Takes of Marine Mammals Incidental to Specified Activities; Taking Marine Mammals Incidental to Phase 2 Construction of the Vineyard Wind 1 Offshore Wind Project Off Massachusetts
Federal RegisterApr 23, 2024
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DEPARTMENT OF COMMERCE
National Oceanic and Atmospheric Administration
[RTID 0648-XD687]
Takes of Marine Mammals Incidental to Specified Activities; Taking Marine Mammals Incidental to Phase 2 Construction of the Vineyard Wind 1 Offshore Wind Project Off Massachusetts
AGENCY:
National Marine Fisheries Service (NMFS), National Oceanic and Atmospheric Administration (NOAA), Commerce.
ACTION:
Notice; proposed incidental harassment authorization; request for comments on proposed authorization.
SUMMARY:
NMFS has received a request from Vineyard Wind LLC (Vineyard Wind) for authorization to take marine mammals incidental to the completion of the construction of a commercial wind energy project offshore Massachusetts in the northern portion of Lease Area OCS-A 0501. Pursuant to the Marine Mammal Protection Act (MMPA), NMFS is requesting comments on its proposal to issue an incidental harassment authorization (IHA) to incidentally take marine mammals during the specified activities; which consists of a subset of activities for which take was authorized previously, but which Vineyard Wind did not complete within the effective dates of the previous IHA. NMFS will consider public comments prior to making any final decision on the issuance of the requested MMPA authorization and agency responses will be summarized in the final notice of our decision. The IHA would be valid for 1 year from date of issuance.
DATES:
Comments and information must be received no later than May 23, 2024.
ADDRESSES:
Comments should be addressed to Jolie Harrison, Chief, Permits and Conservation Division, Office of Protected Resources (OPR), NMFS and should be submitted via email to
ITP.taylor@noaa.gov.
Electronic copies of the application and supporting documents, as well as a list of the references cited in this document, may be obtained online at:
https://www.fisheries.noaa.gov/national/marine-mammal-protection/incidental-take-authorizations-other-energy-activities-renewable.
In case of problems accessing these documents, please call the contact listed below (see
FOR FURTHER INFORMATION CONTACT
).
Instructions:
NMFS is not responsible for comments sent by any other method, to any other address or individual, or received after the end of the comment period. Comments, including all attachments, must not exceed a 25-megabyte file size. All comments received are a part of the public record and will generally be posted online at
https://www.fisheries.noaa.gov/national/marine-mammal-protection/incidental-take-authorizations-other-energy-activities-renewable
without change. All personal identifying information (
e.g.,
name, address) voluntarily submitted by the commenter may be publicly accessible. Do not submit confidential business information or otherwise sensitive or protected information.
FOR FURTHER INFORMATION CONTACT:
Jessica Taylor, OPR, NMFS, (301) 427-8401.
SUPPLEMENTARY INFORMATION:
Background
The MMPA prohibits the “take” of marine mammals, with certain exceptions. Sections 101(a)(5)(A) and (D) of the MMPA (16 U.S.C. 1361
et seq.
) direct the Secretary of Commerce (as delegated to NMFS) to allow, upon request, the incidental, but not intentional, taking of small numbers of marine mammals by U.S. citizens who engage in a specified activity (other than commercial fishing) within a specified geographical region if certain findings are made and either regulations are proposed or, if the taking is limited to harassment, a notice of a proposed IHA is provided to the public for review.
Authorization for incidental takings shall be granted if NMFS finds that the taking will have a negligible impact on the species or stock(s) and will not have an unmitigable adverse impact on the availability of the species or stock(s) for taking for subsistence uses (where relevant). Further, NMFS must prescribe the permissible methods of taking and other “means of effecting the least practicable adverse impact” on the affected species or stocks and their habitat, paying particular attention to rookeries, mating grounds, and areas of similar significance, and on the availability of the species or stocks for taking for certain subsistence uses (referred to in shorthand as “mitigation”); and requirements pertaining to the mitigation, monitoring and reporting of the takings are set forth. The definitions of all applicable MMPA statutory terms cited above are included in the relevant sections below.
National Environmental Policy Act
To comply with the National Environmental Policy Act of 1969 (NEPA; 42 U.S.C. 4321
et seq.
) and NOAA Administrative Order (NAO) 216-6A, NMFS must review our proposed action (
i.e.,
the issuance of an IHA) with respect to potential impacts on the human environment. NMFS participated as a cooperating agency on the Bureau of Ocean Energy Management (BOEM) 2021 Environmental Impact Statement (EIS) for the Vineyard Wind 1 Offshore Wind Project.
NMFS' proposal to issue Vineyard Wind the requested IHA constitutes a federal action subject to NEPA (42 U.S.C. 4321
et seq.
). On May 10, 2021, NMFS adopted the Bureau of Ocean Energy Management's (BOEM) Vineyard Wind 1Final Environmental Impact Statement (FEIS), published on March 12, 2021 and available at:
https://www.boem.gov/renewable-energy/state-activities/vineyard-wind-1.
NMFS is currently evaluating if supplementation of the Vineyard Wind 1 EIS is required per 40 CFR 1502.9(d). We will review all comments submitted in response to this notice prior to concluding our NEPA process or making a final decision on the IHA request.
Summary of Request
On December 15, 2023, NMFS received a request from Vineyard Wind for an IHA to take marine mammals incidental to Phase 2 construction of the Vineyard Wind Offshore Wind Project off Massachusetts, specifically wind turbine generator (WTG) monopile foundation installation, in the northern portion of Lease Area OCS-A 0501. Vineyard Wind completed installation of 47 WTG monopiles and 1 electrical service platform (ESP) jacket foundation in 2023 under an IHA issued by NMFS on June 25, 2021 (86 FR 33810) with effective dates from May 1, 2023, through April 30, 2024. Due to unexpected delays, Vineyard Wind was not able to complete pile driving activities before the expiration date of the current IHA (April 30, 2024); thus, Vineyard Wind is requesting take of marine mammals incidental to installing the remaining 15 monopiles to complete foundation installation for the Project. In total, the Project will consist of 62 WTG monopiles and 1 offshore substation.
