine mammal populations with high site fidelity, including displacement and auditory masking, noting that a lack of observed response does not imply absence of fitness costs and that apparent tolerance of disturbance may have population-level impacts that are less obvious and difficult to document. Avoidance of overlap between disturbing noise and areas and/or times of particular importance for sensitive species may be critical to avoiding population-level impacts because (particularly for animals with high site fidelity) there may be a strong motivation to remain in the area despite negative impacts. Forney et al. (2017) stated that, for these animals, remaining in a disturbed area may reflect a lack of alternatives rather than a lack of effects.

A flight response is a dramatic change in normal movement to a directed and rapid movement away from the perceived location of a sound source. The flight response differs from other avoidance responses in the intensity of the response ( e.g., directed movement, rate of travel). Relatively little information on flight responses of marine mammals to anthropogenic signals exist, but observations of flight responses to the presence of predators have occurred (Connor and Heithaus, 1996; Frid and Dill, 2002). The result of a flight response could range from brief, temporary exertion and displacement from the area where the signal provokes flight to, in extreme cases, beaked whale strandings (Cox et al., 2006; D'Amico et al., 2009). However, it should be noted that response to a perceived predator does not necessarily invoke flight (Ford and Reeves, 2008), and whether individuals are solitary or in groups may influence the response. Flight responses of marine mammals have been documented in response to mobile high intensity active sonar ( e.g., Tyack et al., 2011; DeRuiter et al., 2013; Wensveen et al., 2019), and more severe responses have been documented when sources are moving towards an animal or when they are surprised by unpredictable exposures (Watkins, 1986; Falcone et al., 2017). Generally speaking, however, marine mammals would be expected to be less likely to respond with a flight response to either stationery pile driving (which they can sense is stationery and predictable) or significantly lower-level HRG surveys, unless they are within the area ensonified above behavioral harassment thresholds at the moment the source is turned on (Watkins, 1986; Falcone et al., 2017).

Changes in dive behavior in response to noise exposure can vary widely. They may consist of increased or decreased dive times and surface intervals as well as changes in the rates of ascent and descent during a dive ( e.g., Frankel and Clark, 2000; Costa et al., 2003; Ng and Leung, 2003; Nowacek et al., 2004; Goldbogen et al., 2013a; Goldbogen et al., 2013b). Variations in dive behavior may reflect interruptions in biologically significant activities ( e.g., foraging) or they may be of little biological significance. Variations in dive behavior may also expose an animal to potentially harmful conditions ( e.g., increasing the chance of ship-strike) or may serve as an avoidance response that enhances survivorship. The impact of a variation in diving resulting from an acoustic exposure depends on what the animal is doing at the time of the exposure, the type and magnitude of the response, and the context within which the response occurs ( e.g., the surrounding environmental and anthropogenic circumstances).

Nowacek et al. (2004) reported disruptions of dive behaviors in foraging North Atlantic right whales when exposed to an alerting stimulus, an action, they noted, that could lead to an increased likelihood of ship strike. The alerting stimulus was in the form of an 18-minute exposure that included three 2-minute signals played three times sequentially. This stimulus was designed with the purpose of providing signals distinct to background noise that serve as localization cues. However, the whales did not respond to playbacks of either right whale social sounds or vessel noise, highlighting the importance of the sound characteristics in producing a behavioral reaction. Although source levels for the proposed pile driving activities may exceed the received level of the alerting stimulus described by Nowacek et al. (2004), proposed mitigation strategies (further described in the Proposed Mitigation section) will reduce the severity of response to proposed pile driving activities. Converse to the behavior of North Atlantic right whales, Indo-Pacific humpback dolphins have been observed to dive for longer periods of time in areas where vessels were present and/or approaching (Ng and Leung, 2003). In both of these studies, the influence of the sound exposure cannot be decoupled from the physical presence of a surface vessel, thus complicating interpretations of the relative contribution of each stimulus to the response. Indeed, the presence of surface vessels, their approach, and speed of approach, seemed to be significant factors in the response of the Indo-Pacific humpback dolphins (Ng and Leung, 2003). Low-frequency signals of the Acoustic Thermometry of Ocean Climate (ATOC) sound source were not found to affect dive times of humpback whales in Hawaiian waters (Frankel and Clark, 2000) or to overtly affect elephant seal dives (Costa et al., 2003). They did, however, produce subtle effects that varied in direction and degree among the individual seals, illustrating the equivocal nature of behavioral effects and consequent difficulty in defining and predicting them.

Disruption of feeding behavior can be difficult to correlate with anthropogenic sound exposure, so it is usually inferred by observed displacement from known foraging areas, the cessation of secondary indicators of foraging ( e.g., bubble nets or sediment plumes), or changes in dive behavior. As for other types of behavioral response, the frequency, duration, and temporal pattern of signal presentation, as well as differences in species sensitivity, are likely contributing factors to differences in response in any given circumstance ( e.g., Croll et al., 2001; Nowacek et al., 2004; Madsen et al., 2006; Yazvenko et al., 2007; Southall et al., 2019b). An understanding of the energetic requirements of the affected individuals and the relationship between prey availability, foraging effort and success, and the life history stage of the animal can facilitate the assessment of whether foraging disruptions are likely to incur fitness consequences (Goldbogen et al., 2013b; Farmer et al., 2018; Pirotta et al., 2018a; Southall et al., 2019a; Pirotta et al., 2021).

Impacts on marine mammal foraging rates from noise exposure have been documented, though there is little data regarding the impacts of offshore turbine construction specifically. Several broader examples follow, and it is reasonable to expect that exposure to noise produced during the 5 years that the proposed rule would be effective could have similar impacts. Visual tracking, passive acoustic monitoring, and movement recording tags were used to quantify sperm whale behavior prior to, during, and following exposure to airgun arrays at received levels in the range 140-160 dB at distances of 7-13 km, following a phase-in of sound intensity and full array exposures at 1-13 km (Madsen et al., 2006; Miller et al., 2009). Sperm whales did not exhibit horizontal avoidance behavior at the surface. However, foraging behavior may have been affected. The sperm whales exhibited 19 percent less vocal (buzz) rate during full exposure relative to post exposure, and the whale that was approached most closely had an extended resting period and did not resume foraging until the airguns had ceased firing. The remaining whales continued to execute foraging dives throughout exposure; however, swimming movements during foraging dives were 6 percent lower during exposure than during control periods (Miller et al., 2009). Miller et al. (2009) noted that more data are required to understand whether the differences were due to exposure or natural variation in sperm whale behavior. Balaenopterid whales exposed to moderate low-frequency signals similar to the ATOC sound source demonstrated no variation in foraging activity (Croll et al., 2001), whereas five out of six North Atlantic right whales exposed to an acoustic alarm interrupted their foraging dives (Nowacek et al., 2004). Although the received SPLs were similar in the latter two studies, the frequency, duration, and temporal pattern of signal presentation were different. These factors, as well as differences in species sensitivity, are likely contributing factors to the differential response. The source levels of both the proposed construction and HRG activities exceed the source levels of the signals described by Nowacek et al. (2004) and Croll et al. (2001), and noise generated by US Wind's activities at least partially overlap in frequency with the described signals. Blue whales exposed to mid-frequency sonar in the Southern California Bight were less likely to produce low-frequency calls usually associated with feeding behavior (Melcón et al., 2012). However, Melcón et al. (2012) were unable to determine if suppression of low-frequency calls reflected a change in their feeding performance or abandonment of foraging behavior and indicated that implications of the documented responses are unknown. Further, it is not known whether the lower rates of calling actually indicated a reduction in feeding behavior or social contact since the study used data from remotely deployed, passive acoustic monitoring buoys. Results from the 2010-2011 field season of a behavioral response study in Southern California waters indicated that, in some cases and at low received levels, tagged blue whales responded to mid-frequency sonar but that those responses were mild and there was a quick return to their baseline activity (Southall et al., 2011; Southall et al., 2012b; Southall et al., 2019).

Information on or estimates of the energetic requirements of the individuals and the relationship between prey availability, foraging effort and success, and the life history stage of the animal will help better inform a determination of whether foraging disruptions incur fitness consequences. Foraging strategies may impact foraging efficiency, such as by reducing foraging effort and increasing success in prey detection and capture, in turn promoting fitness and allowing individuals to better compensate for foraging disruptions. Surface feeding blue whales did not show a change in behavior in response to mid-frequency simulated and real sonar sources with received levels between 90 and 179 dB re 1 µPa, but deep feeding and non-feeding whales showed temporary reactions including cessation of feeding, reduced initiation of deep foraging dives, generalized avoidance responses, and changes to dive behavior (DeRuiter et al., 2017; Goldbogen et al., 2013b; Sivle et al., 2015). Goldbogen et al. (2013b) indicate that disruption of feeding and displacement could impact individual fitness and health. However, for this to be true, we would have to assume that an individual whale could not compensate for this lost feeding opportunity by either immediately feeding at another location, by feeding shortly after cessation of acoustic exposure, or by feeding at a later time. There is no indication that individual fitness and health would be impacted, particularly since unconsumed prey would likely still be available in the environment in most cases following the cessation of acoustic exposure.

Similarly, while the rates of foraging lunges decrease in humpback whales due to sonar exposure, there was variability in the response across individuals, with one animal ceasing to forage completely and another animal starting to forage during the exposure (Sivle et al., 2016). In addition, almost half of the animals that demonstrated avoidance were foraging before the exposure, but the others were not; the animals that avoided while not feeding responded at a slightly lower received level and greater distance than those that were feeding (Wensveen et al., 2017). These findings indicate the behavioral state of the animal and foraging strategies play a role in the type and severity of a behavioral response. For example, when the prey field was mapped and used as a covariate in examining how behavioral state of blue whales is influenced by mid-frequency sound, the response in blue whale deep-feeding behavior was even more apparent, reinforcing the need for contextual variables to be included when assessing behavioral responses (Friedlaender et al., 2016).

Marine mammals vocalize for different purposes and across multiple modes, such as whistling, production of echolocation clicks, calling, and singing. Changes in vocalization behavior in response to anthropogenic noise can occur for any of these modes and may result directly from increased vigilance or a startle response, or from a need to compete with an increase in background noise (see Erbe et al., 2016 review on communication masking), the latter of which is described more below.

For example, in the presence of potentially masking signals, humpback whales and killer whales have been observed to increase the length of their songs (Miller et al., 2000; Fristrup et al., 2003; Foote et al., 2004) and blue whales increased song production (Di Iorio and Clark, 2009), while North Atlantic right whales have been observed to shift the frequency content of their calls upward while reducing the rate of calling in areas of increased anthropogenic noise (Parks et al., 2007). In some cases, animals may cease or reduce sound production during production of aversive signals (Bowles et al., 1994; Thode et al., 2020; Cerchio et al., 2014; McDonald et al., 1995). Blackwell et al. (2015) showed that whales increased calling rates as soon as airgun signals were detectable before ultimately decreasing calling rates at higher received levels.

