Spatial and temporal trends and sex differences in ∑PCB exposure

The temporal trend in ln ∑PCB lipid concentrations in UK HPs (1990–2012) and SDs from the western Mediterranean Sea (1990–2009) is shown in Fig. 1A,B. The temporal trend in both populations was highly statistically significant (p < 0.001). In UK-stranded HPs, ln ∑PCB concentrations declined slowly from 1990 to 1998 and then remained relatively stable from 1998 to 2012 against the null hypothesis of no trend (p < 0.001, F = 11.76, residual df = 701.97, trend df = 3.03) (Fig. 1A). In Mediterranean SDs, ln ∑PCB concentrations showed a marked decline from an initial peak in 1990, but then stabilized from 2003 to 2008, but with blubber ∑PCB concentrations still consistently exceeding all mammalian toxicity thresholds. The trend is statistically significant (p < 0.001, F = 55.45, residual df = 212.03, trend df = 6.97) against the null hypothesis of no trend (Fig. 1B).

Figure 1 Temporal trends in ∑PCBs in UK-stranded harbour porpoise (Phocoena phocoena) and striped dolphins (Stenella coeruleoalba) in the western Mediterranean Sea. To assess the ∑PCBs trend we have fitted a Generalised Additive Model (GAM) to the data using the R (R Development Core Team, 2013) package mgcv. We used thin plate regression splines to do the smoothing and the degree of smoothing was determined by generalised cross validation. (A) Ln ∑PCBs (sum 18–25CB) mg/kg lipid concentrations in UK harbour porpoise blubber against date for all data for 1990–2012 (n = 706). The continuous line represents the smoothed trend from a Generalized Additive Model fitted to the data. The trend is statistically significant (p < 0.001, F = 11.76, residual df = 701.97, trend df = 3.03) against the null hypothesis of no trend. The dashed lines represent the 95% bootstrapped Confidence Intervals. The yellow line represents ln ∑PCBs equivalent to 20.0 mg/kg lipid and the red line 40 mg/kg lipid. (B) Ln ∑PCBs (sum 18–25CB) lipid concentrations in biopsied striped dolphin blubber from the Mediterranean Sea against date for all data for 1990–2009 (n = 220). Figure shows the natural logs (ln) of the whole data set plotted against the date found. The continuous line represents the smoothed trend from a Generalized Additive Model fitted to the data. The trend is statistically significant (p < 0.001, F = 55.45, residual df = 212.03, trend df = 6.97) against the null hypothesis of no trend. The dashed lines represent the 95% bootstrapped Confidence Intervals. The yellow line represents ln ∑PCBs equivalent to 20.0 mg/kg lipid and the red line 40 mg/kg lipid. Full size image

Some geographical regions in this study were global PCB “hotspots” for cetaceans. Mean ∑PCB lipid concentrations in BNDs and KWs from the NE Atlantic and in BNDs and SDs from the Mediterranean were among the highest recorded in cetaceans globally9, with males and females markedly exceeding all known PCB toxicity thresholds for marine mammals (Figs 2, 3, 4, 5). Results of kernel smoothing estimates show very high risk of PCB toxicity for KWs in NE Atlantic, BNDs off SW Iberia and in the northern Adriatic Sea and BNDs/SDs in the western Mediterranean (Fig. 6). ∑PCB data from biopsies of 152 free-living individuals constituted 38/131 (29.0%) of all BNDs; 99/220 (45.0%) of the SDs; and 15/24 (62.5%) of the KWs in this study. All HPs in the UK were stranded. Although biases can potentially occur with opportunistically sampled stranded cetaceans, the study population used here consisted of both compromised diseased individuals as well as healthy trauma cases and no clear differences between mean ∑PCB concentrations in stranded and biopsied BNDs (1990–2012), SDs (1991–2009) and KWs (1994–2012) were observed (Fig. 7). More specifically, mean ∑PCB concentrations in stranded male and female SDs were slightly higher than biopsied male and female SDs but biopsied male BNDs had slightly higher ∑PCB concentrations than stranded male BNDs. For KWs, biopsied males and stranded females had higher ∑PCB concentrations than stranded male and biopsied female KWs, respectively (Fig. 7). Box and whisker plots (male v female) were generated for ∑PCB and ln ∑PCB concentrations in all stranded and biopsied HPs (1990–2012), BNDs (1990–2012), SDs (1990–2008) and KWs (1994–2012) (Fig. 8A–D).