Following NMFS' review of the December 2023 application, Vineyard Wind submitted multiple revised versions of the application, and it was deemed adequate and complete on March 13, 2024. Vineyard Wind's request is for take of 14 species of marine mammals, by Level B harassment and, for 6 of these species, Level A harassment. Neither Vineyard
Wind nor NMFS expect serious injury or mortality to result from this activity and, therefore, an IHA is appropriate.
Vineyard Wind previously conducted high resolution geophysical (HRG) site characterization surveys within the Lease Area and associated export cable corridor in 2016, 2018-2021, and June-December 2023 (ESS Group
Inc.,
2016; Vineyard Wind 2018, 2019; EPI Group, 2021; RPS, 2022; Vineyard Wind 2023a-f). During the 2023 construction season, NMFS coordinated closely with Vineyard Wind to ensure compliance with their IHA. In a few instances, NMFS raised concerns with Vineyard Wind regarding their implementation of certain required measures. NMFS worked closely with Vineyard Wind throughout the construction season to course correct, where needed, and ensure compliance with the requirements (
e.g.,
mitigation, monitoring, and reporting) of the previous IHA, and information regarding their monitoring results may be found in the Estimated Take of Marine Mammals section.
Description of Proposed Activity
Overview
Vineyard Wind proposes to construct and operate an 800-megawatt (MW) wind energy facility, the Project, in the Atlantic Ocean in Lease area OCS-A 0501, offshore of Massachusetts. The project would consist of up to 62 offshore wind turbine generators (WTGs), 1 electrical service platform (ESP), an onshore substation, offshore and onshore cabling, and onshore operations and maintenance facilities. The onshore substation and ESP are now complete. Installation of 47 monopile foundations was completed under a current IHA (86 FR 33810, June 25, 2021), effective from May 1, 2023, through April 30, 2024. However, due to unexpected, Vineyard Wind will not be able to complete pile driving activities before the expiration date of the current IHA (April 30, 2024). Take of marine mammals, in the form of behavioral harassment and limited instances of auditory injury, may occur incidental to the installation of the remaining 15 WTG monopile foundations due to in-water noise exposure resulting from impact pile driving. The remaining 15 monopile foundations would occur within a Limited Installation Area (LIA) (64.3 square kilometers (km
2
; 15,888.9 acres)) within the Lease Area (264.4 km
2
(65,322.4 acres)). Installation of the remaining 15 monopile foundations is expected to occur in 2024.
Dates and Duration
The proposed pile driving activities are planned to occur in 2024 after the IHA is issued and, while not planned, may occur in June or July in 2025. Pile driving activities are estimated to require approximately 15 nonconsecutive days (30 nonconsecutive hours of pile driving). Given vessel availability, weather delay, and logistical constraints, these 15 days for installation of the remaining monopile foundations could occur close in time or spread out over months.
Although installation of a single monopile may last for several hours, active pile driving for installation of a single monopile is expected to last for a maximum of 2 hours. Up to 1 monopile may be installed per day, based upon the average pile driving time (up to 2 hours) for the installation of the currently installed 47 monopiles. Monopile foundations would be installed in batches of three to six monopiles at a time as this represents the maximum batch size that the installation vessel can carry to the LIA. After installation of a batch of three to six monopiles, there would be a 4 to 7 day pause in monopile installation to allow time for the installation vessel to return with a new batch of monopiles. No concurrent monopile installation is proposed. Vineyard Wind has proposed, and NMFS would require, that pile driving activities be prohibited from January 1 through May 31 due to the increased presence of North Atlantic right whales (NARWs) in the LIA and the timing of the project (
i.e.,
pile driving in May is not practicable). NMFS is also proposing to restrict pile driving in December to the maximum extent practicable.
Specific Geographic Region
Vineyard Wind's would construct the Project in within Federal waters off Massachusetts, in the northern portion of the Vineyard Wind Lease Area OCS-A 0501 (figure 1). This area is also referred to as the Wind Development Area (WDA). The 15 remaining monopiles would be installed in a LIA within a portion of the southwest corner of the WDA. The LIA is approximately 70.5 km
2
(17,420.9 acres) in size, as compared to the overall size of the Lease Area (264.4 km
2
(63,322.4 acres)). At its nearest point, the LIA is approximately 29 kilometers (km; 18.1 miles (mi)) from the southeast corner of Martha's Vineyard and a similar distance from Nantucket. Water depths in the WDA range from approximately 37 to 49.5 meters (m; 121-162 feet (ft)). Water depth and bottom habitat are similar throughout the Lease Area (Pyc
et al.,
2018).
Vineyard Wind's specified activities would occur in the Northeast U.S. Continental Shelf Large Marine Ecosystem (NES LME), an area of approximately 260,000 km
2
from Cape Hatteras in the south to the Gulf of Maine in the north. Specifically, the LIA is located within the Mid-Atlantic Bight subarea of the NES LME, which extends between Cape Hatteras, North Carolina, and Martha's Vineyard, Massachusetts, extending westward into the Atlantic to the 100-m isobath. The specific geographic region includes the LIA as well as the crew transfer vessel transit corridors (see Proposed Mitigation section) and cable laying routes. The installation vessel and support vessels would conduct approximately three trips to Canada during the period of the IHA, transiting from New Bedford and nearby ports. Figure 1 shows the LIA and planned locations for the remaining 15 monopiles to be installed.
EN23AP24.040
Detailed Description of the Specified Activity
Monopile Installation
Vineyard Wind proposes to install 15 monopile WTG foundations in the LIA (figure 1) to complete the Vineyard Wind Offshore Wind Project (84 FR 18346, April 30, 2019; 86 FR 33810, June 25, 2021). Vineyard Wind assumes all monopile foundations would be installed using an impact hammer. Individual monopile installation would be sequenced according to the numbers in the cross-hatched area in figure 1.