Sound can disrupt behavior through masking, or interfering with, an animal's ability to detect, recognize, or discriminate between acoustic signals of interest ( e.g., those used for intraspecific communication and social interactions, prey detection, predator avoidance, or navigation) (Richardson et al., 1995; Erbe and Farmer, 2000; Tyack, 2000; Erbe et al., 2016). Masking occurs when the receipt of a sound is interfered with by another coincident sound at similar frequencies and at similar or higher intensity and may occur whether the sound is natural ( e.g., snapping shrimp, wind, waves, precipitation) or anthropogenic ( e.g., shipping, sonar, seismic exploration) in origin. The ability of a noise source to mask biologically important sounds depends on the characteristics of both the noise source and the signal of interest ( e.g., signal-to-noise ratio, temporal variability, direction), in relation to each other and to an animal's hearing abilities ( e.g., sensitivity, frequency range, critical ratios, frequency discrimination, directional discrimination, age, or TTS hearing loss), and existing ambient noise and propagation conditions.

Masking these acoustic signals can disturb the behavior of individual animals, groups of animals, or entire populations. Masking can lead to behavioral changes including vocal changes ( e.g., Lombard effect, increasing amplitude, or changing frequency), cessation of foraging or lost foraging opportunities, and leaving an area, to both signalers and receivers, in an attempt to compensate for noise levels (Erbe et al., 2016) or because sounds that would typically have triggered a behavior were not detected. In humans, significant masking of tonal signals occurs as a result of exposure to noise in a narrow band of similar frequencies. As the sound level increases, the detection of frequencies above those of the masking stimulus decreases. This principle is expected to apply to marine mammals as well because of common biomechanical cochlear properties across taxa.

Therefore, when the coincident (masking) sound is man-made, it may be considered harassment when disrupting behavioral patterns. It is important to distinguish TTS and PTS, which persist after the sound exposure, from masking, which only occurs during the sound exposure. Because masking (without resulting in threshold shift) is not associated with abnormal physiological function, it is not considered a physiological effect, but rather a potential behavioral effect.

The frequency range of the potentially masking sound is important in determining any potential behavioral impacts. For example, low-frequency signals may have less effect on high-frequency echolocation sounds produced by odontocetes but are more likely to affect detection of mysticete communication calls and other potentially important natural sounds such as those produced by surf and some prey species. The masking of communication signals by anthropogenic noise may be considered as a reduction in the communication space of animals ( e.g., Clark et al., 2009; Matthews, 2017) and may result in energetic or other costs as animals change their vocalization behavior ( e.g., Miller et al., 2000; Foote et al., 2004; Parks et al., 2007; Di Iorio and Clark, 2009; Holt et al., 2009). Masking can be reduced in situations where the signal and noise come from different directions (Richardson et al., 1995), through amplitude modulation of the signal, or through other compensatory behaviors (Houser and Moore, 2014). Masking can be tested directly in captive species ( e.g., Erbe, 2008), but in wild populations it must be either modeled or inferred from evidence of masking compensation. There are few studies addressing real-world masking sounds likely to be experienced by marine mammals in the wild ( e.g., Branstetter et al., 2013; Cholewiak et al., 2018).

The echolocation calls of toothed whales are subject to masking by high-frequency sound. Human data indicate low-frequency sound can mask high-frequency sounds ( i.e., upward masking). Studies on captive odontocetes by Au et al. (1974, 1985, 1993) indicate that some species may use various processes to reduce masking effects ( e.g., adjustments in echolocation call intensity or frequency as a function of background noise conditions). There is also evidence that the directional hearing abilities of odontocetes are useful in reducing masking at the high-frequencies these cetaceans use to echolocate, but not at the low-to-moderate frequencies they use to communicate (Zaitseva et al., 1980). A study by Nachtigall and Supin (2008) showed that false killer whales adjust their hearing to compensate for ambient sounds and the intensity of returning echolocation signals.

Impacts on signal detection, measured by masked detection thresholds, are not the only important factors to address when considering the potential effects of masking. As marine mammals use sound to recognize conspecifics, prey, predators, or other biologically significant sources (Branstetter et al., 2016), it is also important to understand the impacts of masked recognition thresholds (often called “informational masking”). Branstetter et al. (2016) measured masked recognition thresholds for whistle-like sounds of bottlenose dolphins and observed that they are approximately 4 dB above detection thresholds (energetic masking) for the same signals. Reduced ability to recognize a conspecific call or the acoustic signature of a predator could have severe negative impacts. Branstetter et al. (2016) observed that if “quality communication” is set at 90 percent recognition the output of communication space models (which are based on 50 percent detection) would likely result in a significant decrease in communication range.

As marine mammals use sound to recognize predators (Allen et al., 2014; Cummings and Thompson, 1971; Curé et al., 2015; Fish and Vania, 1971), the presence of masking noise may also prevent marine mammals from responding to acoustic cues produced by their predators, particularly if it occurs in the same frequency band. For example, harbor seals that reside in the coastal waters off British Columbia are frequently targeted by mammal-eating killer whales. The seals acoustically discriminate between the calls of mammal-eating and fish-eating killer whales (Deecke et al., 2002), a capability that should increase survivorship while reducing the energy required to attend to all killer whale calls. Similarly, sperm whales (Curé et al., 2016; Isojunno et al., 2016), long-finned pilot whales (Visser et al., 2016), and humpback whales (Curé et al., 2015) changed their behavior in response to killer whale vocalization playbacks; these findings indicate that some recognition of predator cues could be missed if the killer whale vocalizations were masked. The potential effects of masked predator acoustic cues depend on the duration of the masking noise and the likelihood of a marine mammal encountering a predator during the time that detection and recognition of predator cues are impeded.

Redundancy and context can also facilitate detection of weak signals. These phenomena may help marine mammals detect weak sounds in the presence of natural or manmade noise. Most masking studies in marine mammals present the test signal and the masking noise from the same direction. The dominant background noise may be highly directional if it comes from a particular anthropogenic source such as a ship or industrial site. Directional hearing may significantly reduce the masking effects of these sounds by improving the effective signal-to-noise ratio.

Masking affects both senders and receivers of acoustic signals and, at higher levels and longer duration, can potentially have long-term chronic effects on marine mammals at the population level as well as at the individual level. Low-frequency ambient sound levels have increased by as much as 20 dB (more than three times in terms of sound pressure level (SPL)) in the world's ocean from pre-industrial periods, with most of the increase from distant commercial shipping (Hildebrand, 2009; Cholewiak et al., 2018). All anthropogenic sound sources, but especially chronic and lower-frequency signals ( e.g., from commercial vessel traffic), contribute to elevated ambient sound levels, thus intensifying masking.

In addition to making it more difficult for animals to perceive and recognize acoustic cues in their environment, anthropogenic sound presents separate challenges for animals that are vocalizing. When they vocalize, animals are aware of environmental conditions that affect the “active space” (or communication space) of their vocalizations, which is the maximum area within which their vocalizations can be detected before it drops to the level of ambient noise (Brenowitz, 2004; Brumm et al., 2004; Lohr et al., 2003). Animals are also aware of environmental conditions that affect whether listeners can discriminate and recognize their vocalizations from other sounds, which is more important than simply detecting that a vocalization is occurring (Brenowitz, 1982; Brumm et al., 2004; Dooling, 2004; Marten and Marler, 1977; Patricelli and Blickley, 2006). Most species that vocalize have evolved with an ability to adjust their vocalizations to increase the signal-to-noise ratio, active space, and recognizability/distinguishability of their vocalizations in the face of temporary changes in background noise (Brumm et al., 2004; Patricelli and Blickley, 2006). Vocalizing animals can adjust their vocalization characteristics such as the frequency structure, amplitude, temporal structure, and temporal delivery (repetition rate), or ceasing to vocalize.

Many animals will combine several of these strategies to compensate for high levels of background noise. Anthropogenic sounds that reduce the signal-to-noise ratio of animal vocalizations, increase the masked auditory thresholds of animals listening for such vocalizations, or reduce the active space of an animal's vocalizations impair communication between animals. Most animals that vocalize have evolved strategies to compensate for the effects of short-term or temporary increases in background or ambient noise on their songs or calls. Although the fitness consequences of these vocal adjustments are not directly known in all instances, like most other trade-offs animals must make, some of these strategies likely come at a cost (Patricelli and Blickley, 2006; Noren et al., 2017; Noren et al., 2020). Shifting songs and calls to higher frequencies may also impose energetic costs (Lambrechts, 1996).

Marine mammals are also known to make vocal changes in response to anthropogenic noise. In cetaceans, vocalization changes have been reported from exposure to anthropogenic noise sources such as sonar, vessel noise, and seismic surveying ( e.g., Gordon et al., 2003; Di Iorio and Clark, 2009; Hatch et al., 2012; Holt et al., 2009; Holt et al., 2011; Lesage et al., 1999; McDonald et al., 2009; Parks et al., 2007; Risch et al., 2012; Rolland et al., 2012), as well as changes in the natural acoustic environment (Dunlop et al., 2014). Vocal changes can be temporary or can be persistent. For example, model simulation suggests that the increase in starting frequency for the North Atlantic right whale upcall over the last 50 years resulted in increased detection ranges between right whales. The frequency shift, coupled with an increase in call intensity by 20 dB, led to a call detectability range of less than 3 km to over 9 km (Tennessen and Parks, 2016). Holt et al. (2009) measured killer whale call source levels and background noise levels in the 1 to 40 kHz band and reported that the whales increased their call source levels by 1-dB SPL for every 1-dB SPL increase in background noise level. Similarly, another study on St. Lawrence River belugas reported a similar rate of increase in vocalization activity in response to passing vessels (Scheifele et al., 2005). Di Iorio and Clark (2009) showed that blue whale calling rates vary in association with seismic sparker survey activity, with whales calling more on days with surveys than on days without surveys. They suggested that the whales called more during seismic survey periods as a way to compensate for the elevated noise conditions.

In some cases, these vocal changes may have fitness consequences, such as an increase in metabolic rates and oxygen consumption, as observed in bottlenose dolphins when increasing their call amplitude (Holt et al., 2015). A switch from vocal communication to physical, surface-generated sounds such as pectoral fin slapping or breaching was observed for humpback whales in the presence of increasing natural background noise levels, indicating that adaptations to masking may also move beyond vocal modifications (Dunlop et al., 2010).

While these changes all represent possible tactics by the sound-producing animal to reduce the impact of masking, the receiving animal can also reduce masking by using active listening strategies such as orienting to the sound source, moving to a quieter location, or reducing self-noise from hydrodynamic flow by remaining still. The temporal structure of noise ( e.g., amplitude modulation) may also provide a considerable release from masking through comodulation masking release (a reduction of masking that occurs when broadband noise, with a frequency spectrum wider than an animal's auditory filter bandwidth at the frequency of interest, is amplitude modulated) (Branstetter and Finneran, 2008; Branstetter et al., 2013). Signal type ( e.g., whistles, burst-pulse, sonar clicks) and spectral characteristics ( e.g., frequency modulated with harmonics) may further influence masked detection thresholds (Branstetter et al., 2016; Cunningham et al., 2014).

