Patterns and causes of non-natural mortality in the Iberian
lynx during a 40-year period of range contraction
Alejandro Rodr‘ıguez *, Miguel Delibes
Department of Applied Biology, Estacio‘n Biolo‘gica de Don~ana, CSIC, Avda. Mar‘ıa Luisa s/n, 41013 Seville, Spain
Abstract
We analyse the spatial and temporal variation in non-natural mortality during a 40-year period of strong contraction of the
geographic range of the Iberian lynx (Lynx pardinus), which shrank from 40,600 to 22,300 km2 . We recorded 1258 lynx deaths, an
average of 31.5 losses per year over the study period. Given the reduced lynx population size, especially later in the period (around
1100 individuals), this level of non-natural mortality may have contributed significantly to the quick decline of the Iberian lynx.
Non-natural mortality was not spatially correlated with, and probably did not shape the pattern of, relative abundance of lynx
across its core range, but may have reduced its absolute density. Lynx losses were caused mainly by traps set not only for predator
control but also for rabbits (Oryctolagus cuniculus), the lynx’s staple food. We did not find evidence that non-natural mortality was
higher in small lynx populations through edge effects. The highest mortality levels were recorded in regions where small game was a
valuable economic resource compared with other activities. Mortality decreased throughout the period because of changes in the
prevailing game regimes rather than because of legal protection. The Iberian lynx is now critically endangered and effective protection should be urgently enforced, especially in small game estates, which are environmentally favourable for rabbits but risky for
lynx due to predator control. Lynx reintroductions would be better attempted in traditional rabbit hunting areas. Some big game
estates where small game is not exploited and predators are not controlled may be good candidates for lynx reintroduction too,
provided that habitat is managed towards a suitable interspersion of woody cover and grassland.
Keywords: Extinction; Game management; Land use; Lynx pardinus; Non-natural mortality
1. Introduction
Many ongoing population declines of vertebrates
have their roots in increased mortality prompted by one
or more forms of human intervention, i.e., non-natural
mortality. These forms include deliberate extraction of
individuals for subsistence hunting (Bodmer et al.,
1997), commercial exploitation (Leader-Williams et al.,
1990) and persecution (Breitenmoser, 1998), as well as
indirect sources of mortality related to habitat alteration
(Green and Stowe, 1993), pollution (Beebee et al., 1990)
or introduced predators, parasites, and diseases (Sinclair
et al., 1998; Tompkins et al., 2002). Reconstructing
the spatial patterns, causes, and timing of mortality is
required for diagnosing declines (Caughley and Gunn,
*
Corresponding author. Fax: +34-954-621125.
E-mail address: alrodri@ebd.csic.es (A. Rodr‘ıguez).
1996). Moreover, determining associations between
mortality sources and type of human activity (e.g., land
use or game management) may help to design conservation actions.
In a classification of species based on traits that make
them prone to extinction, the Iberian lynx (Lynx pardinus) has been highlighted as the most vulnerable felid in
the world (Nowell and Jackson, 1996). Its small geographic range was influential in scoring high in this list.
In 1950, the Iberian lynx was already largely confined to
the east-west oriented mountains of southwestern Spain
and southern Portugal, and by 1988 it had lost a further
81% of its range (Rodr‘ıguez and Delibes, 1990). Nonnatural mortality has been presumed to be a major
factor of decline, both at the range scale (Delibes, 1979;
Rodr‘ıguez and Delibes, 1990; Delibes et al., 2000) and
in regional studies (Ferreras et al., 1992; Gonz‘alez,
1998; Garc‘ıa Perea, 2000). However, this perception of
causality has not been supported by objective information at the range scale. It is even unknown whether the
same causes of mortality have operated in all lynx
populations, or whether mortality rates have differed
among them.
In this paper we first analyse the spatial variation of
non-natural mortality across the Iberian lynx range
during a period of 40 years. With this information we
examine whether mortality was related to the distribution of lynx abundance. During the contraction period,
and within continuous populations along the mountain
chains of the core range, lynx abundance increased from
west to east (Rodr‘ıguez and Delibes, 2002). The Iberian
lynx strictly depends upon a single prey species, the
European rabbit (Oryctolagus cuniculus) (Delibes et al.,
2000) whose density shows a similar spatial pattern of
increasing abundance from west to east in southwestern
Iberia (Blanco and Villafuerte, 1993). Here we test the
hypothesis that higher lynx abundance in eastern areas
was also associated with lower intensity of non-natural
mortality.
Second, we examine whether non-natural mortality
co-varied with the size of lynx populations, especially
whether small lynx populations were subjected to increased non-natural mortality compared to large populations, as a result of edge effects (Woodroffe and
Ginsberg, 1998). In a previous study, we concluded that
demographic stochasticity played a role in the extinction
of small lynx populations (Rodr‘ıguez and Delibes,
2003). This conclusion was reached under the assumption that the strength of deterministic agents of extinction was similar in lynx populations of all sizes. We test
this assumption regarding non-natural mortality.
Third, we document temporal changes in rates of
non-natural mortality. We explore whether the legal
protection of the Iberian lynx in Spain in 1973 halted
or substantially reduced the number of losses. Since
large predators were absent in southern Spain, except a
small relict wolf (Canis lupus) population (Blanco et al.,
1990), one might expect little interest in performing
predator control in regions where hunting was oriented
towards big game. Law transgressions after 1973 may
have concentrated in small game estates (Villafuerte
et al., 1998), where we predict higher rates of lynx
mortality.
Finally, we describe how the relative importance of
different causes of mortality has varied across space and
over time. Since dominant land uses and game management were consistent across large regions, we test
whether high mortality levels were associated with particular management styles. We also analyse whether
land uses determine the most likely cause of death. In
small game estates we predict that, due to predator
control and the use of leg-hold traps for rabbit harvesting, traps had a higher impact than firearms.
2. Methods
The material of our analyses was a data base of lynx
reports, obtained during a field survey, which covered
the Spanish fraction of the Iberian lynx range (45,950
km2 , >95% of its total geographic range). Reports dated
between 1950 and 1989 were obtained by interviewing
hunters and gamekeepers, and were subjected to careful
filtering according to their accuracy and reliability.
