O R I G I NA L A RT I C L E
Eawag_09354
doi:10.1111/evo.12859
Rapid evolutionary loss of metal resistance
revealed by hatching decades-old eggs
Patrick Turko,1,2,3 Laura Sigg,4 Juliane Hollender,5 and Piet Spaak1,2
1
Department of Aquatic Ecology, Swiss Federal Institute of Aquatic Science and Technology (Eawag), CH-8600 Dübendorf,
Switzerland
2
Institute of Integrative Biology, ETH Zurich, CH-8092 Zurich, Switzerland
3
4
E-mail: [email protected]
Department of Environmental Toxicology, Swiss Federal Institute of Aquatic Science and Technology (Eawag), CH-8600
Dübendorf, Switzerland
5
Department of Environmental Chemistry, Swiss Federal Institute of Aquatic Science and Technology (Eawag), CH-8600
Dübendorf, Switzerland
Received November 17, 2014
Accepted January 4, 2016
We investigated the evolutionary response of an ecologically important freshwater crustacean, Daphnia, to a rapidly changing
toxin environment. From the 1920s until the 1960s, the use of leaded gasoline caused the aquatic concentration of Pb to increase
at least fivefold, presumably exerting rapid selective pressure on organisms for resistance. We predicted that Daphnia from this
time of intense pollution would display greater resistance than those hatched from times of lower pollution. This question was
addressed directly using the resurrection ecology approach, whereby dormant propagules from focal time periods were hatched
and compared. We hatched several Daphnia genotypes from each of two Swiss lakes, during times of higher (1960s /1980s) and
lower (2000s) lead stress, and compared their life histories under different laboratory levels of this stressor. Modern Daphnia
had significantly reduced fitness, measured as the population growth rate (λ), when exposed to lead, whereas those genotypes
hatched from times of high lead pollution did not display this reduction. These phenotypic differences contrast with only slight differences measured at neutral loci. We infer that Daphnia in these lakes were able to rapidly adapt to increasing lead concentrations,
and just as rapidly lost this adaptation when the stressor was removed.
KEY WORDS:
Adaptation, Daphnia, paleobiology, Pb.
It is important to understand the evolutionary response of natural populations to anthropogenic environmental perturbation, as
these perturbations constitute a large and growing fraction of
the selective forces experienced by natural populations (HilleRisLambers et al. 2011). To investigate such evolutionary responses,
remains of organisms (tissues, eggs, seeds, environmental DNA;
Brendonck and Meester 2003) can be investigated over long time
periods by using paleogenetic reconstructions from environments
with a known environmental change. In aquatic systems, such reconstructions have been used, for example, to correlate changing
taxon composition (Rellstab et al. 2011) and population genetic
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structure (Frisch et al. 2014) through time with cultural eutrophication. Reconstructive studies are informative about how populations have changed over time but are unable to address the
selection process on phenotypes that involve complex metabolic
pathways or plastic gene expression (Jeppesen et al. 2001). Controlled evolution experiments (Kawecki et al. 2012) allow the
measurement of such relevant phenotypic traits, but suffer from
genetic drawbacks related to low experimental population size.
Namely, they can fail to capture relevant natural standing genetic
variation, their populations may be too small to produce many
new mutations, and deleterious alleles may become fixed by drift
(Barrett and Schluter 2008). These problems may be partially
C 2016 The Society for the Study of Evolution.
2016 The Author(s). Evolution Evolution 70-2: 398–407
R A P I D L O S S O F L E A D R E S I S TA N C E
overcome if microbial populations are kept at high effective population sizes long enough to capture chance beneficial mutation
(Wiser et al. 2013). However, such experiments do mostly not account for important sources of genetic diversity such as invasions,
hybridization, or introgression common in natural environments
(Magalhães and Matos 2012).
The above problems may be overcome by “resurrection ecology” (Kerfoot et al. 1999), which allows the simultaneous investigation of genotypic and phenotypic variation of natural populations decades or centuries apart by hatching or germinating
dormant eggs or seeds (Orsini and Schwenk 2013), which were
exposed to different selective forces.
Evolutionary responses of natural populations have been
studied with respect to anthropogenic disturbances such as climate warming (Deutsch et al. 2008), nutrient pollution (Frisch
et al. 2014), fishing (Miethe et al. 2010), and acidification (Derry
et al. 2010), but one worldwide disturbance that has received little
evolutionary scrutiny is the enormous pulse of lead (Pb) released
to the atmosphere during the 20th century. Emission of lead increased dramatically after World War II, partly due to coal burning
(Settle and Patterson 1980) but mainly attributable to the addition
of tetra-ethyl lead to gasoline (Renberg et al. 2001). This anthropogenic lead reached a maximum fallout into lakes in the 1970s
of approximately two orders of magnitude over background levels
(Renberg et al. 2001), and contaminated the entire biosphere to
the point that the diet of Americans was enriched in lead over 100fold (Settle and Patterson 1980). Past histories of lead pollution
can be reconstructed from lake (Renberg et al. 2001) and marsh
(Shotyk et al. 1996) sediments as it has nearly zero mobility in
those substrates (Thomas et al. 1984).