Masking is more likely to occur in the presence of broadband, relatively continuous noise sources, such as vessels. Several studies have shown decreases in marine mammal communication space and changes in behavior as a result of the presence of vessel noise. For example, right whales were observed to shift the frequency content of their calls upward while reducing the rate of calling in areas of increased anthropogenic noise (Parks et al., 2007) as well as increasing the amplitude (intensity) of their calls (Parks, 2009; Parks et al., 2011). Clark et al. (2009) observed that right whales' communication space decreased by up to 84 percent in the presence of vessels. Cholewiak et al. (2018) also observed loss in communication space in Stellwagen National Marine Sanctuary for North Atlantic right whales, fin whales, and humpback whales with increased ambient noise and shipping noise. Although humpback whales off Australia did not change the frequency or duration of their vocalizations in the presence of ship noise, their source levels were lower than expected based on source level changes to wind noise, potentially indicating some signal masking (Dunlop, 2016). Multiple delphinid species have also been shown to increase the minimum or maximum frequencies of their whistles in the presence of anthropogenic noise and reduced communication space ( e.g., Holt et al., 2009; Holt et al., 2011; Gervaise et al., 2012; Williams et al., 2013; Hermannsen et al., 2014; Papale et al., 2015; Liu et al., 2017). While masking impacts are not a concern from lower intensity, higher frequency HRG surveys, some degree of masking would be expected in the vicinity of turbine pile driving and concentrated support vessel operation. However, pile driving is an intermittent sound and would not be continuous throughout the day.

Habituation can occur when an animal's response to a stimulus wanes with repeated exposure, usually in the absence of unpleasant associated events (Wartzok et al., 2003). Animals are most likely to habituate to sounds that are predictable and unvarying. It is important to note that habituation is appropriately considered as a “progressive reduction in response to stimuli that are perceived as neither aversive nor beneficial,” rather than as, more generally, moderation in response to human disturbance having a neutral or positive outcome (Bejder et al., 2009). The opposite process is sensitization, when an unpleasant experience leads to subsequent responses, often in the form of avoidance, at a lower level of exposure.

Both habituation and sensitization require an ongoing learning process. As noted, behavioral state may affect the type of response. For example, animals that are resting may show greater behavioral change in response to disturbing sound levels than animals that are highly motivated to remain in an area for feeding (Richardson et al., 1995; National Research Council (NRC), 2003; Wartzok et al., 2003; Southall et al., 2019b). Controlled experiments with captive marine mammals have shown pronounced behavioral reactions, including avoidance of loud sound sources ( e.g., Ridgway et al., 1997; Finneran et al., 2003; Houser et al., 2013a; Houser et al., 2013b; Kastelein et al., 2018). Observed responses of wild marine mammals to loud impulsive sound sources (typically airguns or acoustic harassment devices) have been varied but often consist of avoidance behavior or other behavioral changes suggesting discomfort (Morton and Symonds, 2002; Richardson et al., 1995; Nowacek et al., 2007; Tougaard et al., 2009; Brandt et al., 2011; Brandt et al., 2012; Dähne et al., 2013; Brandt et al., 2014; Russell et al., 2016; Brandt et al., 2018).

Stone (2015) reported data from at-sea observations during 1,196 airgun surveys from 1994 to 2010. When large arrays of airguns (considered to be 500 cubic inches (in 3 ) or more) were firing, lateral displacement, more localized avoidance, or other changes in behavior were evident for most odontocetes. However, significant responses to large arrays were found only for the minke whale and fin whale. Behavioral responses observed included changes in swimming or surfacing behavior with indications that cetaceans remained near the water surface at these times. Behavioral observations of gray whales during an airgun survey monitored whale movements and respirations pre-, during, and post-seismic survey (Gailey et al., 2016). Behavioral state and water depth were the best `natural' predictors of whale movements and respiration and after considering natural variation, none of the response variables were significantly associated with survey or vessel sounds. Many delphinids approach low-frequency airgun source vessels with no apparent discomfort or obvious behavioral change ( e.g., Barkaszi et al., 2012), indicating the importance of frequency output in relation to the species' hearing sensitivity.

An animal's perception of a threat may be sufficient to trigger stress responses consisting of some combination of behavioral responses, autonomic nervous system responses, neuroendocrine responses, or immune responses ( e.g., Selye, 1950; Moberg and Mench, 2000). In many cases, an animal's first, and sometimes most economical (in terms of energetic costs), response is behavioral avoidance of the potential stressor. Autonomic nervous system responses to stress typically involve changes in heart rate, blood pressure, and gastrointestinal activity. These responses have a relatively short duration and may or may not have a significant long-term effect on an animal's fitness.

Neuroendocrine stress responses often involve the hypothalamus-pituitary-adrenal system. Virtually all neuroendocrine functions that are affected by stress—including immune competence, reproduction, metabolism, and behavior—are regulated by pituitary hormones. Stress-induced changes in the secretion of pituitary hormones have been implicated in failed reproduction, altered metabolism, reduced immune competence, and behavioral disturbance ( e.g., Moberg, 1987; Blecha, 2000). Increases in the circulation of glucocorticoids are also equated with stress (Romano et al., 2004).

The primary distinction between stress (which is adaptive and does not normally place an animal at risk) and “distress” is the cost of the response. During a stress response, an animal uses glycogen stores that can be quickly replenished once the stress is alleviated. In such circumstances, the cost of the stress response would not pose serious fitness consequences. However, when an animal does not have sufficient energy reserves to satisfy the energetic costs of a stress response, energy resources must be diverted from other functions. This state of distress will last until the animal replenishes its energetic reserves sufficiently to restore normal function.

Relationships between these physiological mechanisms, animal behavior, and the costs of stress responses are well studied through controlled experiments and for both laboratory and free-ranging animals ( e.g., Holberton et al., 1996; Hood et al., 1998; Jessop et al., 2003; Krausman et al., 2004; Lankford et al., 2005). Stress responses due to exposure to anthropogenic sounds or other stressors and their effects on marine mammals have also been reviewed (Fair and Becker, 2000; Romano et al., 2002b) and, more rarely, studied in wild populations ( e.g., Lusseau and Bejder, 2007; Romano et al., 2002a; Rolland et al., 2012). For example, Rolland et al. (2012) found that noise reduction from reduced ship traffic in the Bay of Fundy was associated with decreased stress in North Atlantic right whales.

These and other studies lead to a reasonable expectation that some marine mammals will experience physiological stress responses upon exposure to acoustic stressors and that it is possible that some of these would be classified as “distress.” In addition, any animal experiencing TTS would likely also experience stress responses (NRC, 2003; NRC, 2017). Respiration naturally varies with different behaviors and variations in respiration rate as a function of acoustic exposure can be expected to co-occur with other behavioral reactions, such as a flight response or an alteration in diving. However, respiration rates in and of themselves may be representative of annoyance or an acute stress response. Mean exhalation rates of gray whales at rest and while diving were found to be unaffected by seismic surveys conducted adjacent to the whale feeding grounds (Gailey et al., 2007). Studies with captive harbor porpoises show increased respiration rates upon introduction of acoustic alarms (Kastelein et al., 2001; Kastelein et al., 2006a) and emissions for underwater data transmission (Kastelein et al., 2005). However, exposure of the same acoustic alarm to a striped dolphin under the same conditions did not elicit a response (Kastelein et al., 2006a), again highlighting the importance in understanding species differences in the tolerance of underwater noise when determining the potential for impacts resulting from anthropogenic sound exposure.

The definition for a stranding under title IV of the MMPA is that (A) a marine mammal is dead and is (i) on a beach or shore of the United States, or (ii) in waters under the jurisdiction of the United States (including any navigable waters); or (B) a marine mammal is alive and is (i) on a beach or shore of the United States and is unable to return to the water, (ii) on a beach or shore of the United States and, although able to return to the water, is in need of apparent medical attention, or (iii) in the waters under the jurisdiction of the United States (including any navigable waters), but is unable to return to its natural habitat under its own power or without assistance (16 U.S.C. 1421h).

Marine mammal strandings have been linked to a variety of causes, such as illness from exposure to infectious agents, biotoxins, or parasites; starvation; unusual oceanographic or weather events; or anthropogenic causes including fishery interaction, ship strike, entrainment, entrapment, sound exposure, or combinations of these stressors sustained concurrently or in series. There have been multiple events worldwide in which marine mammals (primarily beaked whales, or other deep divers) have stranded coincident with relatively nearby activities utilizing loud sound sources (primarily military training events), and five in which mid-frequency active sonar has been more definitively determined to have been a contributing factor.

There are multiple theories regarding the specific mechanisms responsible for marine mammal strandings caused by exposure to loud sounds. One primary theme is the behaviorally mediated responses of deep-diving species (odontocetes), in which their startled response to an acoustic disturbance (1) affects ascent or descent rates, the time they stay at depth or the surface, or other regular dive patterns that are used to physiologically manage gas formation and absorption within their bodies, such that the formation or growth of gas bubbles damages tissues or causes other injury, or (2) results in their flight to shallow areas, enclosed bays, or other areas considered “out of habitat,” in which they become disoriented and physiologically compromised. For more information on marine mammal stranding events and potential causes, please see the Mortality and Stranding section of NMFS Proposed Incidental Take Regulations for the Navy's Training and Testing Activities in the Hawaii-Southern California Training and Testing Study Area (50 CFR part 218, volume 83, No. 123, June 26, 2018).

The construction activities proposed by US Wind ( i.e., pile driving) do not inherently have the potential to result in marine mammal strandings. While vessel strikes could kill or injure a marine mammals (which may eventually strand), the required mitigation measures would reduce the potential for take from these activities to de minimus levels (see Proposed Mitigation section for more details). As described above, no mortality or serious injury is anticipated or proposed to be authorized from any Project activities.

Of the strandings documented to date worldwide, NMFS is not aware of any being attributed to pile driving or to the types of HRG equipment proposed for use during the Project. Recently, there has been heightened interest in HRG surveys and their potential role in recent marine mammals strandings along the U.S. east coast. HRG surveys involve the use of certain sources to image the ocean bottom, which are very different from seismic airguns used in oil and gas surveys or tactical military sonar, in that they produce much smaller impact zones. Marine mammals may respond to exposure to these sources by, for example, avoiding the immediate area, which is why offshore wind developers have authorization to allow for Level B (behavioral) harassment, including US Wind. However, because of the combination of lower source levels, higher frequency, narrower beam-width (for some sources), and other factors, the area within which a marine mammal might be expected to be behaviorally disturbed by HRG sources is much smaller (by orders of magnitude) than the impact areas for seismic airguns or the military sonar with which a small number of marine mammal have been causally associated. Specifically, estimated harassment zones for HRG surveys are typically less than 200m (such as those associated with the Project), while zones for military mid-frequency active sonar or seismic airgun surveys typically extend for several kms ranging up to 10s of km. Further, because of this much smaller ensonified area, any marine mammal exposure to HRG sources is reasonably expected to be at significantly lower levels and shorter duration (associated with less severe responses), and there is no evidence suggesting, or reason to speculate, that marine mammals exposed to HRG survey noise are likely to be injured, much less strand, as a result. Last, all but one of the small number of marine mammal stranding events that have been causally associated with exposure to loud sound sources have been deep-diving toothed whale species (not mysticetes), which are known to respond differently to loud sounds.