Details of the field methods and the treatment of the
information have been discussed elsewhere (Rodr‘ıguez
and Delibes, 1990, 1992, 2002). In previous analyses,
two or more different sightings or deaths reported in the
same year by the same source were often counted as a
single report. Now we consider them as separate reports
because we are interested in the kind and frequency of
man-lynx encounters. This resulted in a sample of 3052
reports of which 1258 (41%) were lynx deaths, all of
them attributed to non-natural causes. Reports were
plotted on a 10 km UTM projection grid. We constructed eight maps of distribution and abundance at
five-year intervals, each containing reports obtained
during or after the first year of the interval. The number
of reports per cell, either in five-year intervals or over
the whole period, was an estimate of lynx relative
abundance (Rodr‘ıguez and Delibes, 2002). Lynx populations were defined as groups of occupied cells in the
grid connected by their sides or corners. For each cell,
we recorded the number of lynx deaths and sightings
both in each five-year interval and the whole study period. Assuming that the number of deaths and sightings
was proportional to their respective occurrences, the
relative importance of mortality in each cell was calcu-
Table 1
The distribution of lynx reports in natural regions
Sightings
Deaths
Total
Mortality ratio
WCR
ECR
SSP
WTM
ETM
WSM
CSM
ESM
BM
~
DON
Total
107
70
177
0.40
11
10
21
0.48
36
5
41
0.12
259
199
458
0.43
210
258
468
0.55
45
36
81
0.44
136
155
291
0.53
911
503
1414
0.36
49
12
61
0.20
30
10
40
0.25
1794
1258
3052
0.41
WCR, Western Central Range; ECR, Eastern Central Range; SSP, Sierra de San Pedro; WTM, Western Toledo Mountains; ETM, Eastern
~,
Toledo Mountains; WSM, Western Sierra Morena; CSM, Central Sierra Morena; ESM, Eastern Sierra Morena; BM, Betic Mountains; DON
~ana.
Don
lated as the ratio between death reports and total reports
in a given period, called hereafter ‘‘mortality ratio’’.
Thus, the same absolute number of lynx deaths may
result in either high or low mortality ratios depending
on lynx relative abundance which, in turn, is given by
the total number of reports (see Table 1).
The lynx range was divided into ten natural regions
on the basis of two hierarchical criteria (Fig. 1). We first
considered location in distinct mountain ranges. Under
this criterion we separated lynx populations in the
Central Range, Sierra de San Pedro (SSP), Toledo
Mountains (TM), Sierra Morena, and Betic Mountains
(BM). We then distinguished natural regions within
mountain ranges according to prevalent land uses and
landscapes, as well as the degree of isolation of lynx
populations (Rodr‘ıguez and Delibes, 1992). The western
Central Range (WCR) contained one large lynx population well isolated from two smaller populations in the
east (ECR). Along the Toledo Mountains, big game and
small game prevailed in the west (WTM) and east
(ETM), respectively, during the study period. The Sierra
Morena landscape gradually changes from patchy forest
with little understorey in the west to more dense and
continuous scrubland in the east (Moreira and Fern‘andez Palacios, 1995). In an attempt to capture this geographic variation, we subdivided the long Sierra Morena
into western (WSM), central (CSM), and eastern (ESM)
regions (Fig. 1).
In 1988, we surveyed a sample of 60 estates (total
area 1700 km2 ) along Sierra Morena in order to
quantify the spatial variation in land use attributes,
and to study their association with causes of mortality
within a large fraction of the lynx geographic range.
Estates were assigned to the western (WSM plus CSM)
or eastern sector. We recorded the dominant land use
and hunting regime. From hunting records (also direct
observation when possible) rabbit abundance was
scored as scarce (1), moderate (2), or abundant (3) for
each estate, and a mean score was computed for each
sector. We recorded whether rabbits were exploited,
either by shooting or trapping, and the total rabbit
harvest was divided by the estate area. Major methods
of carnivore control were neck-snares and leg-hold
traps. From keeper reports we assigned a score of 0
(absent), 1 (moderate), or 2 (intense) to each method.
For each estate the sum of snare and trap scores was
taken as an index of intensity of predator control
(range 0–4), and this index averaged across each sector.
Finally, we noted whether fur dealers regularly visited
the estate.
The number of lynx reports recorded in some regions
was small, and so was the number of occupied cells,
which could compromise reliable calculations of mortality ratio (Table 1). Therefore, we limited the analyses
of the spatial variation in mortality to natural regions
which had >30 occupied cells in 1950, with an average
number of reports per cell >2.0. These conditions were
fulfilled by five regions: WCR, WTM, WSM, CSM,
and ESM (Table 1). Calculations were performed with
a Geographic Information System (Eastman, 1999).
Spatio-temporal patterns and causes of mortality were
analysed by means of heterogeneity analysis on subdivided contingency tables, regression analysis, and ordinary mean comparison (Zar, 1984).
3. Results
Fig. 1. The distribution of the Iberian lynx in Spain on a 10 km UTM
grid in 1950. Cells are located in natural regions as follows: 1, Western
Central Range (WCR); 2, Eastern Central Range (ECR); 3, Sierra de
San Pedro (SSP); 4, Western Toledo Mountains (WTM); 5, Eastern
Toledo Mountains (ETM); 6, Western Sierra Morena (WSM); 7,
Central Sierra Morena (CSM); 8, Eastern Sierra Morena (ESM); 9,
~ ). Modified
~ana coastal plain (DON
Betic Mountains (BM); 10, Don
from Rodr‘ıguez and Delibes (2002).
3.1. Spatial patterns of mortality
The mortality ratio varied across natural regions
(Table 1). Contingency table analysis showed that there
were three groups of regions with homogeneous
mortality ratios (G ¼ 34:48, df ¼ 2, P < 0:001). The
mortality ratio was low in Eastern Sierra Morena (0.36),
medium in the western regions (0.40–0.44; WCR, WTM,
WSM), and high in Central Sierra Morena (0.53).
