There is little doubt in anyone's mind that chemical pollution has been the cause of the dramatic, sharp, and in some places ultimate decline of Eurasian otter populations in England and several other countries in Europe. It started in the 1950s, and continued into the 1970s and 1980s (summaries by Jefferies 1989; Mason and Macdonald 1986). Only late in the 1980s and during the 1990s did the trend begin to be reversed, and at the turn of the millennium otters in Europe had recovered in many areas, or at least they were on their way. Similar events took place in North America.
The otters' rapid disappearance from many waters of Europe and North America coincided with a massive increase in the use of organochlorines as insecticides in agriculture, and for various purposes in industry. Several of these substances, such as DDT derivatives, dieldrin, lindane and several toxic heavy metals, were found in damaging concentrations in the tissues of dead otters. Many other carnivores, fish predators, birds of prey and others suffered similar fates, demonstrably due to pesticides and other chemicals (Newton et al. 1993). Most of the compounds that were held responsible for the wildlife deaths have now been taken out of use, and many of the species that were affected by them responded with recovery, or are in the process of returning. However, we still know relatively little about exactly which compounds were responsible for the twentieth century demise of the otter: the possibilities are many (Mason 1989; Mason and Macdonald 1986; Roos etal. 2001).
Moreover, although a great deal is known of the possible effects of various relevant chemicals on individual otters, there is ignorance about their role in populations, which is a very different proposition. Newton (1988), in a seminal paper on the effects of pollution in birds of prey, pointed out that the (often well known) concentration of a chemical that causes death in half of exposed individuals (LD50) bears no relationship to concentrations required to cause a population decline. The reason for this is compensatory mortality or reproduction: some populations may suffer high mortality through pollution, but still maintain their density (possibly limited by availability of resources). Other populations may decline at even very low levels of contamination, because of the loss of just a few individuals that received a high dose. In most of the research on chemical contaminants in otters, of whichever species, there is no information on populations.
The data needed to assess the identity of the pollutants that caused the large decline of Eurasian otter populations in the twentieth century were never collected, and we will never know for sure what happened. However, the otter decline took place at about the same time as that of many raptors and fish-eating birds, which were similarly badly hit, and which were better researched. For their disappearance, at least in Britain, there was good evidence that the use of dieldrin was responsible (Newton etal. 1993). In all probability, therefore, dieldrin was also the culprit in the case of the Eurasian otter, as has been argued strongly by Jefferies (1989).
Dieldrin has now largely disappeared from the environment, but some of the other, potentially seriously nasty, compounds have not. Possibly, they still play a role in depressing otter numbers. For instance, at the time of writing in 2005, Eurasian otters have still not returned naturally to some parts of western Europe (Netherlands, parts of Germany and France, Switzerland). In Shetland and parts of north-east Scotland there is evidence of recent declines in otter numbers. One cannot exclude a potential involvement of contaminants, although other factors may be responsible. A study of such compounds, in their present-day role in mortality of Eurasian otters, may therefore provide further insights in their effects on otter populations (Kruuk and Conroy 1991, 1996; Kruuk etal. 1997a).
There are several good overviews of relevant contaminants, and concentrations in which they are found, for example in Britain by Mason (1989), in Sweden by Olsson et al. (1981) and Roos et al. (2001), in the Netherlands by Broekhuizen (1989) and Traas etal. (2001). In studies with my colleagues we analysed liver samples for organochlorines, first from 113 otters from Shetland, and then from 116 otters from different parts of Scotland. The results were complicated, as expected, but we could draw some tentative conclusions (Kruuk and Conroy 1991, 1996). The samples were analysed for various polychlorinated biphenyls (PCBs), for the organo-chlorines DDE (the breakdown product of DDT), dieldrin (HEOD), lindane (BHC) and mercury (Hg). Some carcasses from Shetland were also analysed for cadmium (Cd), lead (Pb) and selenium (Se).
Almost all of these, with the exception of PCBs and mercury, occurred in concentrations well below the levels known to have any significant lethal or sub-lethal effects in individual mammals or birds (e.g. Bunyan et al. 1975; Jefferies 1969; Robinson 1969; Wren et al. 1988). With the exception of mercury, none was correlated with age of the otter, so they were not likely to accumulate and produce complications later, at least not at the rate at which they were consumed at the time of collection. Further, none of the concentrations of other compounds was correlated with the otters' body condition, and of the substances analysed, therefore, only PCBs and mercury were candidates for causing possible deleterious effects in present-day populations.