The different ways that marine mammals respond to sound are sometimes indicators of the ultimate effect that exposure to a given stimulus will have on the well-being (survival, reproduction, etc. ) of an animal. There are numerous data relating the exposure of terrestrial mammals from sound to effects on reproduction or survival, and data for marine mammals continues to grow. Several authors have reported that disturbance stimuli may cause animals to abandon nesting and foraging sites (Sutherland and Crockford, 1993); may cause animals to increase their activity levels and suffer premature deaths or reduced reproductive success when their energy expenditures exceed their energy budgets (Daan et al., 1996; Feare, 1976; Mullner et al., 2004); or may cause animals to experience higher predation rates when they adopt risk-prone foraging or migratory strategies (Frid and Dill, 2002). Each of these studies addressed the consequences of animals shifting from one behavioral state ( e.g., resting or foraging) to another behavioral state ( e.g., avoidance or escape behavior) because of human disturbance or disturbance stimuli.

Attention is the cognitive process of selectively concentrating on one aspect of an animal's environment while ignoring other things (Posner, 1994). Because animals (including humans) have limited cognitive resources, there is a limit to how much sensory information they can process at any time. The phenomenon called “attentional capture” occurs when a stimulus (usually a stimulus that an animal is not concentrating on or attending to) “captures” an animal's attention. This shift in attention can occur consciously or subconsciously (for example, when an animal hears sounds that it associates with the approach of a predator) and the shift in attention can be sudden (Dukas, 2002; van Rij, 2007). Once a stimulus has captured an animal's attention, the animal can respond by ignoring the stimulus, assuming a “watch and wait” posture, or treat the stimulus as a disturbance and respond accordingly, which includes scanning for the source of the stimulus or “vigilance” (Cowlishaw et al., 2004).

Vigilance is an adaptive behavior that helps animals determine the presence or absence of predators, assess their distance from conspecifics, or to attend cues from prey (Bednekoff and Lima, 1998; Treves, 2000). Despite those benefits, however, vigilance has a cost of time; when animals focus their attention on specific environmental cues, they are not attending to other activities such as foraging or resting. These effects have generally not been demonstrated for marine mammals, but studies involving fish and terrestrial animals have shown that increased vigilance may substantially reduce feeding rates (Saino, 1994; Beauchamp and Livoreil, 1997; Fritz et al., 2002; Purser and Radford, 2011). Animals will spend more time being vigilant, which may translate to less time foraging or resting, when disturbance stimuli approach them more directly, remain at closer distances, have a greater group size ( e.g., multiple surface vessels), or when they co-occur with times that an animal perceives increased risk ( e.g., when they are giving birth or accompanied by a calf).

The primary mechanism by which increased vigilance and disturbance appear to affect the fitness of individual animals is by disrupting an animal's time budget and, as a result, reducing the time they might spend foraging and resting (which increases an animal's activity rate and energy demand while decreasing their caloric intake/energy). In a study of northern resident killer whales off Vancouver Island, exposure to boat traffic was shown to reduce foraging opportunities and increase traveling time (Holt et al., 2021). A simple bioenergetics model was applied to show that the reduced foraging opportunities equated to a decreased energy intake of 18 percent while the increased traveling incurred an increased energy output of 3-4 percent, which suggests that a management action based on avoiding interference with foraging might be particularly effective.

On a related note, many animals perform vital functions, such as feeding, resting, traveling, and socializing, on a diel cycle (24-hour cycle). Behavioral reactions to noise exposure (such as disruption of critical life functions, displacement, or avoidance of important habitat) are more likely to be significant for fitness if they last more than one diel cycle or recur on subsequent days (Southall et al., 2007). Consequently, a behavioral response lasting less than 1 day and not recurring on subsequent days is not considered particularly severe unless it could directly affect reproduction or survival (Southall et al., 2007). It is important to note the difference between behavioral reactions lasting or recurring over multiple days and anthropogenic activities lasting or recurring over multiple days. For example, just because certain activities last for multiple days does not necessarily mean that individual animals will be either exposed to those activity-related stressors ( i.e., sonar) for multiple days or further exposed in a manner that would result in sustained multi-day substantive behavioral responses. However, special attention is warranted where longer-duration activities overlay areas in which animals are known to congregate for longer durations for biologically important behaviors.

There are few studies that directly illustrate the impacts of disturbance on marine mammal populations. Lusseau and Bejder (2007) present data from three long-term studies illustrating the connections between disturbance from whale-watching boats and population-level effects in cetaceans. In Shark Bay, Australia, the abundance of bottlenose dolphins was compared within adjacent control and tourism sites over three consecutive 4.5-year periods of increasing tourism levels. Between the second and third time periods, in which tourism doubled, dolphin abundance decreased by 15 percent in the tourism area and did not change significantly in the control area. In Fiordland, New Zealand, two populations (Milford and Doubtful Sounds) of bottlenose dolphins with tourism levels that differed by a factor of seven were observed and significant increases in traveling time and decreases in resting time were documented for both. Consistent short-term avoidance strategies were observed in response to tour boats until a threshold of disturbance was reached (average of 68 minutes between interactions), after which the response switched to a longer-term habitat displacement strategy. For one population, tourism only occurred in a part of the home range. However, tourism occurred throughout the home range of the Doubtful Sound population and once boat traffic increased beyond the 68-minute threshold (resulting in abandonment of their home range/preferred habitat), reproductive success drastically decreased (increased stillbirths) and abundance decreased significantly (from 67 to 56 individuals in a short period).

In order to understand how the effects of activities may or may not impact species and stocks of marine mammals, it is necessary to understand not only what the likely disturbances are going to be but how those disturbances may affect the reproductive success and survivorship of individuals, and then how those impacts to individuals translate to population-level effects. Following on the earlier work of a committee of the U.S. NRC (NRC, 2005), New et al. (2014), in an effort termed the Potential Consequences of Disturbance (PCoD), outline an updated conceptual model of the relationships linking disturbance to changes in behavior and physiology, health, vital rates, and population dynamics. This framework is a four-step process progressing from changes in individual behavior and/or physiology, to changes in individual health, then vital rates, and finally to population-level effects. In this framework, behavioral and physiological changes can have direct (acute) effects on vital rates, such as when changes in habitat use or increased stress levels raise the probability of mother-calf separation or predation; indirect and long-term (chronic) effects on vital rates, such as when changes in time/energy budgets or increased disease susceptibility affect health, which then affects vital rates; or no effect to vital rates (New et al., 2014).

Since the PCoD general framework was outlined and the relevant supporting literature compiled, multiple studies developing state-space energetic models for species with extensive long-term monitoring ( e.g., southern elephant seals, North Atlantic right whales, Ziphiidae beaked whales, and bottlenose dolphins) have been conducted and can be used to effectively forecast longer-term, population-level impacts from behavioral changes. While these are very specific models with very specific data requirements that cannot yet be applied broadly to project-specific risk assessments for the majority of species, they are a critical first step towards being able to quantify the likelihood of a population level effect. Since New et al. (2014), several publications have described models developed to examine the long-term effects of environmental or anthropogenic disturbance of foraging on various life stages of selected species ( e.g., sperm whale, Farmer et al., 2018; California sea lion, McHuron et al., 2018; blue whale, Pirotta et al., 2018a; humpback whale, Dunlop et al., 2021). These models continue to add to refinement of the approaches to the PCoD framework. Such models also help identify what data inputs require further investigation. Pirotta et al. (2018b) provides a review of the PCoD framework with details on each step of the process and approaches to applying real data or simulations to achieve each step.

Despite its simplicity, there are few complete PCoD models available for any marine mammal species due to a lack of data available to parameterize many of the steps. To date, no PCoD model has been fully parameterized with empirical data (Pirotta et al., 2018a) due to the fact they are data intensive and logistically challenging to complete. Therefore, most complete PCoD models include simulations, theoretical modeling, and expert opinion to move through the steps. For example, PCoD models have been developed to evaluate the effect of wind farm construction on the North Sea harbor porpoise populations ( e.g., King et al., 2015; Nabe-Nielsen et al., 2018). These models include a mix of empirical data, expert elicitation (King et al., 2015) and simulations of animals' movements, energetics, and/or survival (New et al., 2014; Nabe-Nielsen et al., 2018).

PCoD models may also be approached in different manners. Dunlop et al. (2021) modeled migrating humpback whale mother-calf pairs in response to seismic surveys using both a forwards and backwards approach. While a typical forwards approach can determine if a stressor would have population-level consequences, Dunlop et al. demonstrated that working backwards through a PCoD model can be used to assess the “worst case” scenario for an interaction of a target species and stressor. This method may be useful for future management goals when appropriate data becomes available to fully support the model. In another example, harbor porpoise PCoD model investigating the impact of seismic surveys on harbor porpoise included an investigation on underlying drivers of vulnerability. Harbor porpoise movement and foraging were modeled for baseline periods and then for periods with seismic surveys as well; the models demonstrated that temporal ( i.e., seasonal) variation in individual energetics and their link to costs associated with disturbances was key in predicting population impacts (Gallagher et al., 2021).

Behavioral change, such as disturbance manifesting in lost foraging time, in response to anthropogenic activities is often assumed to indicate a biologically significant effect on a population of concern. However, as described above, individuals may be able to compensate for some types and degrees of shifts in behavior, preserving their health and thus their vital rates and population dynamics. For example, New et al. (2013) developed a model simulating the complex social, spatial, behavioral, and motivational interactions of coastal bottlenose dolphins in the Moray Firth, Scotland, to assess the biological significance of increased rate of behavioral disruptions caused by vessel traffic. Despite a modeled scenario in which vessel traffic increased from 70 to 470 vessels a year (a six-fold increase in vessel traffic) in response to the construction of a proposed offshore renewables' facility, the dolphins' behavioral time budget, spatial distribution, motivations, and social structure remain unchanged. Similarly, two bottlenose dolphin populations in Australia were also modeled over 5 years against a number of disturbances (Reed et al., 2020), and results indicated that habitat/noise disturbance had little overall impact on population abundances in either location, even in the most extreme impact scenarios modeled.

