Inter- and intra-specific variation on sensitivity of larval amphibians
to nitrite
C. Shinn a, A. Marco
a
b,*
,
L. Serrano c
EDB (Laboratoire Evolution et Diversité Biologique), UMR 5174, CNRS – Université Paul Sabatier, 118 route de Narbonne,
31062 Toulouse cédex 4, France
b
Estación Biológica de Doñana – CSIC, Apartado 1056, 41013 Sevilla, Spain
c
Departamento de Biologı́a Vegetal y Ecologı́a, C/ Profesor Garcı́a González s/n, 41012 Sevilla, Spain
Abstract
Several authors have suggested that nitrogen-based fertilizers may be contributing to the global amphibian decline. We have studied
the impact of sodium nitrite on early aquatic stages of Epidalea calamita, Pelophylax perezi and Hyla meridionalis larvae from Doñana
National Park (coastal wetland) and P. perezi from Gredos Mountain (high mountain ponds), exposed during 10 to 16 days.
After 8 days of exposure all P. perezi larvae from Doñana presented 100% mortality at 5 mg l—1 N–NO—2 while E. calamita larvae
mortality rates were significantly lower at that concentration after 15 days. However, for H. meridionalis at day 15 no deaths were registered at 5 mg l—1 N–NO— and at 20 mg l—1 N–NO— presented intermediate mortality rates. In Doñana the 10 d LC50 of older H. merid2
2
ionalis larvae was between 20 and 30 mg l—1 N–NO—2 whilst for P. perezi it was below 5 mg l—1 N–NO—2. These results indicate interspecific variation of the sensitivity of larval amphibians to nitrite.
Gredos Mountain P. perezi larvae exposed since the egg stage were highly sensitive to nitrite, with a 16 d LC50 below 0.5 mg l—1
N–NO—2 . The same species in Doñana had a 15 d LC50 between 5 and mg l—1 N–NO—2 . These results suggest that there is also intra-specific variation in sensitivity of amphibian larvae to nitrite: mountain amphibian populations appear to be more sensitive to polluted environments than coastal populations. Geographic and genetic variation and evolutionary adaptation of tolerance may also be the keys to
variation amongst populations of the same species.
Keywords: Nitrite; Amphibian larvae sensitivity; Epidalea calamita; Pelophylax perezi; Hyla meridionalis; Inter- and intra-specific variation
1. Introduction
The number of amphibian species that make the present
threatened species lists is alarming (Houlahan et al., 2000;
Stuart et al., 2004). Global environmental changes, pollution, ultraviolet radiation, diseases and habitat fragmentation are the main causes of such a precarious situation
(Wyman, 1990; Berger et al., 1999; Blaustein and Kiesecker, 2002; Daszak et al., 2003). Nitrogen pollution of aquatic ecosystems is a recurrent phenomenon and is at present
*
Corresponding author. Tel.: +34 954232340; fax: +34 954621125.
E-mail address: amarco@ebd.csic.es (A. Marco).
considered a major influence on the global decline of
amphibian populations (Berger, 1989; Rouse et al., 1999).
Nitrite (NO—2 ) is a natural component of the nitrogen
cycle in ecosystems. Bacteria (Nitrosomones and Nitrobacter) oxidize ammonia and nitrite, respectively, contributing
to the equilibrium between levels of toxic and non-toxic
nitrogen compounds in aquatic environments. Nonetheless, this equilibrium can suffer major alterations due to
too big an input of nitrogen compounds or disequilibrium
in the proportion of the bacteria (Wetzel, 2001). The largest
sources of these nitrogen compounds are anthropogenic:
industrial and urban effluents, runoff and lixiviates from
rubbish dumps, agricultural fields (pesticides and fertilizers) and cattle/pig farms (Carpenter et al., 1998).
Closed aquatic systems like ponds, lakes and aquaculture farms accumulate toxic and sometimes lethal levels
of nitrites. Recommended levels of nitrites in potable water
have been found to be higher than the maximum level tolerable by some amphibian species (Marco et al., 1999).