Lead is acutely and chronically toxic to both vertebrates
(Scheuhammer 1987) and invertebrates (Williams and Dusenbery
1990), and if it reached aqueous concentrations imposing
even slight reductions to survivorship or fecundity, it should
have imposed significant selective pressure for tolerance or
detoxification. Despite this, the evolution of lead resistance has
been poorly studied, although adaptation to other heavy metals
has been investigated. These studies have typically been short
term, and typically involved plants (Klerks and Weis 1987;
Macnair 1987; Prasad 1995) or benthic organisms (Levinton
et al. 2003) at locally highly polluted sites. Indeed, Antonovics
and Bradshaw (1970) showed rapid evolution of Zn tolerance
over short distances. In contrast, in our study we investigated
adaptation to a global pollution force experienced over decades.
Members of the genus Daphnia produce hatchable propagules (Brendonck and Meester 2003) that can be retrieved from
accurately datable lake sediments (Krishnaswamy and Lal 1971).
These sediments can also be used to reconstruct lakes’ physicochemical histories (Rellstab et al. 2011). Daphnia (Crustacea:
Anomopoda) are planktonic filter feeders that consume algae and
bacteria and are consumed by fish and invertebrate predators,
making them a central species in many aquatic food webs worldwide. Their reproductive mode is cyclically parthenogenic: periods of asexual reproduction during favorable conditions alternate
with male production and sexual reproduction for brief periods
when conditions worsen. Sexually produced eggs are deposited in
a sclerotenized sac (ephippium) that sinks to the sediments, where
they may hatch the following spring or be buried intact. Ephippial eggs remain viable in sediments for an unknown amount of
time, and have often been experimentally hatched after burial for
decades (e.g., Weider et al. 1997), whereas Frisch et al. (2014)
reported hatching after a period of centuries. Chemical analysis
of same sediments allows us to couple genotypes and phenotypes
with the conditions under which they were produced, even if very
widely separated in time. Daphnia are widely used in ecotoxicology studies (OCDE 2004) and have been identified by the
NIH as one of the 13 model organisms for biomedical research
(http://www.nih.gov/science/models/) to understand the genomic
responses to environmental stressors.
Pb is toxic to cladocerans (Arambašić et al. 1995), but it is
very unclear at what concentration this toxicity should be expected
to impose selection for resistance. Estimates of LC50 are made
over inconsistent time periods and often use different species (cf.
Cooper et al. 2009 vs. Mager et al. 2011), and are not designed to
detect the subtle fitness differences that can result in adaptive evolution. Even the longest term and most sensitive tests (e.g., Cooper
et al. 2009) may lack the statistical power to detect impacts much
smaller than a halving of survivorship or reproduction. Complicating the matter, all recently tested animals should be considered
post-pollution and may have a complex evolutionary history of
resistance and loss. Despite this lack of clarity, we note that researchers have found major reductions in seven-day fecundity at
Pb concentrations ranging as low as 4.5 μg/l (Cooper et al. 2009)
to 55 μg/l (Nys et al. 2014). As the toxic effect of Pb appears to
be approximately log-linear (Nys et al. 2014), and as fitness differences much smaller than detectable in short experiments may
have major evolutionary consequences, we considered it worth
investigating whether Pb toxicity may have exerted selection for
resistance at naturally observed concentrations (e.g., 2 μg/l Chau
et al. 1970). Toxin resistance may impose fitness costs in unpolluted environments (Sibley and Calow 1989). For example,
copper-resistant clones of Daphnia longispina were found to have
higher respiration and lower reproduction than nonresistant clones
when raised in unpolluted media (Agra et al. 2011). Therefore,
we predicted that Pb resistance, if present, would decline as the
selective pressure of Pb pollution was relaxed between the 1970s
and the present.
Here, we investigated the evolution of lead resistance of the
ecologically important crustacean Daphnia, by hatching dormant
eggs that were produced after decades of increasing lead pollution
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and comparing their toxicological response to those genotypes
produced after 30–40 years of lead decline. We use paleolimnological methods to reconstruct past Pb loading. We hypothesize that Daphnia evolved some degree of lead resistance by
the 1970s, and, assuming that this resistance carries a cost, that
this resistance has been lost by now. We test whether this loss of
resistance can be attributed to evolutionary change versus replacement by comparing the phenotypic change to change at neutral
markers.
