Notice2024-08434
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
Primary source
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Published
April 23, 2024
Issuing agencies
Commerce DepartmentNational Oceanic and Atmospheric Administration
Abstract
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.
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[Federal Register Volume 89, Number 79 (Tuesday, April 23, 2024)]
[Notices]
[Pages 31008-31064]
From the Federal Register Online via the Government Publishing Office [<a href="http://www.gpo.gov">www.gpo.gov</a>]
[FR Doc No: 2024-08434]
[[Page 31007]]
Vol. 89
Tuesday,
No. 79
April 23, 2024
Part V
Department of Commerce
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National Oceanic and Atmospheric Administration
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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; Notice
Federal Register / Vol. 89, No. 79 / Tuesday, April 23, 2024 /
Notices
[[Page 31008]]
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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.
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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 <a href="/cdn-cgi/l/email-protection#d49d8084faa0b5adb8bba694babbb5b5fab3bba2"><span class="__cf_email__" data-cfemail="c78e9397e9b3a6beaba8b587a9a8a6a6e9a0a8b1">[email protected]</span></a>.
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: <a href="https://www.fisheries.noaa.gov/national/marine-mammal-protection/incidental-take-authorizations-other-energy-activities-renewable">https://www.fisheries.noaa.gov/national/marine-mammal-protection/incidental-take-authorizations-other-energy-activities-renewable</a>. 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 <a href="https://www.fisheries.noaa.gov/national/marine-mammal-protection/incidental-take-authorizations-other-energy-activities-renewable">https://www.fisheries.noaa.gov/national/marine-mammal-protection/incidental-take-authorizations-other-energy-activities-renewable</a> 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: <a href="https://www.boem.gov/renewable-energy/state-activities/vineyard-wind-1">https://www.boem.gov/renewable-energy/state-activities/vineyard-wind-1</a>. 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
[[Page 31009]]
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.
[[Page 31010]]
[GRAPHIC] [TIFF OMITTED] TN23AP24.040
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
[[Page 31011]]
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
----------------------------------------------------------------------------------------------------------------
Max piling Max piling
Number of time time
Pile type Project Max hammer hammer duration duration Number
component energy (kJ) strikes per pile per day piles/day
(min) (min)
----------------------------------------------------------------------------------------------------------------
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
[[Page 31012]]
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
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Expected
Maximum number maximum number
Vessel type Vessel role of vessels of transits Port
per month
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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 1 4
vessel.
Safety vessel........ 4 2
Motor vessel...................... Crew transfer vessel. 2 12
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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; <a href="https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments">https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments</a>) and more general information about these species
(e.g., physical and behavioral descriptions) may be found on NMFS'
website (<a href="https://www.fisheries.noaa.gov/find-species">https://www.fisheries.noaa.gov/find-species</a>).
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: <a href="https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments">https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments</a>.
[[Page 31013]]
Table 3--Marine Mammal Species That May Occur in the LIA and Be Taken by Harassment
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ESA/MMPA Stock abundance (CV,
status; Nmin, most recent Annual M/SI
Common name \a\ Scientific name Stock strategic (Y/N) abundance survey) PBR \d\
\b\ \c\
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Order Artiodactyla--Cetacea--Mysticeti (baleen whales)
--------------------------------------------------------------------------------------------------------------------------------------------------------
Family Balaenidae:
NARW......................... Eubalaena glacialis............ Western Atlantic.... E, D, Y 340 (0; 337; 2021) 0.7 27.2 \f\
\e\.
Family Balaenopteridae
(rorquals):
Fin whale.................... Balaenoptera physalus.......... Western North E, D, Y 6,802 (0.24, 5,573, 11 2.05
Atlantic. 2021).
Sei whale.................... Balaenoptera borealis.......... Nova Scotia......... E, D, Y 6,292 (1.02, 3098, 6.2 0.6
2021).
Minke whale.................. Balaenoptera acutorostrata..... Canadian Eastern -, -, N 21,968 (0.31, 170 9.4
Coastal. 17,002, 2021).
Humpback whale............... Megaptera novaeangliae......... Gulf of Maine....... -, -, Y 1,396 (0, 1,380, 22 12.15
2016).
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Superfamily Odontoceti (toothed whales, dolphins, and porpoises)
--------------------------------------------------------------------------------------------------------------------------------------------------------
Family Physeteridae:
Sperm whale.................. Physeter macrocephalus......... North Atlantic...... E, D, Y 5,895 (0.29, 4,639, 9.28 0.2
2021).
Family Delphinidae:
Long-finned pilot whale...... Globicephala melas............. Western North -, -, N 39,215 (0.3, 30,627, 306 5.7
Atlantic. 2021).