Nitrogen-based fertilizers have been directly linked to
severe decreases in population size in amphibians inhabiting areas adjacent to agricultural fields. However, there
are populations of certain species that persist in these locations, possibly indicating a higher inherent tolerance to
these pollutants (Blaustein and Wake, 1990).
The contamination of ponds and lakes with nitrogen
compounds has the potential to alter the chemical properties of the breeding sites of many species of amphibians
(Rouse et al., 1999). As many amphibian species are dependent on aquatic habitats for larval development, nitrogen
toxicity poses a constant and unregulated threat via cutaneous and brachial absorption from the surrounding water
(Cooke, 1981). Nitrites present innumerous negative effects
on the physiology and development of amphibian larvae.
Inhaled, swallowed or absorbed nitrites will end up accumulating in the blood stream (Kross et al., 1992). Once
in the blood, these ions react with the hemoglobin, producing methemoglobin (metHb) (Jensen, 2003). This complex
lacks the capacity to take-up or transport O2, causing
hypoxia in body tissues (Bodansky, 1951; Wedemeyer
and Yasutake, 1978). High levels of metHb can lead to
methemoglobinemia (Johnson et al., 1987), already
observed in fish (Cameron, 1971) and amphibians (Huey
and Beitinger, 1980a, 1980b) and is sometimes lethal.
Marco et al. (1999) observed that even at low levels of
nitrite, newly hatched larvae of various anuran species presented a high sensibility. In the presence of nitrite the larvae
had a reduced feeding activity, swam less vigorously,
showed disequilibrium and paralysis, suffered abnormalities
and edemas, and frequently died. These effects increased
with exposure and concentration. Other studies have highlighted the sub-lethal and lethal effects of nitrite on the larval stages of some amphibian species. The activity of the
tadpoles is a response to the surrounding environmental
conditions. Xu and Oldham (1997) found effects upon the
central nervous system with repercussions on the motor
activity resulting in uncoordinated movements, disequilibrium and paralysis. A reduced activity rate may lead to a
higher risk of predation and smaller capacity to face inter
and intra-specific competitive pressures, reducing chances
of survival. The feeding rate of amphibian tadpoles in the
presence of nitrites is also attenuated (Hecnar, 1995). Larvae that eat less and consequently grow at a lower rate reach
metamorphosis with a higher handicap, at a critical phase in
the transition to the terrestrial life-stage.
Some amphibian species reveal a high tolerance to
nitrite exposure (Smith et al., 2004), whilst others are
highly sensitive (Marco et al., 1999). Variations in sensitivity to nitrate between populations of the same species of
amphibians have also been reported (Johansson et al.,
2001). Such scenarios can be explained by the fact that species and populations vary in the genetic basis of their tolerance to environmental stressors like chemical contaminants
(Semlitsch et al., 2000). Genetic variation can lead to geographical adaptation to the ambient levels of certain compounds present in the natural habitat (Miaud and Merilä,
2001).
The purpose of this study was to test the effect of nitrites
on the development, growth, activity and survival of larvae
of species of amphibians from populations of two contrasting natural areas – Doñana National Park and Gredos
Mountain – in controlled laboratory conditions.
2. Material and methods
2.1. Experimental animals
In March 2005 we collected newly hatched (3 to 4 days
old) Epidalea calamita larvae from a temporary pond in
Puebla del Rı́o, Doñ ana National Park (DNP, Sevilla,
Spain; altitude 20; N37° 12 0 1700 , W6°10 0 600 ). The tadpoles
were transported in plastic boxes to the Limnology laboratory in Seville University where the samples were pooled to
gain a genetically varied and representative sample of individuals. Eight concentrations of sodium nitrite were prepared using water collected from a well in DNP (El Bolı́n
Laboratory), and we randomly allocated 14 tadpoles to 3
replicates of each treatment (see Table 1 for setup details
of all experiments). Larvae were exposed to the chemical
during 15 consecutive days.
