Full text of DOI:10.1002/etc.5620210317

Environmental Toxicology and Chemistry, Vol. 21, No. 3, pp. 584–589, 2002
2002 SETAC
Printed in the USA
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GENETIC EFFECTS OF MERCURY CONTAMINATION ON AQUATIC SNAIL
POPULATIONS: ALLOZYME GENOTYPES AND DNA STRAND BREAKAGE
M
ICHAEL J. BENTON,*† MICHELLE L. MALOTT,‡ JAN TRYBULA,§ DEBORAH M. DEAN,§ and
S
HELDON I. GUTTMAN§
†Department of Environmental Health, East Tennessee State University, Johnson City, Tennessee 37614, USA
‡Department of Biomedical Sciences, Wright State University, Dayton, Ohio 45435, USA
§Department of Zoology, Miami University, Oxford, Ohio 45056, USA
(
Received 16 January 2001; Accepted 1 September 2001)
Abstract—Allozyme data and DNA strand break frequencies were compared among populations of Pleurocera canaliculatum from
five sites with varying mercury contamination on the North Fork Holston River (NFHR) in southwestern Virginia, USA. Allozyme
genotype frequencies for four loci were significantly different between populations from the three most highly contaminated sites
and those from two lesser contaminated sites. In addition, heterozygosity at three of these loci was significantly lower in the
populations from the most highly contaminated sites. The DNA strand break frequency was significantly correlated to whole-body
total mercury concentration in snails from three sites. These data add to the evidence supporting the use of DNA strand breakage
as an indicator of chemical contamination and the use of allozyme analysis as a marker of contamination and possible selection
for pollution resistance. However, the relationship between contaminant-induced changes in the genetic variation of enzymes of
central metabolism and the functionalities upon which selection for resistance may act remain unclear, and mechanisms other than
selection for resistance must be considered. Use of enzymes from other biochemical pathways may be appropriate for other species
or for those under other chemical pollution pressures.
Keywords—Mercury
Allozymes
DNA strand breakage
Pleurocera canaliculatum
INTRODUCTION
lihood of inherited genotoxic disease. Although DNA strand
breakage has been examined in some freshwater vertebrate
and marine invertebrate species [21–25], few studies have been
conducted on DNA strand breakage in freshwater inverte-
brates.
Populations of organisms occupying uncontaminated hab-
itats typically exhibit a relatively high level of genetic diver-
sity. In comparison, those populations occupying polluted hab-
itats, although often possessing increased tolerance to contam-
ination, also often exhibit decreased genetic diversity. This has
been demonstrated in studies of enzymes of central metabolism
in a variety of aquatic organisms, by using allozyme frequen-
cies as a measure of genetic diversity [1–10]. Although the
importance of physical barriers to gene flow and stochastic
events cannot be discounted, the argument has been made that
this reduced genetic variability is the result of selection for
resistant allozyme genotypes, as demonstrated in laboratory
bioassays with many of the same enzyme systems and animal
species [11–18]. Of concern is that such genetically con-
strained populations may be more sensitive to other stressors
and so less able to adapt to future environmental conditions,
and that selection for contaminant resistance may be related
to negative effects on other characteristics [16,19].
Breakage of DNA strands in aquatic organisms has gained
increasing attention in recent years as a biological marker of
environmental genotoxicity. When strand breakage is not re-
paired, it may become fixed in the genetic material and cause
genotoxic disease syndrome, including protein dysfunction,
oncogene activation, chromosomal aberrations, or other del-
eterious manifestations that reduce the fitness of the affected
organism [20]. The DNA damage in gametic cells may be
passed down to subsequent generations, increasing the like-
In this study, freshwater snails from both clean and con-
taminated sites on a mercury-impacted river were examined
for differences in population genetic structure, as measured by
allozyme genotype frequencies, and in relative levels of DNA
damage, as measured by strand breakage. Heavy metals in
general have been shown to be selective agents for resistant
genotypes [26,27], and mercury in particular is known to be
cytotoxic and genotoxic in bacterial, plant, and mammalian
cells [28,29]. Snails were considered to be ideal organisms for
this investigation because of their distribution over the river’s
entire course; their limited locomotion, which decreases the
likelihood of immigration to and emigration from sites of in-
terest; and their benthic locomotory and feeding habits, which
put them in intimate contact with mercury-contaminated sed-
iments and periphyton.