During the study period the Toledo Mountains and
Sierra Morena made up the core of the lynx range
(Fig. 1). Before 1985, there was a positive correlation
between the cell position in a linear transect along Toledo Mountains (lowest value in the western end, highest
in the eastern end) and the mortality ratio calculated in
the same cells, i.e., the mortality increased towards the
east (Table 2(a)). The opposite trend appeared along
Sierra Morena but the correlation was weak (Table
2(a)). Since lynx relative density increased towards the
east along the same transects (Rodr‘ıguez and Delibes,
2002), the relationship between mortality ratio and lynx
density was positive or neutral (Table 2(b)), but not
negative as would be expected if non-natural mortality
was responsible for the geographic pattern of lynx relative abundance.
3.2. Mortality versus population size
Average mortality ratio in cells was independent of
the size of the population they belonged to. In 1950,
seven natural regions contained small lynx populations
(65 cells; Fig. 1), the only ones whose probability of
extinction during the whole period was >0 (Rodr‘ıguez
and Delibes, 2003). The average (±SD) mortality ratio
Table 2
Spearman rank correlations between mortality ratio and (a) cell position along three linear transects through mountain chains in the core of the lynx
geographic range, and (b) lynx density along the same transects, as estimated by the number of reports (see Rodr‘ıguez and Delibes, 2002)
Transect
(a) Cell position
Toledo Mountains
Sierra San Pedro + Toledo Mountains
Sierra Morena
(b) Local density
Toledo Mountains
Sierra San Pedro + Toledo Mountains
Sierra Morena
Period
rs
n
P
<1985
P1985
<1985
P1985
<1985
P1985
0.553
0.245
0.689
)0.007
)0.335
)0.127
23
16
13
13
32
28
0.006
0.360
0.009
0.981
0.061
0.521
<1985
P1985
<1985
P1985
<1985
P1985
0.325
0.567
0.634
0.230
)0.101
0.395
23
16
13
13
32
28
0.130
0.022
0.020
0.450
0.582
0.038
Cell position was expressed by integers increasing monotonically from the western end (value ¼ 1) to the eastern end of mountain ranges
(value ¼ number of cells in the transect). Transects were spaced at least 50 km. Only cells with reports during the period indicated were considered.
Legal protection
0.9
W CR
TM
W SM
CSM
ESM
All reports
0.8
Mortality ratio
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0.0
1950-1959
1960-1964
1965-1969
1970-1974
1975-1979
1980-1984
1985-1989
Fig. 2. Lynx mortality ratios over large natural regions in each five-year period (the two first periods were merged). WCR, Western Central Range;
TM, Toledo Mountains; WSM, Western Sierra Morena; CSM, Central Sierra Morena; ESM, Eastern Sierra Morena.
in cells of small populations (0.36 ± 0.48) was similar
(t ¼ 0:297, df ¼ 241, P ¼ 0:766) to that of cells of large
populations (0.38 ± 0.39). Comparisons yielded nonsignificant differences in any of the seven regions
(t < 1:6, P > 0:120).
3.3. Temporal patterns of mortality
The mean mortality ratio remained fairly stable
above 60% until 1970–1974, then it showed a steady
decrease down to 13% after 1985 (arcsin transformed
ratios; for the whole period: r ¼ —0:876, P ¼ 0:010; after 1970–1974: r ¼ —0:998, P ¼ 0:002; Fig. 2). This
significant decrease was found in all natural regions. In
Sierra Morena, the change in trend coincided with the
legal protection of the Iberian lynx in 1973. In the
Central Range and Toledo Mountains, the decline in
mortality ratio started earlier, in 1965–1969 and 1960–
1964, respectively (Fig. 2). Until 1970–1974 the variation
in mortality ratio between regions was higher (Mann–
Whitney U test on standard deviations; Z ¼ 2:121,
n ¼ 7, P ¼ 0:034) than after that period (Fig. 2).
3.5. Mortality risk associated with land use and game
management
3.4. Causes of mortality
Lynx deaths were assigned to one of four non-natural
causes of mortality, namely trapping, shooting, road
casualties, and other causes. Trapping and shooting
accounted together for >85% of deaths in six five-year
periods, and total percentages were 62% and 26%, respectively. The relative importance of each cause of
mortality remained the same over the 40 years (Fig. 3).
After 1975–1979 losses due to traps decreased whereas
the proportion of lynx shot increased, but these varia-
0.7
tions were not significant. Furthermore, there were no
significant differences in the proportion of causes of
death neither in the whole lynx range nor in any of the
large regions. Among traps the importance of leg-hold
traps tended to decrease over time (arcsin transformed
proportion, r ¼ —0:710, P ¼ 0:079), while the number
of deaths in neck-snares increased (r ¼ 0:915, P ¼ 0:004;
Fig. 4). Road casualties made up 1% of deaths, with a
slight increase in recent times (Fig. 3).
The relative importance of each cause of mortality
varied greatly across large natural regions (G ¼ 101:75,
df ¼ 12, P < 0:001). Three groups of regions with different characteristics could be distinguished. In the western
regions WSM and WCR, the frequencies of hunting/
poaching and ‘‘other causes’’ (mainly kills by dogs) were
highest, whereas losses to leg-hold traps were low (Fig. 5).
In Eastern Sierra Morena, we found the opposite pattern.
The third regional group included Toledo Mountains,
especially the eastern section, and Central Sierra Morena,
where there was a high frequency of deaths by leg-hold
traps, a medium frequency from gunshots, and little relevance of snares and dog kills.
TRAPPING
ROAD KILLS
The geographic variation of land use measurements
along Sierra Morena is summarised in Table 3. Livestocking was the prevalent use in the western sector,
whereas big game hunting dominated in the east. In the
west, the most common hunting regime was small game.