Figure 2 Mean ∑PCBs concentrations in male and female cetaceans (four species; all ages) The blue bars are males and the grey bars are females. The lower line is the equivalent ∑PCBs concentrations threshold (9.0 mg/kg lipid) for onset of physiological effects in experimental marine mammal studies6. The upper line is the equivalent ∑PCBs concentrations threshold (41.0 mg/kg lipid) for the highest PCB toxicity threshold published for marine mammals based on marked reproductive impairment in ringed seals in the Baltic Sea6. Mean ∑PCBs concentrations in male (n = 388) and female (n = 318) UK-stranded harbour porpoises (HPs) in 1990–2012. Mean blubber ∑PCBs (mg/kg lipid) concentrations in subsets of male (n = 201) and female (n = 144) UK-stranded HPs that died of acute physical trauma and male (n = 120) and female (n = 132) HPs that died of infectious disease from the same 1990–2012 period. Mean blubber ∑PCBs (mg/kg lipid) concentrations (1990–2012) shown for stranded/biopsied male (n = 29) and female (n = 17) bottlenose dolphins (BNDs) from UK and Ireland; male (n = 28) and female (n = 24) BNDs from Atlantic coast of Spain and Portugal and male (n = 9) and female (n = 11) BNDs from western Mediterranean Sea. Male (n = 50) and female (n = 39) striped dolphins from western Mediterranean Sea (1991–2009) and male (n = 5) and female (n = 19) KWs from NE Atlantic (1994–2012). Error bars = 1Standard Error of the Mean (SEM). Full size image

Figure 3 Mean ∑PCBs concentrations in male and female cetaceans (four species; adults only). The blue bars are adult males and the grey bars are adult females. The lower line is the equivalent ∑PCBs concentrations threshold (9.0 mg/kg lipid) for onset of physiological effects in experimental marine mammal studies6. The upper line is the equivalent ∑PCBs concentrations threshold (41.0 mg/kg lipid) for the highest PCB toxicity threshold published for marine mammals based on marked reproductive impairment in ringed seals in the Baltic Sea6. Mean ∑PCBs concentrations (mg/kg lipid) in adult male (n = 146) and female (n = 134) UK-stranded harbour porpoises in 1990–2012. Mean blubber ∑PCBs (mg/kg lipid) concentrations (1990–2012) shown for stranded/biopsied adult male (n = 20) and female (n = 14) bottlenose dolphins from UK and Ireland and the Atlantic coast of Spain and Portugal (NE Atlantic). Mean blubber ∑PCBs (mg/kg lipid) concentrations shown for adult male (n = 8) and female (n = 14) striped dolphins from the western Mediterranean Sea (1991–2009). Finally, mean blubber ∑PCBs (mg/kg lipid) concentrations shown for adult male (n = 3) and female (n = 18) killer whales from the NE Atlantic (1994–2012). Error bars = 1SEM. Full size image