Bottlenose dolphin........... Tursiops truncatus............. Western North -, -, N 64,587 (0.24, 507 28
Atlantic Offshore. 52,801, 2021) \g\.
Common dolphin............... Delphinus delphis.............. Western North -, -, N 93,100 (0.56, 1,452 414
Atlantic. 59,897, 2021).
Risso's dolphin.............. Grampus griseus................ Western North -, -, N 44,067 (0.19, 307 18
Atlantic. 30,662, 2021).
Atlantic white-sided dolphin. Lagenorhynchus acutus.......... Western North -, -, N 93,233 (0.71, 544 28
Atlantic. 54,443, 2021).
Family Phocoenidae (porpoises):
Harbor porpoise.............. Phocoena phocoena.............. Gulf of Maine/Bay of -, -, N 85,765 (0.53, 649 145
Fundy. 56,420, 2021).
--------------------------------------------------------------------------------------------------------------------------------------------------------
Order Carnivora--Pinnipedia
--------------------------------------------------------------------------------------------------------------------------------------------------------
Family Phocidae (earless seals):
Harbor seal.................. Phoca vitulina................. Western North -, -, N 61,336 (0.08, 1,729 339
Atlantic. 57,637, 2018).
Gray seal \h\................ Halichoerus grypus............. Western North -, -, N 27,911 (0.2, 23,924, 1,512 4,570
Atlantic. 2021).
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\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
(<a href="https://marinemammalscience.org/science-and-publications/list-marine-mammal-species-subspecies">https://marinemammalscience.org/science-and-publications/list-marine-mammal-species-subspecies</a>; 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: <a href="https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments">https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments</a>. CV is the
coefficient of variation; Nmin 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 (Nbest 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
[[Page 31014]]
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 (<a href="https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments">https://www.fisheries.noaa.gov/national/marine-mammal-protection/marine-mammal-stock-assessments</a>), and NMFS' website (<a href="https://www.fisheries.noaa.gov/species-directory/marine-mammals">https://www.fisheries.noaa.gov/species-directory/marine-mammals</a>), 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; <a href="https://oceannoise.noaa.gov/biologically-important-areas">https://oceannoise.noaa.gov/biologically-important-areas</a>) (Van Parijs, 2015)).
There are no ESA-designated critical habitats for any species within
the LIA (<a href="https://www.fisheries.noaa.gov/resource/map/national-esa-critical-habitat-mapper">https://www.fisheries.noaa.gov/resource/map/national-esa-critical-habitat-mapper</a>). 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 <a href="https://www.fisheries.noaa.gov/national/marine-life-distress/active-and-closed-unusual-mortality-events">https://www.fisheries.noaa.gov/national/marine-life-distress/active-and-closed-unusual-mortality-events</a>.
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
[[Page 31015]]
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 <a href="https://www.fisheries.noaa.gov/national/marine-life-distress/2017-2024-north-atlantic-right-whale-unusual-mortality-event">https://www.fisheries.noaa.gov/national/marine-life-distress/2017-2024-north-atlantic-right-whale-unusual-mortality-event</a>.
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
[[Page 31016]]
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 <a href="https://www.fisheries.noaa.gov/national/marine-life-distress/active-and-closed-unusual-mortality-events">https://www.fisheries.noaa.gov/national/marine-life-distress/active-and-closed-unusual-mortality-events</a>.
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
[[Page 31017]]
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 <a href="https://www.fisheries.noaa.gov/national/marine-life-distress/2017-2022-minke-whale-unusual-mortality-event-along-atlantic-coast">https://www.fisheries.noaa.gov/national/marine-life-distress/2017-2022-minke-whale-unusual-mortality-event-along-atlantic-coast</a>.
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 <a href="https://www.fisheries.noaa.gov/2022-2023-pinniped-unusual-mortality-event-along-maine-coast">https://www.fisheries.noaa.gov/2022-2023-pinniped-unusual-mortality-event-along-maine-coast</a>.
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
[[Page 31018]]
in this UME, which is pending closure. Information on this UME is
available online at: <a href="https://www.fisheries.noaa.gov/new-england-mid-atlantic/marine-life-distress/2018-2020-pinniped-unusual-mortality-event-along">https://www.fisheries.noaa.gov/new-england-mid-atlantic/marine-life-distress/2018-2020-pinniped-unusual-mortality-event-along</a>.
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 7 Hz to 35 kHz.
whales).
Mid-frequency (MF) cetaceans 150 Hz to 160 kHz.
(dolphins, toothed whales, beaked
whales, bottlenose whales).
High-frequency (HF) cetaceans (true 275 Hz to 160 kHz.
porpoises, Kogia, river dolphins,
Cephalorhynchid, Lagenorhynchus
cruciger & L. australis).
