The Rapid Population Collapse of a Key Marine Predator in the Northern Antarctic Peninsula Endangers Genetic Diversity and Resilience to Climate Change

Krause, Douglas J.; Bonin, Carolina A.; Goebel, Michael E.; Reiss, Christian S.; Watters, George M.

doi:10.3389/fmars.2021.796488

ORIGINAL RESEARCH article

Front. Mar. Sci., 03 January 2022
Sec. Marine Conservation and Sustainability
Volume 8 - 2021 | https://doi.org/10.3389/fmars.2021.796488

The Rapid Population Collapse of a Key Marine Predator in the Northern Antarctic Peninsula Endangers Genetic Diversity and Resilience to Climate Change

Douglas J. Krause^1*

¹Antarctic Ecosystem Research Division, Southwest Fisheries Science Center, NOAA Fisheries, La Jolla, CA, United States
²Marine and Environmental Science Department, Hampton University, Hampton, VA, United States

Antarctic fur seals (AFS) are an ecologically important predator and a focal indicator species for ecosystem-based Antarctic fisheries management. This species suffered intensive anthropogenic exploitation until the early 1900s, but recolonized most of its former distribution, including the southern-most colony at Cape Shirreff, South Shetland Islands (SSI). The IUCN describes a single, global AFS population of least concern; however, extensive genetic analyses clearly identify four distinct breeding stocks, including one in the SSI. To update the population status of SSI AFS, we analyzed 20 years of field-based data including population counts, body size and condition, natality, recruitment, foraging behaviors, return rates, and pup mortality at the largest SSI colony. Our findings show a precipitous decline in AFS abundance (86% decrease since 2007), likely driven by leopard seal predation (increasing since 2001, p << 0.001) and potentially worsening summer foraging conditions. We estimated that leopard seals consumed an average of 69.3% (range: 50.3–80.9%) of all AFS pups born each year since 2010. AFS foraging-trip durations, an index of their foraging habitat quality, were consistent with decreasing krill and fish availability. Significant improvement in the age-specific over-winter body condition of AFS indicates that observed population declines are driven by processes local to the northern Antarctic Peninsula. The loss of SSI AFS would substantially reduce the genetic diversity of the species, and decrease its resilience to climate change. There is an urgent need to reevaluate the conservation status of Antarctic fur seals, particularly for the rapidly declining SSI population.

Introduction

Rapid changes in the abundance of, or prey selection by, apex predators can fundamentally alter ecosystems through dynamic predator-prey interactions (Hairston et al., 1960; Paine, 1966; Dayton, 1971; Estes et al., 2011). Marine mammalian apex predators, in particular, may drive rapid declines in coastal prey populations because of their large body sizes and proportionally high energetic demands (Williams et al., 2004; Pagano et al., 2018). Despite this potential, regional-scale examples of such effects have been rare (Estes et al., 1998; Springer et al., 2003), and sometimes controversial (DeMaster et al., 2006; Wade et al., 2009); in part, because direct observations of predation are difficult to obtain in remote marine environments (Estes et al., 1998).

Long-term monitoring of marine predators is fundamental to ecosystem-based management of marine ecosystems (Boyd et al., 2006; Moore, 2008). The population success of central place foragers, like some pinnipeds, depends upon predictable prey availability. Therefore, estimates of pinniped abundance (e.g., Southwell et al., 2012; Trukhanova et al., 2013; Lowry et al., 2014), and characterizations of their foraging behavior (Trillmich and Dellinger, 1991; Melin et al., 2008) and body condition (Costa et al., 1989) are used to inform conservation and fisheries management (Weise and Harvey, 2008; Melin et al., 2010). The Antarctic fur seal (AFS, Arctocephalus gazella) is a key indicator species for the ecosystem-based management of the Antarctic krill (Euphausia superba) fishery (Boyd and Murray, 2001).