Materials and Methods
SAMPLE SITES AND SEDIMENT RETRIEVAL
Greifensee (47.355659, 8.674454 Switzerland) is a eutrophic,
peri-alpine lake of medium size (8.5 km2 surface area; 33 m maximum depth) that has undergone a sharp decline in total phosphorus concentration for the last 30 years (Keller et al. 2002),
indicating a shift from hyper-eutrophic (polluted) to eutrophic
(less polluted). Sediment cores (14 cm diameter) were obtained
from the deepest point of the lake in June 2009. These were sliced
on-site into 1.4 cm sections, while excluding an outer sediment
ring of approximately 1 cm. These cores were dated by comparison to reference cores obtained by Albrecht et al. (1998) and
dated using 210 Pb and 137 Cs. Hallwilersee (47.279713, 8.216452)
is a limnologically similar lake (Elber et al. 2001) located 36 km
from Greifensee. Sediment cores (6 cm diameter) were obtained
from the deepest point of the lake (46 m) in July 2012. These
were transported intact to the laboratory, and kept in the dark at
4°C until processing. Cores were sliced longitudinally and photographed. Because the sediments of this lake form annual layers,
we dated the core visually by comparison against reference cores
obtained by Ambühl (1985) and Elber et al. (2001), which were
dated radiographically as above.
In the laboratory, each sediment slice was passed through
a 224 μm sieve to obtain Daphnia ephippia, which were immediately placed in Petri dishes filled with filtered lake water
(Greifensee; 0.45 μm), and incubated under hatching conditions:
20° C with a 16:8 light: dark regimen. Hatchlings were placed
in 100 ml jars filled with filtered lake water and fed chemostatgrown Acutodesmus obliquus, cultured in WC medium (Guillard
1975) at a concentration of 1 mg C/l. Algae was added every
day and water was changed twice per week. After these hatchlings reproduced, a single neonate was chosen from each clone
for identification.
We attempted to hatch eggs from before the period of historic
Pb pollution, from near the peak of pollution, and from the recent
period of recovery. Unfortunately, Daphnia did not hatch from
the older, prepollution sediments. We therefore chose test clones
from times of historical high pollution and recent low pollution,
as described below.
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DAPHNIA IDENTIFICATION
DNA of hatchlings was isolated using the HOTSHOT protocol
(Montero-Pau et al. 2008), and was used to identify them using
microsatellite analysis and/or ITS-RFLPs. From the Greifensee
hatchlings, 10 loci were amplified in a multiplex protocol (Brede
et al. 2006) and analyzed using a 3130 XL sequencer (Applied
Biosystems, Foster City). Microsatellite peaks were scored using Peak Scanner (Applied Biosystems), and the data were analyzed using two Bayesian clustering algorithms as implemented
in STRUCTURE 1.1 (Pritchard et al. 2000) and NEWHYBRIDS
(Anderson and Thompson 2002), with model parameters as in
Brede et al. (2009). Results from these two algorithms were compared and only those animals identified as D. galeata with probability greater than 0.95 were retained. This identification was
confirmed using ITS-RFLPs (Billiones et al. 2004), and again,
only unambiguous D. galeata were retained. Finally, D. galeata
and D. longispina mitochondrial haplotypes were discriminated
by 16S-RFLP (Schwenk 1993; Brede et al. 2009). These three
identification methods were combined, and random lines were
chosen for experimentation from those identified unambiguously
as D. galeata.
As our hatching success was more modest in Hallwilersee,
identification and clone choice was less stringent. We amplified a panel of nine microsatellite markers in a multiplex protocol (DaB10/14, Dp512, SwiD1, SwiD10, SwiD12,
SwiD14, SwiD4, and SwiD5; (Brede et al. 2006), and analyzed them on an ABI 3130 XL sequencer (Applied Biosystems). The lengths of these fragments were inferred using
STRand software (Toonen and Hughes 2001) version 2.4.59
(http://www.vgl.ucdavis.edu/STRand) and binned into integer alleles using the R package MsatAllele. (Alberto 2009). The binned
data were combined with our panel of reference genotypes made
up of 57 D. galeata, 30 D. longispina, 31 D. cucullata, and 49
D. galeata ×longispina hybrids drawn from lakes covering much
of western Europe North and South of the Alps (Möst et al. 2012).
This combined dataset was clustered using Discriminant Analysis
of Principal Components (DAPC) in R using the package “adegenet” (Jombart 2008). These analyses revealed that the Daphnia
hatched from the “high lead” period clustered most closely with
D. galeata ×longispina hybrids, whereas those hatched from the
“low lead” period clustered with D. longispina.
SEDIMENT LEAD MEASUREMENTS
Sediment cores were sliced with acid-washed zinc blades into
segments three to five years of deposition. Approximately
1.5 ml of wet sediment was extracted from each slice, avoiding
the sediments within 0.5 cm of the outside of the core, and freeze
dried. A total of 40–60 mg of dried sediments were weighed
and digested with 4 ml 65% HNO3 and 1 ml 30% H2 O2 in a
high-performance microwave digestion unit (Ultraclave, MLS,
R A P I D L O S S O F L E A D R E S I S TA N C E
Pb concentration in the sediments of two Swiss lakes,
Greifensee and Hallwilersee, deposited over the last century.