Phocid pinnipeds (PW) (underwater) 50 Hz to 86 kHz.
(true seals).
Otariid pinnipeds (OW) (underwater) 60 Hz to 39 kHz.
(sea lions and fur seals).
------------------------------------------------------------------------
* 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
(Hemil[auml] 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
<a href="https://www.dosits.org">https://www.dosits.org</a>. 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
[[Page 31019]]
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
([mu]Pa). For in-air sound, the reference pressure is 20 microPascal
([mu]Pa). 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) 1[mu]Pa.
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 [mu]Pa\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<INF>ss</INF>) or calculated over periods containing multiple pulses
(SEL<INF>cum</INF>). 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
[[Page 31020]]
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; G[ouml]tz 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
[[Page 31021]]
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 1[micro]Pa
(microPascal)) for exposures to 3-4 kHz sonar signals, while deep
feeding and non-feeding whales showed a clear response at exposures at
lower
[[Page 31022]]
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 1[micro]Pa) 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 1[micro]Pa)
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; D[auml]hne 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; D[auml]hne 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
[[Page 31023]]
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 S[uuml]d/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;
D[auml]hne 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
[[Page 31024]]
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
[[Page 31025]]
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 (Melc[oacute]n et al., 2012).
However, Melc[oacute]n 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 [micro]Pa, 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
[[Page 31026]]
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; Cur[eacute] 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 (Cur[eacute] et al.,
2016; Isojunno et al., 2016), long-finned pilot whales (Visser et al.,
2016), and humpback whales (Cur[eacute] 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
[[Page 31027]]
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.
[[Page 31028]]
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; D[auml]hne 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
[[Page 31029]]
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
[[Page 31030]]
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 to seismic surveys using both a forwards and backwards
approach. While a typical forwards approach can determine if a stressor
would have population-level consequences, Dunlop et al. demonstrated
that working backwards through a PCoD model can be used to assess the
most unfavorable scenario for an interaction of a target species and
stressor. This method may be useful for future management goals when
appropriate data becomes available to fully support the model. In
another example, harbor porpoise PCoD model investigating the impact of
seismic surveys on harbor porpoise included an investigation on
underlying drivers of vulnerability. Harbor porpoise movement and
foraging were modeled for baseline periods and then for periods with
seismic surveys as well; the models demonstrated that temporal (i.e.,
seasonal) variation in individual energetics and their link to costs
associated with disturbances was key in predicting population impacts
(Gallagher et al., 2021).
Behavioral change, such as disturbance manifesting in lost foraging
time, in response to anthropogenic activities is often assumed to
indicate a biologically significant effect on a population of concern.
However, as described above, individuals may be able to compensate for
some types and degrees of shifts in behavior, preserving their health
and thus their vital rates and population dynamics. For example,
[[Page 31031]]
New et al. (2013) developed a model simulating the complex social,
spatial, behavioral, and motivational interactions of coastal
bottlenose dolphins in the Moray Firth, Scotland, to assess the
biological significance of increased rate of behavioral disruptions
caused by vessel traffic. Despite a modeled scenario in which vessel
traffic increased from 70 to 470 vessels a year (a six-fold increase in
vessel traffic) in response to the construction of a proposed offshore
renewables' facility, the dolphins' behavioral time budget, spatial
distribution, motivations, and social structure remain unchanged.
Similarly, two bottlenose dolphin populations in Australia were also
modeled over 5 years against a number of disturbances (Reed et al.,
2020), and results indicated that habitat/noise disturbance had little
overall impact on population abundances in either location, even in the
most extreme impact scenarios modeled.
By integrating different sources of data (e.g., controlled exposure
data, activity monitoring, telemetry tracking, and prey sampling) into
a theoretical model to predict effects from sonar on a blue whale's
daily energy intake, Pirotta et al. (2021) found that tagged blue
whales' activity budgets, lunging rates, and ranging patterns caused
variability in their predicted cost of disturbance. This method may be
useful for future management goals when appropriate data becomes
available to fully support the model. Harbor porpoise movement and
foraging were modeled for baseline periods and then for periods with
seismic surveys as well; the models demonstrated that the seasonality
of the seismic activity was an important predictor of impact (Gallagher
et al., 2021).
In their table 1, Keen et al. (2021) summarize the emerging themes
in PCoD models that should be considered when assessing the likelihood
and duration of exposure and the sensitivity of a population to
disturbance (see table 1 from Keen et al., 2021, below). The themes are
categorized by life history traits (movement ecology, life history
strategy, body size, and pace of life), disturbance source
characteristics (overlap with biologically important areas, duration
and frequency, and nature and context), and environmental conditions
(natural variability in prey availability and climate change). Keen et
al. (2021) then summarize how each of these features influence an
assessment, noting, for example, that individual animals with small
home ranges have a higher likelihood of prolonged or year-round
exposure, that the effect of disturbance is strongly influenced by
whether it overlaps with biologically important habitats when
individuals are present, and that continuous disruption will have a
greater impact than intermittent disruption.