Figure 4 Mean ∑PCBs concentrations in male and female cetaceans (four species; juveniles only). The blue bars are juvenile males and the grey bars are juvenile females. The lower line is the equivalent ∑PCBs concentrations threshold (9.0 mg/kg lipid) for onset of physiological effects in experimental marine mammal studies6. The upper line is the equivalent ∑PCBs concentrations threshold (41.0 mg/kg lipid) for the highest PCB toxicity threshold published for marine mammals based on marked reproductive impairment in ringed seals in the Baltic Sea6. Mean ∑PCBs concentrations (mg/kg lipid) in juvenile male (n = 233) and female (n = 180) UK-stranded harbour porpoises (HPs) in 1990–2012. Mean blubber ∑PCBs (mg/kg lipid) concentrations (1990–2012) shown for stranded/biopsied juvenile male (n = 20) and female (n = 14) bottlenose dolphins from UK and Ireland and the Atlantic coast of Spain and Portugal (NE Atlantic). Mean blubber ∑PCBs (mg/kg lipid) concentrations shown for juvenile male (n = 17) and female (n = 15) striped dolphins from western Mediterranean Sea (1991–2009). Finally, mean blubber ∑PCBs (mg/kg lipid) concentrations shown for juvenile male (n = 2) and female (n = 0) killer whales from NE Atlantic (1994–2012). Error bars = 1SEM. Full size image

Figure 5 Mean ∑PCBs concentrations in male and female cetaceans (three species; sexual maturity unknown). The blue bars are males and the grey bars are females. The lower line is the equivalent ∑PCBs concentrations threshold (9.0 mg/kg lipid) for onset of physiological effects in experimental marine mammal studies6. The upper line is the equivalent ∑PCBs concentrations threshold (41.0 mg/kg lipid) for the highest PCB toxicity threshold published for marine mammals based on marked reproductive impairment in ringed seals in the Baltic Sea6. Mean ∑PCBs concentrations (mg/kg lipid) in male (n = 9) and female (n = 4) UK-stranded harbour porpoises of unknown sexual maturity (1990–2012). Mean blubber ∑PCBs (mg/kg lipid) concentrations (1990–2012) shown for stranded/biopsied male (n = 16) and female (n = 13) bottlenose dolphins (BNDs) from the Atlantic coast of Spain and Portugal and male (n = 8) and female (n = 11) BNDs from the western Mediterranean Sea (all sexual maturity unknown). Mean blubber ∑PCBs (mg/kg lipid) concentrations shown for male (n = 25) and female (n = 10) striped dolphins from western Mediterranean Sea (1991–2009) (sexual maturity unknown). Error bars = 1SEM. Full size image

Figure 6 (A–D) Distribution map (smooth mean density kernel plots) of ∑PCBs data points in Europe – all cetacean species (all ages) from 1996–2012. (A) – HPs (n = 548); (B) – BNDs (n = 110); (C) – SDs (n = 71) and (D) – KWs (n = 21). Spatial distribution of ∑PCB lipid concentrations produced in Esri ArcMap 10.1 (www.esri.com). Maps are displayed in the WGS84 co-ordinate system. Data points are shown along with local averages. These averages were calculated by kernel smoothing using a polynomial order 5 kernel with power = 0, ridge parameter = 50 and bandwidth based on the spatial distribution of the observations for each species: bottlenose dolphin 0.75 degrees; harbour porpoise 0.5 degrees; killer whale 1.2 degrees; striped dolphin 0.5 degrees. Both the data points and the local averages are displayed in three colours: yellow (∑PCB concentration = < 20 mg/kg); orange (∑PCB concentration = 20–40 mg/kg lw); and red (∑PCB concentration = > 40 mg/kg lw). Full size image

Figure 7 Mean ∑PCB concentrations in stranded and biopsied BNDs (1990–2012), SDs (1991–2009) and KWs (1994–2012) – all cetacean species (all ages). The blue bars are males and the grey bars are females. The lower line is the equivalent ∑PCBs concentrations threshold (9.0 mg/kg lipid) for onset of physiological effects in experimental marine mammal studies6. The upper line is the equivalent ∑PCBs concentrations threshold (41.0 mg/kg lipid) for the highest PCB toxicity threshold published for marine mammals based on marked reproductive impairment in ringed seals in the Baltic Sea6. Error bars = 1SEM. Full size image