In April 2005 we collected Hyla meridionalis larvae and
Pelophylax perezi eggs and larvae from large ponds in the
DNP (Almonte, Huelva, Spain; altitude 10; N36°59 330 0 ,
W6°26 579 0 ) and transported them to the El Bolı́n labora-
Table 1
Resume of the experimental setup in each test
Species (origin)
—1
[NO—
N–NO2—
2 ] mg l
Gosner stage at
day 0
No. of individuals
per replica
Average temperature
(°C) ± St. Dev.
E. calamita (DNP)
P. perezi (DNP)
P. perezi (DNP)
P. perezi (GM)
H. meridionalis (DNP)
H. meridionalis (DNP)
0.0, 0.1, 0.25, 0.5, 1.0, 2.5, 5.0, 7.5
0, 5, 15
0, 5, 20, 40, 60, 80,100, 130, 165, 200
0.0, 0.5, 1.0, 2.5, 5.0, 10.0, 20.0
0, 5, 20, 40, 60, 80, 100, 130, 165, 200
0, 5, 15
25
14–18 (eggs)
25
18/19 (eggs)
25
25 (advanced)
14 larvae
12 eggs
5 larvae
19 eggs
4 larvae
11 larvae
21.78 ± 0.74
18.00 ± 0.80
18.60 ± 0.32
14.58 ± 1.32
18.22 ± 0.42
17.54 ± 0.83
(DNP – Doñana National Park; GM – Gredos Mountain).
tory (DNP). Three replicates of twelve P. perezi eggs and
eleven H. meridionalis Gosner (1960) stage 25 larvae were
exposed separately to three levels of NaNO2 during 15
days. Two replicates of four P. perezi and five H. meridionalis larvae at Gosner stage 25 were separately placed in
another array of 10 NaNO2 concentrations for 10 days.
In June 2005 we collected P. perezi egg masses from a
temporary pond in Gredos Mountain (GM) (Navalperal
de Tormes, Avila, Spain; altitude 1910; N40°16 0 34.700 ,
W5°14 0 14.500 ). The eggs were transported to the experimental room at the same altitude and were equally distributed
between seven levels of NaNO2, i.e. the same number of
eggs from each clutch in each replica. Larvae were subjected to NaNO2 for 16 days. We obtained the water used
for preparing exposure solutions from a spring that fed the
pond from where the egg masses were collected.
respected the assumptions for parametric analysis. The
effect of NaNO2 level was assessed across sampling events
for transformed variables for each species using repeated
measures Analysis of Variance (ANOVA) (STATISTICA
6.0®). A one-way ANOVA was performed for specific days
of relevant importance (where significant effects were
observed) in each experiment for each species. Post-hoc
Tukey tests were performed to assess which treatments
were significantly responsible for the effect on the variables.
Median lethal concentrations (LC50) were calculated via
the Probit method using StatsDirect® (version 2, 4, 5). In
all analysis a probability level of p < 0.05 was considered
significant.
3. Results
3.1. Physical and chemical conditions
2.2. Experimental design
We used white opaque plastic cups as holding tanks,
each containing 0.5 l of prepared exposure solution. Disposition of recipients in the experimental zone was done in a
randomly fashion so as to minimize possible external influences. The experimental replicates were kept under conditions of natural photoperiod and lighting. We took water
samples on a regular basis to monitor NO—2 levels. The
Shinn method (Shinn, 1941) was used to determine the concentration of NO—2 in each sample spectrophotometrically
(Hitachi U-1000®). At the mid point of each experiment
we refreshed the mediums with the same initial concentrations. Dissolved oxygen (Hanna HI 9142®) pH (Hanna HI
99100®) and conductivity (Hanna HI 9033 Multi-range®)
of the water of all recipients were measured regularly.
The temperature was registered by means of two Data Loggers® at 30 min intervals and downloaded into a p.c. using
commercial software (BoxCar Pro®) (Table 1). We fed the
tadpoles with boiled lettuce ad libitum.