Nearly all PCoD studies and experts agree that infrequent exposures
of a single day or less are unlikely to impact individual fitness, let
alone lead to population level effects (Booth et al., 2016; Booth et
al., 2017; Christiansen and Lusseau, 2015; Farmer et al., 2018; Wilson
et al., 2020; Harwood and Booth, 2016; King et al., 2015; McHuron et
al., 2018; National Academies of Sciences, Engineering, and Medicine
(NAS), 2017; New et al., 2014; Pirotta et al., 2018a; Southall et al.,
2007; Villegas-Amtmann et al., 2015). As described through this notice
for the proposed IHA, NMFS expects that any behavioral disturbance that
would occur due to animals being exposed to construction activity would
be of a relatively short duration, with behavior returning to a
baseline state shortly after the acoustic stimuli ceases or the animal
moves far enough away from the source. Given this, and NMFS' evaluation
of the available PCoD studies, and the required mitigation discussed
later, any such behavioral disturbance resulting from Vineyard Wind's
activities is not expected to impact individual animals' health or have
effects on individual animals' survival or reproduction, thus no
detrimental impacts at the population level are anticipated. Marine
mammals may temporarily avoid the immediate area but are not expected
to permanently abandon the area or their migratory or foraging
behavior. Impacts to breeding, feeding, sheltering, resting, or
migration are not expected nor are shifts in habitat use, distribution,
or foraging success.
Potential Effects From Vessel Strike
Vessel collisions with marine mammals, also referred to as vessel
strikes or ship strikes, can result in death or serious injury of the
animal. Wounds resulting from ship strike may include massive trauma,
hemorrhaging, broken bones, or propeller lacerations (Knowlton and
Kraus, 2001). An animal at the surface could be struck directly by a
vessel, a surfacing animal could hit the bottom of a vessel, or an
animal just below the surface could be cut by a vessel's propeller.
Superficial strikes may not kill or result in the death of the animal.
Lethal interactions are typically associated with large whales, which
are occasionally found draped across the bulbous bow of large
commercial ships upon arrival in port. Although smaller cetaceans are
more maneuverable in relation to large vessels than are large whales,
they may also be susceptible to strike. The severity of injuries
typically depends on the size and speed of the vessel (Knowlton and
Kraus, 2001; Laist et al., 2001; Vanderlaan and Taggart, 2007; Conn and
Silber, 2013), although Kelley et al. (2020) found, through the use of
a simple biophysical model, that large whales can be seriously injured
or killed by vessels of all sizes. Impact forces increase with speed,
as does the probability of a strike at a given distance (Silber et al.,
2010; Gende et al., 2011).
The most vulnerable marine mammals are those that spend extended
periods of time at the surface in order to restore oxygen levels within
their tissues after deep dives (e.g., the sperm whale). In addition,
some baleen whales seem generally unresponsive to vessel sound, making
them more susceptible to vessel collisions (Nowacek et al., 2004).
These species are primarily large, slow-moving whales. Marine mammal
responses to vessels may include avoidance and changes in dive pattern
(NRC, 2003).
An examination of all known ship strikes from all shipping sources
(civilian and military) indicates vessel speed is a principal factor in
whether a vessel strike occurs and, if so, whether it results in
injury, serious injury, or mortality (Knowlton and Kraus, 2001; Laist
et al., 2001; Jensen and Silber, 2003; Pace and Silber, 2005;
Vanderlaan and Taggart, 2007; Conn and Silber, 2013). In assessing
records in which vessel speed was known, Laist et al. (2001) found a
direct relationship between the occurrence of a whale strike and the
speed of the vessel involved in the collision. The authors concluded
that most deaths occurred when a vessel was traveling in excess of 13
kn.
Jensen and Silber (2003) detailed 292 records of known or probable
ship strikes of all large whale species from 1975 to 2002. Of these,
vessel speed at the time of collision was reported for 58 cases. Of
these 58 cases, 39 (or 67 percent) resulted in serious injury or death
(19 of those resulted in serious injury as determined by blood in the
water, propeller gashes or severed tailstock, and fractured skull, jaw,
vertebrae, hemorrhaging, massive bruising, or other injuries noted
during necropsy and 20 resulted in death). Operating speeds of vessels
that struck various species of large whales ranged from 2 to 51 kn. The
majority (79 percent) of these strikes occurred at speeds of 13 kn or
greater. The average speed that resulted in serious injury or death was
18.6 kn. Pace and Silber (2005) found that the probability of death or
serious injury increased rapidly with increasing vessel speed.