MATERIALS AND METHODS
Study site
The North Fork Holston River (NFHR) flows 209 km from
southwestern Virginia to northeastern Tennessee, USA. In-
dustrial activities at a now-defunct chloralkali plant in Salt-
ville, Virginia, USA, from 1951 to 1972 resulted in severe
mercury contamination of the NFHR. In addition, net flux of
mercury from adjacent disposal sites into the river has been
a continuing source of contamination [30]. Recent investiga-
tions have found dry-weight mercury concentrations in the
* To whom correspondence may be addressed
(bentonmi@mnstate.edu). The current address of M.J. Benton is the
Department of Biology, Minnesota State University-Moorhead, Moor-
head, MN, USA.
immediate vicinity of Saltville as high as 14 ␮g/g in sediment,
584

Mercury pollution: Effect on allozymes and DNA damage
Environ. Toxicol. Chem. 21, 2002
585
Table 1. Total mercury (Hg) concentrations (dry wt) for whole-body
Pleurocera canaliculatum from five sites on the North Fork Holston
River, Saltville, Virginia, USA
1.4
2.5
␮
g/g in periphyton, 1.1
␮g/g in benthic invertebrates, and
␮
g/g in fishes, with decreasing concentrations with distance
downstream and generally very low concentrations upstream
[31,32].
Total Hg
Site
(
␮
g/g dry wt)
Study organism
1
2
3
4
5
0.695
0.678
4.257
2.984
2.282
Snails were collected by hand at five sites on the NFHR.
Sites 1 and 2 were 17 and 2 km, respectively, upstream of the
original impact site in Saltville, and sites 3, 4, and 5 were 5,
27, and 48 km downstream of Saltville, respectively. Snails
and all other benthic invertebrate animals are rare in the im-
mediate area of the impact, almost certainly because of the
extremely high sediment mercury load that remains there. Col-
lected snails were immediately placed on dry ice, transported
handheld homogenizer. Each snail was lysed in 1 ml Triton
lysis buffer (100 mM Tris-Cl, 100 mM ethylenediaminete-
traacetic acid [EDTA], 1% Triton X-100 [Sigma], 30
RNase A) followed by proteinase K (800 g/ml) digestion for
6 h at 37 C. Genomic DNA was extracted from the lysed tissue
with a series of phenol–chloroform extractions, with wide-bore
pipet tips to avoid artificial shearing of the DNA. The DNA
then was precipitated with a one-tenth volume of 5 M sodium
acetate and 95% ethanol, pelleted by centrifugation, and re-
suspended in tris-EDTA (10 mM Tris-Cl, 1 mM EDTA, pH
␮g/ml
to the laboratory, and stored at
Ϫ70ЊC. Representative samples
␮
from each site have been identified as Pleurocera canalicu-
latum (Say) (The Academy of Natural Sciences of Philadelphia
[ANSP], Philadelphia, PA, USA, ANSP A13029–13033; G.M.
Davis, M.A. Garback, personal communication). Snails from
all five sites were used in the allozyme study; snails from sites
1, 3, and 4 were used in the DNA strand breakage study.
Њ
Mercury analysis
7.4) overnight at 4ЊC. The DNA samples from all snails used
in this study were prepared simultaneously.
Equal quantities of DNA from 18 individuals from sites 1,
3, and 4 were loaded into the wells of 0.7% neutral-pH agarose
Whole-body tissue samples for P. canaliculatum from sites
1, 3, and 4 were digested and analyzed for total mercury burden
by cold vapor atomic absorption spectrophotometry (Varian
Spectra AA/20 with model 76 Cold Vapor Generation Acces-
sory, Varian, Palo Alto, CA, USA), according to the methods
of Evans et al. [33]. Mercury analyses for P. canaliculatum
from sites 2 and 5 were recently performed by identical meth-
ods for another study [32] and results of that study were used
here. All tissue mercury concentrations are expressed as dry
weight values.
gels. Five hundred micrograms of
␭ Hind III marker DNA
(Gibco-BRL, Carlsbad, CA, USA) was loaded into two wells
on each gel as a quantitative standard. After electrophoresis
(50 V, 5 h), gels were stained in 10 ␮g/ml ethidium bromide
and photographed under ultraviolet (UV) light. Relative levels
of DNA strand breakage were assessed with NIH Image, a
software package that allows quantification of DNA fluores-
cence [40]. By obtaining fluorescence values for the band of
intact DNA in a given gel lane (I) and for the smear of frag-
mented DNA below that band (F), a unitless measure of DNA
strand break frequency (B) can be calculated by the equation:
Allozyme electrophoresis
One hundred snails from each site (except site 3, where
n
ϭ
110) were scanned for enzymatic activity at 13 polymorphic
B
ϭ F/I. Because electrophoresis was performed at neutral pH,
loci with 10 buffer systems. Standard methodologies for starch
gel and cellulose acetate plate electrophoresis were followed,
with minor modifications [34–36]. Snail tissue was removed
from shells and ground in equal volumes of 2% 2-phenoxy-
ethanol, then centrifuged at 12,000 rpm. Starch gel electro-
phoresis samples consisted of tissue supernatant applied to
filter paper wicks, which were loaded into 13% hydrolyzed
potato starch gels (Sigma Chemical, St. Louis, MO, USA).
For cellulose acetate electrophoresis, sample supernatant was
applied by double-striking Titan-III plates (Helena Laborato-
ries, Beaumont, TX, USA).