A WTG monopile foundation typically consists of a coated single steel tubular section, with several sections of rolled steel plate welded together. Each 13-MW monopile would have a maximum diameter of 9.6 m (31.5 ft). WTGs would be arranged in a grid-like pattern within the LIA with spacing of
1.9 km (1 nautical mile (nmi)) between turbines, and driven to a maximum penetration depth of 28 m (92 ft) to 35 m (115 ft) below the seafloor (Vineyard Wind, 2023). Monopile foundations would consist of a monopile with a separate transition piece.
Monopile foundations would be installed by a heavy lift vessel. The installation vessel would upend the monopile with a crane and place it in a gripper frame before lowering the monopile foundation to the seabed (see figure 4 in IHA application). Vineyard Wind would use a Monopile Installation Tool (MPIT) to seat the monopile foundation and protect against pile gripper damage as well as risks to human safety associated with pile run. The MPIT creates buoyancy within the monopile foundation using air pressure to control lowering the monopile through the pile run risk zone (Vineyard Wind, 2023). As the monopile foundation is lowered, air is released from the top of the foundation above the water surface until the pile is stabilized within the seabed. Once the monopile is lowered to the seabed, the crane hook would be released. A hydraulic impact hammer would be placed on top of the monopile and used to drive the monopile into the seabed to the target penetration depth (28-35 m). Monopile foundations would be installed using a maximum hammer energy of 4,000 kilojoules (kJ) (table 1). Pile driving would begin with a 20-minute soft-start at reduced hammer energy (see Proposed Mitigation). The hammer energy would gradually be increased based upon resistance experienced from sediments. Prior to pile driving, the MPIT process may last from 6 to 15 hours and is dependent upon local soil conditions at each monopile foundation (Vineyard Wind, 2023). Vineyard Wind anticipates that one monopile would be installed per day at a rate of approximately 2 hours of active pile driving time per monopile (table 1). Rock scour protection would be applied after foundation installation. The scour protection would be 1-2 m high (3-6 ft), with stone or rock sizes of approximately 10-30 centimeters (4-12 inches).
While post-piling activities could be ongoing at one foundation position as pile driving is occurring at another position, no concurrent/simultaneous pile driving of foundations would occur (see
Dates and Duration
section). Installation of monopile foundations is anticipated to result in the take of marine mammals due to noise generated during pile driving. Proposed mitigation, monitoring, and reporting measures are described in detail later in this document (please see Proposed Mitigation and Proposed Monitoring and Reporting).
Table 1—Impact Pile Driving Schedule
Pile type
Project component
Max
hammer
energy
(kJ)
Number of hammer strikes
Max piling time
duration per
pile
(min)
Max piling time
duration per
day
(min)
Number piles/day
9.6-m monopile
WTG
a
4000
b
2,884
117
117
1
a
Maximum hammer energy for representative monopiles installed during the 2023 Vineyard Wind Offshore Wind Project construction ranged from 3,227 to 3,831 kJ.
b
Number of hammer strikes based upon the AU-38 representative monopile installed during the 2023 Vineyard Wind Offshore Wind Project construction period at a maximum hammer energy of 3,825 kJ.
After monopile installation, transition pieces, containing work platforms and other ancillary structures, and WTGs, consisting of a tower and the energy-generating components of the turbine, would be installed. Transition pieces and WTGs would be installed on top of monopile foundations using jack-up vessels. However, installation of transition pieces and WTGS on monopile foundations is not expected to result in take of marine mammals and, therefore, are not discussed further.
Vineyard Wind has developed a sequencing plan for installation of monopiles throughout the LIA, as shown in figure 1. The sequencing plan will allow for several of the monopiles located in the northeast corner of the LIA and highest density area of NARWs, to be installed first.
Vineyard Wind anticipates that it is possible for the 15 WTGs to become operational within the effective period of the IHA. Nine of the 47 WTGs previously installed in 2023 are currently operational.
Vessel Operation
Vineyard Wind would use various types of vessels over the course of the 1-year proposed IHA for foundation installation and transporting monopile batches between ports and the LIA (table 2). Construction-related vessel activity is anticipated to include approximately 20 vessels operating throughout the specified geographic area on any given work day. Many of these vessels would remain in the LIA for days or weeks at a time, making infrequent trips to port for bunkering and provisioning, as needed. Table 2 shows the type and number of vessels Vineyard Wind would use for various construction activities as well as the associated ports. Vineyard Wind would utilize ports in New London, Connecticut and New Bedford, Massachusetts (table 2) to support offshore construction, crew transfer and logistics, and other operational activities. In addition, monopile foundations would come from a Canadian port in Halifax. Monopile foundations would be transported on an installation vessel to the LIA from Canada, and would be installed in batches of three to six monopiles at a time. Upon completion of installation of a batch of monopiles, the installation vessel would return to Canada to load an additional batch of monopiles (Vineyard Wind, 2023). For the proposed activities, it is expected that the installation vessel would need to make a maximum of three trips between Canada and the LIA.
As part of vessel-based construction activities, dynamic positioning thrusters would be utilized to hold vessels in position or move slowly during monopile installation. 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. Construction-related vessel activity, including the use of dynamic positioning thrusters, is not expected to result in take of marine mammals. While a vessel strike could cause injury or mortality of a marine mammal, Vineyard Wind proposed and NMFS is proposing to require, extensive vessel strike avoidance measures that would avoid vessel strikes from occurring (see Proposed Mitigation and Proposed Monitoring and Reporting). Vineyard Wind did not request, and NMFS
neither anticipates nor proposes to authorize, take associated with vessel activity, and this activity is not analyzed further.
Table 2—Type and Number of Vessels Anticipated During Construction
Vessel type
Vessel role
Maximum number of vessels
Expected
maximum
number of transits per month
Port
Heavy lift vessel
Pile driving
1
2
Halifax, Canada.
Trans-shipment vessel
Bubble curtain
2
4
New London, CT.
Fishing vessel
PSO support vessel
2
3
New Bedford, MA.