By integrating different sources of data ( e.g., controlled exposure data, activity monitoring, telemetry tracking, and prey sampling) into a theoretical model to predict effects from sonar on a blue whale's daily energy intake, Pirotta et al. (2021) found that tagged blue whales' activity budgets, lunging rates, and ranging patterns caused variability in their predicted cost of disturbance. This method may be useful for future management goals when appropriate data becomes available to fully support the model. Harbor porpoise movement and foraging were modeled for baseline periods and then for periods with seismic surveys as well; the models demonstrated that the seasonality of the seismic activity was an important predictor of impact (Gallagher et al., 2021).

In their table 1, Keen et al. (2021) summarize the emerging themes in PCoD models that should be considered when assessing the likelihood and duration of exposure and the sensitivity of a population to disturbance (see table 1 from Keen et al., 2021, below). The themes are categorized by life history traits (movement ecology, life history strategy, body size, and pace of life), disturbance source characteristics (overlap with biologically important areas, duration and frequency, and nature and context), and environmental conditions (natural variability in prey availability and climate change). Keen et al. (2021) then summarize how each of these features influence an assessment, noting, for example, that individual animals with small home ranges have a higher likelihood of prolonged or year-round exposure, that the effect of disturbance is strongly influenced by whether it overlaps with biologically important habitats when individuals are present, and that continuous disruption will have a greater impact than intermittent disruption.

Nearly all PCoD studies and experts agree that infrequent exposures of a single day or less are unlikely to impact individual fitness, let alone lead to population level effects (Booth et al., 2016; Booth et al., 2017; Christiansen and Lusseau, 2015; Farmer et al., 2018; Wilson et al., 2020; Harwood and Booth, 2016; King et al., 2015; McHuron et al., 2018; National Academies of Sciences, Engineering, and Medicine (NAS), 2017; New et al., 2014; Pirotta et al., 2018a; Southall et al., 2007; Villegas-Amtmann et al., 2015). As described through this proposed rule, NMFS expects that any behavioral disturbance that would occur due to animals being exposed to construction activity would be of a relatively short duration, with behavior returning to a baseline state shortly after the acoustic stimuli ceases or the animal moves far enough away from the source. Given this, and NMFS' evaluation of the available PCoD studies, and the required mitigation discussed later, any such behavioral disturbance resulting from US Wind's activities is not expected to impact individual animals' health or have effects on individual animals' survival or reproduction, thus no detrimental impacts at the population level are anticipated. Marine mammals may temporarily avoid the immediate area but are not expected to permanently abandon the area or their migratory or foraging behavior. Impacts to breeding, feeding, sheltering, resting, or migration are not expected nor are shifts in habitat use, distribution, or foraging success.

Vessel collisions with marine mammals, also referred to as vessel strikes or ship strikes, can result in death or serious injury of the animal. Wounds resulting from ship strike may include massive trauma, hemorrhaging, broken bones, or propeller lacerations (Knowlton and Kraus, 2001). An animal at the surface could be struck directly by a vessel, a surfacing animal could hit the bottom of a vessel, or an animal just below the surface could be cut by a vessel's propeller. Superficial strikes may not kill or result in the death of the animal. Lethal interactions are typically associated with large whales, which are occasionally found draped across the bulbous bow of large commercial ships upon arrival in port. Although smaller cetaceans are more maneuverable in relation to large vessels than are large whales, they may also be susceptible to strike. The severity of injuries typically depends on the size and speed of the vessel (Knowlton and Kraus, 2001; Laist et al., 2001; Vanderlaan and Taggart, 2007; Conn and Silber, 2013). Impact forces increase with speed, as does the probability of a strike at a given distance (Silber et al., 2010; Gende et al., 2011).

The most vulnerable marine mammals are those that spend extended periods of time at the surface in order to restore oxygen levels within their tissues after deep dives ( e.g., the sperm whale). In addition, some baleen whales seem generally unresponsive to vessel sound, making them more susceptible to vessel collisions (Nowacek et al., 2004). These species are primarily large, slow-moving whales. Marine mammal responses to vessels may include avoidance and changes in dive pattern (NRC, 2003).

An examination of all known ship strikes from all shipping sources (civilian and military) indicates vessel speed is a principal factor in whether a vessel strike occurs and, if so, whether it results in injury, serious injury, or mortality (Knowlton and Kraus, 2001; Laist et al., 2001; Jensen and Silber, 2003; Pace and Silber, 2005; Vanderlaan and Taggart, 2007; Conn and Silber, 2013). In assessing records in which vessel speed was known, Laist et al. (2001) found a direct relationship between the occurrence of a whale strike and the speed of the vessel involved in the collision. The authors concluded that most deaths occurred when a vessel was traveling in excess of 13 kn.

Jensen and Silber (2003) detailed 292 records of known or probable ship strikes of all large whale species from 1975 to 2002. Of these, vessel speed at the time of collision was reported for 58 cases. Of these 58 cases, 39 (or 67 percent) resulted in serious injury or death (19 of those resulted in serious injury as determined by blood in the water, propeller gashes or severed tailstock, and fractured skull, jaw, vertebrae, hemorrhaging, massive bruising, or other injuries noted during necropsy and 20 resulted in death). Operating speeds of vessels that struck various species of large whales ranged from 2 to 51 kn. The majority (79 percent) of these strikes occurred at speeds of 13 kn or greater. The average speed that resulted in serious injury or death was 18.6 kn. Pace and Silber (2005) found that the probability of death or serious injury increased rapidly with increasing vessel speed. Specifically, the predicted probability of serious injury or death increased from 45 to 75 percent as vessel speed increased from 10 to 14 kn and exceeded 90 percent at 17 kn. Higher speeds during collisions result in greater force of impact and also appear to increase the chance of severe injuries or death. While modeling studies have suggested that hydrodynamic forces pulling whales toward the vessel hull increase with increasing speed (Clyne, 1999; Knowlton et al., 1995), this is inconsistent with Silber et al. (2010), which demonstrated that there is no such relationship ( i.e., hydrodynamic forces are independent of speed).

In a separate study, Vanderlaan and Taggart (2007) analyzed the probability of lethal mortality of large whales at a given speed, showing that the greatest rate of change in the probability of a lethal injury to a large whale as a function of vessel speed occurs between 8.6 and 15 kn. The chances of a lethal injury decline from approximately 80 percent at 15 kn to approximately 20 percent at 8.6 kn. At speeds below 11.8 kn, the chances of lethal injury drop below 50 percent, while the probability asymptotically increases toward 100 percent above 15 kn.

The Jensen and Silber (2003) report notes that the Large Whale Ship Strike Database represents a minimum number of collisions, because the vast majority probably goes undetected or unreported. In contrast, the Project's personnel are likely to detect any strike that does occur because of the required personnel training and lookouts, along with the inclusion of Protected Species Observers (as described in the Proposed Mitigation section), and they are required to report all ship strikes involving marine mammals.

There are no known vessel strikes of marine mammals by any offshore wind energy vessel in the United States. Given the extensive mitigation and monitoring measures (see the Proposed Mitigation and Proposed Monitoring and Reporting section) that would be required of US Wind, NMFS believes that a vessel strike is not likely to occur.

US Wind's proposed activities could potentially affect marine mammal habitat through the introduction of impacts to the prey species of marine mammals (through noise, oceanographic processes, or reef effects), acoustic habitat (sound in the water column), water quality, and biologically important habitat for marine mammals.

Sound may affect marine mammals through impacts on the abundance, behavior, or distribution of prey species ( e.g., crustaceans, cephalopods, fish, and zooplankton). Marine mammal prey varies by species, season, and location and, for some, is not well documented. Here, we describe studies regarding the effects of noise on known marine mammal prey.

Fish utilize the soundscape and components of sound in their environment to perform important functions such as foraging, predator avoidance, mating, and spawning ( e.g., Zelick and Mann, 1999; Fay, 2009). The most likely effects on fishes exposed to loud, intermittent, low-frequency sounds are behavioral responses ( i.e., flight or avoidance). Short duration, sharp sounds (such as pile driving or airguns) can cause overt or subtle changes in fish behavior and local distribution. The reaction of fish to acoustic sources depends on the physiological state of the fish, past exposures, motivation ( e.g., feeding, spawning, migration), and other environmental factors. Key impacts to fishes may include behavioral responses, hearing damage, barotrauma (pressure-related injuries), and mortality. While it is clear that the behavioral responses of individual prey, such as displacement or other changes in distribution, can have direct impacts on the foraging success of marine mammals, the effects on marine mammals of individual prey that experience hearing damage, barotrauma, or mortality is less clear, though obviously population scale impacts that meaningfully reduce the amount of prey available could have more serious impacts.

Fishes, like other vertebrates, have a variety of different sensory systems to glean information from ocean around them (Astrup and Mohl, 1993; Astrup, 1999; Braun and Grande, 2008; Carroll et al., 2017; Hawkins and Johnstone, 1978; Ladich and Popper, 2004; Ladich and Schulz-Mirbach, 2016; Mann, 2016; Nedwell et al., 2004; Popper et al., 2003; Popper et al., 2005). Depending on their hearing anatomy and peripheral sensory structures, which vary among species, fishes hear sounds using pressure and particle motion sensitivity capabilities and detect the motion of surrounding water (Fay et al., 2008) (terrestrial vertebrates generally only detect pressure). Most marine fishes primarily detect particle motion using the inner ear and lateral line system while some fishes possess additional morphological adaptations or specializations that can enhance their sensitivity to sound pressure, such as a gas-filled swim bladder (Braun and Grande, 2008; Popper and Fay, 2011).

Hearing capabilities vary considerably between different fish species with data only available for just over 100 species out of the 34,000 marine and freshwater fish species (Eschmeyer and Fong, 2016). In order to better understand acoustic impacts on fishes, fish hearing groups are defined by species that possess a similar continuum of anatomical features, which result in varying degrees of hearing sensitivity (Popper and Hastings, 2009a). There are four hearing groups defined for all fish species (modified from Popper et al., 2014) within this analysis, and they include: fishes without a swim bladder ( e.g., flatfish, sharks, rays, etc. ); fishes with a swim bladder not involved in hearing ( e.g., salmon, cod, pollock, etc. ); fishes with a swim bladder involved in hearing ( e.g., sardines, anchovy, herring, etc. ); and fishes with a swim bladder involved in hearing and high-frequency hearing ( e.g., shad and menhaden). Most marine mammal fish prey species would not be likely to perceive or hear mid- or high-frequency sonars. While hearing studies have not been done on sardines and northern anchovies, it would not be unexpected for them to have hearing similarities to Pacific herring (up to 2-5 kHz) (Mann et al., 2005). Currently, less data are available to estimate the range of best sensitivity for fishes without a swim bladder.