Rapid climatic warming and the associated loss of sea ice (Meredith and King, 2005; Vaughan, 2006; Turner et al., 2014) in the western Antarctic Peninsula (WAP) are expected to alter pelagic communities and shift their geographical distributions, including those of Antarctic krill and their manifold dependent predators (Massom and Stammerjohn, 2010; Ducklow et al., 2013; Klein et al., 2018). As expected, ice-dependent, and ice-associated penguins are in decline throughout the Peninsula region (Hinke et al., 2007; Trivelpiece et al., 2011; Lynch et al., 2012). Ice seals, like crabeater (Lobodon carcinophagus) and leopard seals (Hydrurga leptonyx), have also likely experienced substantial changes in abundance and distribution (Forcada et al., 2012; Hückstädt et al., 2020). Conversely, at the high-latitude extreme of their thermal tolerance (Costa et al., 1989; Reid and Croxall, 2001), populations of AFS were expected to flourish in a warmer climate due to increased habitat (Siniff et al., 2008; Costa et al., 2010).

Extensive harvesting of AFS between the late 1700s and early 1900s extirpated the species from much of its range, and reduced the global population to near extinction (Bonner, 1968; Weddell, 1970; McCann and Doidge, 1987). By the late 1990s the global population of AFS was estimated to have recovered to near historical levels (Hofmeyr, 2016; Foley and Lynch, 2020). The current conservation paradigm for AFS hypothesizes that a single population survived near the subantarctic island of South Georgia, expanded through the mid-1900s and then recolonized other subantarctic island groups as well as the only Antarctic breeding population of AFS in the South Shetland Islands (SSI; Laws, 1973; Payne, 1977; Shaughnessy and Goldsworthy, 1990), where the only Antarctic breeding population of AFS exists. This hypothesis was consistent with the lack of population structure detected using mitochondrial DNA (mtDNA) markers (Wynen et al., 2000). Therefore, based on the best available science in 2014, the International Union for Conservation of Nature (IUCN) determined that AFS is a species of “least concern” and “neither this species as a whole, nor any separate colonies, are likely to become extinct in the near future” (Hofmeyr, 2016).

A suite of recent studies, however, suggest that the existing conservation paradigm may be inaccurate on two key points. First, genetic analyses using large sample sizes and nuclear markers demonstrate that rather than a single genetic stock, there are at least four genetically distinct subpopulations from South Georgia, Bouvetøya, SSI, and the remaining subantarctic islands (Bonin et al., 2013; Humble et al., 2018; Cleary et al., 2019; Paijmans et al., 2020). Second, three of these four subpopulations have declined in abundance. In addition to reductions (2.8–5.6% per annum) in the number of breeding females at South Georgia and Bouvetøya since the early 2000s (Forcada and Hoffman, 2014; Hofmeyr, 2016), recent reports suggest a rapid, potentially catastrophic decline in the SSI population (Krause et al., 2020; Krause and Hinke, 2021).

Initial concern for AFS related to climate change focused on population and habitat effects due to redistribution of their prey resources (Siniff et al., 2008; Kovacs et al., 2012). Indeed, the use of AFS as indicator species for fisheries management (Boyd and Murray, 2001) and the bulk of population dynamics research to date both assumed and demonstrated bottom-up population control (Boyd, 1993; Reid and Croxall, 2001; Forcada et al., 2005; Forcada and Hoffman, 2014). While predation of AFS pups by leopard seals was reported at South Georgia as early as the 1980s, analyses indicated this top-down forcing had no effect on AFS population dynamics (McCann and Doidge, 1987; Forcada et al., 2009).

As the only breeding population of AFS south of the Antarctic Convergence and the highest-latitude otariid population on Earth, the SSI colonies face a unique array of environmental and ecological challenges. Adaptation to colder air and water temperatures, for example, are likely drivers for the shorter provisioning trip durations and larger body sizes observed for SSI AFS compared with their subantarctic counterparts (this study). Further, AFS from SSI colonies are closer to the core habitat of leopard seals (Forcada et al., 2012). Top-down control of a small SSI AFS colony by leopard seals was modeled in the late 1990s (Boveng et al., 1998), however, broad population recovery continued across the archipelago at that time (Goebel et al., 2003). The recent population crash at Cape Shirreff may be driven by leopard seal predation, and the magnitude could be substantially larger than previously reported (Krause et al., 2015, 2020).

AFS colonies on and around Cape Shirreff, Livingston Island produce 60–85% of all pups born in the SSI (Bengtson et al., 1990; Krause and Hinke, 2021). Further, since the late 1990s the National Oceanic and Atmospheric Administration has annually operated a comprehensive, individual-based monitoring program tracking foraging behavior, body condition, pup survival, observations of predation, and population dynamics of AFS at Cape Shirreff. Here our objectives are to: (1) evaluate the current evidence of AFS population structure, particularly the assignment of the SSI AFS as a genetically distinct subpopulation, (2) summarize the population trends of AFS in the SSI over the last 20 years, and, (3) examine standardized behavioral and demographic parameters of SSI AFS to explain observed trends and inform future research.