Allozymes were scored so that the farthest-migrating anodal
allele was designated A, the next farthest-migrating allele was
designated B, and so on. For multiple loci, the farthest anodal
locus was designated 1, the next farthest was designated 2,
and so forth.
this is a measure of double-strand breakage only. Snail DNA
that had been deliberately sheared by passing it repeatedly
through a 21-gauge surgical needle was loaded on preliminary
gels. These samples allowed us to confirm that the ratio F/I
did indeed increase proportionately with increased DNA break-
age. The relationship between mean whole-body total mercury
concentration and B was tested by simple correlation analysis
[37].
RESULTS
Mercury analysis
Whole-body dry-weight total mercury concentrations for P.
canaliculatum from the five study sites are given in Table 1.
Allozyme electrophoresis
Allozyme allele, genotype, and heterozygote frequencies
were tested for differences among sites, and for differences
between upstream (site 1) and downstream (sites 3 and 4
pooled) locales, with the chi-square test [37]. Allele frequen-
cies were analyzed for relatedness between sites by Nei’s un-
biased genetic distance (D) [38], with BIOSYS-1 [39].
Electrophoretic analysis of 13 enzyme systems revealed 3
alleles and 4 genotypes of acid phosphatase-3 (ACP-3), 3 al-
leles and 6 genotypes of alkaline phosphatase-1 (ALP-1), 4
alleles and 7 genotypes of esterase-1 (EST-1), 3 alleles and 4
genotypes of esterase-2 (EST-2), 2 alleles and 3 genotypes of
esterase-3 (EST-3), 3 alleles and 4 genotypes of esterase-4
(EST-4), 2 alleles and 3 genotypes of UV esterase-4 (EUV-4),
2 alleles and 3 genotypes of glucose phosphate isomerase-1
(GPI-1), 2 alleles and 3 genotypes of malate dehydrogenase-
2 (MDH-2), 4 alleles and 5 genotypes of mannose phosphate
Analysis of DNA strand breakage
Snails were removed intact from their shells with forceps
and macerated in ice-cold phosphate-buffered saline with a

586
Environ. Toxicol. Chem. 21, 2002
M.J. Benton et al.
Table 2. Allele frequencies of all electrophoretically examined loci
for Pleurocera canaliculatum from five sites on the North Fork
Holston River, Saltville, Virginia, USA
among the five collection sites for EST-2, MDH-2, PEP-3, and
2
TPI-1 (
␹
;
p
Ͻ
0.05, 0.01, 0.005, and 0.05, respectively). Ge-
notype frequency differences among sites for PEP-2 were no-
2
table but not statistically significant (␹ ; p ϭ 0.057). Allozyme
Allele
genotype frequency differed significantly between upstream
Locusa
n
A
B
C
D
and downstream locales for MDH-2, PEP-2, PEP-3 and TPI-
2
1 (
4).
␹
;
p
Ͻ
0.05, 0.001, 0.001, and 0.05, respectively; Table
ACP-3
ALP-1
EST-1
EST-2
EST-3
EST-4
EUV-4
GPI-1
MDH-2
MPI-1
PEP-2
PEP-3
TPI-1
386
445
428
427
426
445
460
499
466
479
368
427
430
0.951
0.875
0.412
0.161
0.015
0.002
0.019
0.141
0.836
0.006
0.094
0.306
0.996
0.005
0.033
0.087
0.834
0.985
0.991
0.981
0.859
0.164
0.691
0.905
0.243
0.004
0.024
0.092
0.500
0.005
—
0.006
—
—
—
—
0.001
—
—
—
—
—
—
0.001
—
The frequency of heterozygous genotypes differed signif-
icantly among the five collection sites at the EST-2, ALP-1,
2
TPI-1, MDH-2, and PEP-2 loci (␹ ; p ϭ 0.013, 0.018, 0.045,
0.004, and 0.042, respectively). The frequency of heterozy-
gotes was significantly different between upstream and down-
2
—
stream locales for MDH-2, PEP-2, and TPI-1 (
␹
;
p
ϭ
0.020,
0.301
0.001
0.168
—
0.002, and 0.021, respectively), and in each case heterozygote
frequency was lower downstream than upstream (3, 9, and
12%, respectively; Table 5).
0.284
—
Nei’s (1978) unbiased genetic distance values (D) for all
possible population pairs ranged from 0.002 to 0.007 (Table
6).
a ACP-3
ϭ
acid phosphatase-3; ALP-1
esterase-2; EST-3
ultraviolet esterase-4; GPI-1
ϭ
alkaline phosphatase-1; EST-
1
ϭ
esterase-1; EST-2
esterase-4; EUV-4
phate isomerase-1; MDH-2
ϭ
ϭ
esterase-3; EST-4
ϭ
ϭ
ϭ
glucose phos-
ϭ
malate dehydrogenase-2; MPI-1
ϭ
ϭ
Analysis of DNA strand breakage
mannose phosphate isomerase-1; PEP-2
ϭ peptidase-2; PEP-3
peptidase-3; TPI-1 triose phosphate isomerase-1.