Service operations vessel
1
4
Safety vessel
4
2
Motor vessel
Crew transfer vessel
2
12
Inter-Array Cable Laying
Inter-array cables would be installed to connect WTGs to the ESP. In 2023, Vineyard Wind completed approximately 40 percent of the installation of inter-array cables in the Lease Area. Vineyard Wind anticipates approximately 50 percent of the inter-array cable laying to take place during the effective period of the IHA. Vineyard Wind would perform a pre-lay grapnel run to remove any obstructions, such as fishing gear, from the seafloor. The cable would be laid on the seafloor and buried using a jet trencher with scour added for cable protection near the transition pieces and ESPs. The sounds associated with cable laying are consistent with those of routine vessel operations and not expected to result in take of marine mammals. Inter-array cable laying activities are, therefore, not discussed further.
Other Activities
Vineyard Wind would not conduct high-resolution geophysical (HRG) surveys, UXO/MEC detonation, or fishery research surveys under this IHA.
Description of Marine Mammals in the Area of Specified Activities
Thirty-eight marine mammal species, comprising 39 stocks, under NMFS' jurisdiction have geographic ranges within the western North Atlantic OCS (Hayes
et al.,
2023). However, for reasons described below, Vineyard Wind has requested, and NMFS proposes to authorize, take of only 14 species (comprising 14 stocks) of marine mammals. Sections 3 and 4 of the 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, instead of reprinting the information. See
ADDRESSES
. 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/marine-mammal-stock-assessments
) 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
).
Table 3 lists all species or stocks for which take is expected and proposed to be authorized for this activity 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 by the MMPA 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 (as described in NMFS' SARs; 16 U.S.C. 1362(20)). While no serious injury or mortality is anticipated or proposed to be authorized here, 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. Four of the marine mammal species for which take is requested are listed as endangered under the ESA, including the NARW, fin whale, sei whale, and sperm whale.
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 comprise 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. 2023 draft SARs and NMFS' U.S. 2022 SARs. 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 United States 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 specific geographic region. All values presented in table 3 are the most recent available at the time of publication and are available online at:
https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments.
Table 3—Marine Mammal Species That May Occur in the LIA and Be Taken by Harassment
Common name
a
Scientific name
Stock
ESA/MMPA status;
strategic
(Y/N)
b
Stock abundance
(CV, N
min
, most recent abundance survey)
c
PBR
Annual M/SI
d
Order Artiodactyla—Cetacea—Mysticeti (baleen whales)
Family Balaenidae:
NARW
Eubalaena glacialis
Western Atlantic
E, D, Y
340 (0; 337; 2021)
e
0.7
27.2
f
Family Balaenopteridae (rorquals):
Fin whale
Balaenoptera physalus
Western North Atlantic
E, D, Y
6,802 (0.24, 5,573, 2021)
11
2.05
Sei whale
Balaenoptera borealis
Nova Scotia
E, D, Y
6,292 (1.02, 3098, 2021)
6.2
0.6
Minke whale
Balaenoptera acutorostrata
Canadian Eastern Coastal
-, -, N
21,968 (0.31, 17,002, 2021)
170
9.4
Humpback whale
Megaptera novaeangliae
Gulf of Maine
-, -, Y
1,396 (0, 1,380, 2016)
22
12.15
Superfamily Odontoceti (toothed whales, dolphins, and porpoises)
Family Physeteridae:
Sperm whale
Physeter macrocephalus
North Atlantic
E, D, Y
5,895 (0.29, 4,639, 2021)
9.28
0.2
Family Delphinidae:
Long-finned pilot whale
Globicephala melas
Western North Atlantic
-, -, N
39,215 (0.3, 30,627, 2021)
306
5.7
Bottlenose dolphin
Tursiops truncatus
Western North Atlantic Offshore
-, -, N
64,587 (0.24, 52,801, 2021)
g
507
28
Common dolphin
Delphinus delphis
Western North Atlantic
-, -, N
93,100 (0.56, 59,897, 2021)
1,452
414
Risso's dolphin
Grampus griseus
Western North Atlantic
-, -, N
44,067 (0.19, 30,662, 2021)
307
18
Atlantic white-sided dolphin
Lagenorhynchus acutus
Western North Atlantic
-, -, N
93,233 (0.71, 54,443, 2021)
544
28
Family Phocoenidae (porpoises):
Harbor porpoise
Phocoena phocoena
Gulf of Maine/Bay of Fundy
-, -, N
85,765 (0.53, 56,420, 2021)
649
145
Order Carnivora—Pinnipedia
Family Phocidae (earless seals):
Harbor seal
Phoca vitulina
Western North Atlantic
-, -, N
61,336 (0.08, 57,637, 2018)
1,729
339
Gray seal
h
Halichoerus grypus
Western North Atlantic
-, -, N
27,911 (0.2, 23,924, 2021)
1,512
4,570
a
Information on the classification of marine mammal species can be found on the web page for The Society for Marine Mammalogy's Committee on Taxonomy (
https://marinemammalscience.org/science-and-publications/list-marine-mammal-species-subspecies;
Committee on Taxonomy, 2023).
b
ESA status: Endangered (E), Threatened (T)/MMPA status: Depleted (D). A dash (-) indicates that the species is not listed under the ESA or designated as depleted under the MMPA. Under the MMPA, a strategic stock is one for which the level of direct human-caused mortality exceeds PBR, or which is determined to be declining and likely to be listed under the ESA within the foreseeable future. Any species or stock listed under the ESA is automatically designated under the MMPA as depleted and as a strategic stock.
c
NMFS 2022 marine mammal SARs online at:
https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments.
CV is the coefficient of variation; N
min
is the minimum estimate of stock abundance.
d
These values, found in NMFS's SARs, represent annual levels of human-caused mortality plus serious injury from all sources combined (
e.g.,
commercial fisheries, ship strike).
e
The draft 2023 SAR includes an estimated population (N
best
340) based on sighting history through December 2021 (89 FR 5495, January 29, 2024). In October 2023, NMFS released a technical report identifying that the NARW population size based on sighting history through 2022 was 356 whales, with a 95 percent credible interval ranging from 346 to 363 (Linden, 2023).
f
Total annual average observed NARW mortality during the period 2017-2021 was 7.1 animals and annual average observed fishery mortality was 4.6 animals. Numbers presented in this table (27.2 total mortality and 17.6 fishery mortality) are 2016-2020 estimated annual means, accounting for undetected mortality and serious injury.
g
As noted in the draft 2023 SAR (89 FR 5495, January 29, 2024), abundance estimates may include sightings of the coastal form.
h
NMFS' stock abundance estimate (and associated PBR value) applies to the U.S. population only. Total stock abundance (including animals in Canada) is approximately 394,311. The annual M/SI value given is for the total stock.