Points are the center of measurement intervals on the date axis,
Figure 1.
and the mean of two measurements on the Y axis.
Leutkirch, Germany). After dilution with nanopure water, the
concentration of Pb (isotope 206 Pb) was measured by HR-ICP-MS
(Element 2 High-Resolution ICP-MS, Thermo-Finnigan, Thermo
Fisher, Waltham, MA). This measured concentration was used
to calculate the actual concentrations in the sediments based on
the above subsamplings and dilutions. Each sediment slice was
sampled twice, and all samples were randomized prior to measurement. If the two Pb concentration measurements of a slice
differed by more than 10% of the mean, the data from these samples were discarded. This occurred in three of 16 sediment layers
from Greifensee, and two of 20 sediment layers from Hallwilersee.
Otherwise, the mean concentration was used.
CHRONIC Pb EXPOSURE LIFE-HISTORY EXPERIMENT
From Greifensee, five pure clonal lineages of D. galeata were
randomly selected from each of three time periods: 1977–1981,
1989–1992, and 2001–2004. These periods corresponded to different levels of Pb pollution in the Greifensee: 71 μg/g (in
sediments) in the 1970s, 40 μg/g in the 1990s, and 17 μg/g in
the 2000s (Fig. 1). From Hallwilersee, three clones were chosen
from the period 1983–1987 (1980s), and three from the period
2000–2012 (2000s). These dates correspond to Pb levels of 30–
40 μg/g and 10–20 μg/g, respectively (Fig. 1). Daphnia were
raised under common conditions for three generations before the
experiment started, to avoid any maternal effects.
For the experiment with Greifensee clones in 2009, five
neonates from each clone line were individually exposed to three
concentrations of lead: 0, 0.1, and 0.2 mg/l (control, low, high).
The “high” exposure corresponds to 5% of the LC50 for Pb measured in D. carinata (Cooper et al. 2009); these data were chosen
as a baseline in the absence of information on our study species.
Exposure medium was prepared by diluting a stock solution of
Pb(II)NO3 in filtered Greifensee lake water. Under these conditions, we calculate that Pb may precipitate as PbCO3 (s) and
Pb5 (PO4 )3 Cl(s), leaving dissolved Pb as PbCO3 (aq) and Pb2+
with a concentration of 0.01–0.02 μM. Daphnia were therefore
likely exposed to both dissolved and precipitated Pb(II). Control water was filtered and handled the same as exposure water,
without the addition of Pb. Every 24 h, 100 ml of exposure or
control medium was added to acid-washed, autoclaved jars, algae
was added to a concentration of 1 mg/l C, and Daphnia were
transferred to this renewed medium via large bore pipette. Births
and deaths were recorded for 14 days. Thus, we performed a full
factorial life-history experiment: 3 time periods × 5 clones × 3
lead levels × 5 replicates (minus one lost clone line) = 210 jars.
In 2012, this experiment was repeated using hatched Daphnia
from Hallwilersee. We exposed three clone lines from each of
two time periods to a single Pb treatment of 0.15 mg/l and a
control treatment. Each clone × Pb combination was replicated
four times. One clone failed to reproduce in all replicates and was
discarded for analysis. Therefore, the study design was 2 time
periods × 3 clone lines × 2 lead levels × 4 replicates (minus one
nonreproductive clone) = 40 jars. Procedures were the same as
above.
STATISTICAL ANALYSIS
Birth and death data over 14 days were used to perform analyses
of survivorship and population growth rate. All analyses were
performed in R statistical software (R Core Team 2013).
Survivorship
When the experiments were terminated after 14 days, less than
50% of animals had died. We therefore compared each treatment’s
Kaplan–Meier product limit estimator, which incorporates data
from individuals who outlived the experiment to estimate each
group’s survivorship function, using the “survival” R package.
Survival was regressed against lead, sediment age, clone (nested
in sediment age), and the interaction between lead and sediment
age, with a logistic hazard distribution.
Although the effect of lead was significant in this model,
survivorship in the two lead media “high” and “low” did not significantly differ, so we attempted to merge them for analysis as
a single Pb treatment. We similarly merged Greifensee sediment
depths “1989–1992” and “2001–2004” as a single “recent” hatching stratum. Whether these mergers were statistically appropriate
was assessed using AIC values and likelihood ratio tests (L-test,
Zuur et al. 2009); significant differences between merged and
unmerged models would indicate unacceptable information loss.