ϭ
Putative double-strand breakage frequency (B) in P. can-
aliculatum was significantly and strongly linearly correlated
to mean whole-body total mercury burden (simple correlation,
isomerase-1 (MPI-1), 3 alleles and 4 genotypes of peptidase-
2 (PEP-2), 4 alleles and 10 genotypes of peptidase-3 (PEP-3),
and 2 alleles and 2 genotypes of triose phosphate isomerase-1
(TPI-1) (Table 2).
F
ϭ ϭ 0.532, p Ͻ 0.001; Table 7 and Fig. 1).
27.278, r2
DISCUSSION
Four enzyme loci (TPI-1, MDH-2, PEP-2, and PEP-3) ex-
Allozyme frequencies
hibited significant differences in both allele and genotype fre-
quencies between the two relatively clean upstream sites and
the three more heavily mercury-contaminated downstream
sites. Heterozygote frequency was significantly lower down-
stream for three of these loci (TPI-1, MDH-2, and PEP-2).
Two downstream sites also exhibited a greater level of putative
double-strand DNA breaks than a single upstream site, and
strand break frequency was highly correlated to mean whole-
body total mercury concentration.
Allele frequencies differed significantly among the five col-
lection sites for ALP-1, EST-2, EST-4, GPI-1, MDH-2, MPI-
2
1, PEP-2, PEP-3, and TPI-1 (
␹
;
p
Ͻ
0.05, 0.005, 0.05, 0.05,
0.005, 0.05, 0.01, 0.005, and 0.05, respectively). When col-
lection sites were grouped as upstream and downstream lo-
cales, allele frequencies differed significantly between locales
2
for TPI-1, MDH-2, PEP-2, and PEP-3 (
0.005, and 0.005, respectively; Table 3).
Allozyme genotype frequencies differed significantly
␹
;
p
Ͻ
0.025, 0.005,
The shift in allozyme genotype frequency from clean sites
Table 3. Allele frequencies of four electrophoretically examined loci for Pleurocera canaliculatum from five sites on the North Fork Holston
River, Saltville, Virginia, USA
Locale
Site
Upstream
(sites 1, 2)
Downstream
(sites 3–5)
Locusa
1
2
3
4
5
MDH-2
n
ϭ
79
n
ϭ
90
n
ϭ
98
n
ϭ
99
n
ϭ
100
n
ϭ
169
n ϭ 297
A
B
0.722
0.278
0.856
0.144
0.883
0.117
0.869
0.131
0.850
0.150
0.795
0.205
0.865
0.135
PEP-2
n
ϭ
64
n
ϭ
59
n
ϭ
86
n
ϭ
80
n
ϭ
79
n
ϭ
123
n ϭ 245
A
B
C
0.156
0.844
0.000
0.144
0.856
0.000
0.058
0.942
0.000
0.050
0.944
0.006
0.063
0.937
0.000
0.15
0.85
0.00
0.057
0.941
0.002
PEP-3
n
ϭ
80
n
ϭ
84
n
ϭ
83
n
ϭ
79
n
ϭ
91
n
ϭ
164
n ϭ 263
A
B
C
D
0.338
0.015
0.275
0.238
0.387
0.304
0.137
0.173
0.215
0.242
0.156
0.387
0.335
0.228
0.101
0.335
0.253
0.291
0.170
0.286
0.365
0.230
0.205
0.200
0.265
0.255
0.145
0.335
TPI-1
n
ϭ
100
n
ϭ
100
n
ϭ
60
n
ϭ
100
n
ϭ
70
n
ϭ
200
n ϭ 230
A
B
0.995
0.005
0.980
0.020
1.000
0.000
1.000
0.000
1.000
0.000
0.988
0.012
1.000
0.000
a MDH-2
ϭ
malate dehydrogenase-2; PEP-2
ϭ
peptidase-2; PEP-3
ϭ
peptidase-3; TPI-1 ϭ triose phosphate isomerase-1.