As indicated above, all 14 species (with 14 managed stocks) in table 3 temporally and spatially co-occur with the activity to the degree that take is expected to occur. The following species are not expected to occur in the LIA due to their known distributions, preferred habitats, and/or known temporal and spatial occurrences: the blue whale (
Balaenoptera musculus
), northern bottlenose whale (
Hyperoodon ampullatus
), false killer whale (
Pseudorca crassidens
), pygmy killer whale (
Feresa attenuata
), melon-headed whale (
Peponocephala electra
), dwarf and pygmy sperm whales (
Kogia
spp.), killer whale (
Orcinus orca
), Cuvier's beaked whale (
Ziphius cavirostris
), four species of Mesoplodont whale (
Mesoplodon densitostris, M. europaeus,
M. mirus,
and
M. bidens
), Fraser's dolphin (
Lagenodelphis hosei
), Clymene dolphin (
Stenella clymene
), spinner dolphin (
Stenella longirostris
), rough-toothed dolphin (
Steno bredanensis
), Atlantic spotted dolphin (
Stenella frontalis
), pantropical spotted dolphin (
Stenella attenuata
), short-finned pilot whale (
Globicephala macrorhynchus
), striped dolphin (
Stenella coeruleoalba
), white-beaked dolphin (
Lagenorhynchus albirostris
), and hooded seal (
Crysophora cristata
). None of these species were observed during the 2023 construction season or during previous site assessment/characterization surveys (Vineyard Wind, 2018, 2019, 2023a-f). Due to the lack of sightings of these species in the MA Wind Energy Area (WEA) (Kenney and Vigness-Raposa, 2010; ESS Group,
Inc.,
2016; Kraus
et al.,
2016; Vineyard Wind, 2018; 2019; O'Brien
et al.,
2020, 2021, 2022, 2023; EPI Group, 2021; Palka
et al.,
2017 2021; RPS, 2022; Vineyard Wind, 2023a-f; Hayes
et al.,
2023) as well as documented habitat preferences and distributions, we have determined that
each of these species will not be considered further. Furthermore, the northern limit of the northern migratory coastal stock of the common bottlenose dolphin (
Tursiops truncatus
) does not extend as far north as the LIA. Thus, take is only proposed for the offshore stock which may occur within the LIA. Although harp seals (
Pagophilus groenlandicus
) are expected to occur within the WDA, no harp seals were observed by Protected Species Observers (PSOs) during Vineyard Wind's site characterization surveys (2016, 2018-2021; ESS Group,
Inc.,
2016; Vineyard Wind, 2018, 2019) nor during the 2023 construction campaign (Vineyard Wind, 2023a-f). Thus, Vineyard Wind did not request, and NMFS is not proposing to authorize, take for this species.
In addition to what is included in sections 3 and 4 of Vineyard Wind's ITA application (Vineyard Wind, 2023), 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 unusual mortality events (UMEs) 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 LIA (
https://www.fisheries.noaa.gov/resource/map/national-esa-critical-habitat-mapper
). Any areas of known biological importance (including the BIAs identified in LaBrecque
et al.,
2015) that overlap spatially (or are adjacent) with the LIA are addressed in the species sections below.
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 January 2024, three UMEs are occurring along the U.S. Atlantic coast for NARWs, humpback whales, and minke whales. Of these, the most relevant to the LIA are the NARW and humpback whale UMEs given the prevalence of these species in Southern New England (SNE). 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. 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.
North Atlantic Right Whale
The NARW 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 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 NARW, and NMFS completed 5-year reviews of the species in 2012, 2017, and 2022, which concluded no change to the listing status is warranted.
The NARW 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 NARW 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). NARW 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 (Linden, 2023). 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 NARW 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).
NMFS' regulations at 50 CFR 224.105 designated Seasonal Management Areas (SMAs) for NARWs in 2008 (73 FR 60173, October 10, 2008). SMAs were developed to reduce the threat of collisions between vessels and NARWs. A portion of the Block Island SMA, which occurs off Block Island, Rhode Island, is near the LIA (approximately 4.3 km (2.7 mi) southwest of the OCS-A 0501 Lease Area at the closest point), but does not overlap spatially with the Lease Area or LIA. This SMA is active from November 1 through April 30 of each year, and may be used by NARWs for migrating and/or feeding. As noted below, NMFS is proposing changes to the NARW speed rule (87 FR 46921, August 1, 2022). NMFS has designated critical habitat for NARWs (81 FR 4838, January 27, 2016), along the U.S. southeast coast for calving as well as in the northeast, just east of the LIA. The LIA both spatially and temporally overlaps a portion of a migratory corridor BIA (LaBrecque
et al.,
2015). Due to the current status of NARWs and the spatial proximity of the proposed project with areas of biological significance, (
i.e.,
a migratory corridor, SMA), the potential impacts of the proposed project on NARWs warrant particular attention.
NARWs range from calving grounds in the southeastern United States to feeding grounds in New England waters and into Canadian waters (Hayes
et al.,
2023). Surveys have demonstrated the existence of seven areas where NARWs congregate seasonally in Georges Bank, off Cape Cod, and in Massachusetts Bay (Hayes
et al.,
2023). In late fall (
i.e.,
November), a portion of the NARW population (including pregnant females) typically departs the feeding grounds in the North Atlantic, moves south along the migratory corridor BIA, including through the LIA, to calving grounds off Georgia and Florida. This movement is followed by a northward migration (primarily mothers with young calves) into northern feeding areas in March and April (LaBrecque
et al.,
2015; Van Parijs, 2015). Recent research indicates our 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). NARWs may migrate through the LIA to access more northern feeding grounds or southern calving grounds.