Materials and Methods

Study Site

Research was conducted at Cape Shirreff (62.47° S, 60.77° W) on the north shore of Livingston Island, in the South Shetland Islands archipelago (Figure 1). Historically, the southern boundary of the southern Antarctic Circumpolar Current front overlaps with the continental shelf break approximately 50 km north of the Cape (Orsi et al., 1995). That association enhances the availability of Antarctic krill near the Cape (Atkinson et al., 2009), and results in a persistent foraging hotspot for AFS and other Antarctic predators (Santora and Veit, 2013). Since 1996–1997 (hereafter austral summers are referred to by the second year, e.g., 1997), the U.S. Antarctic Marine Living Resources (U.S. AMLR) Program has conducted long-term ecological monitoring of Antarctic pinnipeds, including AFS, as indicator species to inform management of the regional krill fishery (Agnew, 1997; Goebel et al., 2008). All methods reported here were implemented each summer between 2001 and 2020 unless otherwise indicated.

FIGURE 1

Figure 1. Locations of the western Antarctic fur seal breeding centers including: South Georgia, Bouvetøya, and the South Shetland Islands.

All animal research described herein was conducted in accordance with Marine Mammal Protection Act Permit Nos. 1024, 774-1649, 774-1847, 16472, and 20599, Antarctic Conservation Act Permit Nos. 97-016, 2002-007, 2007-003, 2012-005 and 2017-012, and NMFS-SWFSC Institutional Animal Care and Use Committee Permit Nos. IACUC SWPI2011-02, SWPI 2014-03, and SWPI 2017-03.

Antarctic Fur Seal Breeding Ecology and Dispersal

AFS are polygynous breeders in which males establish territories on breeding beaches (October-November) and accumulate harems of females as the latter arrive to the colonies. Pregnant females give birth within 1–2 days of arrival and most pupping occurs in late November to early December (Bonner, 1968). After suckling their pups for 5–7 days they begin a cyclical series of foraging trips to sea interspersed with onshore pup-provisioning visits (Payne, 1977). During the austral summer, AFS are dependent upon Antarctic krill as their main food source, with myctophid fishes and squid as alternates (Polito and Goebel, 2010). While provisioning pups, female AFS foraging trips vary in duration depending upon the availability of prey, but typically last 4–5 days near South Georgia (Costa et al., 1989; Boyd, 1999). Pups that survive through the summer are weaned in late March to early April. After weaning, females from Cape Shirreff disperse widely and forage in pelagic habitats, including off the coasts of Chile and Argentina (Arthur et al., 2017; Hinke et al., 2017). Conversely, weaned pups depart from Cape Shirreff, but typically remain in the South Shetland Islands and northern Antarctic Peninsula during their first year. Subadult and adult males move south along the Antarctic Peninsula in the fall and winter.

Antarctic Fur Seals Identification Marking and Daily Observations

We deployed bilateral identification (ID) flipper tags on every study animal captured since 1998 (N_{ID–tagged adults} = 613). The individually numbered Dalton cattle tags (40 mm × 20 mm × 2 mm, 10 g) were deployed in the cartilaginous tissue at the proximal trailing edge of the fore flippers to maximize visibility and reduce hydrodynamic drag. As part of our mark re-capture study we also deployed ∼500 ID flipper tags per annum on AFS pups between 1998 and 2011. Thereafter, we decreased the tagging rate to target 10% of the pups born each year through 2020 (N_{ID–tagged pups} = 8,776).

Attendance Behavior

In addition to daily observations of behavior, each year we captured and closely monitored approximately 30 ( $\bar{x}$ = 28.9, range: 16–34) adult female-pup pairs to study inter- and intra-annual changes in female body condition, pup attendance and foraging behavior. We captured visibly healthy females with hoop nets (Animal Equipment by Stoney, LLC) during the perinatal period (1–2 days postpartum during late November—early December) following procedures described by Gentry and Holt (1982) and Majluf and Goebel (1992). Some animals were restrained on capture boards, however, the majority (>82%) were anesthetized using isoflurane gas (Gales, 1989; Gales and Mattlin, 1998) and midazolam (dosage: 0.1–0.25 mg/kg, Haulena, 2014). Pups were manually restrained, and reunited with their mother post-anesthetic recovery. We measured adult females and pups for standard length and axillary girth to the nearest 0.5 cm (Scheffer, 1967); we also weighed them using suspension scales (± 0.1 kg). We marked pups with unique bleach patterns in their fur, and applied identification tags to all adult females that were not previously tagged.