Mercury pollution: Effect on allozymes and DNA damage
Environ. Toxicol. Chem. 21, 2002
587
Table 4. Genotype frequencies of four electrophoretically examined loci for Pleurocera canaliculatum from five sites on the North Fork Holston
River, Saltville, Virginia, USA
Locale
Site
Upstream
(sites 1, 2)
Downstream
(sites 3–5)
Locusa
1
2
3
4
5
MDH-2
n
ϭ
79
n
ϭ
90
n
ϭ
98
n
ϭ
99
n
ϭ
100
n
ϭ
169
n
ϭ
297
AA
AB
BB
0.532
0.380
0.089
0.767
0.178
0.056
0.796
0.174
0.031
0.779
0.182
0.040
0.750
0.200
0.050
0.66
0.27
0.07
0.77
0.19
0.04
PEP-2
n
ϭ
64
n
ϭ
59
n
ϭ
86
n
ϭ
80
n
ϭ
79
n
ϭ
123
n ϭ
245
AA
AB
BB
BC
0.047
0.219
0.734
0.000
0.034
0.220
0.746
0.000
0.012
0.093
0.895
0.000
0.013
0.075
0.900
0.013
0.000
0.127
0.873
0.000
0.04
0.22
0.74
0.00
0.008
0.098
0.890
0.004
PEP-3
n
ϭ
80
n
ϭ
84
n
ϭ
83
n
ϭ
79
n
ϭ
91
n
ϭ
164
n ϭ 263
AA
AB
AC
AD
BB
BC
BD
CC
CD
DD
0.163
0.050
0.125
0.175
0.063
0.088
0.038
0.150
0.038
0.113
0.238
0.083
0.107
0.107
0.179
0.083
0.083
0.036
0.012
0.071
0.075
0.075
0.032
0.172
0.140
0.065
0.065
0.065
0.086
0.226
0.203
0.089
0.038
0.139
0.089
0.051
0.139
0.025
0.063
0.165
0.132
0.088
0.066
0.088
0.187
0.011
0.110
0.088
0.088
0.143
0.20
0.07
0.12
0.14
0.12
0.09
0.06
0.09
0.02
0.09
0.13
0.09
0.05
0.13
0.14
0.04
0.10
0.06
0.08
0.18
TPI-1
n
ϭ
100
n
ϭ
100
n
ϭ
60
n
ϭ
100
n
ϭ
70
n
ϭ
200
n ϭ 230
AA
AB
0.990
0.010
0.960
0.040
1.000
0.000
1.000
0.000
1.000
0.000
0.975
0.025
1.00
0.00
a MDH-2
ϭ
malate dehydrogenase-2; PEP-2
ϭ
peptidase-2; PEP-3
ϭ
peptidase-3; TPI-1 ϭ triose phosphate isomerase-1.
to contaminated sites is consistent with an evergrowing body
of evidence for selection by anthropomorphic stressors at sin-
gle polymorphic loci for enzymes of central metabolism [1–
10]. Further evidence of such selection for pollution tolerance
comes from studies that have demonstrated differential sur-
vival among allozyme genotypes in aquatic animals under tox-
icity bioassay conditions [11–18]. Although the functional re-
lationship of certain allozyme genotypes to fitness under con-
taminated conditions remains unclear, Eanes [41] pointed out
that several studies of specific polymorphisms in enzymes of
central metabolism in invertebrate and vertebrate animals have
indeed demonstrated an association with functional variation.
However, mechanisms other than selection for resistance
may play a role in the kind of results presented here. Stochastic
events, such as genetic drift or genetic bottlenecks, may have
been important factors in the reduced genetic variation ob-
served downstream, although no evidence exists of a decrease
in P. canaliculatum density at any downstream NFHR study
site. The possible influence of interpopulational distance on
gene flow also cannot be discounted, particularly in organisms
such as snails, which have relatively limited mobility and dis-
persal capabilities, but no evidence was found in this study
that nearest-neighbor populations are more closely related to
one another than more widely separated ones (Table 6). Genetic
distance values for all population pairs were low (D
Յ 0.007),
and no relationship between interpopulational geographic and
genetic distances could be discerned. In fact, the greatest mea-
sured genetic distance value (0.007) was between the popu-
lation from the least contaminated site (site 1) and those from
the two most contaminated sites (sites 3 and 4). This suggests
that geographic separation alone has not resulted in a mea-
surable level of reproductive isolation among populations.
More likely to be a factor, perhaps, may be reduced gene flow
between upstream and downstream populations across the con-
taminant barrier created by the high sediment mercury levels
at the original impact site. The benthic zone immediately ad-
jacent to the old chloralkali plant and waste ponds remains
extremely contaminated and markedly depauperate [32]. If
Table 5. Frequency of heterozygous genotypes for Pleurocera
canaliculatum from upstream (sites 1, 2) and downstream (sites 3–
5) locales on the North Fork Holston River, Saltville, Virginia, USA
Locale
Table 6. Nei’s unbiased genetic distance (D) for populations of
Pleurocera canaliculatum from five sites on the North Fork Holston
River, Saltville, Virginia, USA. Sites 1 and 2 are upstream of mercury
impact; sites 3, 4, and 5 are downstream of mercury impact
Upstream
(sites 1, 2)
Downstream
(sites 3–5)
Locusa
MDH-2
PEP-2
TPI-1
0.27
0.18
Population
1
2
3
4
5
(
(
(
n
n
n
ϭ
0.22
ϭ
0.03
169)
(
(
(
n
n
n
ϭ
297)
0.10
1
2
3
4
5
—
0.005
—
0.007
0.006
—
0.007
0.002
0.004
—
0.004
0.003
0.003
0.003
—
123)
ϭ
245)
0.00
ϭ 230)
ϭ
200)
a MDH-2
triose phosphate isomerase-1.