[[Page 31032]]
Specifically, the predicted probability of serious injury or death
increased from 45 to 75 percent as vessel speed increased from 10 to 14
kn and exceeded 90 percent at 17 kn. Higher speeds during collisions
result in greater force of impact and also appear to increase the
chance of severe injuries or death. While modeling studies have
suggested that hydrodynamic forces pulling whales toward the vessel
hull increase with increasing speed (Clyne, 1999; Knowlton et al.,
1995), this is inconsistent with Silber et al. (2010), which
demonstrated that there is no such relationship (i.e., hydrodynamic
forces are independent of speed).
In a separate study, Vanderlaan and Taggart (2007) analyzed the
probability of lethal mortality of large whales at a given speed,
showing that the greatest rate of change in the probability of a lethal
injury to a large whale as a function of vessel speed occurs between
8.6 and 15 kn. The chances of a lethal injury decline from
approximately 80 percent at 15 kn to approximately 20 percent at 8.6
kn. At speeds below 11.8 kn, the chances of lethal injury drop below 50
percent, while the probability asymptotically increases toward 100
percent above 15 kn.
The Jensen and Silber (2003) report notes that the Large Whale Ship
Strike Database represents a minimum number of collisions, because the
vast majority probably goes undetected or unreported. In contrast, the
Project's personnel are likely to detect any strike that does occur
because of the required personnel training and lookouts, along with the
inclusion of PSOs (as described in the Proposed Mitigation section),
and they are required to report all ship strikes involving marine
mammals.
There are no known vessel strikes of marine mammals by any offshore
wind energy vessel in the United States. Given the extensive mitigation
and monitoring measures (see the Proposed Mitigation and Proposed
Monitoring and Reporting section) that would be required of Vineyard
Wind, NMFS believes that a vessel strike is not likely to occur.
Potential Effects to Marine Mammal Habitat
Vineyard Wind's proposed activities could potentially affect marine
mammal habitat through impacts on the prey species of marine mammals
(through noise, oceanographic processes, or reef effects), acoustic
habitat (sound in the water column), water quality, and biologically
important habitat for marine mammals.
Effects on Prey
Sound may affect marine mammals through impacts on the abundance,
behavior, or distribution of prey species (e.g., crustaceans,
cephalopods, fish, and zooplankton). Marine mammal prey varies by
species, season, and location and, for some, is not well documented.
Here, we describe studies regarding the effects of noise on known
marine mammal prey.
Fish utilize the soundscape and components of sound in their
environment to perform important functions such as foraging, predator
avoidance, mating, and spawning (e.g., Zelick and Mann, 1999; Fay,
2009). The most likely effects on fishes exposed to loud, intermittent,
low-frequency sounds are behavioral responses (i.e., flight or
avoidance). Short duration, sharp sounds (such as pile driving or
airguns) can cause overt or subtle changes in fish behavior and local
distribution. The reaction of fish to acoustic sources depends on the
physiological state of the fish, past exposures, motivation (e.g.,
feeding, spawning, migration), and other environmental factors. Key
impacts to fishes may include behavioral responses, hearing damage,
barotrauma (pressure-related injuries), and mortality. While it is
clear that the behavioral responses of individual prey, such as
displacement or other changes in distribution, can have direct impacts
on the foraging success of marine mammals, the effects on marine
mammals of individual prey that experience hearing damage, barotrauma,
or mortality is less clear, though obviously population scale impacts
that meaningfully reduce the amount of prey available could have more
serious impacts.
Fishes, like other vertebrates, have a variety of different sensory
systems to glean information from ocean around them (Astrup and Mohl,
1993; Astrup, 1999; Braun and Grande, 2008; Carroll et al., 2017;
Hawkins and Johnstone, 1978; Ladich and Popper, 2004; Ladich and
Schulz-Mirbach, 2016; Mann, 2016; Nedwell et al., 2004; Popper et al.,
2003, 2005). Depending on their hearing anatomy and peripheral sensory
structures, which vary among species, fishes hear sounds using pressure
and particle motion sensitivity capabilities and detect the motion of
surrounding water (Fay et al., 2008) (terrestrial vertebrates generally
only detect pressure). Most marine fishes primarily detect particle
motion using the inner ear and lateral line system while some fishes
possess additional morphological adaptations or specializations that
can enhance their sensitivity to sound pressure, such as a gas-filled
swim bladder (Braun and Grande, 2008; Popper and Fay, 2011).
Hearing capabilities vary considerably between different fish
species with data only available for just over 100 species out of the
34,000 marine and freshwater fish species (Eschmeyer and Fong, 2016).