Throughout the experimental period we registered the
number of hatched embryos (in experiments that begun
with embryos still in the egg stage), dead larvae and abnormalities, as well as the activity of the larvae (number of
individuals in the water column). Dead larvae and excess
organic waste (egg jelly, for example) were removed daily
to prevent oxygen depletion and excessive bacterial and
algal growth. At the end of each experiment we took photos of the surviving tadpoles by placing them on a petri
dish with millimetric paper underneath for a reference
scale. These photos were then digitally analyzed (ArcView
3.2®) on a personal computer to determine the size of the
tadpoles after being subjected to each treatment (body
length: tip of head to cloaca; total length: tip of head to
tip of tail).
2.3. Statistical analysis
Data for mortality, hatching, activity and abnormality
rates, and growth were ARCSIN transformed and
Throughout the experimental period, nitrite levels of the
treatment solutions remained within a 20% deviation of the
established nominal concentrations. This applied for all
treatments except the lower concentrations (0.1, 0.25 and
0.5 mg l—1N–NO2— ) of the E. calamita experiment. In these
treatments nitrite concentrations deviated almost 50%
below the nominal concentrations, most probably a consequence of error in the analytical procedure or reduction of
NO—2 due to bacterial or algal activity.
Levels of dissolved O2, pH and conductivity of the different treatments did not oscillate significantly throughout
each experiment. Average water conditions were within the
natural ranges of the breeding sites. The conductivity of the
water from GM (15 lS/cm) was considerably lower in comparison to that of the water from DNP (400 lS/cm). This
was most likely a consequence of the main water source
being the thawing of snow and ice. The pH levels of GM
waters were also lower (5.6–6.7) than DNP waters (8.0–
8.2). Dissolved oxygen never went below 70% in the test
water at either location. Background nitrite levels in the
water used for the experiments were between 0.000 and
0.021 mg l—1 N–NO—2 .
3.2. Mortality and LC50s
E. calamita was considerably resistant to all NaNO2 levels. After 15 days of exposure the highest mortality rate
was 38% in 7.5 mg l—1 N–NO2— . The 15d LC50 for this species was >7.5 mg l—1 N–NO—2 (Table 2).
The P. perezi eggs collected in Doñ ana were significantly
more sensitive to the highest level of nitrite: 100% mortality
of larvae in 15 mg l—1 N–NO2— after 13 days of exposure
(F2,6 = 61.08, p = 0.000) . No mortality occurred at 5 or
0 mg l—1 N–NO—2 . The 15d LC50 for P. perezi in this experiment was therefore between 5 and 15 mg l—1 N–NO—2
(Table 2).
Gosner stage 25 P. perezi larvae exposed to NaNO2
showed a higher sensitivity than the larvae that were
exposed since the egg stage. Time of exposure also had a
Table 2
Repeated measures ANOVA on the effect of NaNO2 on mortality of larvae in each experiment
Species (origin)
Wilks test
LC50
F
d.f. (effect, error)
P
Days of exposure
Concentration (mg l—1 N–NO2— )
15
15
6
7
10
10
12
16
5
7
10
15
>7.5
5 < LC50 < 15
127.596 (0.0033) §
46.004 (0.0049) §
<5
14.609 (0.0198) §
2.176 (0.2121) §
<0.5
116.744 (0.0045) §
43.583 (0.0137) §
20 < LC50 < 31.665
>15
E. calamita (DNP)
P. perezi (DNP)
P. perezi (DNP)
0.8361
8.0249
2.7860
42, 55
6, 8
18, 18
0.7098
0.0049
0.0179
P. perezi (GM)
2.0129
30, 42
0.0182
H. meridionalis (DNP)
0.5000
2, 6
0.6297
H. meridionalis (DNP)
14.1180
27, 24
0.0000
LC50 values are presented for each experiment, with corresponding number of days of exposure. Probit calculated LC50s are those resultant from a
sigmoid curve with a slope with a significant t, i.e. p < 0.05 (t-test). DNP, Doñ ana National Park; GM, Gredos Mountain; §, calculated via the Probit
method. Standard error in brackets.
ate mortality rates (around 40% and 67%, respectively)
(Fig. 2). Therefore, the 16d LC50 for this species in this
experiment was below 0.5 mg l—1 N–NO—2 .