We attached marine VHF radio transmitters (164 Mhz, 55 mm × 23 mm × 10 mm, 23 g, Advanced Telemetry Systems Inc.) to the dorsal pelage of all females included in the attendance study using standard 5-min marine epoxy. VHF receivers integrated with automated data loggers recorded the presence or absence of each female every 30 min throughout the breeding season. Each absence longer than 7 h of a female who later returned to a living pup was classified as a foraging trip. We corroborated VHF presence data with alternate VHF receivers, visual observations, and animal-borne data loggers when available. We calculated mean trip durations and standard errors for the first six foraging trips per season (Agnew, 1997). As a metric for describing unusually good or poor foraging conditions, we defined seasons with average trip lengths above or below one standard deviation of the mean for the entire study period as extreme.

Annual Pup Census

Due to regular absences by foraging females throughout the breeding season, and the irregular haul out patterns of males and subadults, the most informative measure of fur seal population size is to annually count pups (Payne, 1979; Bengtson et al., 1990). The U.S. AMLR Program conducted a synoptic census of AFS pups born at Cape Shirreff each year between 2008 and 2020 (for methods used in other years see references listed in Supplementary Table 1). Annual pup censuses were done in late December when over 99% of pups were born, before pups explore widely beyond their birth beach, and before seasonal rates of pup predation increased. In each season at least three field biologists surveyed every breeding beach independently using hand-held binoculars. Counts of pups (live and dead) were independently reviewed, and recounts were conducted as necessary to ensure that all beach-specific counts were within 5% across all observers.

Female Return and Natality Rates

Population dynamics are essentially controlled by the balance between birth and death rates; however, in polygynous otariid populations such rates are disproportionately affected by key factors such as adult female survival, and natality rates (Boyd et al., 1995). Long-term pinniped monitoring programs rely on regular observations of a subset of individuals from their respective breeding populations to provide indices of these key demographic parameters (Boyd, 1993; Melin et al., 2012; DeLong et al., 2017).

Each season every previously ID-tagged adult female was observed daily from arrival through early March. We recorded arrival date and location for all returning females, and parturition status. Because rates of philopatry are exceptionally high for AFS females (Payne, 1977; Hoffman and Forcada, 2012), we considered return rate a proxy for adult female survival and calculated natality rates as the annual proportions of returned adult females that bore pups.

Pup Mortality

A key bottleneck to recruitment for Antarctic predators is survival through the first year (Payne, 1977; Hinke et al., 2020). Unfortunately, it is challenging to assess pup survival for the entire population due to the likelihood that carcasses will be missed, washed away, or double-counted (Hofmeyr et al., 2005). Therefore, to estimate rates of survival to weaning, we tracked the daily health, nutritive status, and survival of all pups born to ID-tagged females at Cape Shirreff. We categorized pup mortality as: (1) neonate mortality (pups that died within 21 days of birth, still birth, being injured by adults, or starvation), (2) assumed leopard seal predation (pups observed being killed by a leopard seal, pup carcasses found with characteristic indicators of leopard seal predation, e.g., crushed or missing skull, inverted body cavity, or healthy pups that disappeared suddenly with no other explanation), or (3) other (all other known causes of death, e.g., starvation beyond 21 days after birth, predation by skuas or giant petrels, potential disease, or other injuries).

Age Determination

Many demographic and health parameters are affected by age. For example, body size and natality rates are significantly lower for young females (Payne, 1979), and pregnancy rates decline in females over the age of 17 at Cape Shirreff (Goebel and Reiss, 2014). Therefore, in addition to all known-age animals tagged as pups, we derived the age of some adult females by analyzing one of their teeth. Characteristic growth layers in the cementum can accurately reflect the age of pinnipeds (Laws, 1952; Arnbom et al., 1992; Childerhouse et al., 2004). For a subset of our anesthetized attendance study females, we extracted a lower post-canine tooth (N = 345). We stored teeth in 90% ethanol. We decalcified the teeth, cut a longitudinal section (0.4 mm) using a diamond-edged saw, and mounted each section to a glass microscope slide (Childerhouse et al., 2004). A tooth-aging laboratory (Matson’s Lab, Missoula, MT) provided independent counts of dentition growth layers from each slide by multiple observers.