In terms of physiology, multiple scientific studies have documented a lack of mortality or physiological effects to fish from exposure to low- and mid-frequency sonar and other sounds (Halvorsen et al., 2012a; Jørgensen et al., 2005; Juanes et al., 2017; Kane et al., 2010; Kvadsheim and Sevaldsen, 2005; Popper et al., 2007; Popper et al., 2016; Watwood et al., 2016). Techer et al. (2017) exposed carp in floating cages for up to 30 days to low-power 23 and 46 kHz source without any significant physiological response. Other studies have documented either a lack of TTS in species whose hearing range cannot perceive sonar (such as Navy sonar), or for those species that could perceive sonar-like signals, any TTS experienced would be recoverable (Halvorsen et al., 2012a; Ladich and Fay, 2013; Popper and Hastings, 2009a, 2009b; Popper et al., 2014; Smith, 2016). Only fishes that have specializations that enable them to hear sounds above about 2,500 Hz (2.5 kHz), such as herring (Halvorsen et al., 2012a; Mann et al., 2005; Mann, 2016; Popper et al., 2014), would have the potential to receive TTS or exhibit behavioral responses from exposure to mid-frequency sonar. In addition, any sonar induced TTS to fish whose hearing range could perceive sonar would only occur in the narrow spectrum of the source ( e.g., 3.5 kHz) compared to the fish's total hearing range ( e.g., 0.01 kHz to 5 kHz).

In terms of behavioral responses, Juanes et al. (2017) discuss the potential for negative impacts from anthropogenic noise on fish, but the author's focus was on broader based sounds, such as ship and boat noise sources. Watwood et al. (2016) also documented no behavioral responses by reef fish after exposure to mid-frequency active sonar. Doksaeter et al. (2009; 2012) reported no behavioral responses to mid-frequency sonar (such as naval sonar) by Atlantic herring; specifically, no escape reactions (vertically or horizontally) were observed in free swimming herring exposed to mid-frequency sonar transmissions. Based on these results (Doksaeter et al., 2009; Doksaeter et al., 2012; Sivle et al., 2012), Sivle et al. (2014) created a model in order to report on the possible population-level effects on Atlantic herring from active sonar. The authors concluded that the use of sonar poses little risk to populations of herring regardless of season, even when the herring populations are aggregated and directly exposed to sonar. Finally, Bruintjes et al. (2016) commented that fish exposed to any short-term noise within their hearing range might initially startle but would quickly return to normal behavior.

Pile driving noise during construction is of particular concern as the very high sound pressure levels could potentially prevent fish from reaching breeding or spawning sites, finding food, and acoustically locating mates. A playback study in West Scotland revealed that there was a significant movement response to the pile driving stimulus in both species at relatively low received sound pressure levels (sole: 144-156 dB re 1μPa Peak; cod: 140-161 dB re 1 μPa Peak, particle motion between 6.51 * 10 3 and 8.62 * 10 4 m/s 2 peak) (Mueller-Blenkle et al., 2010). The swimming speed of sole increased significantly during the playback of construction noise when compared to the playbacks of before and after construction. While not statistically significant, cod also displayed a similar behavioral response during before, during, and after construction playbacks. However, cod demonstrated a specific and significant freezing response at the onset and cessation of the playback recording. In both species, indications were present displaying directional movements away from the playback source. During wind farm construction in the eastern Taiwan Strait, Type 1 soniferous fish chorusing showed a relatively lower intensity and longer duration while Type 2 chorusing exhibited higher intensity and no changes in its duration. Deviation from regular fish vocalization patterns may affect fish reproductive success, cause migration, augmented predation, or physiological alterations.

Occasional behavioral reactions to activities that produce underwater noise sources are unlikely to cause long-term consequences for individual fish or populations. The most likely impact to fish from impact and vibratory pile driving activities at the project areas would be temporary behavioral avoidance of the area. Any behavioral avoidance by fish of the disturbed area would still leave significantly large areas of fish and marine mammal foraging habitat in the nearby vicinity. The duration of fish avoidance of an area after pile driving stops is unknown, but a rapid return to normal recruitment, distribution and behavior is anticipated. In general, impacts to marine mammal prey species are expected to be minor and temporary due to the expected short daily duration of individual pile driving events and the relatively small areas being affected.

SPLs of sufficient strength have been known to cause fish auditory impairment, injury, and mortality. Popper et al. (2014) found that fish with or without air bladders could experience TTS at 186 dB SEL cum . Mortality could occur for fish without swim bladders at >216 dB SEL cum . Those with swim bladders or at the egg or larvae life stage, mortality was possible at >203 dB SEL cum . Other studies found that 203 dB SEL cum or above caused a physiological response in other fish species (Casper et al., 2012; Halvorsen et al., 2012a; Halvorsen et al., 2012b; Casper et al., 2013a; Casper et al., 2013b). However, in most fish species, hair cells in the ear continuously regenerate and loss of auditory function likely is restored when damaged cells are replaced with new cells. Halvorsen et al. (2012a) showed that a TTS of 4-6 dB was recoverable within 24 hours for one species. Impacts would be most severe when the individual fish is close to the source and when the duration of exposure is long. Injury caused by barotrauma can range from slight to severe and can cause death and is most likely for fish with swim bladders. Barotrauma injuries have been documented during controlled exposure to impact pile driving (Halvorsen et al., 2012b; Casper et al., 2013a).

As described in the Proposed Mitigation section below, US Wind would utilize a sound attenuation device which would reduce potential for injury to marine mammal prey. Other fish that experience hearing loss as a result of exposure to impulsive sound sources may have a reduced ability to detect relevant sounds such as predators, prey, or social vocalizations. However, PTS has not been known to occur in fishes and any hearing loss in fish may be as temporary as the timeframe required to repair or replace the sensory cells that were damaged or destroyed (Popper et al., 2005; Popper et al., 2014; Smith, 2006). It is not known if damage to auditory nerve fibers could occur, and if so, whether fibers would recover during this process. In addition, most acoustic effects, if any, are expected to be short-term and localized. Long-term consequences for fish populations, including key prey species within the project area, would not be expected.

Required soft-starts would allow prey and marine mammals to move away from the source prior to any noise levels that may physically injure prey and the use of the noise attenuation devices would reduce noise levels to the degree any mortality or injury of prey is also minimized. Use of bubble curtains, in addition to reducing impacts to marine mammals, for example, is a key mitigation measure in reducing injury and mortality of ESA-listed salmon on the U.S. west coast. However, we recognize some mortality, physical injury and hearing impairment in marine mammal prey may occur, but we anticipate the amount of prey impacted in this manner is minimal compared to overall availability. Any behavioral responses to pile driving by marine mammal prey are expected to be brief. We expect that other impacts, such as stress or masking, would occur in fish that serve as marine mammals prey (Popper et al., 2019); however, those impacts would be limited to the duration of impact pile driving, and, if prey were to move out the area in response to noise, these impacts would be minimized.

In addition to fish, prey sources such as marine invertebrates could potentially be impacted by noise stressors as a result of the proposed activities. However, most marine invertebrates' ability to sense sounds is limited. Invertebrates appear to be able to detect sounds (Pumphrey, 1950; Frings and Frings, 1967) and are most sensitive to low-frequency sounds (Packard et al., 1990; Budelmann and Williamson, 1994; Lovell et al., 2005; Mooney et al., 2010). Data on response of invertebrates such as squid, another marine mammal prey species, to anthropogenic sound is more limited (de Soto, 2016; Sole et al., 2017). Data suggest that cephalopods are capable of sensing the particle motion of sounds and detect low frequencies up to 1-1.5 kHz, depending on the species, and so are likely to detect airgun noise (Kaifu et al., 2008; Hu et al., 2009; Mooney et al., 2010; Samson et al., 2014). Sole et al. (2017) reported physiological injuries to cuttlefish in cages placed at-sea when exposed during a controlled exposure experiment to low-frequency sources (315 Hz, 139 to 142 dB re 1 μPa 2 ; 400 Hz, 139 to 141 dB re 1 μPa 2 ). Fewtrell and McCauley (2012) reported squids maintained in cages displayed startle responses and behavioral changes when exposed to seismic airgun sonar (136-162 re 1 μPa 2 ·s). Jones et al. (2020) found that when squid ( Doryteuthis pealeii ) were exposed to impulse pile driving noise, body pattern changes, inking, jetting, and startle responses were observed and nearly all squid exhibited at least one response. However, these responses occurred primarily during the first eight impulses and diminished quickly, indicating potential rapid, short-term habituation.

Cephalopods have a specialized sensory organ inside the head called a statocyst that may help an animal determine its position in space (orientation) and maintain balance (Budelmann, 1992). Packard et al. (1990) showed that cephalopods were sensitive to particle motion, not sound pressure, and Mooney et al. (2010) demonstrated that squid statocysts act as an accelerometer through which particle motion of the sound field can be detected. Auditory injuries (lesions occurring on the statocyst sensory hair cells) have been reported upon controlled exposure to low-frequency sounds, suggesting that cephalopods are particularly sensitive to low-frequency sound (Andre et al., 2011; Sole et al., 2013). Behavioral responses, such as inking and jetting, have also been reported upon exposure to low-frequency sound (McCauley et al., 2000; Samson et al., 2014). Squids, like most fish species, are likely more sensitive to low-frequency sounds and may not perceive mid- and high-frequency sonars.

With regard to potential impacts on zooplankton, McCauley et al. (2017) found that exposure to airgun noise resulted in significant depletion for more than half the taxa present and that there were two to three times more dead zooplankton after airgun exposure compared with controls for all taxa, within 1 km of the airguns. However, the authors also stated that in order to have significant impacts on r-selected species ( i.e., those with high growth rates and that produce many offspring) such as plankton, the spatial or temporal scale of impact must be large in comparison with the ecosystem concerned, and it is possible that the findings reflect avoidance by zooplankton rather than mortality (McCauley et al., 2017). In addition, the results of this study are inconsistent with a large body of research that generally finds limited spatial and temporal impacts to zooplankton as a result of exposure to airgun noise ( e.g., Dalen and Knutsen, 1987; Payne, 2004; Stanley et al., 2011). Most prior research on this topic, which has focused on relatively small spatial scales, has showed minimal effects ( e.g., Kostyuchenko, 1973; Booman et al., 1996; Sætre and Ona, 1996; Pearson et al., 1994; Bolle et al., 2012).

A modeling exercise was conducted as a follow-up to the McCauley et al. (2017) study (as recommended by McCauley et al. ), in order to assess the potential for impacts on ocean ecosystem dynamics and zooplankton population dynamics (Richardson et al., 2017). Richardson et al. (2017) found that a full-scale airgun survey would impact copepod abundance within the survey area, but that effects at a regional scale were minimal (2 percent decline in abundance within 150 km of the survey area and effects not discernible over the full region). The authors also found that recovery within the survey area would be relatively quick (3 days following survey completion) and suggest that the quick recovery was due to the fast growth rates of zooplankton, and the dispersal and mixing of zooplankton from both inside and outside of the impacted region. The authors also suggest that surveys in areas with more dynamic ocean circulation in comparison with the study region and/or with deeper waters ( i.e., typical offshore wind locations) would have less net impact on zooplankton.