The presence of NaNO2 did not induce any mortality
among the H. meridionalis larvae that began being exposed
at a more advanced 25 Gosner stage. The 15d LC50 for this
species was higher than 15 mg l—1 N–NO2— . H. meridionalis
larvae that were initially exposed from an earlier 25 Gosner
stage demonstrated significant mortality rates after 5 days
of exposure in the higher NaNO2 concentrations (100–
200 mg l—1 N–NO—2 ; F9,10 = 7.76, p = 0.002). At the end
of the experiment all control larvae survived, whilst intermediate mortality rates were observed in treatments with
40 and 20 mg l—1 N–NO—2 (about 75% and 25%, respectively). All other more concentrated treatments caused a
1.0
1.0
0.8
0.8
Mortality rate
Mortality rate
significant effect on mortality rate. After 6 days of exposure
there were significantly different mortality rates in larvae
exposed to the higher concentrations of NaNO2 (60–
200 mg l—1 N–NO—2 ; F9,10 = 3.17, p = 0.043). After 8 days
of exposure only control larvae (0 mg l—1 N–NO—2 ) survived
(Fig. 1). The LC50 for 8, 9 and 10 days of exposure was
between 0 and 5 mg l—1 N–NO—2 .
Compared with the DNP population, P. perezi from
Gredos Mountain showed significantly different mortality
rates between the various treatments after 9 days of exposure (F6,14 = 4.86, p = 0.007). This mortality commenced
among the higher levels of NaNO2 (5, 10 and 20 mg l—1
N–NO—2 ). At the end of the experiment, treatments with
1–20 mg l—1 N–NO—2 presented a mortality rate of 100%
whilst those with 0 and 0.5 mg l—1 N–NO—2 had intermedi-
0.6
0.4
0.2
0.6
0.4
0.2
0.0
0.0
0
5
20
40
60
80
100
130
165
200
Nitrite concentration [mg L-1 N-NO2-]
Days of
exposure
4
5
6
7
8,9,10
Fig. 1. Effect of NaNO2 on the mortality rate of Doñana National Park P.
perezi larvae, after 4–10 days of exposure. Nominal concentrations have
been used.
0.0
0.5
1.0
2.5
5.0
-1
10.0
Nitrite concentration [mg L N-NO2 ]
12
10
8
9
11
14
Days of
exposure
20.0
-
16
Fig. 2. Effect of NaNO2 on the mortality rate of Gredos Mountain P.
perezi tadpoles, after 8–12, 14 and 16 days of exposure. Nominal
concentrations have been used.
4. Discussion
1.0
4.1. Nitrite levels
Mortality rate
0.8
0.6
0.4
0.2
0.0
0
5
20
40
60
80
100
-1
130
Nitrite concentration [mg L N-NO2
Days of
exposure
4
5
6
7
165
200
-]
8,9,10
Fig. 3. Effect of NaNO2 on the mortality rate of Doñ ana National Park
H. meridionalis tadpoles, after 4–10 days of exposure. Nominal concentrations have been used.
100% mortality rate (Fig. 3), resulting in a 10 d LC50
between 31.665 and 20 mg l—1 N–NO—2 .
It was not possible to calculate the LC50 for certain days
via the Probit® method due to lack of response within some
treatment levels. The LC50 values not calculated in this
way were deduced from the mortality rates at the end of
each experiment (Table 2).
3.3. Sub-lethal effects
P. perezi embryos from DNP exposed to 5 mg l—1
N–NO—2 hatched out prior to those exposed to 0 and
15 mg l—1 N–NO2— . An average hatching rate of 80% was
reached by day 7. The hatching rate of the P. perezi
embryos in the GM was not affected by the presence of
NaNO2 and an average hatching rate of 80% was reached
by the 9th day of exposure.