Based on a multi-year analysis of known-age females at Cape Shirreff, we defined females in our study between the ages of 7 and 17 as prime breeders, due to their substantially higher natality and pup growth rates (Goebel and Reiss, 2014).

Data Analysis

We conducted all analyses using R (R-Core-Team, 2021). All values are listed as mean ( $\bar{x}$ ) ± standard deviation, and all inferences were based on a significance level of p ≤ 0.05 unless otherwise indicated. We tested for predictable trends in measurements over the study period using ordinary least squares regressions, verified test assumptions were met, and plotted all figures using ggplot2 (Wickham, 2009).

Inter-Regional Body Size Comparison

In addition to reviewing literature on genetic comparisons between known AFS populations, we compared standard lengths and masses between published measurements from the South Georgia population collected during 1973 and 1974 (Payne, 1979), and the Cape Shirreff population collected between 2001 and 2020. We matched the range of capture dates (late January—early March), and selected only breeding age females (≥ 3 years). We tested for differences using a Welch’s t-test.

Female Arrival Body Condition

The relative proportion of fat stores to body size, or body condition, of otariids is related to quality of the foraging environment before measurement (Boyd and McCann, 1989; Costa et al., 1989). Further, body condition is positively related to natality rate (Guinet et al., 1998). For each season we calculated a body condition index (BCI) for all females included in our attendance study by linearly regressing body mass against standard length and using the residuals from the fitted line as BCI following Guinet et al. (1998). We controlled for the variable weight of fetus and placenta in females just after arrival by only selecting females who were 1–2 days postpartum.

Results

Antarctic Fur Seal Population Structure

Between 2001 and 2020 we captured, measured, and weighed adult female AFS during the second half of the breeding season. Mean standard length was 129.98 ± 6.0 cm (N = 376), and mean mass was 42.12 ± 5.7 kg (N = 377). Based on captures at South Georgia during the 1973 and 1974 summer seasons, Payne (1979) reported standard lengths from 190 ( $\bar{x}$ = 126.30 ± 6.4 cm), and masses from 166 ( $\bar{x}$ = 32.31 ± 5.7 kg) adult female AFS. Adult females from Cape Shirreff were both longer (t = 6.6176, df = 357.01, p << 0.001, Figure 2A), and more massive (t = 18.392, df = 293.93, p << 0.001, Figure 2B) than those from South Georgia.

FIGURE 2

Figure 2. The (A) standard length and (B) mass values from adult Antarctic fur seals measured at South Georgia (N_{Standard length} = 190, N_mass = 166) and Cape Shirreff (N_{Standard length} = 376, N_mass = 377) between 1973–74 and 2000–2020, respectively. Boxes indicate upper and lower quartiles, horizontal lines indicate median values, whiskers represent 10th and 90th percentiles, and dots represent outliers.

South Shetland Islands Population Trends

All known, synoptic counts of AFS born in the South Shetland Islands archipelago, including from Cape Shirreff and the San Telmo Islets, are listed in Supplementary Table 1 (Figure 3). From pup counts, we estimated the total population of all ages using the ratio 1:4.1 (Payne, 1979). The highest pup production for the entire South Shetlands population was measured in 2002 at 10,057 pups born, or 41,233 total AFS. The most recent count from the San Telmo Islets (333 pups in 2019) represents a 90% reduction in pup production since a peak in 1997 (Krause and Hinke, 2021). Similarly, pup production at Cape Shirreff has decreased by 86% since 2007 (Table 1).

FIGURE 3

Figure 3. Total synoptic pup counts (live and dead) from the South Shetland Islands archipelago (gold triangles), and from the subset populations at Cape Shirreff (blue circles) and the San Telmo Islets (green squares), between the 1959 and the 2020 Antarctic seasons. Raw data and associated citations listed in Supplementary Table 1.

TABLE 1

Table 1. The time periods and associated rates of population change (% per annum), based on annual pup production unless otherwise indicated, for AFS in several population centers.