PCBs alter the metabolism of vitamin A in the body (as does dieldrin), causing developmental irregularities (including fetal resorption and abortion) and increasing the risk of infections and cancer. Some authors have suspected PCBs of playing a major role in the demise of otters in Britain and elsewhere in Europe (Mason 1989; Mason and Macdonald 1986; Mason and Madsen 1993; Roos et al. 2001). They may be distributed as aerial or aquatic (industrial) pollutants, so they could have an
Geometric mean PCB (ppm wet, in liver)
Figure 12.10 Concentration of PCBs in the liver of Eurasian otters from different parts of Scotland: Shetland otters carried the highest PCB burden. Number of carcasses analysed: Shetland, 14; Orkney, 14; north/north-west, 10; north central, 10; north-east, 33; west, 18; south-west, 13. (After Kruuk and Conroy 1996.)
impact almost anywhere. PCBs occur as a complex of different congeners with different values of chlori-nation; in particular, those with high chlorination values (for example, those numbered 118, 128, 138, 153, 170, 180 and higher, which were also common in our studies) tend to affect the physiology of animals. We undertook analyses for all the compounds separately, but for the present purpose I use the total value of PCB (following Roos etal. 2001).
It was shown in the laboratory that PCBs cause failure of reproduction in female mink (at concentrations in liver lipids of 50 parts per million (ppm); Jensen etal. 1977), and later this value was accepted uncritically as a yardstick for damaging concentrations in otters. Mason (1989) suggested that otter populations with PCB values of 50 ppm (lipid in liver) or more should be declining. This prediction was not borne out in our work: the apparently dense and healthy population of otters in Shetland in the 1990s showed PCB values with a mean greater than 210 ppm lipid (5.5 ppm wet weight) in the liver (Fig. 12.10). This included lactating (i.e. successfully reproducing) females, with values up to 20 times higher than the concentration causing reproductive failure in mink. On mainland Scotland we found a lactating female otter, shot by a keeper in 1987, with 1097 ppm PCB (lipid).
There was no correlation between PCBs and age, so otters appeared to be able to metabolize or excrete this pollutant, at least at the level of concentrations found in Shetland (in fact, they appear to mostly metabolize it; Smit et al. 1994). There was a significant negative correlation between the body condition of otters and the PCB concentration in the liver; the most likely explanation for this was that, in animals in bad condition, body fats are mobilized and consequently the contaminants of those lipids are moved into the liver.
There was no evidence that the Shetland otter population was suffering from the burden of PCBs, and our results, from the apparently healthy, dense and in the 1990s increasing Shetland otter numbers, do not support the Mason hypothesis. Similarly, otters in north-east Scotland showed no obvious decline in density over some 20 years, although their PCB levels were such that they would be called 'of concern' when compared with those in the mink studies (Smit et al. 1994). PCBs at the above concentrations appeared to be relatively harmless in otter populations, but it is possible that they have deleterious effects on individual otters. Whatever the case, the results show that one should not extrapolate results obtained from individual mink in captivity to populations of otters in the wild.
In the south-west of England, strong negative correlations were found between levels of vitamin A in Eurasian otter livers and the presence of several PCBs as well as dieldrin, between 1988 and 1996 (Simpson et al. 2000). This was a period of major recovery of the otter population in Britain, and of significant declines in pollution levels in the animals. Especially remarkable was that otters with highest levels of dieldrin and lowest levels of vitamin A also showed serious eye malfunction, with detachment of the retina in 26 of 131 otters (Williams etal. 2004b); this ailment could easily have been overlooked in many previous post-mortem examinations.
Eurasian otters from other areas in Scotland, including Orkney, the west and south-west of Scotland, and the north-east, showed lower PCB values than those from Shetland. The explanation is that the waters and sediments of the northern North Sea and the north-east Atlantic Ocean are known to contain high concentrations of PCBs, brought by the North Atlantic currents and through atmospheric deposition (references in Wang-Andersen et al.
1993). Animals such as pilot whales around Faroe are now deemed dangerous for human consumption, because of their high levels of PCBs (Simmonds etal.
1994). Pregnant women in the Faroes had dangerously high concentrations of PCBs (Fangstrom et al. 2002), and arctic foxes in Svalbard carry a similarly high burden (Wang-Andersen et al. 1993). Yet, in none of these species is there evidence that numbers are declining. However, there is also no indication in the area that PCB concentrations in the animals have decreased over the last few decades.