Figure 8 (A–D) Box and whisker plots for male (M) v female (F) v unknown sex (U) were generated for ∑PCB and ln ∑PCB concentrations for (A) all stranded HPs (1990–2012); (B) stranded or biopsied SDs (1990–2008); (C) stranded or biopsied BNDs (1990–2012); and (D) stranded or biopsied KWs (1994–2012). The lower line (blue) is the lower PCB toxicity threshold (=9.0 mg/kg lipid, as ∑PCB) for onset of physiological effects in experimental marine mammal studies6. The upper line is the highest PCB toxicity threshold (=41.0 mg/kg lipid, as ∑PCB) published for marine mammals based on marked reproductive impairment in ringed seals in the Baltic Sea6. Full size image

We compared ∑PCB concentrations among regions within the NE Atlantic and Mediterranean Sea. Since our ∑PCB data is lognormal, we are able to use the sample geometric mean to estimate the population median (Tables 1 and 2) and that this estimator is unbiased for doing this10. Mean ∑PCB concentrations and geometric means (with upper and lower 95% confidence intervals) were generated for males and females of each cetacean species and region for all available data (n = 1,009) (Table 1) and for adults only (n = 401) (Table 2). These tabulated data were presented along with minimum and maximum values (i.e. range) for males and females for each species and region (Tables 1 and 2). In general, the ∑PCB sample geometric means (as estimator of population median values) were 25–65% lower than the (arithmetic) means for the same species, but still exceeded even the very highest marine mammal PCB toxicity threshold (∑PCB = 41 mg/kg lipid) for some BND, SD and KW groups in both NE Atlantic and Mediterranean Sea regions (Tables 1 and 2).

Table 1 ∑PCBs exposure (mean; sample median; geometric mean with upper and lower 95% CI; range) (mg/kg lipid) by species, region and sex (all ages). Full size table

Table 2 ∑PCB exposure (mean; sample median; geometric mean with upper and lower 95% CI range) (mg/kg lipid) by species and region (sexually matures only). Full size table

For males and females (all ages), high mean ∑PCB concentrations (>100.0g/kg lw) were found for BNDs off Iberian Peninsula (NE Atlantic), western Mediterranean and northern Adriatic Sea; for SDs in western Mediterranean and for KWs (NE Atlantic) (Fig. 2). Among adults, high individual female ∑PCB concentrations (>100mg/kg lw) were found in KWs from all NE Atlantic regions, in BNDs from England and Wales and in SDs from the western Mediterranean Sea (Table 2). Mean ∑PCB concentrations in adult female KWs (176.8 mg/kg lipid) in the NE Atlantic greatly exceeded mean concentrations in adult female southern resident (55.4 mg/kg lipid) and transient (58.8 mg/kg lipid) KWs10 off British Columbia (NE Pacific). Adult female KWs in the UK and Ireland had the highest mean ∑PCB concentrations (224.8 mg/kg lipid) followed by KWs in the Strait of Gibraltar (215.4 mg/kg lipid). Mean ∑PCB concentrations in adult female BNDs (61.8–158.3 mg/kg lipid) and SDs (84.5–523.7 mg/kg lipid) in some western European populations also markedly exceeded those in southern resident and transient KWs off British Columbia10 (Fig. 3)(Table 2). Mean ∑PCB concentrations greatly exceeded all marine mammal PCB toxicity thresholds in mature (Fig. 3) and immature (Fig. 4) BNDs, SDs and KWs. Individuals for which the maturity status could not be established, mean ∑PCB concentrations still exceeded all marine mammal toxicity thresholds in BNDs and SDs (Fig. 5).

The toxicological data show that these populations greatly exceed concentrations at which severe toxic effects are known to occur. Pathological findings on necropsy consistent with immunosuppression and increased susceptibility to disease included macro-parasitic and bacterial pneumonias, high lung and gastric macro-parasite burdens and generalised bacterial infections (septicaemias). These were regularly found on necropsy in stranded HPs5, BNDs and KWs. In Mediterranean SDs distemper due to cetacean morbillivirus (CeMV) infection9 was frequently seen. Several stranded KWs also had multiple dental infections leading to large mandibular abscesses identified on necropsy12. High ∑PCB contamination can cause immunosuppression7 and may be a significant contributing factor in the death of many of the stranded individuals that had fatal infectious diseases on necropsy.