Drivers of Population Change

Age-Specific Adult Female Return Rates

The rates at which adult females at Cape Shirreff survived the winter and successfully returned to the breeding colony decreased between 2001 and 2020. This was the case for the entire population (N = 3,149, slope = −0.66%/year, p = 0.003) and for prime-age females only (N = 2,154, slope = −0.75%/year, p = 0.004, Figure 4A).

FIGURE 4

Figure 4. Proportions of (A) all ID-tagged adult female Antarctic fur seals that departed the breeding colony at Cape Shirreff the previous season that returned per annum, (B) all ID-tagged adult females present that pupped in a given season, between 2001 and 2020 at Cape Shirreff, (C) Mean Body Condition Index at colony arrival for adult female Antarctic fur seals at Cape Shirreff during each season between 2001 and 2018. Prime age females (gray triangles) are between 7 and 17 years old; non-prime age females (orange circles) are between 3 and 6 years old.

Age-Specific Natality Rates

The proportion of adult females that returned to Cape Shirreff and pupped in a given season significantly decreased over the study period (N_{Females observed} = 3,060, slope = −0.69%/year, p < 0.001). However, when only prime-age females were considered, there was no significant trend over the study period (N = 1,917, slope −0.47%/year, p = 0.133, Figure 4B). The mean natality rates between 2001 and 2020 were $\bar{x}$ = 81.7 ± 5.8%, and $\bar{x}$ = 86.7 ± 7.9% for all females, and only prime-age females, respectively.

Body Condition Index

The regression of AFS body mass (M) on standard length (sl) used to calculate BCI, M = −33.02 + 0.627 × sl, was highly significant (N = 624, p << 0.001), and the arrival BCI of females included in our attendance study increased over the study period. This trend was significant for both prime-aged (N = 419, slope = 0.55 BCI/year, p << 0.001), and females aged 3–6 years (N = 46, slope = 0.36 BCI/year, p = 0.01, Figure 4C). None of the females captured and measured during the 2019 and 2020 seasons were of known-age and therefore were not included in calculations of BCI.

Foraging Trip Lengths

Mean foraging trip lengths from 2003 ( $\bar{x}$ = 170.1 ± 4.5 h) were anomalously long (>5.5 standard deviations above the long term mean), and were unique during the 20-year study period. As such, 2003 trip lengths are an overly influential outlier in the regression analysis (Cook’s D >> 1). Therefore, we present our analysis with and without data from 2003 included. The mean foraging trip length over the study period was $\bar{x}$ = 84.79 ± 15.31 h (N_females = 484, N_Trips = 2,904) and ranged from 54.99 to 112.03 h without data from 2003, and $\bar{x}$ = 89.05 ± 16.72 h with data from 2003 (N_2003females = 15, N_2003Trips = 90, range₂₀₀₃ = 56.64–333.06 h). With 2003 excluded, mean trip length per season increased between 2001 and 2020 (slope = 1.38 h/year, p = 0.022, Figure 5A), however there was no linear trend when data from 2003 were included (p = 0.761). The incidence of seasons with extremely short foraging trips decreased from 50% in the first half of the study ( $\bar{x}$ = 75.54 ± 14.01 h) to 0% in the second half ( $\bar{x}$ = 95.16 ± 8.61 h, Figure 5B).

FIGURE 5

Figure 5. Mean foraging trip lengths from the first 6 trips in a given season by attendance study AFS females from Cape Shirreff (A) excluding 2003 data (N_females = 484, N_Trips = 2,904), the blue line is a significant (P = 0.022) regression, trip length = -2692 + 1.38 season; and (B) including the anomalously long foraging trip lengths from 2003, the solid orange line marks the mean trip length and dashed orange lines mark standard deviation lines for the entire study period.

Population Age Structure

The mean age of all adult females present within a season increased steadily throughout the study period (N = 3,044, Adjusted R² = 0.932, p << 0.001, Figure 6A). The overall average age across the study period was $\bar{x}$ = 13.39 ± 1.79 years, increasing from 10.48 ± 3.63 years in 2001 to 17.03 ± 4.86 years in 2018.

FIGURE 6

Figure 6. (A) The mean age, with standard deviation bars, of all adult Antarctic fur seal females present at Cape Shirreff in a given season. (B) The proportion of individually-monitored Antarctic fur seal pups born in a given year at Cape Shirreff that were consumed by leopard seals. All panels cover the period between 2001 and 2020.