The level of rabbit exploitation varied accordingly. The
estimates of rabbit abundance had similar values all
along the mountain range, and there was agreement
between such estimates and figures of rabbit harvest. In
estates where rabbits were exploited, their mean abundance score was higher (west: t ¼ 3:466, P ¼ 0:002; east:
t ¼ 3:864, P < 0:001) than in those where they were not
OTHER
SPRING TRAPS
0.5
SNARES
0.6
0.4
0.3
0.2
0.1
0.0
Proportion of deaths
Proportion of deaths
0.6
0.5
0.4
0.3
0.2
0.1
Fig. 3. Proportion of lynx deaths attributed to different causes in each
five-year period (the two first periods were merged). Trapping includes
spring traps and neck-snares used in predator control as well as spring
traps set for commercial harvest of rabbits. The category ‘‘Other’’
includes lynx taken or treed by hounds or shepherd dogs, removed
litters and, exceptionally, poisoning.
0.0
Fig. 4. The proportion of lynx losses to spring traps and neck-snares in
five-year periods (the two first periods were merged).
FIREARM
SPRING TRAPS
SNARES
OTHER
Proportion of deaths
0.6
0.5
0.4
0.3
0.2
0.1
0.0
Fig. 5. The relative importance of major causes of mortality across natural regions over the whole study period. The category ‘‘Other’’ includes lynx
taken by shepherd dogs or hounds, lynx shot when treed by dogs, removed litters and, exceptionally, poisoning. Sample sizes are the number of
deaths given on Table 1.
Table 3
Land use, rabbit harvest, and intensity of predator control in a sample of 60 properties along Sierra Morena
West
n
Dominant land use
Livestock
Forestry
Small game
Big game
Hunting regime
Small game
Big game
Rabbit exploitation
Rabbit abundance
Exploited
Not exploited
Rabbit harvest (ind./ha)
Rabbit trapping
Predator control
Occurrence
Intensity
Use of snares
Absent
Moderate
Intense
Use of leg-hold traps
Absent
Moderate
Intense
Fur trade
t
East
Mean
SD
%
27
n
Mean
SD
27
16
11
16
27
27
27
26
2.63
1.73
1.74
70
30
59
33
33
12
21
12
33
0.62
0.72
1.33
63
0.89
0.93
33
33
32
54
31
15
23
31
27
73
36
2.67
1.76
1.84
0.65
1.76
2.52
14.63
3
0.002
13.21
1
<0.001
3.15
1
0.076
1.82
26
30
26
1
0.864
0.891
0.882
0.178
8.67
1
58
2
<0.001
0.121
0.013
7.69
2
0.021
0.95
1
0.330
0.172
0.139
0.150
18
21
0.48
1.03
11.06
1.572
88
6
6
32
74
17
9
50
P
27
3
9
61
33
27
df
%
33
55
15
15
15
G
22
hunted or trapped (Table 3). The proportion of estates
where rabbits were trapped did not vary significantly
across sectors. In contrast, in the western sector, where
small game was the major hunting regime, the occurrence
of predator control was three times higher than in the
east. Furthermore, in estates with a small game hunting
regime the mean (±SD) intensity score of predator
84
0
16
36
control (1.07 ± 1.21) was much higher (t ¼ 3:039,
P ¼ 0:004) than in estates oriented towards big game
(0.33 ± 0.65). Differences in predator persecution were
even more marked between estates where rabbits were or
were not exploited (exploited: 1.14 ± 1.18; not exploited:
0.25 ± 0.57; t ¼ 3:816, P < 0:001), irrespective of their
location within Sierra Morena (Table 3). Neck-snares
were used more frequently in western estates, whereas
intense usage of leg-hold traps was often recorded in
eastern small game estates. The proportion of estates
involved in fur trade was higher in the west but differences were not significant.
4. Discussion
4.1. Assessment of assumptions
The total number of lynx reports increased from old
to recent times: between 300 and 400 reports per period
before 1980; over 500 per five-year period after 1980.
This trend reflects two facts: the number of respondents
able to report old encounters was limited, and the date
and location of a larger proportion of old reports were
not remembered precisely enough to be considered
(Rodr‘ıguez and Delibes, 2002). Rarity of old reports
affected the whole lynx range and probably did not bias
comparisons between natural regions. However, meaningful mortality ratios require reported sightings and
deaths proportional to their actual frequencies. If either
sightings or deaths were selectively remembered, temporal trends in the relative importance of mortality
could be biased. This is unlikely because lynx were rare
and difficult to observe. Indeed most people reported
just one or a few personal encounters with lynx. Lynx
density has been low in more than 80% of its geographic
range throughout the study period (Rodr‘ıguez and
Delibes, 1992, 2002). Lynx deaths and sightings would
be events remarkable enough in an observer’s lifetime to
be remembered (or forgotten) at similar rates. We conclude that mortality ratios may have not been seriously
biased by selective reporting of deaths or sightings.
To our knowledge there are no suitable data to estimate either the natural or total mortality rate of any
lynx population in the past. Albeit imperfect, calculating
mortality ratios was probably the only feasible way to
reconstruct patterns of mortality during the contraction
in the range of the Iberian lynx. The relationship between mortality ratio and the actual mortality rate of
each lynx population is of course unknown, although we
assume a positive correlation. Therefore, we only draw
qualitative conclusions from our analyses.
4.2. Spatial patterns of mortality
Geographic variation in lynx mortality was fairly
consistent with risks derived from dominant land uses.
In eastern Sierra Morena, big game hunting, mostly red
deer (Cervus elaphus) and wild boar (Sus scrofa), was the
main source of income in most properties. Shooting and
trapping of rabbits, and intensive predator control
measures, were regarded as incompatible with keeping
big game. Many game managers believe that these ac-
tivities may disturb and damage wild ungulates. Moreover, in big game estates thick scrubland was allowed to
grow over large areas to provide refuge to red deer,
which might hamper rabbit trapping and hunting. Perhaps more importantly, the substantial economic returns of big game during the second half of the study
‘ pez Ontiveros, 1991) may have compensated
period (Lo
for the partial or complete abandonment of rabbit
hunting. Consequently, a smaller proportion of big
game estates was subjected to predator control, and lynx
mortality in eastern Sierra Morena was lower than in
other large regions. In contrast, high levels of non-natural mortality were recorded where small game hunting
was a valuable resource, i.e., in the Toledo Mountains
(especially the eastern sector; Rodr‘ıguez and Delibes,
1990), western and central Sierra Morena. Small game
was in practice the only hunting option left across many
western rangelands or other places where big game
hunting (e.g., production of deer trophies) was not as
profitable as in the east, because land management was
oriented towards livestocking.