P. perezi tadpoles from GM that were not exposed to
nitrite had a significantly higher activity rate than those
exposed to the chemical at day 9 (F6,14 = 166.77,
p = 0.000). The presence of NaNO2 did not significantly
affect the growth or activity of E. calamita tadpoles. P.
perezi tadpoles from Doñ ana grew significantly less in the
presence of NaNO2 (F1,63 = 49.04, p = 0.000), whilst the
growth of H. meridionalis tadpoles was not affected. In
the Gredos Mountain P. perezi larvae subjected to
0.5 mg l—1 N–NO—2 grew significantly less than those raised
in pure water (F1,49 = 122.95, p = 0.000).
There was a significant increase of the frequency of abnormalities (mainly arched back) (F6,14 = 3.12, p = 0.037) and
an apparent larger number of edemas in GM P. perezi larvae
exposed to the higher levels of NaNO2.
The nitrite concentrations that were applied in the current study can be found as peak levels in intensive agricultural lands and ponds with organic over-load (McCoy,
1972). Within the Doñ ana Scientific Reserve, nitrite levels
are usually very low, occasionally reaching up to 0.5 mg l—1
N–NO—2 (Serrano and Toja, 1995). However, due to extensive agriculture in the surrounding areas, sandy thus easily
permeable soils, and a highly polluted river bordering one
side (Guadalquivir River), nitrogenous compounds are
found to be at higher levels in some areas at certain times
of the year. During the long, dry summer season, ponds
can be significantly reduced in volume, increasing the concentration of already present substances. In the Tarelo
lagoon (Doñana National Park) nitrite concentrations
ranged from 0.05 to 13.25 mg l—1 N–NO—2 in 2001–2002
(Serrano et al., 2004). Up to 0.1 mg l—1 N–NO—2 has been
detected in streams flowing into the Doñana marshland
(Serrano et al., 2006).
Analysis of nitrite content of the pond water, from
where the clutches were taken reveals low levels (0–
0.021 mg l—1 N–NO—2 ). The year in which the present study
was performed, spring rainfall was late in the season and
scarce. The ponds had recently been replenished when the
water samples (and clutches) were taken, thus not enough
time had yet elapsed for nitrite build-up. Nitrite builds
up in the shallow water table of ponds and stream beds
when oxygen concentration is too low to continue nitrate
reduction during nitrification processes. Nitrate pollution
due to excessive use of fertilizers is an actual problem in
the surroundings of the Doñana National Park. The concentration of nitrates in the Tarelo lagoon has been
detected at 20 mg l—1 N–NO—2 (Serrano et al., 2004).
Nitrate concentration in streams flowing into the Doñana
marshland has increased in recent decades (Serrano et al.,
2006). Nitrogen levels in Doñana and areas other than
the National Park may increase furthermore in the near
future, therefore it is important to assess the possible effect
if they do so. Doñana is one example of a common scenario
in many agricultural-impacted areas.
Gredos Mountain populations are not exposed to significantly increasing levels of nitrites. Furthermore, the highmountain ponds are far from any agricultural influence,
being cow grazing during the summer the only important
(albeit negligible at present) extra input of nitrogenous
waste.
4.2. Inter-specific variation
Data in this study indicated significantly different
degrees of tolerance to NaNO2 levels between species and
within species between habitats. The species that reproduce
in temporary ponds (E. calamita and H. meridionalis) were
more tolerant than the one that breeds in permanent ponds
(P. perezi). Such can be expected as species that are temporary pond breeders are continually subjected to changes in
their habitat. Thus, they would most probably have an
increased genetic tolerance to stress factors that result from
larvae developmental time limitations and deterioration of
the environmental conditions of the pond.