ϭ malate dehydrogenase-2; PEP-2 ϭ peptidase-2; TPI-1 ϭ

588
Environ. Toxicol. Chem. 21, 2002
M.J. Benton et al.
Table 7. Whole-body total mercury burden and putative DNA double-
strand breakage (B) in Pleurocera canaliculatum from one upstream
and two downstream sites on the North Fork Holston River, Saltville,
Virginia, USA
DNA synthesomes has shown that the mutagenicity apparently
is due to inhibition of DNA replication, DNA polymerase ac-
tivity, and DNA replication fidelity [45].
The sum of these genetic studies demonstrates that popu-
lations exposed to sublethal concentrations of chemicals such
as mercury, which is both cytotoxic and mutagenic, may face
multiple and distinct, but interrelated and interdependent, chal-
lenges to their viability and persistence. The first challenge is
that of reduced genetic variability, which has been found in a
number of field studies. When using enzymes of central me-
tabolism as biomarkers, contaminants often seem to select for
individuals of the most resistant allozyme genotypes, resulting
in the decrease or loss of certain alleles or genotypes from the
population, which is manifested by a decrease in allozyme
genotype heterozygosity. Because enzymes of central metab-
olism have been most commonly studied, it has been postulated
(but not well demonstrated) that the mechanism of selection
is the elimination of allozyme genotypes associated with the
greatest contaminant-induced enzyme inhibition, and, there-
fore, the poorest cellular respiration capabilities under con-
taminated conditions. Thus, enzymes of central metabolism
may well serve as effective biomarkers of pollution in many
species under many different types of chemical contamination.
However, several studies suggested that selection in some
species under some contamination regimes actually may result
in increased heterozygosity [3–6,15,18,46]. Moreover, a recent
study by Theodorakis and Shugart [46] made it clear that not
only may enzymes of biochemical pathways outside central
metabolism be important in selection in polluted environments
(and therefore as biomarkers), but that allozyme effects may
be linked to DNA strand breakage. In a study of strand break-
Total Hg
Site
(
␮
g/g dry wt)
B
Upstream
1
0.695
0.2810
Downstream
3
4
4.257
2.984
0.7884
0.5822
some level of genetic exchange among NFHR P. canalicu-
latum had been maintained either by downstream drift or by
upstream movement before the introduction of mercury, the
contaminant barrier between upstream and downstream study
sites may have altered this.
The correlation between whole-body total mercury con-
centration and level of putative DNA strand breakage in P.
canaliculatum is consistent with the results of Meyers-Scho¨ne
et al. [21] and Sugg et al. [22], who reported higher frequencies
of DNA breakage in freshwater turtles (Tachemys scripta and
Chelydra serpentina) and largemouth bass (Micropterus sal-
moides), respectively, from sites contaminated by mercury in
combination with radionuclides than in the same species from
reference sites. The data also are consistent with studies of
redbreast sunfish (Lepomis auritis) and marine mussels (My-
tilus edulis), which found higher levels of strand breakage in
individuals occupying sites most heavily polluted with mer-
cury in combination with other heavy metals and organic pol-
lutants [24,42].
age and allozyme genotype in mosquitofish (Gambusia affinis
)
exposed to radionuclide contamination, the frequency of nu-
cleoside phosphorylase heterozygotes was higher at contam-
inated sites and heterozygotes had fewer DNA strand breaks
than did homozygotes. Coupled with human DNA synthesome
research that demonstrates the inhibitory action of mercury on
enzymes of DNA replication and repair [45], this provides
new avenues of thought and research about the mechanisms
of selection under contaminated conditions. The mode of se-
lection in the presence of a chemical with multisystemic tox-
icities will be dependent upon the relative selective pressures
those toxicities impart upon the individual. Therefore, the val-
ue of a given enzyme system as a biomarker of pollution will
be similarly dependent. Additional research involving enzymes
of central metabolism, DNA replication and repair, signal
transduction, neurotransmitter synthesis and hydrolysis, and
other biochemical pathways will help clarify the relationship
between genetic markers of pollution and the functionalities
with which they are associated.
The genotoxicity of various mercury forms in bacterial,
plant, and animal systems is well documented. Mercury ex-
posure has been shown to be associated with increased levels
of both single- and double-strand DNA breaks [28,29,43], and
mercury compounds have been shown to enhance photocleav-
age of DNA [44]. The mechanism of mercury’s mutagenicity
is difficult to determine in living organisms, but work with
REFERENCES
1. Kopp R, Guttman SI, Wissing TE. 1992. Genetic indicators of
environmental stress in central mudminnow (Umbra limi) pop-
ulations exposed to acid deposition in the Adirondack Mountains.
Environ Toxicol Chem 11:665–676.
2. Gillespie RB, Guttman SI. 1989. Effects of contaminants on the
frequencies of allozymes in populations of the central stoneroller.
Environ Toxicol Chem 8:309–317.