For necropsied cetaceans stranded in the UK (1990–2012), ∑PCB data was obtained for 706 HPs, 38 BNDs and 7 KWs (Tables 1 and 2). For BNDs (UK), causes of death included trauma (n = 5 male; n = 2 female); infectious disease (n = 2 male; n = 4 female). For the UK-stranded KWs, causes of death included starvation (n = 2), infectious disease (n = 2) and live stranding (n = 2). A cause of death could not be determined for 25 HPs, 11 BNDs and 1 KWs. ∑PCB lipid concentration in UK-stranded male HPs of all ages from 1990–2012 was significantly higher (sample mean = 19.4 mg/kg lw; sample median = 11.5 mg/kg lw; n = 388) than for females of all ages (sample mean = 13.8 mg/kg lw; sample median 8.35 mg/kg lw; n = 318) (ANOVA, P < 0.001) (Fig. 2).

∑PCB lipid concentrations in subgroups of “healthy” male UK-stranded HPs that were generally in good nutritive condition and died of acute physical trauma (sample mean = 13.8 mg/kg lw; sample median = 9.57 lw; n = 201) had significantly lower ∑PCB than male HPs that were in poorer condition and died of a range of infectious diseases (sample mean = 26.5 mg/kg lw; sample median = 18.9 mg/kg lw; n = 120) (ANOVA, P < 0.001). In female UK-stranded HPs that died of acute physical trauma, ∑PCB lipid concentrations (sample mean = 8.50 mg/kg lw; sample median = 6.41 mg/kg lw; n = 144) were significantly lower than female HPs that died due to a range of infectious diseases (sample mean = 16.8 mg/kg lw; sample median = 12.2 mg/kg lw; n = 132) from the same period (ANOVA, P < 0.001) (Fig. 2). Male and female HPs that died of starvation were all in a poor or emaciated condition. Both male HP starvation cases (sample mean = 20.7 mg/kg lw; sample median = 12.5 mg/kg lw; n = 26) and female HP starvation cases (sample mean = 27.9 mg/kg lipid; sample median = 9.74 mg/kg lw; n = 16) had higher ∑PCB concentrations than the physical trauma group. For UK-stranded HPs (both sexes), mean ∑PCB mg/kg lipid was 150–328% greater in infectious disease or starvation cases as compared to trauma cases. The increase in ∑PCB on a sample median basis was 130–198% greater for HP disease or starvation cases as compared to HP trauma cases.

Population Dynamics

Such sustained and elevated PCB burdens are likely to have significant effects at the population level in European BNDs, SDs and KWs. Although we cannot directly and causally link high ∑PCB exposures to cetacean population declines, many of these populations are either very small, or show evidence of major and long-term declines or a significant contraction of range. Only very small KW populations are now found in industrialised regions of Europe13,14,15. There is only one resident population in southern Europe14: a tuna-feeding KW subpopulation in the strait of Gibraltar comprising two pods totalling 36 individuals13. These animals have been monitored every year since 1999. Only six mature females were reported to have calved between 1999 and 2011 and these produced only five calves living more than 1 year. This 6.4% annual female fecundity rate over a 13-year period is one of the lowest recorded reproductive rates for KWs globally and the subpopulation was recommended for listing as “Critically Endangered” due to small population size14. Around NW Scotland and Western Ireland, a small group of only nine KWs is regularly seen15. This group of marine-mammal-eating KWs comprises four adult males, two adult females and three adult female or sub-adult males. Although studied over a 19-year period, no calves have ever been reported within this group15. No other resident or coastal KW groups are found off the Iberian Peninsula or in the Bay of Biscay, although occasional offshore KW sightings occur (e.g.16). Long-term cetacean stranding records in the Netherlands show that KWs were first recorded on the Dutch North Sea coastline in 1783, observed regularly from 1918 to 1963 and completely ceased thereafter until two single KW strandings in 2009 and 201017.