In order to better understand acoustic impacts on fishes, fish hearing
groups are defined by species that possess a similar continuum of
anatomical features, which result in varying degrees of hearing
sensitivity (Popper and Hastings, 2003). There are four hearing groups
defined for all fish species (modified from Popper et al., 2014) within
this analysis, and they include: fishes without a swim bladder (e.g.,
flatfish, sharks, rays, etc.); fishes with a swim bladder not involved
in hearing (e.g., salmon, cod, pollock, etc.); fishes with a swim
bladder involved in hearing (e.g., sardines, anchovy, herring, etc.);
and fishes with a swim bladder involved in hearing and high-frequency
hearing (e.g., shad and menhaden). Most marine mammal fish prey species
would not be likely to perceive or hear mid- or high-frequency sonars.
While hearing studies have not been done on sardines and northern
anchovies, it would not be unexpected for them to have hearing
similarities to Pacific herring (up to 2-5 kHz) (Mann et al., 2005).
Currently, less data are available to estimate the range of best
sensitivity for fishes without a swim bladder.
In terms of physiology, multiple scientific studies have documented
a lack of mortality or physiological effects to fish from exposure to
low- and mid-frequency sonar and other sounds (Halvorsen et al., 2012a;
J[oslash]rgensen et al., 2005; Juanes et al., 2017; Kane et al., 2010;
Kvadsheim and Sevaldsen, 2005; Popper et al., 2007, 2016; Watwood et
al., 2016). Techer et al. (2017) exposed carp in floating cages for up
to 30 days to low-power 23 and 46 kHz source without any significant
physiological response. Other studies have documented either a lack of
TTS in species whose hearing range cannot perceive sonar (such as Navy
sonar), or for those species that could perceive sonar-like signals,
any TTS experienced would be recoverable (Halvorsen et al., 2012a;
Ladich and Fay, 2013; Popper and Hastings, 2009a, 2009b; Popper et al.,
2014; Smith, 2016). Only fishes that have specializations that enable
them to hear sounds above about 2,500 Hz (2.5 kHz), such as herring
(Halvorsen et al., 2012a; Mann et al., 2005; Mann, 2016; Popper et al.,
2014), would have the potential to receive TTS or exhibit behavioral
responses from exposure to
[[Page 31033]]
mid-frequency sonar. In addition, any sonar induced TTS to fish whose
hearing range could perceive sonar would only occur in the narrow
spectrum of the source (e.g., 3.5 kHz) compared to the fish's total
hearing range (e.g., 0.01 to 5 kHz).
In terms of behavioral responses, Juanes et al. (2017) discuss the
potential for negative impacts from anthropogenic noise on fish, but
the authors' focus was on broader based sounds, such as ship and boat
noise sources. Watwood et al. (2016) also documented no behavioral
responses by reef fish after exposure to mid-frequency active sonar.
Doksaeter et al. (2009, 2012) reported no behavioral responses to mid-
frequency sonar (such as naval sonar) by Atlantic herring;
specifically, no escape reactions (vertically or horizontally) were
observed in free swimming herring exposed to mid-frequency sonar
transmissions. Based on these results (Doksaeter et al., 2009, 2012;
Sivle et al., 2012), Sivle et al. (2014) created a model in order to
report on the possible population-level effects on Atlantic herring
from active sonar. The authors concluded that the use of sonar poses
little risk to populations of herring regardless of season, even when
the herring populations are aggregated and directly exposed to sonar.
Finally, Bruintjes et al. (2016) commented that fish exposed to any
short-term noise within their hearing range might initially startle but
would quickly return to normal behavior.
Pile driving noise during construction is of particular concern as
the very high sound pressure levels could potentially prevent fish from
reaching breeding or spawning sites, finding food, and acoustically
locating mates. A playback study in west Scotland revealed that there
was a significant movement response to the pile driving stimulus in
both species at relatively low received sound pressure levels (sole:
144-156 dB re 1[mu]Pa Peak; cod: 140-161 dB re 1 [mu]Pa Peak, particle
motion between 6.51 x 10\3\ and 8.62 x 10\4\ m/s\2\ peak) (Mueller-
Blenkle et al., 2010). The swimming speed of sole increased
significantly during the playback of construction noise when compared
to the playbacks of before and after construction. While not
statistically significant, cod also displayed a similar behavioral
response during before, during, and after construction playbacks.
However, cod demonstrated a specific and significant freezing response
at the onset and cessation of the playback recording. In both species,
indications were present displaying directional movements away from the
playback source. During wind farm construction in the eastern Taiwan
Strait, type 1 soniferous fish chorusing showed a relatively lower
intensity and longer duration while type 2 chorusing exhibited higher
intensity and no changes in its duration. Deviation from regular fish
vocalization patterns may affect fish reproductive success, cause
migration, augmented predation, or physiological alterations.