Regional differences in lynx mortality could also be
indirectly related to the distribution of rabbits. It has
been shown that in altered habitats, with a shortage of
natural food or with attractive anthropogenic food
sources, carnivores may prey on livestock or wander
close to human habitations, where the risk of being
killed increases (Hoogesteijn et al., 1993; Sunde et al.,
1998). Similarly, areas with low rabbit density are not
suitable for Iberian lynx breeding, but allow the movement of dispersing individuals (Palomares, 2001) often
at the cost of being exposed to a high risk of non-natural
mortality (Ferreras et al., 1992, in press). In this regard,
the geographic distribution of scores of rabbit abundance (Table 3) should be considered with caution.
These scores were mainly based on records during the
1988 hunting season in exploited populations, or earlier
records in unexploited populations. In estates of eastern
Sierra Morena, where big game was the prime economic
activity, the rabbit harvest was not optimised, i.e., actual
harvest was lower than the potential harvest, and the
scores we assigned were probably conservative. Hence,
rabbit abundance in eastern Sierra Morena may have
been underestimated. Assuming an increasing pattern of
rabbit abundance towards the east along Sierra Morena
and Toledo Mountains (Blanco and Villafuerte, 1993;
Villafuerte et al., 1998), a relative prey scarcity may have
contributed to increased lynx mobility and non-natural
mortality in western regions.
4.3. Did mortality influence lynx relative density?
Persistent exploitation or persecution may seriously
deplete carnivore populations (Reynolds and Tapper,
1996; Helldin, 2000). If the pattern of increasing lynx
density towards the eastern sectors of the range core had
been determined by non-natural mortality, then lynx
density and mortality ratio should have been inversely
related along mountain ranges. In contrast, this relationship was either not significant or positive. Moreover,
whereas absolute lynx density decreased over time, the
pattern of relative density over space has remained fairly
stable (Rodr‘ıguez and Delibes, 2002) in spite of a
marked reduction of direct human pressure over the
study period. Therefore, we conclude that non-natural
mortality may have contributed to reduce absolute
density across the lynx range, but it has not been an
important factor shaping the pattern of local abundance.
One possible explanation is that the Iberian lynx has
not been an explicit target. Predator control may have
been directed towards more conspicuous species such as
raptors (Bijleveld, 1974) and towards carnivores abundant enough to maintain a pelt market (e.g., red fox
Vulpes vulpes, Ruiz-Olmo et al., 1992). Devices employed by Spanish trappers were rather unspecific (Novak, 1987), which indicates that predator control was
not precisely aimed at killing lynx. Until its protection,
the Iberian lynx was an appreciated trophy, but it was
shot only opportunistically during ungulate hunting
(Fern‘andez de Can~ete, 1969). Thus, lynx losses could be
simply proportional to their abundance, in agreement
with the correlation between density and mortality along
large scale transects.
4.4. Was non-natural mortality higher in small populations?
trend towards reduction of non-natural mortality could
have begun partly due to changes in game management,
principally driven by economic factors, as discussed
above. Besides, one would expect that an effective legal
protection would reduce mortality close to zero, whereas
we recorded 495 lynx deaths after 1973 (39.3% of the
data base).
Although no predator species has been free from
deliberate destruction in historical times (Reynolds and
Tapper, 1996), the high mortality ratios detected during
the first half of the study period could have been a new
phenomenon rather than the follow up of a previous
trend. Until 1950, in southern Spain sport hunting was
restricted to an elite who did not seek a direct economic
‘pez Ontiveros, 1981). Therefore,
profit from game (Lo
before 1950 it could be little motivation for intensive
predator control and other expensive game management
if the estate owners were not expecting an established
level of economic returns from hunting activity. The
gradual economic development that took place after
1950 brought spare time and economic resources to a
broader fraction of the population, providing opportu‘ pez
nity to transform hunting into a worthy business (Lo
‘ , 1997). Simultaneously, the
Ontiveros, 1981; Peiro
public administration promoted an intensive predator
control programme based on bounties (see Bijleveld,
1974). The rise of sport hunting as an established economic activity undoubtedly helped to preserve natural
habitats but also provided a stimulus for depleting
populations of lynx and other predators.
4.6. Causes of mortality
It has been suggested that populations living in small
areas experience increased human pressure through edge
effects, which has been argued to explain the higher
vulnerability of small populations to extinction
(Woodroffe and Ginsberg, 1998). For the Iberian lynx,
however, we found similar levels of non-natural mortality in small populations and in adjacent large ones.
Therefore, potential edge effects in small lynx populations did not result in increased mortality. This has been
correctly assumed by Rodr‘ıguez and Delibes (2003), and
gives support to their conclusion that the differential
extinction of small populations was partly due to demographic stochasticity.
4.5. Temporal patterns of mortality and the effect of legal
protection
Mortality ratios were consistently lower after 1973
than before, which might suggest that legal protection of
the Iberian lynx has helped to reduce deaths from nonnatural causes. However, there is indication that the
decrease of non-natural mortality ratios was not necessarily a consequence of legal protection, since in some
regions such decrease started long before banning. This
The relative importance of each cause of mortality
roughly corresponded to dominant land uses. Traps, especially leg-hold traps, were an important cause of death
in rabbit-rich eastern areas as well as in CSM. In these
regions, not only was predator control more intense but
grids of leg-hold traps set for rabbits also caused many
lynx losses as a side effect. Deaths in which dogs were
involved (mainly shepherd dogs but also hounds) were
more frequent in livestocking areas. Due to the prevailing
hunting system in big game estates (hounds chasing deer
towards guns; only one hunting day per year), the proportion of lynx shot was relatively low.