The 10 d LC50 for H. meridionalis was between 20 and
30 mg l—1 N–NO—2 , whilst for P. perezi (in DNP) it was
below 5 mg l—1 N–NO—2 . H. meridionalis therefore seems
to be more resistant to chronic NO—2 exposures than P.
perezi. Natural amphibian habitats do not often register
such high nitrite concentrations for long periods of time,
therefore not being a significant threat to the more tolerant
species (H. meridionalis) under natural circumstances. On
the other hand, US EPA limit for warm water fishes has
been established at 5 mg l—1 N–NO—2 (US EPA, 1986)
equivalent to the 10 d LC50 of DNP P. perezi.
We conclude that the acute exposure to NO—2 has a negative impact on the survival and development of amphibian
larval stages. Toxicity of NaNO2 increased with time and
concentration, as expected. 15 d LC50 values for P. perezi
in GM were not far below those obtained in other studies
with amphibians. Newly hatched larvae of Ambystoma
gracile, Bufo boreas (Anaxyrus boreas), Hyla regilla
(Pseudacris regilla) and Rana aurora also show a high sensitivity to low levels of NO—2 – 15 d LC50 between 1.01 and
1.75 mg l—1 N–NO—2 (Marco et al., 1999). On the other
hand, Smith et al. (2004) found that Gosner stage 26 Rana
catesbeiana (Lithobates catesbeiana) larvae are highly tolerant to nitrite exposure, showing no detectable effects after
15 days of exposure at 10 mg l—1 N–NO—2 .
Inter-specific variation in the sensitivity of amphibians to
contaminated environments has been encountered in a number of previous studies. Marco et al. (2001) observed significant differences in sensitivity to urea fertilizer
concentrations between three species of forest-dwelling
amphibians. In an experiment to assess the effect of nitrite
and nitrate on amphibians from the Pacific Northwest, significant variations in sensitivity were detected among the five
studied species (Marco et al., 1999). Macı́as et al. (2007)
showed that during the first days of exposure to 1 and
3.5 mg l—1 N–NO—2 Bufo bufo embryos were more sensitive
than P. perezi embryos, both collected from the same mountainous field in central Spain. In another study, the salamander Ambystoma tigrinum tigrinum and the wood frog Rana
sylvatica (Lithobates sylvaticus) showed different developmental responses to nitrite exposure. Significant mortality
was caused by exposure to 4.6 mg l—1 N–NO—2 and
0.5 mg l—1 N–NO—2 during the early larval stages of salamander and wood frog, respectively (Griffis-Kyle, 2005, 2007).
4.3. Intra-specific variation
The present study also reveals the existence of different
sensitivity levels between geographically separated populations of the same species. The P. perezi population from
the mountainous area seems to be less tolerant to the pres-
ence of nitrite in the water. The experiments in which exposure was performed since the egg stage, the GM P. perezi
population revealed a higher sensitivity to NaNO2 (15 d
LC50 < 0.5 mg l—1 N–NO2— ) that the DNP population (15
d LC50 between 5 and 15 mg l—1 N–NO2— ). Only by day 9
had 80% of the embryos hatched out from GM clutches,
whilst that same percentage in the DNP clutches had
hatched out by the 7th day of exposure. In Gredos the exposure commenced with embryos at a more advanced Gosner
stage. Egg jelly protects amphibian embryos from certain
environmental aggressions, until hatching out, possibly
protecting them from the toxic proprieties of the chemicals.
The kind of intra-specific variability to nitrite exposure
found here may be partly explained by the different water
temperatures between experiments and resultant variation
in the rate of development. Higher temperatures in the
lowlands could have been an important factor in inducing
faster hatching and development rates, as an increase in
temperature accelerates metabolism and, consequently,
development (Berven, 1982; McDiarmid and Altig, 1999).
Another factor could be due to the low mineral concentration (conductivity) of the water from Gredos Mountain.
Lower total mineral content, therefore lower cation concentrations, may provide a lower buffering capacity in comparison to the water in Doñana.