We believe that both mergers are both statistically and biologically justified: the two lead levels are both above the level at
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PAT R I C K T U R KO E T A L .
which chronic effects are expected, and the Daphnia from the
1990s and 2000s were born during periods of relatively low lead
pollution compared to the 1970s. After the above mergers, the
significance of model terms was assessed by serially refitting the
model without the least significant term from the previous model,
and subjecting these nested models to likelihood ratio testing
(Zuur et al. 2009). Once the best model was obtained, each treatment was compared against each other treatment by refitting the
model while adjusting which treatment was considered the baseline. Where appropriate, the resulting P-values were Bonferonni
corrected.
Population growth rate (λ)
The population growth rate was calculated by producing a daystructured Leslie matrix for each clone × lead level treatment; the
real component of this matrix’s dominant eigenvalue corresponds
to λ. We used the bootstrap method to estimate SEs around this
value to enable statistical comparison. Each bootstrap replicate
began by randomly selecting individuals (with replacement) to
form a pseudo-population the same size as the experimental population. A Leslie matrix was constructed from these data, and λ
was extracted as above and saved. This procedure was repeated
1000 times for each clone in each Pb treatment.
Statistical comparison of these pseudo-values was done using a modified ANOVA. The treatment and error sums of squares
(SS) were calculated as usual, but the mean square error (MSE)
was calculated by dividing SS error by the degrees of freedom produced by the number of individuals in the experiment
rather than the number of bootstrap pseudo-values. This MSE was
used in the calculation of F-values, and the significance of these
F-values was again assessed against distributions with degrees of
freedom produced by the true size of the experiment. Because both
the treatment and error sums of squares were inflated at the same
rate, the mean squares, F-values, and P-values were unaffected
by the number of bootstrap iterations. Because the Greifensee
survivorship data were better explained by merging the two lead
treatments and the two most recent sediment depths, and because
the merger of these factors had no influence on their significance
in the fecundity models, we merged also them for the calculation
of population growth rates.
The bootstrapped population growth rate was regressed
against lead, sediment age, clone (nested in sediment age), and
the interaction between lead and sediment age. Differences between treatments were assessed using Tukey’s Honest Significant
Differences (HSD).
Neutral genetic change over time
Genetic differentiation between subpopulations separated in time
was quantified using several measures of genetic distance. We
calculated Hedricks’ Gst (Hedrick 2005) and Jost’s D (Jost 2008)
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using our microsatellite markers. These statistics were used to
create a genetic distance matrix for each lake, in which the populations from each time period were compared to each other.
In Greifensee, we performed these tests on ephippial eggs
retrieved from sediment cores and genotyped as above. These
eggs were taken from sediment depths corresponding to the years
1966, 1974, and 2004. After genotyping, we rejected any individuals not positively identified as D. galeata, leaving 4, 30, and
14 individuals in the individual years. In Hallwilersee, as our test
clones were drawn randomly from the whole hatched population,
we performed the genetic distance analysis on an unselected sample. We therefore used microsatellite data from 26, 18, 43, and 47
eggs, isolated from 1950, 1960, 1970, and 2000.
Results
Pb LEVELS IN SEDIMENTS
Lead in the Greifensee sediments was measured from 1920 until
2009, and that of Hallwilersee was measured from 1923 to 2012
(Fig. 1). In Hallwilersee, Pb rose from background levels in 1920
(1 μg/g dry sediment) to a peak in 1974, and began declining in
1976, but has still not reached pre-Pb levels. The peak concentration in Greifensee was reached the same time as in Hallwilersee,
however Pb levels in Greifensee were almost twice as much as in
Hallwilersee. Pb decline in Greifensee was more precipitous than
in Hallwilersee, and levels have reached those similar to those of
Hallwilersee in the 1920s.
SURVIVORSHIP
The survivorship of Greifensee hatchlings was affected by Pb
exposure (deviance = 5.18, df = 207, P (χ2 ) = 0.027), sediment
age (deviance = 8.35, df = 206, P (χ2 ) = 0.0037), as well as their
interaction (deviance = 3.72, df = 205, P (χ2 ) = 0.054). Despite
the Pb × sediment age interaction P-value of 0.054, removal of
this term resulted in a higher AIC value (486.7 vs. 488.4), so it
was retained. There was no significant clone × Pb interaction
(deviance = 11.27857, df = 0, P (χ2 ) = 0.99), nor did the
clones significantly differ (deviance = –28.25046, df = –26,
P (χ2 ) = 0.35).
Survivorship of the “old” Greifensee Daphnia (i.e., those
hatched from 1977–1981), did not significantly differ between
the controls and the lead treatments (z = 0.651, P = 0.57).