The proportion of deaths due to leg-hold traps halved
in 40 years. This may reflect the progressive abandonment of rabbit extraction with traps, as a consequence of
the decrease of rabbit density (Delibes et al., 2000). In
1953, myxomatosis, a rabbit viral disease, entered and
spread throughout the Iberian peninsula. The subsequent crash of many rabbit populations probably had
direct negative effects on lynx populations (Rodr‘ıguez
and Delibes, 2002), through reduced natality and juvenile survival. It has also been suggested that rabbit
scarcity reduced the profit of rabbit harvesting and
forced rabbit trappers to stop their activity (Delibes et
al., 2000). As a result, the total number of traps set in
the field may have diminished and this may have alleviated lynx mortality to some extent.
The decreasing importance of traps as a source of
lynx mortality over time may also indicate a concurrent
reduction in the intensity of predator control after the
orientation of many estates towards big game, and the
drop of prices in the pelt market. Snares may damage
carnivore pelts (Hall and Obbard, 1987) and, without
the incentive of the pelt trade, keepers may have switched to snares as the main method of predator control.
Snares were cheaper than leg-hold traps and required
less maintenance effort. Late in our study period, snares
were the only method authorised by the administration.
The gradual abandonment of trapping may have automatically raised the relative importance of firearms after
1979, although the steady increase in the number of
hunting licenses until 1990 may have played a role too
(Rodr‘ıguez and Delibes, 1990). Finally, road casualties
started to be measurable only at the end of the study
period. This probably reflected the development of the
road network and traffic density, in accordance with the
results of previous analyses in the Iberian lynx
(Rodr‘ıguez and Delibes, 1990; Ferreras et al., 1992) and
other carnivore species (Clarke et al., 1998; Philcox et
al., 1999).
5. Conservation implications
From 1950 to 1989 we recorded an average of 31.5
lynx losses per year due to non-natural causes. Since we
were not aware of every death, this value underestimates
the actual figure. It is hardly possible to quantify the
potential consequences of such mortality levels on past
lynx population dynamics. Nevertheless, a loss rate of
this magnitude may have been unsustainable if the average total population size over the whole study period
were close to the 1100 individuals estimated for its last
quarter (Rodr‘ıguez and Delibes, 1992). This is especially
true in an unfavourable context of reduced reproduction
due to persistent rabbit scarcity (Rodr‘ıguez and Delibes,
2002). Synergies between food shortage and non-natural
mortality might be responsible for one of the clearest
accounts of quick decline amongst carnivores. Pronounced declines of large carnivores are usually attributed to deliberate persecution (Woodroffe, 2001). The
case of the Iberian lynx is remarkable in that apparently
high levels of non-natural mortality took place even if
predator persecution was not targeted towards lynx,
which were hunted opportunistically and caught in traps
set mostly for other predators and rabbits.
The data we present indicate that many Iberian lynx
were killed during the first 16 years of legal protection.
Non-natural mortality has been demonstrated to be a
serious threat to the few remaining lynx populations
(Ferreras et al., 2001), especially because most of them
are very small (Rodr‘ıguez and Delibes, 1992). Moreover, this threat persists today. Lynx losses due to traps
and traffic have been regularly reported during the last
decade (Delibes et al., 2000; Ferreras et al., 2004). Legal
protection has little effect if not followed by appropriate
monitoring and law enforcement, accompanied by information campaigns to increase the awareness of people living or hunting in lynx areas. Among other
enforcement measures (Delibes et al., 2000), game
managers should be liable for any lynx deaths occurring
in their estates, especially if the cause was related to the
use of leg-hold traps or other illegal practices. Parallel
efforts should be directed to develop standardised, effective, and selective methods of predator control
(Reynolds and Tapper, 1996) for exceptional use where
authorised.
Our results have direct relevance to the conservation
of remnant lynx populations as well as the selection of
sites for reintroduction. In both cases a relatively high
density of rabbits is needed (Palomares et al., 2001).
However, this condition can be satisfied under quite
different land uses and hunting regimes, which also differ
in the associated risk of mortality.
Small game estates, especially those of the eastern
half of the lynx range, are usually located in places
naturally favourable for rabbits. These sites are also the
most suitable and the most risky for lynx conservation.
Even a few small game estates, embedded in a landscape
largely managed for big game, produced an annual lynx
mortality rate of 25% through predator control (authors’ unpublished data). Therefore, predator control
should be banned in lynx areas without exception or,
alternatively, intensive surveillance should ensure that
only selective methods are used. In small game estates
containing lynx populations, the best urgent conservation option could be buying the hunting rights (i.e.,
suppress hunting and predator control) until a self-sustainable formula is found. At a later stage, hunting
could be restricted to other small game species (e.g., redlegged partridge, Alectoris rufa), with the aim of increasing the amount of rabbits available to predators. In
this respect, it has been suggested that the Iberian lynx
may outcompete other rabbit predators (Palomares et
al., 1996). In the long-term, management could ideally
lead to the recovery and exploitation of rabbit populations in coexistence with lynx.
A medium risk of non-natural mortality occurs in
landscapes devoted to livestocking, but overgrazing may
not allow rabbits to reach high densities (Soriguer et al.,
2001). The lowest risk occurs in big game estates because
(1) there is little motivation for predator control, (2)
hunting activity is minimal during the year, and (3) the
method of hunting allows lynx to easily escape guns.
Besides, hunting activity is concentrated in one or a few
days, which provides the best opportunities to enforce
predator protection. On the question of drawbacks, big
game management often results in homogenisation of
habitats and reduction of patchiness favourable to
rabbits and lynx (Delibes et al., 2000). Since big game
estates tend to be large, one solution could be preserving
patchy lots with a suitable vegetation structure for lynx
foraging (Palomares, 2001), which may not entail substantial economic losses for the main land use. Management oriented towards compatible mixed hunting
regimes within the same estate may not only diversify
sources of income but would also benefit lynx populations.