Notably, a recently described study produced results inconsistent with those of McCauley et al. (2017). Researchers conducted a field and laboratory study to assess if exposure to airgun noise affects mortality, predator escape response, or gene expression of the copepod Calanus finmarchicus (Fields et al., 2019). Immediate mortality of copepods was significantly higher, relative to controls, at distances of 5 m or less from the airguns. Mortality 1 week after the airgun blast was significantly higher in the copepods placed 10 m from the airgun but was not significantly different from the controls at a distance of 20 m from the airgun. The increase in mortality, relative to controls, did not exceed 30 percent at any distance from the airgun. Moreover, the authors caution that even this higher mortality in the immediate vicinity of the airguns may be more pronounced than what would be observed in free-swimming animals due to increased flow speed of fluid inside bags containing the experimental animals. There were no sub-lethal effects on the escape performance, or the sensory threshold needed to initiate an escape response, at any of the distances from the airgun that were tested. Whereas McCauley et al. (2017) reported an SEL of 156 dB at a range of 509-658 m, with zooplankton mortality observed at that range, Fields et al. (2019) reported an SEL of 186 dB at a range of 25 m, with no reported mortality at that distance.

The presence of large numbers of turbines has been shown to impact meso- and sub-meso-scale water column circulation, which can affect the density, distribution, and energy content of zooplankton and thereby, their availability as marine mammal prey. Topside, atmospheric wakes result in wind speed reductions influencing upwelling and downwelling in the ocean while underwater structures such as WTG and OSS foundations may cause turbulent current wakes, which impact circulation, stratification, mixing, and sediment resuspension (Daewel et al., 2022). Overall, the presence of structures such as wind turbines is, in general, likely to result in certain oceanographic effects in the marine environment and may alter marine mammal prey, such as aggregations and distribution of zooplankton through changing the strength of tidal currents and associated fronts, changes in stratification, primary production, the degree of mixing, and stratification in the water column (Chen et al., 2021; Johnson et al., 2021; Christiansen et al., 2022; Dorrell et al., 2022).

US Wind intends to install up to 114 WTG and 4 OSS foundations, with turbine operations commencing in 2025 and all turbines being operational in 2027. As described above, there is scientific uncertainty around the scale of oceanographic impacts (meters to kilometers) associated with turbine operation. The Project is located offshore of Maryland along the mid-Atlantic Bight, and the project area does not include key foraging grounds for marine mammals with planktonic diets ( e.g., North Atlantic right whale), as all known prime foraging habitat is located much further north, off southern New England and north into Canada. This foraging area is approximately 544.1 km (338.1 mi) north of the project area, and it would be highly unlikely for this foraging area to be influenced by activities related to the proposed Project.

Although the project area does not provide high-quality foraging habitat for plankton-feeding marine mammals, such as North Atlantic right whales, coastal Maryland may provide seasonal high-quality foraging habitat for piscivorous marine mammals, such as humpback whales. Generally speaking, and depending on the extent, impacts on prey could impact the distribution of marine mammals in an area, potentially necessitating additional energy expenditure to find and capture prey. However, at the temporal and spatial scales anticipated for this activity, any such impacts on prey are not expected to impact the reproduction or survival of any individual marine mammals. Although studies assessing the impacts of offshore wind development on marine mammals are limited, the repopulation of wind energy areas by harbor porpoises (Brandt et al., 2016; Lindeboom et al., 2011) and harbor seals (Lindeboom et al., 2011; Russell et al., 2016) following the installation of wind turbines are promising. Overall, any impacts to marine mammal foraging capabilities due to effects on prey aggregation from the turbine presence and operation during the effective period of the proposed rule is likely to be limited.

In general, impacts to marine mammal prey species are expected to be relatively minor and temporary due to the expected short daily duration of individual pile driving events and the relatively small areas being affected. In addition, NMFS does not expect HRG acoustic sources to impact fish and most sources are likely outside the hearing range of the primary prey species in the project area.

Overall, the combined impacts of sound exposure and oceanographic impacts on marine mammal habitat resulting from the proposed activities would not be expected to have measurable effects on populations of marine mammal prey species. Prey species exposed to sound might move away from the sound source, experience TTS, experience masking of biologically relevant sounds, or show no obvious direct effects.

The presence of monopile, post-piled jacket, and pin pile foundations, scour protection, and cable protection will result in a conversion of the existing sandy bottom habitat to a hard bottom habitat with areas of vertical structural relief. This could potentially alter the existing habitat by creating an “artificial reef effect” that results in colonization by assemblages of both sessile and mobile animals within the new hard-bottom habitat (Wilhelmsson et al., 2006; Reubens et al., 2013; Bergström et al., 2014; Coates et al., 2014). This colonization by marine species, especially hard-substrate preferring species, can result in changes to the diversity, composition, and/or biomass of the area thereby impacting the trophic composition of the site (Wilhelmsson et al., 2010; Krone et al., 2013; Bergström et al., 2014; Hooper et al., 2017; Raoux et al., 2017; Harrison and Rousseau, 2020; Taormina et al., 2020; Buyse et al., 2022a; ter Hofstede et al., 2022).

Artificial structures can create increased habitat heterogeneity important for species diversity and density (Langhamer, 2012). The WTG, OSS, and meteorological tower foundations will extend through the water column, which may serve to increase settlement of meroplankton or planktonic larvae on the structures in both the pelagic and benthic zones (Boehlert and Gill, 2010). Fish and invertebrate species are also likely to aggregate around the foundations and scour protection which could provide increased prey availability and structural habitat (Boehlert and Gill, 2010; Bonar et al., 2015). Further, instances of species previously unknown, rare, or nonindigenous to an area have been documented at artificial structures, changing the composition of the food web and possibly the attractability of the area to new or existing predators (Adams et al., 2014; de Mesel, 2015; Bishop et al., 2017; Hooper et al., 2017; Raoux et al., 2017; van Hal et al., 2017; Degraer et al., 2020; Fernandez-Betelu et al., 2022). Notably, there are examples of these sites becoming dominated by marine mammal prey species, such as filter-feeding species and suspension-feeding crustaceans (Andersson and Öhman, 2010; Slavik et al., 2019; Hutchison et al., 2020; Pezy et al., 2020; Mavraki et al., 2022).

Numerous studies have documented significantly higher fish concentrations including species like cod and pouting ( Trisopterus luscus ), flounder ( Platichthys flesus ), eelpout ( Zoarces viviparus ), and eel ( Anguilla anguilla ) near in-water structures than in surrounding soft bottom habitat (Langhamer and Wilhelmsson, 2009; Bergström et al., 2013; Reubens et al., 2013). In the German Bight portion of the North Sea, fish were most densely congregated near the anchorages of jacket foundations, and the structures extending through the water column were thought to make it more likely that juvenile or larval fish encounter and settle on them (Rhode Island Coastal Resources Management Council, 2010; Krone et al., 2013). In addition, fish can take advantage of the shelter provided by these structures while also being exposed to stronger currents created by the structures, which generate increased feeding opportunities and decreased potential for predation (Wilhelmsson et al., 2006). The presence of the foundations and resulting fish aggregations around the foundations is expected to be a long-term habitat impact, but the increase in prey availability could potentially be beneficial for some marine mammals.

The most likely impact on marine mammal habitats from the project is expected to be from pile driving, which may affect marine mammal food sources such as forage fish and could also affect acoustic habitat effects on marine mammal prey ( e.g., fish).

Temporary and localized reduction in water quality will occur as a result of in-water construction activities. Most of this effect will occur during pile driving and installation of the cables, including auxiliary work such as dredging and scour placement. These activities will disturb bottom sediments and may cause a temporary increase in suspended sediment in the project area. Currents should quickly dissipate any raised total suspended sediment (TSS) levels, and levels should return to background levels once the project activities in that area cease. No direct impacts on marine mammals are anticipated due to increased TSS and turbidity; however, turbidity within the water column has the potential to reduce the level of oxygen in the water and irritate the gills of prey fish species in the proposed project area. However, turbidity plumes associated with the project would be temporary and localized, and fish in the proposed project area would be able to move away from and avoid the areas where plumes may occur. Therefore, it is expected that the impacts on prey fish species from turbidity, and therefore on marine mammals, would be minimal and temporary.

Equipment used by US Wind within the project area, including ships and other marine vessels, potentially aircrafts, and other equipment, are also potential sources of by-products ( e.g., hydrocarbons, particulate matter, heavy metals). All equipment is properly maintained in accordance with applicable legal requirements. All such operating equipment meets Federal water quality standards, where applicable. Given these requirements, impacts to water quality are expected to be minimal.

Acoustic habitat is the soundscape, which encompasses all of the sound present in a particular location and time, as a whole when considered from the perspective of the animals experiencing it. Animals produce sound for, or listen for sounds produced by, conspecifics (communication during feeding, mating, and other social activities), other animals (finding prey or avoiding predators), and the physical environment (finding suitable habitats, navigating). Together, sounds made by animals and the geophysical environment ( e.g., produced by earthquakes, lightning, wind, rain, waves) make up the natural contributions to the total acoustics of a place. These acoustic conditions, termed acoustic habitat, are one attribute of an animal's total habitat.

Soundscapes are also defined by, and acoustic habitat influenced by, the total contribution of anthropogenic sound. This may include incidental emissions from sources such as vessel traffic or may be intentionally introduced to the marine environment for data acquisition purposes (as in the use of airgun arrays) or for Navy training and testing purposes (as in the use of sonar and explosives and other acoustic sources). Anthropogenic noise varies widely in its frequency, content, duration, and loudness and these characteristics greatly influence the potential habitat-mediated effects to marine mammals (please also see the previous discussion on Masking), which may range from local effects for brief periods of time to chronic effects over large areas and for long durations. Depending on the extent of effects to habitat, animals may alter their communications signals (thereby potentially expending additional energy) or miss acoustic cues (either conspecific or adventitious). Problems arising from a failure to detect cues are more likely to occur when noise stimuli are chronic and overlap with biologically relevant cues used for communication, orientation, and predator/prey detection (Francis and Barber, 2013). For more detail on these concepts, see: Barber et al., 2009; Pijanowski et al., 2011; Francis and Barber, 2013; Lillis et al., 2014.

The term “listening area” refers to the region of ocean over which sources of sound can be detected by an animal at the center of the space. Loss of communication space concerns the area over which a specific animal signal, used to communicate with conspecifics in biologically important contexts ( e.g., foraging, mating), can be heard, in noisier relative to quieter conditions (Clark et al., 2009). Lost listening area concerns the more generalized contraction of the range over which animals would be able to detect a variety of signals of biological importance, including eavesdropping on predators and prey (Barber et al., 2009). Such metrics do not, in and of themselves, document fitness consequences for the marine animals that live in chronically noisy environments. Long-term population-level consequences mediated through changes in the ultimate survival and reproductive success of individuals are difficult to study, and particularly so underwater. However, it is increasingly well documented that aquatic species rely on qualities of natural acoustic habitats, with researchers quantifying reduced detection of important ecological cues ( e.g., Francis and Barber, 2013; Slabbekoorn et al., 2010) as well as survivorship consequences in several species ( e.g., Simpson et al., 2014; Nedelec et al., 2014).