Historic strandings data suggest that multiple BND resident or coastal groups in Europe became depleted or locally extinct in the late-1960s to mid-1970s, including those in the UK (e.g. Morecambe Bay; East coast of England)18 and the North Sea Dutch coast19. The last member of a resident BND population at Arcachon, France, died in 200320 and the small resident BND group (current census n = 27) in the Sado Estuary, Portugal, showed decline due to low calf survival over several decades21. Blubber PCB levels for two females and one male BND (sampled between 1995 and 1997) from this population ranged from 37.1 to 114.0 mg/kg lw. The Mediterranean Sea remains a global PCB hotspot1 where most of the extant marine mammal species, including SDs, BNDs and short-beaked common dolphins (CDs)(Delphinus delphis) have declined over decades14,22. The current IUCN Red List conservation status of BNDs and SDs in the Mediterranean Sea is “Vulnerable”. The Mediterranean CD subpopulation is now classified as “Endangered” after experiencing major and generalized decline over the last 40–50 years, particularly in the northern Adriatic Sea and in the eastern Ionian Sea8. Blubber PCB concentrations are significantly higher in CDs and SDs in the Mediterranean Sea compared to the much more abundant NE Atlantic populations23.

For HPs in the NE Atlantic, a single continuous population ranges from France to northern Norway. A separate much smaller Iberian population is estimated to comprise only 4,398 (CV = 0.92) porpoises24. In the Baltic Sea, two separate HP subpopulations exist: one in the western waters (Kattegat, the Belt Seas) estimated at 40,475 individuals (CV = 0.235) and listed as “Vulnerable” (by HELCOM)25 and another in the Baltic Sea proper, which is now very small and has suffered decades of decline and is listed as “Critically Endangered”8. A history of hunting, high chemical pollutant exposure and accidental entrapment in commercial fishing gear (bycatch) are thought to have been responsible for this decline24. A total of 375,358 (95% CI = 256,304–549,713) porpoises was estimated for EU Atlantic continental shelf waters including the Iberian and western Baltic (sub-)populations (July 2005)16. For the NE Atlantic (UK) HP population, a longer calving interval, lower pregnancy rate and later maturation and higher rates of reproductive abnormalities were also recently identified in a necropsy study of 329 female UK-stranded HPs (1990–2012), as compared to HP populations in much less PCB-polluted regions like Iceland and the Gulf of Maine/Bay of Fundy (NW Atlantic)26. Direct observations of reproductive failure (foetal death, abortion, dystocia or stillbirth) were observed in 25/127 (19.7%) of necropsied mature female HPs in the same study26.

Consideration of alternative causes of mortality

Accidental entrapment in commercial fishing gear (bycatch) is often considered to be the greatest conservation threat globally to small cetaceans8,27. In UK strandings and necropsy data (1990–2013 inclusive) bycatch was diagnosed as the cause of death for 296/615 (48.1%) of UK-stranded common dolphins (Delphinus delphis) (CDs) and 338/1983 (17.0%) of HPs. In contrast, only 9/130 (6.9%) UK-stranded SDs; 4/71 (5.6%) BNDs and 0/7 (0.0%) KWs were diagnosed as bycatch during the same period. Bycatches recorded under the 2013 UK dedicated bycatch sampling programme included 101 days on pelagic trawls and 346 days on static gear vessels and included observed bycatches of 18 HPs, 11 CDs, two SDs and a single BND. Six common dolphins were recorded in pelagic trawls. All other cetacean bycatches were recorded from static net fisheries. Preliminary bycatch estimates for the entire UK fleet during 2013 from systematic observer-based studies provided estimates of HP bycatch of around 1917 animals assuming no pingers (acoustic bycatch deterrent devices) were used (CV = 0.126) and 1652 animals if all UK boats over 12 m used pingers correctly (under EU regulation 812)(CV = 0.147)28. Other estimates for the entire UK fleet in 2013 indicate around 320 CDs and around 470 seals were bycaught28. Bycatch of BNDs, however, was considered too rare to estimate for the entire UK fleet and KW bycatch was not recorded28. A bycatch observer study for the Irish albacore tuna (Thunnus alalunga) drift-net fishery undertaken in offshore waters of the NE Atlantic in the 1990s found the highest bycatch rates in CDs followed by SDs, with much lower numbers of bycaught BNDs and no KW bycatch29. The highest bycatch rates generally occur in the most numerically abundant species in the NE Atlantic region, HPs and CDs, but with much lower rates of bycatch in BNDs and bycatch unrecorded in KWs in recent years30. Very low numbers of bycatches were found in the long-term observer-based studies of the resident European BND populations:- BNDs in Shannon estuary, Ireland (population size = 107–140; zero bycatch recorded since 1993)[S. Berrow, unpublished data]; Sado estuary, Portugal (population size = 27; zero bycatch recorded from 1983–2013) (Manuel E. Dos Santos, pers. com.) and Slovenia, northern Adriatic Sea (population size estimate = 74; 95% CI = 57–90; 2 cases of bycatch since 2002) [T. Genov, unpublished data, see also31]. Collectively, these necropsy and fisheries bycatch observer studies from approximately 1990 onwards indicate that bycatch is unlikely to be driving population declines of BNDs or KWs around the UK or in the wider NE Atlantic during this period.