The decline of the Iberian lynx has continued during
the past decade. The number of known populations
have decreased from nine in 1989 to two in 2002, and the
geographic range has contracted a further 88%, from
223 cells in a map for the period 1985–1989 to 26 cells in
a map built with information collected between 1999
and 2001 (Rodr‘ıguez and Delibes, 1992, 2002;
Rodr‘ıguez, 2002). With just two lynx populations left,
there is not much choice regarding where to focus in situ
conservation efforts. Regarding the selection of localities
for lynx reintroduction, we would recommend estates
naturally (historically) productive for rabbit hunting.
These areas may have a favourable combination of
abiotic factors and landscape structure, allowing rabbit
populations to thrive. These conditions may be difficult
to mimic artificially in less suitable places (Calvete et al.,
1997; Moreno et al., 2004), whereas in big game estates
costly habitat management will be also required. Rabbit
hunting may coexist with the reintroduced lynx population as long as non-natural causes of mortality will be
completely removed.
Acknowledgements
We are grateful to the numerous people who provided lynx reports, to Giulia Crema for efficiently pro~uela and Anja
cessing the data base, and to Javier Vin
Molinari-Jobin for commenting on the manuscript. The
field work was funded by an agreement between ICONA
and CSIC. Our research was supported by the Ministry
of Science and Technology through the grant BOS20012391-C02-01.
References
Beebee, T.J.C., Flower, R.J., Stevenson, A.C., Patrick, S.T., Appleby,
P.G., Fletcher, C., Marsh, C., Natkanski, J., Rippey, B., Battarbee,
R.W., 1990. Decline of the natterjack toad Bufo calamita in Britain:
palaeoecological, documentary and experimental evidence for
breeding site acidification. Biological Conservation 53, 1–20.
Bijleveld, M., 1974. Birds of Prey in Europe. The MacMillan Press,
London.
Blanco, J.C., Rodr‘ıguez, A., Cuesta, L., Reig, S., del Olmo, J.C., 1990.
El lobo en Sierra Morena. In: Blanco, J.C., Cuesta, L., Reig, S.
(Eds.), El lobo (Canis lupus) en Espan
~a. Situacio
‘n, problem‘atica y
apuntes sobre su ecolog‘ıa. ICONA, Madrid, pp. 61–68.
Blanco, J.C., Villafuerte, R., 1993. Factores ecolo
‘gicos que influyen
sobre las poblaciones de conejos. Incidencia de la enfermedad
hemorr‘agica. Report to Instituto Nacional para la Conservacio
‘n de
la Naturaleza, Madrid.
Bodmer, R.E., Eisenberg, J.F., Redford, K.H., 1997. Hunting and the
likelihood of extinction of Amazonian mammals. Conservation
Biology 11, 460–466.
Breitenmoser, U., 1998. Large predators in the Alps: the fall and rise of
man’s competitors. Biological Conservation 83, 279–289.
Calvete, C., Villafuerte, R., Lucientes, J., Osacar, J.J., 1997. Effectiveness of traditional wild rabbit restocking in Spain. Journal of
Zoology, London 241, 271–277.
Caughley, G., Gunn, A., 1996. Conservation Biology in Theory and
Practice. Blackwell, Cambridge, MA.
Clarke, G.P., White, P.C., Harris, S., 1998. Effects of roads on badger
Meles meles populations in south-west England. Biological Conservation 86, 117–124.
Delibes, M., 1979. Le lynx dans la P‘eninsule Ib‘erique: r‘epartition et
r‘egression. Bulletin Mensuel de l’Office National de la Chasse, No.
sp. Sci. Tech. Le lynx, 41–57.
Delibes, M., Rodr‘ıguez, A., Ferreras, P., 2000. Action Plan for the
Conservation of the Iberian Lynx in Europe (Lynx pardinus).
Council of Europe Publishing, Strasbourg.
Eastman, J.R., 1999. Idrisi32. Guide to GIS and Image Processing.
Clark University, Worcester.
Fern‘andez de Can
~ete, J., 1969. Gu‘ıa de la caza en Espan
~a. Ministerio
de Informacio
‘n y Turismo, Madrid.
Ferreras, P., Aldama, J.J., Beltr‘an, J.F., Delibes, M., 1992. Rates and
causes of mortality in a fragmented population of Iberian lynx Felis
pardina Temminck, 1824. Biological Conservation 61, 197–202.
Ferreras, P., Delibes, M., Palomares, F., Fedriani, J.M., Calzada, J.,
Revilla, E., 2004. Proximate and ultimate causes of dispersal in the
Iberian lynx Lynx pardinus. Behavioral Ecology 15, in press.
Ferreras, P., Gaona, P., Palomares, F., Delibes, M., 2001. Restore
habitat or reduce mortality? Implications from a population
viability analysis of the Iberian lynx. Animal Conservation 4,
265–274.
Garc‘ıa Perea, R., 2000. Survival of injured Iberian lynx (Lynx
pardinus) and non-natural mortality in central-southern Spain.
Biological Conservation 93, 265–269.
Gonz‘alez, J.A., 1998. Non-natural mortality of the Iberian lynx in the
fragmented population of Sierra de Gata (W Spain). Miscel-l’ania
Zoolo
’gica 21, 31–35.
Green, R.E., Stowe, T.J., 1993. The decline of the corncrake Crex crex
in Britain and Ireland in relation to habitat change. Journal of
Applied Ecology 30, 689–695.
Hall, G.E., Obbard, M.E., 1987. Pelt preparation. In: Novak, M.,
Baker, J.A., Obbard, M.E., Malloch, B. (Eds.), Wild Furbearer
Management and Conservation in North America. Ministry of
Natural Resources, Toronto, pp. 842–861.
Helldin, J.O., 2000. Population trends and harvest management of pine
marten Martes martes in Scandinavia. Wildlife Biology 6, 111–120.
Hoogesteijn, R., Hoogesteijn, A., Mondolfi, E., 1993. Jaguar predation
and conservation: cattle mortality caused by felines on three
ranches in the Venezuelan Llanos. In: Dunstone, N., Gorman,
M.L. (Eds.), Mammals as Predators. Oxford University Press,
Oxford, pp. 391–407.