Old Greifensee Daphnia from the lead-polluted time period and
recent individuals from the relatively unpolluted time period
(i.e., hatched from 1989–1992 or 2001–2003) showed similar
responses to control treatments (z = 0.368, P = 0.73). In contrast,
survivorship of the recent Daphnia exposed to lead was significantly less than the control treatment (z = 2.65, P = 0.007, or
0.0478). These results indicate that Greifensee Daphnia from the
1960s survive as well as recent Daphnia whether they are exposed
R A P I D L O S S O F L E A D R E S I S TA N C E
Greifensee
Hallwilersee
1.0
0.9
Fraction Surviving
0.8
0.7
0.6
Old
Recent
0.5
Pb
0.4
Control
0.3
0
2
4
6
8
10
12
14
0
2
4
Day
6
8
10
12
14
Day
Figure 2. Survivorship of Daphnia hatched from old (lead pollution) and recent (after lead pollution) Greifensee and Hallwilersee
sediments exposed to Pb or control water.
to chronic lead stress, but the survival of Daphnia from the 1990s
and 2000s is significantly reduced by lead exposure (Fig. 2).
When fitting the survivorship model to Hallwilersee hatchlings, we rejected the clone × Pb exposure interaction (deviance
= –5.524126, df = –10, P (χ2 ) = 0.85), as well as the main
effect of clone (deviance = –12.28225, df = –10, P (χ2 ) =
0.27). There was also no significant interaction between Pb exposure and sediment age (deviance = –2.5998, df = +1, P (χ2 )
= 0.107), nor a main effect of Pb (deviance = –0.613, df = –1,
P (χ2 ) = 0.433). The effect of sediment age, however, remained
significant (deviance = 16.368, df = 47, P (χ2 ) < 0.0001, Fig.
2). In other words, recent Hallwilersee hatchings survived worse
than older hatchlings, but neither was significantly affected by Pb.
POPULATION GROWTH RATE
Greifensee clones significantly differed in their response to Pb
exposure (clone × Pb interaction, F12, 209 = 152, P < 0.0001), as
well as implicitly (clone main effect, F12, 209 = 581, P < 0.0001).
These differences were significantly attributable to the main effects of Pb exposure (F1, 206 = 1990.5, P < 0.0001), sediment
age (F1, 206 = 1850, P < 0.0001), and their interaction (F1, 206 =
4301, P < 0.0001). The population growth rates of the old clones
were slightly increased with the addition of Pb (0.05 ± 0.006, P <
0.0001, Tukey’s HSD), whereas the growth rates of recent clones
were severely diminished (–0.22 ± 0.005, P > 0.0001, Tukey’s
HSD; Fig. 3).
As in Greifensee, the population growth rate of Hallwilersee
Daphnia significantly differed by clone (F3, 59 = 1376, P <
0.0001), and these clones differed in their response to Pb (F3, 59 =
142, P < 0.0001). The main effect of Pb exposure was highly significant (F1, 46 = 243, P < 0.0001), as was the effect of sediment
age (F1, 46 = 2264, P < 0.0001), and their interaction (F1, 46 =
434, P < 0.0001). As in Greifensee, “new” Daphnia’s population
growth rate was significantly reduced by Pb exposure (–0.44 ±
0.003, P < 0.0001), whereas that of the “old” Daphnia increased
slightly but was basically unaffected (0.06 ± 0.002, P < 0.0001;
Fig. 3).
Genetic differentiation
Our microsatellites show that there is essentially no genetic differentiation through time in either of these lakes, perhaps because of
the buffering and lagging effect of the large ephippial pool (Brendonck and Meester 2003). In Greifensee, Gst was 0.29 between
1966 and 1974, 0.42 between 1974 and 2004, and 0.30 between
1966 and 2004. Jost’s D over those periods was 0.21, 0.30, and
0.21, respectively. In Hallwilersee, Gst was 0.18 between 1950
and 1960, 0.36 between 1960 and 1970, 0.32 between 1970 and
2000, and 0.27 between 1950 and 2000. D was 0.13, 0.28, 0.25,
and 0.27 between those periods, respectively.
Discussion
In this study, we demonstrated dramatic fitness differences between old and modern phenotypes when confronted with a
widespread historical environmental stressor. Daphnia clones
hatched from two lakes, which were produced during a time of
peak Pb pollution, did not display reduced fitness when experimentally confronted with this metal, whereas recent hatchlings
were severely impacted. By employing the techniques of resurrection ecology, we were able to show clear phenotypic change
over decades, where more traditional analysis of, for example,
unhatched sedimentary resting eggs or other Daphnia remains
preserved in the sediment could not have detected these fitness differences under different environments. These differing
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PAT R I C K T U R KO E T A L .
Hallwilersee
Greifensee
1.5
Population growth rate (λ)
1.3
1.1
0.9
0.7
Old
Recent
0.5
Control
Figure 3.
Pb added Control
Pb added
Population growth rate of Greifensee and Hallwilersee hatchlings exposed to Pb-treated and control water. Error bars are
99% confidence limits.
phenotypes were detected in populations separated in times over
which little neutral genetic change occurred, suggesting that the
changes were adaptive and directional.