In Spain, as in Scotland, Eurasian otter populations have returned or increased over the past few decades, in the face of sometimes still considerable burdens of PCBs (Ruiz-Olmo et al. 1998c). Populations of North American river otters in Ontario, as measured by annual numbers trapped, did not show evidence of an effect of the high PCB levels in the nearby Great Lakes (Wren 1991). All of this does not, of course, exclude the possibility that in some countries (such as the Netherlands) PCB contamination is such that it would affect otter reproduction to the extent of preventing a self-sustaining population. So far, however, there is no firm evidence for this.
Apart from PCBs, the presence of mercury in Eurasian otters in Scotland and Shetland could also be of concern (Kruuk etal. 199 7a). Mercury, in its methylated form, has potentially lethal effects on many wild carnivores, causing damage to the central nervous system with symptoms of lassitude, loss of coordination and paralysis (Wren 1986). It occurs naturally in the environment; in the waters near Shetland, for instance, there are remarkably high concentrations of methyl mercury in fish, due to submarine volcanic activity (Carr etal. 1974; Davies 1981). Because of mercury, many freshwater eels in these islands are unfit for human consumption, by World Health Organization standards (Kruuk et al. 1997a), so their effect on a fish-eating carnivore could be even more serious. Elsewhere, naturally occurring mercury is far outweighed by that produced by agriculture and by industrial sources such as mining, smelting, use of fossil fuels, waste incineration and others, with much of the produced inorganic mercury converted into the lethal methyl mercury (Lindquist and Rohde 1985). Its occurrence is declining (Newton etal. 1993).
Methyl mercury affects individual otters. In river otters in the USA some unpleasant, but useful, experiments by O'Connor and Nielsen (1981) showed that river otters fed with mercury-laced fish died after 6 months, with a mercury level of 110 ppm (dry weight) in the liver. By comparing mercury levels in road-killed Eurasian otters of different ages in Shetland, we found that the animals accumulate it over the years, with a mean concentration of 10 ppp and a maximum of 65 ppm in the liver (Fig. 12.11).
Many sea mammals absorb mercury in relatively high quantities, but are able to use selenium to counteract its toxicity (Koeman et al. 1975; Reijnders 1980; Smith and Armstrong 1978). Thase species accumulate and store selenium at the same rate as mercury, and on analysis the above authors found a near-perfect correlation between the two elements in species such as seals and dolphins. Wren (1984) suggested a similar relationship in semi-aquatic freshwater mammals in Canada. However, in our
Age of otter (years)
Figure 12.11 Older Eurasian otters in Shetland have significantly more methyl mercury in their liver (rs = 0.63, P < 0.001). (After Kruuk and Conroy 1991.)
Age of otter (years)
Figure 12.11 Older Eurasian otters in Shetland have significantly more methyl mercury in their liver (rs = 0.63, P < 0.001). (After Kruuk and Conroy 1991.)
Shetland otters there was no correlation between mercury and selenium levels, suggesting that in these animals no such selenium mechanism has evolved to cope with the mercury problem (Kruuk and Conroy 1991).
On the Scottish mainland, we found a close correlation between annual rainfall and mercury concentrations in Eurasian otters, consistent with an industrial and then atmospheric origin of the mercury. Here, the mean mercury concentrations in otter livers was 9 ppm, similar to that found by Wren (1984) in river otters in Canada. Just as with PCBs, there was a negative correlation between mercury concentration and physical condition (K) of the otter, probably with the same explanation.
Summarizing the occurrence of contaminants in Eurasian otters: the use of laboratory tests on mink to set critical danger levels for contaminants in populations of wild otters is suspect, and should be avoided. At present there is no evidence to suggest that any of the organochlorine compounds or heavy metals is exercising a deleterious effect on otter populations. Even if individuals are affected, this does not necessarily translate into a population impact. We have no good data on the identity of pollutants in past, dramatic, population declines of otters; most evidence points to a significant role for dieldrin, although others, such as PCBs and mercury, cannot be exonerated completely.
For the North American river otter there have been similar concerns about the effects of chemical pollution, although the role of pollution in twentieth century population declines was probably not as dramatic as for the Eurasian otter. Dieldrin was widely used in the USA; high levels were found in river otters around 1980, for example in Georgia (Clark 1981), and are still present in some areas, in fish and water birds in Florida (Rumbold etal. 1996; Marburger etal. 2002). Kinlaw (2005) suggested that physical abnormalities, associated with dieldrin and its subsequent vitamin A deficiency, perhaps cause river otters to be more prone to be killed on roads.