Leader-Williams, N., Albon, S.D., Berry, P.S.M., 1990. Illegal
exploitation of black rhinoceros and elephant populations: patterns
of decline, law enforcement and patrol effort in Luangwa Valley,
Zambia. Journal of Applied Ecology 27, 1055–1087.
‘pez Ontiveros, A., 1981. El desarrollo creciente de la caza en
Lo
‘ reas
Espan
~a. In: Actas del Coloquio Hispano-Franc‘es sobre las A
~a. Servicio de Publicaciones Agrarias, Ministerio de
de Montan
Agricultura, Madrid, pp. 271–297.
Lo
‘pez Ontiveros, A., 1991. Algunos aspectos de la evolucio
‘n de la caza
en Espan
~a. Agricultura y Sociedad 58, 13–51.
Moreira, J.M., Fern‘andez Palacios, A., 1995. Usos y coberturas
vegetales del suelo en Andaluc‘ıa. Seguimiento a trav‘es de im‘agenes
de sat‘elite. Consejer‘ıa de Medio Ambiente, Junta de Andaluc‘ıa,
Sevilla.
Moreno, S., Villafuerte, R., Cabezas, S., Lombardi, L., 2004. Wild
rabbit restocking for predator conservation in Spain. Biological
Conservation, in press.
Novak, M., 1987. Traps and trap research. In: Novak, M., Baker, J.A.,
Obbard, M.E., Malloch, B. (Eds.), Wild Furbearer Management
and Conservation in North America. Ministry of Natural Resources, Toronto, pp. 941–969.
Nowell, K., Jackson, P., 1996. Wild cats. Status survey and conservation action plan. IUCN, Gland.
Palomares, F., 2001. Vegetation structure and prey abundance
requirements of the Iberian lynx: implications for the design of
reserves and corridors. Journal of Applied Ecology 38, 9–18.
Palomares, F., Delibes, M., Revilla, E., Calzada, J., Fedriani, J.M.,
2001. Spatial ecology of Iberian lynx and abundance of European
rabbits in southwestern Spain. Wildlife Monographs 148, 1–36.
Palomares, F., Ferreras, P., Fedriani, J.M., Delibes, M., 1996. Spatial
relationships between Iberian lynx and other carnivores in an area
of south-western Spain. Journal of Applied Ecology 33, 5–13.
Peiro
‘, V., 1997. Gestio
‘n Ecolo
‘gica de los recursos cineg‘eticos.
Universidad de Alicante, Alicante.
Philcox, C.K., Grogan, A.L., Macdonald, D.W., 1999. Patterns of
otter Lutra lutra road mortality in Britain. Journal of Applied
Ecology 36, 748–762.
Reynolds, J.C., Tapper, S.C., 1996. Control of mammalian predators in
game management and conservation. Mammal Review 26, 127–156.
Rodr‘ıguez, A., 2002. Lince ib‘erico. In: Palomo, L.J., Gisbert, J. (Eds.),
Atlas de los mam‘ıferos terrestres de Espan
~a. DGCN-SECEMSECEMU, Madrid, pp. 302–305.
Rodr‘ıguez, A., Delibes, M., 1990. El lince ib‘erico (Lynx pardina) en
Espan
~a. Distribucio
‘n y problemas de conservacio
‘n. ICONA,
Madrid.
Rodr‘ıguez, A., Delibes, M., 1992. Current range and status of the
Iberian lynx Felis pardina Temminck, 1824 in Spain. Biological
Conservation 61, 189–196.
Rodr‘ıguez, A., Delibes, M., 2002. Internal structure and patterns of
contraction in the geographic range of the Iberian lynx. Ecography
25, 314–328.
Rodr‘ıguez, A., Delibes, M., 2003. Population fragmentation and
extinction in the Iberian lynx. Biological Conservation 109, 321–
331.
Ruiz-Olmo, J., Grau, J.M.T., Puig, R., 1992. Comparacio
‘n de la
evolucio
‘n de las poblaciones de zorro (Vulpes vulpes L., 1758) en el
NE ib‘erico en base a datos histo
‘ricos (siglos XVIII–XIX) y actuales
(siglo XX). Miscel-l’ania Zoolo
’gica 14, 225–231.
Sinclair, A.R., Pech, R.P., Dickman, C.R., Hik, D., Mahon, P.,
Newsome, A.E., 1998. Predicting effects of predation on conservation of endangered prey. Conservation Biology 12, 564–575.
Soriguer, R.C., Rodr‘ıguez Sierra, A., Dom‘ınguez, L., 2001. An‘alisis de
la incidencia de los grandes herb‘ıvoros en la marisma y vera del
Parque Nacional de Don
~ana. Organismo Auto
‘nomo de Parques
Nacionales, Madrid.
Sunde, P., Overskaug, K., Kvam, T., 1998. Culling of lynxes Lynx lynx
related to livestock predation in a heterogeneous landscape.
Wildlife Biology 4, 169–175.
Tompkins, D.M., Parish, D.M.B., Hudson, P.J., 2002. Parasitemediated competition among red-legged partridges and other
lowland gamebirds. Journal of Wildlife Management 66, 445–
450.
Villafuerte, R., Vin
~uela, J., Blanco, J.C., 1998. Extensive predator
persecution caused by population crash in a game species: the case
of red kites and rabbits in Spain. Biological Conservation 84, 181–
188.
Woodroffe, R., 2001. Strategies for carnivore conservation: lessons
from contemporary extinctions. In: Gittleman, J.L., Funk, S.M.,
Macdonald, D., Wayne, R.K. (Eds.), Carnivore Conservation.
Cambridge University Press, Cambridge, pp. 61–92.
Woodroffe, R., Ginsberg, J.R., 1998. Edge effects and the extinction of
populations inside protected areas. Science 280, 2126–2128.
Zar, J.H., 1984. Biostatistical Analysis, second ed. Prentice Hall,
Englewood Cliffs.