Other studies have employed resurrection techniques on
Daphnia, but usually have investigated evolutionary change in
response to changing nutrient environment (Frisch et al. 2014) or
the ecological consequences thereof (Hairston et al. 1999). In toxicology, we are aware only of the hatching studies of Derry et al.
(2010), which demonstrated evolutionary change of copepods to
acidification. In contrast, the experiments that we present here
represent some of the first evidence of evolutionary adaptation to
the worldwide fluctuation in anthropogenically released Pb, using
genotypes that were actually produced during times of high pollution levels. Remarkably, given the global scope of this pollution
and its contamination of much of the biosphere, its evolutionary
consequences have been little investigated. The few studies that
explicitly examine adaptation to Pb have usually been done on
small sites of intense pollution (Macnair 1987), with sessile organisms (Klerks and Levinton 1989), over short time periods, and
have generally inferred adaptation by comparison to plants taken
from nonpolluted sites (Klerks et al. 2011). In contrast, by hatching decade-old resting stages, we may be able to infer a direct
evolutionary chain.
In Greifensee, we believe that the past and present populations that we examined were linked through time. The population genetic structure of the Daphnia community in this lake has
been investigated in some detail (Brede et al. 2009). The resident
D. longispina was hybridized with and slowly replaced by the
invasive D. galeata during a time of intense nutrient enrichment,
which closely preceded our target hatching period. Assuming
that preindustrial Daphnia had little resistance to Pb, and noting
the widespread nature of Pb pollution, we think it is reasonable
to suppose that our hatched D. galeata acquired their resistance
404
EVOLUTION FEBRUARY 2016
over time in their source region and not abruptly upon entering
Greifensee. Even so, given the time frame of Pb pollution (Shotyk
et al. 2003), they would not likely have begun to acquire this resistance much sooner than the 1920s—an adaptation period of only
four decades. We believe that the modern genotypes we hatched
are the descendants of this population, indicating that the loss of
resistance we observed may be adaptive change. These past and
present genotypes were identical at one nuclear and one mitochondrial marker, and were clustered together by 10 microsatellite loci. Further, the “old” and “modern” clones tested in this
study did not differ in their population growth rate in the control
treatment.
This identity may also arise by infiltration into the lake of
other, non-Pb-resistant D. galeata genotypes over time, but this
is unlikely given the small degree of differentiation we have measured over the study period. This population continuity accords
with Daphnia’s well-established high level of spatial differentiation (De Meester 1996), thought to be due to local adaptation
(Brendonck and Meester 2003) and selection against migrants,
except of course when changing environmental conditions alter
the selective landscape (Brede et al. 2009). Second, it is unlikely
that nonadapted genotypes could have persisted in the landscape,
as lead contamination was a general feature of lakes in Switzerland (Moor et al. 1996; Von Gunten et al. 1997). The extent of
Pb contamination may have somewhat differed in various lakes,
depending on the relative importance of atmospheric and wastewater inputs, but the typical time course of lead as illustrated in the
sediments of Greifensee and Hallwilersee is also found elsewhere
(Moor et al. 1996; Von Gunten et al. 1997). The physicochemical
conditions of lakes in this region mostly differ over depth and over
seasons in the water column of the lakes, with anoxic conditions
in the hypolimnion during summer, but were similar in lakes of
the same region. Given the above, it seems to us a reasonable
R A P I D L O S S O F L E A D R E S I S TA N C E
supposition that the fitness differences we detected are two points
on a continuous evolutionary chain.
The environmental history of Hallwilersee is more complex.
In addition to the eutrophication and re-oligotrophication experienced by many lakes in the region (Correll 1998), managers
of Hallwilersee began oxygenating the hypolimnion in the year
1986 (Stöckli 2010). The population genetic consequences for the
Daphnia community of these changes have not yet been investigated. Based on our admittedly small sample size of hatchlings,
it appears that the lake was dominated by D. galeata xlongispina
hybrids during the time of high Pb pollution and is now dominated
by D. longispina. These hatched populations differed at more than
a genetic level: the population growth rates of the recent clones
were significantly lower than that of the old clones in the control
treatments, indicating innate differences in life history.
This historical snapshot—hybrids in the past and pure species
now—may have come about via at least three processes, with different implications for the evolutionary history that we can infer.
First, the past hybrid population may have been replaced by nonresident D. longispina from other lakes in the region. This would
imply that there is no evolutionary chain uniting our studied populations. Second, the past hybrid population may have dwindled
via competition or asexual incompetence, without gene flow to
the parental D. longispina population, again resulting in a temporal genetic discontinuity. Finally, the past hybrid population
may have continued to breed sexually with the parental species,
but with selection favoring those genotypes that are more nearly
D. longispina, until the population became indistinguishable from
“pure” parental. Because our species identifications are based on
allele frequencies at neutral markers, and are therefore probabilistic, we cannot distinguish between the above scenarios. We
note, however, that it is at least possible that the two populations studied here were linked through time and that the observed
lower Pb resistance of modern genotypes is in fact an adaptive
loss.