NARWs may occur year-round in SNE, near Martha's Vineyard and Nantucket Shoals as well as throughout the Massachusetts and Rhode Island/Massachusetts Wind Energy Areas (MA and RI/MA WEAs) (Quintan-Rizzo
et al.,
2021; O'Brien
et al.,
2023; Van Parijs
et al.,
2023). Kraus
et al.
(2016) found acoustic detections in SNE to peak during the winter and early spring (January through March). Visual surveys (Quintana-Rizzo
et al.,
2021) have also confirmed the abundance of NARWs in SNE to be the highest during the winter and spring (January through May), although peaks in acoustic detections may vary seasonally across years (Quintana-Rizzo
et al.,
2021; Estabrook
et al.,
2022). Distribution throughout SNE may vary seasonally with NARW occurrence being closest to the LIA during the spring (Quintana-Rizzo
et al.,
2021). Van Parijs
et al.
(2023) monitored acoustic detections of baleen whales throughout SNE and detected NARWs near the LIA from January through May. Acoustic detections began to increase near the LIA in November and further increased into December (Van Parijs
et al.,
2023).
An 8-year analysis of NARW sightings within SNE showed that the NARW distribution has been shifting (Quintana-Rizzo
et al.,
2021). NARWs 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 NARW 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; Pendleton
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). Pendleton
et al.
(2022) observed shifts in the timing of NARW peak habitat use in Cape Cod Bay during the spring, likely in response to changing seasonal conditions, and characterized SNE as a “waiting room” for NARWs in the spring, providing sufficient, although sub-optimal, prey choices while the NARWs wait for foraging conditions in Cape Cod Bay (and other primary foraging grounds such as the Great South Channel) to optimize as seasonal primary and secondary production progresses.
While Nantucket Shoals is not designated as critical NARW habitat, its importance as a foraging habitat is well established (Leiter
et al.,
2017; Quintana-Rizzo
et al.,
2021; Estabrook
et al.,
2022; O'Brien
et al.,
2022). Nantucket Shoals' unique oceanographic and bathymetric features, including a persistent tidal front, help sustain year-round elevated phytoplankton biomass, and aggregate zooplankton prey for NARWs (Quintana-Rizzo
et al.,
2021). SNE serves as a foraging habitat throughout the year, although not to the extent provided seasonally in more well-understood feeding habitats like Cape Cod Bay in late spring, the Great South Channel, and the Gulf of St. Lawrence (O'Brien
et al.,
2022). A BIA for foraging (LaBrecque
et al.,
2015) within Cape Cod Bay is approximately 71 km (44.1 mi) north of the LIA, while critical habitat northeast of Martha's Vineyard and Nantucket Island is within 56 km (34.8 mi). SNE also represents socializing habitat for NARWs as Leiter
et al.
(2017) documented surface active groups (SAGs), indicative of socializing behavior, year-round in SNE.
Observations of NARW transitions in 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). In addition, extensive data gaps that were highlighted in a recent report by the National Academy of Sciences (NAS, 2023) have prevented development of a thorough understanding of NARW foraging ecology in the Nantucket Shoals region. However, it is clear that the habitat was historically valuable to the species, given that the whaling industry capitalized on consistent NARW occurrence there, and has again become increasingly so over the last decade.
Since 2017, 125 dead, seriously injured, or sublethally injured or ill NARWs along the United States and Canadian coasts have been documented, necessitating a UME declaration in 2017 and subsequent investigation. The leading category for the cause of death for this ongoing UME is “human interaction,” specifically from entanglements or vessel strikes. As of April 9, 2024, there have been 39 confirmed mortalities, 1 pending mortality (dead, stranded, or floaters), and 34 seriously injured free-swimming whales for a total of 73 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 125. 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; Pirotta
et al.,
2024). Pirotta
et al.
(2024) found an association between the decreased mean length of female NARWs and reduced calving probability. More information about the NARW UME is available online at
https://www.fisheries.noaa.gov/national/marine-life-distress/2017-2024-north-atlantic-right-whale-unusual-mortality-event.
On August 1, 2022, NMFS announced proposed changes to the existing NARW vessel speed regulations to further reduce the likelihood of mortalities and serious injuries to endangered right whales from vessel collisions, which are a leading cause of the species' decline and a primary factor in the ongoing Unusual Mortality Event (87 FR 46921, August 1, 2022). Should a final vessel speed rule be issued and become effective during the effective period of this IHA (or any other MMPA incidental take authorization), the authorization holder would be required to comply with any and all applicable requirements contained within the final rule. Specifically, where measures in any final vessel speed rule are more protective or restrictive than those in this or any other MMPA authorization, authorization holders would be required to comply with the requirements of the rule. Alternatively, where measures in this or any other MMPA authorization are more restrictive or protective than those in any final vessel speed rule, the measures in the MMPA authorization would remain in place. These changes would become effective immediately upon the effective date of any final vessel speed rule and would not require any further action on NMFS's part.
Humpback Whale
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 LIA. Bettridge
et al.
(2015) estimated the size of the West Indies DPS population at 12,312 (95 percent confidence interval 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).
In New England waters, feeding is the principal activity of humpback whales, and their distribution in this region has been largely correlated to abundance of prey species, although behavior and bathymetry are factors influencing foraging strategy (Payne
et al.,
1986, 1990). Humpback whales are frequently piscivorous when in New England waters, feeding on herring (
Clupea harengus
), sand lance (
Ammodytes
spp.), and other small fishes, as well as euphausiids in the northern Gulf of Maine (Paquet
et al.,
1997). During winter, the majority of humpback whales from North Atlantic feeding areas (including the Gulf of Maine) mate and calve in the West Indies, where spatial and genetic mixing among feeding groups occurs, though significant numbers of animals are found in mid- and high-latitude regions at this time and some individuals have been sighted repeatedly within the same winter season, indicating that not all humpback whales migrate south every winter (Hayes
et al.,
2018).
Kraus
et al.