Nonetheless, water physical–chemical properties were
by no means the only probable factors to create such
intra-specific variation. Morpho–physiological traits may
be yet another explanation for the observed intra-specific
variation in nitrite sensitivity. For example, the existence
of a more sensitive developmental stage, as when the gills
are more exposed to the aquatic medium. Specimens from
the lowland populations, in the laboratory at higher temperatures, may have experienced the most sensitive phase
before the toxic compound had time to take effect. In
colder mountain areas the sensitive phase may have
occurred after the chemical started to take effect, resulting
in a higher toxicity to the larvae. Ortiz-Santaliestra et al.
(2006) have found a higher sensitivity of Pelobates cultripes
to ammonium nitrate during the first 4 days of exposure. In
addition, they found a higher sensitivity of exposed Gosner
stage 19 individuals compared to exposed Gosner stage 13
embryos. Only a small difference in the developmental
stages at the beginning of the exposure was enough to
result in a visible differential effect. Future environmental
impact studies with amphibian species must necessarily
focus on the effects on early life stages.
4.4. Geographic variation
Further still, genetic and geographic variations have
been reported to explain such variability of tolerance
between populations of the same species. Many studies
on the evaluation of the impact of chemical stressors on
amphibian population species result in contradictory conclusions. Variations in response may be due to genetic differences, maternal effects or even geographic adaptation to
changing local environmental conditions (Hecnar, 1995;
Lande and Shannon, 1996). Lowland DNP populations
of P. perezi might have adapted to higher ambient levels
of nitrites in their habitat, whilst the Gredos Mountain
population has not been subjected to such levels over time,
being geographically displaced from major natural or
anthropogenic sources of NO—2 .
Other studies indicate a similar trend. Johansson et al.
(2001) tested the tolerance of different Scandinavian populations of Rana temporaria to nitrate. In Scandinavia, the
nitrate concentrations in lakes and ponds increase from
north to south as a result of a naturally higher productivity
of the habitats, and also due to higher supplementation of
anthropogenic nitrogen (Johansson et al., 2001). The
northern population suffers from reduced growth rates
and metamorphic size in the presence of high concentrations of nitrate, whilst the southern population does not
reveal to be as significantly affected.
Hatch and Blaustein (2003) described differences in
response to nitrate exposure coupled with UV-B radiation
in two species of amphibians at two different elevations. In
the low elevation experiment, UV-B and nitrate together
reduced the mass of larval H. regilla (P. regilla). In the high
elevation experiment the combination reduced the survival
of larval H. regilla (P. regilla). The authors were therefore
able to highlight the potential differences in response to
UV-B radiation between populations that have historically
been exposed to different levels of UV-B.
Macı́as et al. (2007), in a study on the combined effect of
UV-B radiation and nitrite exposure, showed that a P.
perezi lowland population (altitude 10) was less sensitive
than a highland (altitude 1900) population. After 15 days
of exposure to 1 mg l—1 N–NO2— , the lowland population
presented 15.3% survival whilst the highland population
presented 0% survival, at the lowest UV-B radiation (4%
of natural incidence).
Available data indicate that some species have an
increased capacity to adapt to constantly changing environmental conditions (Bridges and Semlitsch, 2000; Semlitsch et al., 2000; Rä sä nen et al., 2003), as those that are
contaminated and suffer the impacts of global climate
changes. Variation in tolerance to chemicals of amphibians
is essential for persistence of populations through survival
and successful reproduction in an increasingly contaminated world (Semlitsch et al., 2000). In order to shield
declining species from chemical contamination, conservation efforts must consider individual, population and geographic variation in tolerance to such contamination.
Overlapping information of geographical distribution of
past and present chemical contamination with the distribution of amphibian population declines is increasingly necessary (Bridges and Semlitsch, 2000). To clearly understand
the processes leading to species declines, this type of information should be accompanied with genetic and fitness
tradeoff studies. Global amphibian population decline is
a fact and the implementation of efficient conservation
and protection plans is urgent.
Acknowledgments
The authors wish to thank Guadalupe Macı́as and Luca
Moura for their practical assistance. This study was funded
by the Spanish Ministry of Education and Science.
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