Occasional behavioral reactions to activities that produce
underwater noise sources are unlikely to cause long-term consequences
for individual fish or populations. The most likely impact to fish from
impact and vibratory pile driving activities at the LIAs would be
temporary behavioral avoidance of the area. Any behavioral avoidance by
fish of the disturbed area would still leave significantly large areas
of fish and marine mammal foraging habitat in the nearby vicinity. The
duration of fish avoidance of an area after pile driving stops is
unknown, but a rapid return to normal recruitment, distribution and
behavior is anticipated. In general, impacts to marine mammal prey
species are expected to be minor and temporary due to the expected
short daily duration of individual pile driving events and the
relatively small areas being affected.
Occasional behavioral reactions to activities that produce
underwater noise sources are unlikely to cause long-term consequences
for individual fish or populations. The most likely impact to fish from
impact pile driving activities at the LIA would be temporary behavioral
avoidance of the area. Any behavioral avoidance by fish of the
disturbed area would still leave significantly large areas of fish and
marine mammal foraging habitat in the nearby vicinity. The duration of
fish avoidance of an area after pile driving stops is unknown, but a
rapid return to normal recruitment, distribution and behavior is
anticipated. In general, impacts to marine mammal prey species are
expected to be minor and temporary due to the expected short daily
duration of individual pile driving events and the relatively small
areas being affected.
As described in the Proposed Mitigation section below, Vineyard
Wind would utilize a sound attenuation device which would reduce
potential for injury to marine mammal prey. Other fish that experience
hearing loss as a result of exposure to impulsive sound sources may
have a reduced ability to detect relevant sounds such as predators,
prey, or social vocalizations. However, PTS has not been known to occur
in fishes and any hearing loss in fish may be as temporary as the
timeframe required to repair or replace the sensory cells that were
damaged or destroyed (Popper et al., 2005, 2014; Smith, 2006). It is
not known if damage to auditory nerve fibers could occur, and if so,
whether fibers would recover during this process. In addition, most
acoustic effects, if any, are expected to be short-term and localized.
Long-term consequences for fish populations, including key prey species
within the LIA, would not be expected.
Required soft-starts would allow prey and marine mammals to move
away from the source prior to any noise levels that may physically
injure prey and the use of the noise attenuation devices would reduce
noise levels to the degree any mortality or injury of prey is also
minimized. Use of bubble curtains, in addition to reducing impacts to
marine mammals, for example, is a key mitigation measure in reducing
injury and mortality of ESA-listed salmon on the U.S. west coast.
However, we recognize some mortality, physical injury and hearing
impairment in marine mammal prey may occur, but we anticipate the
amount of prey impacted in this manner is minimal compared to overall
availability. Any behavioral responses to pile driving by marine mammal
prey are expected to be brief. We expect that other impacts, such as
stress or masking, would occur in fish that serve as marine mammals
prey (Popper et al., 2019); however, those impacts would be limited to
the duration of impact pile driving, and, if prey were to move out the
area in response to noise, these impacts would be minimized.
In addition to fish, prey sources such as marine invertebrates
could potentially be impacted by noise stressors as a result of the
proposed activities. However, most marine invertebrates' ability to
sense sounds is limited. Invertebrates appear to be able to detect
sounds (Pumphrey, 1950; Frings and Frings, 1967) and are most sensitive
to low-frequency sounds (Packard et al., 1990; Budelmann and
Williamson, 1994; Lovell et al., 2005; Mooney et al., 2010). Data on
response of invertebrates such as squid, another marine mammal prey
species, to anthropogenic sound is more limited (de Soto, 2016; Sole et
al., 2017). Data suggest that cephalopods are capable of sensing the
particle motion of sounds and detect low frequencies up to 1-1.5 kHz,
depending on the species, and so are likely to detect airgun noise
(Kaifu et al., 2008; Hu et al., 2009; Mooney et al., 2010; Samson et
al., 2014). Sole et al. (2017) reported physiological injuries to
cuttlefish in cages placed at-sea when exposed during a controlled
exposure experiment to low-frequency sources (315 Hz, 139 to 142 dB re
1 [mu]Pa\2\; 400 Hz, 139 to 141 dB re 1 [mu]Pa\2\).
[[Page 31034]]
Fewtrell and McCauley (2012) reported squids maintained in cages
displayed startle responses and behavioral changes when exposed to
seismic airgun sonar (136-162 re 1 [mu]Pa\2\ x s). Jones et al. (2020)
found that when squid (Doryteuthis pealeii) were exposed to impulse
pile driving noise, body pattern changes, inking, jetting, and startle
responses were observed and nearly all squid exhibited at least one
response. However, these responses occurred primarily during the first
eight impulses and diminished quickly, indicating potential rapid,
short-term habituation.