Although this proposed rulemaking primarily covers the noise produced from construction activities relevant to the Maryland Offshore Wind Project offshore wind facility, operational noise was a consideration in NMFS' analysis of the project, as all turbines would become operational within the effective dates of the rule (if issued). It is expected that all turbines would be operational in Q1 2028. Once operational, offshore wind turbines are known to produce continuous, non-impulsive underwater noise, primarily below 1 kHz (Tougaard et al., 2020; Stöber and Thomsen, 2021).

In both newer, quieter, direct-drive systems and older generation, geared turbine designs, recent scientific studies indicate that operational noise from turbines is on the order of 110 to 125 dB re 1 μPa root-mean-square sound pressure level (SPL rms ) at an approximate distance of 50 m (Tougaard et al., 2020). Recent measurements of operational sound generated from wind turbines (direct drive, 6 MW, jacket foundations) at Block Island wind farm (BIWF) indicate average broadband levels of 119 dB at 50 m from the turbine, with levels varying with wind speed (HDR, Inc., 2019). Interestingly, measurements from BIWF turbines showed operational sound had less tonal components compared to European measurements of turbines with gear boxes.

Tougaard et al. (2020) further stated that the operational noise produced by WTGs is static in nature and lower than noise produced by passing ships. This is a noise source in this region to which marine mammals are likely already habituated. Furthermore, operational noise levels are likely lower than those ambient levels already present in active shipping lanes, such that operational noise would likely only be detected in very close proximity to the WTG (Thomsen et al., 2006; Tougaard et al., 2020). Similarly, recent measurements from a wind farm (3 MW turbines) in China found at above 300 Hz, turbines produced sound that was similar to background levels (Zhang et al., 2021). Other studies by Jansen and de Jong (2016) and Tougaard et al. (2009) determined that, while marine mammals would be able to detect operational noise from offshore wind farms (again, based on older 2 MW models) for several kilometers, they expected no significant impacts on individual survival, population viability, marine mammal distribution, or the behavior of the animals considered in their study (harbor porpoises and harbor seals). In addition, Madsen et al. (2006) found the intensity of noise generated by operational wind turbines to be much less than the noises present during construction, although this observation was based on a single turbine with a maximum power of 2 MW.

More recently, Stöber and Thomsen (2021) used monitoring data and modeling to estimate noise generated by more recently developed, larger (10 MW) direct-drive WTGs. Their findings, similar to Tougaard et al. (2020), demonstrate that there is a trend that operational noise increases with turbine size. Their study predicts broadband source levels could exceed 170-dB SPL rms for a 10-MW WTG; however, those noise levels were generated based on geared turbines; newer turbines operate with direct drive technology. The shift from using gear boxes to direct drive technology is expected to reduce the levels by 10 dB. The findings in the Stöber and Thomsen (2021) study have not been experimentally validated, though the modeling (using largely geared turbines) performed by Tougaard et al. (2020) yields similar results for a hypothetical 10 MW WTG.

Recently, Holme et al. (2023) cautioned that Tougaard et al. (2020) and Stöber and Thomsen (2021) extrapolated levels for larger turbines should be interpreted with caution since both studies relied on data from smaller turbines (0.45 to 6.15 MW) collected over a variety of environmental conditions. They demonstrated that the model presented in Tougaard et al. (2020) tends to potentially overestimate levels (up to approximately 8 dB) measured to those in the field, especially with measurements closer to the turbine for larger turbines. Holme et al. (2023) measured operational noise from larger turbines (6.3 and 8.3 MW) associated with three wind farms in Europe and found no relationship between turbine activity (power production, which is proportional to the blade's revolutions per minute) and noise level, though it was noted that this missing relationship may have been masked by the area's relatively high ambient noise sound levels. Sound levels (RMS) of a 6.3-MW direct-drive turbine were measured to be 117.3 dB at a distance of 70 m. However, measurements from 8.3 MW turbines were inconclusive as turbine noise was deemed to have been largely masked by ambient noise.

Finally, operational turbine measurements are available from the Coastal Virginia Offshore Wind (CVOW) pilot pile project, where two 7.8 m monopile WTGs were installed (HDR, 2023). Compared to BIWF, levels at CVOW were higher (10-30 dB) below 120 Hz, believed to be caused by the vibrations associated with the monopile structure, while above 120 Hz levels were consistent among the two wind farms.

Overall, noise from operating turbines would raise ambient noise levels in the immediate vicinity of the turbines; however, the spatial extent of increased noise levels would be limited. NMFS proposes to require US Wind to measure operational noise levels. US Wind did not request, and NMFS is not proposing to authorize, take incidental to operational noise from WTGs. Therefore, the topic is not discussed or analyzed further herein.

This section provides an estimate of the number of incidental takes proposed for authorization through the regulations, which will inform both NMFS' consideration of “small numbers” and the negligible impact determination.

Harassment is the only type of take expected to result from these activities. Except with respect to certain activities not pertinent here, section 3(18) of the MMPA defines “harassment” as any act of pursuit, torment, or annoyance, which has the potential to injure a marine mammal or marine mammal stock in the wild (Level A harassment) or has the potential to disturb a marine mammal or marine mammal stock in the wild by causing disruption of behavioral patterns, including, but not limited to, migration, breathing, nursing, breeding, feeding, or sheltering (Level B harassment).

Authorized takes would primarily be by Level B harassment, as noise from pile driving and HRG surveys, could result in behavioral disturbance of marine mammals that qualifies as take. Impacts such as masking and TTS can contribute to the disruption of behavioral patterns and are accounted for within those takes proposed for authorization. There is also some potential for auditory injury (Level A harassment) of all marine mammals except North Atlantic right whales. However, the amount of Level A harassment that US Wind requested, and NMFS proposes to authorize, is low. While NMFS is proposing to authorize Level A harassment and Level B harassment, the proposed mitigation and monitoring measures are expected to minimize the amount and severity of such taking to the extent practicable (see Proposed Mitigation and Proposed Monitoring and Reporting).

As described previously, no serious injury or mortality is anticipated or proposed to be authorized incidental to the specified activities. Even without mitigation, both pile driving activities and HRG surveys would not have the potential to directly cause marine mammal mortality or serious injury. However, NMFS is proposing measures to more comprehensively reduce impacts to marine mammal species. While, in general, there is a low probability that mortality or serious injury of marine mammals could occur from vessel strikes, the mitigation and monitoring measures contained within this proposed rule are expected to avoid vessel strikes (see Proposed Mitigation section). No other activities have the potential to result in mortality or serious injury.

For acoustic impacts, we estimate take by considering: (1) acoustic thresholds above which the best available science indicates marine mammals will be behaviorally harassed or incur some degree of permanent hearing impairment; (2) the area or volume of water that will be ensonified above these levels in a day; (3) the density or occurrence of marine mammals within these ensonified areas; and (4) the number of days of activities. We note that while these factors can contribute to a basic calculation to provide an initial prediction of potential takes, additional information that can qualitatively inform take estimates is also sometimes available ( e.g., previous monitoring results or average group size). Below, we describe the factors considered here in more detail and present the proposed take estimates.

As described below, there are multiple methods available to predict density or occurrence and, for each species and activity, the largest value resulting from the three take estimation methods described below ( i.e., density-based, PSO-based, or mean group size) was carried forward as the amount of take proposed for authorization, by Level B harassment. The amount of take proposed for authorization, by Level A harassment, reflects the density-based exposure estimates and, for some species and activities, consideration of other data such as mean group size.

Below, we describe NMFS' acoustic thresholds, acoustic and exposure modeling methodologies, marine mammal density calculation methodology, occurrence information, and the modeling and methodologies applied to estimate take for each of the Project's proposed construction activities. NMFS has carefully considered all information and analysis presented by US Wind, as well as all other applicable information and, based on the best available science, concurs that the estimates of the types and amounts of take for each species and stock are reasonable, and is proposing to authorize the amount requested. NMFS notes the take estimates described herein for foundation installation can be considered conservative as the estimates do not reflect the implementation of clearance and shutdown zones for any marine mammal species or stock.

NMFS recommends the use of acoustic thresholds that identify the received level of underwater sound above which exposed marine mammals would be reasonably expected to be behaviorally harassed (Level B harassment) or to incur PTS of some degree (Level A harassment). A summary of all NMFS' thresholds can be found at https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-acoustic-technical-guidance.

Though significantly driven by received level, the onset of behavioral disturbance from anthropogenic noise exposure is also informed to varying degrees by other factors related to the source or exposure context ( e.g., frequency, predictability, duty cycle, duration of the exposure, signal-to-noise ratio, distance to the source, ambient noise, and the receiving animal's hearing, motivation, experience, demography, behavior at time of exposure, life stage, depth) and can be difficult to predict ( e.g., Southall et al., 2007, 2021; Ellison et al., 2012). Based on what the available science indicates and the practical need to use a threshold based on a metric that is both predictable and measurable for most activities, NMFS typically uses a generalized acoustic threshold based on received level to estimate the onset of behavioral harassment.

NMFS generally predicts that marine mammals are likely to be behaviorally harassed in a manner considered to be Level B harassment when exposed to underwater anthropogenic noise above the received sound pressure levels (SPL RMS ) of 120 dB for continuous sources ( e.g., vibratory pile driving, drilling) and above the received SPL RMS 160 dB for non-explosive impulsive or intermittent sources ( e.g., impact pile driving, scientific sonar). Generally speaking, Level B harassment take estimates based on these behavioral harassment thresholds are expected to include any likely takes by TTS as, in most cases, the likelihood of TTS occurs at distances from the source less than those at which behavioral harassment is likely. TTS of a sufficient degree can manifest as behavioral harassment, as reduced hearing sensitivity and the potential reduced opportunities to detect important signals (conspecific communication, predators, prey) may result in changes in behavioral patterns that would not otherwise occur.

The proposed Project's construction activities include the use of impulsive or intermittent sources ( i.e., impact pile driving, some HRG acoustic sources); therefore, the 160-dB re 1 μPa (rms) threshold is applicable to our analysis.

NMFS' Technical Guidance for Assessing the Effects of Anthropogenic Sound on Marine Mammal Hearing (Version 2.0, Technical Guidance) (NMFS, 2018) identifies dual criteria to assess auditory injury (Level A harassment) to five different marine mammal groups (based on hearing sensitivity) as a result of exposure to noise from two different types of sources (impulsive or non-impulsive). As dual metrics, NMFS considers onset of PTS (Level A harassment) to have occurred when either one of the two metrics is exceeded ( i.e., metric resulting in the largest isopleth). As described above, US Wind's proposed activities include the use of impulsive sources. NMFS' thresholds identifying the onset of PTS are provided in table 8. The references, analysis, and methodology used in the development of the thresholds are described in NMFS' 2018 Technical Guidance, which may be accessed at https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-acoustic-technical-guidance.