Other potential causes of cetacean mortality include ship-strike; acoustic disturbance from high-intensity/anthropogenic acoustic sources; nutritional limitation caused by reduced prey availability (e.g. overfishing or climate change); biotoxins (harmful algal blooms) and infectious disease (e.g. cetacean morbillivirus, CeMV)32. Using CSIP UK strandings and necropsy data (1990–2013 inclusive) ship-strike was a relatively infrequent cause of death accounting for only 17/1983 (0.8%) stranded HPs; 0/130 (0.0%) SDs; 0/71 (0.0%) BNDs and 0/7 (0.0%) KWs necropsied in the UK. All tests for biotoxins in UK-stranded cetacean tissues were negative (or at trace levels) since 199032,33.

The potential impacts of anthropogenic high-intensity acoustic sources on cetaceans include mass stranding events (MSEs) that have occurred on a global basis, predominantly involving beaked whales exposed to mid-frequency active sonars in naval exercises34,35,36. Non-beaked whale MSEs with a probable acoustic cause have occurred, such as a 2008 CD MSE in the UK, but are very rare in European waters32. No acoustically-induced MSEs were recorded in HPs, SDs, BNDs or KWs in European waters. Acoustically-induced cetacean MSEs have been recorded in the Mediterranean Sea but also primarily involved beaked whales36,37.

In the Mediterranean and Black Seas, all cetacean species where population assessments have occurred are considered to be declining8,14,22. A range of threats has been proposed including hunting (historically); bycatch; overfishing/prey depletion; habitat degradation; morbillivirus infection and chemical and acoustic pollution8,14,22. Empirical evidence for the effects of prey depletion (e.g. due to overfishing or climate change effects) leading to nutritional limitation is limited for cetaceans, but diet/nutrition effects have been linked to recruitment in KWs and fin whales (Balaenoptera physalus) [reviewed in26]. Such processes would potentially increase PCB toxicity in malnourished animals by mobilising lipophilic PCBs from blubber to other body compartments7,23. A recent global review of Cetacean Morbillivirus (CeMV) found that SDs have been negatively impacted by CeMV epizootics in 1990–1992 and 2006–200838 and probably again in 2011 in the western Mediterranean39. The largest CeMV epizootic in 1990–1992 was associated with very high PCB concentrations9. A CeMV epizootic also occurred in long-finned pilot whales (Globicephala melas) in the Mediterranean Sea in 2006–200740. In contrast, only small periodic CeMV-related mortalities have occurred in individual cetaceans in NE Atlantic region38. No large scale epizootics of CeMV38 or of any other pathogens are known to have occurred in any cetacean species in NE Atlantic region since 1990.