(2016) conducted aerial surveys from 2011-2015 in SNE and observed humpback whales during all seasons, yet humpback whales were observed most often during the spring and summer. The greatest number of sightings occurred during the month of April (
n
=33) (Kraus
et al.,
2016). Calves, feeding behavior, and courtship behavior were observed as well. More recent studies (O'Brien
et al.,
2020, 2021, 2022, 2023) confirm that humpback whales peak in abundance in the LIA during spring and summer, with the majority of sightings year-round occurring in the eastern portion of the MA and RI/MA WEAs and near the Nantucket Shoals area (O'Brien
et al.,
2020). O'Brien
et al.
(2022) identified seasonal distribution patterns of humpback whales throughout SNE with more concentrated sightings near Nantucket Shoals in the fall and sightings being distributed more evenly across the MA and RI/MA WEAs during spring and summer. As observed during the 2011-2015 surveys, O'Brien
et al.
(2023) also observed feeding behavior and mother/calf pairs throughout the spring and summer. Van Parijs
et al.
(2023) detected humpback whales near the LIA mainly from November through June. During the Vineyard Wind 2023 construction campaign, visual and acoustic detections of humpback whales occurred mainly from June through October, with the greatest detections occuring in October (Vineyard Wind, 2023).
The LIA 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). This BIA is located approximately 73 km (45.5 mi) northeast of the Lease Area and would not likely be impacted by project activities.
Since January 2016, elevated humpback whale mortalities along the Atlantic coast from Maine to Florida led to the declaration of a UME in April 2017. As of April 9, 2024, 218 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. Since January 1, 2023, 43 humpbacks have stranded along the east coast of the United States (7 of these whales have stranded off Massachusetts). 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 Whale
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 north 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).
New England waters represent a major feeding ground for fin whales, and 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
). Aerial surveys conducted from 2011-2015 in SNE documented fin whale occurrence in every season, with the greatest numbers of sightings during the spring (
n
=35) and summer (
n
=49) months (Kraus
et al.,
2016). Fin whale distribution varied seasonally, with fin whales occurring in the southern regions of the MA and RI/MA WEAs during spring and closer to northern regions of the WEAs during summer (Kraus
et al.,
2016). More recent surveys have documented fin whales throughout winter, spring, and summer (O'Brien
et al.,
2020, 2021, 2022, 2023) with the greatest abundance occurring during the summer and clustered in the western portion of the WEAs (O'Brien
et al.,
2023). Acoustic detection of fin whales in SNE indicate fin whale presence in the area from August through April and, sporadically, from May through July (Parijs
et al.,
2023). During the 2023 construction campaign, Vineyard Wind detected fin whales from June through December (with the exception of August), with the most detections occurring in October (Vineyard Wind, 2023). Based upon observations of feeding behavior and the close proximity of the Lease Area to the
feeding BIAs (8.0 km (5.0 mi) and 76.4 km (47.5 mi) to the Montauk Point and southern Gulf of Maine BIAs, respectively) fin whales may use the LIA for foraging as well as migrating.
Minke Whale
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 LIA, 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). The nearest BIA is approximately 44.0 km (27.3 mi) northeast of the Lease Area. Due to the close proximity of the BIA, minke whale feeding may occur within the LIA.
Although minke whales are sighted in every season in SNE (O'Brien
et al.,
2022), minke whale use of the area is highest during the months of March through September (Kraus
et al.,
2016; O'Brien
et al.,
2023). Large feeding aggregations of humpback, fin, and minke whales have been observed during the summer (O'Brien
et al.,
2023), suggesting the LIA may serve as a supplemental feeding grounds for these species. Acoustic detections data support visual sighting data, and indicate minke whale presence in SNE from March through June and August through late November/early December and, sporadically, in January (Parijs
et al.,
2023). During the 2023 construction campaign, Vineyard Wind detected minke whales from June through August (Vineyard Wind, 2023).
From 2017 through 2024, elevated minke whale mortalities detected along the Atlantic coast from Maine through South Carolina resulted in the declaration of a UME in 2018. As of April 9, 2024, a total of 166 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.
Sei Whale
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). Sei whale migratory movements are not well understood. In June and July, sei whales are believed to migrate north from SNE to feeding areas in eastern Canada, 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, approximately 101.4 km (63 mi) east of the LIA (LaBrecque
et al.,
2015).
Aerial surveys conducted from 2011-2015 in SNE observed sei whales between March and June, with the greatest number of sightings occurring in May (
n
=8) and June (
n
=13), and no sightings from July through January (Kraus
et al.,
2016). Acoustic detections confirm peak occurrences of sei whales in SNE from early spring and through mid-summer (March through July) (Davis
et al.,
2020). In addition, Van Parijs
et al.
(2023) acoustically detected sei whales near the LIA during the months of February and August. However, Davis
et al.
(2020) acoustically detected sei whales in SNE year-round, suggesting this area is an important habitat for sei whales. As sei whales are known to target the prey such as copepods (
C. finmarchicus
), which are abundant in SNE waters (Quintana-Rizzo
et al.,
2018), SNE likely represents a supplemental foraging area for sei whales as well.
Phocid Seals
Harbor and gray seals have experienced multiple UMEs since 2018. From June through July 2022, elevated numbers of harbor seal and gray seal mortalities occurred across the southern and central coast of Maine. This event was declared a UME. During the event, 181 seals stranded. Based upon necropsy, histopathology, and diagnostic findings, this UME was attributed to spillover events of the highly pathogenic avian influenza from infected birds to harbor and gray seals. While the UME did not occur in the LIA, the populations that were affected by the UME are the same as those potentially affected by the project. This UME has recently been closed. 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.
Marine Mammal Hearing
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, 2019) 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 4.