3. Gillespie RB, Guttman SI. 1993. Correlations between water qual-
ity and frequencies of allozyme genotypes in spotfin shiner (No-
tropis spilopteris) populations. Environ Pollut 81:147–150.
4. Benton MJ, Diamond SA, Guttman SI. 1994. A genetic and mor-
phometric comparison of Helisoma trivolvis and Gambusia hol-
brooki from clean and contaminated habitats. Ecotoxicol Environ
Saf 29:20–37.
Fig. 1. Putative DNA strand break frequency (B) as a function of
whole-body total mercury concentration for Pleurocera canalicula-
tum from three sites on the North Fork Holston River, Saltville, Vir-
ginia, USA. Site 1 is upstream of mercury impact; sites 3 and 4 are
downstream of mercury impact.

Mercury pollution: Effect on allozymes and DNA damage
Environ. Toxicol. Chem. 21, 2002
589
5. Fore´ SA, Guttman SI, Bailer AJ, Altfater DJ, Counts BV. 1995.
Exploratory analysis of population genetic assessment as a water
KC, Adams SM. 2000. Relationship between genotoxicity, mu-
tagenicity, and fish community structure in a contaminated stream.
J Aquat Ecosyst Stress Recov 7:131–143.
quality indicator. I. Pimephales notatus. Ecotoxicol Environ Saf
30:24–35.
25. Mitchelmore CL, Birmelin C, Livingstone DR, Chipman JK.
1998. Detection of DNA strand breaks in isolated mussel (Mytilus
edulis L.) digestive gland cells using the ‘‘Comet’’ assay. Eco-
toxicol Environ Saf 41:51–58.
26. Guttman SI. 1994. Population genetic structure and ecotoxicol-
ogy. Environ Health Perspect 102:97–100.
27. Gillespie RB, Guttman SI. 1999. Chemical-induced changes in
the genetic structure of populations: Effects on allozymes. In
Forbes V, ed, Genetics and Ecotoxicology. Taylor & Francis,
London, UK, pp 55–77.
28. Ariza ME, Holliday J, Williams MV. 1994. Mutagenic effect of
mercury (II) in eukaryotic cells. In Vivo 8:559–564.
29. Stohs SJ, Bagchi D. 1995. Oxidative mechanisms in the toxicity
of metal ions. Free Radic Biol Med 18:321–336.
6. Fore´ SA, Guttman SI, Bailer AJ, Altfater DJ, Counts BV. 1995.
Exploratory analysis of population genetic assessment as a water
quality indicator. I. Campostoma anomalum. Ecotoxicol Environ
Saf 30:36–46.
7. Beaumont AR, Toro JE. 1996. Allozyme genetics of Mytilus ed-
ulis subjected to copper and nutritive stress. J Mar Biol Assoc
UK 76:1061–1071.
8. Roark S, Brown K. 1996. Effects of metal contamination from
mine tailings on allozyme distribution of populations of Great
Plains fishes. Environ Toxicol Chem 15:921–927.
9. Heithaus MR, Laushman RH. 1997. Genetic variation and con-
servation of stream fishes: Influence of ecology, life history, and
water quality. Can J Fish Aquat Sci 54:1822–1836.
10. Tanguy A, Fernandez CN, Marhic A, Moraga D. 1999. Effects
of an organic pollutant (tributyltin) on genetic structure in the
30. Turner RR, Lindberg SE. 1978. Behavior and transport of mercury
in river–reservoir system downstream of inactive chloralkali
plant. Environ Sci Technol 12:918–923.
Pacific oyster Crassostrea gigas. Mar Pollut Bull 38:550–559.
11. Diamond SA, Newman MC, Mulvey M, Dixon PM, Martinson
D. 1989. Allozyme genotypes and time to death of mosquitofish,
Gambusia affinis (Baird and Girard), during acute exposure to
inorganic mercury. Environ Toxicol Chem 8:613–622.
12. Benton MJ, Guttman SI. 1990. Relationship of allozyme genotype
to survivorship of mayflies (Stenonema femoratum) exposed to
copper. J North Am Benthol Soc 9:271–276.
13. Benton MJ, Guttman SI. 1992. Allozyme genotype and differ-
ential resistance to mercury pollution in the caddisfly, Nectop-
syche albida. I. Single-locus genotypes. Can J Fish Aquat Sci
49:142–146.
14. Benton MJ, Guttman SI. 1992. Allozyme genotype and differ-
ential resistance to mercury pollution in the caddisfly, Nectop-
syche albida. II. Multilocus genotypes. Can J Fish Aquat Sci 49:
147–149.
15. DeNicola M, Gambardella C, Guarino SM. 1992. Interactive ef-
fects of cadmium and zinc pollution on PGI and PGM polymor-
31. Hildebrand SG, Strand RH, Huckabee JW. 1980. Mercury ac-
cumulation in fish and invertebrates of the North Fork Holston
River, Virginia and Tennessee. J Environ Qual 9:393–400.