Where the species did disappear in the 1960s to 1980s, it has been quite slow in returning, despite generally low levels of contaminant organochlorines and heavy metals (Elliott et al. 1999; Halbrook et al. 1996). As for the Eurasian otter, there is no good evidence to indicate which compounds were responsible for previous population crashes. In Tennessee, of all the bird and mammal fish predators studied, only the otter was deemed to be at some risk because of the presence of PCBs and mercury. However, this assumed the same sensitivity as found in mink (Sample and Suter 1999), which is just as unlikely as for the Eurasian otter.
In recent years several marine populations of sea otter, river otter and Eurasian otter have been hit by large oil spills, in well publicised incidents. By far the most substantial was caused by the stranding of the Exxon Valdez, in March 1989, in the Prince William Sound of Alaska, which killed both sea and river otters. Oil affects otters in three different ways: it destroys the insulation of the fur, poisons the animals when they ingest it, and kills potential prey. Once oil has entered the ecosystem, effects may be felt for many years because of the presence of various breakdown products, and some of the oil itself may be buried and released later.
The effects of the Exxon Valdez spill on coastal river otters were studied over several years in detail, by Terry Bowyer and his team, who compared the oiled area with a clean one nearby (Bowyer et al. 2003). In the oiled area 12 river otters were found dead immediately after the spill, but this could have been only a fraction of the total. Over the following years 27 radio-otters died, and of those only four would have been found if they had not been equipped with radio-transmitters. Consequently the number of river otters killed by the oil could well have been 80 or more, mostly perishing inside dens or under cover somewhere away from the coast.
River otters that came in contact with the oil, but did not die immediately, were likely to have a more difficult struggle to survive than clean ones. Their blood was shown to have lower haemoglobin levels, their oxygen consumption increased, they dived less often and their prey capture rate decreased (BenDavid etal. 2000).
The effects of the oil on river otters wore off quite quickly. In the first year after the accident, their body condition was lower than that of otters in a control area, but 3 years later this difference had disappeared. The presence of oil derivatives in the otters themselves was assessed from the presence of characteristic components, 'biomarkers' such as haptoglo-bin (which do not necessarily translate into deleterious effects on otters). These had largely disappeared after 3 years, and there were no further measurable differences in otter diet, density or home range size.
Sea otters were probably much harder hit by this same oil disaster. There were a number of different estimates, with the latest taking into account several possible biases: a total of some 750 sea otters died in the Prince William Sound (confidence interval 600-1000; Garshelis 1997). This was considerably lower than some of the earlier assessments (e.g. 2650; Garrott etal. 1993). It proved difficult to determine what proportion this constituted of the population in the sound.
Fortunately, until now, no oil spills of similar magnitude have occurred in the coastal habitats of Eurasian otters. The only recorded otter deaths from such a source were in Shetland in 1978, when the tanker Esso Bernicia spilled oil whilst docking. At least 13 otters died, with symptoms of gastrointestinal bleeding, probably after ingesting the oil during grooming. There were no noticeable and significant effects on populations (Baker etal. 1981).
Apart from the North American river otter and the Eurasian otter, the only other otter species for which chemical contaminants have been cause for concern is the giant otter. In that case, mercury pollution of rivers is caused by gold mining, and in Amazonian giant otter habitats many fish carry concentrations of methyl mercury that far exceed recommended maximum levels for human consumption (Gutleb etal. 1997; Uryu etal. 2001). However, at least in the Peruvian Amazon, the scats of the giant otters contained relatively low levels of mercury (Gutleb et al. 1997), which is open to various interpretations. It is not known whether populations of the animals are affected. Any potential effect of this pollution may well be hidden, for instance by the devastations caused by poaching (Schenck 1997).
In general, therefore, we do not have good evidence that present-day populations of otters, anywhere, are diminished by any specific contaminant. If individual animals are affected and die early (which is quite possible), compensatory mechanisms in populations may mask the effects of this, as explained above. Nevertheless we should be extremely cautious, because, as the past has shown, we may notice the pernicious influence of pollutants only well after the otters have disappeared.
Recently (since the late 1990s), I and other observers have noted a marked decline in otter numbers in Shetland, and, similarly, numbers are well down in some freshwater areas on the Scottish mainland. In such cases it is possible that pollution is rearing its ugly head again, and this should be checked. It is more likely, however, that declining fish stocks are the culprit (see below).
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