Table 4—Marine Mammal Hearing Groups
[NMFS, 2018]
Hearing group
Generalized hearing range *
Low-frequency (LF) cetaceans (baleen whales)
7 Hz to 35 kHz.
Mid-frequency (MF) cetaceans (dolphins, toothed whales, beaked whales, bottlenose whales)
150 Hz to 160 kHz.
High-frequency (HF) cetaceans (true porpoises,
Kogia,
river dolphins,
Cephalorhynchid,
Lagenorhynchus cruciger
&
L. australis
)
275 Hz to 160 kHz.
Phocid pinnipeds (PW) (underwater) (true seals)
50 Hz to 86 kHz.
Otariid pinnipeds (OW) (underwater) (sea lions and fur seals)
60 Hz to 39 kHz.
* Represents the generalized hearing range for the entire group as a composite (
i.e.,
all species within the group), where individual species' hearing ranges are typically not as broad. Generalized hearing range chosen based on the ~65-dB threshold from normalized composite audiogram, with the exception for lower limits for LF cetaceans (Southall
et al.,
2007) and PW pinniped (approximation).
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
et al.,
2013).
For more detail concerning these groups and associated frequency ranges, please see NMFS (2018) for a review of available information.
Potential Effects of Specified Activities on Marine Mammals and Their Habitat
This section provides a discussion of the ways in which 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 whether those impacts are reasonably expected to, or reasonably likely to, adversely affect the species or stock through effects on annual rates of recruitment or survival.
Vineyard Wind has requested, and NMFS proposes to authorize, the take of marine mammals incidental to the construction activities associated with the LIA. In their application, Vineyard Wind presented their analyses of potential impacts to marine mammals from the acoustic sources. NMFS carefully reviewed the information provided by Vineyard 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 15 permanent foundations to support WTGs. 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.
Description of Sound Sources
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 is 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 IHA, 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 permanent threshold shift (PTS) and temporary threshold shift (TTS).
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 a 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.
Potential Effects of Underwater Sound on Marine Mammals
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 Vineyard 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 Vineyard 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).
Hearing Threshold Shift
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, 2015; Nachtigall
et al.,
2016a-c, 2018; Finneran, 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).
Behavioral Effects
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 (2.1-5.9 mi) 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, or 73.3 mi, 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-b, 2018; Falcone
et al.,
2017; 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 IHA 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 five-fold 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 and Displacement
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 (3.1-6.2 mi) of the pile driving site, with declines at up to 20-30 km (12.4-18.6 mi) of the pile driving site documented in some cases. Similar results demonstrating the long-distance displacement of harbor porpoises (18-25 km; 11.1-15.5 mi) and harbor seals (up to 40 km (24.9 mi)) 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; 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), or the manner in which species use the habitat in the LIA, are likely the driving factors of this variation.
NMFS notes that the aforementioned European studies involved installing much smaller monopiles than Vineyard Wind proposes to install (Brandt
et al.,
2016) and, therefore we anticipate noise levels from impact pile driving to be louder. However, we do not anticipate any greater severity of response due to harbor porpoise and harbor seal habitat use off Massachusetts 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 Massachusetts, harbor porpoises and seals are more transient, and a very small percentage of the harbor seal population are only seasonally present with no rookeries established (Hayes
et al.,
2022). 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 (2.1-5.9 mi) away while showing no such reaction to the same signal when the signal was 118 km (73.3 mi) 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 NARWs, gray whales migrate close to shore (approximately +2 km (+1.2 mi)) and are low-frequency hearing specialists. The LFA sonar source was placed within the gray whale migratory corridor (approximately 2 km (1.2 mi) offshore) and offshore of most, but not all, migrating whales (approximately 4 km (2.5 mi) 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 levels were 170 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 (+0.6 mi). 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 (12.4 mi) from the area (Boyd
et al.,
2008; Southall
et al.,
2009; Tyack
et al.,
2011). In contrast, the sounds produced by pile driving activities 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 stationery pile driving (which they can sense is stationery and predictable), unless they are within the area ensonified above behavioral harassment thresholds at the moment the pile driving begins (Watkins, 1986; Falcone
et al.,
2017).
Diving and Foraging
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 NARWs 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 NARWs, 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 year that the proposed IHA 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 (4.3-8.1 mi), following a phase-in of sound intensity and full array exposures at 1-13 km (0.6-8.1 mi) (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 NARWs 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 noise generated by Vineyard Wind's proposed activities would at least partially overlap in frequency with signals described by Nowacek
et al.
(2004) and Croll
et al.
(2001). 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 of tagged blue whales in Southern California waters indicated that, in some cases and at low received levels, the 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, 2012b, 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 by an activity that influences foraging disruption, 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).
Vocalizations and Auditory Masking
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 NARWs 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; Sorensen
et al.,
2023). 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. Even when animals attempt to compensate for masking, such as by increasing the amplitude or duration of their signals, this may still be insufficient to maintain behavioral coordination between individuals necessary for complex behaviors, foraging, and navigation (Sorensen
et al.,
2023). 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, 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 NARW 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 (1.9 mi) to over 9 km (5.6 mi) (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, 2011). Clark
et al.
(2009) observed that right whales' communication space decreased by up to 84 percent in the presence of vessels due to an increase in ambient noise from vessels in proximity to the whales. Cholewiak
et al.
(2018) also observed loss in communication space in Stellwagen National Marine Sanctuary for NARWs, 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, 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 and Sensitization
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). Habituation is considered 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). Animals are most likely to habituate to sounds that are predictable and unvarying. 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-b; 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, 2012, 2014, 2018; Dähne
et al.,
2013; Russell
et al.,
2016).
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 before, during, and after seismic surveys (Gailey
et al.,
2016). Behavioral state and water depth were the best “natural” predictors of whale movements and respiration, and after accounting for 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.
Physiological Responses
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 response (in terms of energetic costs) 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 specifically 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 NARWs.
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, 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, 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.
Stranding
The definition for a stranding under 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 Stranding and Mortality discussion in NMFS' proposed rule for the Navy's Training and Testing Activities in the Hawaii-Southern California Training and Testing Study Area (83 FR 29872, 29928; June 26, 2018).
The construction activities proposed by Vineyard Wind (
i.e.,
pile driving) are not expected to result in marine mammal strandings. Of the strandings documented to date worldwide, NMFS is not aware of any being attributed to pile driving. While vessel strikes could kill or injure a marine mammal (which may then eventually strand), the required mitigation measures would reduce the potential for take from these activities to de minimis 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.
Potential Effects of Disturbance on Marine Mammal Fitness
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), outlined 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, NARWs, 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
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