32. Dye SK. 1999. Mercury contamination in abiotic and biotic com-
partments in the North Fork of the Holston River, VA/TN. MS
thesis. East Tennessee State University, Johnson City, TN, USA.
33. Evans SJ, Johnson MS, Leah RT. 1986. Determination of fish
tissue, a rapid, automated technique for routine analysis. AA-60.
Varian Techtron Pty., Mulgrave, Victoria, Australia.
34. Selander RK, Smith MH, Yang SY, Johnson WE, Gentry JB.
1971. Biochemical polymorphism and systematics in the genus
Peromyscus. I. Variation in the old field mouse (Peromyscus po-
lionotus). Studies in genetics VI. Univ. Tex. Publ. 7103:49–90.
35. Harris H, Hopkinson DA. 1976. Handbook of Enzyme Electro-
phoresis in Human Genetics. North Holland, Amsterdam, The
Netherlands.
36. Hebert PDN, Beaton M. 1989. Methodologies for Allozyme Anal-
ysis Using Cellulose Acetate Electrophoresis. Helena Labora-
tories, Beaumont, TX, USA.
phisms in Idotea baltica. Mar Pollut Bull 24:619–621.
16. Schlueter MA, Guttman SI, Oris JT, Bailer AJ. 1995. Survival of
copper-exposed juvenile fathead minnows (Pimephales prome-
las) among allozyme genotypes. Environ Toxicol Chem 14:1727–
1734.
17. Snyder CD, Hendricks AC. 1997. Genetic responses of Isonychia
bicolor (Ephemeroptera:Isonychiidae) to chronic mercury pol-
lution. J North Am Benthol Soc 16:651–663.
18. Duan Y, Guttman SI, Oris JT, Huang X, Burton GA. 2000. Ge-
notype and toxicity relationships among Hyalella azteca: II.
Acute exposure to fluoranthene-contaminated sediment. Environ
Toxicol Chem 19:1422–1426.
19. Schlueter MA, Guttman SI, Oris JT, Bailer AJ. 1997. Differential
survival of fathead minnows, Pimephales promelas, as affected
by copper exposure, prior population stress, and allozyme ge-
notype. Environ Toxicol Chem 16:939–947.
20. Kurelec B. 1993. The genotoxic disease syndrome. Mar Environ
Res 35:341–348.
21. Meyers-Scho¨ne L, Shugart LR, Beauchamp JJ, Walton BT. 1993.
Comparison of two freshwater turtle species as monitors of ra-
dionuclide and chemical contamination: DNA damage and residue
analysis. Environ Toxicol Chem 12:1487–1496.
22. Sugg DW, Chesser RK, Brooks JA, Grasman BT. 1995. The as-
sociation of DNA damage to concentrations of mercury and ra-
diocesium in largemouth bass. Environ Toxicol Chem 14:661–
668.
23. Theodorakis CW, Blaylock BG, Shugart LR. 1996. Genetic eco-
toxicology I: DNA integrity and reproduction in mosquitofish
exposed in situ to radionuclides. Ecotoxicology 5:1–14.
24. Theodorakis CW, Swartz CD, Rogers WJ, Bickham JW, Donnelly
37. SAS Institute. 1985. SAS User’s Guide: Statistics. Cary, NC,
USA.
38. Nei M. 1978. Estimation of average heterozygosity and genetic
distance from a small number of individuals. Genetics 89:583–
590.
39. Swofford DL, Selander RK. 1981. BIOSYS-1: A FORTRAN pro-
gram for the comprehensive analysis of electrophoretic data in
population genetics and systematics. J Hered 72:281–283.
40. Rasband WS, Bright DS. 1995. NIH Image: A public domain
image processing program for the Macintosh. Microbeam Anal
Soc J 4:137–149.
41. Eanes WF. 1999. Analysis of selection on enzyme polymor-
phisms. Annu Rev Ecol Syst 30:301–326.
42. Steinart SA, Streib-Montee R, Leather JM, Chadwick DB. 1998.
DNA damage in mussels at sites in San Diego Bay. Mutat Res
399:65–85.
43. Bucio L, Garcia C, Souza V, Hernandez E, Gonzalez C, Betan-
court M, Gutierrez-Ruiz C. 1999. Uptake, cellular distribution
and DNA damage produced by mercuric chloride in a human fetal
hepatic cell line. Mutat Res 423:65–72.
44. Tabata M, Sarker AK, Watanabe K. 1998. Enhanced DNA pho-
tocleavage by a subnanomolar amount of mercury (II) porphyrin.
Chem Lett 1998:325–326.
45. Sekowski JW, Malkas LJ, Wei Y, Hickey RJ. 1997. Mercuric ion
inhibits the activity and fidelity of the human cell DNA synthe-
some. Toxicol Appl Pharmacol 145:268–275.
46. Theodorakis CW, Shugart LR. 1998. Genetic ecotoxicology III:
The relationship between DNA strand breaks and genotype in
mosquito fish exposed to radiation. Ecotoxicology 7:227–236.