By what mechanism could Pb have exerted its selective pressure? Pb uptake in Daphnia may occur over water and food intake
(algae) and is expected to be proportional to the Pb concentrations
in the water column. Although we have no measurements from
the high-Pb time periods, aqueous measurements in Greifensee
from 1999–2000 were 0.02–0.06 μg/l Pb (Odzak et al. 2002). If
we assume that aquatic concentrations were proportional to those
in the sediments, the highest Pb concentrations would have been
in the range of 0.1– 0.3 μg/l (0.5–1.5 nM). Dissolved Pb in the
water column of these lakes is expected to be mostly bound to
dissolved organic matter (DOC = 2–4 mg/l) and to carbonate
(alkalinity = 2–4 mM) under oxic conditions. Dissolved Pb may
be taken up by algae and accumulated (Stewart et al. 2015). Once
Pb has entered a cell, its divalent cation masquerades as Ca in
metabolic pathways (Settle and Patterson 1980). Daphnia have
particularly high Ca demands, which, as an essential component
of the carapace, controls their growth and molting (Hessen and
Rukke 2000). Pb could disrupt these cellular processes and lead
to reduced growth and fecundity and early mortality. This interpretation suggests that Pb toxicity may be ameliorated by high
Ca levels. Addition of Ca and Na cations has been found to reduce the toxic impact of Ni and Cu in some highly polluted lakes
(Celis-Salgado et al. in press). If Pb toxicity is similarly affected,
the selection for Pb resistance inferred in our relatively hardwater lakes may have been even more pronounced in soft water
lakes. Additionally, aside from selection on the free-swimming
life stages, metal pollution has been found to decrease egg hatching rates and juvenile survivorship (Rogalski 2015). As our study
lakes are, at least partly, repopulated every spring from the egg
bank, selection on hatching success could provide a strong force
for the acquisition of resistance.
There are several possible metal detoxification mechanisms
that could plausibly account for Daphnia’s evolution of Pb
resistance. One possible mechanism by which Daphnia may adapt
to this threat is via the ATP-binding cassette (ABC) transporter
system (Sturm et al. 2009), which functions as a nonspecific defense against foreign materials. As Daphnia are known to have
a very high rate of gene duplication (Colbourne et al. 2011), we
suggest that Pb resistance may have been conferred by an increased copy number of ABC genes. This also suggests how the
presumably adaptive loss of resistance may have come about.
ABC transporters are expensive to manufacture and require ATP
for their operation. These costs could have selected against those
genotypes with a high number of gene copies. This suggestion
accords with the recent findings of Agra et al. (2011), who measured increased respiratory rates in Cu-resistant D. longispina,
even in unpolluted media. This increased respiration suggests
that resistance carries a measureable and innate energetic cost,
which over time would be selected against as the environmental
toxin concentration declines.
The population-level mechanisms of selection were not
detectable in this study. To wit, intense pollution may cause
“genetic erosion” (Ribeiro et al. 2012) of populations, in which
genetic diversity is reduced via the loss of sensitive genotypes.
Such erosion could have far-reaching consequences such as
lowered resilience to other perturbations. In contrast, if toxic
effects are sublethal, directional selection may occur via superior
reproduction of resistant genotypes. Although we detected
severe lethality of Pb in nonresistant modern genotypes, we
have no evidence that the Daphnia of the early 20th century
were so affected. Future paleo-ecological studies may effectively
examine the dynamics of genetic diversity over the last century,
but considering the strength and variety of environmental
perturbations that occurred during this time, it will be difficult to
implicate single stressors in major demographic changes.
EVOLUTION FEBRUARY 2016
405
PAT R I C K T U R KO E T A L .
These results are a first step toward understanding the evolutionary response of natural systems to the worldwide pulse of Pb.
We have shown that modern Daphnia species in two lakes have
reduced resistance to Pb as compared to those that were born
during the peak of pollution. From these data we infer that these
populations developed resistance from a nonresistant background,
but were unable to test this hypothesis due to a lack of hatching.
Future work could focus on establishing the resistance of pre-Pb
genotypes in lakes where hatching of older clones is possible, and
could very profitably apply the techniques of resurrection ecology
to other taxa.
ACKNOWLEDGMENTS
We thank D. Kistler and P. Ganesanandamoorthy for assistance with Pb
measurement, and N. Brede for all her support. Three anonymous reviewers greatly improved our analysis and this manuscript. This work
was supported by an SNF–DFG grant (SNF 310030L 135750) and an internal grant from Eawag (EvoChemTox). The authors declare no conflict
of interest.
DATA ARCHIVING
The doi for our data is 10.5061/dryad.1pd88.
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