Cephalopods have a specialized sensory organ inside the head called
a statocyst that may help an animal determine its position in space
(orientation) and maintain balance (Budelmann, 1992). Packard et al.
(1990) showed that cephalopods were sensitive to particle motion, not
sound pressure, and Mooney et al. (2010) demonstrated that squid
statocysts act as an accelerometer through which particle motion of the
sound field can be detected. Auditory injuries (lesions occurring on
the statocyst sensory hair cells) have been reported upon controlled
exposure to low-frequency sounds, suggesting that cephalopods are
particularly sensitive to low-frequency sound (Andre et al., 2011; Sole
et al., 2013). Behavioral responses, such as inking and jetting, have
also been reported upon exposure to low-frequency sound (McCauley et
al., 2000; Samson et al., 2014). Squids, like most fish species, are
likely more sensitive to low-frequency sounds and may not perceive mid-
and high-frequency sonars.
With regard to potential impacts on zooplankton, McCauley et al.
(2017) found that exposure to airgun noise resulted in significant
depletion for more than half the taxa present and that there were two
to three times more dead zooplankton after airgun exposure compared
with controls for all taxa, within 1 km (0.6 mi) of the airguns.
However, the authors also stated that in order to have significant
impacts on r-selected species (i.e., those with high growth rates and
that produce many offspring) such as plankton, the spatial or temporal
scale of impact must be large in comparison with the ecosystem
concerned, and it is possible that the findings reflect avoidance by
zooplankton rather than mortality (McCauley et al., 2017). In addition,
the results of this study are inconsistent with a large body of
research that generally finds limited spatial and temporal impacts to
zooplankton as a result of exposure to airgun noise (e.g., Dalen and
Knutsen, 1987; Payne, 2004; Stanley et al., 2011). Most prior research
on this topic, which has focused on relatively small spatial scales,
has showed minimal effects (e.g., Kostyuchenko, 1973; Booman et al.,
1996; S[aelig]tre and Ona, 1996; Pearson et al., 1994; Bolle et al.,
2012).
A modeling exercise was conducted as a follow-up to the McCauley et
al. (2017) study (as recommended by McCauley et al., 2017), in order to
assess the potential for impacts on ocean ecosystem dynamics and
zooplankton population dynamics (Richardson et al., 2017). Richardson
et al. (2017) found that a full-scale airgun survey would impact
copepod abundance within the survey area, but that effects at a
regional scale were minimal (2 percent decline in abundance within 150
km (93.2 mi) of the survey area and effects not discernible over the
full region). The authors also found that recovery within the survey
area would be relatively quick (3 days following survey completion) and
suggest that the quick recovery was due to the fast growth rates of
zooplankton, and the dispersal and mixing of zooplankton from both
inside and outside of the impacted region. The authors also suggest
that surveys in areas with more dynamic ocean circulation in comparison
with the study region and/or with deeper waters (i.e., typical offshore
wind locations) would have less net impact on zooplankton.
Notably, a recently described study produced results inconsistent
with those of McCauley et al. (2017). Researchers conducted a field and
laboratory study to assess if exposure to airgun noise affects
mortality, predator escape response, or gene expression of the copepod
Calanus finmarchicus (Fields et al., 2019). Immediate mortality of
copepods was significantly higher, relative to controls, at distances
of 5 m or less from the airguns. Mortality 1 week after the airgun
blast was significantly higher in the copepods placed 10 m from the
airgun but was not significantly different from the controls at a
distance of 20 m from the airgun. The increase in mortality, relative
to controls, did not exceed 30 percent at any distance from the airgun.
Moreover, the authors caution that even this higher mortality in the
immediate vicinity of the airguns may be more pronounced than what
would be observed in free-swimming animals due to increased flow speed
of fluid inside bags containing the experimental animals. There were no
sub-lethal effects on the escape performance, or the sensory threshold
needed to initiate an escape response, at any of the distances from the
airgun that were tested. Whereas McCauley et al. (2017) reported an SEL
of 156 dB at a range of 509-658 m, with zooplankton mortality observed
at that range, Fields et al. (2019) reported an SEL of 186 dB at a
range of 25 m, with no reported mortality at that distance.
Airguns and impact pile driving are similar in that they both
produce impulsive and intermittent noise and typically have higher
source levels than other sources (e.g., vibratory driving). We
anticipate marine mammal prey exposed to impact pile driving would
demonstrate similar physical consequences and behavioral impacts
compared to exposure to airguns; however, the spatial extent of these
impacts during impact pile driving is dependent upon source levels and
use of noise attenuation systems (NA
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